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The entire process depicted in Figure 15.2.1 is equivalent to the following set of equations.
$\frac{dS}{dt}\,=\,b(S+I+R)\,-\beta\,I\frac{S}{S+I+R}\,-\delta\,S$
$\frac{dI}{dt}\,=\beta\,I\frac{S}{S+I+R}\,-\gamma\,I\,-\alpha\,I$
$\frac{dR}{dt}\,=\gamma\,I\,-\delta\,R$
At the left in each equation is the net rate of change of each box, accounting for all arrows transferring individuals out of one box and into another. Again, $S$ is the density of susceptible individuals, $I$ the density of infected individuals, and $R$ the density of recovered individuals. Note that the terms are balanced— the term $\gamma\,I$, for example, representing individuals entering the recovered box in the last equation, is balanced by the complementary term $-\gamma\,I$, leaving the infected box in the middle equation.
The SIR model is another “macroscale model.” With recent changes in computation, “microscale models,” processing tens or hundreds of millions of individual hosts, are becoming more widely used. They can reliably take you beyond what purely mathematical formulations can do. More about them in later chapters.
15.04: The SI Equations
SIR is also a “frequency-dependent” model, at one end of a spectrum which has “density-dependent” models at the other end. Frequency dependence approximates situations in which infection propagules are limited, while density dependence approximates situations in which potential victims are limited.
But do not be concerned with the full SIR model for now. We shall simplify it here to reveal its basic properties. First, suppose there is no recovery— this is an incurable disease that, once contracted, stays with its victim forever. Many viral diseases approximate this situation—herpes and HIV, for example. In gray below are all the terms that will drop out if there is no recovery.
$\frac{dS}{dt}\,=\,b(S+I\color{grey}{+R}\color{black}\,)\,-\beta\,I\frac{S}{S+I\color{grey}{+R}}\,-\delta\,S$
$\frac{dI}{dt}\,=\beta\,I\frac{S}{S+I\color{grey}{+R}}\,\color{grey}{-\gamma\,I}\color{black}{\,-\alpha\,I}$
$\color{grey}{\frac{dR}{dt}\,=\gamma\,I\,-\delta\,R}$
Removing those terms gives an “SI” model.
$\frac{dS}{dt}\,=\,b(S+I)\,-\beta\,I\frac{S}{S+I}\,-\delta\,S$
$\frac{dI}{dt}\,=\beta\,I\frac{S}{S+I}\,-\alpha\,I$
But we won’t be concerned with this model just now.
15.05: The I Equation
We shall simplify the SI model further to an “I” model, the most basic epidemiological formulation. This can be done by considering a constant population, with births always matching deaths. To accomplish this, the birth term, $b(S +I)$, can be made equal to the death term, $\delta\,(S + I)$, giving
\begin{align*} \frac{dS}{dt}\, &=\delta\,(S+I)\,-\beta\,I\frac{S}{S+I}\,-\delta\,S \[4pt] \frac{dI}{dt}\, &=\beta\,I\frac{S}{S+I}\,-\alpha\,I \end{align*}
However, since the total population—call it $N$—is constant, $S$ is not needed in the equations at all. It is always equal to the total population $N$ minus the number infected: $S\,=\,N\,−\,I$. You can forget about the $S$ equation and substitute $S\,=\,N\,−\,I$ into the $I$ equation. This gives
\begin{align} \frac{dI}{dt} &=\beta\,I\frac{N-1}{(N-1)+I}\,-\alpha\,I \[4pt] \dfrac{dI}{dt} &=\beta\,I\frac{N-I}{N}\,-\alpha\,I \[4pt] \dfrac{dI}{dt} &=\beta\,I(1-\frac{I}{N}\,)\,-\alpha\,I \[4pt] \dfrac{1}{I}\frac{dI}{dt} &=\beta\,(1-\frac{I}{N}\,)\,-\alpha \end{align}
This is the beginning of the I model. It needs a little more work before proceeding with its analysis, but first some additional terminology:
• Infection: A term applied as soon as a pathogen has taken hold in a host.
• Disease: A term often applied when an infection starts producing symptoms in a host.
• Incidence: The number of new infections or cases of disease appearing in a population per time unit. Often expressed as a fraction of the total population.
• Prevalence: The total number of infections or cases of disease existing in a population. Often expressed as a fraction of the total population.
With this terminology in mind, let us put the I model into the form of prevalence, as a fraction of the entire population. In this case, prevalence is just p = I/N. Start where we left off, with
$\frac{1}{I}\frac{dI}{dt}\,=\beta\,(1-\frac{I}{N}\,)\,-\alpha$
multiply through by $I$,
$\frac{dI}{dt}\,=\beta\,I(1-\frac{I}{N}\,)\,-\alpha\,I$
divide through by $N$,
$\frac{d \frac{I}{N}}{dt} = \beta I(1-\frac{I}{N}) -\alpha I$
substitute prevalence
$\frac{dp}{dt}\,=\beta\,p(1-p)\,-\alpha\,p$
and, finally, divide through by $p$,
$\frac{1}{p}\frac{dp}{dt}\,=\beta\,p(1-p)\,-\alpha$
Now think about the $(1−p)$ term. This “one minus the prevalence” represents the fraction of the population that is susceptible to the disease. But some fraction of the population may have natural immunity to the disease, and another may have been successfully vaccinated against the disease. Let us call this fraction v and subtract it too from the fraction that is susceptible, finally giving
$\frac{1}{p}\frac{dp}{dt}\,=\beta(1-v-p)\,-\alpha$
Whew, that is the final I model—the starting place for analysis!
You will see in an upcoming chapter that Equation \ref{15.5.12} is identical to a model of habitat destruction in which plants are isomorphic to “infections” of the landscape, and that habitat destruction, which “protects” the landscape from infection by plants, is isomorphic to vaccination.
Here is the equation with term-by-term explanations:
$\frac{1}{p}\frac{dp}{dt}\,=\beta(1-v-p)\,-\alpha$
$\frac{dp}{dt}\Rightarrow\,\,$ Relative growth in prevalence under conditions prevailing at time $t$
$\beta\Rightarrow\,\,$ Number of new infections induced by each infected individual in each time unit in a completely uninfected population.
$v\Rightarrow\,\,$ Probability of encountering an individual who cannot be infected.
$p\Rightarrow\,\,$ Probability of encountering an infected individual.
$\alpha\Rightarrow\,\,$ Fraction of infected individuals lost per time unit.
Keep in mind that this is a constant-population approximation. Whenever an individual dies of the disease, a new susceptible individual enters the population. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/15%3A_Theory_of_Disease/15.03%3A_The_SIR_Equations.txt |
With $v$ as the prevalence of vaccination in a population, what will the I model reveal about vaccination? In an uninfected population the prevalence of infection will be zero $(p\,=\,0)$, so the number of new infections produced per infected individual per time unit will be $\beta\,(1\,−\,v)$.
On average, a fraction $\alpha$ will die per time unit, so the average duration of infection will be 1/$\alpha$, assuming complete randomness. If 1/10 die per year, for example, the average duration of infection will be 10 years.
This makes $R_0\,(v)\,=\,\beta\,(1\,−\,v)\,\times\,(1/\alpha)\,=\,(\beta/\alpha)\,(1\,−\,v)$. And the disease will decline to extinction if $R_0\,(v)\,\lt\,1$ —that is, if $R_0\,(v)\,=\,(\beta/\alpha)\,(1\,−\,v)\,\lt\,1$, which you can work out algebraically in a few steps to $v\,\gt\,1\,−\,\alpha/\beta$.
Look what this means. A disease that infects 4 individuals per year in a totally uninfected population $\beta\,=\,4), and which remains infectious for one year \(1/\alpha\,=\,1), will decline to extinction if \(v\,\gt\,1\,−\,1/4\,=\,3/4$. If only slightly more than 3/4 of the population is vaccinated, that disease will eventually vanish. Remarkably, a disease can be eradicated even if the whole population cannot be vaccinated! Largely because of this, society can develop programs striving toward the conquest of disease.
What does the $I$ model reveal about the evolution of infectious disease? The pathogen has many more generations and can therefore evolve biologically more rapidly than the host, and $\alpha$ and $\beta$ can evolve to benefit the pathogen.
Because $\beta$ enters the equation with a plus sign and $\alpha$ enters with a minus sign, the disease will spread more rapidly— $(1/p)\,(dp/dt)$ will be larger—if $\beta$ increases and $alpha$ decreases.
This means that, genetics permitting, a successful disease operating according to this or any similar model will tend to become more infectious (higher $\beta$) and less virulent (lower $alpha$) over time. At the limit, nearly everybody will be infected but the effect on anybody will be minimal. Polio in humans before the 20th century, and SIV in monkeys, are examples.
In the ultimate limit, a disease could evolve to negative virulence—that is, to be a mutualism with the host. Rhizobial bacterial in legumes may be an example.
As usual, there are refinements to this idea, in part because infectivity and virulence are not independent. Diseases that evolve to be more infectious may have to use more of their victims’ metabolic resources, and consequently may become more virulent in the process. Again, such refinements can be addressed in more specific models.
15.07: 15.7 The rsN model again
Let’s return to shorter time frames without evolution. Notice that prevalence $p$ changes with time as the epidemic spreads, but $\beta$, $\alpha$, and $v$ remain constant. So starting with
$\frac{1}{p}\frac{dp}{dt}\,=\beta(1-v-p)\,-\alpha$
rearranging terms,
$\frac{1}{p}\frac{dp}{dt}\,=(\beta(1-v)-\alpha)-\beta\,p$
and substituting $r\,=\,\beta(1\,-\,v)\,−\,\alpha,\,s\,=\,−\beta\,,\,and\,N\,=\,p$,
$\frac{1}{N}\,\frac{dN}{dt}\,=\,r\,+\,sN$
Voilà, the epidemiological I model is revealed to be just the standard model of ecology in another disguise! But now, with mechanisms included (infectivity, virulence, vaccination), deeper conclusions can be reached.
In getting to this standard model, for example, we set s equal to −$\beta$. Because $\beta$ is positive, the density dependence term $s$ is negative, Negative $s$ implies logistic growth (positive $s$ is orthologistic), meaning that $N$ will reach a carrying capacity—an equilibrium, a steady state. But to arrive at the standard model, we set $N$ equal to $p$, so since $N$ reaches an equilibrium in the standard model, the prevalence $p$ will reach an equilibrium in the $I$ model.
Without further analysis you can therefore conclude that a disease will not necessarily infect an entire population, but that its prevalence will level out when it reaches a carrying capacity, at the equivalent of −$r/s$. Substitute backwards $(r\,=\,\beta\,(1\,−\,v)\,−\,\alpha\,and\,s\,=\,−\beta\,)$ and you will find the carrying capacity of the disease:
$\hat{p} = 1 − \frac{α}{b} − v$
The small hat atop the $p$ is just a reminder that this is not the variable prevalence $p$, but rather the fixed value of the equilibrium prevalence, $\hat{p}$.
Think about the approach in 15.6, which used $R_0$. Here is another way to get the result. Since $v$, the proportion of the population vaccinated, appears in the equation with a minus sign, that means that the greater the proportion vaccinated, the lower the equilibrium prevalence $\hat{p}$. In fact, setting $\hat{p}\,=\,0$ and solving for $v$, when $v\,=\,1\,−\,\alpha\,/\beta$, the prevalence $p$ will be zero and the disease will be eradicated. (Actually, for a margin of error, when $v\,\ge\,1\,−\,\alpha\,/\beta$. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/15%3A_Theory_of_Disease/15.06%3A_Analysis_of_the_I_model.txt |
An ominous outbreak of Ebola in West Africa became widely known in 2014, with the rate of deaths repeatedly doubling and redoubling. By the fall of that year cases started appearing on other continents.
Ebola apparently enters human populations from wild animals such as bats, in whom it is not particularly virulent. It has not, however, adapted to the human body. It becomes such a dread disease in humans because it exploits almost all of the body’s exit portals— not simply by finding limited passages through them, but by completely destroying them. Some diseases may induce vomiting or diarrhea, for example, as a means of for the pathogen to exit the alimentary canal, but with Ebola, entire tissue systems are destroyed and chunks of intestine accompany the exit.
At the time of the outbreak, two of us (CL, SW) were jointly teaching courses in the United States in quantitative ecology and in the ecology of disease, using tools illustrated thus far in this book. A level of fear prevailed in the country because the disease had just reached the U.S., with a few deaths in U.S. hospitals. Along with our students, we decided to make Ebola a case study for application of the disease equations, applying the principles week-by-week as the outbreak was advancing. Hundreds of thousands of deaths had been predicted by health organizations. This section describes what we did and what we discovered.
Real-time data. Data from the World Health Organization (WHO) and other official sources had been tabulated on a website about Ebola in West Africa, and you can select a date there to see exactly which data were available when we started tracking the outbreak, or at any subsequent time. The site listed both the number of individuals infected with Ebola and the number who had died, but in the early days of the outbreak we guessed that the number of deaths would be more reliable. Moreover, the world was paying greatest attention to deaths, so we hypothesized that these numbers would most strongly influence social efforts to combat the disease.
We thus started with total deaths, plotting them as in Figure $1$. The number of deaths was not only increasing but accelerating, seemingly on an exponential course. The calculated doubling time was 31.5 days, not quite what U.S. Centers for Disease Control (CDC) had found earlier, but within reasonable correspondence. They estimated 15–20 days to double in one country and 30–40 in another (Meltzer et al., 2014).
Calculating from these doubling times and extending for several more months results in the number of deaths shown in Figure $2$. Before five months with the estimates we made during class (blue curve in the figure) and before three months with the earlier estimates (red curve), the number of deaths was predicted to exceed 100,000.
But there is a flaw in this approach. As shown by the calculations on bacteria and Darwin’s calculations on elephants (Chapter 3), exponential growth models cannot be extended very far. They can be quite accurate a limited number of time units in the future. But when applied to biological populations both exponential and orthologistic models inevitably fail when extended indefinitely. Actually, one view is that they do not really fail—they just warn that some other growth model will supplant them before populations grow too large.
Table $1$. Consequences of unlimited doubling in Ebola deaths.
Days Years Total Deaths
0 0.00 892
100 0.27 8,412
200 0.55 79,335
300 0.82 74,8267
400 1.10 7,057,440
500 1.37 66,563,800
600 1.64 627,811,000
700 1.92 5,921,330,000
710 1.94 7,411,030,000
In fact, assuming the unrestrained doublings of exponential growth is tantamount to assuming a disease will kill the entire world, with the only question being when. Table $1$ shows the results of doubling at the rate illustrated by the blue curve in Figure $2$, extended farther. At this rate the entire human population would be extinguished in less than two years!
Expected moderation. Of course, the entire human population will not be extinguished by a manageable disease. Draconian measures would be put in place long before— isolating those infected, closing borders, and more. In effect, the growth rate $r$ will be moderated by strong negative social pressure, term $s$.
When would such negative pressure appear in the data? Could it be seen early in the Ebola outbreak, when we and the students began watching? Because the basic disease equation is equivalent to the $rsN$ model, we thought early trends might appear if we examined the data in terms of $r$ and $s$. The individual growth rate in deaths, $(1/N)dN/dt$, could be examined and plotted against the total number of deaths, $N$. This is Figure $3$, with the same data as Figure $1$, just reformulated.
The data show quite a bit of noise, but with a clear downward trend, with the rate in the total number of new deaths decreasing as deaths increase. Both the upper and lower ranges of points are decreasing (gray lines). The green line through the averages (least-square regression line, solid, with $r$ and $s$ as shown in the figure) projects forward (dashed) to about 12,000 deaths before the outbreak would be over—an extensive human tragedy, but far below the straightforward projections of Figure $2$.
If this decline in death rate was real, it was likely developing from more and more attention to moderating the disease—medical workers expanding hospitals, populations practicing more careful burials, governments cautioning only justified travel, and the like. When we made the projection of 12,000 in the fall of 2014, we had no certainty about what would happen; we were simply looking at the data, which showed not unrestrained exponential growth, but moderating growth instead.
The next step was to plug the $r$ and $s$ derived from the fitted curve and project forward six months or a year. This projection is shown in Figure $4$. The green curve is the projection from $r\,=\,0.028$ and $s\,=\,−0.0024$. It is markedly different than the other two curves, leveling off early and reaching about 12,000 deaths. Also in the figure are three additional points of actual data, in yellow, not sufficiently advanced to tell which of the three curves—red, green, or blue—will be the real one.
In only a few weeks, however, it became apparent that the $r+sN$ curve was the most accurate. By the end of the winter semester (about day 140 in the graphs) it was clear that the outbreak was coming under control, and that the number of deaths was fairly close to our initial projection. After tracking the outbreak with our students through the winter and spring semesters, we could see how remarkable that early projection was (Figure $5$). Reading the data carefully early in the crisis gave an accurate projection of the outcome, from a simple model indeed!
Doubling times. This example provides a good place for us to reconsider doubling times, introduced with exponential growth. Recall that exponential growth is the infinitely thin dividing line between logistic and orthologistic growth, and has the property of a fixed “doubling time.” In other words, there is a specific time interval—call it tau ($\tau$)—during which the population exactly doubles. In logistic growth the doubling time constantly decreases, while in orthologistic growth it constantly increases. Earlier the doubling time is shown to be the natural logarithm of 2 divided by r— $(ln2)/r$, or approximately $0.693/r$. This is in years if $r$ is measured per year, days if per day, an so forth.
Thus the doubling time for the exponential curve of Figure $1$, with $r\,=\,0.022$, is $0.693/0.022\,=\,31.5$ days. The logistic growth of Figure $3$ has no fixed doubling time. However, at all times there will be an “instantaneous doubling time,” which will hold approximately for a short time. In particular, near the very beginning of the outbreak—when $N$ is close to 0—the growth rate is $r+sN\,=\,r+s\,\cdot\,0\,=\,r$. For the Ebola data early in the outbreak, we found $r\,=\,0.028$ and $s\,=\,0.0024$. The Ebola doubling time, averaged over the countries in which it spread, therefore started out at $0.693/r\,=\,0.693/0.028\,=\,25$ days. This is in accord with early estimates made by health organizations.
When we began following the outbreak, about 4.5 thousand deaths had occurred, meaning that the growth rate was $r+sN\,=\,0.028−0.0024\,\cdot\,4.5\,=\,0.0172$, and the doubling time was $0.693/0.0172\,=\,40$ days.
The doubling time continued to decrease until the outbreak was conquered and all deaths from Ebola ceased.
Caveats and considerations. There was some fortuity in the timing of our initial projection. Shortly after we started the death rate dropped, possibly from increased efforts following intense world-wide attention. Had we made the projection a few weeks later we would have seen two slopes and would have had to guess which would prevail. It turns out that the first slope prevailed, returning about 40 days after our initial projection. But we had no way to know this from the data at the time.
Our approach could also be criticized because the $r+sN$ equation we used is a hybrid—a single equation trying to represent two different things, in this case infections and deaths. It is reasonable here to base the amelioration parameter s on the total number of deaths, since deaths were the parameter of world concern. Deaths influence social attention to the disease and efforts to control it. However, what sense is there in saying that deaths grow at rate $r$, based on the total number of deaths thus far? Ebola can be transmitted to others shortly after death, but in general deaths do not cause new deaths. Infections cause new infections, which in turn cause new deaths. If $N$ represents total deaths, it seems the equations should also have an $I$ to represent infections, and some fraction of infections should result in deaths—as in this two-dimensional system of equations.
$\frac{dI}{dt}\,=\,\beta\,I\,-\,\gamma\,I\,-\,\alpha\,I\,-\,sNI$
$\frac{dN}{dt}\,=\,\alpha\,I$
The two dimensions are $I$, the number of existing infections, and $N$, the cumulative number of deaths. Infectivity is $\beta$, the rate of recovery from infection is $\gamma$, and the rate of death from the disease is $\alpha$. These correspond to the notation in Figure 15.2.1. In addition, the term $sN$ moderates the growth of the infection by representing all the cautions and infrastructure put in place against the disease as the number of deaths increased.
No factor like $1−v−p$, representing the fraction of susceptible individuals, is needed to multiply the $\beta$ term here, because both in the early stage and the outbreak the prevalence was low and there was no vaccine, so almost everyone was susceptible. And with this rapidly progressing disease the population remained almost constant, so births could be considered negligible. Equation is thus still a simplified equation.
As simplified as it is, however, Equation carries more parameters than can be perceived in the raw data. The death rate from the disease $\alpha$, the recovery rate $\gamma$, the infectivity $\beta$, and even the number of infections $I$, can only be discovered through the use of special research programs. But even at the beginning of an outbreak, and even for a poorly understood disease, there may be enough data to determine the doubling time for deaths from the disease, and how the doubling time is changing—enough data to determine $r$ and $s$, even if little more.
This is fortunate, because Equation can be reduced in dimension and approximated by the $r+sN$ form. Without the amelioration term $sNI$, the two dimensions are independent and growth of each is exponential, with $N$ being the integral of $I$. Because the integral of an exponential is still exponential, the two can be approximated by a single equation of one fewer dimension, with the amelioration term $sNI$ reinstated. In this way, total deaths become a legitimate surrogate for infections in low-prevalence outbreaks, as in the present example of Ebola.
And as we noted, this gave an accurate projection on the course of a dread disease, from a very simple model indeed!
15.09: Conflicting ethics
Fundamental conflicts in ethical behavior are hidden just below the surface of the cold mathematics in these equations. Virulence $\alpha$ appearing in the numerator of a term with a minus sign means that the greater the virulence, the lower the equilibrium level of disease, $\hat{p}$. In the absence of vaccination $(v\,=\,0)$, setting $\hat{p}$ = 0 and solving for $\alpha$ yields $\alpha\,=\,\beta$. Artificially increasing the virulence $\alpha$ of a disease to equal or exceed the infectivity $\beta$ will therefore drive the disease extinct.
The enduring ethical conflict here is between the individual and the population. With all other things being equal, working to reduce virulence benefits the individual but may cause more individuals in the population to become infected. Working to increase virulence, in contrast, harms individuals but may reduce the number who become infected.
The ethics of modern medicine emphasizes individuals— working to cure disease and reduce virulence, ameliorating symptoms, reducing discomfort, and recognizing patient needs. Increasing the virulence of a disease in a human patient to reduce its spread is unthinkable, both in medicine and public health. The ethics of modern agriculture, however, are the diametric opposite. If a crop is infected with a destructive communicable disease, entire fields of the crop may be mowed, burned, or otherwise disposed of. Infected populations of poultry and livestock are treated similarly, killed en masse and buried or burned to contain the disease.
Artificially altering the infectivity $\beta$ is also a possibility. In Equation 15.4, $\beta$ appears in the denominator of a term having a minus sign—meaning that decreasing $\beta$ will decrease the equilibrium level of the disease, $\hat{p}$. Ethical conflicts also arise here, though they are not as stark as the conflicts connected with $\alpha$. During the influenza epidemic of 1918–19, San Francisco leaders required citizens to wear breathing masks—for “conscience, patriotism, and self-protection,” wrote the mayor.
This lowered $\beta$ by containing respiratory droplets from infected individuals, and lowered the chance of infected droplets entering the respiratory systems of susceptible individuals, in turn reducing the infectivity $\beta$. Some citizens, however, refused to wear the masks.
During the Ebola outbreak of 2014–15, amidst fears and warnings of the disease becoming established around the world, some U.S. governors ordered temporary quarantine of returning medical workers who had been in direct proximity with Ebola, until it was clear that they were not infected. At least one refused the quarantine based on individual rights, and the courts upheld the refusal.
These ethical conflicts surrounding $\beta$ are not as grim as those surrounding $\alpha$, with options currently practiced for domestic plants and animals but so extreme that they are never proposed for human populations.
It is curious that these high-level social dilemmas are perceptible within the most basic equations of ecology. Science can inform such ethical issues, but society must decide them. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/15%3A_Theory_of_Disease/15.08%3A_Application_to_an_actual_outbreak.txt |
Ecological competition is a potent force driving organic evolution. When Charles Darwin reached the Galapagos Islands at age 26, he studied an assortment of fifteen similar species that are now called “Darwin’s Finches” (Figure \(2\)). His observation that various species had different beak structures, specialized for different foods, led him to question the stability of species. Indeed, such “character displacement” is one of several consequences of ecological competition.
Seeing the gradation and diversity of structure in one small, intimately related group of birds, one might really fancy that from an original paucity of birds in this archipelago, one species had been taken and modified for different ends. —Charles Darwin
16.02: The niche concept
A species’ niche is the range of environmental factors that allow that species to survive and reproduce. A particular tree species, for example, may be able to live where temperatures do not drop below −40, and where yearly precipitation is at least 750 mm. Perhaps it also needs open sunlight and an appropriate collection of root fungi. Such are the parameters of a niche.
G. Evelyn Hutchinson, one of the great ecologists of the twentieth century, envisioned the parameters that form a niche as an “n-dimensional hyperspace.” The “fundamental niche” is the set of conditions allowing the species to survive if there are no other species interfering. Physical conditions are chief among those. The “realized niche” is the real life niche—where species are restricted by interactions with other species.
Consider latitude on the earth’s surface, which is connected to several parameters such as sunlight and temperature. And consider two species that can thrive anywhere between 40 and 60 latitude, and whose density drops slowly with increasing latitude (Figure \(1\)).
At the top of the figure are two nearly horizontal lines representing the abundance you might observe of the two species as you travel north. If free of Species 2 (its competitor), Species 1 (blue line) declines slowly in abundance in more northerly climates. Species 2 similarly declines in abundance (red line), but compared with Species 1 fares a little better in the north and a little worse in the south.
When these two species are together they compete with each other—each suppressing the other. Using competition equations like those presented in this chapter, we see that in the south, where Species 1 fares better, it takes over and dominates. In the north, in contrast, where Species 2 fares better, it dominates instead (Figure \(1\), bottom).
You see that there can be a sharp change in abundance even with only very slight changes in species characteristics. A range of one species can end and that of a new species can begin, even though you may not be able to discover anything from either species alone as to why they switch their dominance. And the switch-over point need not correspond to the exact place in which their dominance switches. Here the actual switch-over point is a few degrees to the north because of the migration simulated in the model. This phenomenon is called “competitive exclusion” and, when it occurs over space like this, “zonation.”
Any environmental gradient can induce zonation. Figure \(2\) illustrates this on a restored prairie in the North American Upper Midwest. Though less than ten meters, the variation in elevation is enough to induce a mild moisture gradient. The entire area had been converted from a corn field to a restored prairie and seeded uniformly with a mixture of prairie grasses and prairie flowers, but distinct boundaries arose only three years after restoration.
All the upland areas, labeled 1, contained standard restored prairie flora such as Andropogon gerardi (Big Bluestem grass) and rapidly emerging flowers such as Rudbekia hirta (Black-eyed Susan). But in successive zones surrounding mild depressions in the landscape, labeled 2 to 4, there were sharp transitions to moisture-loving genera such as Typha latifolia (cattail) and Stachys palustris (Smartweed).
Definitions
Fundamental niche: The conditions under which a species can live, absent interference from other species.
Realized niche: The conditions to which a species is restricted by interactions with other species. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/16%3A_Competition/16.01%3A_Chapter_Introduction.txt |
Competition among two species occurs when the interaction terms $s_{1,2}$ and $s_{2,1}$ in Equation 8.1 are both negative. This rather abstract approach does lead to broad insights, but for other kinds of insights let us proceed to a more mechanistic view. Instead of abstract coefficients representing inhibition among species, let us consider resources which species need to thrive and survive. The species will not interact directly— they never even need to come into contact—but will influence each other through their use of a common resource, which they both need for maintenance and growth of their populations.
Resource competition is one of the oldest parts of ecological theory, introduced in the late 1920s by mathematician Vivo Volterra. We will start where he started, considering what have been called “abiotic resources”.
Species require sunlight, space, nitrogen, phosphorous, and other resources in various amounts. If a resource is too rare, populations cannot grow, and in fact will decline. In Figure $1$ this is shown in the region to the left of the arrow marked A, in which the individual growth rate $1/N\,dN/dt$ is negative.
At higher resource levels the growth rate increases and, at point A, the population can just barely maintain itself. Here the individual growth rate $1/N\,dN/dt$ is zero. This level of resources is called $R^{\ast}$, pronounced “are star”. At higher levels of resource, above $R^{\ast}$, the population grows because $1/N\,dN/dt$ becomes positive.
Once resources are abundant—approximately above B in the figure—needs are satiated and the addition of more resources does not make any large difference. Population growth stays approximately the same between marks B and C.
At very high levels, too much resource can actually harm the population. Too much sunlight can burn leaves, for example, while too much nitrogen can damage roots. At this point, above C in Figure $1$, the growth rate starts falling. By D the species can again just barely hold its own, and to the right of D the species is killed by an overabundance of resources.
Such high resource levels, however, are not usually observed, because species draw resources concentrations down by using them up. Unless extreme environments are being modeled, only the left dashed linear piece of Figure $1$ needs to be modeled, as shown in Figure $2$.
At this point, it is helpful to review various forms for the equation of a straight line. The usual slope–intercept form, $y\,=\,mx+b$, which is a y-intercept form—is not as useful here. It’s the x-intercept form, $y\,=\,m(x−a)$, that comes into play for writing a mechanistic resource model for single species population growth.
1. Slope–intercept form: $y\,=\,mx\,+\,b$
(slope m, y-intercept b)
2. Intercept–intercept form: $y\,=\,b(1\,−\,x/a)$
(x-intercept a, y-intercept b)
(m = −b/a)
3. Slope–x-intercept form: $y\,=\,m\,(x − a)$
(slope m, x-intercept a)
(b = −m/a)
The zero-growth point, $R^{\ast}$, is important in the theory of resource competition. It is the amount of resource that just barely sustains the species. If the resource level is less than $R^{\ast}$, the species dies out; if it is greater, the species grows and expands. The resource level in the environment therefore is expected to be at or near the $R^{\ast}$ value of the dominant species. If it is above that level, the population grows, new individuals use more resource, and the resource level is consequently reduced until growth stops.
$R^{\ast}$ can be measured in the greenhouse or the field. In the greenhouse, for example, you might arrange plants in 20 pots and give them different amounts of nitrogen fertilizer in sterile, nutrient-free soil. In the complete absence of fertilizer, the plants will die. With larger amounts of fertilizer the plants will be luxuriant, and when there is too much fertilizer, the resource will become toxic, again leading to dieback in the plants. You can thus measure the curve of Figure $1$ fairly easily and find the point on the left where the plants just survive. This is their $R^{\ast}$.
You can also measure this value for different species independently, and from the results estimate how plants will fare living together. To start, suppose that one resource is the most limiting. Represent the amount of that limiting resource available in the environment at time $t$ by the symbol $R(t)$ or, for shorthand, simply $R$. The amount in excess of minimal needs is $R−R^{\ast}$, and that amount of excess will determine the rate of growth. A tiny excess will mean slow growth, but a larger excess can support faster growth. So the equation in Figure $2$ shows the individual growth rate, $1/N\,dN/dt$, being proportional to how much resource exists in excess.
As before, $N$ measures the size of the population at time $t$, in number of individuals, total biomass, or whatever units are relevant to the species being studied. $R^{\ast}$ is the smallest amount of resource that can support a viable population, and m tells how the individual growth rate, $(1/N)dN/dt$, depends on the amount of resource available in excess of minimal needs.
Now let us say that $R_{max}$ is the maximum amount of resource available in the environment, in absence of any organisms, and u is the amount of resource used by each living organism in the population. Then $uN$ is the amount of resource tied up in the population at time $t$. $R_{max}\,−\,uN\,=\,R$ is the amount of resource not used by the population. This is the basis of a resource theory that assumes resources are released immediately upon death of an organism, and it has many of the important properties of more complex resource models.
Start with the above statements in algebraic form,
$\frac{1}{N}\frac{dN}{dt}\,=\,m(R-R^{\ast})\,\,\,and\,\,\,R\,=\,R_{max}\,-uN$
then substitute the equation on the right above into the one on the left. This gives
$\frac{1}{N}\frac{dN}{dt}\,=\,m\,(R_{max}\,-\,uN\,-\,R^{\ast})$
Multiplying through on the right by $m$ and rearranging terms gives
$\frac{1}{N}\frac{dN}{dt}\,=\,m(R_{max}\,-\,R^{\ast})\,-\,umN$
Notice that the first term on the right is a constant and the second term is a constant times N. Does this look familiar? This is just density-regulated population growth in disguise— the $r\,+\,sN$ model again!
$\frac{1}{N}\frac{dN}{dt}\,=\,r+sN,\,\,\,r\,=\,m(R_{max}\,-\,R^{/ast}),\,\,\,s\,=\,-um$
Recall that this also happened for the epidemiological $I$ model. And it will arise occur again in future models. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/16%3A_Competition/16.03%3A_Resource_competition.txt |
Consider what will happen with two species using the same resource, such as light or space or nitrogen fertilizer. The amount of resource available will be the maximum amount, $R_{max}$, minus what is tied up in all individuals of all species. With $u_i$ being the amount of resource tied up in each individual of species $i$, the resource remaining at any time will be
$R\,=\,R_{max}\,-\,u_1N_1\,-\,u_2N_2$
Or for many species
$R\,=\,R_{max}\,-\,u_1N_1\,-\,u_2N_2\,-u_3N_3\,-\cdots\,-u_hN_h$
$R\,=\,R_{max}\,-\sum_{i=1}^h u_iN_i$
Each species has its own growth equation, identical in form for all species, but different in the critical level of resource, $R_i^{\ast}$, and the growth coefficient, $m_i$:
$\frac{1}{N_i}\frac{dN_i}{dt}\,=\,m_i(R\,-\,R_i^{\ast})$
What remains is to consider how the growth coefficient $m_i$ relates to the minimal level of resource tolerated, $R_i^{\ast}$. It turns out to be a tradeoff between the two. Consider, for example, a plant species that is limited by the amount of nitrogen available, as plants are. And to have a large growth coefficient $m_i$ the plant must produce abundant seed. To have superior nitrogen use, measured by a low value of $R^{\ast}$, it needs abundant roots. But it cannot do both. There is a limited amount of solar energy to exploit, so if the plant allocates more to roots there is less to allocate to seeds, and vice versa.
It therefore turns out that species which are good colonizers, producing abundant seed, are poorer competitors for resources, having a higher value of $R^{\ast}$. This idea is illustrated by measurements reported in Figure $1$.
In Figure $2$, tradeoffs are formulated for modeling. Species 2 grows more rapidly when resources are abundant. This is the case at time 0, marked with $t_0$ on the top axis.
As the populations grow they reduce the amount of resource available in the environment. At time 1, marked with $t_1$ on the upper axis, Species 2 can continue to grow faster than Species 1, though the margin is deteriorating. But there comes a point at which the resources become depleted enough that the characteristics of Species 2 do not let it gather enough resources to maintain its advantage. This is the crossing point of the blue and red lines in the figure. At time 2, both species are still growing, but Species 1 is growing faster. At time 3, with still lower resource levels—drawn down by Species 1—the resource falls below the minimal level for Species 2, $R_2^{\ast}$. The growth rate of Species 2 falls negative and Species 2 starts to die out.
Finally, at time 4, Species 1 depletes the resource to the level that it can just barely survive, and it stands alone, having wiped out its competitor. This process is called “competitive exclusion.”
How this plays out over time is illustrated in Figure $3$. At the top, Species 2 alone does just fine, rapidly rising to its carrying capacity of 50 and pulling the resource down to its $R^{\ast}\ ) of 2. In the middle, Species 1 alone also does just fine, rapidly rising to its carrying capacity of 60 and pulling the resource down to its \(R^{\ast}$ of 1.
But grown together, Species 2 makes an initial splash and then declines. This is due to the incessant growth of Species 1, which outcompetes it. Species 1 simply draws the resource down below the level at which Species 2 can survive.
Competitive exclusion, which assumed that no more species could exist than there were resources, was treated as an inviolable law of ecology for over fifty years. In the 1970s, however, this was shown not to be the case (Armstrong and McGehee 1980). More about that later in the chapter. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/16%3A_Competition/16.04%3A_Competitive_exclusion.txt |
A similar process for more than two species results in a succession of species taking over, one after the other, in an ecological process known as “succession.”
In natural systems many species compete, with tradeoffs between their \(R^{\ast}\) values and their growth rates, as in Figure \(1\). The following is a program to simulate the differential equations for five species competing for the same resource and producing the curves of Figure \(2\). With only two species, this same program can produce the curves of Figure 16.3.3.
```# SIMULATE ONE YEAR
# This routine simulates competition differential equations through one time
# unit, such as one year, taking very small time steps along the way.
# Accuracy should be checked by reducing the size of small time steps until
# the results do not significantly change.
# This routine implements <Q>Euler’s Method</Q> for solving differential
# equations, which always works if the time step is small enough.
#
# ENTRY: ’N1’ to ’N5’ are the starting populations for species 1-5.
# ’m1’ to ’m5’ specify the sensitivity of the corresponding species
# to the available amount of resource.
# ’u1’ to ’u5’ specify the resource tied up in each species.
# ’R1star’ to ’R5star’ are the minimum resource levels.
# ’Rmax’ is the greatest amount of resource possible.
# ’dt’ is the duration of each small time step to be taken throughout
# the year or other time unit.
#
# EXIT: ’N1’ to ’N5’ are the estimated populations of species 1-5 at the
# end of the time unit.
# ’R’ is the estimated resource level at the end of the time step.
```
```Rmax=R=7;
R1star=1.0; R2star=2.0; R3star=3.0; R4star=4.0; R5star=5.0;
N1=0.000001; N2=0.000010; N3=0.000100; N4=0.001000; N5=0.010000;
m1=0.171468; m2=0.308642; m3=0.555556; m4=1.000000; m5=1.800000;
u1=0.001000; u2=0.001000; u3=0.001000; u4=0.001000; u5=0.001000;
# SIMULATE ONE YEAR
SimulateOneYear = function(dt)
{ for(v in 1:(1/dt)) # Advance a small time step.
{ R=Rmax-u1*N1-u2*N2-u3*N3- # Compute resource remaining.
u4*N4-u5*N5;
dN1=m1*(R-R1star)*N1*dt; # Estimate the change in the
dN2=m2*(R-R2star)*N2*dt; # population of each species.
dN3=m3*(R-R3star)*N3*dt;
dN4=m4*(R-R4star)*N4*dt;
dN5=m5*(R-R5star)*N5*dt;
N1=N1+dN1; N2=N2+dN2; # Add the estimated change to
N3=N3+dN3; N4=N4+dN4; N5=N5+dN5; } # each population and repeat.
assign("N1",N1, envir=.GlobalEnv); # At the end, export the results and return
assign("N2",N2, envir=.GlobalEnv);
assign("N3",N3, envir=.GlobalEnv);
assign("N4",N4, envir=.GlobalEnv);
assign("N5",N5, envir=.GlobalEnv); }
# SIMULATE ALL YEARS
for(t in 0:100) # Advance one year.
{ print(round(c(t,N1,N2,N3,N4,N5))); # Display results.
SimulateOneYear(1/(365*24)); } # Repeat.
```
An environment can change because species living in it have effects that can “feed back” and change the environment itself. In this case the feedback is change in the resource level, which each successive species changes in a way that is compatible with its own existence. There is nothing teleological in this; any species that change the environment in ways not compatible with their own existence simply do not persist, and hence are not observed. When the program runs, it produces a file excerpted below, which is graphed in Figure \(2\).
t N1 N2 N3 N4 N5
1 0 0 0 0 0
2 0 0 0 0 13
3 0 0 0 7 391
4 0 0 0 45 1764
5 0 0 1 127 1894
6 0 0 3 329 1743
7 0 0 9 790 1392
: : : : : :
60 1649 3536 0 0 0
61 1891 3324 0 0 0
62 2157 3094 0 0 0
63 2445 2864 0 0 0
64 2751 2584 0 0 0
65 3070 2313 0 0 0
66 3397 2039 0 0 0
67 3725 1767 0 0 0
: : : : : :
96 5999 1 0 0 0
97 5999 0 0 0 0
98 6000 0 0 0 0
99 6000 0 0 0 0
100 6000 0 0 0 0
At the beginning in Figure \(2\), from time 0 to about time 3, the resource is at its maximum level, \(R_{max}\), and the abundances of all species are at very low levels. Between times 3 and 5—Species 5, the one with the highest growth rate when resources are abundant—increases rapidly while resources drop accordingly. But near the end of that time the next in the series, Species 4, starts to increase, pulling the resource down below the level that allows Species 5 to survive. Species 5 therefore declines while Species 4 increases.
This process continues in succession, with Species 3 replacing 4, 2 replacing 3, and, finally, Species 1 replacing 2. The resource falls in stages as each successive species gains dominance. Finally, when no more superior species exist, the system reaches what is called its “climax condition” at about time 90, with resources at a low level.
There is nothing especially remarkable about Species 1. It is simply (1) the best competitor living in the region, meaning better competitors cannot readily arrive on the scene, or (2) the best competitor that evolutionary processes have yet produced. In either case, it is subject to replacement by another—for example, by an “invasive species” arriving by extraordinary means.
Of course, succession in complex natural systems may not be as clear-cut as in our simple models. Multiple resources are involved, species may be very close to each other in their ecological parameters, and stochastic events may intervene to add confusion.
16.06: Single-resource phase space
Some aspects of competition for a resource are clarified by looking at the phase space, as introduced in Chapter 10. Combining Equations 16.3.1 and 16.3.2 gives the following as a starting point:
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,m_1(R_{max}\,-\,R_1^{\ast})\,-\,u_1m_1N_1\,-\,u_2m_2N_2$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,m_2(R_{max}\,-\,R_2^{\ast})\,-\,u_2m_2N_2\,-\,u_1m_1N_1$
As before, $m_i$ is the rate of growth of Species $i$ for each level of resource above its minimum resource requirement $R_i^{\ast}$, and $u_i$ is the amount of resource tied up in each individual of Species $i$. For reference, here is the assignment of parameters in terms of $r_i$ and $s_{i,j}$.
$r_1\,=\,m_1(R_{max}\,−\,R_1^{\ast}),\,\,\,s_{1,1}\,=\,−u_1m_1,\,\,\,s_{1,2}\,=\,−u_2m_2$
$r_2\,=\,m_2(R_{max}\,−\,R_2^{\ast}),\,\,\,s_{2,1}\,=\,−u_1m_1,\,\,\,s_{2,2}\,=\,−u_2m_2$
Where in the phase space will the growth rate be 0 for each species? For Species 1 it will be where
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,0\,=\,r_1\,+\,s_{1,1}N_1\,+\,s_{1,2}N_2$
Solving for $N_2$ gives
$N_2\,=\,−\frac{r_1}{s_{1,2}}\,-\frac{s_{1,1}}{s_{1,2}}N_1\,\,\,\,\,\leftarrow\,Species\,1\,isocline$
Anywhere along that line, the population of Species 1 will not change, but on either side of the line it will (Figure $1$). Formulae for the four possible equilibria and their stability are in Table 10.1. The vertical intercept of the isocline, where $N_1\,=\,0$, is $-r_1/s_{1,2}$, and the horizontal intercept, where $N_2\,=\,0$, is $-r_1/s_{1,1}$. The slope is $-s_{1,1}/s_{1,2})\,=\,(u_1m_1)/(u_2m_2)$.
Similarly, growth of Species 2 will be 0 where
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,0\,=\,r_2\,+\,s_{2,2}N_2\,+\,s_{2,1}N_1$
Solving for $N_2$ gives
$N_2\,=\,-\frac{r_2}{s_{2,2}}\,-\frac{s_{2,1}}{s_{2,2}}N_1\,\,\,\,\,\,\leftarrow\,Species\,2\,isocline$
Again, anywhere along that line the population of Species 2 will not change, but on either side of the line it will (Figure $2$). The vertical intercept of that line, where $N_1\,=\,0$, is $-r_2/s_{2,2}$, the horizontal intercept, where $N_2\,=\,0$, is -$r_2/s_{2,1}$, and the slope is $-s_{2,1}/s_{2,2}\,=\,(u_1m_1)/(u_2m_2)$.
Notice this: In terms of the resource, the slope of the isocline for Species 2 is identical to the slope for Species 1—both are equal to ($u_1m_1)/(u_2m_2$). What does this mean? It means that the two isoclines are parallel. And that, in turn, means that the two species cannot permanently coexist.
The populations can fall into only one of the three regions of Figure $3$. If they start in the upper region, they decrease until they enter the middle region. If they start in the lower region, they increase until they also enter the middle region. Once in the middle region, only Species 2 increases. That means the population of Species 1 is driven leftward, toward lower values of $N_1$, while the population of Species 2 is driven upward, toward higher values of $N_2$.
These dynamics show up in the flow diagram of Figure $4$. The origin (0,0) is an unstable equilibrium. In this single-resource system, any populations near extinction, but not completely extinct, increase until they hit the middle region. The horizontal axis has another unstable equilibrium, where Species 1 is at its carrying capacity and Species 2 is extinct $(−r_1/s_{1,1},$0). Any populations near that unstable equilibrium soon arrive in the middle region. All populations not precisely on one of those two unstable equilibria converge on the red disc on the vertical axis, where Species 2 is at its carrying capacity, $K_2\,=\,−r_2/s_{2,2}$, and Species 1 is extinct (0,$−r_2/s_{2,2}$). This equilibrium is called a “global attractor.”
Phase spaces thus provide another view of competitive exclusion, the theory of which applies at least to two species competing for a single resource at equilibrium. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/16%3A_Competition/16.05%3A_Succession.txt |
As the next step, consider the case of two essential resources. This can be done mathematically following the approach we used earlier for a single resource.
Call the two resources $R_A$ and $R_B$. Suppose that of the total of these two resources used by Species 1, a proportion $p_1$ is Resource A and therefore a proportion $q_1\,=\,1−p_1$ is Resource B. Likewise, for Species 2 a proportion $p_2$ is Resource A and $q_2\,=\,1−p_2$ is Resource B.
This is confusing, so for a clarifying example, suppose $R_A$ is phosphate, PO4, and $R_B$ is silicate, SiO2, both essential to two species of algae in a waterway. Take Species 1 to be an Asterionella species and Species 2 to be a Cyclotella species, as in a pioneering study by David Tilman (1977). In this case, Species 1 needs phosphate and silicate resources in about a 1:99 ratio, while Species 2 needs them in about a 6:94 ratio. If silicate is low, Species 1 will thus suffer first, since it needs a larger proportion of it, while Species 2 will suffer first if phosphate is low, for the related reason. Here it would be $p_1\,=\,0.01,\,q_1\,=\,1−p_1\,=\,0.99$ for the use of Resources A and B by Species 1, and $p_2\,=\,0.06,\,q_1\,=\,1−p_1\,=\,0.94$ for use by Species 2.
With $u_1$ being the total amount of resource tied up by each individual of Species 1, $p_1u_1$ will be the amount of Resource A tied up by Species 1 and $q_1u_1$ the amount of Resource B tied up the same way. Similarly, $p_2u_2$ will be the amount of Resource A tied up by each individual of Species 2 and $q_2u_2$ the amount of Resource B. Also, assume as before that the resources under consideration disappear from the environment when they are taken up by individuals newly born, are released immediately when individuals die.
With this in mind, the resources remaining at any time, as functions of the maximum resource and the abundance of each species, will be
$R_A\,=\,R_{Amax}\,-\,p_1u_1N_1\,-\,p_2u_2N_2$
$R_A\,=\,R_{Bmax}\,-\,q_1u_1N_1\,-\,q_2u_2N_2$
Suppose that populations of Species 1 and 2 grow based on which resource is closest to the $R^{\ast}$ experienced by that species for the resource. This can be represented by the “min” function, min($a,b$), which selects the smaller of two values. For example, min(200,10) = 10, min(−200,10) = −200. Now the two-species, two-resource growth equations, generalizing the single-species, single-resource growth in Equation 16.1.1, are
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,m_1\,min(R_A\,-\,R_{1A}^{\ast},\,R_B\,-\,R_{1B}^{\ast})$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,m_2\,min(R_A\,-\,R_{2A}^{\ast},\,R_B\,-\,R_{2B}^{\ast})$
This could be refined, so that the growth rates $m_1$ and $m_2$ would depend on which resource was limiting, but this does not matter in the present analysis. If the species are similar enough and the resource level is such that they are limited by the same resource, one will tend to be competitively excluded, as in the previous section. But if the two species are quite different, they can be limited by different resources and the equations can be simplified.
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,m_1(R_A\,-\,R_{1A}^{\ast})$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,m_2(R_B\,-\,R_{2B}^{\ast})$
Some algebra will reveal the basic properties. If you substitute the expressions for $R_1$ and $R_2$ from Equations 16.6.1 and 16.6.2 into Equations 16.6.3 and 16.6.4 you will get
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,m_1(R_{Amax}\,-\,p_1u_1N_1\,-\,p_2U_2N_2\,-\,R_{1A}^{\ast})$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,m_2(R_{Bmax}\,-\,q_1u_1N_1\,-\,q_2u_2N_2\,-\,R_{2B}^{\ast})$
Now if you expand, collect, and rearrange terms, you get this equivalent form:
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,m_1(R_{Amax}\,-\,R_{1A}^{\ast})\,-\,m_1p_1u_1N_1\,-\,m_1p_2u_2N_2$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,m_2(R_{Bmax}\,-\,R_{2B}^{\ast})\,-\,m_2q_2u_2N_2\,-\,m_2q_1u_1N_1$
Notice that, again, a mechanistic model with measurable parameters is just the general ecological RSN model in disguise.
The RSN formulation can expose the possibilities of two-resource situations in phase space. The isoclines are Equations 16.5.5 and 16.5.6 of this chapter, with slopes −$s_{1,1}/s_{1,2}$ and −$s_{2,1}/s_{2,2}$ for Species 1 and 2, respectively.
These two slopes can be written in terms of the resource. With the values for $s_{i,j}$ from Equations 16.6.7 and 16.6.8 $(s_{1,1}\,=\,−m_1p_1u_1,\,s_{1,2}\,=\,−m_1p_2u_2,\,s_{2,1}\,=\,−m_2q_1u_1,$ and $s_{2,2}\,=\,−m_2q_2u_2$), the slopes of the isoclines become
$\frac{s_{1,1}}{s_{1,2}}\,=\,\frac{-m_1p_1u_1}{-m_1p_2u_2}\,=\,\frac{u_1}{u_2}\frac{p_1}{p_2}$
$\frac{s_{1,1}}{s_{1,2}}\,=\,\frac{-m_1p_1u_1}{-m_1p_2u_2}\,=\,\frac{u_1}{u_2}\frac{1-p_1}{1-p_2}$
First notice that if the two species use the two resources in equal proportions (if $p_1\,=\,p_2$), both slopes become $u_1/u_2$. The slopes are parallel, as in the single-resource case in Figures 16.5.1 through 16.5.4. Therefore, if two species use two different resources identically—that is, in equal proportions— they do not coexist. Coexistence requires some difference in how they use resources.
However, using the resources differently does not guarantee coexistence. Depending on their $p$’s and $q$’s, one isocline could still enclose the other completely. Figure $1$ has the same properties as Figure 16.5.4. Everywhere below the red isocline, Species 2 will increase, including the broad band between the red and blue isoclines where Species 1 will decrease.
If the blue and red isoclines are reversed, the result is similar, but with Species 1 excluding Species 2. Figure $2$ shows this, with the arrows reversed as Species 1 increases everywhere below the blue isocline, including the broad band between the isoclines where Species 2 decreases.
In all three cases, from Figures 16.5.4 to $2$, the system has three equilibria—at the origin (0,0), where both species are absent, at the carrying capacity $K_1$ for Species 1 alone (−$r_1/s_{1,1}$,0), and at the carrying capacity $K_2$ for Species 2 alone (0,$−r_2/s_{2,2}$). The origin is unstable and only one of the other two equilibria is stable, depending on which isocline encloses the other.
Figures $1$ and $2$ can be combined to give each species a chance to exclude the other, depending on circumstances. This means not allowing one isocline to completely enclose the other, as in Figure $3$. Intersecting isoclines introduce a fourth equilibrium at the interior of the phase space. This equilibrium is unstable, marked with an open circle, and the two equilibria for the individual species, on the axes, are stable. They are no longer “globally stable,” however, since only local regions of the phase space lead to either of them.
Depending on where the populations start, one of the two species will exclude the other. A curve called a “separatrix,” dividing these starting points according to which species will exclude the other, is shown with the dashed black line in Figure $4$. That separatrix corresponds to a long curved ridge in any surface above the phase space, as described in Chapter 10. It necessarily passes through the unstable interior equilibrium. Here it is a simple curve, though in some cases (such as the Mandelbrot system, not representing competition), related curves can be infinitely complex.
Finally, the isoclines can intersect in the opposite way, as in Figure $5$. In this case, neither species has any region where its isocline encloses the other, as seen in each of the phase spaces of Figures 16.5.4 through $4$. What will happen when neither species can exclude the other in any part of the phase space? They are forced to coexist. The individual equilibria on the axes become unstable and the interior equilibrium becomes stable—indeed, globally stable.
In summary, the isoclines in competitive systems have four different configurations, as in Figure $6$. Cases 1 and 2 can represent competition for a single resource or, equivalently, competition for two different resources that the two species handle identically. One of the two is a superior competitor that excludes the other. Case 3 is “bistable,” where either of the species can exclude the other, depending on how the system starts out. Finally, Case 4 is globally stable, where neither species can exclude the other, and stable coexistence prevails. Cases 3 and 4 can represent competition for two different resources. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/16%3A_Competition/16.07%3A_Multiple_resource_phase_space.txt |
Competition equations are usually presented in textbooks as the Lotka–Volterra competition model. This first appeared in the ecological literature in the 1920s and is defined not just in terms of the interactions among species, but also in terms of the species’ carrying capacities, as follows.
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,r_1(1-\frac{N_1+a_{1,2}N_2}{K_1})$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,r_2(1-\frac{N_2+a_{2,1}N_1}{K_2})$
$K_1$ and $K_2$ are the carrying capacities for Species 1 and 2, respectively. Parameters $a_{1,2}$ and $a_{2,1}$ represent the interference by each species on the other. If $a_{1,2}=2$, for example, each individual of Species 2 interferes with the growth of Species 1 as if it were two individuals of Species 1. If, on the other hand, $a_{1,2}=1/2$, it takes two individuals of Species 2 to have the same negative effect on the growth of Species 1 as one individual of Species 1 itself.
To compare this with the RSN formulation, represent Equations 16.7.1 and 16.7.2 with $i$ and $j$ subscripts
$\frac{1}{N_i}\frac{dN_i}{dt}\,=\,r_i(1-\frac{N_i+a_{i,j}N_j}{K_i})$
and multiply the right hand side through by $r_i$,
$\frac{1}{N_i}\frac{dN_i}{dt}\,=\,r_i\,-\frac{r_i}{K_i}N_i\,-\frac{a_{i,j}r_i}{K_i}N_j$
This shows that the Lotka–Volterra formulation is isomorphic to the RSN formulation. All conclusions about competitive systems examined thus far apply to the Lotka–Volterra formulation as well, with the appropriate translation of parameters. Parameter $r_i$ is the same in both in the Lotka–Volterra formulation and the RSN formulation, but $s_{i,i}= −r_i/K_i$ and $s_{i,j}=−a_{i,j}r_i/K_i$.
Beware, however, of a widely quoted statement derived from this formulation that appears throughout the ecological literature and textbooks. Statements like “Coexistence requires that each species inhibit itself more than it inhibits the other species” are abundant in textbooks and in the ecological literature. Alas, those statements are not correct.
box $2$
Coexistence in the Lotka–Volterra model requires that each species be able to increase from low densities when the other species is at its single-species equilibrium.
To see this, examine Figure 16.21. Here Species 1 inhibits itself with s11 =−1, while it inhibits the other species more strongly with s12 =−1.153. Yet there is global coexistence. Evidently, coexistence does not require that each species inhibit itself more than it inhibits the other, as in the conventional wisdom. The confusion in the literature has apparently emerged from the presence of the carrying capacity terms, K1 and K2, in the Lotka–Volterra formulation. These terms obscure the effects of the interaction terms, a1,2 and a2,1, when carrying capacities differ between the species.
What, then, is a correct statement about coexistence? It can be put in terms of increasing from low densities, as shown in Box $2$. The qualifying phrase “from low densities” is required, because the species not present can increase from high densities in a bistable system, as in Figure16.20, and flip it to the other state, even though coexistence cannot occur.
Another way in which coexistence has been explained is by Vandermeer (1981 Bioscience), connecting coexistence with a certain kind of “overyielding,” wherein two crops require less land for the same annual productivity when growing together than when growing apart.
The test is whether the joint equilibrium is above a line connecting the single-species equilibria (dashed gray in Figure $3$A ) or below the line (Figure $3$B).
This view is correct for models we have been considering with straight-line isoclines, but incorrect for more general models with curved isoclines (Figure $3$C,D ). The statement in Box $2$, however, holds true in each of these cases.
All things considered, rather than relying on rules of thumb, it can be better to evaluate a system directly, for example with the methods of eigenvectors and eigenvalues described in Chapter 10. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/16%3A_Competition/16.08%3A_Lotka-Volterra_formulation.txt |
• 17.1: Chapter Introduction
• 17.2: Phase Diagrams
• 17.3: Quasi-species
A new phenomenon arises in equations for unrestrained mutualism, one not possible for competition or predation. It can be shown that unrestrained mutualists are pulled into fixed ratios which are based on how their interaction terms differ from their self-limiting terms. As populations grow and move toward these fixed ratios, the two equations governing the individual species collapse, in effect, into an equation of a single-species.
• 17.4: Our own species
17: Mutualism
In Equation 8.1, mutualism occurs when the interaction terms, \(s_{1,2}\) and \(s_{2,1}\), are both positive. This is the defining feature of mutualism. The self-engagement terms, \(s_{1,1}\) and \(s_{2,2}\), may be positive or negative, and the intrinsic growth terms, \(r_1\) and \(r_2\), may also be positive or negative.
If both growth terms \(r_i\) are positive, then either species can survive on its own, without the presence of the other. This is called “facultative mutualism.” An example is the mutualism between oxpecker and warthog (Figure 9.2 left). Oxpeckers have other sources for food, and warthogs may suffer but likely not die from larger parasite loads.
If the \(r_i\) are both negative, then neither species can survive on its own without the presence of the other. This is called “obligate mutualism.” An example is lichen, composed of an algae and a fungus. If the fungus alone is placed on a rock, it will die. Although it can eat rock—or, more accurately, can dissolve nutrients from rock—it cannot produce carbon compounds from rock. And if the algae alone is placed on a rock, it will die. Although it can produce carbon compounds from the air through photosynthesis, it cannot eat rock. But the two together become a perfect partnership, each covering the other’s weaknesses.
When \(r_1\) is positive and \(r_2\) is negative, or vice versa, it is a partially obligate mutualism. One species depends on a second, but the second can get along without the first. This is a mathematical possibility, and some cases such as chloroplast– euglena may be examples (Figure \(1\) left), where the chloroplast cannot live without the protist, but the protist can.
Besides obligate versus facultative mutualism, there is another major distinction. This is between “restrained” and “unrestrained” mutualism. This distinction is not typically made in textbooks because it is related to the difference between logistic and orthologistic population growth, which typically is not covered.
When self-limiting effects of terms \(s_{1,1}\) and \(s_{2,2}\) are stronger than the interspecies enhancement terms \(s_{1,2}\) and \(s_{2,1}\), the mutualistic pair has a carrying capacity that can be computed from the properties of the individual species. They reach their joint carrying capacity along a path that looks like logistic growth, and is identical in form to that of two competing species that coexist and reach a joint carrying capacity. However, when the self-limiting effects of \(s_{1,1}\) and \(s_{2,2}\) are weaker than the inter-species enhancement terms \(s_{1,2}\) and \(s_{2,1}\), the carrying capacity of the pair of species cannot be computed from the value of those parameters, and they increase ever more rapidly until some other limitation is hit.
Think about lichen again. The carrying capacity of each species is essentially zero when they are living separately on the surface of a rock. Together, however, they can cover the rock, and their joint carrying capacity on Earth is related to the total area of rocks on Earth—plus other suitable habitat such as the bark of trees, the exterior surface of an automobile junked and neglected for decades, and myriad other surfaces otherwise uninhabitable. The area of rocks in the world, and other suitable surfaces, certainly is not part of the ri and si,j parameters! This is quite unlike the situation for competition and predation, in which the joint carrying capacity is encoded in the parameters. Why this should be is revealed by some diagrams and corresponding mathematics. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/17%3A_Mutualism/17.01%3A_Chapter_Introduction.txt |
The principles can be visualized in phase diagrams, with arrows showing how populations change. Earlier, Figure 10.11 showed strong competition with coexistence at equilibrium, as could be the case if the species were competing for two different resources. Nonetheless, each species restricts the other to much lower levels than it could maintain on its own. Weaker competition means that each species is restricted less so each can maintain higher levels, as in Figure 17.1.2. You can tell this is competition because the two parameters $s_{1,2}$ and $s_{2,1}$ are both negative. This is shown in the negative slopes of the two diagrams in the upper left. The carrying capacity of each species together is only slightly reduced from the individual carrying capacities, which would be about 1.2 for Species 1 living alone and about 0.8 for Species 2 living alone. (for example, an average of 1.2 individuals per square meter, or 1,200,000 individuals per square mile if measured in millions). However, for species living together, the carrying capacity of each is slightly reduced, perhaps 10 to 20 percent.
In this case, by the way, the two species together have a higher total population than would be the case if either was living alone. This is called “over yielding,” and is a recurrent theme in studies of plant communities.
Figure 17.1.3 shows a similar situation, but now with the inter-species interaction terms $s_{1,2}$ and $s_{2,1}$ both positive, shown by the positive slopes in the two upper-left diagrams of the figure. It looks quite similar to Figure 17.1.2, but the two together are each more abundant than they would be apart— the joint equilibrium is larger than the individual carrying capacities.
This joint equilibrium can be computed from the $r_i$ and $s_{i,j}$ parameters. It will occur where the growth of each species simultaneously reaches 0. You can find the numerical value for this equilibrium with pencil and paper by setting the first species growth rate to 0, solving for the populations of Species 1, substituting that into the equation for Species 2, and solving for when the growth of that species reaches 0. Alternatively, you can pose the problem to a symbolic mathematics program and ask it to solve the two equations simultaneously. In any case, you would start with both growth rates set to zero at equilibrium,
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,r_1\,+\,s_{1,1}N_1\,+\,s_{1,2}N_2\,=\,1.2\,-\,1N_1\,+0.1N_2\,=\,0$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,r_2\,+\,s_{2,2}N_2\,+\,s_{2,1}N_1\,=\,0.8\,-\,1N_2\,+0.1N_1\,=\,0$
and end up with $N_1$ = 1.2929 and $N_2$ = 0.929.
As mutualisms become stronger—meaning that the interspecific interactions become more positive—the equilibrium point moves further out. It can be very large, as in in Figure 17.1.4, but, in restrained mutualism, the equilibrium is finite and computable from the parameters of the individual species.
On the other hand, when the inter-species enhancement terms are stronger still, a bifurcation occurs and the joint equilibrium ceases to exist at all. (Figure 17.1.5). The calculated equilibrium point has, in effect, moved to infinity, or in a sense beyond, meaning that the carrying capacity cannot be computed from the parameters of the species and their interactions. Some further information is needed about the system.
Beyond this, the mutualists can become more dependent on each other, so that the $r_i$ terms become smaller, as in Figure 17.1.6, or negative, as in Figure 17.1.7. The mutualism can be unrestrained even if the intrinsic growth rates $r_i$ are negative. What arises is a kind of Allee point, where the populations run away if they start above that point, but decline to extinction if they start below. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/17%3A_Mutualism/17.02%3A_Phase_Diagrams.txt |
A new phenomenon arises in equations for unrestrained mutualism, one not possible for competition or predation. It can be shown that unrestrained mutualists are pulled into fixed ratios which are based on how their interaction terms differ from their self-limiting terms. As populations grow and move toward these fixed ratios, the two equations governing the individual species collapse, in effect, into an equation of a single-species. This single species grows orthologistically, a form of growth you saw earlier in Figure 4.3. The system behaves mathematically as if only one species were participating.
What does the mathematical collapse to a single equation mean biologically? It suggests that the biological world could respond analogously—that two actual species growing as unrestrained mutualists could intermingle into a single quasi-species—at least into the best approximation of a single species that biology could accomplish using genetically different entities. Lichen and the eukaryotic cell are examples, genetically separate but biologically merged. Indeed, lichens were thought to be individual species until the nineteenth century, and the eukaryotic cell was only accepted as the result of mutualistic combinations late in the twentieth.
It therefore appears that natural selection has not overlooked this possibility. In a community model with mutualistic species locked in approach to a singularity, the two mutualists may be replaced in the model by a new quasi-species, representing the two species jointly but ultimately having a non-singular form of population growth. The resulting quasi-species may grow without obvious inhibition toward a singularity, then switch to a different model, as you have seen for single-species models in Figure 4.4 and for our own species in Figure 6.3.
17.04: Our own species
We reached our present numbers through a complex of mutualisms. A variety of increasingly refined crops and increasing domesticated animals acted as mutualists, switching our population dynamics from boom-and-bust predator–prey cycling (Figure \(2\)A) to prolonged runaway growth (\(2\)B). Through cultivation of plants and animals that were formerly prey, agriculture increased the carrying capacity of the cultivated species, which in turn increased human carrying capacity, forming a positive feedback loop. It may seem dubious to consider domesticated plants and animals as our mutualists, but their parameters meet the mathematical requirements of mutualism. And both populations can expand so rapidly that they would soon exceed all former bounds.
Mutualisms in nature typically do not long follow the unrestrained growth toward astronomical levels shown in Figure \(2\)B. Instead, natural mutualisms increase only until they are checked by some other force, such as predators or disease. Our ancestors, however, kept their mutualisms pure by eliminating third-species interferences. They hunted and killed predators of domesticated animals, weeded crops to eliminate plant competitors, and fenced crops to exclude herbivores. Indeed, the human species does not now dominate the earth, as often suggested. The mutualisms dominate. The joint biomass of cows, horses, pigs, chickens, and dogs exceeds the biomass of humans, to say nothing of the biomass of crops. We did not get here alone.
We have recently outgrown the need for animal mutualists to power plows and support locomotion, substituting engines consuming fossil fuels instead. These are like synthetic mutualists. We also outgrew the need for animals to fertilize crops, substituting artificial nitrogen fertilizer created in Haber–Bosch process furnaces from the vast supply of nitrogen in the air. The need for animals and plants for clothing diminished with synthetic cloth and furs. All of these brought us to a partially non-biological world, still developing, where not even the carrying capacities of our living mutualists need be limiting.
However, advancing medicine and public health eventually decreased child mortality. Death rates fell and, later, birth rates began to fall even faster. The mutualisms remain, but—unexpectedly and abruptly— twelve millennia of rapidly accelerating population growth ended (Figures 6.3, \(2\)C).
As a closing note, in the twenty-first century, after only about fifty years of decelerating growth, the global population continues to increase. But, births in many societies have dropped below replacement levels in a final stage characterized by declines in the resident populations, compensated in many countries by immigration. The relative rapidity of this change may spill over into social disjunction between conditions of the present and memories of times past. It may also foretell a future time when our species may voluntarily reduce populations to sustainable values yet unknown (dotted lines in Figure \(2\)C). | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/17%3A_Mutualism/17.03%3A_Quasi-species.txt |
The RSN formulation of Equations 4.1 and 8.1 incorporate the classical competition, predation, and logistic equations of ecology, plus mutualistic and orthologistic equations that are not part of the classical set. For this, the parameters can remain constant.
But they are not constant in all cases. Volterra (1928) considered parameters that vary through time, but more generally the parameters can be functions of population density as well. In this way the RSN formulation can be extended to cover not just the classical equations of ecology and related equations, but any equations of ecology whatever, as with Equation 4.3, proposed by Kolomogorov.
18.02: One-species Example
To see the basic idea, start with a single species. With variable parameters, population growth for a single species can be written as follows, with the growth rate $r$ and the density dependence term $s$ depending on the population level $N$ and optionally on the time $t$.
$\frac{1}{N}\frac{dN}{dt}\,=\,r(N,t)\,+\,s(N,t)N$
Because anything can be embedded in the term $r(N,t)$, or in the corresponding $s(N,t)$, the equation is perfectly general and can cover any ecological situation for a single species modeled by a differential equation. A basic example, which you have already seen for human population growth, is where the parameters are approximately constant for long periods, but change at certain “bifurcation” events. The parameters for human population growth changed abruptly at the beginning of the modern era, resulting in overall population dynamics that were neither orthologistic nor logistic, but a piecewise combination of the two. In that case the two were blended by varying the parameters as follows.
$\frac{1}{N}\frac{dN}{dt}\,=\begin{cases}-&0.001185\,&+\,0.00684N,\,&\text{when}\,N\leq\,3.28\,\text{billion}\&0.03077\,&-\,0.00289N,\,&\text{when}\,N\gt\,3.28\,\text{billion}\end{cases}$
This led to the population growth curve of Figure 6.3, which nicely modeled human population growth over the centuries.
Human population growth called for a piecewise blending of the parameters, because the parameters changed rather abruptly from one constant set to another. The parameters can also be blended continuously for parameters that change gradually.
For example, take an orthologistic equation, $1/N\,dN/dt\,=\,−2+2N$, and a logistic equation, $1/N\,dN/dt\,=\,4−2N$, and consider $N$ as it ranges from 0 to 1. Have the orthologistic equation apply exactly as $N$ approaches 0, and the logistic equation apply exactly when N reaches 1. Then let $r$ change uniformly from −2 to +4 and $s$ change uniformly from +2 to −2 as $N$ goes from 0 to 1, as follows.
$r(N,t)\,=6N-2\s(N,t)\,=-4N+2$
Plugging that into Equation 18.1 gives
\begin{align}\frac{1}{N}\frac{dN}{dt}&=\,(6N-2)\,+\,(-4N+2)N\&\,=\,-2\,+\,8N\,-4N^2\&\,=\,r\,+\,sN\,=s_2N^2\end{align}
This kind of blending between orthologistic and logistic has simply added one more term to the population growth equation, an $N^2$ term—one of the terms proposed by Hutchinson (Equation 4.2). The result is graphed in Figure $1$.
This smoothly blends the orthologistic, which has an Allee point but no carrying capacity, with the logistic, which has a carrying capacity with no Allee point, providing both in the blended curve of Figure $1$. The curve has an Allee point at about $N\,=\,0.3$ and a carrying capacity at about $N\,=\,1.7$. Compare this with the piecewise blending depicted earlier in Figure 4.4.
18.03: Two-species blending
Blending two-species systems is a similar process, but has more options in the parameters. Equation 18.3.1 is an example with limited options that produced the phase spaces in Figures 10.1.3 through 10.1.5.
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,r_1(N_1)\,+\,s_{1,1}N_1\,+\,s_{1,2}(N_1)N_2\\frac{1}{N_2}\frac{dN_2}{dt}\,=\,r_2(N_2)\,+\,s_{2,2}N_2\,+\,s_{2,1}(N_2)N_1$
Changing the parameters uniformly from $a\,b$ value when the corresponding $N$ value is 0 to an $a\,+\,b$ value when the corresponding $N$ value is 1 is analogous to the blending that produced Figure 4.4.1. The parameters would vary as follows, using four distinct $a$ values ($a_1,\,a_2,\,a_{1,2},\,a_{2,1}$), plus four distinct $b$ values with matching subscripts ($b_1,\,b_2,\,b_{1,2},\,b_{2,1}$).
$r_1(N_1)\,=\,a_1N_1\,+\,b_1,\qquad\,s_{1,2}(N_1)\,=\,a_{1,2}N_1\,+\,b_{1,2}$
$r_2(N_2)\,=\,a_2N_2\,+\,b_2,\qquad\,s_{2,1}(N_2)\,=\,a_{2,1}N_2\,+\,b_{2,1}$
Substituting the above into Equation 18.3.1 and collecting terms gives an equation having all the RSN terms present, but now with a cross-product in terms of $N_1N_2$ added at the end:
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,b_1\,+\,(a_1\,+\,s_{1,1})N_1\,+\,b_{1,2}N_2\,+\,a_{1,2}N_1N_2\\frac{1}{N_2}\frac{dN_2}{dt}\,=\,b_2\,+\,(a_2\,+\,s_{2,2})N_2\,+\,b_{2,1}N_1\,+\,a_{2,1}N_1N_2$
In the specific case of Figures 10.1.3 through 10.1.5, we used $s_{1,1}\,=\,s_{2,2}\,=\,−0.98$ and
$r_1(N_1)\,=\,0.75N_1\,−0.5\qquad\,s_{1,2}(N_1)\,=\,−1.15N_1\,+\,2.5$
$r_2(N_2)\,=\,0.75N_2\,−0.5\qquad\,s_{2,1}(N_2)\,=\,−0.45N_2\,+\,1.3$
which gave
$\frac{1}{N_1}\frac{dN_1}{dt}\,=\,-0.5\,-\,0.23N_1\,+\,2.50N_2\,-\,1.15N_1N_2$
$\frac{1}{N_2}\frac{dN_2}{dt}\,=\,-0.5\,-\,0.98N_2\,+\,2.50N_1\,-\,0.45N_1N_2$
for the flow in the figures.
18.04: A way of organizing
The methods we have described in this chapter provide an organized way of designing higher-order models of ecological interactions. Characteristics of the individual models purely in RSN form can be mixed and matched as needed to produce the higher-order model, then the resulting parameters can be fit to observations or modeling needs.
For example, the bistable phase space in Figures 10.1.3 through 10.1.5, with an unstable Allee point combined with a stable carrying capacity, was easy to achieve by combining sub-models, each with one of those characteristics. The result was simply an RSN system with an \(N_1N_2\) cross-product term added, but how easy would it have been to conceive that term a priori and perceive its sufficiency?
Moreover, a blended model in the form of Equation 18.3.1 can be visualized in terms of (1) the intrinsic growth rate of each species, (2) how each species affects the other, (3) how each species affects itself, and (4) how those effects change with population densities. In a way, any higher-order system constructed in this way can be viewed as an RSN system at every point, yet as a higher-order system overall.
In summary, we cautiously recommend this method of creating and organizing ecological models, and are eager to hear of successes and failures you might have with it.
The characters acting in this book—the equations of population growth, the parameters and their relationships, the graphs and phase planes—live in variations and disguises within an enormous variety of actual ecological applications, detailed in thousands of scientific papers you will be able to sample as your journey continues. We hope you have enjoyed this introductory book and that it has been useful. We welcome your comments on any aspects at all. An ebook is never finished but can constantly be improved.
—C.L., S.L., A.C. | textbooks/bio/Ecology/Book%3A_Quantitative_Ecology_-_A_New_Unified_Approach_(Lehman_Loberg_and_Clark)/18%3A_Higher-order_models/18.01%3A_Chapter_Introduction.txt |
Popular interest in protecting biological diversity—which describes the amazing range of species, genetic diversity within each species, and the multitude of Earth’s complex biological communities with their associated ecosystem processes—has intensified during the past few decades. During this time, scientists and the public have recognized that biological diversity (often shortened to biodiversity) is being lost at increasing rates. Across the world, human activities are destroying ecological communities that have developed over millions of years. Over the next several decades, thousands of species and millions of populations will likely go extinct.
The fundamental driver of all the biodiversity losses we are currently witnessing is a rapidly expanding human population coupled with increased consumptive needs. In 1850, after roughly 300,000 years of Homo sapiens on the planet, there were around 1 billion people on Earth. By 1987, not even 140 years later, the world’s human population had surpassed 5 billion. By 2017, there were 7.5 billion humans globally, of which over 1 billion lived in Sub-Saharan Africa (World Bank, 2019). With this many people, the human population grows by tens of millions of people each year, even with modest population growth (Figure 1.1). To make matters worse, Sub-Saharan Africa has the fastest population growth rate in the world, with a projected human population estimate of over 4 billion people by the year 2100—a number that is well beyond the ecological capacity of the region to support.
To survive and prosper, people use natural resources. They harvest and use oil, water, and wildlife products, and convert natural ecosystems for agriculture, cities, roads, and industrial activities. This consumption, which reduces natural habitat and the associated wildlife populations, is intensifying because of the demands of a rapidly increasing human population. Consumption of resources also increases as countries develop and industrialize: the average citizen of the USA uses five times more resources than the average global citizen, 11 times more than the average Chinese citizen, and 32 times more than the average Kenyan citizen (Worldwatch Institute, 2015). This growth in the number of humans, together with their ever-more-intensive use of natural resources, is the fundamental driver behind most current species extinctions.
For conservation biologists and other nature lovers, the widespread extinction of species and destruction of natural ecosystems are incredibly discouraging.
For conservation biologists and other nature lovers, the widespread extinctions of species and destruction of natural ecosystems are incredibly discouraging. Perhaps nowhere in the world is this issue as dramatic as in Africa with its rich and spectacular wildlife, but also its significant socio-economic challenges, such as a rapidly increasing human population, persistent poverty, weak governance structures, and many people’s near-obligate dependence on natural resources. Many Africans are also confused by the importance and need for conservation actions, pointing to the romanticised but inaccurate notion that humans have been living in relative harmony with nature since humans first made an appearance on Earth (see Box 8.1). But it is possible, and indeed necessary, to find ways to ensure the persistence of biodiversity. Actions taken, or not taken, during the next few decades will determine how many species and natural areas will continue to survive. Someday, people will likely look back and say that this time—the first half of the 21st century—was an important and exciting time when people worked together, and acted locally and globally, to prevent the extinction of many species and ecosystems. Examples of successful conservation efforts are described throughout this textbook. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/01%3A_What_is_Conservation_Biology/1.00%3A_Prelude_to_What_is_Conservation_Biology.txt |
As a distinct scientific field, conservation biology is an integrated, multidisciplinary subject that developed in response to the challenge of preserving populations, species, ecosystems, and biological interactions. The main aim of conservation biology is to ensure the long-term preservation of biodiversity. To achieve its aim, conservation biology has set three goals:
• To document Earth’s biological diversity.
• To investigate how humans influence species, evolution, and ecosystem processes.
• To investigate practical approaches to protect and restore biological communities, maintain genetic diversity, and prevent the extinction of species.
The first two goals describe typical scientific research investigating objective facts. The third goal, however, is a part of what makes conservation biology a normative discipline; that is, conservation biology incorporates human values, not just facts, to understand and achieve its value-laden goals (Lindenmayer and Hunter, 2010). In this sense, conservation biology is related to environmentalism, in which people aim to protect the natural environment for its own sake (see Section 4.3.2). However, conservation biology is at its core a scientific discipline; it is founded on scientific principles. This is not to say you must be a scientist to practice conservation biology; there are many people who are not scientists who apply the principles of conservation biology in their professional and personal lives.
The emergence of conservation biology as a distinct scientific field in the 1970s has given rise to the formation of various formal societies representing the field in a united voice. Most notable among these is the Society for Conservation Biology (SCB, Figure 1.2), which is a non-profit international professional organization with a mission to advance “the science and practice of conserving the Earth’s biological diversity”. To facilitate opportunities where like-minded people can share ideas locally, the SCB has regional branches, including an active Africa Section (http://conbio.org/groups/sections/africa) which hosts regular conferences. In addition to the SCB, a great number of other local, national, and regional conservation organizations also exist and act as mouthpieces for grassroots movements and as custodians of nature. Many of these groups focus on specific animals or local protected areas. Others organically adapt their missions and visions in response to a specific need or threat. For example, established in 1913 as an exchange forum between collectors of rare plants, the Botanical Society of South Africa now actively works toward protecting those rare plants in their natural habitats.
In recent years, conservation practice has evolved from just a plan to save the environment to a vision that includes sustainable development and social justice.
Conservation biology also has a history of adapting to new challenges. The very first conservation activities, in Africa and beyond, were geared towards securing the rights to valuable natural resources for people in powerful positions, such as kings and tribal chiefs, enforced through a strictly adherence to cultural norms and customary laws (Section 2.2). But as a growing human population expanded its influence on the environment, and wildlife started to decline, earliest conservation models gradually shifted towards fortress conservation approaches (Wilshusen et al., 2002) which aimed to shield wildlife from people by setting aside protected areas where human activities were strictly controlled.
Today, however, as human populations are exploding, and consumption is increasing, even protected areas are increasingly unable to withstand the multitude of threats to biodiversity that ignore property boundaries and political borders. In response, fortress conservation approaches are beginning to make way for large-scale integrated activities that highlight the social and economic benefits of biodiversity conservation. To do this, new alliances are being formed and new agendas are being established, such as those that directly link human health with environmental health (Box 1.1). These integrated conservation philosophies that pursue strategies that benefit both humans and biodiversity show much promise because they focus on fundamental extinction drivers, and advocate for more inclusive sustainable development. In this way, the practice of conservation has evolved from just a plan to save the environment to a vision that accomplishes its goals through sustainable development and social justice.
Yet, as we consider how to best invest limited conservation resources, some difficult questions arise. With seemingly more work to be done than can be accomplished, should we let some species go extinct (Bottrill et al., 2008)? Which species? Who decides? How can we even dare to think that we can play god? Such questions predictably bring about strongly opinionated and emotional debate (Soulé, 2013 vs. Marvier, 2014; Tallis and Lubchenco, 2014). Given the successful track record of fortress conservation initiatives in preventing extinctions despite limited budgets (Young et al., 2014), as well as the promising progress of more complex people-centred initiatives (Pooley et al., 2014), it seems clear that conservation relies on some balance between these two conservation philosophies (Sodhi et al., 2011). Conservation biologists of tomorrow will be able to fine-tune the balance between these strategies by closely inspecting the successes and failures of our actions today.
Box 1.1 Conservation Through Public Health: A Case Study
Gladys Kalema-Zikusoka
Conservation Through Public Health,
Kampala, Uganda.
Conservation Through Public Health (CTPH) is a grassroots non-governmental organization (NGO) and non-profit that promotes biodiversity conservation by enabling people, wildlife, and livestock to coexist. The organization was founded in 2003 after fatal scabies skin disease outbreaks in mountain gorillas (Gorilla beringei beringei, EN) were traced to people living around Bwindi Impenetrable National Park, Uganda, who had limited access to basic health services (Kalema-Zikusoka et al., 2002). Since then, CTPH has contributed to conservation and sustainable development in Africa by improving human and animal health and welfare in and around protected areas.
One of the main goals of CTPH is to reduce disease transfer between humans and gorillas. We accomplish this through an integrated population, health, and environment (PHE) program that was established in 2007 with funding from the US Agency for International Development (USAID). As a first step, piloted around Bwindi, CTPH held consultative meetings with local leaders, during which at least one Village Health and Conservation Team (VHCT) volunteer was selected from each village and two from each parish (consisting of 11 villages) to oversee distribution of family planning supplies. This initiative rapidly expanded into a sustainable social service delivery network that promotes family planning, hygiene, and sanitation. The network resulted in a 20% to 60% (national average is 30%) increase in new users to modern family planning, and a 10% to 60% increase in adoption of hand washing facilities at homes visited by VHCTs. VHCT volunteers also refer people suffering from infectious diseases and malnutrition to local health centres and promote more sustainable alternative livelihoods. Another group of community volunteers, the “Human and Gorilla Conflict Resolution” (HUGO) team, in turn collect gorilla faecal samples left on communal land to monitor their health (Figure 1.A), and visually monitor gorillas for clinical signs of disease inside and outside protected areas (Gaffikin and Kalema-Zikusoka, 2010). In the process, we have seen reduced disease incidences in the gorillas, reduced conflict between people and gorillas, and improved attitudes toward conservation. One unintended outcome has been increased gender equality: men are now more involved in family planning, and women are more involved in natural resource management.
Our experience in initiating and managing PHE programs for the past 10 years has taught us several lessons. One of the most important lessons to ensure project sustainability is to regularly engage with local leaders and the government. The Uganda Wildlife Authority, Uganda’s Ministry of Health, and local health centres all attend CTPH meetings with VHCTs. Attendance by and representation of these groups not only informs them of our activities, but also provides a platform to inform or train the VHCTs in what they would like them to disseminate to the local communities.
We have also learnt that PHE-implementing partners and projects need to be well-suited to each other and each site; this remains true even though health needs are often the same, regardless of the location. For example, at Mount Elgon National Park in Uganda, we found that training VHCTs in reducing conflict with park management played a key role in changing community attitudes toward conservation. In contrast, at Virunga National Park, Democratic Republic of the Congo (DRC), we found that VHCTs needed to work more closely with local health centres to prevent disease transmission between people and gorillas, and to promote family planning in a largely Catholic country.
Lastly, we found that establishing income-generating projects for groups rather than individuals was key to sustaining VHCT networks and program goals beyond donor funding cycles where we have had no volunteer dropouts in the first 10 years of initiating the PHE program. These key components were accomplished by initiating livestock group enterprises and by encouraging VHCT volunteers to invest generated income into Village Saving and Loan Associations (see http://www.care.org/vsla). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/01%3A_What_is_Conservation_Biology/1.01%3A_Conservation_Biology_is_Still_Evolving.txt |
While there are a few extinctions that have only one cause, more generally, extinctions occur because several factors acted simultaneously and/or sequentially. Blaming a certain industry or specific group of people for an extinction (or other biodiversity loss) is thus simplistic, ineffective, and often counter-productive. Though challenging, a better approach would be to better understand how local, national, and international links led to those losses, and to find viable alternatives to prevent it from happening again. To succeed in this challenge, conservation biologists should strongly consider taking on one or more of several roles:
• Conservation biologists should be curious. The world around us is full of natural wonders waiting to be discovered. These discoveries underpin conservation action, by allowing us to define all the different components of biodiversity, enabling us to better understand the needs of different species, and providing us with opportunities to celebrate our conservation successes.
• Conservation biologists must be good listeners. Sometimes, the only difference between attracting a new ally and making an enemy, or between developing a landscape and saving a species from extinction, is the way we communicate. Conservationists must be careful and respectful listeners, especially to opposing perspectives. Careful listening is particularly important in rural areas, where villagers often have practical concerns related to their daily contact with wildlife, such as staying safe and preventing crop damage and livestock loss. Quite often, those villagers may also have unique insights into wildlife ecology that could prove valuable in local conservation measures.
• Conservation biologists must be law-abiding citizens. Activities that involve wildlife and ecosystems are regulated by laws and regulations. These laws are important because ethical boundaries differ from person to person—activities acceptable to one group of people may be considered harmful by another. As conservation biologists, abiding by environmental laws is especially important if we want others to take those laws seriously.
Laws are important because ethical boundaries differ from person to person—activities acceptable to one person may be immensely harmful to another.
• Conservation biologists should become effective communicators. They should be able to discuss the problems facing biodiversity in depth, as well as the consequences of losing biodiversity, to as broad a range of people as possible. Groups like hunters, community leaders and organisers, and church leaders may be interested in participating in conservation efforts once they recognize that their activities, health, and emotional well-being depend on conservation action.
• Conservation biologists could become politically active leaders, so that they can influence public opinion and policy. As a starting point, those interested in this role can join a conservation organization to learn more about broader issues. They could also use their personal networks to form alliances with lawyers, citizen groups, and politicians.
• Conservation biologists could become pro-active land managers. Those taking on this task must be willing to walk on the land and go out on the water to find out what is really happening. They should also talk with local people to communicate their knowledge to others in ways that are clear and easily understood.
• Above all, a conservation biologist must be honest. To encourage effective action, both from the public and through policy, conservationists must present arguments backed by reliable evidence. To do otherwise, conservation biologists could lose credibility, which would very likely delay or even compromise conservation efforts.
It is worth taking a moment to distinguish between two important pillars of conservation action, namely conservation advocacy and conservation science. Conservation advocacy describes the roles that conservation biologists adopt to guide social, political, and economical systems towards a personally-preferred outcome—adopting environmentally-friendly practices; incorporating these activities makes conservation biology a normative discipline. Conservation science, in contrast, describes activities that conservation biologists undertake to generate knowledge, like objectively describing biodiversity and measuring biodiversity’s response to stressors and safeguards. While conservation advocacy and conservation science often support and inform each other as to the next steps required for “doing conservation”, it is important to distinguish between these two pillars to ensure that policy makers and other stakeholders in the environment understand when we advocate for personal preferences and when we offer objective findings (Rykiel, 2001; Lackey, 2007; Nelson and Vucetich, 2009). The next section will further expand on the importance of science in conservation biology. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/01%3A_What_is_Conservation_Biology/1.02%3A_The_Role_of_Conservation_Biologists.txt |
The field of conservation biology applies scientific methods to achieving its goals. Like the medical sciences, which apply principles from physiology, anatomy, and genetics to problems of human health, conservation biologists solve biodiversity problems using principles from fields, such as mathematics, veterinary medicine, social sciences, and several natural sciences (Figure 1.3). Conservation biology differs from these and other component disciplines in that its primary goal is the long-term preservation of biodiversity. Unlike many other scientific fields, conservation biology can also be described as a crisis discipline (Soulé, 1985; Kareiva and Marvier, 2012). That is, conservation biologists are often required to take creative steps to respond to imminent threats, typically without a complete knowledge of the systems requiring attention. Conservation scientists must also articulate long-term visions for conservation beyond solving immediate problems.
To be effective, conservation biologists must demonstrate the relevance of their findings to a range of stakeholders. To be successful in this task, the importance of sound scientific principles cannot be over-emphasised. Nature is a complex network of many interdependent connections and feedback loops. Science is underpinned by principles that provide conservationists the necessary quantitative and qualitative tools to better measure and control for all these different aspects of biodiversity. Such measurements allow us to gain a better understanding of complex natural systems, and the consequences of human activities. Reliable, unbiased data obtained from sound and transparent scientific methods also facilitate policy making that is too often based on value judgments by non-experts who must balance many needs and different sources of information (Ntshotsho et al., 2015).
One of the cornerstones of modern science is to identify a hypothesis (a proposed explanation for a specific observation) to evaluate. The best hypotheses, often expressed as goals or objectives, are usually those that are SMART:
Setting specific, measurable, realistic and timebound goals and objectives is essential for effective conservation.
Specific: not overly general;
Measurable: has both units and a method of measurement;
Attainable: realistic to achieve;
Relevant: related to what needs to be accomplished;
Time-bound: achievable within a specific timeframe.
Identifying SMART goals and objectives is an essential aspect of conservation biology. Without such benchmarks, practitioners cannot know whether their tasks were successful, or when management actions should be adjusted to achieve success. While this may seem obvious, many previous conservation projects have failed because biologists neglected to set SMART goals and objectives (Tear et al., 2005). While lofty, “We’re going to save all species” is not a SMART conservation goal because it is overly general, hard to measure, unrealistic, and not time bound. In contrast, “We want to protect 25% of our country’s wetlands within the next 10 years” is a SMART goal because it sets a very clear and measurable objective. In general, it is wise to set smaller short-term (e.g. quarterly), and medium-term (e.g. annual) goals as one works towards long-term (e.g. 5–10 years) objectives; this allows one to constantly assess progress, which in turn provides opportunities for celebrations and strategic adjustments as and when needed.
Too many conservation decisions are based on biased anecdotes, personal intuition, and myths.
Another scientific standard that conservation practitioners must adopt at a larger scale is the transfer of knowledge gained from unique and specialist experiences. Conservation activities are too often hampered by the lack of guidance from credible and available sources. This forces conservationist managers to base important decisions on biased anecdotes, personal intuition, and even myths (Sutherland et al., 2004). Successful conservation actions on the other hand often rely on results and guidelines that were disseminated to the broader community by practitioners who faced similar challenges earlier. To maximise this learning from each other’s successes and mistakes, it is crucial for conservation scientists and managers to make every effort to ensure knowledge transfer, by carefully tracking their activities, and publishing their results and experiences in scientific journals and reports.
Public outreach builds on the public’s existing connection to nature and helps them better understand the value of local biodiversity.
Biodiversity conservation, however, is not accomplished by simply setting SMART goals, measuring outcomes, and publishing results in scientific journals and reports. It is also important for conservation biologists to engage in public outreach activities, during which they can build on the public’s existing connection to nature, help them better understand the value of biodiversity in their local area, and enable them to actively contribute in conservation projects. When interacting with the public, conservation practitioners must be sensitive to the complicated emotions and diverging interests of different groups of people (Milfont et al., 2017), especially vulnerable peoples who may be negatively impacted—hopefully only in the short term—by conservation actions. This requires a sense of emotional awareness, because the words we choose matter when we encourage others to care for and reduce their impact on nature. Equally important, conservation biologists, as with any field of science, should be sceptical of their results. The process of generating data is not equal to generating facts, because data can be fraught with bias, imprecision, and uncertainty. This is perhaps even more important when sharing findings with lay people, as scientists have rigorous training in understanding uncertainty and connecting cause and effect. Putting scientific findings in context with adequate and clear explanation is a challenge to all scientists, but it is necessary, especially when partnering with conservationists not specifically trained as scientists. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/01%3A_What_is_Conservation_Biology/1.03%3A_The_Value_of_Scientific_Methods.txt |
Most human societies today aim to protect biodiversity through rules and regulations (Chapter 12). An alternative approach is to change the fundamental materialistic values of modern society to values that prioritise genuine and lasting human well-being. This is the goal of environmental ethics, a discipline within philosophy that emphasises the ethical values of biodiversity. The foundation of environmental ethics lies in the philosophical principle that every organism of Earth has a right to exist, regardless of its usefulness to humans, so any action that negatively impacts biodiversity would be considered unethical.
Because human quality of life is intricately linked to the ability of the natural world to prosper (Chapter 4), the ethical arguments for biodiversity conservation hold even for people who value only human life. Or, in other words, respect for human life—even our instincts for self-preservation—should compel us to preserve biodiversity. In contrast, if we neglect our assumed responsibility to act as guardians of life on Earth, future generations will suffer with a lower quality of life. We can already see signs of this today: as species are lost and natural ecosystems replaced with sprawling cities, children are increasingly deprived of the wonderful experience of seeing a ‘new’ animal (Figure 1.4) or pretty flower. We can imagine that we are borrowing Earth from future generations, and that it is our responsibility to ensure that they receive it in good condition.
Because of this close link between nature and human well-being, the concept of nature preservation has permeated through the value systems of most human cultures, philosophies, and religions throughout history. This is especially relevant in Africa, where most (if not all) traditional societies have a deep connection with nature that is woven into their spiritual beliefs and customs (Figure 1.5). Our responsibility to protect animals is also explicitly described in Jewish, Christian, and Islamic traditions. Other major religions, including Hinduism, Buddhism, and Taoism also strongly support the preservation of non-human life. In light of accelerated biodiversity losses, faith-based groups have recently started playing a more active role in conservation, particularly among urbanised people. They do this by informing adherents that it is wrong to allow the destruction of nature, and that such destructive activities can have negative consequences for all people on Earth. These links between faith-based organizations and conservation have given rise to consortiums such as the Forum on Religion and Ecology (http://fore.yale.edu), the Alliance of Religions and Conservation (http://www.arcworld.org), and the SCB’s Religion and Conservation Working Group (https://twitter.com/ReligionConBio), as well as the emerging field of spiritual ecology (Vaughan-Lee, 2016).
Environmental ethics has strong links to the environmental justice movement and has recently established strong ties to the social justice movement. Some of the most exciting developments in this direction involve initiatives that combine protection activities with community upliftment programs that improve the well-being of local peoples (Box 1.2; see also Section 14.3). These developments have shown that when poor and marginalised people are empowered to protect the environment, they may act as strong local guardians of forests, coastal areas, and other ecosystems that may have been destroyed otherwise.
Box 1.2 The Okapi Wildlife Reserve: Protecting Nature and Providing for People
Rosmarie Ruf & Marcel Enckoto
Okapi Conservation Project,
Okapi Wildlife Reserve,
Epulu, DRC.
The Okapi Wildlife Reserve, a World Heritage Site in peril, is located within the dense, tropical Ituri Forest in north-eastern DRC. The reserve was created to protect the okapi (Okapia johnstoni, EN) (Stephenson and Newby, 1997). Researchers estimate that there are between 10,000 and 25,000 okapi (Figure 1.B) left in the wild, but with populations appearing to have declined by more than 50% over the last 15 years (Kümpel et al., 2015). The reserve also protects charismatic species like forest elephants (Loxodonta cyclotis), 14 species of primates, including chimpanzees (Pan troglodytes, EN), leopards, forest buffalo (Syncerus caffer nanus), and bongo antelope (Tragelaphus eurycerus, NT).
Since 1987, the Okapi Conservation Project has partnered with the Congo Institute for Conservation of Nature, the government agency responsible for the Okapi Wildlife Reserve’s management, to provide financial and technical support for the operation of the reserve and preservation of the surrounding Ituri Forest. The project is partially managed and funded by Wildlife Conservation Global, a non-profit NGO based in Florida, USA.
Despite the support, conservation managers in this region face various challenges due to political instability since the 1990s. This has led to the deaths and displacement of millions of residents and rampant poaching in the area and beyond. Epulu Station, the reserve’s headquarters, was tragically attacked in 2012, resulting in hostage taking, destruction and looting of the headquarters’ infrastructure, and the deaths of staff and families, as well as 14 okapis at the captive breeding station.
To meet its goals in this difficult environment, the Okapi Conservation Project has seven objectives aimed at enhancing conservation, safety, and community:
• Financially supporting the operation of the reserve, paying warden and guard bonuses, building informer and monitoring networks, and providing necessities, such as food rations for patrols, fuel and spare parts for travel and field equipment.
• Maintaining and building infrastructure in the region that includes an airstrip, okapi pens, patrol posts for rangers, and high-quality tourism facilities. The project has also outfitted 20 health centres with necessities and has overseen the refurbishment and setup of territory offices in Mambasa and Wamba, the construction of a primary school in Epulu, and rehabilitation of medical dispensaries in Sondo and Koki.
• Breeding okapi in captivity for release into their natural habitat to boost non-captive population numbers and genetic diversity. Between 1987 and 2012, the project succeeded in producing 11 calves.
• Promoting environmental education, public engagement, and public awareness. Project staff achieves this by developing and implementing school programs in and around the reserve, presenting seminars in primary and secondary schools, and producing radio broadcasts. They also facilitate focus group meetings with women and farmers, public meetings in villages, work with local committees, and produce outreach materials such as conservation films, calendars, leaflets, and posters. The Okapi Conservation Project has already supplied 112 schools with educational materials.
• Facilitating tourism activities, including visits around the Epulu Station and zoo, forest walks, and participation in traditional hunting.
• Promoting food security in local communities by providing seeds and farming tools to more than 900 farmers in and around the reserve. In addition, more than 142 women have benefited from sewing and embroidery materials, and participation in community farm fields.
• Offering medical care to more than 300 families, a total of more than 1,500 family members, who work at the reserve.
The work of staff at the Okapi Conservation Project and Congo Institute for Conservation of Nature is not easy due to political instability, breakdown of law and order, and lack of financial security. The Reserve and surrounding area also face increasing pressures from mining, poaching, and logging interests. But the project is necessary to help preserve the unique biodiversity of this Global Biodiversity Hotspot. In coming years, we hope that governance of the region will continue to improve and restore peace, justice, and the socio-economic status of local people. In such a situation, local communities and ecosystems, including the okapi, will benefit.
1.4.1 Conservation biology’s ethical principles
Conservation biology rests on a set of underlying ethical principles that is generally agreed upon (Soulé, 1985) and can be summarised as follows:
• The diversity of species and biological communities should be preserved: Most people appreciate biodiversity. Hundreds of millions of people visit national parks, game reserves, zoos, botanical gardens, and aquaria each year. They spend money and take actions to protect these places and species. People also recognize that biodiversity has economic value, whether through tourism, consumption, or other services.
• The untimely extinction of populations and species should be prevented: Throughout history, species have occasionally died off as a result of natural, non-human causes. The loss of a local population was generally temporary, until a new population established itself through dispersal. However, human activities have increased the rate at which species are going extinct by more than a hundredfold (Box 1.3). Meanwhile, there is no similar increase in the rate at which new populations and species are being created.
• Ecological complexity should be maintained: In complex natural environments, biodiversity expresses many of its most valuable features and interactions. Although the biodiversity of species may be partially preserved in captivity, maintaining ecological complexity requires that natural areas be preserved.
• Evolution should continue: Evolution creates new species, increases biodiversity over time, and facilitates adaptation to changing environmental conditions. People can help preserve these evolutionary processes by maintaining genetic diversity in wild populations and allowing populations to exchange genetic material. In captivity, many natural evolutionary processes do not occur, which can hamper survival when species are reintroduced in the wild.
• Biodiversity has intrinsic value: The value of species, communities, and ecosystems does not depend on their utility to people. They are intrinsically valuable on their own, with unique evolutionary histories and ecological roles. There are certain iconic species that people simply want to have around, but other, lesser-known species or species seen as problematic to people are not less valuable.
Box 1.3 Biodiversity: Can Humanity be Saved?
Nkengifor Nkeshia Valery
Regina International Cameroon,
Member of Union Farms of Africa,
Yaoundé, Cameroon.
What happened over the past 200 years that we have arrived where we are? How did we get to this modern paradox? A society where we cherish comfort at the cost of the ever-increasing destruction of our planet. Never in the history of humanity has the environment been degraded to the point that even the air we breathe has become cancerous. Animals are exploited by industries at an alarming rate and those remaining are killed to enrich a privileged few. And all this evil happens with our complicity as indirect consumers. Our inheritance from God, the source of all our nourishment, does not belong to us. Yet it has been bought and exploited by multinational corporations and financial markets that hinder us from cultivating sustainably. We are pushed to feed ourselves and our crops with chemical products that are dangerous to our long-term health. We are also experiencing the start of the sixth mass extinction episode of biodiversity (Ceballos et al., 2017). As a result, the natural world has declared World War III against humanity. This is a war fought not by nation against nation, but that the environment has declared against the whole human race.
This war condemns us to live in an illusion of freedom; we are, in fact, destroyed at an increasing rate by different dangerous diseases and rendered slaves of the polluted environments that we blindly accept. The question we need to ask is not whether we should act to save our planet, but what future and meaning we are going to give the word “HUMANITY”. We are all actors in a civilisation that we are constructing; to quote the Indian leader Mahatma Gandhi: “If we could change ourselves, the tendencies in the world would also change. As a man changes his own nature, so does the attitude of the world change towards him. […] We need not wait to see what others do”. Let us pause and ask ourselves what we want the future to say of us. Are we a destructive generation, or a generation that is ready to sustainably preserve its biodiversity? It is a question every reader needs to ponder. The future is judging no one and blaming no one, but it needs us to change our habits towards protecting the world’s biodiversity.
To change our attitude and make the world a better place, I drafted the following poem with passion to see my words become action for every lover of biodiversity
WORLD CHANGERS
WE ARE A PEOPLE OF PEACE CALLED FORTH OUT OF HUMANITY INTO RESTORING LIFE TO OUR NATURAL HABITAT.WE ARE GOVERNED AND GUIDED BY A SENSE OF SUSTAINABILITY.CONSERVATION AND PROTECTION IS OUR PRIORITY IN ALL THINGS AT ALL TIMES WE ARE LED AND DRIVEN BY THE SPIRIT OF AN ENVIRONMENT FREE OF POLLUTION WE ARE CALLED TO EFFECT AND AFFECT EVERY LIFE THAT WE COME IN CONTACT WITH TOWARDS THE SUSTAINABLE DEVELOPMENT OF THE ENVIRONMENT
WE ARE CALLED BY HUMANITY TO BE WORLD CHANGERS WE REFUSE TO CONFORM WITH THE THINKING PATTERN OF THE WORLD SYSTEM BECAUSE WE ARE WORLD CHANGERS
These principles are not absolute, nor are conservation biologists required to agree with them—they are actively discussed and debated. But many individuals and organizations agree with two, three, or all the principles, and support conservation efforts. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/01%3A_What_is_Conservation_Biology/1.04%3A_Environmental_Ethics.txt |
1. Conservation biology has three goals: (a) to document Earth’s biological diversity; (b) to investigate how humans influence species, evolution, and ecosystem processes; and (c) to investigate approaches to protect and restore biological communities, maintain genetic diversity, and prevent the extinction of species.
2. Because conservation is multidisciplinary and requires a deep understanding of natural processes and human society, conservation biologists must take on multiple roles. Specifically, conservation biologists, as a group, must be curious, good communicators, effective educators, law-abiding citizens, and effective managers and practitioners of conservation projects.
3. Conservation biology relies on scientific evidence and ethical principles that underpins the preservation of biodiversity. Conservation biologists generally agree that biodiversity should be preserved, untimely extinctions should be prevented, ecological complexity should be maintained, evolution should continue, and biodiversity has intrinsic value.
4. Science provides conservation biologists with useful tools that guide the setting of clear, achievable, and measurable goals, to monitor conservation actions to assess whether goals have been met, and to communicate in a clear and unbiased manner.
5. Environmental ethics appeals to people of different walks to life to preserve biodiversity. It holds that biodiversity must also be protected because human well-being and economic opportunities are linked to a healthy environment.
1.06: Topics for Discussion
1. Is conservation biology substantially different in its approach from other fields of science, such as physics, chemistry, or medicine? If so, how? How is it linked to but different from environmentalism and social justice?
2. Looking at the titles of Chapters 5–7, what do you think are the biggest threats to biodiversity near where you live? Explain your answers.
3. There have been three broad conservation approaches throughout history: early models that secured natural resources for powerful people, fortress conservation, and integrated conservation. How do these approaches differ and how do they complement each other? How do you think each of these approaches can contribute to conservation today?
4. Which statements about the ethical principles of conserving biodiversity in this chapter do you agree with? Which do you disagree with? Explain your answers.
1.07: Suggested Readings
Bottrill, M.C., L.N. Joseph, J. Carwardine, et al. 2008. Is conservation triage just smart decision making? Trends in Ecology and Evolution 23: 649–54. https://doi.org/10.1016/j.tree.2008.07.007 Should we let some species go extinct?
Elosegi, A., M.O. Gessner, and R.G. Young. 2017. River doctors: Learning from medicine to improve ecosystem management. Science of the Total Environment 595: 294–302. https://doi.org/10.1016/j.scitotenv.2017.03.188 Comparing conservation to the health sciences
Illingworth, S. 2017. Delivering effective science communication: Advice from a professional science communicator. Seminars in Cell and Developmental Biology 70: 10–16. https://doi.org/10.1016/j.semcdb.2017.04.002 Tips to help you better communicate with public audiences.
Kareiva, P., and M. Marvier. 2012. What is conservation science? BioScience 62: 962–69. https://doi.org/10.1525/bio.2012.62.11.5 A review of an article published in 1985, offering a revised set of core principles that should guide conservation biology today.
Meyer, J.L., P.C. Frumhoff, S.P. Hamburg, et al. 2010. Above the din but in the fray: environmental scientists as effective advocates. Frontiers in Ecology and the Environment 8: 299–305. https://doi.org/10.1890/090143 Scientists have a responsibility to be advocates as well as researchers.
Sutherland, W.J., A.S. Pullin, P.M. Dolman, et al. 2004. The need for evidence-based conservation. Trends in Ecology and Evolution 19: 305–08. https://doi.org/10.1016/j.tree.2004.03.018 How do conservationists make decisions?
Tallis, H., and J. Lubchenco. 2014. Working together: A call for inclusive conservation. Nature 515: 27–28. https://doi.org/10.1038/515027a There is a risk that conservation can be polarising. We should rather be working together.
Tear, T.H., P. Kareiva, P.L. Angermeier, et al. 2005. How much is enough? The recurrent problem of setting measurable objectives in conservation. BioScience 55: 835–49. https://doi.org/10.1641/0006-3568(2005)055[0835:HMIETR]2.0.CO;2 Guidelines for setting goals in conservation. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/01%3A_What_is_Conservation_Biology/1.05%3A_Summary.txt |
Biraben J.-N. 2003. The rising numbers of humankind. Populations and Societies 394: 1–4. https://www.ined.fr/en/publications/population-and-societies/the-rising-numbers-of-humankind-en
Bottrill, M.C., L.N. Joseph, J. Carwardine, et al. 2008. Is conservation triage just smart decision making? Trends in Ecology and Evolution 23: 649–54. https://doi.org/10.1016/j.tree.2008.07.007
Ceballos, G., P.R. Ehrlich, and R. Dirzo. 2017. Biological annihilation via the ongoing sixth mass extinction signalled by vertebrate population losses and declines. Proceedings of the National Academy of Sciences 114: E6089–E6096. https://doi.org/10.1073/pnas.1704949114
Gaffikin, L., and G. Kalema-Zikusoka. 2010. Integrating human and animal health for conservation and development: Findings from a program evaluation in southwest Uganda (Kampala: CTPH; Stanford: EARTH, Inc.; Boston: John Snow, Inc). https://www.jsi.com/JSIInternet/Inc/Common/_download_pub.cfm?id=11196&lid=3
Kalema-Zikusoka, G., R.A. Kock, and E.J. Macfie. 2002. Scabies in free-ranging mountain gorillas (Gorilla beringei beringei) in Bwindi Impenetrable National Park, Uganda. Veterinary Record 150: 12–15. http://doi.org/10.1136/vr.150.1.12
Kareiva, P., and M. Marvier. 2012. What is conservation science? BioScience 62: 962–69. https://doi.org/10.1525/bio.2012.62.11.5
Lackey, R.T. 2007. Science, scientists, and policy advocacy. Conservation Biology 21: 12–17. https://doi.org/10.1111/j.1523-1739.2006.00639.x
Lindenmayer, D., and M. Hunter. 2010. Some guiding concepts for conservation biology. Conservation Biology 24: 1459–68. https://doi.org/10.1111/j.1523-1739.2010.01544.x
Marvier, M. 2014. New conservation is true conservation. Conservation Biology 28: 1–3. https://doi.org/10.1111/cobi.12206
Milfont, T.L., P.G. Bain, Y. Kashima, et al. 2017. On the relation between social dominance orientation and environmentalism: A 25-nation study. Social Psychological and Personality Science 2017: 1948550617722832. https://doi.org/10.1177/1948550617722832
Nelson, M.P., and J.A. Vucetich. 2009. On advocacy by environmental scientists: what, whether, why, and how. Conservation Biology 23: 1090–101. https://doi.org/10.1111/j.1523-1739.2009.01250.x
Ntshotsho, P., H.E. Prozesky, K.J. Esler, et al. 2015. What drives the use of scientific evidence in decision making? The case of the South African Working for Water program. Biological Conservation 184: 136–44. https://doi.org/10.1016/j.biocon.2015.01.021
Pooley, S., J.A. Mendelsohn, and E.J. Milner-Gulland. 2014. Hunting down the chimera of multiple disciplinarity in conservation science. Conservation Biology 28: 22–32. http://doi.org/10.1111/cobi.12183
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Not only is Africa the second most-populous continent in the world, its human population is also incredibly diverse. Consider, for example, that over 2,000 native languages are spoken across the continent (Lewis et al., 2014). (Interestingly, there are strong positive correlations between linguistic diversity and biodiversity, as well as between the loss of species and languages; Gorenflo et al., 2012). Africa is also economically diverse; the continent contains some of the poorest nations in the world but also some of the fastest growing economies (World Bank, 2017). Herein also lies a major challenge: Africa’s diverse human population—already over 1 billion people—is expected to double over the next 25 years (World Bank, 2019). To stimulate economic growth and provide resources for a growing and upwardly mobile human population, once unending wildernesses are constantly being cleared for agriculture, timber, expanding cities, and other human activities. In the process, the remaining natural areas are being polluted, overharvested, and fragmented, particularly in areas of outstanding conservation value (Balmford et al., 2011).
This environmental destruction we are witnessing across Africa holds negative consequences for all people on the continent. Among the most vulnerable are traditional peoples who rely on natural products such as firewood, wild animals, and wild edible fruits and roots to maintain their way of life. The destruction of the environment also makes it more challenging for city dwellers to access basic needs such as clean drinking water, clean air, and wilderness areas where they can fulfil their spiritual and emotional needs. With Africa’s human population and consumption expected to grow substantially for many years to come, there is an urgent need to find ways to ensure that the region’s unique environmental treasures are preserved, for the benefit of current and future generations.
2.01: Sub-Saharan Africas Natural Environment
Much of the African continent encompasses the Afrotropical ecoregion, which is separated from other ecoregions by the Indian Ocean to the East, the Atlantic Ocean to the West, and the Saharan Desert to the North. These major geographic features have acted as barriers to movement since the African continent first took its current shape, enabling species and ecosystems characteristic of the region to evolve in relative isolation from those of other ecoregions. The Afrotropical ecoregion can be further subdivided into eight terrestrial biomes (Figure 2.1), each with its own distinct climate, geology, and biota (Burgess et al., 2004):
• Tropical and subtropical savannahs and grasslands: Sub-Saharan Africa’s largest biome is a mosaic of grasslands, woodlands, bushlands, thickets, and semi-arid drylands that are maintained by fire and grazing. East and Southern Africa’s miombo and mopane savannah-woodland ecosystems are included in this ecosystem.
• Deserts and arid scrublands: A biome of areas where evaporation exceeds precipitation, generally with rainfall < 250 mm/year. Generally associated with searing daytime temperatures and wind-swept sand dunes, this biome contains scrub deserts rich in succulent plants, rocky mountain deserts, and arid grassland-savannah mosaics, such as the Sahel region located just south of the Sahara.
• Tropical moist forests: Lowland broadleaf ecosystems with near-continuous canopies that run as a broad band across equatorial Africa. This biome is characterized by high rainfall ( 2 m/year), low variability in temperatures, and very high species diversity.
• Montane grasslands and scrublands: A patchily distributed biome that occurs at altitudes 800 m and has enough rainfall that a variety of grasses can thrive. Generally lacking trees except along some rivers and streams, it includes high altitude heathlands and other Afro-alpine areas.
• Mediterranean scrub: A scrubland ecosystem of limited extent, better known as the Fynbos or Cape Floristic Region, that is situated at Africa’s southwestern tip. Characterised by hot dry summers and cool moist winters, it contains one of Earth’s richest concentrations of endemic plant species.
• Flooded grasslands and savannahs: Grasslands, marshes, and shallow lakes that are periodically flooded by water that can be fresh, brackish, or hypersaline. When flooded, these areas host some of the largest water bird congregations in the region.
• Tropical dry forests: A highly restricted forest type that can be found in western Zambia and adjacent Angola, as well as on Cabo Verde. While these areas may receive high rainfall, they are characterized by seasonal droughts that can last several months.
• Mangroves: Coastal wetlands of tropical climates characterized by distinctive woody plants with aerial roots that can tolerate saltwater. Typically associated with intertidal zones and muddy bottoms, mangroves provide nursery grounds for many aquatic animal species.
In addition to these terrestrial biomes, Sub-Saharan Africa also contains several aquatic biomes. Prominent freshwater biomes include several large rivers along with their headwaters and deltas, numerous small rivers, multiple large and small lakes, as well as a variety of wetland ecosystems such as swamps, bogs, and salt marshes (WWF/TNC, 2013). Prominent marine biomes include tropical coral reefs along Africa’s east coast, as well as temperate continental shelves and seas along South Africa and Namibia (Spalding et al., 2007). There are also several important oceanic upwellings—areas of high productivity where surface waters are fertilised by nutrient-rich waters that “wells up” from below; these include the tropical Gulf of Guinea upwelling along West Africa, and the Benguela upwelling ecosystem along Africa’s southwest coast.
The variety of biomes present in Sub-Saharan Africa is the result of variable geology and a long history of changes in climate and ecological communities. For example, when Earth’s climate was warmer, tropical moist forests were more widely distributed. As the planet cooled during glacial periods, forests contracted and became fragmented while grasslands expanded; some new biomes developed as the climate changed and species moved around. Even today, biome boundaries are still shifting: for example, over the last few decades the boundary between the Sahara Desert and Sahel has shifted by hundreds of kilometres southward (Foley et al., 2003). The development, fragmentation, and movement of these and other biomes, as well as the influence of major dispersal barriers, such as large rivers and mountain ranges, have stimulated speciation, as different populations became specialised to conditions that were restricted to their particular elevations or on certain sides (wet or dry, sunny or shady) of mountain ranges.
Sub-Saharan Africa boasts tremendous species richness, the result of a complex geological and environmental history.
Due to this dynamic geological, climatic, and environmental history, as well as all the factors that have promoted speciation, Sub-Saharan Africa boasts tremendous species richness. The region is particularly well known for its mammals, particularly its charismatic terrestrial megafauna and other large mammals that attract millions of tourists from all around the world each year (Figure 2.2). Among the most famous are the Big Five animals—lions (Panthera leo, VU), savannah elephants (Loxodonta africana, VU), African buffalo (Syncerus caffer, NT), African leopards (P. pardus, VU), and black rhinoceros (Diceros bicornis, CR). Other notable mammals include cheetahs (Acinonyx jubatus, VU), the fastest mammal on Earth; Maasai giraffes (Giraffa camelopardalis tippelskirchii, VU), the world’s tallest mammal; the giant eland (Tragelaphus derbianus, VU), the world’s largest antelope; and Africa’s four species of great apes. Many small mammals are also noteworthy. For example, East Africa’s naked mole-rat (Heterocephalus glaber, LC) is the world’s only mammalian thermoconformer—meaning it is almost entirely cold-blooded; like reptiles their body temperature tracks ambient temperatures (Buffenstein and Yahav, 1991). The naked mole-rat and Southern Africa’s Damaraland mole-rat (Fukomys damarensis, LC) are the only known eusocial mammals; like some ants and bees, only one female (the queen) reproduces with one to three breeding males, while all the other colony members are sterile workers (Jarvis et al., 1994).
While Africa’s large mammals are a major tourist drawcard, the region hosts many other rich and noteworthy wildlife assemblages. With more than 2,100 bird species, 1,400 of them found nowhere else on Earth (Sinclair and Ryan, 2011), the Afrotropics may be the most taxonomically diverse bird region on Earth (Lotz et al., 2013). Among the many bird species that call Africa home is the world’s largest extant species of bird, the red-necked ostrich (Struthio camelus camelus); standing up to 2.74 m tall, it is in dire need of conservation attention (Miller et al., 2011). Africa is also home to the world’s heaviest extant flying animal, the kori bustard (Ardeotis kori, NT), which can weigh over 20 kg (Dunning, 2008). Over 100,000 insects have been described in Sub-Saharan Africa (Miller and Rogo, 2001), which include the world’s smallest butterfly, the dwarf blue (Oraidium barberae, LC) of Southern Africa, and the aptly named goliath beetles (Goliathus spp.), which can be found throughout much of tropical Africa. The region also hosts a great number of noteworthy endemic amphibians and reptiles, which include the world’s largest frog, the Goliath frog (Conraua goliath, EN) of Cameroon and Equatorial Guinea, and the black mamba (Dendroaspis polylepis, LC), arguably the world’s most feared snake, which is widespread across Africa’s savannahs. Lastly, Africa is home to Jonathan the Aldabra giant tortoise (Aldabrachelys gigantea, VU); having hatched in 1832, he is considered the oldest living terrestrial animal in the world.
Species that have survived previous mass extinction events are unable to withstand the current onslaught of human activities.
The region’s plant richness, estimated at over 45,000 species (Klopper et al., 2007), is also important from a global perspective. Many plant species have high economic value, particularly those that have been domesticated in the region, and are now important crops across the world. Primary among these are coffee—second only to tea in worldwide popularity as a beverage—which is native to West and Central Africa (Coffea robusta) and Ethiopia (Coffea arabica). Other important crops that originated in in the Afrotropics include okra, black-eyed peas, watermelon, and African oil palm. Conserving the wild genetic diversity of these domesticated plants in their native ranges is important because they may serve as “insurance” for today’s crops that may be less productive in future due to anthropogenic climate change (Davis et al., 2012). Others, such as the wide variety of plants utilised in traditional medicine to treat malaria, may one day lead to new antimalarial drugs (Chinsembu, 2015). Similarly, many plant species also have high evolutionary value. These include relict species that survived previous mass extinction events, such as cycads (Encephalartos spp.) (unfortunately several cycad species are now Extinct in the Wild), and Lazarus species that were once believed to be extinct, such as the unique jellyfish tree (Medusagyne oppositifolia, CR) of the Seychelles.
A few small and isolated African ecosystems are particularly rich in species. Particularly noteworthy is the Rift Valley lakes, such as Lake Victoria, Lake Malawi, and Lake Tanganyika, which hold the richest freshwater fish diversity in the world. For example, nearly 14% of the world’s freshwater fish species occur in Lake Malawi (also known as Lake Nyasa). Moreover, over 90% of Lake Malawi’s 500–1,000 (numbers vary by source) fish species (Figure 2.3) are endemic, and thus found nowhere else on Earth. The Cape Floristic Region is home to the greatest concentration of non-tropical endemic species in the world, including speciose well-known plant genera like Protea and Erica. The Succulent Karoo, directly north of the Cape Floristic Region, may be the most floristically rich desert in the world (Mittermeier et al., 2004). Africa has deservedly received international acclaim for these and many other natural wonders. Prominently, more than 37 sites in Sub-Saharan Africa have already been recognized as natural World Heritage Sites. One such site is also Africa’s oldest national park, Virunga National Park in eastern DRC, which contains at least 218 mammal and 706 bird species (WHC, 2007). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/02%3A_Introduction_to_Sub-Saharan_Africa/2.00%3A_Prelude_to_Introduction_to_Sub-Saharan_Africa.txt |
Traditional communities have long held a belief that humans are physically and spiritually connected to nature, and that communal needs outweighed individual desires. This also extended to natural resources, which were considered communal property that must also be shared with the spirits of the ancestors and future generations. Managing natural resources this way required strict adherence to customary law systems that imposed controls on the collection of animal and plant products. Some animals and plants were also worshiped, which leads to mythical superstitions and taboos that prohibited the killing of culturally and spiritually important animals, as well as totem species that bond families and villages together. Customary laws also created Africa’s first protected areas, such as royal hunting grounds (areas where kings and traditional chiefs had exclusive hunting rights) and areas of spiritual significance (Box 2.1), where access and harvesting of natural resources were restricted.
Traditional African communities have long shared the belief that humans are physically and spiritually connected to nature, and that communal needs outweigh individual desires.
Box 2.1 Sacred Spaces: A Tradition of Forest Conservation in Benin
Emile N. Houngbo
School of Agribusiness and Agricultural Policies,
National University of Agriculture,
Cotonou, Republic of Benin.
[email protected]
The importance of forests for human life has been recognized for millennia. That is why public approaches have historically been adopted for their protection. Today, some of the most effective programs are those that integrate local communities and their traditional knowledge with scientific forest management. The Convention on Biological Diversity (CBD) recognizes the value of the cultural practices of traditional peoples for (a) practicing conservation and maintaining biodiversity and (b) promoting sustainable use. The sacred forests of Benin are recognized as a tangible heritage, both natural and cultural; their management by the local community is a major achievement in modern conservation.
Forest protection, an ancient reality in Benin
The life of traditional communities of Benin is closely linked to conservation of its forests, also known as Zoun in the local Goun language. Many social practices of Beninese communities rely on leaves, animals, water, stones, and other resources; the areas that provide these natural resources are called sacred forests because they are inhabited by deities or spirits, serve as spaces for rituals, or represent the seat of past kings. Monitoring of sacred forests is often entrusted to members of a certain lineage. For example, custody of the forest of the city of Abomey is the responsibility of traditional chief Dah Djagba, whose ancestors were installed near the sacred spring Didonou by King Houegbadja of Abomey in the 17th century. A sacred forest is a point of contact between a community and a spirit or deity, and between the visible and the invisible. The value and protection of the sacred forest is passed down from generation to generation, as are the rules and regulations. Typically hunting and setting fires in sacred forests are prohibited, while logging for timber and gathering plants for food and medicine are strictly regulated, with these products shared between priests and caretakers of the site (Juhe-Beaulaton and Roussel, 2002). The Aloe vera plant, for example, has long been used by vodun (spirit) priests during religious ceremonies to heal the wounds of new initiates.
Sacred forests today
Sacred forests have significant spiritual capital, or the power to influence the communities that revere them. They influence the collective consciousness regarding experiences as basic as rain, health, and the collection of spring water (in the case of the Abomey forest), or as complex as religious ceremonies, fertility, and overall happiness. Sacred sites also play an important role in cult practices (Roussel, 1994): funeral rites, ceremonies for dead infants, rites for accidental deaths (Laine, 1990; Sokpon et al., 1998), and healing ceremonies with medicinal plants. Meetings of secret societies such as the Zangbeto, Kuvito, and Oro, and religious or social ceremonies and ordeals are held in sacred forests. They also play an important role in the exercise of justice and social cohesion; disobeying the traditional rules and damaging the sacred forest can cause harm to the whole community (bad harvests, epidemics, drought, and mosquito infestations) or the person responsible (accidents, illness, or misfortune). The wrongdoer may need to perform a rite of reparation, such as an animal sacrifice or offering to repair the damage that they have caused.
Resistance to human pressures
One difficulty of managing sacred forests today is that they are often not well delineated. Under the influence of population growth, the area occupied by a sacred forest sometimes diminishes to a minimum size under communal protection. Some sacred forests in Benin, such as the 32 km2 Birni forest, 11 km2 Tanekas forest, and 2 km2 Natitingou forest, have vanished due to human pressure on the land. The peripheral zone of Gbevozoun sacred forest in which the Gbevo deity is believed to dwell is currently encroached by agriculture (Figure 2.A), and only a central core of 0.5 km2 of the forest’s original 1.6 km2 is still protected. The Honhoue sacred forest, meanwhile, still retains an area of 0.04 km2 that has not shrunk over time. This is due the local community’s belief in the power of the Honhoue divinity and 40 other deities that dwell in the forest.
Sacred forests are based on traditions of safeguarding religious ceremonies and nature for the future, and they continue to be a means of protecting biodiversity. They may be a resource for conservation of rare plant species for medicinal purposes, and even future improvement of agro-biodiversity. The preservation of sacred forests is crucial to community involvement in conservation.
This culturally driven system of checks and balances was greatly disrupted with the arrival of European settlers in the 17th century. Armed with guns, and little thought given to sustainability, the earliest colonists killed thousands of animals for food, trophies, sport, and profit. Following concerns about declining wildlife populations, particularly at the southern tip of South Africa, Sub-Saharan Africa’s first formal environmental legislation was introduced in 1657, followed by the region’s first formal environmental law in 1684 (MacKenzie, 1997). Significantly, this first law separated protected species, such as the common hippopotamus (Hippopotamus amphibious, VU), from pest species (which at the time included lions). Unfortunately, these early laws and regulations were of little consequence as an increasing number of colonists, lured by the promise of unlimited hunting on unexplored lands, arrived in the region. Consequently, by 1700, populations of every animal over 50 kg within 200 km from Cape Town were extirpated (Rebelo, 1992). These developments also led to Africa’s first modern human-caused mammal extinctions. First to disappear was the bluebuck (Hippotragus leucophaeus, EX) around 1798. Nearly a century later, in 1871, the Cape warthog (Phacochoerus aethiopicus aethiopicus, EX)—more closely related to East Africa’s desert warthog (Phacochoerus aethiopicus delamerei LC) than the widespread common warthog (Phacochoerus africanus LC)—disappeared, followed by the quagga (Equus quagga quagga, EX) around 1878 (the last captive individual died in 1883). Elsewhere, bontebok (Damaliscus pygargus pygargus, NT), Cape mountain zebra (Equus zebra zebra, VU), southern white rhinoceros (Ceratotherium simum simum, NT), and black wildebeest (Connochaetes gnou, LC) were all reduced to about a dozen individuals at one or two locations.
Ecosystems—forests in particular—near early European settlements similarly suffered as early colonists perceived them as an “inexhaustible” supply of fuel and timber. This widespread overharvesting prompted the Cape Colony’s Governor in 1778 to appoint its first professional nature conservator, Johann Fredrick Meeding, to exercise some control over deforestation. But, like controls on hunting large mammals, these efforts generally only had a local and temporary impact.
The 1800s and launching of formal conservation efforts
Interest in the formal protection of Africa’s biodiversity started to intensify during the 19th century. Most of the initial steps were taken in South Africa, which had the largest early colonial settlements and, hence, the most species threatened by human activities. First, in 1822, the Game Law Proclamation introduced hunting licence fees and closed seasons for selected species, followed by regulations to protect ‘open spaces’ in 1846 and forests in 1859. A major step towards ecosystem protection was taken in 1876 with the creation of the Cape Colony’s Department of Forests and Plantations, while the appointment of a Superintendent of Woods and Forests in 1881 led to initial efforts towards the scientific management of ecosystems. Then, in 1886, the British government passed the Cape Act for the Preservation of Game (in 1891 extended to other British South African Territories), followed by the Cape Forest Act of 1888. The Cape Forest Act played an instrumental role in the proclamation of the Cape Colony’s first formally protected areas, namely the Tsitsikamma and Knysna Forest Reserves, in 1888; today these lands are incorporated into South Africa’s Garden Route National Park (Figure 2.4). These were followed by the appointment of Southern Africa’s first formal game warden, H. F. van Oordt, in 1893, to manage Pongola Nature Reserve, proclaimed in 1894. (Pongola was degazetted and converted into agriculture land in 1921 but re-established in 1979). Thereafter, protected areas were established at regular intervals across South Africa, starting with Groenkloof Nature Reserve in February 1895, then Hluhluwe Valley and Umfolozi Junction Game Sanctuaries (today the Hluhluwe-iMfolozi Park) in April 1895. (St Lucia Game Reserve, today part of iSimangaliso Wetland Park, was also established sometime in 1895.)
West and Central Africa saw its first steps towards formal conservation efforts in 1885, with the establishment of forest reserves to protect valuable timber products (Brugiere and Kormos, 2009). The region’s first game reserves were gazetted as early as 1889 in the DRC to protect elephants. Unfortunately, these efforts were of little consequence as ivory hunters continued to slaughter the region’s elephant populations. It was only after colonial governments raised concerns about declining ivory revenues that the region passed its first formal environmental law in 1892, with the ratification of the Congo Basin Convention to regulate the ivory trade in French, Portuguese, and Belgian territories (Cioc, 2009).
In East Africa, colonial authorities passed its first formal environmental legislations in 1888. These initial laws called for game reserve establishment, hunting quotas for common species, strict protection for breeding females and immature animals, and hunting bans for rare species (Prendergast and Adams, 2003). While protected area establishment was initially slow, a circular from Lord Salisbury (the UK’s Prime Minister at the time) in which he called for protected areas and hunting restrictions to prevent large mammal extinctions, prompted the passing of the German East African Game Ordinance of 1896. That same year, East Africa saw the proclamation of its first modern protected areas, both in Tanzania: one along the Rufiji river (today included in Selous Game Reserve), and one west of Mount Kilimanjaro.
Initial laws and regulations to protect Africa’s environment were greatly expanded in 1900, with the signing of the Convention on the Preservation of Wild Animals, Birds, and Fish in Africa, during the International Conference of the African Colonial Powers held in London, UK. The most innovative agreement of this treaty was the establishment of Schedules that afforded different species different levels of protection. Species on Schedule 1 included rare and valuable species for which all hunting was prohibited; Schedule 2 and 3 included species for which hunting of young animals and accompanying females was prohibited; Schedule 4 included species for which hunting was allowed ‘in limited numbers’; and Schedule 5 included ‘harmful’ species whose populations needed to be reduced. While this convention never went into force (because not enough parties ratified it), several signatories continued to follow the convention’s agreements by establishing wildlife reserves. Among the first to act were Ghana and Sierra Leone, which took their first formal steps towards conserving the environment in 1901. Soon afterwards, in 1903, Africa’s first conservation non-governmental organization (NGO) was established, namely the Society for the Preservation of Wild Fauna of the Empire (today known as Fauna & Flora International, or FFI).
In 1925, Africa’s first national park, the Albertine Rift’s Albert National Park (today divided into the DRC’s Virunga and Rwanda’s Volcanoes National Parks) was proclaimed. The following year, South Africa’s Sabie Game Reserve (which was originally gazetted in 1898) was renamed and expanded as Kruger National Park. Although most early laws focused on protecting rare and ‘valuable’ mammals, birds, tortoises, and timber forests, the welwitschia (Welwitschia mirabilis) (Figure 2.5) was the first African plant to enjoy formal protection after colonial powers ratified the 1933 Convention Relative to the Preservation of Fauna and Flora in the Natural State (often referred to as the London Convention).
From the outset however, colonial governments managed Africa’s earliest protected areas with policies more representative of Western values, which emphasised the need for nature to be shielded from human activities, and conservation management to be centralised. This top-down, protectionist “fines and fences” strategy, also known as “fortress conservation”, showed little regard for the rights and cultural practices of local communities. In fact, local peoples were more likely seen as a threat to the environment. Consequently, many of Africa’s first formally protected areas were established on land forcibly taken from communal ownership, and access to natural resources on which the local peoples previously relied upon was prohibited. Paradoxically, hunting privileges were reserved for wealthy elites on protected areas set aside for colonists’ enjoyment (Figure 2.6). These practices, termed eco-colonialism for the similarity to the abuses of native rights by colonial powers, caused a growing rift between conservation authorities and deeply offended local peoples.
Conservation efforts after colonialism
Following World War II (1939–1945), after which many African countries regained independence, there was an urgent need for new conservation treaties that also addressed the needs of local peoples. Tanzania’s first president, Julius Nyerere, most vividly expressed this at the 1961 Pan-African Symposium on the Conservation of Nature and Natural Resources in Modern African States (Watterson, 1963), in a speech that became known as the Arusha Manifesto:
The survival of our wildlife is a matter of grave concern to all of us in Africa. These wild creatures amid the wild places they inhabit are not only important as a source of wonder and inspiration, but are an integral part of our natural resources and our future livelihood and well-being. In accepting the trusteeship of our wildlife we solemnly declare that we will do everything in our power to make sure that our children’s grand-children will be able to enjoy this rich and precious inheritance. The conservation of wildlife and wild places calls for specialist knowledge, trained manpower, and money, and we look to other nations to cooperate with us in this important task – the success or failure of which not only affects the continent of Africa but the rest of the world as well.
Soon after the Arusha Manifesto, the African Charter for the Protection and Conservation of Nature was established in 1963. This was followed by the African Convention on the Conservation of Nature and Natural Resources (Algiers Convention in short), which was adopted by member states of the Organisation of African Unity (which preceded the African Union) in 1968. The Algiers Convention provided a major break from colonial conservation models by acknowledging the principle that environmental management is a common responsibility among all Africans, while it also called for conservation of soil and water, and for environmental research and conservation (IUCN, 2004).
Despite the progress and extended scope of the Algiers Convention, conservation policies implemented by early post-colonial governments unfortunately continued to resemble those of colonial governments, notably the centralised and authoritarian style of decision-making. Similarly, the visions of well-funded international conservation organizations operating in the region generally reflected the perceptions and policies of developed nations, and thus lacked adequate consideration of local cultures (Abrams et al., 2009). Consequently, in the years following Africa’s decolonisation, conservation largely remained a polarising endeavour that continued to uproot the lives of tens of millions of conservation refugees over time (Geisler and de Sousa, 2001). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/02%3A_Introduction_to_Sub-Saharan_Africa/2.02%3A_History_of_Conservation_in_Sub-Saharan_Africa.txt |
Building on the environmental laws and protected areas system Africans have inherited from the tumultuous past has not been easy. The scars left in the collective psyche by forced relocations and exclusions have been difficult to mend, with many conservation initiatives still struggling to shake the unfortunate association. Nevertheless, Africa’s passionate conservation biologists and the broader public have shown tremendous fortitude and initiative in advancing the biodiversity conservation agenda over the last few decades. Much of this progress can be attributed to a growing realisation that conserving biodiversity is best achieved when combined with the social and economic upliftment of local people.
Conservation initiatives continue to struggle to shake the unfortunate association from past actions taken with a centralised and authoritarian style of decision-making.
Perhaps the first true step to conservation reform came at the 1975 World Parks Congress hosted in the DRC, when the International Union for Conservation of Nature and Natural Resources (IUCN) adopted its first resolution that recognized the rights and needs of traditional peoples. Over the next few decades, conservation policies of national governments followed, many of which included local people in very explicit terms. One example is Namibia’s Constitution, passed in 1990, stating that:
The State shall actively promote and maintain the welfare of the people by adopting, inter alia, policies aimed at the following: maintenance of ecosystems, essential ecological processes and biological diversity of Namibia and utilisation of living natural resources on a sustainable basis for the benefit of all Namibians, both present and future”.
As the previous centralised and authoritarian style of conservation policy making has made way for more inclusive conservation activities (Abrams et al., 2009), an increasing number of local communities have become active participants in environmental programs and policy development inside and on the periphery of protected areas. Two notable examples are biosphere reserves (Section 13.5.2) and transfrontier conservation areas (TFCA, Box 2.2), both pioneering strategies in promoting human-wildlife coexistence. Several governments are also expanding their protected areas networks by experimenting with private ownership of protected areas (Box 2.3) and co-management partnerships (Section 13.1.4), a land tenure model in which local people share the decision-making and other responsibilities of protected areas management with public institutions (Borrini-Feyerabend et al., 2004). In recent years, integrated conservation and development projects (ICDPs, Section 14.3) have also emerged as viable options to link conservation and socio-economic development.
Box 2.2 Why Go Transfrontier? (And Why Not?)
Tamar Ron
Biodiversity Conservation consultant.
[email protected]
The past two decades have brought high praise and gaining momentum for TFCAs in Southern Africa, as in other parts of the world (e.g. Vasilijevic et al., 2015; Zunckel, 2014). While Africa’s first TFCA, the W National Park, was established already in 1954 by the governments of Benin, Burkina-Faso, and Niger, it was only after the Kgalagadi Transfrontier Park was established in 1999 (between the governments of South Africa and Botswana) that TFCAs have become a prominent component of the concepts driving biodiversity conservation and tourism development in Southern Africa, and across the continent.
TFCAs can support biodiversity conservation in several ways. They help protect large conservation areas and ecological corridors, facilitate cross-border knowledge exchange and cooperation in conservation and enforcement efforts, and promote mainstreaming conservation considerations into land-use planning. These benefits, in turn, offer socio-economic advantages through eco-tourism, sustainable use of natural resources, increased attraction for investors and donors, and, in some cases, supporting peace-building efforts.
Establishing a TFCA, however, entails challenges and risks (Vasilijevic et al., 2015; Zunckel, 2014; Ron, 2007). These processes are often top-down in nature, involving long and costly high-level negotiations between governments with critical conservation funds being spent on multiple cross-border meetings of senior officials and coordination efforts. Due to financial and political considerations, too often the focus remains at the central governments’ level, with limited engagement with local stakeholders and on-the-ground impact. At times, many residents in the concerned area are not even aware that they live in a TFCA, or how this can change their lives.
Political and financial challenges at the local, national and regional levels may hinder the establishment of TFCAs. National inter-agency competition, disagreements within and between local communities, and conflict between international agencies, NGOs, and supporting donors may all have negative consequences. Facilitated cross-border movement of people and goods can cause security challenges and other risks, such as disease transfer, spread of invasive species, increased human-wildlife conflict, and increased illegal wildlife traffic and other criminal activities. In establishing a TFCA, it is thus essential to consult and engage all key stakeholders, and especially local communities, beginning in the planning phase, as well as to prioritise investment in on-the-ground impact-generating activities, to achieve conservation, social and development goals.
My experience in developing the Mayombe Transfrontier Initiative, between Angola, Republic of the Congo, DRC, and Gabon was most revealing (Ron, 2011a). In 2000, we initiated conservation efforts in the Angolan component of the Mayombe forest. From the start, it became clear that the striking difference in the level of degradation between the countries that share the Mayombe forest (Figure 2.B) could not be sustainable. Moreover, uncontrolled logging for timber and poaching of primates, elephants, parrots, pangolins, and other threatened species were driven, to a large extent, by illegal cross-border wildlife traffickers. It was evident that cooperation between the four countries that shared the forest was essential (Ron, 2003), so we solicited financial support from several international organizations. Initial support focused on high-level meetings and negotiations (Ijang et al. 2012). Unfortunately, local stakeholders perceived these mediation attempts as unbalanced. Finally, through governmental leadership, a Memorandum of Understanding was signed between the first three countries in 2009, with Gabon joining in 2013. A study was implemented through extensive consultation with stakeholders, and a strategic plan focusing on the most needed on-the-ground activities was adopted (Ron, 2011b). While conservation efforts have progressed at the national level, the same originally identified threats are still prominent throughout the TFCA, so it is now critical that substantial funding be allocated to the strategy’s actual on-the-ground implementation.
So, what is the conclusion? Go transfrontier? The answer is yes—but not in every case—and very carefully. Perspective must be kept through long term planning, while keeping the focus on local-level priorities.
Box 2.3 Privately Owned Lands for African Conservation
Graeme Cumming
ARC Centre of Excellence in Coral Reef Studies,
James Cook University,
Townsville, Australia.
[email protected]
With rates of species loss increasing and natural communities under pressure worldwide from human demands, the creation and maintenance of protected areas continues to be a vitally important conservation strategy. At the 2014 World Parks Congress in Sydney, Australia, there was widespread recognition of the need to increase the total amount of land and ocean under protection. However, this cannot be achieved by governments simply setting aside more land. Protected areas are ultimately created by people for people, and if they are to be successful, they must be created and managed in a way that is socially acceptable and sustainable.
Committing more land to biodiversity conservation means achieving a consensus between political, economic, societal, and ecological forces. This is particularly important in heavily populated landscapes, especially in Africa where local communities still bear the scars of a history of colonialism and top-down decision-making. One possible solution is to provide incentives that encourage private landowners to engage voluntarily in conservation. The area of land in private nature reserves in South Africa (both individually- and community-owned) is already estimated to be nearly twice the extent of government-owned protected areas (de Vos et al., 2019). The dynamics of privately protected areas and their overall contributions to biodiversity are, however, largely undocumented and poorly understood.
The number of privately protected areas in South Africa has increased rapidly since the end of apartheid in 1994 (de Vos et al., 2019). This increase can be partly attributed to increased tourism in South Africa and partly to the removal of perverse subsidies that kept marginal agricultural land in production (see also Section 4.5.3). Unlike statutory reserves, privately protected areas receive little or no financial support from the government and must ensure their own survival by generating revenue. They can be economically self-sufficient only if they can generate enough income from tourism. Two models appear to be particularly effective: either offering a high-cost, high-investment Big Five game viewing experience (i.e. staying in a comfortable bungalow, being guided by knowledgeable individuals), or providing a cheaper, lower-investment experience that focuses on affordable accommodation with access to hiking trails, striking scenery, and outdoor recreational opportunities (Clements et al., 2016). These models may be particularly effective in areas adjacent to national parks. For example, Shamwari Private Game Reserve, one of the more successful upper-end privately protected areas (Figure 2.C), is adjacent to Addo Elephant National Park in the Eastern Cape.
The conservation value of private lands, and particularly those that stock large herbivores, has been questioned in South Africa because of concerns about economic influences on their management. For example, tourist demand for wildlife viewing experiences can drive the overstocking of large animals, such as elephants, in small Southern African protected areas, even though higher densities of elephants do not necessarily provide a better tourism experience (Maciejewski and Kerley, 2014). Overstocking of large mammals can also lead to the conversion of woodlands to thickets, decreasing both conservation and tourism value (Cumming et al., 1997). Conversely, many private lands in the Western and Eastern Cape of South Africa have high conservation potential; many private lands in the Cape sit lower in the landscape than parks, which, in water-scarce South Africa, have been focused on mountainous water catchment areas, and many harbours threatened lowland vegetation (Winter et al., 2007). Lowland ecosystems with their richer soils are under higher pressure from agriculture and settlement, meaning that well-managed private areas may make a disproportionately large contribution to the conservation of globally rare and endemic fynbos plants and animals (e.g. proteas, heathers, reptiles, and birds). Several governmentally supported programs, such as the stewardship program of the South African National Biodiversity Institute (SANBI), have been created to foster biodiversity conservation on private lands by providing information and encouraging good management practices (Rouget et al., 2014).
The owners and managers of privately protected areas could potentially interact with one another, and with the leadership of provincial and national parks, on a wide range of issues. But the managers of private lands are often poorly connected in these networks and may not benefit from knowledge sharing in the same way as managers of established reserves (Maciejewski and Cumming, 2015). In addition, many privately protected areas are not profitable, with the result that financial demands may push managers to make short-term decisions that attract revenue (e.g. overstocking large herbivores or suppressing wildfires) but have harmful long-term ecological consequences. Possible measures to ensure that private conservation efforts are both sustainable and effective include governmental interventions through tax breaks and support, and improved integration of private lands with national and provincial parks and their managers. Private conservation has considerable promise as a strategy for Africa, but its full potential will only be realised if it is achieved equitably with secure land tenure and supportive governments.
Through these different conservation partnerships models (see also Chapter 13), Africans have surpassed expectations in how rapidly they have expanded their conservation areas network. Illustrating the progress, Cameroon has augmented its existing protected areas system with nine new national parks between 2000 and 2015, with an additional nine in the proposal phase (UNEP-WCMC, 2019). The new parks include Takamanda National Park, which connects with Nigeria’s Cross River National Park to form one of West Africa’s largest continuous formally protected areas; it also plays a critical role in protecting the world’s last remaining Cross River gorillas (Gorilla gorilla diehli, CR), of which fewer than 300 remain. As of mid-2019, protected areas covered over 38% of Tanzania’s land area (more than 361,000 km2, an area larger than Germany or Côte d’Ivoire [UNEP-WCMC, 2019])!
Sub-Saharan Africa’s marine protected areas (MPA, Section 13.4.1) are similarly also expanding. For example, in 2017, Gabon declared 26% of its territorial waters protected, offering a haven to at least 20 species of whales and dolphins, and 20 species of sharks and rays (Parker, 2017). More recently, the Seychelles created two new MPAs that cover an area of 210,000 km2—an area the size of Great Britain. The South African government, in collaboration with World Wide Fund For Nature (WWF), has taken the addition step by creating a forum (http://mpaforum.org.za) to improve MPA governance, and a website (https://www.marineprotectedareas.org.za) to teach the public more about South Africa’s rapidly expanding MPA system.
Many of Africa’s protected areas are nothing more than paper parks, areas that are protected on paper but not in reality.
It is important, however, to keep in mind that protecting a certain area of land and water should not in itself be the only goal in conservation. Even when a country has numerous protected areas, certain unique ecosystems may remain unprotected. Being safeguarded in name is not enough, protected areas must also be maintained and managed to achieve meaningful conservation success. Too many protected areas are nothing more than paper parks, areas that are protected on paper but not in reality. Two of the most important causes of protected area failure are lack of buy-in from local people, and lack of investment, financially or otherwise, from local and national governments (Watson et al., 2014; McClanahan et al., 2016; Gill et al., 2017).
Fortunately, African conservation biologists regularly employ a can-do attitude, shown in a long history of resourcefulness in the face of resource constraints. For example, conservationists from all over the region have established, and are partnering with, non-profit NGOs to facilitate a variety of innovative mechanisms to advance biodiversity conservation (see also Section 15.3). One notable example is the African Parks Network; as of mid-2019, African Parks, in partnership with its host governments, are managing 15 national parks in nine countries, covering 10.5 million hectares. Through this collaboration, which includes extensive community engagement and law enforcement, several once-declining parks are now seeing their wildlife prospering. For example, lions were reintroduced to Rwanda in 2016 after a 20-year absence, elephant strongholds in Chad and the DRC are been secured, and populations of threatened large mammals on Zambia’s Liuwa Plains have increased by 50% to over 100% in just a few years (African Parks, 2016). Not only do recovering wildlife populations here and elsewhere attract more tourists, they also provide opportunities to attract new people to conservation, through environmental education (Figure 2.7), public health services, and other community upliftment programs that improve the well-being of local peoples (see Box 1.2). These benefits then provide additional positive feedback towards wildlife conservation, for example by encouraging an increasing number of poachers to transition into new fulfilling lives as conservation professionals (Cooney et al., 2017).
By seeing and being exposed to all the social and economic benefits biodiversity conservation efforts offer, many local communities have been inspired to take the lead in protecting wildlife on their own lands. For example, community efforts have successfully safeguarded Mali’s savannah elephants (Canney and Ganamé, 2015) and Rwanda’s mountain gorillas (Gorilla beringei beringei, EN) (Kalpers et al., 2003) through periods of conflict. Locally managed forest reserves now protect more than 36,000 km2 of land in Tanzania (Roe et al., 2009), while conservation efforts on community conserved areas in Kenya have renewed hope for the future of the world’s rarest antelope, the hirola (Beatragus hunter, CR) (King et al., 2016). These examples have set a positive, enterprising tone that has enabled conservation to play in increasingly prominent role in multiple economies through the creation of job opportunities while also improving Africans’ overall quality of life. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/02%3A_Introduction_to_Sub-Saharan_Africa/2.03%3A_Conservation_in_Sub-Saharan_Africa_Today.txt |
Despite many examples of progress, conservation challenges and conflicts persist across Africa. As a result, the region lags in several aspects with regards to safeguarding our natural heritage (Table 2.1). The causes are many and vary by region. Below is a discussion of some of the more prominent impediments to effective conservation action in Africa.
Persistent poverty
Poverty can drive desperate people to illegal actions, even though they understand the detriment these actions may have on society at large and their own futures.
There is a direct link between poverty and conservation failure (Oldekop et al., 2016; Hauenstein et al., 2019). This is a problem particularly in Africa, where millions of people live in extreme poverty that is difficult to escape. Faced with hard choices to ensure there is food on the table, poverty can drive desperate people to illegally collect natural products from protected areas, even though they likely understand the detriment these actions may have on society at large and their own futures. Other vulnerable peoples that live close to the land, such as traditional hunter-gatherers and pastoralists, are increasingly pushed into wildlife sanctuaries by mining, deforestation, agricultural expansion, and development that encroach on their traditional lands. Lacking the resources to defend their land and/or support to transition to new lifestyles, these marginalised communities are often left desolate, with few if any legal options to support their livelihoods.
Table 2.1 A comparison between the number of species and number of threatened species for several major groups of animals and plants present in Sub-Saharan Africa.
Group
Species assessed by IUCNa
Species threatened with extinction
Data deficient species
Numberb
Percentage
Vertebrate animals
10,463
1,464
14
1,427
Mammals
1,226
203
17
196
Primates
97
39
40
1
Carnivores
85
13
15
3
Bats
248
23
9
55
Birds
2,265
233
10
18
Birds of preyc
141
32
23
0
Vultures
10
7
70
0
Amphibians
840
213
25
152
Reptiles
736
109
15
123
Ray-finned fishes
5,650
637
11
846
Cichlids
1,026
232
23
146
Arthropods
2,368
637
27
334
Arachnids
186
142
76
2
Insects
1,796
396
22
246
Ants
8
6
100
0
Butterflies
305
72
24
32
Dragonflies
737
70
10
67
Plants
4,916
2,165
44
294
Cycads
68
48
71
0
Ferns
115
47
41
3
Source: IUCN, 2019, current as of April-2019
a Low species richness generally reflects inadequate data because only a few species were evaluated. For example, 100% of ants are listed as threatened, but only eight species have been evaluated; there are more ant species in many African towns and villages.
b Categories included: Extinct in the Wild, Critically Endangered, Endangered, Vulnerable
c Includes raptors, falcons, and owls
Further complicating matters, many well-intentioned citizens and organizations from western countries continue to have overly simplistic views of Africa. By imposing their outsider views on rare species management in Africa, these groups exacerbate the impacts of poverty, by cutting off funding sources of well-functioning conservation programs. A good example comes from regulating trophy hunting of rare animals. Some African mammals, such as lions and elephants, are globally rare, but locally common in well-managed private game reserves and community conserved areas. Due to their global rarity, land managers of such well-managed populations can earn large fees from foreign hunters targeting these sought-after trophy species; the money earned supports local communities by boosting the local economy and conservation efforts (Lindsey et al., 2007; IUCN/PACO, 2009; Cooney et al., 2017). Unfortunately, the hunting of rare species remains controversial because many people dislike seeing charismatic animals killed. Consequently, campaigns from western countries (e.g. Hance, 2018) have significantly impeding the African trophy hunting industry, with no exemption for effective self-supporting land managers. By limiting and threatening the benefits regulated trophy hunting can bring to well-managed conservation areas and poor communities (Mbaiwa, 2018), there is fear that these campaigns will achieve the opposite of their intended purposes, by removing the incentive to protect those rare and/or charismatic species. Conservation requires all parties involved to weigh the benefits as well as unintended consequences of wildlife trade—i.e. overharvesting and black markets (Lenzen et al., 2012; Hsiang and Sekar, 2016), land grabbing (see Section 5.2), corruption, and terrorism (Christy and Stirton, 2015) and adapt as and when needed. Section 14.3 provides some solutions on how to link conservation with development.
Obstructive mindsets
Colonial Africa has provided many examples showing that conservation activities implemented in an authoritarian manner are bound to fail. Yet, authoritarian mindsets continue to impede conservation efforts throughout the region. Work from Guinea-Bissau has shown that authoritarian conservation actions that disempower or displace local communities are more likely to worsen than overcome conservation challenges in post-colonial Africa (Cross, 2015). Conservation in Africa is as much about people as it is about wildlife; this book provides many examples to show how human welfare and conservation are tied to one another.
At the same time, integrating diverging cultural beliefs about the natural environment into conservation practices also remains an obstacle (Dickman et al., 2015). Many Africans continue to fixate on cultural justifications (“We have been hunting for many generations”, Figure 2.8) without acknowledging that human population growth, more sophisticated weapons, and increased levels of consumption are putting unsustainable pressure on natural landscapes. Others believe that the destruction of nature is simply not possible because their ancestors will intervene before this happens, effectively removing individual or community responsibility from conservation management and planning. Breaking down such barriers is hard, frustrating, and takes a long time to achieve. It requires an interdisciplinary approach (Section 1.1) bringing together aspects of conservation science and the social sciences to find common ground. Despite the challenges to putting effective conservation into practice, it is important to remember that fortress conservation models—telling people how they should act, with little to no local input—are more likely to produce enduring counter-productive results.
Weak governance/institutional structures
Africa’s natural environment and its people often fall victim to weak governance and institutional structures. It is well-known that weak policies, failing governments, and civil conflict hamper conservation efforts and drive biodiversity declines (Nackoney et al., 2014; Brito et al., 2018; Daskin and Pringle, 2018). But even in well-functioning countries, government officials turning a blind eye (either willingly, or because they lack capacity) may enable corporations to cut corners for increased profits at the cost of the environment. Corruption and greed also fuel land grabbing (Section 5.2), black markets (Hauenstein et al., 2019), and unwarranted protected area degazettement (Section 13.7.3). There is broad interest to challenge these behaviors which benefit only a handful of people at the cost of thousands of others (Box 2.4). Fixing these issues will rely on strengthening institutional capacity on multiple levels (Amano et al., 2018).
Weak policies, failing governments, and civil conflict hamper conservation efforts and drive biodiversity declines.
Box 2.4 Malawi: No Longer a Weak Link in the Elephant Ivory Trafficking Chain?
Jonathan Vaughan
Lilongwe Wildlife Trust,
Lilongwe, Malawi.
http://www.lilongwewildlife.org
International efforts to combat illegal wildlife trade—now the fourth largest transnational crime in the world (Nellemann et al., 2016)—have intensified in recent years, but Malawi has escaped public scrutiny due to its small size and relatively small wildlife numbers. Despite these factors, Malawi’s wildlife populations have been decimated by poaching in the last few decades. For example, Kasungu National Park’s wildlife was so abundant in the 1980s that animals were translocated to the Kruger National Park in South Africa. Back then, elephants numbered as high as 2,000. Today, there are no more than 60.
Southern Africa’s principle transit hub for wildlife trafficking
In 2016, CITES identified Malawi as a “country of primary concern”, and Southern Africa’s principle transit hub for ivory trafficking. Malawi’s own Illegal Wildlife Trade Review (Waterland et al., 2015), published a year earlier, had come to similar conclusions, uncovering evidence of large-scale international trafficking of bushmeat, carnivore pelts, tortoises, pangolins, orchids, ivory, and rhino horn. The revelations served a wake-up call for urgent action to protect not just Malawi’s own wildlife but also wild populations throughout Southern Africa.
Central to region’s poaching hotspots
Why is Malawi such a significant link in the trafficking chain? The first clue is geography. Malawi is surrounded by Africa’s biggest elephant poaching hotspots. Selous Game Reserve in Tanzania reportedly lost 25,000 elephants between 2009 and 2013, while 1,000 elephants were killed in Mozambique’s Niassa Province in 2011, alone (Booth and Dunham, 2016). Poaching in Zambia’s Luangwa Valley is well above the CITES average (Nyirenda et al., 2015). Wasser et al. (2015) found that all the study samples of ivory seized from consignments weighing more than half a tonne between 2006 and 2014 originated from ecosystems immediately bordering Malawi.
Malawi has already been implicated in some of the largest ivory seizures in the world. The biggest impoundment ever—at 7.5 tonnes, equivalent to over 1,500 elephants—was made in Singapore in 2002 and had been shipped from Malawi’s capital, Lilongwe (Wasser et al., 2007, 2015). In 2013, 2.6 tonnes of ivory were confiscated from a container within Malawi’s borders at Mzuzu. Fifty cases were recorded between 2010 and 2014, including numerous smaller examples of ivory trafficking. With an estimated 10% interception rate, the true scale of ivory trafficking was evidently much larger than previously thought (Waterland et al., 2015).
Risk-reward ratio in favour of criminals
Malawi’s weak wildlife legislation was another significant factor. Coupled with under-resourced law enforcement and high levels of corruption, this offered an attractive risk-to-reward ratio for wildlife criminals. The individuals convicted of trafficking in the 2013 Mzuzu case escaped with a fine of just US \$5,000 for a 2.6-tonne haul. This paled in comparison to the penalties handed out in other countries. For example, during the same period, a Zambian man was sent to prison for five years for trafficking 12.5 kg of ivory, a South African man received 10 years and a US \$392,000 fine for trafficking one tonne of ivory and, in Kenya, a man was fined US \$233,000 for trafficking a single tusk weighing 3.4 kg.
While sentencing in the Mzuzu case was hampered to some extent by limitations of the law, it was also indicative of the fact that, historically, trafficking was not treated as a serious crime in Malawi. Most wildlife prosecutions had taken place in lower courts and have been prosecuted by lower-ranked officials. The average fine for ivory trafficking was found to be just US \$40 between 2011 and 2014. This is an extremely low amount given the potential profits from the trade of ivory and, thus, provides virtually no deterrent to traffickers. Awareness, motivation, and cooperation within and between departments like the police and border forces were found to be severely lacking. Government resources to combat wildlife crime are also limited, with many other causes competing for funding and attention.
Management of wildlife crime data also made life easier for wildlife criminals. Take the case of a Chinese national who was arrested and prosecuted for an ivory trafficking offence under one name, deported under a second name, and reported by the INTERPOL country office to INTERPOL headquarters under a third. This shows the ease with which criminals are circumventing the weak systems currently in place.
Turning the Tide
Today, however, things are changing. Recommendations from the 2015 Illegal Wildlife Trade Review were swiftly executed, strengthening the process from investigations and arrest right through to prosecution and sentencing. As a result, in just four years, over 1.5 tonnes of ivory were confiscated, average monthly arrests for wildlife crime jumped from 0.7 to 9.5, and custodial sentence rates rose to over 90%, with judgments passed of up to 18 years. Remember that, in comparison, no-one convicted of a wildlife crime between 2010 and 2015 had been put behind bars and the average fine was just \$40.
Other initiatives included improving protected area management, launching the country’s first wildlife crime investigations unit, and establishing an Inter-Agency Committee to Combat Wildlife Crime to improve cooperation and information sharing. Critical amendments to wildlife legislation were also passed in record time, and technical expertise from partners was harnessed to maximise impact. Lilongwe Wildlife Trust is currently the only NGO sanctioned to prosecute wildlife crime cases in partnership with an African government. In short, Malawi has strengthened each stage of the enforcement chain.
These successes came about largely as a result of a collaborative, innovative, and holistic approach that moved beyond traditional wildlife conservation to incorporate practices used in combatting serious organised crime.
Support from the very top
Strengthened legislation and enforcement will continue to be a critical deterrent, but advocacy has also been a critical tool for securing high-level political will and turning it into action. The President of Malawi, himself, His Excellency Peter Mutharika, backed the nation’s “Stop Wildlife Crime” campaign (Figure 2.D) and the Malawi Parliamentary Conservation Caucus continues to raise awareness through the media, essentially holding stakeholders such as the police or judiciary to account by highlighting both successes and questionable outcomes.
Focus on trafficking
Poaching has been a major focus of conservation efforts elsewhere in Africa, and local poachers can still expect to feel the full weight of the Malawian law. However, bringing traffickers to justice is proving a more effective use of limited resources. After all, it is members of organised international crime syndicates that ultimately exploit local communities, incite corruption, threaten our national security, and provide the routes to overseas markets.
What’s next for Malawi?
Sustaining Malawi’s astounding turnaround will be no easy feat. But with continued determination, as well as local and international cooperation and support, we believe that these criminal networks can be disrupted enough to halt the impending extinction of one of Africa’s most iconic species.
The same tenacity and high-level commitment we have witnessed in the last five years must now be applied to other conservation challenges, as attention is being turned to the protection of Malawi’s wider biodiversity. In 2018, a further 216 species of animals, plants and trees were placed under legal protection, and lessons from combatting wildlife crime can now be applied to other illegal or unsustainable practices, such as trades in timber, charcoal, and fish.
When it comes to pioneering conservation, Malawi is one to watch. Let’s hope that there are more achievements to celebrate in another five years’ time.
Skills shortages
Scientific advances depend on increased or updated knowledge. That is also true for conservation biology—effective conservation depends on local experts who can design and implement monitoring and research projects, apply adaptive management (Section 10.2.3) when needed, act as managers and advocates for conservation activities, and increase awareness of the importance of the environment (Laurance, 2013). It is thus of great concern that conservation in Africa continues to face an enduring skills shortage (Wilson et al., 2016). Illustrating the problem, a recent review found that, over the past three decades, only 129 of the scientific articles focussed on West African birds were produced in international journals by local authors. This productivity contrasts strongly with Europe, where 12,380 ornithological articles were produced over the same time (Cresswell, 2018). Another review, covering all of Africa, found that less than 30% of the continent’s birds received attention in international journals (Beale, 2018). While high-impact publications are not the only metric to estimate conservation success, they provide an accurate accounting of persistent knowledge gaps, as well as skills shortages further down the hierarchy, from researchers and teachers to rangers and other fieldworkers down to citizen scientists.
There are many reasons for these skills shortages. Some of the most prominent foundational issues include a fragmented communication network that limits skills transfer, financial and other resources limitations, a shortage of quality training institutes, and overburdened teachers at existing educational facilities. Fortunately, many of these shortfalls are currently being addressed. For example, new people are being involved in conservation activities through citizen science projects (see Box 15.3), innovative funding mechanisms are being developed (Section 15.3), legal and organizational structures are being adapted to foster increased collaboration (Section 15.4) and freely-accessible resources such as this textbook are being made available.. It is important to continue to build on this progress by supporting such initiatives, and continuously highlighting to others the importance of nature to their own well-being.
2.4.5 Competing interests
Like stock market investments, the benefits to be gained from conservation may take years to materialise.
Because of competing interests (for land, natural resources, etc.), there is always a risk that a wealthy business will threaten a conservation initiative with competing offers that typically include promises of jobs and development (Koohafkan et al., 2011). Local peoples, especially those in poverty, may find it hard to turn down such attractive counteroffers, even if they recognize that those offers rarely live up to the promises made. Conservation biologists should carefully consider what such offers on the table might look like and factor in how their conservation programs compete and bring better results for all.
People concerned with the environment have worked hard to better highlight that conservation has the potential to be profitable and to spur sustainable development. These activities have seen the emergence of fields such as environmental economics, and methods to put a market value on ecosystem services (Section 4.5). Unfortunately, some conservation biologists have fallen into a trap of (over)emphasising the economic benefits that conservation can bring, without a realistic representation of the upfront investment required or the length of time required for a tangible return on investment. Like stock market investments, the benefits to be gained from conservation may take years to materialise, sometimes with very little to show for it in the meantime. Given that all investments require either expendable capital or credit, willing stakeholders with neither are essentially being asked to maintain a more restrictive livelihood over an unsustainable (and often undisclosed or unknown) period of time. It is crucial for conservation biologists to set realistic expectations and to offer a balanced approach that provides interim funding/credit options. Such options could perhaps include microloans, village savings and loan associations (http://www.care.org/vsla), or community conservation banks (https://sema.fzs.org/en/conservation-banks) such as those established by Frankfurt Zoological Society (FZS) in Tanzania. It is also important to incorporate benefits beyond immediate financial gain when starting or expanding conservation programs. Conservation actions should also aim to provide concrete benefits, whether financial or otherwise, to local communities from an early stage. In that way, if a project comes to a premature end, one can still point to the progress made, which will make it easier to engage with that community when future opportunities arise. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/02%3A_Introduction_to_Sub-Saharan_Africa/2.04%3A_Ongoing_Conservation_Challenges.txt |
Because of the many challenges that conservation projects continue to face, the list of Sub-Saharan African species and ecosystems that are threatened with extinction and destruction continues to grow every year. In a recent assessment, BirdLife International identified 51 Important Bird and Biodiversity Areas (IBA) in Sub-Saharan Africa—many of them national parks—in danger of ecosystem degradation (BirdLife International, 2019). A United Nations (UN) assessment similarly found that the outlook of 12 natural World Heritage Sites situated in Sub-Saharan African “in danger” (http://whc.unesco.org/en/danger). These are substantial and challenging problems that will keep conservation biologists very busy in the future. These problems need to be faced head-on to ensure that future generations will also be able to enjoy the natural treasures and resources the region has to offer.
2.06: Summary
1. Sub-Saharan Africa supports extremely diverse ecological communities across its eight terrestrial biomes (which include forests, savannahs, woodlands, grasslands, scrublands, deserts, and mangroves) as well as multiple freshwater and marine ecosystems. The region’s complex climate, geology, and history have contributed to the development of its exceptional biodiversity.
2. Conservation in Africa has gone through major changes over the past few centuries including traditional relationships with nature; exploitation of wildlife and natural resources by European settlers in the 17th and 18th centuries; western practices of setting aside land shielded from human influences; and more recently integrated conservation and development practices.
3. Africa’s conservation biologists and the broader public have shown tremendous fortitude and initiative to overcome the various challenges facing biodiversity over the last few decades. This includes greatly expanding the protected areas network, passing laws protecting the environment, and establishing productive partnerships.
4. By reaping the benefits from conservation activities in and around protected areas, many private individuals and local communities have been inspired to take the lead in protecting biodiversity on their own lands.
5. Historical legacies, poverty, greed, weak governance, consumptive needs by an increasing human population, and competing interests remain challenges to conservation in Africa. Many of these challenges lead to threats to the future persistence of many species and ecosystems, including environmental degradation and overharvesting.
2.07: Topics for Discussion
1. The human population of Sub-Saharan Africa is predicted to increase dramatically in coming decades. How do you think this growth will affect the region’s biodiversity? Do you think that the increase in human population will increase consumptive needs?
2. What are the main international and national organizations contributing to conservation in your region? What projects are they working on? What are the most important goals of those projects? What do you think are the biggest challenges facing those projects?
3. Conservation in Africa has gone through several stages through history. Can you summarise each of these stages in two or three sentences? What do you think are the strengths and weaknesses of each stage?
2.08: Suggested Readings
Abrams, R.W., E.D. Anwana, A. Ormsby, et al. 2009. Integrating top-down with bottom-up conservation policy in Africa. Conservation Biology 23: 799–804. https://doi.org/10.1111/j.1523-1739.2009.01285.x Africa requires locally-adapted conservation policies.
African Parks. 2019. Unlocking the value of protected areas. African Parks Annual Report 2018 (Johannesburg: African Parks). https://www.africanparks.org/unlocking-value-protected-areas An overview of activities undertaken by a successful conservation NGO.
Balmford, A., J.L. Moore, T. Brooks, et al. 2011. Conservation conflicts across Africa. Science 291: 2616–19. https://doi.org/10.1126/science.291.5513.2616 Many of the challenges to conservation in Africa are rooted in population growth.
Beale, C.M., S. van Rensberg, W.J. Bond, et al. 2013. Ten lessons for the conservation of African savannah ecosystems. Biological Conservation 167: 224–32. https://doi.org/10.1016/j.biocon.2013.08.025 Learning from past efforts to guide future actions.
Cooney, R., D. Roe, H. Dublin, et al. 2017. From poachers to protectors: Engaging local communities in solutions to illegal wildlife trade. Conservation Letters 10: 367–74. https://doi.org/10.1111/conl.12294 Guidelines for engaging the local community in conservation activities.
Cross, H. 2015. Displacement, disempowerment and corruption: Challenges at the interface of fisheries, management and conservation in the Bijagós Archipelago, Guinea-Bissau. Oryx 50: 693–701. https://doi.org/10.1017/S003060531500040X Authoritative top-down strategies will likely worsen rather than solve conservation conflicts.
Hauenstein, S., M. Kshatriya, J. Blanc, et al. 2019. African elephant poaching rates correlate with local poverty, national corruption and global ivory price. Nature Communications 10: 2242. https://doi.org/10.1038/s41467-019-09993-2 Linking poverty, corruption, and wildlife declines.
Kalpers, J., E.A. Williamson, M.M. Robbins, et al. 2003. Gorillas in the crossfire: Population dynamics of the Virunga mountain gorillas over the past three decades. Oryx 37: 326–37. https://doi.org/10.1017/S0030605303000589 African conservationists have had to deal with a very dynamic, and sometimes dangerous environment.
Prendergast, D.K., and W.M. Adams. 2003. Colonial wildlife conservation and the origins of the Society for the Preservation of the Wild Fauna of the Empire (1903–1914). Oryx 37: 251–60. https://doi.org/10.1017/S0030605303000425 Consider how conservation biologists today face similar challenges to those a century ago. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/02%3A_Introduction_to_Sub-Saharan_Africa/2.05%3A_Conclusion.txt |
Abrams, R.W., E.D. Anwana, A. Ormsby, et al. 2009. Integrating top-down with bottom-up conservation policy in Africa. Conservation Biology 23: 799–804. https://doi.org/10.1111/j.1523-1739.2009.01285.x
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Conservation biology aims to improve the protection of biodiversity—that is, all the species, genetic diversity, and ecosystems on Earth. By this definition, the process of documenting life on Earth requires us to consider biodiversity on three different levels (Figure 3.1):
• Species diversity: The full variety of species, from single-celled organisms like bacteria to larger multicellular organisms like animals and everything in between.
• Genetic diversity: The full range of variability in genetic material within a species. This variation can occur spatially as differences between populations or as differences between individuals of the same population.
• Ecosystem diversity: The full variety of ecosystems—i.e., assemblages of species and the physical environments in which they live.
The relationship between species, genetic, and ecosystem diversities is complex and interdependent. That is, a species cannot exist without genetic diversity or ecosystem diversity, and vice versa. For that reason, it is virtually impossible to affect one aspect of diversity without affecting the other. We can therefore think of species, genetic, and ecosystem diversities simply as different ways to measure the variety of life.
3.01: Species Diversity
In general, the first step in responding to the conservation need of a species or population is to know its identity. For this reason, one of the three main goals of conservation biology is to document all life on Earth or, in plain language, to give each species a name. The task of giving each species a (formal) name falls on specialist scientists known as taxonomists. Taxonomists (and the people assisting them) explore nature, collect specimens of plants, animals, and other organisms, describe/name those specimens, and store the specimens in permanent collections, such as natural history museums and herbaria (there are currently over 6,500 natural history museums in the world). These permanent collections, affectionally called “Libraries of Life”, provide the material and locations that taxonomists use to describe species and to develop systems for biodiversity classifications.
When a species is formally described, it is given a unique two-part name, known as a binomial name. For example, the binomial name for the lion is Panthera leo. The first part of the name, Panthera, identifies the generic epithet (or simply genus); in this case, the panthers or big cats. The second part of the name, leo, identifies a subset within the genus known as the specific epithet (or simply species); in this case, the lion. This binomial system thereby both identifies a lion as its own species and connects it to other closely-related species: Africa and Asia’s leopards (P. pardus, VU); Asia’s snow leopard (P. uncia, VU); Asia’s tigers (P. tigris, EN); and South America’s jaguars (P. onca, NT) (Figure 3.2).
Binomial species names, as well as the taxonomic relationships between different species, form the backbone of taxonomic databases, as compiled and organised by biodiversity informatics projects. Some biodiversity informatics projects focus on one group of species, while others focus on certain regions. For example, all known marine species are listed in the World Register of Marine Species (http://www.marinespecies.org), while the Catalogue of Afrotropical Bees (https://doi.org/10.15468/u9ezbh) collates information of only African bees. In some cases, multiple projects—each using different assumptions to suit different user groups better—may catalogue the same group of species. For example, the world’s fungi, are listed both in Index Fungorum (http://www.indexfungorum.org) and MycoBank (http://www.mycobank.org), while bird names are indexed by at least seven different projects, each a little different from the other. There are even some biodiversity informatics projects that attempt to catalogue all life on Earth; examples include Catalogue of Life (http://www.catalogueoflife.org), Encyclopaedia of Life (http://eol.org), and Wikispecies (https://species.wikimedia.org).
What is a species?
There are three rules of thumb that taxonomists use to describe a species:
• Morphological definition of species: Individuals that are distinct from other groups in their morphology, physiology, or biochemistry.
• Biological definition of species: Individuals that breed (or could breed) with each other in the wild, but do not breed with members of other groups.
• Evolutionary definition of a species: Individuals that share a common evolutionary past, usually indicated by genetic similarities.
In practice, conservation biologists generally rely on the morphological definition to identify species. The ability to recognize physical or morphological differences between organisms is handy even when the actual identity of specimens is unknown. In such cases, field biologists may refer to the unknown species as morphospecies (Figure 3.3), at least until an expert identifies the unknown individuals or a taxonomist gives them an official scientific name. In contrast, the biological definition of species relies on information that is difficult to obtain and thus not readily available. The biological definition also fails to recognize recent speciation, which can cause closely related but distinct species to interbreed. Similarly, it is generally impractical for fieldworkers to measure differences in genetic sequences to distinguish one species from another because these procedures currently require expensive, immovable laboratory equipment.
Taxonomists can use morphological, biological, and genetic information to identify species.
Despite the practical difficulties of applying the biological and evolutionary definitions in the field, both provide important guidelines for conservation efforts. The biological species definition allows us to better understand species biogeography and the mechanisms that prevent two closely-related species to interbreed. The evolutionary species definition in turn allows us to better understand how and why the genetic makeup of populations change over time, through processes such as random mutations, natural selection, emigration, and immigration. It is thus important for conservation biologists to acknowledge the importance of maintaining these dynamic processes in protecting natural systems, and where possible, include them in their fieldwork (Box 3.1).
Box 3.1 Finding a Needle in a Haystack: Monitoring Species Using eDNA
Tammy Robinson and Clova Mabin
Centre for Invasion Biology, Stellenbosch University,
Stellenbosch, South Africa.
Trying to find threatened species in aquatic systems can be like trying to find a needle in a haystack. Traditionally, researchers have set off with nets, buckets, and even snorkels and scuba gear to painstakingly search for threatened species in ecosystems, ranging from streams to coral reefs. While searching in a small system, such as a pond, might not seem too difficult, it can be a real challenge to find tiny, inconspicuous organisms in and amongst the mud, stones, and plants, especially when they are trying their best to remain hidden. Things get even trickier when combing through large ecosystems like lakes or bays. These difficulties make it hard to reliably monitor the status or distribution of threatened aquatic species.
However, scientists have recently developed a new search tool called environmental DNA, (eDNA in short), where researchers collect and search water samples for the DNA of the species they are interested in. The eDNA technique was first developed by a biologist trying to detect organisms in sediment (Willerslev et al., 2003) but is now being used by conservationists working in all kinds of aquatic ecosystems. Organisms continually release small amounts of DNA into the water by sloughing off skin or other cells and releasing bodily wastes. This DNA then mixes in the surrounding environment, allowing those organisms to be detected through genetic analyses without actually sampling them directly.
Researchers have been testing just how useful eDNA is for finding threatened species in ponds and streams (Thomsen et al., 2012). They detected the eDNA of fish, shrimp, dragonflies, and amphibians in most ponds where the species were known to occur and found no trace of the eDNA of these species where they were absent. The most exciting development was their ability to detect eDNA evidence of threatened species in places where they had previously occurred but not been recently recorded by traditional search methods. Field observations and experiments also showed that eDNA can persist for up to two weeks in fresh water, and that concentrations can correspond to population sizes; this suggests that scientists may be able to monitor the abundance of rare aquatic species to a high degree of accuracy using this approach. For example, Lake Victoria could be searched for rare cichlid fish species that may still be present at low numbers even though researchers have not seen them for several years.
eDNA technology also holds considerable promise for the management of aquatic invasive species, if they could be detected as new arrivals before their numbers grow enough to be detected by conventional methods (Takahara et al., 2013). Early detection will give conservation managers a head-start and enable them to react quickly to invasions and increase their chances of preventing the environmental damage associated with invasive species. In a local twist to the tale, ongoing work in South Africa is applying eDNA as a tool for measuring the success of management efforts aimed at removing the invasive marine European shore crab (Carcinus maenas) (Figure 3.A) that could outcompete or threaten native African marine species. It is hoped that eDNA will be able to track the decline in crab numbers as the species is removed and then be used to monitor for any new arrivals should the crabs re-invade.
This exciting new approach in detecting species is rapidly developing and improving our efficiency at monitoring threatened and invasive species. This makes the process less like looking for a needle in a haystack, and more like finding the millions of needles right under your nose. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/03%3A_What_is_Biodiversity/3.00%3A_Prelude_to_What_is_Biodiversity.txt |
Every extant species on Earth consists of at least one population, a group of individuals at a certain place that generally look alike and can potentially breed with each other to produce offspring. These population(s) can be very small (just a few individuals), very large (billions of individuals), or anything in-between. The individuals within each population generally differ genetically from one another to some degree. This genetic variation, a component of genetic diversity (Figure 3.4), exists because the genes—the functional units of hereditary information that provide the blueprint of an organism—in different individuals are made up of slightly different DNA sequences. Different forms of a gene, which arise through mutations that change DNA sequences, are known as alleles. The gene pool, in turn, consists of the total diversity of genes and alleles in a population or species. The particular mix of genes and alleles in an individual is its genotype. The expression of an individual’s genotype, as determined by the environment where an organism has developed, is its phenotype—that is, the organism’s morphology, anatomy, physiology, and biochemistry. Common characteristics to describe a person include height, hair colour, and blood type, which taken together begin to describe that person’s phenotype.
In species which reproduce asexually, the potential for increased genetic diversity is limited to DNA mutations. However, sexual reproduction creates new genetic combinations by bringing together chromosomes from each parent. This process, called recombination, results in offspring that are genetically unique from their parents. Genetic mutations provide the foundation of genetic variation, but sexual reproduction dramatically increases genetic diversity by randomly mixing alleles in different combinations.
Genetic diversity enables species to adapt to environmental change.
Two factors determine a species’ genetic diversity: the number of genes that have multiple alleles (polymorphic genes) and the number of alleles present in a population for each polymorphic gene. If a gene is polymorphic, some individuals will have two different forms of the gene—that is, they will be heterozygous because they received different alleles of the same gene from their parents. Some individuals will have two of the same forms of the gene—they will be homozygous because each parent gave them the same allele. In general, the greater the genetic diversity in a population, especially the greater number of alleles present, the more capable a species will be to adapt to changing circumstances in their environment. Genetics also affect an individual organism’s development, physiology, and fitness—the relative ability of individuals to survive and reproduce. This same principle gives humans the ability to select and breed crops and domestic animals with characteristics that benefit the production and quality of food (Davis et al., 2012). Many rare species have relatively low genetic diversity, especially in populations which have dwindled to small sizes. Low genetic diversity limits small populations’ ability to adapt to changes in environmental conditions and leaves them at risk of extinction when conditions do change. Section 8.7 discusses the importance of maintaining genetic diversity in greater detail.
3.03: Ecosystem Diversity
Those who have climbed Africa’s highest mountains have likely noticed how the plants and animals present gradually change, as one moves from tall lowland forest to moist, mid-elevation forest with a low canopy, then into grassy alpine meadows, and lastly, onto cold, windy, and rocky mountain peaks. We see these changes because, as we move across the landscape, physical conditions (e.g. geology, soil type, temperature, precipitation) change, and so also the species adapted to different environmental niches, as determined by the varying conditions. Thus, one by one, the species present at one location are replaced by new species better suited to the new conditions. We can see how the whole landscape changes in response to dynamic biotic and abiotic components of the environment (Figure 3.5). The variety of life resulting from these environmental changes is what gives rise to ecosystem diversity.
Ecosystem diversity describes the full variety of ecosystems of an area, while the term “ecosystem” describes all the organisms in an area, as well as the physical and chemical environment with which those organisms interact. An important component of any ecosystem is its biological community (or ecological community), defined as all the living individuals, populations, and species of a place, as well as all the biological interactions among those organisms. The abiotic (or physical) environment, especially climate, energy, and nutrients availability, greatly affects the structure, composition, and characteristics of an area’s biological community (or biotic environment), and ultimately the type of ecosystem present (Figure 3.6). For example, water that evaporates from leaves, the ground, and other surfaces may later become rain or snow that provides drinking water that sustains life. Sunlight energy, in turn, enables photosynthetic plants (or primary producers) to grow; the energy from the sun is later transferred to animals that eat the plants (herbivores, or primary consumers), and then to animals that eat other animals (carnivores, or secondary consumers). The physical environment similarly affects aquatic ecosystems. For example, in freshwater stream, the biological community present is determined in large part by the physical characteristics of the stream, including water chemistry, temperature, flow rate, and substrate.
At local scales, biological communities themselves can play prominent roles in altering the physical environment. For example, the trees present in a forest ecosystem can influence wind speed, light, humidity, soil chemistry, and temperature. Likewise, marine biological communities, such as kelp forests, seagrass beds, and coral reefs, can affect water temperature, water chemistry, sunlight penetration, and wave energy.
Within a biological community, individual species have specific ecological roles and have different requirements for survival. These roles and requirements enable different species to coexist, and in cases of interdependency, necessitate that they do so. For example, a given plant species may grow only in one type of soil, be pollinated by one type of insect, or have its seeds dispersed by only one type of animal. If any one of these requirements restricts the population size or distribution of that plant, it is considered a limiting resource. Even animal dung, usually considered a waste product, may become a limiting resource to species that rely on it for feeding and breeding. For example, studies from Côte d’Ivoire and Southern Africa have linked dung beetle population declines to the extirpation of large herbivores such as elephants and buffaloes (Nichols et al., 2009).
Environmental conditions that regulate the abundance of limiting resources may change over time. Consequently, many ecological communities can undergo major shifts in their composition over time. This is particularly prominent during ecological succession, which describes the gradual process during which ecosystems change after a disturbance. Consider, for example, an old-growth forest that is cleared by a logging operation. Shortly after clearing and abandonment, the soil absorbs more sunlight, resulting in high temperatures and low humidity during the day. These early stages present an ideal environment for pioneer species, such as sun-loving butterflies, annual herbs, and grasses, with wind-dispersed seeds. In a few years’ time, the early successional herb-field or grassland transition to a scrubland, which accommodates a new suite of species. As the shrubs mature, forest trees germinate in the shade provided by the shrubs. Over the course of decades, as the forest trees mature, the forest canopy is gradually re-established which, in turn, provide opportunities for species characteristic of mid- and late-successional stages, such as shade-tolerant wildflowers of moist soils. Eventually, after many decades, climax species representative of mature forests, such as birds that nest in the holes of dead trees, start colonising the area. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/03%3A_What_is_Biodiversity/3.02%3A_Genetic_Diversity.txt |
Developing a strategy to conserve biodiversity requires a firm understanding of where threatened species and populations occur, why they are threatened, what their needs are, and what role they play in their respective ecosystems. By obtaining an understanding of species’ distributions, biologists simultaneously gain an initial rough “estimate” of genetic diversity and ecosystem diversity. While addressing these questions is a critical task, finding appropriate answers can be complex, expensive, and take a long time to solve. This is in no small part because identifying species can, at times, be a very challenging endeavour.
Challenging species identifications
Before biologists can determine a species’ distribution, needs, and population status, it is important to know the identity of the individuals being studied. While this may sound like a straightforward task, the process of identifying (and naming) a species can be deceptively hard, even for professional taxonomist. For example, a recent study found that 58% of 4,500 wild African ginger (Aframomum spp.) specimens that were deposited by professional biologists across 40 herbaria in 21 countries were given the wrong name (Goodwin et al., 2015)!
Describing species can be difficult, in part, because the multiple methods used by biologists to separate species do not always give the same results.
Identifying species can be hard, in part because the three tests biologists use to separate different species—morphology, biology, and evolution—do not always give the same results. That is because the methods and assumptions of each test are different. For example, some species have several varieties with easily observed morphological differences but are biologically and genetically similar enough that all those varieties are still considered a single species. A well-known example is the single species Canis familiaris, or domestic dog, whose wildly variable and numerous breeds can interbreed despite their large morphological differences. In contrast, some butterflies are considered distinct species because they cannot interbreed and have a characteristic genetic makeup, even though they cannot be separated by the naked eye.
Another important aspect complicating species identifications is that speciation—whereby one species evolves into another—is a slow and gradual process; for some species, it may take many thousands of years. Consequently, much controversy exists about where to draw the “new species” line; in other words, when is a species distinct enough to be considered a separate species? Africa’s iconic giraffes (Giraffa camelopardalis, VU) are a case in point. Taxonomists recently suggested that the region’s giraffes—previously considered a single species—may, in fact, consist of four (Fennessy et al., 2016) or even six (Brown et al., 2007) species. Unfortunately, the final number of giraffe species is still disputed because of the different assumptions made by each study and how that impacts the number of species (Bercovitch et al., 2017). Similarly, biologists often struggle to split and identify cryptic species—undescribed species that are wrongly grouped with other similar-appearing species. A recent study estimated that 60% of newly discovered species are cryptic (Ceballos and Ehrlich, 2009). Even well-known groups may suffer from this problem: there is a reasonable chance that the bushbuck (Tragelaphus scriptus. LC) and klipspringer (Oreotragus oreotragus, LC) may in fact consist of several cryptic species yet to be described (Plumptre and Wronski, 2013; Groves et al., 2017).
Hybridisation plays an important role in speciation, but it can also be detrimental to conservation efforts, particularly when it involves rare species and/or human disturbance.
To complicate matters even further, some species are closely related enough that they sometimes mate and produce hybrids. These hybrids blur the distinction between species, particularly those that may be early in the process of speciation. For some taxa, hybridisation naturally occurs in areas where the distribution ranges of related species overlap (e.g. de Jong and Butynski, 2010). Such natural hybridisation plays an important role in speciation (the evolution of new species); for example, it may have contributed to the high diversity of cichlid fishes in Africa’s Rift Valley lakes (Salzburger et al., 2002). But hybridisation can also be detrimental to conservation efforts, particularly when it involves rare species and/or human disturbance. For example, when humans reduce one species’ populations so much that they struggle to find reproductive partners of their own kind (e.g. vaz Pinto et al., 2016), when humans remove dispersal barriers that kept related species apart (e.g. Mondol et al., 2015), or when humans force related species that naturally occupy separate distributions to live together through translocations (e.g. Grobler et al., 2011; Benjamin-Fink and Reilly, 2017; van Wyk et al., 2017). While some hybrids may be sterile and thus unable to reproduce, at other times the resulting offspring can be quite strong in an evolutionary sense—a condition known as hybrid vigour (or heterosis)—and may outcompete their parent species. Such is the case with a land snail from the Seychelles (Pachnodus velutinus, EX), which was recently driven to extinction by hybridisation with a closely-related species (Gerlach, 2009)—hybrid individuals can still be found where P. valutinus used to occur.
Conversely, there may also be times when conservation biologists get it wrong and prioritise a species that does not warrant specific status. The Liberian greenbul (Phyllastrephus leucolepis) is one such example. Known from only a handful of records, this species was considered Critically Endangered until 2016, when geneticists discovered that the Liberian greenbul was the same species as the common icterine greenbul (Phyllastrephus icterinus, LC), but with an unusual coloration due to nutrient deficiencies (Collinson et al., 2018).
Implications of challenging species identifications
The difficulties in distinguishing between species have several practical conservation implications. First, when it is hard to identify a species, it may also be hard to determine that species’ true population size and distribution which, in turn, impacts its conservation status. This was illustrated in a study on bushmeat markets in Guinea-Bissau, which showed how primate misidentifications hide the true impact of hunting on some of the region’s most impacted species (Minhós et al., 2013). It also hampers captive breeding projects, by making the captive populations susceptible to outbreeding depression, which occurs when individuals that are not closely related (i.e. from different populations) breed and produce offspring (Conservation genetics is discussed in more detail in Section 8.7). Lastly, identification challenges with cryptic species can also cause delays in the formal description process, a necessary step in writing effective laws to protect them. The recent controversy among biologists arguing whether Africa’s elephants are one or two species is a case in point. African elephants were already considered threatened when biologists thought they were a single species. This all changed in 2005, when taxonomic authorities officially recognized two elephant species, in effect dividing a single threatened species into two (thus even more imperilled) species (CBD, 2015). Yet, to avoid leaving hybrid elephants (e.g. Mondol et al., 2015) with an uncertain conservation status, the IUCN continues to assess elephants as one single species (Blanc, 2008); thus, their current Vulnerable assessment may not be an accurate reflection of each species’ true conservation status.
Despite these challenges, conservationist biologists need to make every effort to obtain correct identifications. For most studies, morphological methods may be adequate. But when there is doubt, it is important for researchers to confirm their identifications with additional methods. Recent progress in making genetic technology more widely accessible through hand-held devices (Pennisi, 2016; Parker et al., 2017) and techniques such as DNA barcoding has also greatly enhanced our ability to correctly classify cryptic species, allowing us to give those species the conservation attention they deserve (Box 3.2).
Box 3.2 Golden Mole Conservation Requires a Sound Taxonomy
Sarita Maree1,2 and Samantha Mynhardt2
1Department of Genetics, &
2Department of Zoology and Entomology,
University of Pretoria, South Africa.
Golden moles (Chrysochloridae) are small, subterranean insectivores that rank among Africa’s most unique, most threatened, and yet poorly studied mammals thanks to their secretive burrowing lifestyle. Ten of the 21 known species are currently threatened with extinction (IUCN, 2019) as their highly restricted and naturally fragmented sandy soil habitats are under threat from human activities. Current conservation efforts are severely jeopardised by taxonomic uncertainties and ambiguous evolutionary relationships, thus far based on morphological and limited genetic data, which suggest that many distinct but cryptic species remain undescribed (Taylor et al., 2018).
To remedy the dearth in knowledge on two endemic South African golden mole species, we analysed molecular data of individuals collected across the entire distribution range of both Juliana’s golden mole (Neamblysomus julianae, EN) and Hottentot golden mole (Amblysomus hottentotus, LC) (Figure 3.B). In contrast to the widespread Hottentot golden mole, the Juliana’s golden mole counts among South Africa’s most imperilled mammals and is known from only three range-restricted, geographically isolated populations (Maree, 2015, Maree et al., 2016; Maree, 2017; Taylor et al., 2018). These three populations, together covering less than 160 km2 occur in southeastern Pretoria (Gauteng population), the district of Modimolle (Limpopo, ~ 120 km north of Pretoria), and in southwestern Kruger National Park (Mpumalanga, ~ 400 km east of Pretoria) (Figure 3.C).
Using molecular and other genetic methods, we have gained insights about the evolutionary relationships and gene flow between these two golden mole species, which have several conservation implications. First, preliminary findings suggest that the Hottentot golden mole contains several morphologically similar, but evolutionary distinct and genetically divergent lineages, some of which would represent undescribed cryptic species (Mynhardt et al., 2015; Taylor et al., 2018). Similarly, preliminary evidence suggests the Juliana’s golden mole contains pronounced genetic separation between the Mpumalanga population and the Gauteng and Limpopo populations. This also corresponds to morphological differences observed between these populations, which collectively suggest that the Mpumalanga population of Juliana’s golden mole might in fact be a cryptic species (Maree, 2015; Maree et al., 2016; Maree, 2017; Taylor et al., 2018). Unfortunately, in each of these cases the knowledge gaps remaining precluded definitive conclusions. Rigorous geographic sampling and additional molecular/genomic analyses will be needed to confirm the taxonomic status and geographic boundaries of putative new species within these and other golden mole taxa (Taylor et al., 2018).
Our results show that genetic frameworks contribute substantially to informed conservation decision-making. For golden moles and other taxa, some newly described species will undoubtedly be considered more threatened than in their previous species designations. Threat assessments on the Juliana’s golden mole has already identified the Gauteng population as Critically Endangered due to severe habitat loss and transformation within its highly restricted and already fragmented range (~ 22 km2 in extent) caused by rapid urbanisation and opencast sand mining. This pressure is exacerbated by this species’ extreme habitat specificity and poor dispersal capabilities (Jackson and Robertson, 2011; Maree, 2015; Maree et al., 2016; Maree, 2017; Taylor et al., 2018). Species distribution modelling (SDM, discussed in Section 10.1.1) predicted several regions throughout Gauteng, Mpumalanga, and Limpopo Provinces where the species could potentially occur, but subsequent surveys led to the discovery of only two new localities around Modimolle (Jackson and Robertson, 2011). This finding emphasises that the protection of all suitable habitats remaining for the species and the Pretoria population, in particular, would be key to its persistence. Strategies to achieve this ought to be incorporated into current conservation planning (Maree, 2015; Maree et al., 2016; Maree, 2017; Taylor et al., 2018).
We also illustrated the importance of maintaining the integrity of geographically isolated and/or genetically unique populations, lest yet undescribed species be lost to extinction before they could be fully recognized. A sound taxonomy, obtained through genetic analyses, thus contributes substantially to informed conservation decision-making. Even in the absence of such information, it is still crucial that isolated populations be managed as distinct units to conserve the evolutionary history of different species and populations.
Because the demand for expert taxonomists outstrips their availability, there is also a need to train and employ more taxonomists, particularly in the tropics and other species-rich areas. The public can help in this endeavour. In 2015, citizen scientists—volunteers participating in scientific projects—discovered 51 of 60 new dragonfly species from Africa that were described that year (Dijkstra et al., 2015). For conservation biologists, it is also important to not become despondent about the lagging efforts to describe species. They should instead take an example from motivated parrot lovers who were motivated to work even harder to get their study species recognized as distinct (Box 3.3). It is also important to keep in mind that species are never fixed; evolve all the time, albeit at different rates, due to challenges and opportunities presented by their environment.
Box 3.3 Does Tardy Recognition of a Species Hamper its Conservation?
Colleen T. Downs
School of Life Sciences, University of KwaZulu-Natal,
Pietermaritzburg, South Africa.
[email protected]
The usefulness of subspecies in conservation has long been a subject of controversy (Coetzer et al., 2015). Accurately drawing the line between an individual species and other similar animals is important for effective studies of biodiversity, and for planning and implementing official conservation strategies. Across Africa, there are many species with very broad historical distributions that are thought to contain locally adapted varieties. However, the distributions of many of these species are now fragmented and disjointed, mainly because of changes in available habitat. Examples include reptiles, such as the Nile crocodile (Crocodylus niloticus, LC), mammals, such as the common hippopotamus (Hippopotamus amphibious, VU), and a wide range of bird species. As a result of this fragmentation, various subspecies, recognized by morphology and habitat distribution, are now recognized as individual species. Modern DNA technology allows these discoveries to be supported with genetic evidence.
Protecting a newly recognized species can be difficult; genetic testing takes time and funding, and if an animal or plant is threatened before it has full species status, conservation success is that much more difficult. An example is the Cape parrot (Poicephalus robustus, EN), a forest species which was first suggested in 1997 to be a separate species and distinct from the more widespread grey-headed parrot (Poicephalus fuscicollis, LC) of Africa’s savannah ecosystems. Additional support for the Cape parrot (Figure 3.D) being a separate species came from ecological and morphological data in 2002 (Wirminghaus et al., 2002) and separate genetic evidence in 2015 (Coetzer et al., 2015). Although many published bird guides reflect the change, the species was recently recognized as a species by authorities (e.g. BirdLife International, 2017), which affected its ability to receive legal protection. The Cape parrot is endemic to South Africa, with a distribution primarily restricted to southern mist-belt Afromontane forests in the Eastern Cape and southern KwaZulu-Natal plus a relict population in Limpopo Province. Cape parrots are restricted in their distribution by their specialised habitat and dietary requirements for particular fruits. A decrease in this species’ abundance over the past 50 years is a consequence of several factors, including habitat fragmentation and degradation, food and nest site shortages, illegal trade of the birds for pets and aviculture, and disease.
Dedicated researchers have recognized the importance of determining population size and raising the awareness of the plight of the Cape parrot and the forests for which it is a flagship species. Current abundance of the Cape parrot is relatively low but stable, with an estimate of fewer than 1,600 birds in the wild (Downs et al., 2014). Estimates are based on an annual census held since 1998, organized by citizen scientists. For the Cape parrot, tardy genetic recognition of full species status was overcome by conservationists’ perseverance. We must be vigilant if we want to protect other still-hidden species from future extinction.
Measuring species diversity
Biologists have developed three quantitative measures of species diversity as a means of measuring and comparing species diversity (Figure 3.7):
• Alpha diversity (or species richness), the most commonly referenced measure of species diversity, refers to the total number of species found in a particular biological community, such as a lake or a forest. Bwindi Forest in Uganda, with an estimated 350 bird species, has one of the highest alpha diversities of all African ecosystems.
• Gamma diversity describes the total number of species that occur across an entire region, such as a mountain range or continent, that includes many ecosystems. The Albertine Rift, which includes Bwindi Forest, has more than 1,074 species of birds, a very high gamma diversity for such a small region.
• Beta diversity connects alpha and gamma diversity. It describes the rate at which species composition changes across a region. For example, if every wetland in a region was inhabited by a similar suite of plant species, then the region would have low beta diversity; in contrast, if several wetlands in a region had plants communities that were distinct and had little overlap with one another, the region would have high beta diversity. Beta diversity is calculated as gamma diversity divided by alpha diversity. The beta diversity for forest birds of the Albertine Rift is about 3.0, if each ecosystem in the area has about the same number of species as Bwindi Forest.
It is important to note that alpha, beta, and gamma diversity describe only part of what is meant by biodiversity. For example, none of these three terms completely account for genetic diversity, which allows species to adapt as conditions change (Section 8.7.1). It also neglects the importance of ecosystem diversity, which results from the collective response of species to their dynamic environment. However, these diversity measures are useful for comparing different regions, and identifying locations with high concentrations of native species that should be protected.
How many species exist?
To date, taxonomists have described about 1.5 million species that share this planet with us (Costello et al., 2012). While this total may seem impressive, available evidence suggests that this estimate vastly underestimates the true extent of Earth’s biodiversity. In fact, even now, after all the exploration in years gone by, several thousand new species are being described each year. Many new discoveries are made by skilled researchers recognizing new species by being able to discern variation in morphological characters; that includes the discoveries of a new small forest antelope from West Africa (Colyn et al., 2010) and a new species of shark off Mozambique (Ebert and Cailliet, 2011). Such discoveries can also be rather surprising and unexpected. For example, an amateur botanist recently discovered two new flowering plants in the heavily studied Cape Floristic Region (Bello et al., 2015). Similarly, the lesula (Cercopithecus lomamiensis)—a species of monkey long known to local hunters—was only formally described after biologists discovered this “different” monkey on a leash in a remote village of the DRC (Hart et al., 2012). Some recent discoveries even include entire new communities in unexpected places. For example, in 2007, grassland surveys by citizen scientists in an area starting 5 km from South Africa’s Johannesburg metropolitan area found previously unknown populations of five threatened bird species, as well as a number of regionally threatened birds and mammals; these discoveries were instrumental in recognizing this area as the Devon Grasslands Important Birding Area (Marnewick et al., 2015).
New genetic technologies have highlighted that there are many thousands of species yet to be described.
The most exciting and newsworthy discoveries of new species generally involve higher-level taxa, especially living fossils. For example, in 1938, biologists across the world were stunned by the report of a strange fish caught in the Indian Ocean off South Africa. This fish, subsequently named coelacanth (Latimeria chalumnae, CR), belongs to a group of marine fishes that were common in ancient seas but were thought to have gone extinct 65 million years ago. Coelacanths are of interest to evolutionary biologists because they show certain features of muscles and bones in their fins that are comparable to the limbs of the first vertebrates that crawled onto land. Following the initial discovery, coelacanths have been found along Africa’s Indian Ocean coast from South Africa to the Comoros and through to Kenya. Unfortunately, the entire coelacanth population, estimated at fewer than 500 individuals, is currently highly threatened because of ongoing fishing pressures (Musick, 2000).
Although field surveys have proven to be of great importance for discovering new species and populations, perhaps the greatest taxonomic progress has come from advances in genetic analyses which help to separate cryptic species previously lumped under more widespread species. For example, advances in genetic research recently highlighted that the African clawed frog (Xenopus laevis, LC)—a popular model organism in biomedical research—consists of seven distinct species (Evans et al., 2015). Similarly, using new genetic methods, scientists recently confirmed that the slender-snouted crocodile (Mecistops cataphractus, CR) consists of two different species, one endemic to West Africa and the other to Central Africa (Shirley et al., 2018).
Estimates suggest there are somewhere between 1–6 billion distinct species on Earth. The most diverse group of species is bacteria.
The presence of so many undiscovered species and communities makes precise estimates of species diversity incredibly difficult, especially in Africa where so many areas remain scientifically unexplored. Our most recent estimates, combining genetic analysis of well-known groups with mathematical patterns, suggests there are between 1–6 billion distinct species on Earth (Table 3.1) of which there are only about 163 million animals and 340 thousand plants (Larsen et al., 2017)—this is obviously much greater than the current catalogue of 1.5 million species! Given the amount of new species that continue to hide in plain sight, so to speak, there is no doubt that a great number of species and communities are waiting to be discovered by eager African adventurers over the next several decades.
Table 3.1 Estimated living biomass and number of species for each kingdom of life, following the seven-kingdom system (Ruggiero et al., 2015). Note how plants weigh the most, but bacteria have the most species.
Kingdom
Weight (Gt)a
Number of species (in million)
% of all speciesb
Number of described speciesc
% of described species
Animals
2
163
7
1,205,336
< 1
Fungi
12
165
7
135,110
< 0.1
Plants
450
0.382c
< 0.5
364,009
95
Chromista
Unknown
0.025c
< 0.5
23,428
94
Protozoans
4
163
7
2,686
0.1
Archaea
7
0.0005
< 0.5
377
75
Bacteria
70
1,746
78
9,982
0.1
a As gigatonnes of carbon, from Bar-On et al., 2018
b From Larsen et al. (2017)’s Table 1, Scenario 1
Where are most species found?
Because it is so hard to obtain accurate estimates of species numbers, many conservation biologists have recently started to focus their efforts on understanding and planning around patterns of species diversity. This makes sense: regions with many species of one taxon tend to also have many species of other taxa, so protecting one diverse group of species will likely also protect many other species, even if those other species are not well understood. Consequently, many conservation biologists see the forests of the Congo Basin, Albertine Rift, and West Africa as critical conservation priorities because these areas hold Africa’s greatest species concentrations, particularly birds, mammals, and butterflies. But there are very important outliers. For example, due to factors that include the geology and soil characteristics, size and variability of the environment, historical circumstances, or climatic conditions, none of these tropical forest areas have as many plant species as the Cape Floristic Region of South Africa—an area of unparalleled importance for plant diversity. Species diversity relationships may also break down at the local scale; for example, amphibians are likely more diverse in wet, shady riverbeds, whereas reptiles may be more diverse in drier, open habitats even if only tens of metres of space separate the reptiles from a riverbed full of amphibians.
By examining all these patterns of species diversity across the world, biologists have discovered at least two general frameworks governing species richness. The first framework is that stable ecosystems usually have many species, while ecosystems that were subjected to more recent glaciation usually have fewer species. This observation explains why tropical ecosystems are generally considered the world’s most species-rich environments (Table 3.2). While tropical grasslands, wetlands, and other ecosystems all hold relatively high species diversity, species richness of tropical forests are particularly noteworthy; even though these areas occupy only about 7% of Earth’s land surface, they contain over half of the world’s species (Corlett and Primack, 2010). This is, in a large part, due to the relatively large global distribution of the tropical forests and the diversity of geological history between these areas of South and Central America, Africa, Asia, and Australia, which has resulted in unique assemblages of species that have evolved in isolation from each other.
Table 3.2 Number of endemic and native mammal species as a function of the environment, for a range of African countries.
Country
Dominant ecoregion
Area (× 1000 km2)
Number of endemic mammals
Number of native mammals
Mammals per 1000 km2
Seychelles
Oceanic island
0.45
2
24
53.3
Cabo Verde
Oceanic island
4.03
0
29
7.2
Rwanda
Montane forest
26.8
1
189
7.05
Eq. Guinea
Lowland forest
28.1
3
184
6.55
Burundi
Montane forest
27.8
1
144
5.18
Sierra Leone
Varied Forest
71.7
0
197
2.75
Zimbabwe
Savannah
391
0
204
0.52
Zambia
Savannah
753
5
242
0.32
Namibia
Desert
825
3
206
0.25
South Africa
Varied
1,221
31
307
0.25
South Sudan
Sahel
644
1
151
0.24
DRC
Varied
2,345
26
438
0.19
Niger
Sahel
1,267
0
134
0.11
Source: IUCN, 2019.
Tropical forests are not the only species-rich tropical ecosystem. Tropical coral reefs, colonies of tiny aquatic invertebrates that form entire ecosystems (Figure 3.8), are the marine equivalent of tropical forests both in terms of species richness and complexity. These areas not only provide homes for corals, but also for huge numbers of fish, molluscs, and marine mammals that find shelter in these highly productive and sheltered ecosystems. In Africa, tropical coral reefs are most widespread and diverse in coastal East Africa, but unique tropical coral reef communities can also be found along Mozambique and South Africa’s north-eastern coast.
High levels of species diversity, especially among plants, can also be found in ecosystems with a Mediterranean climate, such as southwestern Africa, as well as southwestern Australia, California, central Chile, and the Mediterranean Basin of southern Europe and North Africa. The climate of a Mediterranean-type ecosystem is characterized by cool, moist winters, hot, dry summers, resulting in distinctive plant adaptations such as short twigs and stiff leaves. A combination of special environmental factors, including a considerably old geological age, complex site characteristics (such as varied topography and soils), and frequent fires facilitated rapid speciation and helped to prevent any one species from dominating. Today, although regions with a Mediterranean climate cover only 2% of Earth’s surface, 20% of all plant species are found here (Underwood et al., 2009). The Cape Floristic Region—the only Mediterranean climate in Sub-Saharan Africa—is particularly important to conservationists as it has the highest concentration of higher plant diversity (over 9,000 species) in the world.
The second framework governing pattern of species diversity is that locations with high numbers of species usually hold many endemic species. The Cape Floristic Region, for example, boasts more than 6,200 endemic plant species, which include 12 endemic families and 160 endemic genera. Similarly, Lake Malawi holds nearly 14% of the world’s freshwater fishes (500–1,000 species, totals vary by source), with more than 90% of those being endemic.
Biogeographic transition zones—also known as ecotones—regions where different ecosystems meet and overlap, are a special case of areas that contain great species diversity and high levels of endemicity. These areas share environmental factors of two or more environments, allowing for the mixture of biodiversity from those component environments, while unique features within these areas often also give rise to unique species. A case in point is the Maputaland Centre of Endemism, situated in far southern Mozambique. Here, biological communities from northern tropical and southern temperate ecosystems overlap, resulting in surprisingly high levels of species richness as well as endemism (van Wyk, 1996).
Today is an exciting time of biological exploration. Methods and technologies for exploration are improving rapidly, and we are learning more about the value and function the diversity of life on Earth. As genetic techniques advance and become more accessible, an increasing number of people are participating in recording the presence of species in locations around the world; this includes amateur naturalists and citizen scientists who contribute to bird surveys, plant walks, and other natural history activities. With this increased knowledge of biodiversity also comes an acute awareness that human activities damage ecosystems and reduce diversity. Hopefully this broader awareness will spur more people to take responsibility to protect and restore that biodiversity. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/03%3A_What_is_Biodiversity/3.04%3A__Patterns_of_Biodiversity.txt |
1. Earth’s biodiversity includes the entire range of living species (species diversity), the genetic variation that occurs among individuals within a species (genetic diversity), and, at a higher level, the biological communities in which species live and their associations with the physical and chemical environment (ecosystem diversity).
2. For practical purposes, most ecologists and conservationists identify species in the field according to their morphology, although improvements in genetic techniques are allowing more species to be identified according to their evolutionary past, revealing many cryptic species that people did not realise were there.
3. There are several ways to measure and compare biodiversity. The most popular measurement is species richness in a particular community, such as a forest or grassland (alpha diversity), species richness across a larger landscape, such as a mountain range (gamma diversity), and the rate of change of species composition as one crosses a large region (beta diversity).
4. It is estimated that there may be as many as 2 billion species on Earth. Most species already described are insects, while the best-known species include birds and mammals. The majority of species still need to be discovered.
5. Variation in climate, topography, and geological age are all factors that affect patterns of species richness. Geological age and complexity provide environmental variation, which in turn allows opportunities for genetic isolation, local adaptation, and speciation, given enough time. Tropical forests, coral reefs, and Mediterranean-type ecosystems host a disproportionately large amount of the world’s biodiversity.
3.06: Topics for Discussion
1. Think of any group of species (birds, trees, or maybe insects) that can be found in the area where you live. Do you think it is important to be able to identify these species? Why? How many species can you personally identify at this moment? What steps would you take to learn to identify more species?
2. Which ecosystems in your country are particularly species-rich, and which are species-poor? Describe some factors that make ecosystems species-rich or species-poor.
3. Where in Africa do you think most undescribed species are lurking? Explain your answer.
3.07: Suggested Readings
Ceballos, G., and P.R. Ehrlich. 2009. Discoveries of new mammal species and their implications for conservation and ecosystem services. Proceedings of the National Academy of Sciences 106: 3841–46. https://doi.org/10.1073/pnas.0812419106 Even familiar taxa contain many undescribed species.
Bar-On, Y.M., R. Phillips, and R. Milo. 2018. The biomass distribution on Earth. Proceedings of the National Academy of Sciences 25: 6505–11. https://doi.org/10.1073/pnas.1711842115 Note the impact of humans on the composition of life on Earth.
Ebach, M.C., J.J. Morrone, L.R. Parenti, et al. 2007. International code of area nomenclature. Journal of Biogeography 35: 1153–57. https://doi.org/10.1111/j.1365-2699.2008.01920.x Scientists also grapple with confusing terminology when describing biodiversity.
Gippoliti, S., F.P.D. Cotterill, D. Zinner, et al. 2018. Impacts of taxonomic inertia for the conservation of African ungulate diversity: An overview. Biological Reviews 93: 115–30. https://doi.org/10.1111/brv.12335 Taxonomic inertia, or the delay in recognizing distinct species, slows conservation efforts.
Hart, J.A., K.M. Detwiler, C.C. Gilbert, et al. 2012. Lesula: A new species of Cercopithecus monkey endemic to the Democratic Republic of Congo and implications for conservation of Congo’s Central Basin. PLoS ONE 7: e44271. https://doi.org/10.1371/journal.pone.0044271 Many species known to local people still need to be formally described.
Joppa, L.N., D.L. Roberts, and S.L. Pimm. 2011. The population ecology and social behavior of taxonomists. Trends in Ecology and Evolution 26: 551–53. https://doi.org/10.1016/j.tree.2011.07.010 The number of taxonomists and the number of species described per year are steadily increasing.
Laikre, L., F.W. Allendorf, L.C. Aroner, et al. 2010. Neglect of genetic diversity in implementation of the Convention on Biological Diversity. Conservation Biology 24: 86–88. https://doi.org/10.1111/j.1523-1739.2009.01425.x A greater emphasis on genetic diversity needs to be part of conservation efforts.
Larsen, B.B., E.C. Miller, M.K. Rhodes, et al. 2017. Inordinate fondness multiplied and redistributed: The number of species on Earth and the new pie of life. Quarterly Review of Biology 92: 229–65. https://doi.org/10.1086/693564 We have much to learn about life on Earth.
Minhós, T., E. Wallace, M.J.F. da Silva, et al. 2013. DNA identification of primate bushmeat from urban markets in Guinea-Bissau and its implications for conservation. Biological Conservation 167: 43–49. https://doi.org/10.1016/j.biocon.2013.07.018 Misidentifications could have significant conservation implications. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/03%3A_What_is_Biodiversity/3.05%3A_Summary.txt |
Bar-On, Y.M., R. Phillips, and R. Milo. 2018. The biomass distribution on Earth. Proceedings of the National Academy of Sciences 25: 6505–11. https://doi.org/10.1073/pnas.1711842115
Bello A., C.H. Stirton, S.B.M. Chimphango, et al. 2015. Psoralea diturnerae and P. vanberkelae (Psoraleeae, Fabaceae): Two new species restricted to the core Cape Region of South Africa. PhytoKeys 44: 97–107. https://doi.org/10.3897/phytokeys.44.8999
Benjamin-Fink, N., and B.K. Reilly. 2017. Conservation implications of wildlife translocations: The state’s ability to act as conservation units for wildebeest populations in South Africa. Global Ecology and Conservation 12: 46–58. https://doi.org/10.1016/j.gecco.2017.08.008
Bercovitch, F.B., P.S.M. Berry, A. Dagg, et al. 2017. How many species of giraffe are there? Current Biology 27: R136–R137. https://doi.org/10.1016/j.cub.2016.12.039
BirdLife International. 2017. Poicephalus robustus. The IUCN Red List of Threatened Species 2017: e.T119194858A119196714. http://doi.org/10.2305/IUCN.UK.2017-3.RLTS.T119194858A11919
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Blanc, J. 2008. Loxodonta africana. The IUCN Red List of Threatened Species 2008: e.T12392A3339343. http://doi.org/10.2305/IUCN.UK.2008.RLTS.T12392A3339343.en
Brown, D.M., R.A. Brenneman, K.P. Koepfli, et al. 2007. Extensive population genetic structure in the giraffe. BCM Biology 5: 57. https://doi.org/10.1186/1741-7007-5-57
CBD (Center for Biological Diversity). 2015. Petition to reclassify and uplist African elephants from Threatened to Endangered under the Endangered Species Act as two separate species: forest elephants (Loxodonta cyclotis) and savannah elephants (Loxodonta africana) (Tucson: CBD). https://www.biologicaldiversity.org/species/mammals/pdfs/African_Elephant_Uplisting_Petition.pdf
Ceballos, G., and P.R. Ehrlich. 2009. Discoveries of new mammal species and their implications for conservation and ecosystem services. Proceedings of the National Academy of Sciences 106: 3841–46. https://doi.org/10.1073/pnas.0812419106
Coetzer, W.R., C.T. Downs, M.R. Perrin, et al. 2015. Molecular systematics of the Cape Parrot (Poicephalus robustus): Implications for taxonomy and conservation. PLoS ONE 10: e0133376. https://doi.org/10.1371/journal.pone.0133376
Collinson, J.M., M. Päckert, Y. Lawrie, et al. 2018. Taxonomic status of the Liberian Greenbul Phyllastrephus leucolepis and the conservation importance of the Cavalla Forest, Liberia. Journal of Ornithology 159: 19–27. https://doi.org/10.1007/s10336-017-1477-0
Colyn, M., J. Hulselmans, G. Sonet, et al. 2010. Discovery of a new duiker species (Bovidae: Cephalophinae) from the Dahomey Gap, West Africa. Zootaxa 2637: 1–30. http://doi.org/10.11646/zootaxa.2637.1.1
Corlett, R., and R.B. Primack. 2010. Tropical Rainforests: An Ecological and Biogeographical Comparison (Malden: Wiley-Blackwell). https://doi.org/10.1002/9781444392296
Costello, M.J., S. Wilson, and B. Houlding. 2012. Predicting total global species richness using rates of species description and estimates of taxonomic effort. Systematic Biology 61: 871–83. http://doi.org/10.1093/sysbio/syr080
Davis A.P., T.W. Gole, S. Baena, et al. 2012. The impact of climate change on indigenous Arabica coffee (Coffea arabica): Predicting future trends and identifying priorities. PLoS ONE 7: e47981. http://doi.org/10.1371/journal.pone.0047981
de Jong, Y.A., and T.M. Butynski. 2010. Three sykes’s monkey Cercopithecus mitis × vervet monkey Chlorocebus pygerythrus hybrids in Kenya. Primate Conservation 25: 43–56. https://doi.org/10.1896/052.025.0109
Dijkstra, K.-D.B., J. Kipping, and M. Nicolas. 2015. Sixty new dragonfly and damselfly species from Africa (Odonata). Odonatologica 44: 447–678.
Downs, C.T., M. Pfeiffer, and L. Hart. 2014. Fifteen years of annual Cape Parrots (Poicephalus robustus) censuses: Current population trends and conservation contributions. Ostrich 85: 273–80. https://doi.org/10.2989/00306525.2014.959088
Ebert, D.A., and G.M. Cailliet. 2011. Pristiophorus nancyae, a new species of sawshark (Chondrichthyes: Pristiophoridae) from southern Africa. Bulletin of Marine Science 87: 501–12. https://doi.org/10.5343/bms.2010.1108
Evans B.J., T.F. Carter, E. Greenbaum, et al. 2015. Genetics, morphology, advertisement calls, and historical records distinguish six new polyploid species of African clawed frog (Xenopus, Pipidae) from West and Central Africa. PLoS ONE 10: e0142823. https://doi.org/10.1371/journal.pone.0142823
Fennessy, J., T. Bidon, F. Reuss, et al. 2016. Multi-locus analysis reveals four giraffe species instead of one. Current Biology 26: 2543–49. https://doi.org/10.1016/j.cub.2016.07.036
Gerlach, J. 2009. Pachnodus velutinus. The IUCN Red List of Threatened Species 2009: e.T40091A10304648. http://doi.org/10.2305/IUCN.UK.2009-2.RLTS.T40091A10304648.en
Goodwin, Z.A., D.J. Harris, D. Filer, et al. 2015. Widespread mistaken identity in tropical plant collections. Current Biology 25: R1066–R1067. https://doi.org/10.1016/j.cub.2015.10.002
Grobler, J.P., I. Rushworth, J.S. Brink, et al. 2011. Management of hybridization in an endemic species: decision making in the face of imperfect information in the case of the black wildebeest—Connochaetes gnou. European Journal of Wildlife Research 57: 997–1006. http://doi.org/10.1007/s10344-011-0567-1
Groves, C.P., F.P.D. Cotterill, S. Gippoliti, et al. 2017. Species definitions and conservation: A review and case studies from African mammals. Conservation Genetics 18: 1247–`156. https://doi.org/10.1007/s10592-017-0976-0
Hart, J.A., K.M. Detwiler, C.C. Gilbert, et al. 2012. Lesula: A new species of Cercopithecus monkey endemic to the Democratic Republic of Congo and implications for conservation of Congo’s Central Basin. PLoS ONE 7: e44271. http://doi.org/10.1371/journal.pone.0044271
IUCN, 2019. The IUCN Red List of Threatened Species. http://www.iucnredlist.org
Jackson, C.R., and M.P. Robertson. 2011. Predicting the potential distribution of an endangered cryptic subterranean mammal from few occurrence records. Journal for Nature Conservation 19: 87–94. http://doi.org/10.1016/j.jnc.2010.06.006
Larsen, B.B., E.C. Miller, M.K. Rhodes, et al. 2017. Inordinate fondness multiplied and redistributed: The number of species on Earth and the new pie of life. Quarterly Review of Biology 92: 229–65. https://doi.org/10.1086/693564
Maree S., N.C. Bennett, and G.N. Bronner. 2016. A conservation assessment of Neamblysomus julianae. In: The Red List of Mammals of South Africa, Swaziland and Lesotho, ed. by M.F. Child, et al. (Cape Town and Johannesburg: SANBI and EWT).
Maree S. 2017. Planning for persistence of a Juliana’s Golden Mole (Neamblysomus julliane) subpopulation threatened by urban development on the Bronberg Ridge of Pretoria (Tshwane), South Africa. IUCN/SSC Afrotheria Specialist Group Newsletter 13: 24–33. https://www.afrotheria.net/newsletter.php
Maree, S. 2015. Neamblysomus julianae. The IUCN Red List of Threatened Species 2015: e.T1089A21285354. http://doi.org/10.2305/IUCN.UK.2015-2.RLTS.T1089A21285354.en
Marnewick, M.D., E.R. Retief, N.T. Theron, et al. 2015. Important Bird and Biodiversity Areas of South Africa (Johannesburg: BirdLife South Africa). https://www.birdlife.org.za/images/IBA/Documents/IBA%20Version%202.pdf
Minhós, T., E. Wallace, M.J.F. da Silva, et al. 2013. DNA identification of primate bushmeat from urban markets in Guinea-Bissau and its implications for conservation. Biological Conservation 167: 43–49. https://doi.org/10.1016/j.biocon.2013.07.018
Mondol, S., I. Moltke, J. Hart, et al. 2015. New evidence for hybrid zones of forest and savanna elephants in Central and West Africa. Molecular Ecology 24: 6134–47. https://doi.org/10.1111/mec.13472
Musick, J.A. 2000. Latimeria chalumnae. The IUCN Red List of Threatened Species 2000: e.T11375A3274618. http://doi.org/10.2305/IUCN.UK.2000.RLTS.T11375A3274618.en
Mynhardt, S., S. Maree, I. Pelser, Net al. 2015. Phylogeography of a morphologically cryptic golden mole assemblage from South-Eastern Africa. PLoS ONE 10: e0144995. https://doi.org/10.1371/journal.pone.0144995
Nichols, E., T.A. Gardner, C.A. Peres, et al. 2009. Co‐declining mammals and dung beetles: An impending ecological cascade. Oikos 118: 481–87. https://doi.org/10.1111/j.1600-0706.2009.17268.x
Parker, J., A.J. Helmstetter, D. Devey, et al. 2017. Field-based species identification of closely-related plants using real-time nanopore sequencing. Scientific Reports 7: 8345. https://doi.org/10.1038/s41598-017-08461-5
Pennisi, E. 2016. Pocket DNA sequencers make real-time diagnostics a reality. Science 351: 800–01. https://doi.org/10.1126/science.351.6275.800
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Ruggiero, M.A., D.P. Gordon, T.M. Orrell, et al. 2015. A higher-level classification of all living organisms. PLoS ONE 10: e0119248. http://doi.org/10.1371/journal.pone.0119248
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All of us depend on nature for survival, whether we live off the land, or in a city where we can buy natural resources, transported to us from a distance, at the market. When we do not take care of nature, our quality of life suffers. To illustrate this point, in the book, Collapse (2011), prize-winning author Jared Diamond describes how, throughout history, ineffective responses to ailing environments have contributed to human conflicts. In one case study, Diamond examines how overpopulation contributed to Rwanda’s collapse into genocide in the early 1990s. Prior to the genocide, Rwanda had one of the highest human population densities in the world, putting enormous strain on its natural resources. Widespread deforestation led to erosion, which in turn contributed to famine, further escalating conflict over what resources remain.
Rwanda—situated in the Albertine Rift Biodiversity Hotspot—is not the only country in which environmental degradation has led to human pain and suffering. Between 1950 and 2000, 80% of the world’s armed conflicts occurred within the boundaries of the world’s 36 Global Biodiversity Hotspots (Hanson et al., 2009)—areas with high levels of biodiversity that also suffer from substantial environmental degradation. Even today, environmental degradation continues to play a major role in fuelling ongoing conflicts, such as those of the Middle East (Gleick, 2014), West Africa’s Sahel region (Benjaminsen, 2008), and the Horn of Africa (Markakis, 1995). Preventing these conflicts, which also impact biodiversity negatively (Nackoney et al., 2014; Brito et al., 2018; Daskin et al., 2018), from escalating and new conflicts from developing requires political and societal changes. People in government and local communities must recognize the value of healthy ecosystems and become their champions. After all, complex and adaptive ecosystems provide jobs, food, and other resources, thereby contributing to our overall well-being.
A healthy environment improves our overall wellbeing by enabling us to live healthy and prosperous lives. In other words, it is our life support system.
But what exactly are we losing when we fail to protect biodiversity? Why should we care if a species goes extinct, or an ecosystem becomes degraded? What evidence do we have that the natural world is our life support system? To better understand the importance of biodiversity for human well-being and quality of life, and the variety of benefits people freely gain from biodiversity, the UN brought together a group of leading scientists to study nature’s contributions to people (NCP, Díaz et al., 2018), more commonly referred to as ecosystem services. This group, called the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES), recognizes three categories of ecosystem services, namely material contributions, regulation services, and nonmaterial contributions. Note that there are broad overlaps and interdependence among the three categories; consequently, some contributions and services can easily fit under more than one category.
4.01: Material Contributions
Nature’s material contributions to people, also called provisioning services, commodity values or direct use values, represent contributions derived from the direct extraction and physical consumption of natural resources (Figure 4.1). This category is often the most visible and marketed of all ecosystem services. Also, because of their important contribution to the economy, economists are often interested in calculating the values of material contributions and associated services, which they do by monitoring the cost of each product at several points along its life cycle, as well as the behaviors of target groups of people.
Material contributions can be subdivided into four subcategories. The first subcategory is energy resources, such as firewood and biofuels. The second is food resources, such as drinking water, bushmeat, and edible fruit. The third is materials, companionship, and labour, which include natural products used to make clothes, ornamental resources used for decorations, and animals used for biomedical research, as pets, and for labour. The fourth is medicinal, biochemical, and genetic resources, which include medicinal plants used to cure ailments, psychoactive fungi used in spiritual ceremonies, and genetic stocks used to improve crops.
Many people, especially those in rural areas, obtain many of the material contributions they need for survival from the surrounding environment. These products, which include bushmeat, perfumes from aromatic plants, and firewood, are often assigned to consumptive use values. In contrast, material contributions that are sold at commercial markets, whether locally or internationally, are assigned to productive use values. Because of material contributions’ importance in sustaining people’s material assets and health, it is important to ensure that these products are sustainably harvested (Box 4.1).
Box 4.1 Research on Hunting Underpins Conservation in Central Africa
Katharine Abernethy1,2 and Lauren M. Coad3
1Biological and Environmental Sciences, Faculty of Natural Sciences,
University of Stirling, UK.
2Institut de Recherches en Ecologie Tropicale, CENAREST,
Gros Bouquet, Libreville, Gabon.
3CIFOR, Jalan CIFOR Situ Gede,
Sindang Barang Bogor (Barat) 16115, Indonesia.
A major threat to wildlife in Africa is hunting. Subsistence hunting has been practiced for thousands of years, but new technologies allow hunters to have higher impacts than they had in the past. Improved access routes and vehicles allow hunters to cover more ground and sell to a greater client base, while habitat encroachment from logging or agriculture squeezes wildlife into smaller areas. Growing human populations are pushing the overall demand for wildlife products to a level that the remaining fauna simply cannot support. Yet wild meat is a critically important resource in rural Central Africa, so managing hunting is an important issue for conservation and human welfare (Coad et al., 2010).
Our 20-year research program looked at hunting in Central Africa to determine how conservation may be most effective. We studied how human communities rely on hunting, impacts of hunting on wildlife and ecosystems, law enforcement challenges, and alternative practices. We found that across Central Africa hunters are in the poorer sections of society and hunt for very similar reasons: food and income. In rural villages, most able-bodied men hunt, but usually < 10% of men make most kills and have disproportionately important impacts on wildlife. These hunters have invested most in equipment and local assets; thus, they have the most to lose and are resistant to regulations or alternatives. The more successful a hunter, the more meat he sells (Coad et al., 2013). Only around 40–60% of hunted meat is consumed directly within the community; smoked or frozen meat can be traded up to 1,000 km away. Even remote villages now trade meat as a commodity to buy supplies such as fuel and medicines.
Under subsistence-driven hunting, studies show that larger-bodied species ( 20 kg) are targeted first. As these decline, smaller species are hunted (Ingram et al., 2015). During this process, the wildlife community changes and, as large predators, browsers, and seed dispersers are lost, ecosystem functioning is compromised (Abernethy et al., 2013).
Commercial hunting often targets illegal trophies, which is only lucrative if hunters have access to clients. These illegal hunters are often recruited directly by the buyer and local people may not necessarily participate, or even benefit at all. If profits are high, hunters can access better weaponry and surveillance than law enforcers, making them difficult and dangerous to apprehend. In the past 20 years, species such as elephants, rhinoceros, lions, and gorillas, have suffered drastic declines that authorities have not been able to combat.
Although wildlife protection laws are generally strong in the region, law enforcement is underfunded and complex. Commercial hunting is regulated but subsistence hunting is allowed, making the identification of illegal hunting difficult as most hunters sell only part of their catch. Alternative livelihood projects have been promoted in the hope of reducing hunting without complex enforcement. However, our review of these projects shows negligible impact, as they have generally been on a small scale and were often unreliable in generating better revenues than hunting (Wicander and Coad, 2015).
Our research shows that the effective regulation of hunting is desperately needed to preserve Central Africa’s ecosystems and the sustainability of rural communities. This will require balancing law enforcement and long-term community outreach with policy interventions—such as lobbying—to change laws or awareness campaigns. A conservation practitioner tasked with trying to manage hunting should ask who hunts, why they hunt, where hunting pressure is greatest, and how hunting affects the local ecosystem in order to determine whether they are tackling a subsistence issue or a commercially-driven one, and from there to decide which strategies could be used and who the interventions will affect. This will help to ensure planning for fair, long-term solutions, which have broad local support and the best chance of success. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/04%3A_Why_Should_We_Protect_Biodiversity/4.00%3A_Prelude_to_Why_Should_We_Protect_Biodiversity.txt |
Regulating services maintain nature’s ability to provide material contributions, usually in indirect ways. For example, our ability to obtain food indirectly depends on a variety of subtle yet important ecosystem processes (e.g. energy cycling), as well as more observable services (e.g. pollination). For that reason, regulation services are sometimes called indirect use values. Regulating services (together with nonmaterial contributions, discussed below) are also sometimes referred to as non-consumptive use values because they provide economic benefits without needing to be collected, harvested, consumed, converted, or destroyed during use.
The economic benefits we gain from regulating services are estimated to be larger than all the different kinds of material contributions together.
The economic benefits we gain from regulating services are estimated to be larger than all the different kinds of material contributions together, especially in areas where ecosystems are intact (Costanza et al., 2014). Even so, these benefits do not always appear in descriptions of national economies because those statistics generally focus on material contributions. Nonetheless, maintaining regulating services is very important. When damaged ecosystems cannot provide these benefits, substitute resources must be found—often at great expense—to avoid economic collapse. In Section 4.2.4, we discuss one such example, by considering the value and replacement costs of water maintenance services obtained from forests.
Regulating services can be subdivided into many different subcategories depending on context, each overlapping to varying degrees with one another. Following is a discussion of some prominent subcategories of regulating services.
Maintaining ecosystem stability
Perhaps the most important indirect contribution we gain from biodiversity is its ability to maintain conditions that enable life on Earth to persist. This principle complements the Gaia hypothesis, which proposes that all the biological, physical, and chemical properties on Earth interact to form a complex, self-regulating superorganism, and that these interactions maintain the conditions and processes necessary for life to persist (Lovelock, 1988).
There are two complementary theories that explain the importance of maintaining a variety of different species if one is to conserve this superorganism (Ehrlich and Walker, 1998). Originally proposed by American ecologist Paul Ehrlich, the rivet-popper hypothesis compares biodiversity to the rivets (some of which may be redundant) that hold an aeroplane together. Just as an aeroplane can only lose so many rivets before it falls apart, so will the progressive loss of species systematically weaken an ecosystem until the entire system collapses. A well-known example of the rivet-popper hypothesis is the mutualistic relationships many plants have with all the various pollinators and seed dispersers (Section 4.2.5), in this context representing the rivets holding the system together. We might not immediately notice the systematic loss of pollinators we are currently experiencing (Gallai et al., 2008; Dirzo et al, 2014), but eventually these losses will catch up with us, perhaps in the form of food insecurity.
Keystone species provide such an outsized contribution to ecosystem functioning that their loss will greatly alter ecosystem composition and functioning.
The species redundancy hypothesis, proposed by African ecologist Brian Walker, holds that biodiversity and ecosystem stability is best maintained not by focussing on preserving individual species, but by preserving redundancy in ecosystem functioning, by ensuring that each ecosystem is composed of a variety of (seemingly redundant) species performing similar roles. In other words, we should not focus our efforts on protecting just one or two important pollinating species, but a variety of them, to ensure that a variety of plants (and hence entire ecosystems) can also continue to survive. In this way, if one pollinator is lost due to an environmental disturbance or disease, the system will not collapse because other pollinating species will be able to compensate for the loss of that one species.
It is important to note that there are some individual species that provide such an outsized contribution to ecosystem functioning that their loss will greatly alter ecosystem composition and functioning. These “pilots” of natural ecosystems are generally known as keystone species (Figure 4.2). The keystone species concept was originally proposed after scientists observed that removing sea stars from intertidal zones allowed their prey (mussels) to increase uncontrollably which, in turn, pushed species, such as sea urchins and other shellfish, away, leaving an overall poorer ecosystem (Paine, 1969). Apex predators, such as lions (Panthera leo, VU) and cheetahs (Acinonyx jubatus, VU), are also keystone species because of their role in keeping herbivore populations under control. If these apex predators were to disappear, increasing herbivore populations would lead to overgrazing, and ultimately also herbivore declines. This top-down control predators exert on herbivores also answers one of modern ecology’s oldest questions: “why is the world green?” (Hairston et al., 1960).
An ecosystem engineer is a special type of keystone species that extensively modifies the physical environment, thereby creating and maintaining habitats for other species. Mount-building termites are important ecosystem engineers in many African ecosystems because their activities alter physical, chemical, and biological soil properties (Jouquet et al., 2011), and their massive mounts (some mounts are 10 m high, 20 m across, and may be over 2,000 years old) support distinctive ecological communities and serve as refuges for a large variety of animals and even plants (Loveridge and Moe, 2004; van der Plas et al, 2013). Elephants are also ecosystem engineers; their dramatic foraging habit of pushing over trees provides suitable habitats to countless small animals (Pringle, 2008). Elephants also open up dense vegetation, which allows grasses to thrive, in turn providing food for grazing antelope (Valeix et al., 2011). Holes dug by elephants sometimes make water more accessible, while elephant dung provides food for butterflies and dung beetles and creates an important germination environment for seeds and fungi. But too many elephants can also damage ecosystems by reducing the number of large trees on which other species depend (Cumming et al., 1997). It is important to remember that water is an important limiting resource for elephants (Chamaillé-Jammes et al., 2008), so there is a greater risk for elephants to become overly destructive in areas where humans artificially increase aboveground water availability.
Because so many species depend on ecosystem engineers and other keystone species for survival, their disappearance from an ecosystem can create an extinction cascade—a series of linked extinction events following one another. A related phenomenon known as a trophic cascade describes the situation where one keystone species’ loss has rippling effects at other trophic levels. Some of the best-studied trophic cascades involve apex predators and their role in suppressing prey populations (Estes et al., 2011), but disease pathogens can also be a keystone species that leads to trophic cascades. For example, the introduction of rinderpest from Asia to Africa in the late 1800s caused catastrophic ungulate population declines in East Africa through the early 1900s. With no primary consumers, grasslands were encroached by woody plants; these changes in the primary producer community also increased the intensity and frequency of wildfires, leading to cascading impacts throughout these savannah communities. An extensive vaccination program finally saw the disease eradicated in the 1960s, allowing ungulate population and grasslands to recover; and wildfires to become less destructive (Holdo et al., 2009).
The loss of keystone species from an ecosystem may create an extinction cascade—a series of linked extinction events following one another.
Maintaining ecosystem productivity
Plants and algae—in this context known as primary producers—use photosynthesis to capture and store energy from sunlight in their living tissue. This ability of ecosystems to generate living biomass, starting with plants trapping the sun’s energy, is known as ecosystem productivity. Primary consumers (i.e. herbivores) can then harvest this captured energy by eating plant material. The energy (and nutrition) obtained from plants enable herbivores to generate their own living biomass, in the form of growth and reproduction, before they, themselves, are consumed by secondary consumers (e.g. carnivores, predators, omnivores). This cycle ends (or starts, depending on one’s perspective) when decomposers and detritivores (e.g. fungi, earthworms, and millipedes) that break down complex plant and animal tissues into simple compounds such as nitrates, and phosphates. These simple compounds are then released into the soil and water, from where primary producers can take them up again.
Climate regulation
Many of us were taught from a young age that plants are the “lungs of the planet” (Figure 4.3) because they convert carbon dioxide (CO2) into breathable oxygen (O2) during photosynthesis. This contribution, whereby plants regulate the atmosphere’s CO2/O2 balance through carbon absorption and storage (termed carbon sequestration) forms part of the atmospheric carbon cycle and plays a major role in regulating global climate patterns. The reduction in plant life through deforestation or other human activities is thus of major concern because of the reduced capacity of plants to sequester atmospheric carbon dioxide, a greenhouse gas that contributes to climate change (Chapter 6). The important role of plant communities in the atmospheric carbon cycle is now even being recognized by global markets. For example, the carbon-storing capacity of the Congo Basin’s forests has an estimated value at over US \$2.5 billion per year (Hughes, 2011). As part of the worldwide effort to reduce carbon dioxide emissions and address climate change, industrial countries and corporations have started paying some landowners to preserve and restore ecosystems that store significant amounts of carbon (Section 10.4).
Plants are also important in regulating regional climate conditions by influencing both the water cycle via transpiration, and local heating and cooling via solar radiation absorption. For example, forests and other vegetation often absorb more heat than bare soil due to their respective albedos. Because heat rises, heat absorbed by vegetation enables water vapor released by plants via transpiration to rise higher into the atmosphere, where it subsequently condenses and falls as rain. In contrast, the loss of vegetation is often associated with reduced rainfall (Garcia-Carreras and Parker, 2011), which can in turn reduce agricultural productivity and biodiversity (Lawrence and Vandecar, 2015).
Lastly, trees keep local areas cool by providing shade and releasing water vapor into the atmosphere (Morakinyo et al., 2013; Kardan et al., 2015). This cooling effect increases people’s comfort and work efficiency, and reduces the need for fans or air conditioners, leading to higher productivity and cost savings (Balogun et al., 2014; Ogueke et al., 2017). Trees also act as windbreaks, thereby reducing evaporation and erosion in agricultural areas, and reducing the loss of heat from homes and other buildings in cold weather. The value of shade trees is also recognized in agro-ecosystems, as a strategy for coffee and cacao farmers to increase crop yields (Section 14.1.1) and to adapt to increasing temperatures due to climate change (Jaramillo et al., 2011).
Conserving soil and water quality
Wetlands play a prominent role in flood control. They are also very effective in immobilizing pathogens and toxic pollutants released into the environment by human activities.
Wetlands play a prominent role in regulating soil and water quality, as well as flood control. During heavy rains, wetlands slow the speed of rushing floodwater, which lowers flood height and reduces erosion. Wetlands also act as natural sponges: they absorb vast amounts of floodwater during heavy rains, which is then released more slowly and evenly afterwards, thereby maintaining water sources used for drinking, irrigation, hydropower generation, and transport. Wetlands are also very effective in breaking down and immobilising pathogens, toxic pollutants, and excess nutrients released into the environment from agricultural activities, sewage, industrial wastes, and pesticides. One study from South Africa found that wetlands were almost 100% effective in preventing further spread of highly toxic organophosphorus pesticides (Schulz and Peall, 2001).
Wetlands are, however, not the only ecosystem that maintain soil and water quality and quality. In fact, maintaining complex and adaptive ecological communities of all kinds are of vital importance in buffering ecosystems against flooding and drought, protecting fertile soils, and maintaining water quality (see also Section 10.2.1). In intact ecosystems, plant foliage and dead leaves intercept rain, which slows the flow of water from upper reaches of catchment areas into streams and rivers; this allows for a slow release of water for days or even weeks after rains have ceased. Soil is anchored in place by plant roots and aerated by soil organisms; this combination increases the soil’s capacity to absorb water and hold nutrients. All these aspects together reduce flooding and limit erosion of fertile topsoil which, in turn, limits loss of essential nutrients that would otherwise occur after heavy rains.
The economic benefits of water quality maintenance services provided by intact plant communities are enormous. In the late 1980s, the New York City administration paid US \$1.5 billion to local authorities in rural New York State to protect their water supplies by maintaining forests in the catchment area that surrounded the city’s reservoirs, and by improving agricultural practices in the catchment area. While US \$1.5 billion may seem like a lot of money, at the time it was considered a pittance compared to the US \$9 billion that the man-made water filtration systems—doing the same job—would have cost over just the first 10 years in operation (NRC, 2000).
A situation very similar to the one in New York is currently playing out in Kenya. The Mau Forest Complex is one of East Africa’s largest montane forests and serves as the principle catchment area for waters that flow into the famed Mara River and Lake Victoria. But large-scale deforestation in the Mau Forest Complex over the past few decades (Figure 4.4) has resulted in reduced water storage, flow regulation, groundwater discharge, and water purification, causing annual economic losses of over US \$65 billion to Kenya’s energy, tourism, and agricultural sectors (UNEP, 2012). The situation in Kenya was so severe that the 2008 inauguration of a hydropower station was postponed due to low water levels; this station later achieved only 50% of its production capacity as a result of deforestation in the Mau Complex. To avoid further losses, the Kenyan government initiated a multi-stakeholder taskforce to investigate options to restore the Mau complex’s degraded forests (Prime Minister’s Task Force, 2009). Since then, tens of thousands of trees have been planted to reverse deforestation in the area.
Pollination and seed dispersal
Pollination describes the transfer of pollen grains from male parts of a flower to female parts to allow fertilization and production of offspring. Some plants can be pollinated by wind, but others require animals to pollinate their flowers; examples include birds, bats, bees, flies, butterflies, and other insects (Figure 4.5). These pollination services are important for the persistence of many wild plants, as well as for many fruit, seed, and vegetable crops that we utilise as food (Box 4.2). Research from The Gambia has shown that management practices that increase the abundance of bats and bees to contribute to increased yields and sweetness of African locust bean (Parkia biglobosa) crops (Lassen et al., 2012). In contrast, work done in Zambia, Mozambique, and Uganda showed that pollinator collapse could increase malnutrition rates by over 50% which, in turn, could increase death rates among children and mothers during childbirth (Ellis et al., 2015). Luckily, many agricultural systems in Africa are still friendly to pollinators (see Box 7.4). Given the dependency on animal-assisted pollination in many agricultural systems, it is critical to maintain or expand pollinator-friendly practices. Our ability to continue benefitting from these services will depend on our ability to maintain and expand on those pollinator-friendly activities.
Box 4.2 Are Wild Pollinators Important in African Agriculture?
Abraham J. Miller-Rushing
Acadia National Park, US National Park Service,
Bar Harbor, ME, USA.
Pollinators and food security are so closely tied to one another they should almost be considered synonymous terms. But when people think of pollination, they often only think of honeybees, which people domesticated more than 8,500 years ago for honey production. However, wild pollinators, which include a variety of insects, birds, and mammals, are often more effective at pollinating than honeybees are. One estimate suggests wild pollinators can double fruit production compared to honeybees (Garibaldi et al., 2013). This is most likely because the morphological and behavioral diversity of wild pollinators allow for more specialised pollination relationships with plants. For example, some wild pollinators have longer proboscis (i.e. insect tongues) that enable them to pollinate deeper flowers (Figure 4.A), something honeybees cannot do. African crops rely even more on wild pollinators than do crops in other areas of the world because it can be difficult to maintain aggressive African honeybee hives and prevent them from being damaged by wild animals (African Pollinators Initiative, 2007).
Eggplant, papaya, coffee, and palm oil—crops of huge economic and cultural importance—highlight the value of wild pollinators to local and global economies. Eggplants are hermaphroditic; in other words, they can self-pollinate. Even so, pollination from two wild bee species, namely the doubleband carpenter bee (Xylocopa caffra) and a type of sweat bee (Lipotriches rufipes), increase fruit production far beyond that of self-pollination (Gemmil-Herren and Ochieng, 2008). In contrast, papaya trees are dioecious (i.e. they have separate male and female trees) and thus depend on crosspollination (i.e. pollinators take pollen grains from male flowers on one tree to female flowers on another tree) to produce fruit. While a wide variety of wild bees and butterflies visit papaya flowers, only some hawkmoths and skipper butterflies are effective papaya pollinators, probably because they have long proboscises that can penetrate the deep papaya flowers (African Pollinators Initiative, 2007). A healthy and diverse pollinator community also help coffee plants (which relies on a variety of pollinators, Samnegård et al., 2014) and oil palm (which requires cross pollination by specialist oil palm weevils, African Pollinators Initiative, 2007) produce more fruits, thereby increasing their economic value.
Despite their value to natural ecosystems and food security, wild pollinator populations are declining worldwide (Gallai et al., 2008; Dirzo et al., 2014). To avoid losing them forever, it is important to preserve wild pollinators through the conservation and restoration of native ecosystems (Chapter 10), sustainable agricultural practices, such as the reduced use of pesticides and herbicides (Section 14.1.1), and by communicating the value of pollinators to the general public, land managers, and politicians. Additionally, monitoring and research programs aimed at pollinators could enhance our understanding of their value, ecology, and conservation.
Many fruit and seed-bearing plants also depend on a process called seed dispersal to reproduce, colonise vacant habitats, and avoid competing with parent plants for limiting resources. Seed dispersal describes the physical movement of seeds by fruit-eating and seed-eating birds, large herbivores, primates, and a range of other animals away from the parent plant. Due to specialised features, some seeds can stick to animals’ fur, allowing them to be carried along much further distances than wind could, and different directions than water could. Many animals also consume seeds and fruits, providing opportunities for dispersal when the consumer moves off looking for more food, a resting spot, or mates to interact with. For some plants, seed dispersal involves a critical step required for germination, namely seed scarification. One method of scarification involves an animal breaking the seed’s hard coat by biting it. Alternatively, stomach acids may weaken the consumed seed’s hard coat while it passes through the animal’s digestive tract. Without this step, seeds requiring scarification may not be able to germinate; those plants’ persistence thus depends upon the animals that consume them. While the importance of pollination for food security is well known, the importance of seed dispersal should not be underestimated. A study from Côte d’Ivoire found that primates provided necessary seed dispersal services for at least 25 fruiting plant species important to humans (Koné et al., 2008).
Hazard detection and mitigation
When intact, nature is our first line of defence against many natural disasters. Consider, for example, the contribution of mangrove swamps in protecting us from cyclones/hurricanes (van Bochove et al., 2014), or the contribution of wetlands in flood control (Section 5.3.3). In contrast, degrading the natural environment can have severe consequences. For example, a 2010 landslide in Uganda that buried three villages, killing over 300 people and displacing 8,000 more, was attributed to deforestation activities three years earlier (Gorokhovich et al., 2013). To prevent such disasters, and harness all the other contributions of forests, there are numerous projects across Africa working to reverse deforestation (Section 10.3). Unfortunately, Africa’s tropical forests regenerate very slowly—sometimes requiring more than 100 years (Bonnell et al., 2011). It is thus critical to prevent ecosystem degradation in the first place, rather than having to resort to costly restoration projects.
In addition to keeping us safe, biodiversity can also be used to help track environmental changes. Species used for this purpose, called indicator species or environmental monitors are, by definition, associated with unique environmental conditions or sets of ecosystem processes. Tracking changes in their population sizes, distributions, and behavior of can thus serve as a substitute for expensive detection equipment (Section 10.1). Aquatic filter feeders, such as mussels and clams are particularly useful in this regard because their tissues accumulate chemical pollutants. A study from Senegal’s mangrove swamps detected heavy metal pollution using clams, mussels, and snails after tests barely detected those pollutants in the area’s sediments (Bodin et al., 2013). But even common everyday species can serve as indicator species: for example, conservation authorities around the world are using bird abundances and behaviors to better understand the impact of climate change (http://climatechange.birdlife.org).
Sentinel species are a special type of indicator species that can act as an early warning system for environmental hazards because they are more sensitive to certain conditions than humans are. Lichens are particularly well-known sentinel species. Being sensitive to air pollution and chemicals in rainwater, some lichens cannot survive in polluted areas. Thus, their presence is generally a sign of good air quality, while their absence may signal air pollution (Bako et al., 2008). Another example is seabirds, whose declining populations can serve as an early-warning system for overfishing (Paiva et al., 2015). Some sentinel species can even be used directly for human health purposes. For example, the non-profit NGO APOPO has been taking advantage of the incredibly fine sense of smell of southern giant pouched rats (Cricetomys ansorgei LC)—affectionately called HeroRATs—to detect landmines (Figure 4.6), tuberculosis (Reither et al., 2015), salmonella infections (Mahoney et al., 2014), and even people trapped under collapsed structures (LaLonde et al., 2015).
Lastly, some species can be used to mitigate various sources of pollution. For example, through a process called biosorption, the superior absorption capabilities of some lichens, plants, fungi, and microorganisms offer some of the cheapest and most effective methods for removing toxic heavy metals (Fosso-Kankeu and Mulaba-Bafubiandi, 2014) from the environment. Scientists also recently discovered a plastic-eating fungus (Khan et al., 2017) that may provide a potential solution to plastic pollution.
Pest and disease control
Every day, predators, such as owls and bats, keep us healthy by controlling populations of disease vectors, such as rats and mosquitoes. This process, where predatory (and parasitic) species regulate populations of pests and other nuisance species, is known as biological control, or biocontrol in short (Box 4.3). The use of insectivores (i.e. insect-eating species), such as bats and birds, to control crop pests is well established in traditional farming systems (Abate et al., 2000). But even on commercial crop farms, natural enemies, such as bats and birds, play an important role in keeping pests under control (Taylor et al., 2018). Some plants also play a part in biocontrol efforts: recent research found that some native plants used for intercropping in traditional agricultural systems emit chemical signals that kill and drive pest species away from crops (Khan et al., 2010). With an increasing number of studies illustrating the significant benefits gained from natural pest control systems, enhanced farming practices that facilitate greater ecosystem complexity (Section 14.1.1) will hopefully play a bigger role in food security in future.
Box 4.3 Biological Control Saves the Cassava Crop
Meg Boeni and Richard Primack
Biology Department,
Boston University,
Boston, MA, USA.
As it stands along the farm-plot boundary,
its base appears beautiful like a bride’s feet…
O cassava to whom the bembe drum beats a salute
that never reaches an end…
It is no small service the cassava renders us in this our land
Yoruba Poem (Babalola, 1966)
This traditional song from Nigeria praises the cassava, a South American crop brought to tropical Africa in the 16th century, and upon which millions of Africans have since relied for food and income.
Disaster struck in the 1970s, when an agricultural scientist that brought a new variety of cassava from South America to Africa also accidentally introduced a new pest: the cassava mealybug (Phenacoccus manihoti) (Neuenschwander, 2001). Previously unknown to science, the bug attacked the new shoots of cassava plants, laying its eggs at their tips and stripping them of their leaves. As it spread through Central and West Africa, the mealybug wiped out 80–90% of the productivity of most cassava fields, threatening large parts of Africa with famine.
With so many Africans relying on the cassava as a primary food source, scientists had to find a solution, and quickly. The bug’s waxy coating that protected it from pesticides complicated this effort. With conventional pest-control methods failing, scientists turned to biological control, hoping that introducing a natural predator would counteract the spread of the invasive insect. Researchers searching for the source of the mealybug finally found a candidate in the fields of Paraguay, where cassava, known locally as mandioca, was also an important food staple. Here, investigators discovered that mealybug numbers were kept low by a tiny wasp called Anagyrus lopezi that attack the mealybugs’ eggs and larvae (Figure 4.B). A. lopezi passed laboratory tests for host specificity—it fed and bred exclusively on cassava mealybugs and would not attack other African insects. And so, the International Institute of Tropical Agriculture (IITA) began field tests using the wasp as a biological control agent.
Results were astounding; the quick-spreading Paraguayan wasp reduced crop losses by an impressive 95% (Neuenschwander, 2001), all without the danger of pollution and poisoning associated with traditional pesticides. While identifying and introducing the biocontrol agent required significant resources, estimates suggest gains of 370–740 times the original investment, depending on the region considered (Zeddies et al., 2009), making it well worth the cost. Today, A. lopezi is found everywhere where the cassava mealybug survives in Africa. Bolstered by this success, the IITA has subsequently expanded its biological control programs to fight tropical pests on crops, such as cowpeas, maize, and bananas.
In 2008, the cassava mealybug was discovered in Southeast Asia, where it repeated the damage inflicted in Africa (Graziosi et al., 2016). Scientists are now replicating Africa’s biocontrol efforts to reduce crop failure in Vietnam, Thailand, Cambodia, and China. In conjunction with a number of local parasites, they hope that A. lopezi will halt the spread of the mealybug before it reaches even larger fields in India (Parsa et al., 2012). The control of the cassava mealybug is certainly one case where biological control was able to achieve great success.
Most Africans are familiar with scavengers, such as jackals and vultures, that work as nature’s clean-up crew, picking at food scraps left in the field by large predators. Together with the range of flesh-eating insects, detritivores, and decomposers, scavengers play a crucial role in keeping us healthy by sanitising the environment (O’Bryan et al., 2018). While it is all too easy to take these activities for granted, some people actively welcome these services. For example, in northern Ethiopia, spotted hyenas (Crocuta crocuta, LC) are tolerated in urban settlements because they consume livestock carcasses and sometimes even human corpses, which pose a disease risk (Yirga et al., 2015).
Recent experiences have shown that without proper care, the sanitary services provided by wildlife can collapse over a very short time. For example, during what is known as the Asian vulture crisis of the 1990s, vulture populations in India, Pakistan, and Nepal declined precipitously in a matter of years from secondary poisoning after eating carcasses of dead animals treated with the anti-inflammatory drug diclofenac. With nothing else available to remove carcasses of dead animals as efficiently as vultures, rotting flesh contaminated drinking water and allowed populations of rats and feral dogs (Canis familiaris) to proliferate. While vultures have stomach acids which kill pathogens, dogs and rats do not and thus became major pathogen vectors, spreading deadly diseases such as rabies, anthrax, and plague. The estimated healthcare costs in the face of Asia’s vulture crisis amounted to over US \$1 billion per year (Markandya et al., 2008). Today, Africa is facing its own vulture crisis. But instead of one threat, Africa’s vultures face a multitude of human-made threats, making solving this crisis much more complex (Box 4.4).
Scavengers such as vultures and jackals are nature’s clean-up crew; they keep us healthy by sanitising the environment.
Box 4.4 Conservation Lessons from the Asian and African Vulture Crises
Ara Monadjem
Department of Biological Sciences, University of Eswatini,
Kwaluseni, Eswatini.
[email protected]
A common perception among laypeople and conservationists alike is the idea of safety in numbers for wildlife species. After all, is a widely distributed and abundant species not safe from the threats of extinction? The answer is a firm no! As the collapse of central Asia’s vulture populations (Oaks et al., 2004) demonstrates, species numbering in the millions can disappear in the space of just a few years.
The Asian vulture crisis shares some similarities with the demise of the passenger pigeon (Ectopistes migratorius, EX) in North America a century ago. This pigeon was once the most abundant bird on Earth; yet, despite numbering in the billions, it was driven to extinction in a short span of time, primarily due to hunting over a 20-year period in the late 1800s. In the case of Asian vultures, the threat was not hunting, but rather a nonsteroidal anti-inflammatory drug (NSAID)—diclofenac—which is fed to sick cattle and then ingested by vultures when they feed on dead livestock. As diclofenac is deadly toxic to vultures, the widespread use of this treatment on the Indian subcontinent (which includes India, Nepal, and Pakistan) has seen catastrophic vulture population declines. With one of Asia’s major natural trash disposal systems gone, the area experienced a human health crisis from widespread drinking water contamination and increased incidence of diseases carried by ubiquitous and increasing rat and feral dog populations (Markandya et al., 2008).
The Asian vulture crisis is instructive on several grounds. First, it took a long time to detect and confirm the vulture declines because regular and standardised monitoring of the three affected vulture species had not been conducted. Second, the extent of the decline was extreme, with vulture numbers declining by over 95% within a decade. Third, the declines were due to a single threat—contamination from diclofenac, which were subsequently found to be deadly-poisonous to vultures (Oaks et al., 2004). Thanks to the concerted efforts of conservationists and politicians, and the rapid reactions of the governments of India, Pakistan, and Nepal, diclofenac was removed from the market in 2006. Vulture populations in Asia have since stabilised, with even a cautious suggestion of an increase.
Now, Africa faces its own vulture crisis (Ogada et al., 2015). However, in contrast to the Asian crisis, Africa’s crisis involves a greater number of species, and spans a larger geographical area. Importantly, it also includes a greater number of threats, including poisoning, harvesting for traditional medicine and for food, and electrocution following contact with power lines. Many vultures also die when they scavenge on poisoned carcasses meant to kill problem predators (Figure 4.C). To this list of lethal causes, one should also add the universal threats of habitat loss and persecution of birds of prey.
Thanks to long-term monitoring, the collapse of African vulture populations has been well documented. Of the 95 vulture populations being monitored, 89% were either extirpated or experienced severe declines. Across eight study species, the mean rate of decline is estimated at 4.6% per year (i.e. one out of 20 birds that are dying per year are not being replaced). The charismatic Rüppell’s vulture (Gyps rueppellii, CR) has declined by 85% across its range; consequently, this species is now considered highly threatened by the IUCN, as are the hooded vulture (Necrosyrtes monachus, CR), white-headed vulture (Trigonoceps occipitalis, CR), and African white-backed vulture (Gyps africanus, CR). Only slightly better off, at least for now, are the lappet-faced vulture (Torgos tracheliotos, EN) and Egyptian vulture (Neophron percnopterus, EN), the Cape vulture (Gyps coprotheres, VU), and the bearded vulture (Gypaetus barbatus, NT).
The collapse of Africa’s vulture populations is cause for serious concern among conservation biologists, wildlife and livestock managers, and public health officials. Unlike in Asia, however, workable solutions to Africa’s vulture crisis have not yet been found. This may be due to the multitude of threats, and the complexity of the problem exacerbated by the involvement of individual poachers, local communities, and government structures across more than 40 countries. If conservationists and governments can work together, as they did in Asia, then perhaps Africa’s vultures and the ecosystem services that they provide can still be saved. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/04%3A_Why_Should_We_Protect_Biodiversity/4.02%3A_Regulating_Services.txt |
Nonmaterial contributions from nature, also called cultural services, include the subjective and psychological aspects of nature that influence our perceptions about quality of life. These contributions can be divided into three subcategories: inspiration and learning support, supporting psychological and physical experiences, and supporting individual and group identities.
Inspiration and learning support
Nature has inspired artists and writers throughout history. Consequently, many books, television programmes, movies, and websites produced for entertainment purposes are based on natural themes. This infusion of nature into popular culture is worth billions of dollars per year. To take one example, the 1994 Disney blockbuster The Lion King, based on the lives of a variety of African savannah animals, generated revenues estimated at just under US \$1 billion from theatre attendances alone. It was so successful that three movie sequels, an animated television series, and several video games and books followed. A musical based on The Lion King movie plot continues to be a top-earning title in box-office history for both stage productions and films.
Movies featuring stunning natural landscapes and charismatic wildlife often increase the desire of moviegoers to visit natural areas where they can see these landscapes and animals first-hand. But it can also raise awareness of environmental issues in new audiences. While many documentaries are created with this purpose in mind, such benefits can also extend to blockbuster movies meant for broader audiences (Silk et al., 2018). For example, Disney’s Happy Feet (2006) highlighted the threat of overfishing and plastic pollution to penguins; Avatar (2009) raised awareness of habitat loss and overharvesting; and The Jungle Book (2016) exposed audiences to the plight of pangolins. Such exposure can even lead to environmentally conscious behavioural changes. For example, moviegoers were willing to donate 50% more money to climate mitigation after watching the apocalyptic movie The Day After Tomorrow (2004) (Balmford et al., 2004). Perhaps, in part, due to the influence of environmentally-orientated movies, an increasing number of movie stars (and other celebrities) have started using their stardom as a platform from where they promote biodiversity conservation efforts in Africa (Duthie et al. 2017; see also https://wildfor.life).
Scientists and engineers often seek inspiration from nature for new technologies or to solve innovation challenges.
Scientists and engineers also sometimes turn to nature to seek inspiration for new technologies or to solve innovation challenges. For example, the water-vapor collecting capacity of the racing stripe darkling beetle (Stenocara gracilipes) from the Namib Desert in Namibia (Figure 4.7) inspired engineers who developed self-filling water bottles (Clark, 2012), irrigation systems to overcome drought conditions (Scott, 2011), fog-free windows and mirrors (Parker and Lawrence, 2011), and methods for controlling condensation and frost on aircraft surfaces (Boreyko et al., 2016). While these and other scientific endeavours, collectively known as biomimicry, provide many social and economic benefits, their primary value comes from new knowledge, improved education, and enriched human experiences.
Supporting psychological and physical experiences
While the economic benefits gained from nature incentivises biodiversity protection, many people believe that the aesthetic values of nature provide an even greater incentive for conservation. This principle rests on the fact that nearly everyone enjoys wildlife and landscapes aesthetically. Even city-dwellers who are superficially removed from nature find a sense of relief and well-being when they have opportunities to come in close contact with the natural world. But what if dragonflies and butterflies disappeared? What if our favourite sports team’s mascot ceased to exist in nature? What if there were no more forests filled with bird flocks or monkey troops?
The intangible but desirable aesthetic values people attach to certain aspects of nature are known as amenity values. Amenity values are becoming increasingly important in many local and national economies throughout Africa, in the form ecotourism. At any one time, there are millions of tourists traveling and spending money across Africa to see particular species or to experience unique ecosystems. This includes scuba divers approaching a coral reef (Figure 4.8), birdwatchers visiting a rare species’ stakeout, and people on a safari to view the many flagship species for which Africa’s savannahs are so well known. Ecotourism has long been a major industry in southern and East Africa. For example, ecotourism generated over US \$1 billion in annual revenue in the Cape Floristic Region more than a decade ago (Turpie et al., 2003), and has accounted for over 15% of Kenya’s gross domestic product (GDP) at times (WWF and BSI, 2006). Ecotourism is also becoming increasingly important in other parts of Africa. For example, since overcoming periods of social unrest, Burundi, Rwanda, and Uganda have created profitable local industries charging tourists high fees to visit habituated populations of mountain gorillas (Gorilla beringei beringei, EN). Also, in South Africa, some bird guides earn an average of US \$362 per month by showing tourists the unique birds their local area has to offer (Biggs et al., 2011).
In recent years, volunteer-based ecotourism has emerged as a lucrative industry that combines ecotourism with learning opportunities. These organisations offer aspiring conservationists and citizen scientists hands-on experience while bringing financial and other logistical support to rural and protected areas. Many wildlife sanctuaries and conservation NGOs also offer volunteer opportunities and field courses that combine conservation action with local community outreach and education programmes. The research done by professional scientists and citizen science volunteers can be used in locally-relevant educational materials. Biological field stations (Section 13.1.5) often host these activities; the stations can also provide training and jobs for local community members.
The revenue and jobs generated by ecotourism provides a strong and immediate justification to protect biodiversity and restore areas that have been degraded.
The revenue and jobs generated by ecotourism provides a strong and immediate justification to protect areas rich with biodiversity or to restore areas that have been degraded. Ecotourism can even be integrated directly in plans for future development, protection, and restoration. One such example is integrated conservation and development projects (ICDPs, Section 14.3), which provide models for how empowered rural communities can successfully establish accommodation, develop expertise in nature guiding, and sell local handicrafts at curio stores to obtain multiple stable income streams. The revenue obtained from ecotourism also allows local people to move away from unsustainable hunting, fishing, or grazing practices towards lifestyles that can be maintained in the long term.
Still, many of Africa’s ecotourism resources remain under-utilised. To use one example, only a few locations in Africa cater to people who enjoy the thrill of swimming with sharks in their natural habitat. Beyond removing fear and instilling a healthy respect for sharks, this industry also plays an important role in conservation by showing how living sharks bring greater economic benefit than a once-off catch. For example, shark diving at just one location in South Africa is estimated at US \$4.4 million annually (Hara et al., 2003); similar industries in the Maldives (Cagua et al., 2014) and Palau (Vianna et al., 2012) generate even more revenue. Presenting unique recreational experiences and a growing global ecotourism sector, more and more African countries will hopefully explore these and other opportunities soon. It is worth noting that the long-term effects of shark diving operations are largely unknown, particularly as it relates to possible behavioural changes from using bait to attract sharks to people, and an active area of current research (Gallagher and Huveneers, 2018).
Although ecotourism can provide many valuable conservation and economic benefits (Thiel et al., 2014), care must always be taken that these activities abide by accepted ethical standards (Hayward et al., 2012). It is also worth remembering that wildlife ecotourism is often geared towards wealthy western markets, making it prohibitively costly to the people who live near facilities, and are most vulnerable to factors such as human-wildlife conflict. As such, it is important to consider what portion of the generated funds are invested locally versus reserved to enrich well-compensated shareholders in the far-away capital. Are local people given opportunities to further their training and education, and to advance their careers within ecotourism organisations? Unfortunately, in many areas of Africa, local people continue to receive only the smallest percentage of money spent by tourists. Similarly, even though national parks themselves may receive large numbers of foreign visitors, governments continue to use only a small percentage of the generated funds on park management (Lindsey et al., 2014; Balmford et al., 2015).
Supporting individual and group identities
Many people care deeply about biodiversity. The thought of a charismatic animal or a special landscape (Figure 4.9) may elicit a strong emotional response, which leads to a desire to protect plants, animals, and natural places. For some people, this desire is associated with a hope to someday see those unique species or landscapes in person. Others do not expect or even desire to see these species and landscape themselves, yet they value their existence. In either case, these individuals recognize the existence values of wildlife and nature—the benefit people receive from simply knowing that an ecosystem or species exists. Bequest values (also known as beneficiary values) is a component of existence values, defined as the perceived benefit people receive from preserving a natural resource or species for future generations.
The desire to ensure the protection of biodiversity has prompted a wide range of people to establish, join, or otherwise contribute to conservation organisations. For many people involved in these organisations, their participation stems from the ethical premise that wildlife are equal to human life, and that biodiversity conservation offers genuine and long-lasting well-being. This environmental philosophy is often described as deep ecology, the ethical premise that species and biodiversity have a right to exist independent of their possible benefits to humans, and that humans have an inherent responsibility to protect species and biodiversity (see also Section 1.4). Deep ecology holds that social structures (including politics, economics, technology, and ideology) must change radically to reduce the destruction of Earth’s biodiversity and to enhance people’s quality of life. It emphasises and prioritises the natural environment, aesthetics, religion, and culture, rather than material consumption. Although the ethical appreciation of biodiversity is similar in deep ecology and conservation biology, deep ecology includes broader goals for personal, social, and political change.
Biodiversity forms the basis of spiritual, celebratory, and other social-cohesion experiences for many people.
Biodiversity also forms the basis of spiritual, celebratory, and other social-cohesion experiences for many people. It ensures people experience a sense of place and belonging, reminds them of childhood experiences, and gives a sense of connection when they experience natural sights, sounds and smells. This is especially true for Africans, many of whom attach deeply-held spiritual, cultural, and symbolic values to the environment. Even the money of most, perhaps all, African countries features aspects of nature, as if to add a little extra (if only symbolic) value to those coins and bills. All these factors play a major role in people’s sense of who we are—our identity.
4.04: Option Values
The option values of biodiversity describe nature’s potential to provide currently unknown or unrealised benefits at some point in the future. For example, while many species may not currently have any realised material contributions, a small number of taxa may have enormous potential to support new industries or prevent major agricultural crops from collapsing. For this reason, scientists continuously search for species with hidden uses: entomologists search for insects that can control pest species, microbiologists search for bacteria useful in biochemical manufacturing, and agricultural scientists search for genetic varieties of plants that can produce more food to feed a growing human population. As fears of antibiotic resistance become reality, archaea (widespread single-celled microorganisms with no nucleus which are also thought to be the oldest life forms on Earth) may be used to develop new classes of antibiotic medicine (Metcalf et al., 2014). Some researchers also hope that studying primates—the likely original source of diseases such as HIV/AIDS, Ebola, and malaria (Martin et al., 2005)—may allow us to one day find cures for these diseases. It is worth noting that the effectiveness of using animals to study human diseases remains highly controversial (Archibald and Clotworthy, 2007; Festing and Wilkinson, 2007; Rollin, 2007), and that many people believe that the suffering and death of animals during biomedical research is unethical.
This continuous search for valuable or useful natural products, called bioprospecting, has already contributed a great amount to global economic development, and is expected to become even more important in the coming decades. This is particularly true in the rush to find replacements for climate sensitive crops that may be threatened by climate change. For example, researchers hope that the genetic diversity in wild coffee populations can act as an insurance policy in case our warming planet damages currently popular commercial strains (Davis et al., 2012). There is also much hope that plants from Africa will lead to new medical treatments, for diseases such as malaria, cancer, and high blood pressure (Gurib-Fakim, 2017). It is for reasons such as these that losing even small portions of expansive ecosystems concerns scientists. The extinction of even one valuable species or gene can represent a tremendous loss to humanity, even if many other species are preserved.
Bioprospecting, the search for valuable or useful natural products, has already contributed greatly to global economic development. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/04%3A_Why_Should_We_Protect_Biodiversity/4.03%3A_Nonmaterial_Contributions.txt |
It should be clear from reading this chapter that the well-being of people around the world is fundamentally linked to opportunities for biodiversity to survive and prosper. That means that when we destroy an ecosystem or let a species go extinct, we also put at risk our own ability to survive and prosper. To fully account for these risks, decisions that negatively affect biodiversity must account for all the costs and benefits (hidden or otherwise), including the impacts on ecosystem services, before the decision is implemented.
One of the most popular methods for accounting for potential harm to biodiversity, especially when weighing public policy and commercial decisions, is to attach market (or monetary) values to the ecosystem services. For some ecosystem services, it is rather straightforward to estimate a market value. For example, how much would it cost to replace a natural pollination service with hand pollination by farm workers? But for many services, estimating a monetary value is much more difficult. For example, how do we calculate the value of the Congo Basin’s carbon stocks? Where do we even start to estimate the value of breathing clean air, or knowing that dolphins exists?
Environmental economics strengthen arguments for biodiversity protection by examining the contribution of ecosystem services to global economies.
To examine these kinds of complex questions, conservation biologists look to a sub-discipline within economics called environmental economics. Environmental economics broadly examines the contribution of ecosystem services to global economies. An important component of this examination involves estimating the market value of all the different ecosystem services we benefit from, but it also includes studying the environmental costs of economic transactions, environmental policies, and other decisions that impact the environment.
4.5.1 Placing a price on the natural world
Approximating the market values of ecosystem services is no small feat, in part because nature’s contributions to people vary by location and perspective (Díaz et al., 2018). There also continue to be technical (e.g. Kling et al., 2012) and ethical (e.g. McCauley, 2006; Silvertown, 2015) disagreements about the need and methods used to translate nature’s services into monetary terms. Nevertheless, including such estimates has become a widely accepted norm in economics models and conservation activities (Guerry et al., 2015). To accomplish this task, economists rely on seven main methods to estimate the market values of ecosystem services (Farber et al., 2002):
• Market value: The price a person is willing to pay for a specific product or service. For example, how much is a person willing to pay for a bundle of firewood at a local market?
• Avoidance cost: The cost society avoids paying because a specific ecosystem service exists. For example, how much does society avoid paying for water filtration service otherwise provided by a region’s forests and wetlands?
• Replacement cost: The cost society would have incurred if a specific ecosystem service had to be replaced. For example, how much would society have to pay in extra healthcare costs and in clean-up costs for diseased carcasses to replace the sanitation services provided by vultures?
• Factor income: The additional income generated by the enhancement of an ecosystem service. For example, how much would a reduction in water pollution increase the income of fishermen through healthier fish populations?
• Travel cost: The additional travel cost a person is willing to pay to experience an ecosystem service otherwise not available to them. For example, how much extra is a person willing to pay for transport to participate in recreational activities at a cleaner lake?
• Hedonic pricing: The additional expense a person is willing to pay to experience an ecosystem service. For example, how much extra is someone willing to pay for a house with an ocean view, compared to an inland house?
• Contingent valuation: The additional expenses a person is willing to pay for an alternative hypothetical scenario. For example, how much is someone willing to pay for cleaner air, or the right to catch more fish?
The combined value of all of Earth’s ecosystem services may be double the current value of the global economy.
Using a combination of these methods, a range of ecosystem services have been valued in recent years. For example, the services offered by pollinating insects around the world have been valued at US \$153 billion per year (Gallai et al., 2008). In just South Africa’s Western Cape Province, free pollination services provided by wild insects to the local fruit industry, valued at US \$500 million, has been estimated at nearly US \$360 million per year (Allsopp et al., 2008). The replacement cost of tropical forests is also increasingly appreciated in carbon sequestration markets, where heavy greenhouse gas emitters pay huge sums of money to conserve forests to become more carbon neutral (Section 10.4). For example, the United Nations Environmental Programme (UNEP) has estimated that their forests are worth 4.2 times more intact than the value that could be earned through logging; the value of just Kenya’s remaining Mau forest, if left intact, is estimated at US \$1.3 billion per year (UNEP, 2012). One ambitious study estimated the value of all of Earth’s ecosystem services at US \$145 trillion annually (Costanza et al., 2014), which is almost double the current US \$78 trillion value of the global economy. By comparing the value of ecosystem services over time, Costanza et al. (2014) also estimated that we are losing US \$4.3–20.2 trillion per year in ecosystem services through land degradation.
4.5.2 Environmental economics’ biggest contributions
Since its development, environmental economics has contributed to conservation biology in several very important ways. Perhaps the most important contribution is that it has enabled conservation biologists to better communicate the value of ecosystem goods and services to audiences like government officials and business leaders, who often base decisions on economic considerations. By doing this, environmental economics has also focussed our attention on the wide range of goods and services that biodiversity provides and has elevated these topics into corridors where they were not previously discussed. These efforts have already paid dividends; in 2012, several Africa countries signed the Gaborone Declaration, a pledge to integrate the value of ecosystem services into their economies.
Environmental economics enable us to better communicate the value of biodiversity to those who base decisions on economic considerations.
Environmental economics also enabled conservation biologists to better account for environmental impacts of environmental damaging activities. In doing so, the field highlighted how activities that appear profitable are running at an economic loss when properly accounting for otherwise ignored environmental (and social) damages. While such calculations have traditionally focussed on imbalances in overharvesting of material contributions (see negative externalities, Section 4.5.3), recent developments have also started accounting for damages inflicted on regulating services and nonmaterial contributions, such as the loss of nature’s contribution to climate regulation (Auffhammer et al., 2017; Hsiang et al., 2017).
4.5.3 Environmental economics’ biggest challenges
Despite all the direct and indirect contributions of environmental economics to biodiversity conservation, there are also several challenges facing the field. Some of these challenges relate to methodological complexities of valuing ecosystem services, but many challenges also have their roots in governance failures. Following is a discussion of the most important challenges facing environmental economics.
Accounting for negative externalities
Modern economics is built on the principle of voluntary transactions—that is, a transaction occurs only when it benefits all the stakeholders involved. However, environmental (and social) harm often arises when some hidden costs are passed on to people not directly involved in the transactions. The unregulated use of open-access resources—resources such as water, air, and fish populations that are freely used by many different groups of people—provide many opportunities for this kind of abuse. Consider a company that dumps chemical waste into a river instead of properly disposing of it. While the company may benefit from this cost-cutting measure, people further downstream bear the environmental and social costs of the company’s “free” waste disposal by having to contend with polluted drinking water, loss of swimming and other recreational opportunities, and loss of fish as a safe food source. Damage inflicted on rivers and other open-access resources also represent a classic example of the tragedy of the commons—while some people initially benefit from abusing the “free” ecosystem services, those values are gradually lost to all of society, including those who abused it (NRC, 2002).
Because negative externalities allow a small number of people to benefit at the expense of the rest of society, they often lead to market failures.
The hidden costs of economic transactions that are passed on to people not directly involved are generally known as negative externalities (Figure 4.10). Because negative externalities allow a small number of people to benefit at the expense of the rest of society, they often lead to market failures, characterized by transactions that do not lead to optimal outcomes for all stakeholders. Governments may correct for these kinds of market failures by imposing taxes on activities that are harmful to the environment. Carbon taxes imposed on greenhouse gas emitters (see climate change, Chapter 6) is a common example. But many times, governance structures fail, or even exacerbate, the impact of negative externalities, by artificially maintaining destructive activities with tax incentives, direct payments, and price regulations. For example, subsidies give foreign fishing fleets operating off Africa a competitive advantage over local fisherman and artificially inflate their profitability despite declining fish populations (Brashares et al., 2004; Sumaila and Pauly, 2006; Mallory, 2013). The financial incentives governments provide to maintain destructive activities are more often referred to as perverse subsidies (Myers and Kent, 2001). The size of perverse subsidies is often very large, regularly dwarfing conservation spending. For example, US \$26 billion in subsidies were provided to the Africa’s fossil fuel industry just in 2015 (Whitley and van der Burg, 2015), compared to just US \$381 million spent annually to secure Africa’s protected areas with lions (Lindsey et al., 2018).
There are many reasons why governance structures continue to fail nature and allow market failures to occur. For example, due to the prevailing mindset of pursuing economic growth at all costs, politicians, developers, and industries often skew their cost-benefit analyses by prioritising the short-term benefits gained from destructive sectors over long-term societal well-being and sustainability. Another factor is intense lobbying by industries benefitting from perverse subsidies, which leads to corruption and other questionable decisions. Solving these challenges will rely on a society that prioritises economic development (Section 15.1) and establishes structures (i.e. passing and enforcing environmental laws, Chapter 12) that fully account for negative externalities.
Determining ownership
Another problem that plagues environmental economists and other stakeholders is deciding who owns the commercial rights to biodiversity. Imagine a biochemist from a wealthy country traveling in a rural part of West Africa. The biochemist falls sick, but luckily local villagers help the chemist get better with the aid of a traditional healing plant. Once back home, the biochemist scientifically demonstrates that this plant can be used to synthesise a new effective medicine. Do the profits from this new medicine belong to the biochemist, the organization that sponsored his/her trip, or the local people in the area who helped the biochemist?
Scientists, economists, politicians, and others are currently debating who owns the commercial rights to the world’s biodiversity.
In the past, corporations and scientists (generally from wealthier countries) travelled extensively (often to poorer countries in the tropics) to collect species from which commercially valuable products might be obtained. These new products were then sold, but all profits were kept by the corporations while the people in the poorer source countries received little to no financial benefit. One such example is the production of palm oil, of which Malaysia and Indonesia currently contribute 85% of the global vegetable oil supply. This industry is entirely dependent on the oil palm (Elaeis guineensis, LC), and its specialist pollinator, the oil palm weevil (Elaeidobius kamerunicus), both imported from West Africa. Yet, West Africa have seen little benefit from the profits palm oil generated in Southeast Asia (Mbugua, 2017). (Note this exploitation goes multiple ways; for example, South America has also seen little benefit from profits generated from cacao production in West Africa.)
To combat this unfair exploitation, called biopiracy, many developing countries now require scientists and corporations to obtain permits before they can collect biological material for commercial or research purposes. Also, at the international level, nearly 100 countries have agreed to the fair sharing of benefits arising from the use of biological resources, through the Nagoya Protocol (see Section 12.2.1 for further discussion on international laws). Through these and similar laws and agreements, the hope is that a greater portion of the profits gained from biodiversity will be allocated to people who protect biodiversity and who live in the areas from where it is extracted.
A more inclusive approach
The valuation of ecosystem services has traditionally relied on generalised principles of economics and natural sciences. While this focus enabled scientists to develop broadly applicable themes and metrics in ecosystem evaluation, it also neglected the role of context and culture in understanding nature’s role in people’s lives. Many people have also remained uneasy about commodifying nature (i.e. giving it a market value), because some of the most important contributions of biodiversity are not easily converted into monetary metrics. Consequently, many feared that the transactional approach to ecosystem services would lead to social inequity concerns and alienate people offended by the idea that nature’s metaphysical properties must compete against commercial interests.
To address these concerns, the valuation and classification of ecosystem services are currently undergoing several major transformations. Prominently, the UN’s most recent classification scheme (Díaz et al., 2018) has given a more prominent voice to a wider range of stakeholders, including the social sciences, and recognizes the importance of culture and context in nature’s contributions to people. This exciting area of research is actively developing, and readers are encouraged to track developments and reactions associated with 2019 IPBES Global Assessment at https://www.ipbes.net/news/ipbes-global-assessment-preview. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/04%3A_Why_Should_We_Protect_Biodiversity/4.05%3A_Environmental_Economics.txt |
1. People value biodiversity in many ways. The reasons vary from person to person, and from region to region. But generally, nature’s contributions to people, also called ecosystem services, are divided into three overlapping and interdependent categories, namely material contributions, regulating services, and nonmaterial contributions.
2. Material contributions include benefits people get from consuming natural resources (e.g. drinking water or burning wood for cooking) or using natural resources in production and trade (e.g. timber to build homes or other structures).
3. Biodiversity also provides a large variety of regulating services that enable people to benefit from nature’s material contributions. Some of these contributions include ecosystem productivity, water and soil protection, climate regulation, pollination, seed dispersal, and disaster prevention and detection.
4. People attach nonmaterial values to biodiversity which are difficult to quantify, and thus to account for, in modern economic systems. These values include support for inspiration and learning, support for psychological and physical experiences, and support for personal and group identities.
5. Environmental economics studies the implications of economic transactions, environmental policies, and other decisions that impact the environment. This field has highlighted how damage to the environment, such as pollution caused by industry, are not always fully considered when making political and development decisions, leading to unsustainable economic practices and market failure. Accounting for negative externalities and perverse subsidies can help policymakers design incentives that promote sustainable practices.
4.07: Topics for Discussion
1. Think of a recent infrastructure development near where you live, such as a recently-built road or dam. Try to come up with a list of ecosystem services that were damaged by this development. Who carries the costs of these lost services? Do you think the benefits from the development were worth the costs? Explain your answer.
2. Do individual organisms, populations, species, and biological communities have rights? What about physical features such as lakes, rivers, and mountains? While explaining your answer, also think about where we should draw the line of moral responsibility in how we care for nature.
3. A European botanist on holiday visits your area. During a short hike, you show this botanist a plant used as a traditional treatment for malaria. The botanist takes samples of this plant back to Europe, where subsequent testing shows that it can be used to develop an effective anti-malarial drug. Who do you think should receive the profits from this new drug? The botanist who undertook the trip, and you because you showed the botanist the plant? What about the organization that funded the drug’s development, and the scientists who synthesised the new drug? What about all the people who educated you and your family in the plant’s value? If the profits belong to multiple entities, how should it be divided?
4. More than a decade ago, the shark ecotourism industry at Gansbaai, South Africa, was estimated at US \$4.4 million annually (Hara et al., 2003)—it has been increasing ever since. There are an estimated 900 great white sharks (Carcharodon carcharias, VU) living in Gansbaai (Towner et al., 2013). Assuming the average white shark lives for 70 years (Hamady et al., 2014), what is the value of each shark at Gansbaai? Can you find (or estimate) the price that a single shark sold for food would obtain on the world market? How do these values compare? What do you think is the best use of the sharks?
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Turpie, J.K., B.J. Heydenrych, and S.J. Lamberth. 2003. Economic value of terrestrial and marine biodiversity in the Cape Floristic Region: Implications for defining effective and socially optimal conservation strategies. Biological Conservation 112: 233–51. https://doi.org/10.1016/S0006-3207(02)00398-1
UNEP. 2012. The role and contribution of montane forests and related ecosystem services to the Kenyan economy (Nairobi: UNEP). http://wedocs.unep.org/handle/20.500.11822/8513
Valeix, M., H. Fritz, R. Sabatier, et al. 2011. Elephant-induced structural changes in the vegetation and habitat selection by large herbivores in an African savanna. Biological Conservation 144: 902–12. http://doi.org/10.1016/j.biocon.2010.10.029
van Bochove, J., E. Sullivan, and T. Nakamura. 2014. The Importance of Mangroves to People: A Call to Action (Cambridge: UNEP). http://wedocs.unep.org/handle/20.500.11822/9300
van der Plas, F., R. Howison, J. Reinders, et al. 2013. Functional traits of trees on and off termite mounds: understanding the origin of biotically‐driven heterogeneity in savannas. Journal of Vegetation Science 24: 227–38. https://doi.org/10.1111/j.1654-1103.2012.01459.x
Vianna, G.M.S., M.G. Meekan, D.J. Pannell, et al. 2012. Socio-economic value and community benefits from shark-diving tourism in Palau: A sustainable use of reef shark populations. Biological Conservation 145: 267–77. https://doi.org/10.1016/j.biocon.2011.11.022
Whitley S., and L. van der Burg, 2015. Fossil fuel subsidy reform in sub-Saharan Africa: from rhetoric to reality (London and Washington: New Climate Economy). https://newclimateeconomy.report/workingpapers/workingpaper/fossil-fuel-subsidy-reform-in-sub-saharan-africa-from-rhetoric-to-reality-2
Wicander, S., and L.M. Coad. 2015. Learning our Lessons: A Review of Alternative Livelihood Projects in Central Africa (Gland: IUCN). https://www.iucn.org/content/learning-our-lessons-review-alternative-livelihood-projects-central-africa
WWF, and BSI. 2006. Poverty and environmental issues: Governance institutions, institutional frameworks and opportunities for communities (Nairobi: UNEP). https://www.unpei.org/sites/default/files/e_library_documents/kenya-poverty-environment-issues.pdf
Yirga, G., H. Leirs, H.H. de Iongh, et al. 2015. Spotted hyena (Crocuta crocuta) concentrate around urban waste dumps across Tigray, northern Ethiopia. Wildlife Research 42: 563–69. https://doi.org/10.1071/WR14228
Zeddies, J., R.P. Schaab, P. Neuenschwander, et al. 2001. Economics of biological control of cassava mealybug in Africa. Agricultural Economics 24: 209–19. https://doi.org/10.1111/j.1574-0862.2001.tb00024.x | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/04%3A_Why_Should_We_Protect_Biodiversity/4.08%3A_Bibliography.txt |
The right to shelter, food, and association are basic human needs recognized in many international charters and country constitutions. Like humans, wildlife also needs areas where they can find protection, nourishment, and mates to have any hope of survival. The area where a species can survive and meet their basic needs is known as its habitat. It is often useful to think of a habitat as a multi-dimensional space, characterized by suitable levels of many different environmental variables. Some species, including humans, are highly tolerant of changes in their environmental conditions; consequently, such generalist species find it relatively easy to move to a new area in the unfortunate event that their “home” is destroyed. In contrast, specialist species—those that can only survive within a narrow range of environmental conditions—often do not have anywhere else to go when their habitat is lost, and consequently they go extinct.
In a world where intact natural ecosystems are increasingly being altered by the activities of an ever-increasing human population and its consumptive needs, habitat loss has emerged as the number one threat facing biodiversity today. The expansion of human activity causes massive disturbances to natural ecosystems by altering, degrading, and outright destroying wildlife habitats. A number of specialist species have already been pushed to extinction. But even generalist species are increasingly falling victim to habitat loss: pushed out of their shrinking habitats, they come into conflict with humans while trying to meet their needs near urban centres and on agricultural land. Eventually our own lives will suffer, whether through lost ecosystem services, or sorrow for all the wonderful landscapes and species that have disappeared under our watch. In this chapter we delve into the causes and consequences of this increased competition for space between man and wildlife.
The primary threat to Africa’s biodiversity today is habitat loss and degradation.
5.01: What is Habitat Loss
Habitat loss is defined as the outright destruction of natural ecosystems, an inevitable consequence of expanding human populations and human activities. The theory of island biogeography (MacArthur and Wilson, 1967) offers a good explanation for why habitat loss drives species extinctions. Using oceanic islands as a model system, one of the theory’s main predictions is that large islands have more species than small ones because they can accommodate more individuals, which causes those species to be better buffered against extinctions (Section 8.7). Empirical evidence offers strong support for this observation, also known as the species-area relationship. For example, large African islands generally hold more bird species than small islands (Figure 5.1). In addition, 62 of the 79 (63%) Sub-Saharan Africa’s species that went extinct over the past few centuries (IUCN, 2019) have been confined to oceanic islands, rather than the continental mainland which in effect functions like one very big island.
The species-area relationship underpins much of conservation biology today. By applying the relationship’s principles to “islands” of suitable habitat surrounded by a “sea” of damaged or unsuitable habitat (the “matrix”), conservation biologists know that conserving large areas of suitable habitat is much more effective in protecting biodiversity (Box 5.1). This is especially true when trying to protect species that have large home ranges, and/or occur in low densities: they can only live in habitat patches that are large enough to maintain viable populations (Chapter 9 discusses the relationship between population size and population viability in more detail). Observations of extirpations in differently sized habitat patches support this application. For example, researchers have found that nearly 50% of Ghana’s forest bird species are sensitive to habitat size, with 25% of species never found in forest patches smaller than 0.1 km2 (Beier et al., 2002). One Ghanaian species that seems particularly sensitive to habitat patch size is the icterine greenbul (>Phyllastrephus icterinus, LC); due to habitat loss, this once-common species decreased by 90% during one study’s 15-year period (Arcilla et al., 2015).
Box 5.1 The Importance of Liberia’s Forest Network to the Survival of the Pygmy Hippopotamus
Mary Molokwu-Odozi1 and Kathryn Phillips2
1Fauna & Flora International,
Harmon Compound, Congo Town,
Monrovia, Liberia.
2Fauna & Flora International,
Cambridge,> UK.
Remaining populations of the pygmy hippopotamus (Choeropsis liberiensis, EN) are found predominately within transboundary West African rainforests spanning Côte d’Ivoire, Guinea, Liberia, and Sierra Leone (Ransom et al., 2015). Liberia contains the largest intact blocks (over 40%) of this Upper Guinean rainforest, a Global 200 ecoregion (Olson et al., 2002). An elusive animal, little is known about the pygmy hippo’s distribution, population status, and ecology. Pygmy hippo numbers are currently estimated at fewer than 2,500 individuals across its range, with the expectation of further decline as a result of agricultural expansion, logging, development, and hunting (Ransom et al., 2015).
Within Liberia, pygmy hippopotamus populations are found in the major forest blocks of the southeast and northwest (Figure 5.A), which are separated by an area of degraded land with high human densities (FFI and FDA, 2013). The southeastern forest block is made up of several large chunks of national, communal, and protected forests fragmented by logging routes and concessions. Although populations are well documented within protected areas, recent reports indicate that populations also exist outside formally protected forests (Hillers et al., 2017). The establishment and management of forest corridors linking key habitats is therefore a conservation priority.
In the last few years, huge investment in agriculture, logging, and mining has increased pressure on forests for conversion and from increased human settlements and access roads. Weak law enforcement in Liberia’s protected areas and limited operational capacity has led to increased incursion of illegal activities, such as poaching and mining in these critical habitats. Sapo National Park, Liberia’s only national park and second largest in West Africa after Taï National Park in Côte d’Ivoire, is believed to be a stronghold for the species. However, pygmy hippo numbers remain low at an average encounter rate of 0.12 individuals/km from 2007–2009 (Vogt, 2011) to 0.15 individuals/km (2014–2016 data), i.e. one per 7–8 km, much lower than records from Taï (Vogt, 2011). Sapo National Park has historically suffered from—and continues to suffer from—mining and hunting pressures; hundreds of illegal miners who were evacuated after the civil crises of 2002–2007 and in 2010–2011 and again during the 2014-2015 Ebola crisis, reoccupied a large section of the park, where hunting signs (trails, camps, gun shells) were encountered almost every kilometre of walking within the park’s 1,804 km2 area. The miners have once again been removed from the park, this time with local community support.
Recent efforts to save Liberia’s declining forests have brought about an increase in activities by national and international NGOs and the formation of transboundary collaboration initiatives. For example, an agreement has been established between the Liberian and Sierra Leonean governments, creating the Gola Transboundary Peace Park and significant progress has been made in the development of the Taï-Grebo-Krahn-Sapo Transboundary Forest Complex with Côte d’Ivoire. The Liberian and Guinean government have also commenced a bilateral agreement for the conservation and sustainable management of the Ziama-Wonegizi-Wologizi Transboundary Forest Landscape. Agricultural investments have also evolved to promote public-private partnerships in green-growth and community-based forest protection initiatives. Notable support has also come from the Government of Norway to help Liberia fully halt deforestation by 2020.
Fauna & Flora International (FFI) has worked in Liberia since 1997 focusing on the pygmy hippopotamus as a flagship species. These efforts have contributed to increased knowledge of the species in Liberia, including recording (in collaboration with the Zoological Society of London, ZSL) the first footage of the species in Liberia. FFI also developed a pygmy hippopotamus national action plan and will be revising the regional conservation strategy for the pygmy hippo in 2019. FFI has also established monitoring programs and a training and research centre at Sapo National Park. FFI’s capacity building program saw the development of the first conservation biology curriculum for Liberia’s premier university and engagement of close to 1,000 children in a conservation education program focused on the pygmy hippopotamus. Effective transboundary and protected area law enforcement will be key towards safeguarding and increasing remaining pygmy hippopotamus numbers, whilst awareness raising, collaborative forest management and national/regional policies to reduce deforestation will be needed to secure habitats for pygmy hippopotamus populations to thrive.
It is important to understand that species living in ecosystems that are not conspicuously destroyed may also experience the effects of habitat loss, and hence suffer population declines. This is because habitat loss often manifests itself, at least initially, through less visible but equally threatening habitat degradation. For example, disturbances such as overgrazing do not immediately change the organization of dominant plants and other structural features of an ecological community. First, barely noticeable, a few sensitive habitat specialists disappear, being unable to cope with high levels of grazing. Soon, invasive species that can tolerate trampling start occupying the niches left open by the extirpated sensitive species. Eventually, when livestock eat the last remaining edible morsels of palatable plants not choked out by invasive species, all that is left of the once productive grassland is a field full of dense, unpalatable, invasive shrubbery.
5.1.1 What is habitat fragmentation?
As governments and industries implement measures to accelerate economic growth, ecosystems that formerly covered large, continuous swathes of land are being increasingly subdivided into smaller parcels by roads, farm fields, towns, and other human constructs. A recent study estimated that roads have divided the African continent into more than 50,000 individual units of land; the median unit size was an alarming 6.75 km2 (Ibisch et al., 2016). This process, known as habitat fragmentation, divides once large and widespread wildlife populations—many already suffering from habitat loss—into several increasingly smaller subpopulations. Habitat fragmentation thereby hastens extinctions, as each of these fragmented subpopulations are more exposed to a range of deleterious genetic effects (Section 8.7) than the previously large and connected population.
Habitat fragmentation creates small and isolated subpopulations that have fewer opportunities to find food, water, shelter, and mates.
As if they are victims of double jeopardy, habitat fragmentation also impedes these smaller subpopulations’ dispersal and colonisation abilities. Most species, especially those that occur in low densities, have large home ranges and/or live in ephemeral habitats, and must be able to move freely across the landscape to find shelter, food, water, and mates. A recent global review found that habitat fragmentation has already reduced the average distance of animal movements by two-thirds—from 22 km to 7 km—over the past few decades (Tucker et al., 2018). If they cannot move freely, these individuals cannot fulfil their needs and are at risk of extinction. Habitat interior specialists are particularly vulnerable to habitat fragmentation, as they are often reluctant to disperse over degraded or cleared areas, even if only a few metres wide (Blake et al., 2008; van der Hoeven et al., 2010). And yet, many habitat specialists face barriers much larger than a few metres. This includes Cameroon’s few remaining drill (>Mandrillus leucophaeus, EN) populations, which are facing extinction because individuals are reluctant or unable to disperse over agricultural land that stretches over hundreds of metres (Morgan et al., 2013).
Physical barriers that impede the ability of wildlife to move freely across the landscape also represent a form of habitat fragmentation. Dispersal impeded by human-constructed barriers, such as railways; dams; water-filled ditches; roads; and fences (Figure 5.2), can have disastrous consequences for biodiversity. Consider, for example, Africa’s seasonal drylands. These areas were historically characterized by vast herds of migratory herbivores constantly moving from one area to another after fresh pasture. But as land management systems changed over time, the construction of roads and erection of fences to mark property boundaries impeded the ability of these herds to move freely after the resources they needed to stay alive (Durant et al., 2015; Hopcraft et al., 2015; Stabach et al., 2016). Restricted to only small parts of their range, these once-migratory animals were forced to overgraze the areas they already exploited, leading to extensive population declines. Through this process, Africa has already lost seven mass migrations, each involving millions of animals (Harris et al., 2009). Considering the economic stimulus provided by tourists visiting East Africa’s famous Serengeti-Mara herbivore migration each year, the loss of these seven mass migrations have come at a huge cost to economies elsewhere. Luckily, through diligent conservation efforts, all of Africa’s once-migratory herbivores have managed to persist in small and scattered populations throughout their range (Hoffmann et al., 2015). Section 11.3.1 discusses how some herbivore populations are reverting to their old migration routes after fence removals.
Habitat loss and habitat fragmentation may even threaten the survival of species that are not as obviously dependent on large-scale movements for survival. As discussed in Section 4.2.5, many plants cannot persist without seed dispersal. Unfortunately, many seed dispersers, including forest primates (Estrada et al., 2017) as well as frugivorous birds, such as parrots, orioles, turacos, and hornbills (Lehouck et al., 2009), are sensitive to habitat fragmentation. In one of the few studies looking at this issue in Africa, researchers found that valuable timber trees in Tanzania’s East Usambara Mountains are being extirpated as forest fragments become too small to support viable populations of fruit-eating birds (Cordeiro et al., 2009). The loss of these important seed dispersers will therefore have knock-on effects on the plants that depend on them for survival. Eventually, if enough seed dispersers, or perhaps even a single keystone species, disappear because of habitat fragmentation, entire ecosystems may eventually collapse.
5.1.2 What are edge effects?
Edge effects exacerbate the impact of habitat fragmentation by reducing the functional size of habitat patches.
Edge effects are closely associated with, and exacerbate, the negative effects of habitat loss and fragmentation by altering environmental conditions in the habitat interiors. Dense woodlands, thickets, and forests are especially vulnerable to edge effects. Imagine a tropical forest, especially its large trees forming a continuous leafy canopy. These continuous canopies regulate the microclimate of a forest’s understory by blocking sunlight and wind and maintaining humidity during the day, but also trapping heat rising from the forest floor at night. When the forest’s trees are felled, the continuous canopy is fragmented, which in turn compromises the canopy’s ability to regulate the forest’s microclimate. Cleared areas, as well as forested areas directly adjacent to the cleared areas, will consequently be sunnier, warmer, windier, and dryer during the day, and cooler at night; these climatic changes also disturb nutrient cycles and biomass balances (Haddad et al., 2015). All of these changes further reduce the size of the forest patch to be smaller than the remaining canopy might indicate (Figure 5.3) as the new conditions prevent forest specialists such as shade-loving mosses, seedlings of late-successional trees, and humidity-sensitive amphibians from living in forest edges, leaving them with less interior forest habitat for which they must compete. Importantly, these microclimatic changes can penetrate a forest patch over much greater distances than one might expect. For instance, some forest birds in Uganda are sensitive to edge effects as far as 500 m from cleared areas (Dale et al., 2000).
Edge effects also create several additional threats to the forest species already suffering from altered microclimates. Notably, disturbed edge conditions present a favourable environment for colonisation by fast growing and fast reproducing invasive species. (Threats posed by invasive species are discussed in more detail in Section 7.4). Those forest species that are not displaced by the altered microclimates and invasive species also face elevated predation risk. That is because trees that have died due to altered edge conditions provide suitable perches with clear views from which predatory birds can hunt (Sedláček et al., 2014). The degraded forest edge, sometimes resembling a savannah structure, also provides opportunities for woodland species such as snakes to enter the forests, pushing the remaining forest species even deeper into the forest (Freedman et al., 2009). For this reason, forest edge communities generally consist of widespread generalist species and invasive species, while specialist species that can hang on are, literally and figuratively, living on the edge.
Edge effects beget further edge effects in a positive feedback loop leading to a rapidly disappearing ecosystem.
The most devastating impact of edge effects is that edge effects beget further edge effects in a positive feedback loop leading to a rapidly disappearing ecosystem. First, expanding invasive (and generalist) species populations at habitat edges can easily overwhelm more sensitive habitat specialists. As habitat specialists are displaced at the contact zones, microclimatic conditions change, which allows for invasions even deeper into the fragmented habitat patch. In this way, invasions systematically penetrate deeper and deeper into the forest as microclimates are disturbed, habitat specialists are displaced, and new contact zones are created. The forest plants that die in the process also increase fuel loads, which, combined with drier and windier edge conditions, create an environment increasingly favourable for fire disturbance. Whether from lightning strikes or human activities, subsequent fires burn hotter and over a larger area (van Wilgen et al., 2007), disturbing and destroying more and more habitat each time. Through these mechanisms, edge effects can degrade entire ecosystems over time, harming both the native species and human livelihoods that depend on those areas. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/05%3A_The_Scramble_for_Space/5.00%3A_Prelude_to_the_Scramble_for_Space.txt |
At present, Africa’s biggest driver of habitat loss is agriculture (Potapov et al., 2017). African farmers have always cleared lands to meet their subsistence needs. Much of this clearing was traditionally and historically done in the form of slash-and-burn agriculture (also called shifting cultivation, Figure 5.4). To prepare land for crops, smallholder farmers would first cut down trees to clear the land and to obtain fuel wood. The remaining vegetation would then be burned away to release carbon and other nutrients, which increases land fertility. Farmers would grow crops on these cleared areas for two or three seasons. Then soil fertility would diminish, crop production would decline, and the farmers would abandon the area and clear new land, giving the natural ecosystem on the abandoned land time to regenerate.
Medical and technological advances, and the arrival of colonists, saw Africa’s human population grow considerably since the 1800s. Feeding and accommodating the activities of this growing human population saw an increasing number of natural ecosystems replaced by agricultural land, and less area given the time to regenerate. An increasing number of people also started abandoning their rural subsistence lifestyles for cities in search of jobs, financial freedom, and an easier life. As urbanisation increased (i.e. more people moved to cities) and competition for jobs intensified, an increasing number of city dwellers became dependent on collecting charcoal for cooking and cultivating cash crops, such as yams and cassava (Rudel, 2013). This saw even more natural ecosystems converted, particularly on the outskirts of cities. In the meantime, the remaining rural population became increasingly sedentary due to changing land tenure systems, which forced them into unsustainable farming practices as competition for land increased. These factors not only increased rates of habitat loss, but also changed the nutrient content in the soil which, in turn, reduce the land’s ability to regenerate and to produce food (Drechsel et al., 2001; Wallenfang et al., 2015) which, in turn, leads to even more land clearing for agriculture.
While land clearing for smallholder agricultural needs continues to be an important driver of habitat loss (Tyukavina et al., 2018), its impact is increasingly dwarfed by the demands of commercial interests (Austin et al., 2017). The impact of land grabbing is of particular concern. Foreign companies from Asia and other parts of the world have acquired millions of hectares of land across Africa to stake a claim on the continent’s rich natural resources, and to produce food and biofuels for their own people (von Braun and Meinzen-Dick, 2009). The foreign stakeholders, who often strike these land deals through loan agreements at the governmental level (i.e. with little to no local input), typically prioritise their own needs and profits over local interests with little care for the environment. These deals thus often end with a country saddled with debt they struggle to repay, and environmental damage that will take generations to reverse. Moreover, the foreign companies often employ migrant labourers with fewer protections and rights, compared to local peoples. In the process, while a modest number of local people may benefit from job creation, technology investment, and infrastructure development, a large number of local people become disenfranchised and displaced from the lands that previously supported their livelihoods. These foreign investments are a type of neocolonialism for their resemblance to Africa’s earlier colonial era. They not only drive large-scale habitat loss, but in many instances also leave local people impoverished and desolate (Koohafkan et al., 2011).
The impacts of land clearing for smallholder farms are increasingly dwarfed by the outsized demands of commercial interests.
To understand the impact of land grabbing on Africa’s natural environment, one simply needs to consider their scale. For example, Chinese bioenergy producers recently secured over 48,000 km2 of land in the DRC and Zambia (Smaller et al., 2012). Another deal, between the Ethiopian government and companies from India and Saudi Arabia, saw 5,000 km2 of land (including sections of Gambella National Park) earmarked for commercial agriculture. At the time, this Ethiopian deal threatened both the second largest mammal migration on Earth (Ykhanbai et al., 2014) and the livelihoods of the local pastoralist Anuak community (Abbink, 2011). Fortunately, the Ethiopian government and developers were responsive to concerns raised by conservationists and human rights advocates, and agreed to set some areas aside for conservation, while also putting measures in place to maintain free movement of animals and pastoralists.
Infrastructure developments are also becoming an important driver of habitat loss. Offering access to previously unexploited areas, roads are perhaps the single biggest driver of habitat loss facing Africa’s last remaining wildernesses (Figure 5.5). As prominent tropical biologist Bill Laurance eloquently noted, “Roads usually open a Pandora’s Box of environmental problems—such as illegal fires, deforestation, overhunting and gold mining” (Laurance et al., 2014). A vast, growing body of literature from Africa supports these claims. For instance, research in the Congo Basin has shown how deforestation generally occurs within 2 km from roads (Mertens and Lambin, 1997)>—more roads thus mean more deforestation. Roads also facilitate other drivers of forest loss, including the spread of invasive species, human settlements, fire, and pollution (Kalwij et al., 2008; Potapov et al., 2017). Providing access points for hunters, roads also facilitate unsustainable hunting; a recent review found that the wildlife reductions due to hunting could be detected as far as 40 km from the nearest road (Benítez-López et al., 2017). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/05%3A_The_Scramble_for_Space/5.02%3A_Drivers_of_Habitat_Loss_and_Fragmentation.txt |
5.3.1 Tropical forests
Occupying about 7% of all land surfaces, tropical forests are estimated to contain over 50% of the world’s terrestrial species (Corlett and Primack, 2010). Due to these high levels of biodiversity, the complexity of biological interactions in tropical forests is unparalleled in other ecosystems, and consequently also their importance to humans. On a local scale, the timber and non-timber products from tropical forests sustain the traditions (Box 5.2), livelihoods, and financial well-being of millions of Africans. Tropical forests also have regional importance including protecting catchment areas (Section 4.2.4) and moderating climate (Section 4.2.3). Lastly, as reservoirs of carbon, tropical forests play a globally important role in mitigating the negative effects of anthropogenic climate change (Section 10.4), and with 17% of Earth’s tropical forests, Africa plays a globally important role in tropical forest conservation efforts.
Box 5.2 The Conservation and Exploitation of East African Plants
John R. S. Tabuti
>College of Agricultural and Environmental Sciences,> Makerere University,
>Kampala,> Uganda.
Ethnobotany, as a scientific discipline, studies the relationships between people and plants: how people affect the survival and distribution of plants, and how plants influence human behavior and cultures. For instance, local cuisines are shaped by available plant species, and people cultivate species that they consider useful. Conservation of plant diversity can be aided in many ways by recognizing the importance of plants to people’s livelihood and spiritual practices.
The people of East Africa identify and use a great many plant species that are essential for their well-being (Tabuti, 2006). Native plants are used for food, for construction, to treat the diseases of both people and livestock, and in numerous other ways. Some of the most important species include White‘s ginger (>Mondia whitei) and red stinkwood (>Prunus africana, VU) for medicine, African teak (>Milicia excels, NT) for timber, shea tree (>Vitellaria paradoxa, VU) for food and cosmetics, and African sandalwood (>Osyris lanceolate, LC) as a source of fragrant oil.
Some plant species (and sometimes entire ecosystems, such as forests) are valued for religious or cultural reasons. The plants or forest areas themselves are considered sacred, the site of a deity or spirit, with certain rituals performed using those special plant species or the habitats they occupy. These sacred sites and species are protected by local taboos. For example, the powderbark gardenia (>Gardenia ternifolia) is not harvested for firewood among the Balamogi people of Uganda because it is believed to bring bad luck. Among the Mijikenda people of Kenya, sacred forests known as Kaya are protected because people believe that the forests are inhabited by spirits and are places of prayer and held as a source of ritual power. Cutting down trees, grazing livestock, and farming are prohibited within the Kaya. One protective belief holds that cutting a tree in the Kaya with a machete can result in the machete rebounding and causing injury to the woodcutter. Another belief is that food cooked using wood from these sacred forests can cause sickness, and that a dwelling built with timber drawn from the forest will collapse. Consequently, more than 50 Kaya—ranging in size from 0.3 to 3 km2 and home to 187 plants, 48 birds, and 45 butterfly species—have enjoyed unofficial protection due to religious and cultural beliefs.
Today, however, the plants and their natural communities on which people rely for their well-being are being threatened. By far the greatest threat is land use change and habitat conversion to agriculture to grow food for a growing population. Changing cultural and spiritual values in East Africa, as well as social and economic pressures, are threatening the existence of even sacred forests. For instance, the coronation site of the Paramount Chief of the Balamogi in Uganda was previously protected as a sacred forest by local lore, but it has now been cut down and converted into gardens by local people who no longer follow ancient traditions. Harvesting of plant species, such as the red stinkwood and East African sandalwood, for international markets is also a significant threat no longer held at bay by cultural norms.
Thankfully, several species continue to be actively protected by local communities and governments. According to Greger (2012), traditional healers aid conservation by replanting around 50% of the medicinal plant species that they consider to be important to their practice. For the relationship between people and plants to survive, scientific conservation and local tradition must work together. An example of such collaboration is on display in Uganda’s Budongo Forest Reserve (Figure 5.B), where researchers at the Budongo Conservation Field Station are working with local communities to refine methods for sustainable management and utilisation of the region’s local plants.
Despite the importance of tropical forests, their destruction has become synonymous with the rapid loss of biodiversity (Figure 5.6). Africa had already lost over 65% of its original tropical forests by 1990 (Sayer, 1992); human activities destroyed an additional 308,000 km2 (an area larger than Italy) between 1990 and 2010 (Achard et al., 2014). Losses were particularly severe in Burundi, Benin, and Mozambique, with each country holding less than 5% of its original forest cover (Sayer, 1992). Retaining about half of its original forest cover, the DRC is relatively better off, but current deforestation rates in this country are currently second highest globally (Weisse and Goldman, 2019). Current deforestation rates are so severe in Equatorial Guinea that this country will lose all its forests within the next 20 years if current trends hold (Potapov et al., 2017). Despite these alarming trends—the destruction continues nonstop, particularly in Ghana and Côte d’Ivoire, which saw a 60% and 26% rise in forest loss (the highest rise globally), respectively, between 2017 and 2018 (Weisse and Goldman, 2019). Across Africa, logging is currently the dominant driver of tropical forest loss (causing 77% of total losses over the past decade), followed by agriculture (Potapov et al., 2017).
5.3.2 Rivers and deltas
Due to our dependence on freshwater, humans have always preferred to live near rivers, streams, and lakes. Consequently, these aquatic environments have been destroyed at a scale at least equal to that of terrestrial environments. Rivers have taken a particularly hard hit from human activities, being polluted by industries and dammed to ensure a reliable, year-round supply of water for consumption and irrigation, and to generate hydroelectricity.
Dam construction holds several negative consequences for biodiversity and people. Aquatic organisms that cannot survive the altered river conditions downstream (reduced flow and dissolved oxygen, higher temperatures, and increased turbidity) are most vulnerable. For example, a study from South Africa found that native macroinvertebrate populations (often a good indicator of water quality) were reduced by 50%, and some insect orders virtually extirpated following dam construction (Bredenhand and Samways, 2009). Dams also displace aquatic organisms upsteam. In one well-studied example, back flooding of Mozambique’s Massingir Dam facilitated river substrate changes and the spread of invasive species, which in turn forced sharptooth catfish (>Clarias gariepinus, LC), tiger fish (>Hydrocynus vittatus, LC), and Nile crocodiles (>Crocodylus niloticus, LC) to change their diet. Increased stress levels due to these dietary and environmental changes leave the affected animals susceptible to pansteatitis (a condition where body fat becomes inflamed), leading to mass wildlife mortality events in South Africa’s Kruger National Park (Woodborne et al., 2012). Lastly, dams reduce connectivity in freshwater ecosystems, preventing freshwater organisms from exchanging genetic material, migrating between upsteam and downstream areas, and adapting to changing conditions. For example, in West Africa, the damming of the Senegal River blocked the annual migration path for African river prawns (>Macrobrachium vollenhoveni, LC), a major predator of snails which host schistosomiasis (bilharzia). Once the dam was completed, prawn populations collapsed, leading to a schistosomiasis epidemic in villages upstream from the dam (Sokolow et al., 2015).
Damming rivers harms biodiversity and people both upsteam and downstream from these developments.
Terrestrial ecosystems also suffer from dam construction. Of concern is the direct loss of riverine and palustrine ecosystems downstream from the dam due to reduced waterflow. For example, construction of Nigeria’s Kainji Dam in the Niger River caused the drying of large wetlands and floodplains downstream, in the process displacing nearly 400,000 people who depended on the river’s now-compromised seasonal flood cycles (Drijver and Marchand, 1985). Flooding of upland areas next to dammed rivers also displaces terrestrial wildlife and people. For example, construction of Mali’s Manalati Dam flooded 430 km2 of savannah and 120 km2 of forest, which fractured the migration routes of the region’s nomadic pastoralists, leading to overgrazing and soil erosion of the remaining grazing lands (deGeorges and Reilly, 2006), in addition to a 90% loss of fisheries downstream (Acreman, 1996).
5.3.3 Wetlands
Throughout Africa, wetlands are being mined for valuable peat, or drained and/or filled in for development and agriculture. Through these activities, the region has already lost approximately 43% of its wetlands, with current rates of loss among the highest in the world (Davidson, 2014). This is a major concern because wetlands serve as spawning grounds and nurseries for aquatic and amphibious wildlife and stop-over sites for migratory birds (Box 5.3). Wetlands also provide multiple important ecosystem services. For example, they prevent erosion and runoff by capturing large volumes of floodwater, which is then released slowly over time. This process also allows sediments and nutrients kicked up during flood events to settle out, creating fertile habitats for a wide diversity of animals and plants, as well as for agriculture. Water that leaves after this settling period is cleaner than when it entered, having been filtered by the soil, plants, and microbes of wetlands. This water purification and filtration service is generally cheaper and much more efficient than man-made filtrations systems. The loss of any wetlands, but especially at such large scales, is thus a grave concern not only because of the countless animals and plants threatened with extinction, but also the people that depend on all the valuable ecosystem services they offer.
Box 5.3 Migratory Birds of Africa: The Largest of the Last Great Migrations?
Abraham J. Miller-Rushing1 and John W. Wilson
1>Acadia National Park,> US National Park Service,
>Bar Harbor,> ME, USA.
How are Africa’s bird migrations, the biggest in the world, faring in a rapidly changing world? Each year, about 2.1–5 billion birds (mostly songbirds, but also raptors, waterbirds, and many others) travel back and forth between their wintering grounds in Africa and breeding grounds in Europe and Asia (Figure 5.C). Of the 126 species involved in this migration, over 40% have continuously decreased in abundance since 1970 (Vickery et al., 2014). At first, populations that overwintered in open dry savannahs declined: examples include the Ortolan bunting (>Emberiza hortulana, LC) and European turtle dove (>Streptopelia turtur, VU) which decreased by 84% and 69% between 1980 and 2009, respectively. More recently, species overwintering in the humid Afrotropics also started declining: this includes songbirds, such as the common nightingale (>Luscinia megarhynchos, LC) and river warbler (>Locustella fluviatilis, LC)—populations of both declined by 63%—and waterbirds such as the black-tailed godwit (>Limosa limosa, NT), which declined by 45%.
To survive their long journeys, migratory birds need favourable weather conditions, adequate food sources, and intact habitat not only at the end points where they breed or overwinter, but also along their routes where the migratory animals can rest and refuel (Runge et al., 2015). Disturbances in any of these places can lead to sharp population declines. For example, recent research showed that the habitat quality of a single stop-over site can determine whether a migration is successful or not (Gómez et al., 2017). Illustrating this point, a drought in the Sahel, an important migratory stop-over site, led to food shortages that killed 77% of the world’s common whitethroats (>Sylvia communis, LC); even today, this population has not yet fully recovered (Vickery et al., 2014).
Human activities have greatly contributed to the declines of Africa’s migratory birds (Kirby et al., 2008; Vickery et al., 2014). For example, each year thousands of hectares of wetlands, forests, grasslands, and savannahs are being converted into farmlands and urban areas or polluted by rampant use of pesticides and herbicides. Migratory birds also need to deal with hunters and trappers, and an increasing number of human-made structures, such as high-rise buildings, wind turbines, and power lines that represent collision and electrocution hazards (e.g. Rushworth et al., 2014). Then there is the threat of inconsistent rainfall, which causes food shortages and direct mortality, and climate change, which causes temporal mismatches between migratory movements and abundance of key food resources (Both et al., 2006; Vickery et al., 2014).
Addressing these declines, governments, conservation organizations, and local communities all over Africa have started initiatives to protect migratory birds and their habitats. One such initiative is happening in Kenya’s Tana River Delta, one of the most important stop-over sites along the Asian-East African Flyway. Every year, Basra reed warblers (>Acrocephalus griseldis, EN) return from their Middle Eastern breeding grounds to overwinter in the Delta, which covers 1,300 km2 and supports dozens of threatened species. The area, however, has been under serious threat from development for sugarcane and biofuel crops since 2008. These activities could reduce dry season water flow by up to one third. Local people and conservationists strongly oppose these developments because of its threat to local communities’ ways of life and to wildlife populations. Their efforts gained international attention, and in 2012, Kenyan courts halted development until comprehensive management plans were developed that included environmental impact assessments and local stakeholder engagement (Neville, 2015). Today, local people gain benefit from more sustainable industries, including eco-charcoal audited by the Forest Stewardship Council (FSC), and solar-powered energy to reduce the need for wood.
Also, in West Africa, collaborative conservation initiatives are taking steps to protect the critical East Atlantic Flyway. For example, under the guidance of BirdLife International, Guinea-Bissau residents are now monitoring several wetlands in the Bijagós Archipelago to track how well migratory waterbirds are doing at this critically important stop-over site. Also, in Senegal, where two important stop-over sites (Saloum Delta and Djoudj wetlands) are located, the local non-profit NGO Nature Communautés Développement initiated an extensive conservation education program aimed at safeguarding the region’s birds.
Conserving migratory species that cover huge distances and rely on habitats in many areas is not easy. However, efforts like these in West Africa and Kenya (which combine the interests of local people and wildlife) provide excellent models for others to build from.
Mangrove swamps (sometimes called mangrove forests, though technically a wetland because their function and structure are primarily determined by hydrology, Lewis, 2005; Gopal, 2013) are one of Africa’s most threatened wetland ecosystems. Characterised by woody plants that can tolerate saltwater, mangrove swamps occupy brackish waters in tropical coastal areas, typically where there are muddy bottoms. These areas are sparsely distributed; globally, mangrove swamps cover only 53,000 km2 of land scattered across 118 countries (Dybas, 2015). Protecting Africa’s mangrove swamps, comprising 21% of Earth’s total, is important both biologically and economically. In addition to holding many unique species, mangrove swamps also protect coastal cities and villages from cyclone/hurricane and tsunami damage and provide important breeding and feeding grounds for marine shellfish and fish. One study estimated that mangrove swamps provide an estimated US \$57,000 worth of ecosystem services per hectare (van Bochove et al., 2014). Yet, only 7% of Africa’s mangrove swamps are protected. With so little protection, it comes as no surprise that a large percentage of Africa’s mangrove swamps have been destroyed or damaged by agriculture, urban expansion, pollution, and commercial shellfish farming (Giri et al., 2011). In West Africa, the situation is particularly dire. Wood extraction for commercial fish smoking is one of the biggest drivers of mangrove losses, even within protected areas (Feka et al., 2009). With so much destruction, it should come as little surprise that about 40% of vertebrate species endemic to mangrove swamps are currently threatened with extinction (Luther and Greenberg, 2009).
Mangrove losses around Africa have been extensive despite them providing an estimated US \$57,000 worth of ecosystem services per hectare.
5.3.4 Seasonal drylands
Africa is also rapidly losing its semi-arid savannahs, scrublands, and grasslands through conversion to agriculture (Box 5.4) and desertification—the systematic degradation of formerly complex and adaptive seasonal drylands into barren wastelands (Figure 5.7). When human populations were low, nomadic pastoralism and shifting cultivation enabled people to utilise seasonal drylands in a sustainable way. Today however, population growth, combined with restrictions placed on free movement by administrative borders and competition for land, forces people and animals living on drylands to be more sedentary. While these areas may initially support some agriculture and livestock, unsustainable techniques, such as overgrazing and excessive tilling, lead to soil erosion and the depletion of soil nutrients and natural seed banks. With the cover vegetation gone, the unprotected topsoil is easily lost to wind and flooding, leaving behind the deeper, infertile, and compact subsoil layers with little capacity tfo hold water. The result is something that closely resembles a man-made desert. However, rather than a functional ecosystem characterized by species adapted to life in the desert, these wastelands have lost their original productivity and biological communities, only to be revived through expensive and/or time-consuming land reclamation methods.
Africa is rapidly losing semi-arid ecosystems due to desertification, the conversion of productive ecosystems into barren wastelands.
Box 5.4 Saving Critically Endangered Ground Nesting Birds from Habitat Loss
Bruktawit Abdu Mahamued1,2
1>Biology Department,> Kotebe Metropolitan University,
>Addis Ababa,> Ethiopia.
2>Edge of Existence Fellow,> Zoological Society of London,
>London,> UK.
We are currently witnessing the start of the sixth mass extinction of species on our planet. From here onwards, biodiversity losses are expected to increase rapidly: a recent UN report estimated that about one million species are already threatened with extinction (IBPES, 2019). While the reasons behind these losses vary by region, in Africa, a major driver is habitat loss. With the current push for development, the impacts of habitat loss are increasing dramatically, affecting species both inside and outside of protected areas. Two Ethiopian birds (Figure 5.D), the Liben lark (>Heteromirafra archeri, CR) and white-winged flufftail (>Sarothrura ayresi, CR), exemplify many of the dilemmas associated with protecting biodiversity on unprotected lands where habitat loss is severe.
The Liben Plain is part of the Borana rangelands, managed by Borana pastoralists under their traditional rangeland management system which is generally compatible with conservation ideals. The Borana’s way of life was disrupted about 40 years ago due to pressure from a former Ethiopian government who wanted the Boranas to adopt a more sedentary lifestyle. For example, drilling of water wells in dry season grazing areas disrupted seasonal grazing systems, while fires that the Boranas used to maintain productive grazing lands and prevent shrub encroachment were prohibited. The Boranas also face pressure from changing land tenure systems. The Liben Plain grasslands are located on communal lands upon which nobody can claim ownership. However, if someone wants to farm here, they just pay a tax that in effect assures ownership of the land. The Boranas were initially slow to adopt this farming lifestyle, but when outside settlers started taking advantage of the government’s farming incentives, the Boranas were pushed to do the same to prevent all their ancestral land from being turned over (Mahamued, 2016). The subsequent loss of fire management (and associated shrub encroachment) and cropland expansion, together with increased human and livestock populations, have led to a major loss of the Liben Plains’ natural ecosystem.
The Liben lark is a ground-nesting bird that is near-endemic to Ethiopia (a second population in Somalia may already be extinct; Spottiswoode et al., 2013). Here, its main population is restricted to the open grasslands of the Liben Plain. Although it was previously common in this ecosystem, habitat loss and degradation have reduced the availability of suitable feeding and nesting sites. Further, the reduced population is also increasingly vulnerable to direct threats such as nest predation and trampling of nests by cattle (Spottiswoode et al., 2009). Due to these threats, the lark’s numbers have decreased so dramatically in recent years that it was classified as >Critically Endangered in 2009.
To prevent the extinction of the lark, the Ethiopian Wildlife and Natural History Society (EWNHS), BirdLife International, and other organizations collaborated with local authorities and community leaders in 2016 to establish enclosures for grassland regeneration. These enclosures are in effect communally-managed grassland reserves regulated under a subset of customary laws. These areas not only secure suitable habitats for the Liben lark, they also provide benefits to the Borana community like securing grazing lands for the dry season when the lark is not breeding. This initiative shows early promise—over 350 ha of grassland reserves have already been established, and over 1,000 ha of shrub have been cleared (Kariuki and Ndang’ang’a, 2018). But to truly secure the future of the Liben lark, more support is needed from the Ethiopian government, particularly in preventing further land conversion, supporting ecosystem restoration, and encouraging the Borana pastoralists’ traditional way of life.
Another species facing imminent extinction due to habitat loss is the white-winged flufftail. One of Africa’s most enigmatic birds, the flufftail is an intra-African migrant restricted to a few seasonal high-altitude wetlands in South Africa and Ethiopia. Like the lark, the flufftail is a ground-nester that struggles to find suitable nesting sites relatively free from disturbance. The Berga floodplain, the flufftail’s Ethiopian stronghold, used to be covered by productive grasslands. This unspoiled landscape is now being replaced by settlements, crop farms, and eucalyptus plantations that generate quick profits. This, together with overgrazing, has led to extensive soil erosion, which in turn has altered the structure and grass composition of the floodplain. Today, the floodplain is encroached by invasive weeds and other less desirable vegetation (seen during EDGE project surveys in 2018) which, together with others forms of disturbance, have reduced the amount of suitable habitat available for the flufftail to such an extent that it is now considered >Critically Endangered.
To prevent the extinction of the flufftail, the EWNHS along with the Middlepunt Trust and BirdLife South Africa have taken several steps to improve the outlook for the flufftail. Much of this work involved working with the people at Berga to improve their livelihoods, and to instil a sense of ownership of their local biodiversity. A prominent outcome of this collaboration was a primary school named after the flufftail; results from the project also contributed to a species action plan (Sande et al., 2008). But without continued maintenance, the progress made by this short-term initiative will have limited long-term value. The flufftail’s future thus continues to be dire, as unsustainable land use practices continue to destroy the Berga floodplain. There is an urgent need for joint long-term efforts to reverse the fate of the species, including taking steps to establish protected areas, to initiate carefully-planned ecosystem restoration efforts, and to develop a new species management plan that will provide lasting benefits. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/05%3A_The_Scramble_for_Space/5.03%3A__Habitat_Loss_Impact_on_Africas_Ecosystems.txt |
Until about 150 years ago, the rate of human population growth in Africa had been relatively slow, with the birth rate only slightly exceeding the death rate. Modern medical achievements and more reliable food supplies have changed this balance; they have reduced mortality rates while birth rates remain high. Consequently, Sub-Saharan Africa’s human population has exploded to 1 billion people over the past decade (World Bank, 2019). Today, Sub-Saharan Africa is leading the world in human population growth, projected to increase by four-fold over the next century. Population growth rates for individual countries are similar, if not higher. For example, Ethiopia’s human population has grown from 48 million in 1990—when the region experienced a famine crisis—to nearly 100 million in 2015; current projections forecast a population of 172 million by 2050. The human population of Tanzania’s Dar es Salaam, a coastal city particularly vulnerable to sea level rise (Section 6.3.2), is expected to increase from 4 million to 21 million between 2015 and 2050, while Lagos in Nigeria is expected to grow from 21 million to 39 million people over the same time.
Simple math suggests that more people leads to less space for biodiversity (Figure 5.8), because humans and wildlife compete for the same resources, broadly speaking. With many countries in Africa already facing social, economic, and developmental challenges such as malnutrition, crime, and unemployment, one can almost understand why politicians prioritise socio-economic upliftment over biodiversity conservation. This is a grave mistake; as discussed in Chapter 4, biodiversity and human well-being are intricately linked. It is one of conservation biologists’ most important tasks: to make the link between conservation and human welfare clear to policy scholars and politicians.
In recent years, there has been an increasing tendency of economists, scientists, and politicians to shift the focus from population growth to consumption as the more important underlying driver of biodiversity loss. For many, the emphasis on consumption avoids politically charged topics, such as population control, which most people oppose on ethical or moral grounds, and because it is associated with divisive topics such as xenophobia, racism, and eugenics (Kolbert and Roberts, 2017). Others highlight that it is not the number of people per se, but how natural resources are consumed that is the main cause of environmental decline. Indeed, affluent people and affluent countries have a disproportionate impact on the natural environment because they consume a disproportionately large share of the world’s natural resources. To use one example, the USA accommodates only 5% of the world’s human population but uses 25% of the world’s harvested natural resources each year (WRI, 2019). In fact, decorative Christmas lights in the USA alone use more energy than the annual energy usage of the entirety of Ethiopia or Tanzania (Moss and Agyapong, 2015). And yet, the average USA citizen uses less than half of the energy (measured as carbon emissions) that an average citizen of Qatar uses (World Bank, 2019; see also Figure 5.9), Qatar being a small but wealthy Middle Eastern country.
The major threats to biodiversity are all rooted in expanding human populations and unsustainable consumption patterns.
The global demand for natural resources such as coffee, cacao, palm oil, and timber is helping fuel habitat loss in Africa.
Another important aspect to consider in the consumption argument is that, through increased globalisation, the impacts of consumption in industrialized countries are felt over much greater distances than before (Moran and Kanemoto, 2017). For instance, chocolate consumed in Europe was most likely made with cacao produced in West Africa (Gockowski and Sonwa, 2011); other crops, such as coffee and tea, produced in Africa are similarly enjoyed all over the world. In the best-case scenario, African farmers are satisfying a demand in a global market; at worst, foreign companies are establishing croplands with little benefit trickling down to local people. Supporters of the consumption argument rightfully point out that it would be very unjust to blame the local farmers for cleared forests when they simply produce commodity crops that the international market demands.
As with many other complex challenges, both sides of the population-consumption debate are correct. One method to link the impact (\(I\)) of a human population on the environment is through the formula
\(I = P × A × T\]
This is the IPAT in short, where \(P\) is population size, \(A\) is Affluence (e.g. per capita GPD), and \(T\) is technology (e.g. per capita energy use) (Ehrlich and Goulder, 2007). The IPAT equation is similar in concept to the ecological footprint (Figure 5.9): both illustrate that human populations and consumption patterns interact to exacerbate human impacts on the environment. In other words, many poor Africans can have the same impact on the environment as just a few wealthy Americans, and vice versa.
Both the IPAT equation and ecological footprint concept are insightful as to the challenges facing Africa’s ecosystems and people. Today, Africans are increasingly aspiring to attain the same levels of high consumption as industrial countries. These patterns generally lead to an inefficient, wasteful, and unsustainable use of natural resources (i.e. overconsumption). Population growth rates in many industrial countries are currently slowing; some countries are even experiencing long-term population declines, which allow conservation-minded individuals in those countries to focus their efforts on addressing consumption patterns. The situation is quite different in Africa, where we are faced with increasing per capita consumption and the fastest growing human population on Earth. In the face of the resulting increased competition for space, African conservation biologists must adopt a holistic approach to ensure that welfare standards are upheld or improved while our natural heritage is protected. One of the most important strategies involves championing sustainable economic development over unsustainable economic growth (Section 15.1). While conservation biologists differ in terms of how strongly they argue for addressing the population size issue, most also agree that conservation goals benefit from education, the empowerment of women, and wider access to family planning and reproductive health services. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/05%3A_The_Scramble_for_Space/5.04%3A_Population_Growth_and_Consumption.txt |
There is no doubt that agriculture, forestry, and infrastructure developments—the main drivers of habitat loss and fragmentation—play an important role in socio-economic development across Africa. Nevertheless, many (perhaps most) of these developments are set up to benefit a select few individuals and corporations primarily interested in short-term gains rather than a wide range of stakeholders over the long-term. To maintain biodiversity and improve our quality of life, governments across the region must ensure that the benefits of development are shared fairly across society and that industries are accountable for their fair share of the natural resources they use (Section 4.5.3). Also, the region’s growing number of wealthy people who benefit most from development must re-evaluate their lifestyles (whether willingly or through government interventions, such as taxation) to avoid excessive consumption patterns. Some of the first steps may be relatively easy. For example, the water used to produce Sub-Saharan Africa’s wasted food—a full third of all produced food (FAO, 2013)—equals the annual discharge of the mighty Zambezi River where it enters the Indian Ocean in Mozambique (Beilfuss and dos Santos, 2001). At the same time, we >must all play our part in achieving sustainable development, by encouraging family planning activities and assisting industries to grow in a responsible way (Section 15.1). Neglecting that, we compromise our own futures, and that of our children.
Summary
1. One of the primary threats to biodiversity today are habitat loss and habitat fragmentation. Many species living in tropical forests, freshwater ecosystems, the marine environment, and seasonal drylands are at risk of extinction due to habitat loss.
2. The theory of island biogeography and the species-area relationship can be used to predict the numbers of species that will go extinct due to habitat loss. Both theories predict that large habitat patches are better able to maintain wildlife populations because they accommodate populations better buffered against extinction.
3. Habitat fragmentation describes the process when once large and widespread habitats (and hence wildlife populations) are divided into several increasingly smaller and isolated units. This process leads to extinctions because it impedes dispersal, colonisation, foraging, and reproduction.
4. Edge effects reduce the functional size of habitats because they alter microclimates and expose habitat specialists to displacement by invasive species, predators, and other disturbances.
5. Habitat loss and fragmentation are rooted in expanding human populations and excessive consumption of natural resources. The >IPAT equation illustrates how population size, wealth, and technology together determine our impact on the environment.
5.06: Topics for Discussion
1. Why does oil palm (>Elaeis guineensis, LC) cultivation represent a significant threat to biodiversity in Africa? (In addition to your own research, it might also be useful to read Box 6.1.)
2. Read Harris et al. (2009) on the decline of the world’s mass wildlife migrations. Which lost African migration appeals to you most and why? What species were involved? What numbers of animals were involved? How do you think this migration can be revived?
3. Which ecosystem in your region would you consider the most damaged, and which would you consider the most pristine? Can you explain why these two ecosystems have such different fates?
4. Do you agree with the idea that human population growth is the primary driver of extinctions today? Why? How do we balance protecting biodiversity with providing for a growing human population, and the right of people to have children?
5.07: Suggested Readings
Arcilla, N., L.H. Holbech, and S. O’Donnell. 2015. Severe declines of understory birds follow illegal logging in Upper Guinea forests of Ghana, West Africa. >Biological Conservation 188: 41–49. https://doi.org/10.1016/j.biocon.2015.02.010 Illegal logging is on the increase, and wildlife communities struggle to recover.
Caro, T., J. Darwin, T. Forrester, et al. 2012. Conservation in the Anthropocene. >Conservation Biology 26: 185–88. https://doi.org/10.1111/j.1523-1739.2011.01752.x Even though human activities dominate large areas of the earth, it is important to remember and plan for the many places and ecosystems where human influence is still minimal.
Haddad, N.M., L.A. Brudvig, J. Clobert, et al. 2015. Habitat fragmentation and its lasting impact on Earth’s ecosystems. >Science Advances 1: e1500052. https://doi.org/10.1126/sciadv.1500052 There are many ways that fragmentation hurts biodiversity.
Harris, G., S. Thirgood, J.G.C. Hopcraft, et al. 2009. Global decline in aggregated migrations of large terrestrial mammals. >Endangered Species Research 7: 55–76. https://doi.org/10.3354/esr00173 Habitat loss continues to threaten the world’s remaining mass migrations.
Ibisch, P.L., M.T. Hoffmann, S. Kreft, et al. 2016. A global map of roadless areas and their conservation status. >Science 354: 1423–27. https://doi.org/10.1126/science.aaf7166 Africa still holds large roadless areas; we need to keep it that way.
Laurance, W.F., J. Sayer, and K.G. Cassman. 2014. Agricultural expansion and its impacts on tropical nature. >Trends in Ecology and Evolution 29: 107–16. https://doi.org/10.1016/j.tree.2013.12.001 Agriculture and roads will have severe impacts on Africa’s ecosystems in the coming century.
Rudel, T.K. 2013. The national determinants of deforestation in sub-Saharan Africa. >Philosophical Transactions of the Royal Society B 368: 20120405. https://doi.org/10.1098/rstb.2012.0405 One manuscript in a special issue on deforestation in Africa; other manuscripts in this issue are also worth scanning.
van der Hoeven, C.A., W.F. de Boer, and H.H. Prins. 2010. Roadside conditions as predictor for wildlife crossing probability in a Central African rainforest. >African Journal of Ecology 48: 368–77. https://doi.org/10.1111/j.1365-2028.2009.01122.x Some species are highly reluctant to cross roads, even inside protected areas.
Woodborne, S., K.D.A. Huchzermeyer, D. Govender, et al. 2012. Ecosystem change and the Olifants River crocodile mass mortality events. >Ecosphere 3: 1–17. https://doi.org/10.1890/ES12-00170.1 Damming rivers can lead to ecological disasters
Ykhanbai, H., R. Garg, A. Singh, et al. 2014. >Conservation and “>Land Grabbing”> in Rangelands:> Part of the Problem or Part of the Solution? (Rome: International Land Coalition). http://pubs.iied.org/pdfs/G03853.pdf Conservation biologists should work with local communities to prevent land grabs. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/05%3A_The_Scramble_for_Space/5.05%3A_Concluding_Remarks.txt |
Abbink, J. 2011. ‘Land to the foreigners’: Economic, legal, and socio-cultural aspects of new land acquisition schemes in Ethiopia. >Journal of Contemporary African Studies 29: 513–35. https://doi.org/10.1080/02589001.2011.603213
Achard, F., R. Beuchle, P. Mayaux, Het al. 2014. Determination of tropical deforestation rates and related carbon losses from 1990 to 2010. >Global Change Biology 20: 2540–54. https://doi.org/10.1111/gcb.12605
Acreman, M.C. 1996. Environmental effects of hydro‐electric power generation in Africa and the potential for artificial floods. >Water and Environment Journal 10: 429–35. https://doi.org/10.1111/j.1747-6593.1996.tb00076.x
Arcilla, N., L.H. Holbech, and S. O’Donnell. 2015. Severe declines of understory birds follow illegal logging in Upper Guinea forests of Ghana, West Africa. >Biological Conservation 188: 41–49. https://doi.org/10.1016/j.biocon.2015.02.010
Austin, K.G., M. González-Roglich, D. Schaffer-Smith, et al. 2017. Trends in size of tropical deforestation events signal increasing dominance of industrial-scale drivers. >Environmental Research Letters 12: 054009. https://doi.org/10.1088/1748-9326/aa6a88
Beier, P., M. van Drielen, and B.O. Kankam. 2002. Avifaunal collapse in West African forest fragments. >Conservation Biology 16: 1097–111. https://doi.org/10.1046/j.1523-1739.2002.01003.x
Beilfuss, R., and D. dos Santos. 2001. Patterns of hydrological change in the Zambezi Delta, Mozambique. >Working Paper 2 (Baraboo: International Crane Foundation). https://doi.org/10.13140/RG.2.2.14255.12961
Benítez-López, A., R. Alkemade, A.M. Schipper, et al. 2017. The impact of hunting on tropical mammal and bird populations. >Science 356: 180–83. https://doi.org/10.1126/science.aaj1891
BirdLife International. 2019. >Migratory birds and flyways. https://www.birdlife.org/worldwide/programs/migratory-birds
Blake, S., S.L. Deem, S. Strindberg, et al. 2008 Roadless wilderness area determines forest elephant movements in the Congo Basin. >PloS ONE 3: e3546. https://doi.org/10.1371/journal.pone.0003546
Both, C., S. Bouwhuis, C.M. Lessells, et al. 2006. Climate change and population declines in a long-distance migratory bird. >Nature 441: 81–83. https://doi.org/10.1038/nature04539
Bredenhand, E., and M.J. Samways. 2009. Impact of a dam on benthic macroinvertebrates in a small river in a biodiversity hotspot: Cape Floristic Region, South Africa. >Journal of Insect Conservation 13: 297–307. http://doi.org/10.1007/s10841-008-9173-2
Cordeiro, N.J., H.J. Ndangalasi, J.P. McEntee, et al. 2009. Disperser limitation and recruitment of an endemic African tree in a fragmented landscape. >Ecology 90: 1030–41. http://doi.org/10.1890/07-1208.1
Corlett, R., and R.B. Primack. 2010. >Tropical Rainforests:> An Ecological and Biogeographical Comparison (Malden: Wiley-Blackwell). https://doi.org/10.1002/9781444392296
Dale, S., K. Mork, R. Solvang, et al. 2000. Edge effects on the understorey bird community in a logged forest in Uganda. >Conservation Biology 14: 265–76. https://doi.org/10.1046/j.1523-1739.2000.98340.x
Davidson, N.C. 2014. How much wetland has the world lost? Long-term and recent trends in global wetland area. >Marine and Freshwater Research 65: 934–41. https://doi.org/10.1071/MF14173
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Life-threatening heatwaves, drowning coastal towns, tens of thousands of displaced refugees… These words may very well describe a scene from the latest horror movie. But they also describe the nightmare scenario facing us humans in just a few decades if we continue to leave the threat of climate change under-addressed. This term, climate change (which is shorthand for anthropogenic climate change), refers to the complete set of climate characteristics—temperature; precipitation; pressure systems; wind patterns; and oceanic currents—that are changing both locally and regionally due to human influences. It is closely related to global warming, also called global heating, which describes the general trend of increasing global temperatures we see under climate change.
Climate change has the potential to render Earth unrecognisable from what any human has ever experienced. These changes will have an immense impact on ecosystem services, global economies, and our own quality of life. Yet, while there is much talk about these risks, there is too little action addressing its main causes. Some of the lack of action may be attributed to “climate change” and “future” often being used in the same sentence, giving politicians and industries a false impression that we can deal with climate change once we achieved sufficient economic growth. The reality could however not be further from the truth, as we already see signs of the changes to come here today (Table 6.1), including near-annual crop-failures, record-high temperatures, and record-strength coastal storms.
Table 6.1 Some examples illustrating how climate change is already impacting Africa.
Impact
Evidence
Increased temperatures and incidence of heat waves
Global temperatures in 2016 were the warmest since modern recordkeeping began in 1880; the two previous records were set in 2015 and 2014 (Gillis, 2017). Heat waves are also hotter, longer, and over a larger area than before (Russo et al., 2016).
Widespread droughts
East Africa saw its worst drought in 60 years from mid-2011 to mid-2012. Over 250,000 people died; nearly 10 million more needed humanitarian assistance (Maxwell et al., 2014). The increased intensity of similar droughts in 2016 are directly attributable to climate change (Uhe et al., 2017).
Rising sea levels
Coastal floods disrupt lives and local economies in Ghana, Nigeria, and Benin almost every year. Coastal erosion has damaged commercial properties in The Gambia and Senegal, while the coastline retreated 35 m in some areas of Togo (Fagotto and Gattoni, 2016).
Earlier spring activity
Bloom dates for several plants, including commercially grown apple and pear trees in South Africa, are now between 1.6 and 4.2 days earlier per decade than 35 years ago (Grab and Craparo, 2011).
Shifts in species ranges
Malaria recently appeared in the highlands of Ethiopia, Kenya, Rwanda, and Burundi, in areas where it did not occur before (Siraj et al., 2014).
Wildlife population declines
Reporting rates for some bird species endemic to the Cape Floristic region declined by over 30% over the past 15 years (Milne et al., 2015).
Thankfully, with the increase in understanding that our activities are creating a global crisis of epic proportions, the impacts of climate change are now being actively debated in the corridors of governments and major corporations. Politicians, the media, and others are also increasingly replacing “climate change” with more vivid language, like “climate crisis” and “climate emergency” (e.g. Carrington, 2019). This will hopefully encourage even more governments and industries to come to the table and cooperate like never before to address the fundamental drivers of climate change. Solving this global crisis requires an international multi-pronged approach that should include ecosystem protection and restoration (Chapter 10), direct species management (Chapter 11), and legislative action (Chapter 12). But before we consider the solutions, we will first investigate why climate change is happening, and how it will impact biodiversity over the coming decades. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/06%3A_Our_Warming_World/6.00%3A_Prelude_to_Our_Warming_World.txt |
Greenhouse gases are essential for life on Earth. But too much of them cause Earth to heat up too much too fast, leading to climate change.
The climate change we are experiencing today is driven by human activities that increase greenhouse gas concentrations in Earth’s atmosphere. Although we mainly hear about greenhouse gases in the context of their contribution to climate change, they are in fact essential for life on Earth. Consider for a moment carbon dioxide’s (CO2) critical role in photosynthesis, and water vapour’s role in the formation of rain. Both of these gases are greenhouses gases. Greenhouse gases earn their name because they function much like the glass covering a greenhouse; they allow sunlight to easily pass through the atmosphere but trap the reflected heat energy so that it stays close to Earth’s surface. This greenhouse effect allows all the organisms on Earth, even us humans, to flourish. Without greenhouse gases, temperatures would drop, and our planet would be too cold to sustain life. However, high concentrations of greenhouse gases can also be harmful. Think for a moment of greenhouse gases as “blankets” covering the Earth’s surface: more “blankets” will trap more heat, giving rise to higher temperatures. This is exactly what is happening today—human activities are currently increased greenhouse gas concentrations in the atmosphere so much, and at such a fast pace, that Earth is heating up too fast for biodiversity to adapt to the changes.
Africa’s biggest contribution to climate change comes from the destruction of complex ecosystems, which leads to the loss of important carbon sinks.
At present, the single biggest cause of increased greenhouse gas concentrations is the burning of fossil fuels. Since the Industrial Revolution about 200–250 years ago, humans have become heavily dependent on the energy captured in these fuels—coal, oil, and natural gas—for activities such as transportation, heating, manufacturing, and electricity generation. Fossil fuels contain a high percentage of carbon, so when it is burned, that carbon is released into the atmosphere, generally as CO2. Consequently, since human populations started exploding and have been using fossil fuels at increased rates, the greenhouse effect has been significantly amplified.
While fossil fuel burning is currently the biggest overall driver of climate change, the greatest contribution from Africa is the destruction of carbon sinks, such as tropical forests (Box 6.1) and peatlands. Destroying these ecosystems contributes to rising atmospheric CO2 concentrations directly through burning of vegetation that releases carbon, and indirectly through the loss of vegetation that would otherwise extract CO2 from the atmosphere if they were still alive. The contribution of ecosystem loss to climate change is substantial: 13% of today’s global carbon emissions can be accounted for by tropical deforestation (IPCC, 2014). This impact is much stronger in Africa where deforestation accounts for 35% of the region’s overall climate change impacts (WRI, 2019). In comparison, Africa’s energy and agricultural sectors contribute 30% and 24%, respectively.
Box 6.1 Does Oil Palm Agriculture Threaten Biological Diversity in Equatorial Africa?
Abraham J. Miller-Rushing
Acadia National Park, US National Park Service,
Bar Harbor, ME, USA.
Oil palm (Elaeis guineensis, LC) is among the fastest expanding crops in the world. Native to West Africa, this species produces more oil per hectare than any other cultivated crop in the world. It should thus come as no surprise that it has become world’s most popular source of vegetable oil. Tropical Africa is poised as a hotspot for new oil palm plantations (Linder, 2013; Vijay et al., 2016). Is this a good thing? Will the benefits from jobs and carbon sequestration outweigh the loss of native ecosystems?
To many people, oil palm cultivation presents a win-win situation. The industry provides jobs and economic stimulus (Figure 6.A) and claims that oil palms sequester carbon from the atmosphere (Burton et al., 2017). This could potentially help countries offset carbon emissions; they may even receive funding from carbon markets. Palm oil can also be used for cheap bioenergy production, and as an ingredient in food and household products (e.g. cooking oil, baked goods, salad dressings, shampoo, and soap). Consequently, demand is rapidly growing as sales of processed and packaged foods (today about 50% of packaged foods include palm oil as an ingredient) expand globally.
Oil palm plantations, however, are rarely developed in environmentally friendly ways that allow them to realise their potential value. Rather, it generally comes at a great ecological cost. For example, to ensure net positive carbon sequestration, oil palm plantations must be developed on degraded landscapes, rather than displacing intact ecosystems that are already very effective at sequestering carbon (Burton et al., 2017). In practice however, intact forests are more often logged to make space (and additional revenue) for oil palm plantations (Ordway et al., 2019), resulting in habitat loss and net positive carbon emissions. Oil palm plantations are also often associated with great societal costs, like land grabbing, exploitation of local people, and displacement of traditional activities (Linder and Palkovitz, 2016). The influx of migrant plantation workers puts further strain on the environment through unsustainable hunting of bushmeat. One study found that primate population sizes declined by 25–100% after palm plantation development in Côte d’Ivoire (Gonedelé et al., 2012).
Recently, Herakles Farms/SG Sustainable Oils, an American agribusiness company, attempted to develop a 730 km2 oil palm plantation in Cameroon. This land grab would have been one of the largest palm oil projects in Africa, nestled deep within Cameroon’s lowland tropical forests, one of the continent’s most biologically diverse and threatened ecosystems. The forests threatened by this development is situated adjacent to four protected areas that include two national parks (Linder and Palkovitz, 2016), and host 14 species of threatened primates, including the Nigeria-Cameroon chimpanzee (Pan troglodytes ellioti, EN) (Linder, 2013). Residents and environmental groups opposed the plantation because of possibly illegal activities by the company, the ecological consequences of the project, and because the local people would have received little, if any, benefit from the project. After protracted debate and struggle, including intimidation and the arrest of local social and environmental activists, the company withdrew its Cameroonian plans in 2013.
It seems that there is potential for oil palm plantations to be good for economic development, job creation, and conservation. But in practice, companies establishing these plantations frequently exploit local people and degrade local ecosystems. They sometimes even do this under the auspices of sustainability, arguing that low-impact activities by traditional peoples indicate that the area is already degraded and thus suitable for development. Hopefully, one day we can live in a world where palm oil companies and robust legal systems truly consider the protection of biodiversity and the rights of local people in those operations.
The link between human-induced climate change and atmospheric CO2 concentrations was first highlighted in the late 19th century (Arrhenius, 1896). However, it was not until the mid-1950s (e.g. Kaempffert, 1956) that scientists started to raise concerns about increasing CO2 concentrations in the atmosphere. By the 1980s, as global annual mean temperatures started to rise, consensus about climate change linked to CO2 began to spread among the broader public. Yet concrete steps to curb CO2 emissions would only be initiated decades later (Section 12.2.1). In the meantime, CO2 emissions continue to accelerate (Figure 6.1): more than 37 billion tonnes of carbon, a new record, were released into the atmosphere in 2018 (Jackson et al., 2018; Le Quéré et al., 2018). To put it in another way, during 2018, humans released on average over 100 million tonnes of CO2 into the atmosphere every day.
The second-most important greenhouse gas that drives climate change is methane (CH4). Methane is a natural by-product emitted from decaying organic matter, most notably from wetlands that inhibit the speed of decomposition. These important ecosystem processes release methane into the atmosphere, albeit in relatively low concentrations. However, human activities have boosted methane emissions significantly over the past few centuries, through wasted food decaying at landfills, leaks from natural gas wells, an increase of industrial-scale cattle and dairy farms, and large-scale destruction of swamps and peatlands. Warmer temperatures also result in the drying of wetlands and peatlands; this drying speed up decomposition of organic material, which increases the rate of methane release. Methane currently constitutes 16% of all global greenhouse gas emissions released by humans (IPCC, 2014). This may not seem to be a major contribution; however, methane is 72 times more effective than CO2 in trapping radiation over a 20-year period (Forster et al., 2007), so even small increases in atmospheric methane can have dramatic effects.
The third important greenhouse gas that drives climate change is nitrous oxide (N2O), also known as laughing gas. Nitrous oxide is a by-product of synthetic fertilisers used in agriculture, burning of fossil fuels, and several industrial processes, and accounts for 6% of all human-caused greenhouse emissions (IPCC, 2014). However, it is even more potent than methane, and stays in the atmosphere for about 114 years, so the impact of one tonne of N2O is equivalent to 310 tonnes of CO2 over 100 years (Forster et al., 2007). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/06%3A_Our_Warming_World/6.01%3A_Drivers_of_Climate_Change.txt |
Climate change forecasting is famously complex, with a great amount of uncertainty attached to the task. Most of us have been exposed to short-term (i.e. day-to-day) weather forecasts on television, radio, and newspapers. These daily weather forecasts are derived from current weather measurements while considering the historical record of past events that occurred during similar conditions. Some daily forecasts may also be created for as many as two weeks into the future, but these future outlooks are generally much less detailed. In contrast, forecasting climate change involves predicting novel weather conditions for several decades into the future. The general circulation models (GCM) used for climate change forecasting (Figure 6.2) also need to account for a great number of highly variable components, each affecting one another across the only planet we can adequately measure or examine (we have no other planet where we can test predictions). Among thousands of considerations, climatologists (scientists who study climate) need to account for how human activities might change over time, and how these activities will change the atmosphere’s composition. They also need to account for how much CO2 the world’s oceans and plants will absorb, and how wind and fire might influence these processes. Combining all the component parts, climatologists then need to estimate how increased temperatures will affect the polar ice caps, how the melting ice will affect oceanic conditions and currents which, in turn, will affect terrestrial conditions and weather patterns. Uncertainty also exists over interactive effects of some drivers. For example, higher temperatures increase evaporation and cloud cover which, in turn, will have a cooling effect (a similar short-term cooling effect, caused by an albedo effect, is noted after an ecosystem is cleared due to the bare ground’s ability to reflect more sunlight than it absorbs, Section 4.2.3). Because of the complexity of these and other variables going into climate models, a great number of research groups are encouraged to develop their own climate forecasts, each using a range of different scenarios on how human activity might change in the future.
To further improve upon climate change forecasting, in 1988, the UN appointed a group of leading scientists, collectively known as the Intergovernmental Panel on Climate Change (IPCC), to study the implications of climate change. By regularly doing extensive reviews of all the evidence and climate science literature, the IPCC has found that, despite the complexity of climate models, results of all the models taken together exhibited significant agreement with changes already observed. Climate change models have also proven reliable in predicting responses of biodiversity to climate change (Fordham et al., 2018). Thus, while some fringe groups may continue to deny the validity of climate science, there is broad consensus among the world’s scientists that increased atmospheric greenhouse gases—caused by human activities—are causing the world’s climate to change, and it will continue to change in coming decades. While climatologists continue to improve on the finer details of their models, conservation biologists can and should confidently use the climate forecasts available for general conservation planning purposes.
Assuming human activities continue business as usual, and current greenhouse gas emission rates continue unabated, climatologists predict that average annual temperatures in Sub-Saharan Africa will increase by 0.5°C by 2050, compared to temperatures late in the 20th century (Serdeczny et al., 2017). The increase could be even greater, towards 4°C, if humans emit more greenhouse gases than predicted and Earth’s carbon storage systems underperform. Conversely, temperatures could warm less or more slowly if we manage to slow greenhouse gases emissions and better protect natural carbon sinks. Unfortunately, current evidence suggests that the higher temperature estimates seem more likely. For example, 2016 was the hottest year (since modern record-keeping) globally for the third straight year (Gillis, 2017) with temperatures already 0.9°C above late 20th century averages. Another climate record was set in April 2018, which was Earth’s 400th straight warmer-than-average month (NOAA, 2018c). Also, more locally, scientists observed that temperatures in some South African national parks reached temperature increases predicted for 2035 already in 2015 (van Wilgen et al., 2016). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/06%3A_Our_Warming_World/6.02%3A_Predicting_Earths_Future_Climate.txt |
Climate change is not a new phenomenon. During the past 2 million years, there have been at least 10 cycles of global warming and cooling. When the polar ice caps melted during warming periods, sea levels rose to well above their earlier levels, and a larger portion of Earth experienced tropical climates. During cooling periods, the polar ice caps expanded, sea levels dropped, and tropical species’ ranges contracted. Sometimes these changes occurred gradually, which enabled the affected species to adapt. But the onset of some climate change periods was abrupt, causing major ecosystem disruptions and global mass extinction events (Section 8.1). Yet, nature recovered every time; many of the species we see today are survivors of previous climate change events. It is thus fair to ask why today’s climate change is of such concern to us.
6.3.1 Climate change’s impact on people
History provides us with many lessons to illustrate the impact of climate change on human societies. These lessons start with the earliest well-documented example of a societal collapse—that of the Middle East’s Natufian communities roughly 10,000 years ago—which has been attributed to climatic changes (Weiss and Bradley, 2001). Since then, climate change has regularly contributed to the collapse of complex human societies across the world. Notable examples of such collapses include the Akkadian Empire (the world’s first empire) of the Middle East (Carolin et al., 2019), Egypt’s Old Kingdom (who constructed the pyramids), Central America’s Classic Mayan civilisation, the USA’s first English colony (deMenocal, 2001), several Chinese dynasties (Wang et al., 2010), and the Late Bronze Age societies along the Mediterranean Sea (Kaniewski et al., 2013). Also, in Southern Africa, the fall of the Mapungubwe Kingdom has been attributed to crop failures and declining grazing lands due to regional droughts and warming cycles (O’Connor and Kiker, 2004).
Unlike the unavoidable natural climatic shifts that led to the historical societal collapses discussed above, we have brought today’s climatic change impacts upon ourselves. Because of our general lack of response in addressing the drivers of climate change, thousands of people will suffer the consequences. Prominently, many parts of Africa are already seeing higher temperatures and longer droughts (Engelbrecht et al., 2009). These conditions are compromising our quality of life (Watts et al., 2017) by leading to more intense wildfires (Jolly et al., 2015; Strydom and Savage, 2016), increased incidences of malaria (Siraj et al., 2014), increased crop failures (Myers et al., 2014; Medek et al., 2017), and increased competition for water (Flörke et al., 2018). Many coastal areas are also seeing storms increasing in intensity and frequency, exposing people living near large rivers, deltas, and estuaries to more frequent flooding (Figure 6.3) and storm surges (Fitchett and Grab, 2014). Sea level rise is expected to leave many low-lying oceanic islands uninhabitable within a few decades (Storlazzi et al., 2018). With all these impacts expected to increase the competition for space under an increasing human population, it would be wise for the world’s governments to start preparing for thousands of climate refugees that would need to be relocated in the near future (Merone and Tait, 2018).
To combat climate change, politicians of several countries have started to enact laws to reduce greenhouse gas emissions and habitat destruction (Section 12.2.1). Many industries are also hard at work developing “greener” technologies to enable us to live more sustainable lives. Conservation biologists also play a crucial role in mitigating the negative impacts of climate change. In addition to highlighting the plight of the natural world to society at large, we could work towards reducing the loss of ecosystem services and preventing species extinctions. To accomplish this task, we need to identify which species and ecosystems are most sensitive to climate change and develop strategies that will ensure the continued persistence of as many sensitive species and their habitats as possible. The rest of this chapter is dedicated to methods we can employ to understand which species are sensitive, and how they may respond to climate change, while Chapters 10–15 discuss methods we can employ to better address climate change.
6.3.2 Climate change’s impact on terrestrial ecosystems
Aside from regional variations in temperature and precipitation, Earth’s surface will be a few degrees warmer in future than the temperatures we experience today. In effect, that means that today’s climatic zones will generally shift upslope in mountainous areas and towards the poles on lowlands, plains, and plateaus. To survive, climate-sensitive plants and animals will need to track these shifts so that they remain within their suitable climatic envelopes of temperature and precipitation.
Climate change on mountains
Species that live on mountains are at particular risk from climate change. Because temperatures decrease by roughly 0.65°C for every 100 m in elevation rise (known as temperature lapse rates), a 1°C increase suggests that climate-sensitive species living on a mountain would be displaced by at least 150 m (1.5 m/year) upslope between the years 2000 and 2100. Species that live on the lower slopes of mountains and are mobile enough to make such an adjustment may have opportunities to move to higher ground. However, species that live on or near peaks may have nowhere else to go as the world heats up, resulting in what biologists call mountain-top extinctions. While a mountain-top extinction has yet to be recorded in Africa, we have ample evidence to suggest that the region’s wildlife is vulnerable to it. For example, due to climate change, populations of some bird species endemic to the Cape Floristic region’s mountains have shrunk by 30% over the past two decades (Milne et al., 2015). Species inhabiting Tanzania’s Eastern Arc Mountains (Dimitrov et al., 2012), Albertine Rift (Ponce-Reyes et al., 2017), and the Guinean Forests of West Africa (Carr et al., 2014) appear to have experienced similar declines. Given these observations, it is only a matter of time before one of Africa’s mountain specialists follows the example of Costa Rica’s once abundant Monteverde golden toad (Bufo periglenes, EX), the first known amphibian extinction attributed to climate change (Crump et al., 1992).
Species that live on mountain peaks are vulnerable to climate change because they may have nowhere else to go as the world heats up.
Climate change in the lowlands
The response of species living in lowlands and on plains tend be more variable and complex than those living on mountains. While some species may only need to make minor range adjustments, researchers estimate that some African taxa may need to move 500 km (Barbet-Massin et al., 2009)—maybe even 1,000 km (Hsiang and Sobel, 2016)—to keep up with climate shifts. For species, such as Tanzania’s savannah birds that have already shifted their distributions by 200–300 km (Beale et al., 2013), adapting seems relatively easy thanks to their mobility and largely intact ecosystems. Unfortunately, the rate of climate change will likely outpace the ability for most species to adapt (Jezkova and Wiens, 2016; Wiens, 2016). For example, nearly 62% of Sub-Saharan Africa’s species are predicted to undergo range contractions (Hole et al., 2009), and 37% species are facing extinction if climate forecasts hold true (Thomas et al., 2004). Species living in Southern Africa’s Miombo Woodlands are even more vulnerable, where as many as 90% of amphibians, 86% of birds, and 80% of mammals face extirpation (Warren et al., 2018).
Species of tropical lowland forests and deserts are also highly vulnerable to shifting climates. Many tropical species have narrow tolerances for temperature and rainfall variation, while desert specialists may be at the limits of their physiological heat and desiccation tolerances (Figure 6.4). Consequently, even small changes in the climate of these two ecosystems may have major effects on reproduction, species distributions, and hence ecosystem composition (Box 6.2). One species already impacted is the nocturnal aardvark (Orycteropus afer, LC): a study in Southern Africa’s Kalahari Desert found over 80% mortality rates in this species during recent summers (Rey et al., 2017). The high levels of mortality in this species was attributed to above average temperatures, which subjected the animals to heat stress, leading to behavioral disruptions, declining body conditions, and eventually starvation. The impact of climate change on the aardvark is of concern because it is an ecosystem engineer: their burrows provide denning and refuge sites for multiple other species (Whittington-Jones et al., 2011).
Box 6.2 Desert Birds and Climate Change
Susan Cunningham1 and Andrew McKechnie2,3
1FitzPatrick Institute of African Ornithology, DST-NRF Centre of Excellence,
University of Cape Town, South Africa.
2DST-NRF Centre of Excellence at the FitzPatrick Institute,
Department of Zoology and Entomology,
University of Pretoria, South Africa.
3South African Research Chair in Conservation Physiology, National Zoological Garden,
South African National Biodiversity Institute,
Pretoria, South Africa.
[email protected]
Deserts, with their extreme temperatures and scarce and unpredictable rainfall, are among the most inhospitable environments on the planet. To survive and breed in arid regions, organisms must minimise their energy and water requirements, and avoid exposure to potentially lethal temperatures. Birds are generally small and diurnal; and are therefore among the groups of animals most vulnerable to even small increases in air temperatures associated with climate change. Studies of the effects of temperature on arid-zone birds can thus be highly informative in terms of identifying new conservation challenges posed by global warming, developing mitigation measures, and understanding the management interventions that may become necessary during the 21st Century.
Daytime temperatures in many deserts regularly exceed avian body temperature, creating conditions under which birds can avoid lethal heat stroke only by dissipating heat via evaporation. But rapid rates of evaporation increase the risk of birds becoming lethally dehydrated. Desert birds thus face life-or-death decisions between avoiding hyperthermia by evaporative cooling versus avoiding lethal dehydration by minimising water losses. Mass mortality events occasionally take place during extreme heat waves when air temperatures exceed birds’ physiological tolerance limits. In Australia, for example, there are both historic and contemporary accounts of die-offs sometimes involving millions of birds. As Earth heats up under climate change, the risk of such die-offs in desert birds is expected to increase dramatically for the deserts of Australia and North America during the 21st Century (McKechnie and Wolf, 2010; Albright et al., 2017).
Africa’s arid regions are also experiencing significant temperature increases which are predicted to continue over the next several decades (Conradie et al., 2019). Under these conditions, the impact of air temperature on avian physiology can be mediated by behavior. Birds employ a trio of behavioral adjustments to manage heat load and keep their body temperatures within safe limits. These include shade-seeking, reducing activity to minimise metabolic heat production, and gaping the beak (panting, sometimes accompanied by gular flutter) to facilitate respiratory evaporative cooling (Figure 6.B). Although these behaviors can buffer birds against physiological costs of high temperature, they carry subtle but important costs of the own, notably via their impact on birds’ ability to forage.
For desert birds, foraging is critically important for maintaining both energy and water balance, as most species obtain all their water from food. Reduced activity almost inevitably means reduced food intake via impacts on time available for foraging. Seeking shade also carries costs: for some species, returns on foraging effort in shaded locations are significantly lower than in the sun (e.g. Cunningham et al., 2013). Finally, respiratory evaporative cooling can severely restrict the ability of actively-foraging birds to acquire food due to mechanical constraints on simultaneously gaping the bill and using it for prey capture and handling (e.g. du Plessis et al., 2012).
Under climate change, the implications of these behavioral trade-offs between foraging and thermoregulation are non-trivial. Inability to balance water and energy budgets mean birds progressively lose body condition during heat waves (du Plessis et al., 2012). Compromised foraging also affects birds’ capacity to provision offspring, resulting in reduced nest success and/or smaller, lighter fledglings which may struggle to survive and recruit into the breeding population (e.g. Cunningham et al., 2013, Wiley and Ridley, 2016).
Successfully balancing the trade-offs between foraging and thermoregulation, and between hyperthermia and dehydration, is the secret to success for birds in hot places. As the climate warms, achieving this balance will become ever more challenging. Sublethal behavioral costs of keeping cool kick in at temperatures cooler than those promoting mass mortalities. In some parts of the world, such as Southern Africa, the loss of birds from desert ecosystems may therefore occur through the insidious whittling away of fitness and weakening of populations (Conradie et al., 2019) before we even witness the dramatic die-off events for which Australia is already infamous.
An additional concern for lowland ecosystems is that climate change will likely lead to the creation of novel (i.e. hotter) ecosystems unlike any others currently on Earth (Williams et al., 2007). These changes will lead to biotic attrition. The gradual impoverishment of biological communities of lowland ecosystems as species either go extinct or move away while tracking their climatic envelopes. What is not clear is how the niches left open by the net loss of species, and newly created niches in the novel ecosystems, will be filled. The most likely scenario is that more tolerant, generalist species will fill the empty niches. However, with the inevitable loss of some species, combined with the decoupling of important biological interactions (discussed below), some functions and services associated with lowland ecosystems are likely to eventually collapse. It is important to note that tropical lowland forests and deserts are by no means the only ecosystems vulnerable to biotic attrition. For example, researchers have found that even mild warming would expose the Ethiopian Highlands to biotic attrition (Kreyling et al., 2010).
Climate change and dispersal limitations
Across many diverse ecosystems, a great number of species are threatened by climate change because of their poor dispersal abilities. Because they lack appropriate dispersal mechanisms, species, such as slow maturing plants (Foden et al., 2007), mosses, and flightless insects may simply not be able to keep up with changing climatic conditions. The impacts of climate change on Africa’s dispersal-limited species can already be seen. For example, the once abundant Aldabra banded snail (Rhachistia aldabrae, CR) is today so rare that this Lazarus species (Figure 6.5) was once believed to be extinct due to climate change (Battarbee, 2014). There are also fears that successive droughts in the Cape Floristic Region may have recently driven a rare sorrel species (Oxalis hygrophila, CR) to extinction (Zietsman et al., 2008). Next might be the cave katydid (Cedarbergeniana imperfecta, CR) and Marais’ lace-winged katydid (Pseudosaga maraisi, CR); these highly threatened insects count among Africa’s very few cave specialists, and yet, by living in highly restricted and restrictive ecosystems, they face major challenges in adapting to climate change (Bazelet and Naskrecki, 2014). Dispersal limitations will also greatly affect terrestrial species living on oceanic islands, which will find it near impossible to track their climatic niches as it moves over the ocean. One such species is Cabo Verde’s Raso lark (Alauda razae, CR); with a population size that fluctuates in response to rainfall, climate change induced drought conditions have pushed this bird to the brink of extinction in recent years (BirdLife International, 2016).
Climate change and biological interactions
Species that are highly mobile are not entirely spared from the negative impacts of climate change. Consider migratory species for a moment. The same way the musicians of an orchestra rely on a conductor to remain synchronised, migratory species rely on environmental cues, such as daylength and temperature, to decide when they need to start moving from one area to the next. But because different species rely of different environmental cues to time their life cycles (e.g. breeding), not all species will adjust to climate change at the same rate. There is consequently a high likelihood that climate change will disrupt these synchronous movements that the animal kingdom has developed over thousands of years (Renner and Zohner, 2018). This disruption of timed aspects of species’ life cycle, such as migration and breeding, is called phenological mismatch or trophic asynchrony. Researchers have already seen signs of phenological mismatch: some migratory birds that overwinter in Africa have started to migrate to their European breeding grounds at earlier dates than before (Both et al., 2006; Vickery et al., 2014). If these trends hold, they may soon start breeding before peak food availability, which could lead to lower fitness of offspring.
We can already see evidence of how climate change is disrupting migrations and mutualistic relationships that were developed over thousands of years.
Resident species are also vulnerable to phenological mismatch. While these species might not be known for large-scale movements around the globe, they may still have to adjust their ranges to keep track of their climatic niches. Considering the improbability of different species will adapt at the same pace, there is thus a danger that important mutualistic relationships might be pulled apart these during range adaptations. This is of concern for species with specialised feeding niches, as seen in some pollinators. For example, studies from South Africa have shown how necessary range adjustments under climate change threaten both sunbirds—which show low adaptability (Simmons et al., 2004)—and their host plants, if specialised pollinator niches are left vacant (Huntley and Barnard, 2012). Extinctions arising from this decoupling of mutualistic relationships are referred to as coextinction (Koh et al., 2004), while a series of linked coextinctions is called an extinction cascade (Section 4.2.1).
Climate change and reptiles
One may think that reptiles—often seen basking on sun-drenched rocks to obtain active body temperatures—may benefit from climate change. Yet, as a group, they are also expected to suffer under climate change. One reason is because many reptiles will also have to adapt their ranges to shifting climates (Houniet et al., 2009). Even more important, climate change will increase reptiles’ vulnerability to demographic stochasticity (Section 8.7.2). Many reptiles—and some fish—have their sex determined by temperature during embryonic development, with warmer temperatures often leading to more females (Valenzuala and Lance, 2004). In general, females regulate their offspring’s sex ratios by fine-scale breeding site selection. Under climate change, however, it might be harder for the females to find breeding sites with suitable microclimates. This situation is of concern at South Africa’s iSimangaliso Wetland Park, where Nile crocodiles (Crocodylus niloticus, LC) are already struggling to find suitable breeding sites due to microclimate changes caused by invasive plant encroachment (Leslie and Spotila, 2001). Those species unable to adopt new mechanisms to control for offspring sex ratio bias may eventually go extinct, even under relatively small temperature shifts (Sinervo et al., 2010).
6.3.3 Climate change’s impact on freshwater ecosystems
With Africa’s freshwater ecosystems already strained by the demands of a growing human population, freshwater biodiversity will face several additional stressors associated with climate change. Climate change will impact water temperature, flow volume, and flow variability. Because these variables are three primary predictors of freshwater ecosystem composition (van Vliet et al., 2013; Knouft and Ficklin, 2017), it is expected that climate change will greatly affect freshwater ecosystem composition and functioning in the coming decades.
Warmer rivers and streams
Climatologists and hydrologists predict that freshwater ecosystems will generally experience temperature increases under climate change. These changes are already evident in Africa: for example, Lake Albert on the DRC-Uganda border, and Zambia’s Lake Mweru Wantipa, have experienced surface temperature increases of 0.62°C and 0.56°C respectively over the past decade (O’Reilly et al., 2015). Like their terrestrial counterparts, many freshwater species are sensitive to temperature shifts (e.g. Reizenberg et al., 2019). Warmer water also holds less dissolved oxygen, and increased pollutant toxicity (Whitehead et al., 2009). In addition, longer growing seasons and higher water temperatures will lead to a general increase in primary productivity and decomposition rates, which in turn will lead to increased nutrient loads, algae blooms, and eutrophication (Whitehead et al., 2009). All these factors will force many freshwater species—even those not sensitive to temperature shifts—to adjust their ranges to keep track of suitable conditions. Many of these adjustments will be impeded by habitat fragmentation, notably by dams and other human constructs that block suitable dispersal pathways. As an additional complication, many aquatic organisms cannot travel overland, so are naturally limited to adjust their ranges along the rivers and streams in which they live. But the orientation of these rivers and streams may not follow suitable thermal isolines: consider a cold-water species that needs to disperse to higher elevation—and hence upstream—as its climate niche moves higher up a mountain. For some freshwater species, the impediments to adjusting their ranges as necessary may be insurmountable.
Climate change will alter water temperature, flow volume, and flow variability, the three primary predictors of freshwater ecosystem composition.
Changing flow regimes
Changing precipitation levels will have several impacts on freshwater ecosystems, particularly as it relates to changes in their flow regimes (Thieme et al., 2010; Knouft and Ficklin, 2017). For example, areas that are undergoing decreased precipitation will experience decreased runoff and increased drying of wetlands and small streams, while areas with increased precipitation will experience increased storm surges and flushing. These changes, together with the impacts of increased water extraction rates and evapotranspiration under a warmer world, will cause significant changes in water levels, flow rates, sediment loads, water turbidity, and the structure of the physical environment. With an estimated 80% of Africa’s freshwater fishes predicted to experience significant flow regime changes (Thieme et al., 2010), the region will likely see substantial changes in the composition of freshwater communities in the coming decades.
Given these multiple stressors, there is a reasonable expectation that many freshwater species will go extinct or face significant population declines and range shifts over the next decades. These changes are of major concern in Africa, where so many people depend on fish and related natural resources for their livelihoods. Communities in Uganda, Malawi, Guinea, and Senegal are already finding it more difficult to meet their nutritional needs due to climate-induced freshwater fish declines (Allison et al., 2009). Also, at Lake Tanganyika—which supplies 20–40% of the surrounding countries’ dietary protein—fish yields have decreased by 30% in recent years, also attributed to climate change (O’Reilly et al., 2004).
6.3.4 Climate change’s impact on marine ecosystems
Like tropical forests, the world’s oceans have historically provided a relatively stable environment in which marine organisms have evolved. While this stability promotes species diversity, it also leaves marine species more vulnerable to environmental changes. In fact, a recent study found that cold-blooded marine species are twice as vulnerable to the impacts of warmer oceans than their terrestrial counterparts (Pinsky et al., 2019). In addition to the impacts of storm surges (Figure 6.6) and ocean warming (which leads to rising sea levels and ocean deoxygenation), marine organisms also must deal with ocean acidification. These threats will likely have impacts like those faced in terrestrial and freshwater ecosystems, including range adjustments, biotic attrition, and decoupling or important interactions. Below we discuss the mechanisms that will lead to some of these changes in more detail.
Ocean acidification
As discussed earlier, human activities release massive amounts of CO2 into the atmosphere each day. Although forests and other plant communities get considerable attention for CO2 sequestration, the world’s oceans also play a key role in keeping Earth’s carbon balance in check. In fact, the world’s oceans absorb an estimated 20–25% of our current CO2 emissions (Khatiwala et al., 2009). Now, with more atmospheric CO2 available, oceans absorb more carbon, which dissolves in seawater as carbonic acid. While this absorption may slow climate change, it also increases the acidity (i.e. lowing the pH levels) of the world’s oceans. This process—known as ocean acidification—has several consequences that may directly and indirectly kill marine organisms. For example, it inhibits the ability of coral animals to deposit the calcium used to build their reefs’ structure (Mollica et al., 2018), and prevents shellfish from accumulating adequate amounts of calcium carbonate to develop shells strong enough for survival (Branch et al., 2013). Ocean acidification also disturbs predator-prey dynamics by impairing the senses of prey species (Leduc et al., 2013), and compromising the ability of marine creatures to communicate with conspecifics (Roggatz et al., 2016).
Climate change is causing sea level rise and increased seawater temperatures, with broad implications for marine ecology and people living in coastal areas.
Sea level rise
Over the past 30–40 years, ocean surface temperatures have warmed by about 0.64°C (NOAA, 2016).Ocean warming has several implications, the most well-known being sea level rise, caused by the thermal expansion of ocean water combined with the released water from melting glaciers and polar ice caps. Current predictions suggest that sea levels in Sub-Saharan Africa will rise by 0.2–1.15 m over the next 100 years, compared to 2005 levels (Serdeczny et al., 2017). As the oceans creep further inland, the extent of low-lying coastal ecosystems such as rocky shores or sandy beaches will shrink, and so also the sizes of the wildlife populations living in those areas. The extinction of Australia’s Bramble Cay melomys (Melomys rubicola, EX)—the world’s first documented mammalian extinction caused by anthropogenic climate-change—has been attributed to sea level rise (Gynther et al., 2016).
Coral bleaching
The incredible diversity of corals reef ecosystems is attributable to the relative stability of tropical oceans. Because of this stability, individual coral species have adapted to very specialized niches. Many corals thus tolerate only narrow ranges in temperature, sunlight levels, water opacity, and nutrient loads. Climate change is disrupting this stability, by changing the temperature (ocean warming), depth (sea level rise), sediment and nutrient loads (increased erosion and runoff) of the environments where corals live. These changes are leading to a breakdown of critical mutualistic relationships between photosynthetic algae and corals. In the process, corals also lose their vibrant colours, revealing the corals’ ghostly white skeletons, hence the name coral bleaching (Figure 6.7). This relationship breakdown deprives the corals of essential carbohydrates they obtain from the algae, causing the corals to starve to death if the stressful conditions continue for a prolonged time.
Africa’s tropical oceans have experienced extensive coral bleaching events in recent years. For example, parts of Tanzania and Kenya have seen over 80% of their corals affected (McClanahan et al., 2007; Chauka, 2016). Coral bleaching also affects other species associated with coral reefs. For example, in the Seychelles, where coral bleaching at 70–99% of reefs was observed, butterfly fish exhibited a breakdown in territorial behavior, making it hard for them to breed and feed (Samways, 2005). In Zanzibar, Tanzania, eroded fish communities showed little signs of recovery multiple years after a bleaching event (Garpe et al., 2006).
Ocean deoxygenation
Marine fish and invertebrates rely on dissolved oxygen that enters the water either through the atmosphere, or by photosynthetic plankton. But because warmer water absorbs less oxygen, scientists predict that some areas of the ocean will see a 3–6% drop in dissolved oxygen concentrations under climate change (IPCC, 2014). This process, known as ocean suffocation or ocean deoxygenation (Ito et al., 2017), will leave parts of the ocean unsuitable for marine fishes and invertebrates. The impact of ocean deoxygenation will also be felt by economically important fisheries, notably along West Africa (Long et al., 2016) where climate change is predicted to lead to fisheries-related economic losses upwards of US \$311 million each year (Lam et al., 2012).
6.3.5 Climate change interacts with habitat loss
Habitat loss and climate change each cause negative impacts on biodiversity; however, these threats also interact to have an overall larger negative impact than the sum of these threats independently. Prominently, because of habitat loss, many species will be unable to adequately adjust their ranges to keep track of their shifting climatic niches. For example, some species might not be able to adapt their ranges because suitable habitat in their future ranges will be destroyed by human activity.
Climate change interacts with habitat loss, by impeding species’ ability to adapt, and by bringing dispersing wildlife into conflict with humans.
Range-shift gaps describes a habitat gap that prevents a species from dispersing from its current to future ranges (Figure 6.8). These gaps, which may occur naturally or because of habitat fragmentation, may also impede range adjustments under climate change. While the impact of range-shift gaps is an active area of research, it is expected that mountain-top species may be inherently vulnerable to range-shift gaps, particularly if they are unable to first disperse downslope before they can reach climatically suitable locations at higher elevation elsewhere. For example, over 60% of herbaceous plants living on Ethiopia’s Arsi Mountains might face range-shift gaps soon (Mekasha et al., 2013). But even highly mobile species might be vulnerable, with many African birds expected to face range-shift gaps as they adjust their ranges (La Sorte et al., 2014).
Habitat loss and climate change are also expected to exacerbate human-wildlife conflicts (Section 14.4). Sub-Saharan Africa will face losses of up to 2.5 million km2 in arable land between 2010 and 2100 (Zabel et al., 2014). These losses will see even more natural ecosystems converted for agriculture which, in turn, will further increase competition among and between humans and wildlife for resources such as food, water, and suitable habitat (Serdeczny et al., 2017). As the human footprint expands across Earth, agriculture and infrastructure will impede the ability of specialist species to find food and adapt to changing conditions, while generalist species will be forced into agricultural lands and nearby human habitation as they search for resources and/or disperse across the landscape. Such a scenario will likely exacerbate human-wildlife conflict in areas like Kenya’s Amboseli region, where lions living in fragmented ecosystems with diminishing natural prey populations are increasingly prone to wandering beyond protected area boundaries into ranching areas in search of food (Tuqa et al., 2014). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/06%3A_Our_Warming_World/6.03%3A_The_Impact_of_Climate_Change.txt |
To be clear, not all species will suffer equally from climate change. In fact, there are some species that will be resilient, and others that will even benefit from a warming world. Primary among the beneficiaries are plants in the northern areas of Europe, Asia, and North America (Zabel et al., 2014), and to a lesser extent in southern South America and New Zealand. In these areas, plants will benefit from longer growing seasons (earlier springs and shorter winters) and increased CO2 concentrations (which will increase photosynthesis rates).
Generalist species with high genetic diversity and that reproduce quickly are likely to benefit from climate change. Many species that exhibit these traits carry diseases and are agricultural pests.
Closer to home, a variety of African species are also expected to benefit from climate change. These include generalist species currently limited by interactions with localised specialists that are—at least at present—better competitors for limiting resources. Some tropical species may thrive as their habitats become hotter and wetter. Species with high genetic diversity that reproduce quickly (allowing for rapid adaption to environmental changes) are also likely to benefit. Unfortunately, many species that exhibit these traits carry diseases (Box 6.3) and are agricultural pests (Serdeczny et al., 2017). For example, populations of the coffee berry borer (Hypothenemus hampei)—Africa’s most notorious coffee pest—are expected to greatly increase in a warmer world (Jaramillo et al., 2011). This growing threat is particularly worrying given that higher temperatures have already reduced coffee harvests in countries such as Tanzania by as much as 50% (Craparo et al., 2015).
Box 6.3 Habitat Alteration, Climate Change, and Mosquito-Borne Diseases
Kevin Njabo
Center for Tropical Research,
UCLA Institute of the Environment and Sustainability,
Los Angeles, CA, USA.
[email protected]
With unprecedented climate change looming, mosquito-borne diseases, including malaria and dengue fever, will impact humans and wildlife in novel and unpredictable ways. While climate change is global in nature, changes due to habitat alteration are occurring more rapidly on a local scale, and are having significant effects on mosquito-borne diseases (Figure 6.C). For example, destruction of Peruvian rainforests unleashed more than 120,000 cases of malaria in the late 1990s, compared to fewer than 150 nine years earlier (Vitor et al., 2006).
The rainforests of the Congo Basin harbour roughly 20% of all known plant and animal species on Earth. Yet, habitat alteration continues at an alarming rate (Harris et al., 2012). Exacerbating these threats is the fact that Africa (Boko et al., 2007), and Central Africa in particular (McClean et al., 2006), are predicted to be some of the most severely affected by climate change. Predicted temperature increases would lead to longer seasons of malaria transmission and a 5–7% extension of the disease into higher latitudes (Craig et al., 1999, Boko et al., 2007). Coupled with projected population growth, climate change would nearly double the number of people at risk of infection from dengue fever by 2080. This is concerning because Africa is particularly vulnerable to environmental changes due to its limited adaptive capacity, widespread poverty, and low levels of development.
How then, will habitat alteration and climate change affect mosquito-borne diseases such as malaria? The relationship between disease transmission, habitat alteration, and climate change is complex. Though deforestation increases the risk of disease transmission (Vitor et al., 2006), different malaria-carrying mosquitoes (Anopheles spp.) are adapted to different microclimates. Ironically, our multi-faceted ecosystems both play the role of maintaining transmission cycles with cross-infections to humans and regulating those cycles while controlling spill-over into human populations. The balance between these factors is influenced by the availability of suitable habitat for mosquitoes and of reservoir hosts of infection. In an ideal world, transmission cycles are regulated by density-dependent processes such as acquired immunity to infectious diseases, and by limits on the carrying capacity of the environment to support insects and hosts.
Altered natural habitats and possible increases in disease transmission from animals to people also increase potential risks of new pathogens adapting to human hosts. Only about 2,000 of an estimated 1 million unique viruses carried by wild vertebrate species with potential zoonotic threats have been described. For example, when a lentivirus of chimpanzees first jumped into humans in the 1930s, not many people died. But the disease carved a foothold in the rapidly growing African city of Kinshasa in DRC and evolved into a form that efficiently preyed upon humans. More than 78 million people were infected between 1981 and 2015. To date, the disease it causes, AIDS, has killed more than 39 million people, while another estimated 37 million people are living with HIV.
Today, habitat alteration, such as deforestation, is not only driving species extinct and emitting lots of climate-changing carbon dioxide, it is also increasing opportunities for mosquito-borne diseases, such as malaria and dengue, to infect more humans in new places. Technological advances, including mathematical and computer modelling, genomics, and satellite tracking, will hopefully allow us to predict future disease outbreaks better. But we can also reduce outbreak opportunities by taking better care of our environment.
One group of species currently threatened with extinction that may benefit from a warmer world is marine turtles. Researchers working on Cabo Verde speculate that the island nation’s loggerhead turtle (Caretta caretta VU) populations will benefit from an increasing female-biased sex ratio (as expected under warmer conditions) given that a single male can breed with several females (Laloë et al., 2014). However, the researchers note that this population requires continued monitoring as insurance against demographic stochasticity (Section 8.7.2) that may become a larger threat under climate change.
Climate change has the potential to greatly restructure the world’s ecosystems, ecosystem services, and national economies. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/06%3A_Our_Warming_World/6.04%3A_Beneficiaries_of_Climate_Change.txt |
It should be clear to anyone that climate change has the potential to greatly restructure the world’s ecosystems, ecosystem services, and national economies. Many coastal areas will experience rising sea levels and increased flooding, while inland areas may experience desertification and less favourable crop growing conditions. Poor Africans will suffer the consequences disproportionately because of their limited mobility, high dependence on ecosystem services, and general lack of disaster management infrastructure (Serdeczny et al., 2017).
If we are to mitigate the far-reaching impacts of climate change, we must carefully monitor and study changes in biological communities and ecosystem functioning, and how they relate to changes in climate and other stressors. While we may lose some species in a warmer world, we can also prevent many extinctions with pro-active wildlife management (Section 11.4). It is likely that many existing protected areas will no longer preserve some of the rare and threatened species that currently live in them (Hole et al., 2009; Smith et al., 2016, but see Beale et al., 2013), necessitating careful planning when establishing new protected areas (Section 13.7.2). Even if climate change is not as severe as predicted, the steps we take now to safeguard biodiversity can only help in future.
In 2007, the world economy was close to collapse because of the misdeeds of the financial services industry. Considering climate change’s record in causing societal disruption and suffering, and our increasingly globalised world (in which regional disruptions are felt much wider than before), politicians are rightfully concerned about our ability to adapt to a widespread restructuring of the world’s natural resources (Dietz et al., 2016). While the consequences of climate change are closely associated with the environmental sciences, it is truly, at its core, a human rights concern.
The widespread and dramatic impacts of climate change rightfully deserve much attention. But it is also important to remember that we continue to destroy habitat at a massive scale and increasing pace, and this loss of habitat is currently the main cause of species extinctions. The highest priorities for conservation must continue to be the preservation of healthy, intact, and connected ecosystems, and the restoration of degraded ecosystems. These actions will simultaneously reduce the impacts of climate change, by reducing carbon emissions, increasing carbon sequestration, and giving wildlife more opportunities to adjust their ranges, in their own time, as the world’s climate changes.
6.06: Summary
1. While climate change is often thought of as a future challenge, we can already see its impacts today, as shown by record-high temperatures and changing rainfall patterns. These changes are happening because human activities release large amounts of greenhouse gases into the atmosphere on a daily basis.
2. Habitat loss contributes to climate change directly through the destruction of complex ecosystems (i.e. carbon sinks) which releases stored CO2, and indirectly through the loss of vegetation that would otherwise sequester CO2 from the atmosphere.
3. Some climatic shifts are predicted to be so rapid in coming decades that many species will be unable to adjust their ranges to keep up with environmental changes. Species with dispersal limitations, special habitat requirements, and important mutualistic relationships are at especially high risk of extinction.
4. Mitigating the negative impacts of climate change will require an international multi-pronged approach that includes ecosystem protection and restoration, direct species management, and legislative action.
5. Species are seldom exposed to only one threat; rather, different threats interact with climate change so that their combined impact is greater than their individual effects. A successful conservation strategy needs to deal with these threats collectively.
6.07: Topics for Discussion
1. Think of any particular ecosystem in your region. How do you think climate change will impact that ecosystem? What single measure do you think can be implemented to reduce the impact of climate change on that ecosystem? Can you think of the resources you will need to implement that measure?
2. Which groups of people and wildlife in Africa do you think will benefit the most from climate change and why? Who do you think will suffer the most and why?
3. How should we deal with species that have nowhere to go under climate change? Should we let them go extinct? What if it is an economically important species, like one that supports an important fishery or ecotourism industry? What do you think are our best options to save such species?
6.08: Suggested Readings
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Conservation biologists aim to preserve all the components, interactions, and processes within and between ecosystems, natural communities, species, and populations. The main obstacle to accomplishing this goal is habitat loss, while climate change will also play an increasingly important role. But let us for a moment consider widespread species and migratory populations. These species and populations typically live in different habitats and encounter different climates as they move across the landscape. We might think that tolerance for variety would make these groups robust against habitat loss and climate change. And yet, they are also declining, even in seemingly intact ecosystems and protected areas (Craigie et al., 2010; Lindsey et al., 2014). How can it be that populations apparently buffered from the two main extinction drivers are also subjected to population declines and extirpations?
Comprehensive conservation efforts must recognize that biodiversity faces multiple threats that need to be dealt with at different scales.
While habitat loss and climate change are the most prominent threats facing biodiversity at present, they are not alone. Nearly all human activities place additional pressures on populations, even those that already suffer from habitat loss and climate change. These additional pressures are primarily from pollution, overharvesting, persecution, invasive species, and disease (Maxwell et al., 2016). Because these threats are associated with and/or exacerbated by human activities, they can be dynamic in their nature, develop rapidly, and persist at such large scales that wildlife populations have little opportunity to adapt or move to safer areas. Moreover, these threats may interact with each other, as well as with climate change and habitat loss, so that their combined impact is greater than their individual effects. In this chapter, we explore how each of these threats impact wildlife and natural communities, and how they could push populations and species to extinction. Methods for lessening the impact of these threats are integrated into Chapters 9–15.
7.01: Pollution in Its Many Forms
Rachel Carson’s 1962 book, Silent Spring, described the dangers of pollution—pesticide pollution in particular—with a clarity that captured the public’s attention for many years afterwards. Carson, an American biologist, was particularly successful in drawing attention to biomagnification (also called bioaccumulation), a process through which pesticides and other toxins accumulate and become more concentrated in animals at higher levels of the food chain (Figure 7.1). Her work drew on research that found that dichlorodiphenyltrichloroethane (DDT), sprayed on crops to kill pest insects and on water bodies to kill malaria mosquito (Anopheles spp.) larvae, was also harming non-target organisms that consumed insects and fish exposed to DDT. Of note is that non-target organisms high on food chains, particularly fish-eating birds, such as eagles, pelicans, and egrets, often had high levels of DDT concentrated in their tissues. The affected birds were generally weakened, and the shells of their eggs were thin and prone to cracking during incubation. Consequently, bird populations declined dramatically in areas where DDT was used, as adults died and failed to raise young.
In the 1970s, many industrialized countries recognized the dire situation and banned the use of DDT, which eventually allowed for the partial recovery of the affected bird populations. Unfortunately, while some countries have switched to safer alternatives (e.g. Hargrove, 2003), DDT continues to be widely used in Africa to control malaria mosquito, tsetse fly (Glossina spp.), and other disease vectors. Researchers recently observed complete absences of breeding fish-eating birds in some African wetlands, and some of the highest-ever recorded DDT levels in seed-eating birds (Bouwman et al., 2013). This is cause for concern, not only for wildlife, but also for the long-term effects on people, particularly the consumers of the food products exposed to these chemicals (e.g. Manaca et al., 2011) and the workers who handle these chemicals in the field.
Pollution does not always lead to immediate mortality, but instead can have sublethal impacts that compromise organisms’ fitness over time, with population declines as the end result.
DDT is however not the only form of pollution we battle today. With the impacts of a growing human population becoming gradually more pervasive, pollution is compromising water, soil, and air quality at rates faster than ever before. Some forms of pollution can be highly visible, and with dramatic consequences (Figure 7.2). But importantly, there are many less detectable forms of pollution. While it may not always lead to immediate mortality, these insidious forms of pollution have sublethal impacts that compromise organisms’ fitness over time, with early death and population declines still being the end result. Responding to the silent threats of subtle and easily-overlooked pollution is often delayed, especially when the negative effects are felt only years after exposure. In their totality, pesticides and other pollutants claim 1.4–2.2 million human lives in Africa each year; globally, they claim 9 million lives, which is over three times more than the total impact of AIDS, malaria, and tuberculosis, together (Landrigan et al., 2018). Yet we continue to tolerate these threats, in part because the impact of pollution on our health is not always that apparent, especially when pollution deaths are expressed as a stroke, heart disease, respiratory infections, diarrhoea, or cancer, among other health issues.
One of the most challenging aspects when trying to prevent pollution is identifying the source. Many forms of pollution can easily be transported away from their source through the air, via rivers, even in groundwater. This transport of pollutants (called pesticide drift in the case of pesticides) means that a substantial burden (perhaps as much as 95%, Miller, 2004) of impacts are being felt by non-target species, including economically important non-target organisms. For example, pesticide drift from cotton fields in Benin has caused extirpations of freshwater fish (Agbohessi et al., 2015), while beneficial pollinating insects are also often negatively impacted (Pettis et al., 2013). Studies on fish in Nigeria (Adeogun et al., 2016), large mammals in South Africa (Bornman et al., 2010), and frogs in Kenya (Hayes and Menendez, 1999) have shown that beneficial organisms that survive this secondary pesticide exposure have disrupted reproductive and endocrine systems, and hence reduced fitness. Even humans may be exposed to secondary poisoning from pesticides, as toxic pesticide levels have been found in edible oysters and mussels in Ghana (Dodoo et al., 2013), prawn in Côte d’Ivoire (Roche and Tidou, 2009), and even chickens in South Africa (Barnhoorn et al., 2009).
To make matters worse, many pollutants take many years to biodegrade (i.e. break down in nature), and thus continue to pose a threat to wildlife and humans long after entering the environment. One important class of such long-lived pollutants is persistent organic pollutants (POP). Several types of pesticides qualify as POPs, which are prone to bioaccumulation and drift. The most famous POP is DDT; in the USA, biologist continue to see eggshell thinning and bird deaths, nearly 50 years after DDT was banned in that country (Burnett et al., 2013). This is a concern in places like Ethiopia’s Lake Koka, where recent studies have found DDT residues in every sample of fish tissue (from several different species) tested (Deribe et al., 2011). More information on POPs, many which are banned from use by signatories of the Stockholm Convention on Persistent Organic Pollutants, can be found on the Stockholm Convention website (http://pops.int).
There are also many types of persistent inorganic pollutants that find their way into the environment on a daily basis. One important class of persistent inorganic pollutants that also bioaccumulate is heavy metals; these include mercury, cobalt, copper, lead, and arsenic. A study from Zambia traced cobalt contamination in living trees to soil pollution from mining activities that occurred the mid-1970s (Mihaljevič et al., 2011). Some everyday products can also persist in the environment. For example, an aluminium can takes about 200 years to break down, while a plastic bag takes between 100–1,000 years to break down. The continued use of these products should thus raise alarm to anyone concerned about the environment and human health. But it also provides opportunities for any person to contribute to conservation by reducing use of these products and reusing/recycling those products that find their way into the supply chain.
Many pollutants take many years to biodegrade, and thus continue to pose a threat to wildlife and humans long after entering the environment.
Water pollution
Water pollution, the accidental or intentional dumping of pesticides; herbicides; oil products; fertilisers; sewage; industrial waste; detergents; and other foreign chemicals and objects into aquatic environments, is arguably the biggest current pollution concern in Africa (Prüss-Ustün et al., 2016; Landrigan et al., 2018).
The dumping of products containing heavy metals into aquatic environments is particularly concerning because heavy metals are toxic even in small concentrations, and likely to biomagnify. When aquatic organisms process contaminated water, they absorb or ingest the heavy metals along with other essential nutrients. With each additional step along the food chain, organisms ingest and accumulate increasingly higher concentrations of these toxic elements (see Figure 7.1). In this way, even small amounts of heavy metals can become lethal across several levels of the food web over time. Biomagnification is especially a concern with long-lived predatory marine fishes that people consumed as food, such as swordfish (Xiphias gladius, LC), marlins, sharks, and some tunas and sea basses. For example, mercury (emitted mainly during fossil fuel use), lead, and arsenic have bioaccumulated so much in sharks off South Africa that many species are now considered unsafe for human consumption (McKinney et al., 2016; Bosch et al., 2016; Merly et al., 2019). Recent studies also found unsafe levels of mercury in freshwater fish from regions as wide as Central Africa’s Great Lakes (Campbell et al., 2008), Ethiopia’s Lake Awassa (Desta et al., 2006), and several reservoirs in West Africa (Quédraogo and Amyot, 2013).
Because of biomagnification, many long-lived predatory marine fishes are now considered unsafe for human consumption.
Oil pollution involves the release of petroleum products into the environment, which can originate from damaged ships, failed drilling rigs, leaking offshore platforms, or other unexpected events. The released oil causes mammals and birds to lose the insulating abilities of their fur and feathers, leaving those animals vulnerable to hypothermia and drowning. Other aquatic animals, including fish and shellfish, may ingest oil products, causing them to sicken and die. Because of the way oil is extracted and transported, marine ecosystems are particularly at risk. Furthermore, because of the massive amount of oil that are involved in oil extraction and transport, an oil pollution event often represents a serious ecological disaster (Figure 7.3). Africa has been hit hard by oil spills in recent years, particularly around oil-producing countries like Angola and Nigeria, and along shipping lanes passing along the coasts of Namibia, South Africa, and Mozambique. Nigeria is perhaps the biggest victim of oil spills; between 1976 and 2001, there were an estimated 6,817 oil spills around Africa’s largest wetland, the Niger Delta (UNDP, 2006)! These oil spills have destroyed thousands of hectares of mangrove swamps, estuarine wetlands, and other coastal ecosystems, causing severe hardship to marginalised local communities who depended on those areas for subsistence fishing and farming (Fentiman and Zabbey, 2015).
Plastic pollution is fast becoming a ubiquitous threat to Africa’s environment, its wildlife, and its people. To visualise the magnitude of the problem, consider that there are more than 1.6 trillion pieces of plastic, collectively weighing over 70,000 tonnes, currently floating in the Atlantic and Indian Oceans surrounding Africa (Eriksen et al., 2014). While many of these plastic items were dumped directly in the ocean, many also have a terrestrial origin. For example, if someone throws a plastic wrapper on a sidewalk, there is a good chance that the wrapper will find its way into a nearby stream at some point, carried by wind or rain runoff. From here, the wrapper will float along various streams and rivers until it reaches the ocean. A recent review found that 88–95% of plastics floating into the world’s oceans originated from just 10 rivers, which include West Africa’s Niger River and East Africa’s Nile River (Schmidt et al., 2017; Lebreton et al., 2017). In the process, thousands of seabirds, dolphins, whales, turtles, seals and fish die each year from suffocation or starvation after ingesting plastics and other pieces of trash that they confused with food (Wilcox et al., 2015). This plastic pollution also impacts humans: researchers recently found microfibers (many of which are plastic) in over 80% of tap water samples from Uganda (Kosuth et al., 2017), as well as food-grade commercial sea salt originating from South Africa (Karami et al., 2017).
There are more than 1.6 trillion pieces of plastic, collectively weighing over 70,000 tonnes, currently floating in the Atlantic and Indian Oceans surrounding Africa.
Some of the biggest impacts from plastic pollution are caused not by visible scraps of plastic, but by microplastics, the collective name for plastic particles smaller than 1 mm (some are microscopic). Microplastics may originate from the breakdown of larger pieces of plastic and polystyrene products, or they may be manufactured intentionally small, such as beads added to cosmetics and other personal care products that are flushed down drains after use. Because microplastics are so small, they easily pass through the standard filters used at sewage treatment plants. Consequently, microplastics generally end up in the aquatic environment, where they are unintentionally consumed by crustaceans (crabs, lobsters, and krill), molluscs (mussels, oysters, and clams), echinoderms (sea stars, sea urchins, sea cucumbers), and baby fish. This consumption can block or damage the victim’s digestive and respiratory systems, cause reduced food uptake by creating a false sense of satiation, or even poison animals through leeching of synthetic chemicals. Each of these threats increases death rates and lowers reproductive rates (Sussarellu et al., 2016). Just as with the biomagnification we discussed earlier, the consumption of microplastics also affects other consumers (including humans), because the small organisms that ingest the microplastics are often food for other animals, allowing plastic pollution to move through an entire food chain. For example, a recent study from Lake Victoria found microplastics imbedded in the digestive tracts of perch and tilapia bought at a local market and meant for human consumption (Biginagwa et al., 2016). Because microplastics are so hard to remove once in an ecosystem, the best method for their containment may be to reduce plastic use, to ban products containing microplastics, or to develop microplastics that are biodegradable within a reasonable timeframe. But for this to happen, there is a need to educate the public and lawmakers (Galloway and Lewis, 2016) about the dangers posed by this threat to the environment and local economies.
Nutrient pollution, caused in part by excessive fertiliser use, can led to eutrophication, famous for causing algae blooms, aquatic dead zones, and fish kills.
Nutrient pollution represents another growing threat to Africa’s aquatic environments. Many lakes, streams, and other freshwater and marine environments naturally contain low concentrations of essential nutrients, such as nitrates and phosphates. In order to survive, the species living in these nutrient-poor waters must then be adapted to this natural nutrient scarcity. However, raw sewage, agricultural fertilisers, concentrated animal feeding operations, and industrial processes release large amounts of additional nitrates and phosphates into the environment, which are washed into the aquatic environment. Minor additions of essential nutrients stimulate plant growth, providing more food for organisms at higher trophic levels. However, at high concentrations, the system become subjected to nutrient pollution.
One of the worst outcomes of nutrient pollution is eutrophication. During eutrophication, surface algae grow so rapidly (known as an algae bloom) that it starts blocking sunlight from reaching aquatic organisms below the surface. Because each individual alga is short-lived, their rapid growth also adds large amounts of decaying matter to the environment. In response, decomposers that feed on the dead algae can become so abundant that they consume most of the water’s dissolved oxygen. Without oxygen and sunlight, aquatic plant and animal life may die off in large numbers. The resultant dead zones are sometimes visibly in the form of fish kills, with large numbers of dead fish floating on the surface of the affected water body. The organisms that die during this process is generally also toxic to humans because of bacteria build-up and other imbalances. Eutrophication is an increasingly common problem in Africa; for example, a recent review found that 41–76% of South Africa’s lakes may be eutrophic (Harding, 2015). Eutrophication has already negatively impacted Africa’s tourism and fisheries sectors (Nyenje et al., 2010), and even led a temporary shutdown of water supplies on the Kenyan side of Lake Victoria (Sitoki et al., 2012). Preventing further eutrophication should thus be a high priority—not only will it prevent harmful algae blooms but may even play an important role in controlling invasive aquatic plants such as the water hyacinth (Eichhornia crassipes) (Coetzee and Hill, 2012; Bownes et al., 2013).
Groundwater pollution—the release of pollutants into aquifers and other sources of groundwater—is also becoming a serious issue across Africa. This type of pollution generally originates from landfills, on-site sanitation systems, leaking sewage systems, mining leachate, agriculture runoff (fertiliser, pesticides, animal waste, etc.), and other types of waste dumping. The pollutants may sometimes be released directly into aquifers; however, more often the contaminants and pathogens leak into the soil, from where it seeps into groundwater.
Because fracking poses many serious risks, governments across the world have banned the practice from their lands.
One of the most important emerging threats to groundwater in Africa is hydrological fracturing or fracking, in short. During this process, pressurised liquids that contain suspended particles and thickening agents are blasted into rock formations deep underground to break them open. When the pressure and liquids are removed, the suspended particles keep the fractures open, which enables extraction of natural gas and petroleum. While fracking was initially hailed as a method to access previously inaccessible fossil fuels, scientists subsequently found that it poses a wide variety of very serious environmental and health risks. Most importantly, the liquids used in fracking contain toxic chemicals which pose a high risk for groundwater pollution (Osborne et al., 2011), which in turn lead to miscarriages and birth defects (McKenzie et al., 2014), cancer (McKenzie et al., 2012), as well as skin and respiratory diseases (Rabinowitz et al. 2015). In addition, fracking increases greenhouse gas emissions (Howarth, 2014) and induces infrastructure-damaging earthquakes (Ellsworth, 2013). Because of these myriad serious risks, several national governments in Europe, and several local governments in the USA, UK, Canada, and Australia have banned the practice from their lands (https://keeptapwatersafe.org/global-bans-on-fracking). In contrast, and despite opposition from civil society, several countries in Africa (e.g. South Africa: Roelf, 2016; Botswana: Barbee, 2015) recently approved this harmful practice.
Air pollution
In the past, people and industries thought that the atmosphere was so vast that any gases or particles released into the air would disperse and dilute to the point that they would post no ill effects. But as air quality has diminished over time, scientists have documented that air pollution can cause irreparable harm to ecosystems and human health, often far from the original sources. A striking example comes from West Africa’s Lake Chad, which shrank by 95% between 1963 and 1998 (Figure 7.4). Experts generally thought that the shrinkage was caused by unsustainable water use in the region, but recent evidence suggests that air pollution from Europe which reduced rainfall in the Lake’s catchment area may also have contributed to this ecological disaster (Hwang et al., 2013). The Lake’s water level has risen since 2007, likely due, in part, to clean air regulations implemented by the European Union. Despite this positive turn around, air pollution continues to be a serious problem (Amegah and Agyei-Mensah, 2017) that threatens humans and wildlife throughout Africa.
An important form of air pollution is hydrocarbons, which are released during fossil fuel burning, particularly during transport, power generation, and other industrial activities (Karagulian et al., 2015). Pollution from airborne hydrocarbon compounds can sometimes be sensed without scientific equipment, by the bad smells, high air turbidity, and eye and lung irritation a person may experience in large cities with highly polluted air. When exposed to sunlight, these chemicals can react with other gases and particles in the atmosphere to produce photochemical smog, which is made up of ozone and other secondary compounds. In the upper atmosphere, ozone filters harmful ultraviolet radiation, which benefits most living things; but at ground level, high concentrations of ozone pose several dangers. For example, it damages plant tissues which make them brittle; high surface ozone levels have found to cause crop damage in Botswana and South Africa (Zunckel et al., 2004). Hydrocarbon exposure also poses several threats to humans: it altered some people’s DNA—often a cancer precursor—in Benin (Fanou et al., 2006), caused lung damage in Côte d’Ivoire (Kouassi et al., 2010), and subjected people to carcinogenic compounds in the DRC and Ghana (Tuakuila, 2013; Bortey-Sam et al., 2017). The lack of air monitoring and standards over much of Sub-Saharan (Petkova et al., 2013), and lack of awareness—people often confuse photochemical smog with natural mist and early-morning fog—should thus be of serious concern both to conservation biologists and society at large.
Air pollution from hydrocarbons often manifests itself as photochemical smog. Hanging like a thick cloud over industrial areas, people sometimes confuse it with natural mist and early-morning fog.
Burning fossil fuels also releases sulphur oxides (SOx) and nitrogen oxides (NOx) into the atmosphere, where they combine with water vapor to produce nitric and sulphuric acids. These acids later return to the ground as acid rain, with dramatically low pH relative to normal rainwater. Prevailing winds can transport acid rain clouds over long distances, so the effects of acid rain may occur hundreds of kilometres from its sources. Because the acid rain is closely tied to the water cycle, aquatic and soil organisms are particularly vulnerable to the negative effects of acid rain. Plants exposed to acid rain, either directly or after absorbing contaminated water from the ground, are often left severely weakened or even killed: it has even caused plant extirpations in Zambia (UNEP, 2006).
Another important contributor to air pollution is domestic fuel burning (Karagulian et al., 2015). During these activities, very small pollutant particles are released into the air. Because these particles are so small, they are difficult to filter from the air, and can easily be inhaled. Once inhaled, the particles can pass into the victim’s bloodstream, from where they negatively impact cardiovascular health, neurodevelopment, and cognitive function (WHO, 2013). Despite the harmful impact of these particles in the environment, their monitoring is virtually non-existent in Africa, making it very hard to guide air quality policy decisions and legislations. In contrast, measures that mitigate pollution from domestic fuel burning may even help slow the rate of habitat loss (Chapter 5), as this type of pollution is associated with inefficient wood stoves, slash-and-burn agriculture, and the artisanal charcoal industry.
Soil pollution
Soil pollution occurs when soil meets foreign chemicals and other pollutants. This type of pollution is often associated with industrial activities that extract resources from the earth, agricultural runoff, pesticide use, oil spills, acid rain, improper treatment of sewage, and improper disposal of waste. People and wildlife can then become sick through direct contact with contaminated soils, or through secondary contamination via polluted groundwater or eating food grown in contaminated soil. For example, a recent review reported how soil pollution has left medicinal plants toxic to humans in countries such as Botswana; Ghana; and Mali, at times with fatal consequences (Street, 2012).
The improper disposal of electronic waste (or e-waste in short) is a particularly serious form of soil pollution. Because electronic products contain toxic heavy metal contaminants that are expensive to recycle, discarded electronic products usually end up in dump yards (Figure 7.5). Here, open burning of electronic and other waste materials releases the toxic compounds into the soil, as well as the air and water (Robinson, 2009), from where it also accumulates in the environment.
Light pollution
Light pollution describes the addition of excessive, ill-timed, or poorly designed artificial light to the natural world. A consequence of an increasingly industrialized world (Falchi et al., 2016), light pollution has increased dramatically over the past decades as more people have gained greater access to electricity (Figure 7.6). Behavioural disruption is perhaps the most well-known consequence of increased light pollution—consider all the moths and other nocturnal insects (and insect predators, such as bats and geckos) attracted to artificial night lights. Light pollution also interferes with the navigation abilities of nocturnal species, which often use the stars, moon, and light reflectance from water surfaces to orientate themselves. For example, work in Gabon has shown how artificial lights disorientate sea turtle hatchlings trying to reach the sea (Bourgeois et al., 2009), while others have highlighted the significance of light-induced seabird mortality (Black, 2005). These and other behavioral disruptions—which include attraction to and repelling away from artificial light—may seem to only affect a small number of individuals around a few lights in your home. But the systemic impact of thousands of lights every night has wide-ranging ecosystem impacts when considering the cumulative impact of reduced reproductive performance (Firebaugh and Haynes, 2016), disrupted predator-prey dynamics (Minnaar et al., 2015) and disturbed night-time pollination services (Knop et al., 2017) on the many thousands of organisms impacted every night.
Light pollution also disrupts the natural day-night cycles with which most species evolved. These disruptions interfere with circadian rhythms, which negatively affect living organisms’ physiology. For example, one study showed that night-time light pollution disrupted natural sleep patterns in birds, leaving the affected individuals more susceptible to malaria infections (Ouyang et al., 2017). Circadian rhythm disruptions from light pollution (especially from high frequency “blue” light) also impact humans by increasing stress, fatigue, and anxiety, and susceptibility to obesity (Rybnikova et al., 2016) and cancer (Haim and Portnov, 2013). It is important to note that light pollution does not mean that the use of light is inherently bad—light has and will continue to play an important role in our daily lives. However, it does mean that we need to be more thoughtful about the consequences of light pollution and put measures in place to mitigate its impacts on the natural world and our own lives.
Noise pollution
Many people find a sense of freedom when they are in natural surroundings, with peace and quiet facilitating a much-needed connection to nature. These experiences are increasingly being threatened by noise pollution. However, noise pollution (also called acoustic pollution)—caused by human activities, such as industrial, military, and transportation systems—affects more than just the appealing tranquillity of nature. It also prevents animals from hearing each other, predators, and prey, all which could interfere with feeding, reproduction, navigation, and predator-avoidance behaviors. While African studies on the impact of noise pollution on wildlife are near-absent (Shannon et al., 2015), one study that did investigate the topic found that traffic noise increased dwarf mongooses’ (Helogale parvula, LC) alertness but also reduced responsiveness to alarm calls (Kern and Radford, 2016). Such responses could leave the affected individuals less fit and more vulnerable to predators.
Noise pollution prevents interferes with communication, feeding, reproduction, navigation, and predator-avoidance behaviors; it may even contribute to mass strandings of whales.
One would think that marine organisms living in the vast oceans may be spared from noise pollution, but this is not the case (Koper and Plön, 2012; Kunc et al., 2017). Sound carries much further in salt water than air, so noises from ship propellers; military sonar; seismic activities, and construction have significantly increased the level of ambient noise levels marine organisms experience. This increased level of ambient noise not only disrupts communication in sea animals (e.g. Cerchio et al., 2014), but can even lead to death (some mass whale strandings have been attributed to noise pollution: Morell et al., 2017; Williams et al., 2017). As with light pollution, there is a general need to be more thoughtful about the consequences of sound pollution on the natural world and to put measures in place (see e.g. Koper and Plön, 2012) to mitigate its impacts.
Thermal pollution
Thermal pollution describes localised human-induced temperature changes to the natural world. Aquatic ecosystems represent one of the ecosystems most vulnerable to thermal pollution. For example, when water is released from big dams, it comes from the colder middle and lower strata of the reservoir, leading to rapid cooling of aquatic ecosystems further downstream. The opposite is true at power plants that use river water as a coolant; turbines release their heat to the circulating water and then the warmed water is released back into the environment. These abrupt releases of thermally discordant water often lead to thermal shock which can be lethal to fish and other aquatic organisms. For example, studies from South Africa have shown that thermal shock can kill fish embryos and larvae and caused deformities in the young of Clanwilliam yellowfish (Barbus capensis, VU) (King et al., 1998).
The urban heat island effect represents a terrestrial form of thermal pollution. Urban and other developed areas are generally covered with large swaths of man-made surfaces (e.g. asphalt roads, pavement surfaces, and building roofs), which absorb solar energy rather than reflect it. This absorbed heat, in combination with heat outputs from industrial activities, cause urban areas to function like “islands of heat” that are several degrees warmer (Figure 7.7) than surrounding rural areas (Feyisa et al., 2014; Chakraborty and Lee, 2018). The urban heat island effect reduces the quality of life for people and wildlife by reducing comfort and water availability (due to increasing evaporation). It also increases energy consumption to offset the heat increases which, in turn, contributes to air pollution and climate change. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/07%3A_Pollution_Overharvesting_Invasive_Species_and_Disease/7.00%3A_Prelude_to_Pollution_Overharvesting_Invasive_Species_and_Disease.txt |
People have always hunted, collected, trapped, or otherwise harvested the food and other natural resources they need to survive. When human populations were small, at least relative to the abundance of their resources, and collection methods were relatively unsophisticated, people could sustainably harvest and hunt wildlife in their local environments. However, as human populations have increased, and roads have provided access to previously remote areas, our impact on the environment has escalated. At the same time, our methods of harvesting have become dramatically more efficient. Guns are now used instead of blowpipes, spears, or arrows, while networks of wire snares indiscriminately catch animals of all types, even young and pregnant females. Populations of species that mature and reproduce rapidly can often recover quickly after harvests and can thus be exploited sustainably; however, species that are slow-maturing and slow-reproducing cannot sustain current harvest levels. Consequently, many species are threatened due to overharvesting, the unsustainable collection of natural resources (Maxwell et al., 2016). Overharvesting may take on many forms, including hunting, fishing, logging, and gathering of plants and animals for medicine, captive collections, subsistence, commerce, or recreation purposes (Figure 7.8).
The Bushmeat Crisis
Bushmeat harvesting is one of Africa’s most prominent overharvesting concerns (see Box 4.1). Bushmeat—wild sources of protein obtained on land by hunting and collecting birds, mammals, snails, and caterpillars—provides much of the protein in people’s diets in large parts of Africa. For example, in Nigeria and Cameroon, 12,000 tonnes of bushmeat—two tonnes obtained from bay duiker (Cephalophus dorsalis, NT) alone—are sold at markets in the Cross-Sanaga rivers region each year (Fa et al., 2006). Similarly, more than 9,000 primates are killed annually for a single market in Côte d’Ivoire (Covey and McGraw, 2014); people from Central Africa harvest an astonishing 5.3 million tonnes of mammalian bushmeat annually (Fa et al., 2002). Usually seen as a conservation challenge in Africa’s tropical forests, the bushmeat crisis also impacts savannah regions (reviewed in Lindsey et al., 2013). For example, bushmeat hunters, numbering between 1.500 and 2,000, remove over 600,000 kg of herbivore biomass from Botswana’s Okavango Delta each year, despite the region’s protected status and importance for ecotourism sectors (Rogan et al., 2017).
The massive wildlife declines caused by the bushmeat crisis are also threatening ecosystem services, food security and people’s livelihoods.
Outside influences play a prominent role in the harvesting pressure associated with the bushmeat crisis. In Section 5.2, we discussed neocolonialism, where jobs associated with land-grabbing industries are frequently reserved for migrant labourers. Poorly paid and with limited rights, migrant labourers are often forced to turn to local natural resources to fulfil their basic needs (Thibault and Blaney, 2003). The impact of these migrant labourers on the local environment is massive compared to traditional (and many other local) peoples that prioritise sustainability. For example, immigrants working at logging concessions in the northern parts of the Republic of the Congo hunt 72% of all bushmeat harvested in the region (Poulsen et al., 2009). The increased commercialisation of bushmeat also poses challenges (Lindsey et al., 2013); for example, in the broader Congo Basin, commercial hunters are exploiting bushmeat at scales 27 times that of the area’s traditional peoples (Fa et al., 2016). In addition to hunting for local markets, illegal exports also play an important role. For example, more than 50 tonnes of wild fish and bushmeat enters France from Africa each week (Chaber et al., 2010); similar amounts were also estimated for airports in Switzerland (Wood et al., 2014).
Very few animal populations can withstand such high extraction rates. Consequently, regions dependent on bushmeat have already seen substantially wildlife declines (Lindsey et al., 2013). West Africa, where forest mammal populations are down an estimated 80%, have been hot particularly hard (Benítez-López et al., 2019). These wildlife declines also lead to reduced harvests—some hunters have seen their harvests reduced by over 80%, with impacts to wildlife notable as far as 40 km from hunters’ access points along roads from their villages (Benítez-López et al., 2017). At current exploitation rates, supplies are expected to decrease by an additional 80% within the next 50 years (Fa et al., 2003). Unless more sustainable, alternative sources of protein are found, people dependent upon bushmeat will see increased malnutrition and compromised livelihoods as bushmeat species are pushed to extinction. When that happens, families relying on bushmeat will face even worse food insecurity than that which is driving the current bushmeat crisis.
Exacerbating the risk of food insecurity, people in the affected regions will also suffer from compromised ecosystem services as populations of predators, seed dispersers, and pollinators are reduced (Rosin and Poulsen, 2016). For example, reduced mammal populations have been linked to reduced abundance of fruits and other useful plant products available for human consumption (Vanthomme et al., 2010). Some areas are already suffering from “empty forest syndrome”—a condition where a forest appears to be green and healthy, but is practically devoid of animals, and in which ecological processes have been irreversibly altered such that the forest’s species composition will change over subsequent decades (Nasi et al., 2011; Benítez-López et al., 2019). The bushmeat crisis is thus a major concern to people concerned about biodiversity and/or human well-being.
Overfishing
Pressure on biodiversity in aquatic environments is also increasing as people continue to harvest fish, sea turtles, dolphins, shellfish, and manatees for meat at increasing rates. Modernised fishing methods play a major role. For example, a motorised fishing fleet that faces few restrictions has caused a 75% decline in fish populations at Ethiopia’s Lake Tana in recent years (de Graaf et al., 2004). Also, in the marine environment, motorised fleets and enormous factory ships can now spend months at sea where they catch fish to sell at local and global markets (Ramos and Grémillet, 2013; Pauly et al., 2014). Some estimates suggest that wild-caught seafood could be virtually absent by 2050 if current exploitation levels persist (Worm et al., 2006).
For many aquatic organisms, the indirect impacts of modern commercial fishing methods outweigh direct exploitation (Figure 7.9). One example is ghost fishing, which causes thousands of animals to die each year after becoming entangled in dumped, abandoned, and lost fishing gear. Similarly, approximately 25% of fish harvests are considered bycatch—animals that are accidentally caught, injured, or killed during fishing operations. Recent declines in skates, rays, turtles, sharks, dolphins, and seabirds have all been linked to incidental deaths as bycatch (Cox et al., 2007; Carruthers et al., 2009). Seabird biologists from South Africa have been at the forefront of solving bycatch problems in recent years (Box 7.1).
Box 7.1 Solving Seabird Bycatch Problems: From Theory to Practice
Ross Wanless
DST-NRF Centre of Excellence at the FitzPatrick Institute of African Ornithology,
University of Cape Town, South Africa.
[email protected]
The global problem of seabird bycatch in fisheries—the accidental death of seabirds during fishing—is one of the biggest threats to pelagic seabirds (Croxall et al., 2012). Ironically, it is both one of the easiest and one of the most challenging problems to solve. How so? Simple technical fixes to stop birds from getting snagged on fishing gear and drowning can work amazingly well, but fishermen must be convinced to use them.
Techniques for preventing bycatch break down into two basic approaches. The first approach is to prevent access to the danger point (the baited hook or the cables that birds strike). Fishing only at night eliminates up to 80% of the problem, but still jeopardises nocturnal foragers and diurnal species during a full moon. Another option is bird-scaring lines (Figure 7.A) consisting of a mainline flown from the stern of a boat with hanging streamers that scare birds away from danger areas behind the vessel. The second approach, primarily used in longline fishing, is to remove the risky gear (baited hooks) as quickly as possible; essentially this involves adding weights to lines to sink them faster. It has reduced seabird bycatch on some fisheries by 90–95%.
Despite clear harm to seabirds caused by fisheries and the simple, effective fixes at hand, implementing these mitigation measures has been patchy at best in most fisheries where seabird bycatch occurs. There are some exceptions, and it is useful to examine what elements led certain fisheries to fix the problem. A good case study is in South Africa, where BirdLife South Africa’s Albatross Task Force (ATF) demonstrated in 2006 that trawl fishing for hake (Merluccius spp.) was killing around 18,000 seabirds each year (Watkins et al., 2008). The fishery involved had Marine Stewardship Council (MSC) certification, which gives a fishery access to premium European markets on the condition that it meets environmentally friendly and sustainable metrics, including no significant bycatch impacts. This provided a powerful incentive for fleet-wide implementation of a bycatch mitigation measure; failure to do so would have resulted in a loss of certification, with catastrophic financial implications.
Despite this strong incentive, it required another five years of work from the ATF to assess bird scaring lines and refine the design, overcome resistance to their use, and close loopholes in regulations. In 2014, the ATF published an assessment of the effectiveness of their bird scaring lines—a single measure to prevent the accidental and avoidable deaths of around 10,000 albatrosses and large numbers of other species. When used correctly, the system eliminated 90–95% of seabird bycatch (Maree et al., 2014). Why did it take so long for the fleet to adopt this measure, despite it costing almost nothing, requiring no skill or time to use, and posing no meaningful operational problems? And why have identical fishing industries in many other countries failed to follow suite?
The answer is complex. ATF teams are present in South Africa (and now also in Namibia), providing sustained pressure and constant presence. South Africa had standing legislation, yet compliance from the South African fleet was initially minimal (and remains less than perfect today). MSC certification certainly created an enabling environment (Wanless and Maree, 2014) and incentive to drive change, yet it took more than that to change the entire fleet. Constant lobbying from BirdLife and regular dialogue from deck to boardroom were also critical ingredients. A legislative framework that provides some hope of censure against non-compliant vessels meant that there was internal pressure within the industry to “tow the line”—pun intended. Ultimately, widespread change became possible when there was a credible, independent observer program to verify deck practice and give teeth to agencies when addressing non-compliance.
The impact of traditional medicine
Africa has a long history of sustainable use of traditional medicines. Unfortunately, as the number of people living in Africa has increased, so has the demand for traditional medicine. Today, harvesting for traditional medicine is putting unsustainable pressure on species exploited for this purpose (Williams et al., 2014). One prominent example is vultures: the demand of vulture body parts, believed to bestow clairvoyant abilities, is driving massive vulture population declines across Africa (see Box 4.4). The growth in traditional medicine markets in East Asian countries such as China, Thailand, Cambodia, and Vietnam exacerbates these problems. For example, as tigers (Panthera tigris, EN) and rhinoceros have become scarce in Asia, Asian traditional healers are increasingly targeting African predators and rhinoceros to satisfy their market demands. Another group of species threatened by the Asian traditional medicine trade is sea horses (Hippocampus spp.). Due to population declines from overharvesting, sea horse exports from Kenya and Tanzania to East Asia have halved over recent years; yet, more than 600 kg of dried sea horses (over 254,000 individuals) continue to be exported annually (McPherson and Vincent, 2004). Exploitation for Asian traditional medicine markets has already pushed the western black rhinoceros (Diceros bicornis longipes, EX) to extinction. In a similarly perilous position is the northern white rhinoceros (Ceratotherium simum cottoni, CR); with only two non-reproductive females left in the world, this species is now considered committed to extinction (see Section 8.3; Box 11.4). A group of species sought after by both African and Asian traditional medicine markets is pangolins, thought to be the most heavily poached animals on Earth. For example, between 2012 and 2016, more than 20 tonnes of African pangolin scales (involving up to 30,000 animals) were seized during law enforcement operations across the region (Andersen, 2016). The problem is also getting worse: authorities intercepted 13 tonnes of scales in Singapore in 2019, all from a single shipment believed to travel from Nigeria to Vietnam (Geddie, 2019). With such a large active operational scale, it comes as no surprise that all four African pangolin species are now threatened with extinction (IUCN, 2019).
The impact of live animal trade
Millions of non-domesticated animals are sold as pets around the world each year (Table 7.1). Given that many of these pets were originally collected in the wild, it is no surprise that the most popular species tend to be at a high risk of extinction (Bush et al., 2014). These huge numbers are magnified by the extra millions of animals needed to compensate for deaths during collection and shipping. Collection of wild animals for pets and other purposes has a massive impact of biodiversity in Africa, the world’s largest pet trade exporter (Bush et al. 2014).
Among the most popular groups of wildlife traded are Africa’s parrots (Figure 7.10). For example, 32,000 wild-sourced African grey parrots (Psittacus erithacus, EN) were imported into the European Union in 2005 (UNEP-WCMC, 2007). Combined with habitat loss, the wild bird trade has already caused extirpations of this species in some areas of West Africa (Annorbah et al., 2015). Similarly, 82 of the 291 species of African freshwater fish known to occur in the pet trade are considered threatened with extinction (UNEP-WCMC, 2008). While it is true that collecting wild animals for the pet trade sustains many people’s livelihoods, research on harvesting of ornamental fish in Cameroon has shown that this practice is not sustainable in the long term (Brummet et al., 2010). It is therefore critical to find ways to make these practices more sustainable, for the sake of the pet collectors and biodiversity.
Table 7.1 Examples of groups targeted in global wildlife trade, and their levels of exploitation.
Group
Number traded each year
Notes
Orchids
250 million
Mainly cultivated, but about 10% sourced from the wild. Illegal trade—and mislabelling to avoid regulation—a major problem.
Succulent plants
35 million
Mainly cultivated, but about 15% sourced from the wild. Illegal trade remains a major problem.
Corals
13 million
Collected using destructive methods; used for aquarium decor and jewellery.
Reptiles
7.2 million
Mainly sourced from the wild for zoos and pet trade, but increasingly from farms. Does not include large skin trade.
Birds
2.3 million
Mostly perching birds destined for zoos and pet trade. Also includes legal and illegal trade of parrots.
Ornamental fish
2 million
Most originate from wild reefs, caught by illegal methods that damage the surrounding coral reef and other wildlife.
Primates
148,000
Used for biomedical research, while many also destined for pets, circuses, zoos, and private collections.
Sources: http://cites-dashboards.unep-wcmc.org, data presented as live specimens exported from 2011–2015. Data generally do not include illegal traded specimens, which are usually not reported to CITES.
Overharvesting of plant products
Overharvesting is not restricted to animals and animal products. While legal and illegal timber and firewood extraction is a major source of deforestation throughout Africa, it is also an important extinction driver. In fact, logging and other forms of wood harvesting have already contributed to the extinction of at least six plant species in Sub-Saharan Africa, with an additional 116 species considered Critically Endangered in part due to these threats (IUCN, 2019). Other plant species face extinction due to exploitation for medicines, spices, fragrances, and ornaments. For example, White’s ginger (Mondia whitei)—reputed to have aphrodisiac and antidepressant properties—has been harvested to extirpation in parts of central Kenya and South Africa (Aremu et al., 2011). Similarly, harvesting rates of African blackwood (Dalbergia melanoxylon, NT)—popular for making musical instruments and fine furniture—are currently unsustainable because the tree is slow-growing, has low germination rates, and extractions are seldom offset with planting of new seeds or seedlings (Amri et al., 2009).
Challenges in managing overharvesting
One of the biggest challenges in combatting overharvesting is the non-enforcement and/or outright absence of legal controls to protect exploited species. But even where strong regulatory frameworks exist, the sheer scale of the problem poses practical challenges for effective enforcement (discussed in Chapter 12), as billions of dollars flow among participants in illegal wildlife trade, which include local people trying to make a living, professional poachers, corrupt government officials, unethical dealers, and wealthy buyers who are not concerned about how the wildlife products they use were obtained. The illegal wildlife trade has hit Africa’s megafauna particularly hard. For example, even though there has been an international ban on the ivory trade since 1989, thousands of African elephants continue to be illegally killed on an annual basis (Box 7.2). Similarly, despite a ban on rhinoceros horn trade since 1977, an increasing number of rhinoceros succumb to poaching every year (Figure 7.11). Worse yet, the illegal wildlife trade shares many characteristics and practices with the illegal trade in drugs and weapons; in some cases, the same syndicates run these various criminal enterprises (Christy and Stirton, 2015). Apprehending these criminal networks is generally very dangerous, requiring vast resources.
Box 7.2 Conserving Elephants in the Anthropocene
David H.M. Cumming1,2
1FitzPatrick Institute of African Ornithology, University of Cape Town,
Cape Town, South Africa.
2Tropical Resource Ecology Programme, University of Zimbabwe,
Harare, Zimbabwe.
[email protected]
As our increasingly human-dominated planet enters a new geological era, will there still be room for Earth’s largest land mammals? Or will there be, as happened to mammoths, sabre-toothed cats, and giant sloths during the Pleistocene (see Box 8.1), another hominid-induced extinction of large mammals? Our new human-dominated era has become known as the Anthropocene (Waters et al., 2015), and the animals are, of course, Africa’s elephants.
Elephants encapsulate the dilemmas of conserving large charismatic mammals. They are dominant ecosystem engineers that, depending on their densities, can facilitate or adversely impact species diversity and ecosystem processes (Section 4.2.1). They are also economically important to ecotourism industries and revered by many; ivory ornaments and carvings have been valued highly by many cultures past and present. But elephants are also regarded as dangerous pests by expanding small-scale farming communities, responsible for destroying crops and killing people. While retaliatory killings and habitat loss (primarily through agricultural expansion) certainly contribute to the endangerment of elephants, poaching to supply Asian markets (see Figure 12.1) is the primary cause behind massive population reductions we are currently witnessing (Wittemyer et al., 2014).
Africa’s elephants have, in the past, been greatly exploited, first for their meat and later also for their ivory. In 1887, about 1,000 tonnes of ivory were being exported from Africa (Spinage, 1973) and, by 1900, elephant populations in many African countries had all but collapsed. In Southern Africa, for example, it was feared that they might soon go extinct. However, with effective protection, elephant populations increased twentyfold, to more than 200,000, south of the Zambezi River by the 1970s. Elsewhere, elephant numbers also recovered, and, in the mid-1970s, the continental elephant population was estimated to be more than 1 million (Table 7.B). But a rapid escalation of the illegal killing of elephants for ivory and meat soon followed, accompanied by a steep rise in the price of ivory. In response, African elephants were placed on CITES Appendix II in 1976 to control the international trade in ivory. Elephants in some Southern African countries were well protected, so numbers continued to grow. However, elsewhere poaching and illegal trade in ivory continued and, in 1989, the African elephant was placed on CITES Appendix I, which banned all international trade in elephants and elephant products. The result was a decline in the price of ivory, and recovery of many populations.
Conservationists, and the world at large, traditionally regarded the African elephant as a single species. Recent morphometric and genetic evidence has revealed that forest and savannah elephants represent two distinct species, with forest elephants (Loxodonta cyclotis) occupying the West and Central African forests, and the larger savannah elephants (Loxodonta africana) being widespread in non-forested regions of Sub-Saharan Africa (Roca et al., 2015). The distinction has important implications for their conservation, as each of these elephant species now viewed on its own is even more sensitive to population declines (CBD, 2015).
Since about 2006, poaching of elephants again began to escalate, in part a response to an increase in the price of ivory, poorly funded wildlife agencies, and corruption (Hauenstein et al., 2019). The scale of these killings is extraordinary. For example, an estimated 30,000 elephants were killed in 2013 alone. Forest elephants declined by about 60% (Maisel et al., 2013). While population trends for savannah elephants vary across the region, they too face increased poaching pressure (Chase et al., 2016). Due to these large-scale killings, combined with the impact of habitat loss from agricultural expansion, West Africa’s elephants are today confined to small isolated protected areas with a total population of about 17,000 (Maisel et al., 2013). Elephant population trends in East Africa vary: numbers are increasing in Uganda and Kenya, but Tanzania has lost some 60,000 elephants in the last few years. In Southern Africa, Botswana has the largest elephant population, estimated in 2014 to number at least 130,000. Neighbouring Zimbabwe has a population of 83,000 elephants, much the same number as it had in 2001. However, two of Zimbabwe’s four regional populations declined significantly between 2006 and 2014 with a loss of at least 20,000 elephants (Figure 7.C).
Global and national efforts to curb elephant poaching are currently focused on improving law enforcement on the ground, intercepting ivory shipments to Asia, closing ivory markets in Africa and Asia, and leading campaigns to reduce demand for ivory in major consuming countries in Asia, particularly in China. Importantly, while these initiatives are relieving poaching pressure on African elephant populations, they fail to address core issues relating to the interactions between people and elephants in the rural areas of Africa. A high proportion of elephant ranges lie outside protected areas where they overlap with people. Relieving continued pressures on elephants, outside as well as inside protected areas, would only happen if people who are harmed by elephants derive enough benefits from elephants and other wildlife to outweigh the direct and indirect costs of sharing land with them. Community-based natural resource management (CBNRM) projects, such as those in Namibia (Section 14.3) show that this can be achieved. Establishing secure and sustainable funding streams through payments for ecosystem services (Section 15.3), or payments for co-existing with large dangerous mammals such as elephants (Section 14.4), could extend these promising initiatives even further.
Now, consider a hypothetical conservationist concerned about the bushmeat crisis’ impact on biodiversity. This person may very well think that effective enforcement of a hunting ban would be the best solution to prevent further overharvesting. Unfortunately, solving complex challenges with simplistic steps runs a high risk of being counterproductive. For example, local bushmeat markets provide important contributions to food and financial security in many rural parts of Africa (van der Merwe et al., 2015). Replacing bushmeat with livestock and crops production—two primary drivers of habitat loss (Chapter 5)—also carries risks. For example, an estimated 250,000 km2 of forest will need to be converted to pastureland to replace the bushmeat trade just for the Congo Basin (Nasi et al., 2011). Clearly, there is a need for the bushmeat trade, albeit in a sustainable manner to ensure long-term viability of the local biodiversity, as well as the prosperity of the people who inhabit these areas. We delve deeper into solutions for these kinds of complex challenges from Chapter 9 onwards. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/07%3A_Pollution_Overharvesting_Invasive_Species_and_Disease/7.02%3A_Overharvesting.txt |
Persecution involves the indiscriminate abuse or killing of a group of animals, generally used as a strategy to prevent property damage (e.g. crop-raiding elephants) and livestock depredation (e.g. lions endangering cattle). Some animals are also persecuted because of the real or perceived dangers they pose to humans; this includes the indiscriminate killing of sharks, snakes, and spiders to avoid bites, and culling of bats to prevent spread of zoonotic diseases (Schneeberger and Voigt, 2016). Lastly, local folklore also contributes to persecution: animals, such as moles, chameleons, and owls are sometimes indiscriminately killed because of cultural beliefs that they bring bad luck.
While killing a (potential or perceived) problem animal may bring a certain instant gratification, it is a short-term solution that often causes more harm than good.
While killing a (potential or perceived) problem animal may bring a certain instant gratification, it is a short-term solution that often causes more harm than good. Culling bats, for example, may in fact increase the prevalence of the same zoonotic diseases people are trying to control (Schneeberger and Voigt, 2016). Persecution also leads to the loss of ecosystem services, as it often targets ecosystem engineers and keystone species (Section 4.2.1). Retaliatory poisoning is particularly harmful for all the other useful organisms that may be killed in the process. One study in Namibia estimated that about 100 non-target animals are killed for every target animal, putting harmless species such as aardwolf (Proteles cristatus, LC), bat-eared fox (Otocyon megalotis, LC), and Cape fox (Vulpes chama, LC) also at risk (Brown, 2006). Among the most vulnerable to such poisoning are vultures, which may be accidentally poisoned by bait set out for problem predators, or directly targeted for traditional medicine and to hide poaching activities. In June 2019, 537 vultures (comprising five different Endangered and Critically Endangered species) were killed in Botswana after scavenging from three poisoned elephant carcass (de Greef, 2019). This mass poisoning event was particularly devastating because it was during the vulture breeding season, so many vulture chicks likely also died from starvation if not from eating tainted meat brought back to the nest by their parents.
7.04: Invasive Species
An exotic species is a species that occurs beyond its native range, most often because humans have moved it, whether intentionally or not. Most exotic species do not establish viable populations in the new areas to which they have been moved because the new environments may not meet their needs, because native species may outcompete them or otherwise displace them, or because there may not be a sufficient number of individuals to become established. However, a small number of exotic species go on to become invasive species—exotic species that rapidly spread and increase in abundance at the expense of native species and ecosystems. While there is no definitive list of qualities that predict which exotic species can become invasive, many invasive species have the following in common: (1) they begin to reproduce at an early age; (2) they can reproduce rapidly; (3) they lack sufficient predators in their introduced range; (4) they disperse easily; (5) and they are generalist species, able to survive in a variety of ecosystems.
Invasive species displace native species through competition, predation, and habitat alterations. They often thrive in environments disturbed by human activities.
Spread of invasive species
As stated above, invasive species spread and invade new areas because human activities move them there. Some of the most prominent means by which human activities facilitate the spread of invasive species include:
• Agriculture: Large industries exist to grow agricultural plants for crop production, ornamental plants for gardens, grasses for pastures, and livestock for food. Many of these organisms later escape from cultivation and captivity and go on to invade and harm local ecosystems. Other species spread when industry workers accidentally harvest the seeds of weedy plants along with commercial seeds, and then sow those seeds in new localities, while microbes, parasitic organisms, and insects may be transported with plant leaves and roots, in potting soil, or even attached to transported animals. One of Africa’s worst plant invaders, the triffid weed (Chromolaena odorata), continues to spread because of its sustained use to boost soil fertility on agricultural lands (Uyi et al., 2014).
• Accidental transport: Many invasive species spread to new areas because people transported them unintentionally. Domestic rats (Rattus spp.) and mice (Mus spp.) are classic examples: they have spread around the world as stowaways aboard ships (Box 7.3). A great number of invasive plants that currently occur in South Africa’s Cape Floristic Province arrived by accident by clinging onto the luggage and hiking gear of tourists (Anderson et al., 2015). The ballast tanks of ships are also common hiding places for invasive aquatic species: DNA analyses have shown how a German ship carried tiny snails in this way along the entire length of Africa’s Atlantic coast, providing invasion opportunities across this entire shipping route (Ardura et al., 2015).
• Biological control: Environmental and agricultural organizations sometimes use biological control (Section 4.2.7) to manage the spread of, and harm caused by, invasive species. While this approach can be very effective, in rare cases the biocontrol agent can become invasive and harm native species, rather than its intended target. For example, domestic cats that were introduced to Marion Island off Africa’s south coast feasted on native seabirds—in some cases, even causing seabird extirpations—instead of the rats and mice they were meant to control (Bloomer and Bester, 1992). For this reason, very careful research is necessary to test the appropriateness of a biocontrol agent before being released.
• Deliberate introductions: Soon after their arrival, colonists released hundreds of European birds and mammals into countries like South Africa and Kenya to make the African countryside feel more familiar. Other species, especially fish (e.g. trout, bass, and carp), were released to provide food and recreational opportunities. Many of these species have subsequently become so successful that they harm native species. For example, mesquite (Prosopis juliflora), introduced to Ethiopia from Mexico to reduce soil salinity, proved such a successful invader that it completely displaced local plants; the subsequent encroachment even threatens the viability of Ethiopia’s few remaining Grevy’s zebras (Equus grevyi, EN) (Kebede and Coppock, 2015). The introduction of the Nile perch (Lates niloticus, LC) to the Rift Valley to boost local fisheries likely led to the extinction of hundreds of fish species endemic in Lake Victoria (Pringle, 2005).
• Captive escapees: Many invasive species were originally kept as pets or ornamental plants but have escaped from captivity to establish feral populations that harm local wildlife. Some of the most problematic aquatic invasive species are common pets that people dumped in streams, lakes, or storm drains because they could not care for them anymore. Finding these escapees before they establish should thus be a priority. For example, a recent survey found that 258 alien ornamental plant species growing in South Africa’s Kruger National Park are at risk of becoming invasive—most of these plants were subsequently removed from the park (Foxcroft et al., 2008).
Box 7.3 Aliens on Islands: Damage and Control
Peter Ryan
FitzPatrick Institute of African Ornithology, DST-NRF Centre of Excellence,
University of Cape Town, South Africa.
[email protected]
Invasive species are one of the main threats to biodiversity. Island ecosystems are particularly vulnerable; of the 156 bird species that have gone extinct in the last 500 years, more than 90% lived on oceanic islands (IUCN, 2019). This vulnerability is mainly due to species on oceanic islands evolving in the absence of competing species or predators, and thus lacking adequate defences against introduced species (including humans).
Invasive species pose many threats to island biodiversity. Newly arrived mammalian predators have exacted the greatest toll. Seemingly unable to appreciate the danger posed by these strange new arrivals, the ecologically-naïve adult birds simply remain on their nests to be eaten rather than fleeing. Introduced herbivores can have devastating impacts, because many island plants lack defences like tough leaves or thorns. Most devastating are domestic goats and rabbits, introduced by early island explorers to provide a source of food in the case of shipwreck, that have grazed many once-lush islands down to the ground. Introduced plants can also outcompete native plants. For example, the Mexican thorn (also called mesquite, Prosopis juliflora) has formed dense thickets on the once sparsely vegetated lowlands of Ascensión Island, making those areas unsuitable for both nesting seabirds and sea turtles.
Some of Africa’s least-transformed islands are the sub-Antarctic Prince Edward Islands, 2,000 km southeast of Cape Agulhas, and Gough Island, 2,800 km west of Cape Town. Their small size, isolation, and lack of sheltered harbours prevented human settlement. Nevertheless, their large seal and seabird populations were frequently exploited for oil and skins in the 19th and early 20th centuries. In the 19th century, sealing parties accidentally introduced house mice (Mus musculus) to Marion Island, the larger of the two Prince Edward Islands, and Gough Island. The mice flourished by eating native invertebrates and plants, probably causing the local extinction of one flightless moth on Marion Island. A few domestic cats were brought in to control the mice at Marion Island’s weather station, established in 1948. Instead, the cats targeted the island’s birds, which were easier prey. By the 1970s, some 2,000 cats were killing an estimated 450,000 seabirds each year, greatly reducing the island’s burrow-nesting petrels and even driving some species to local extinction (Bloomer and Bester, 1992). The events on Marion contrasted with nearby predator-free Prince Edward Island that continued to support vast breeding populations of burrowing petrels.
A pioneering initiative eradicated Marion Island’s cats in 1991, using a combination of introduced cat influenza, hunting, trapping, and poisoning (Bloomer and Bester, 1992). Researchers hoped Marion’s seabird populations would recover within a decade but had not considered the impact of mice once the cats were removed. The precedent was set on Gough Island; in 2001, introduced mice were discovered to predate on large numbers of seabird chicks, including Tristan albatross (Diomedea dabbenena, CR) chicks more than 100 times larger than themselves (Davies et al., 2015; Dilley et al., 2015a). It was hypothesised that mice are more likely to attack seabirds when they are the sole introduced predators on an island. Sure enough, the first attacks on Marion’s albatross chicks were recorded in 2003 (Figure 7.D); by 2015, the attacks had increased dramatically (Dilley et al., 2015b).
Fortunately, it is possible to eradicate invasive species from islands. In 2014, Australia removed mice, rats, and rabbits from sub-Antarctic Macquarie Island (Parks and Wildlife Service, 2014), which is almost twice the size of Gough Island. Plans are now also in place to eradicate Gough’s mice in 2019. The island’s isolation facilitates this effort—damage to other species can be minimized, and possible spread of toxins or diseases will be confined. To access areas inaccessible on foot, helicopters will be used to spread poison bait from specially designed hoppers slung under the aircraft. Some poisoning of non-target native individuals is inevitable, but this is a small price to pay compared to extinctions of those species. If adequate measures are put in place to prevent subsequent reintroductions, there is hope of restoring at least part of the island’s natural balance.
Impact of invasive species
Invasive species have many negative consequences for native biodiversity: they displace native species through competition, alter the structure and composition of natural communities, and sometimes also hybridise with native species. These impacts may also translate to financial losses, as invasive species compromise ecosystem services (Figure 7.12), damage infrastructure, and spread infectious diseases.
Invasive species often become pervasive because they outcompete and displace native species. One such example is the Mediterranean mussel (Mytilus galloprovincialis), which was accidentally introduced to South Africa in the mid-1970s via European ships. A superior competitor, the exotic mussel soon started displacing native mussels and limpets, especially in the inter-tidal zone of South Africa’s west coast (Branch and Steffani, 2004). Considered South Africa’s most successful marine invasive, recent evidence suggests that the Mediterranean mussel is continuing to spread north into Namibia and along South Africa’s east coast towards Mozambique.
Another superior competitor is the water hyacinth. A native to South America’s Amazon forest, this species was intentionally introduced as a showy ornamental plant to dams, ponds, and lakes across Africa in the early 20th century. The plant established well, but then started reproducing and spreading at such rapid rates that water bodies across the region were soon covered by a dense mat of leaves. With little surface exposure and water movement, eutrophication and suffocation followed, leading to the deaths of countless fish and other aquatic organisms (Villamagna and Murphy, 2010). A biological control program targeting hyacinth showed promise during the 1990s; however, eutrophication from fertiliser overuse (which stimulate growth of hyacinth and other invasive aquatic plants) may be contributing to this species’ recent resurgence (Coetzee and Hill, 2012; Bownes et al., 2013).
A single gum or pine tree can transpire as much as 50,000 litres of water per year, while plantations of these trees can reduce water resources in an area by as much as 70%.
Natural communities are at particular risk in cases where invasive species change ecosystem structure and functioning so much that native species can no longer survive. Such is the case across many parts of Africa, where invasive Australian gum (Eucalyptus spp.) and pine (Pinus spp.) trees (both widely planted for timber) transpire so much water through their leaves (as much as 50,000 litres of water per tree/year; Dzikiti et al., 2016) that they can reduce the availability of surface- and groundwater in an area by as much as 70% (le Maitre et al., 2016). In addition to creating drought conditions, the closed canopies created by these invasive trees reduce the amount of solar radiation reaching the ground, greatly limiting thermoregulatory opportunities for taxa such as reptiles (Schreuder and Clusella-Trullas, 2016). Habitat degradation caused by invasions of the cinnamon tree (Cinnamomum verum), originally from Sri Lanka, has already caused at least ten invertebrate extinctions in the Seychelles (IUCN, 2019).
While many of the invasive species mentioned earlier originated from outside Africa, it is important to note that African species can also become invasive in other parts of Africa when they are moved outside of their native ranges. When invading nearby areas, non-native species can come in close contact with closely-related species, creating a high risk of genetic mixing—also called genetic pollution or genetic swamping—which describes the hybridisation of invasive species with native species. For example, hybridisation with the widespread banded tilapia (Tilapia sparmanii, LC) threatens the survival of Namibia’s Otjikoto tilapia (Tilapia guinasana, CR), globally restricted to the < 1 km2 Lake Guineas (Bills 2007). The Cape platanna (Xenopus gilli, EN), an endemic to South Africa’s Cape Floristic Region, is similarly threatened by hybridisation with the widespread African clawed frog (Xenopus laevis, LC) (Fogell et al., 2013).
Genetically modified organisms
A topic of conflict among conservation biologists is the increased popularity of genetically modified organisms (GMO). A GMO is an organism whose genetic material has been altered to provide useful or improved products and services. To do this, scientists typically use genome editing technologies to transfer genes from a “source” organism into the DNA of the target organism. For example, scientists can transfer a bacterial gene that produces an insect toxin into a crop, such as maize, to obtain a GMO that can resist insect herbivory. Farmers using this GMO maize would then be able to increase production and reduce pesticide use (Gewin, 2003). While GMOs are usually associated with the development of pest-resistant and drought-resistant crops, uses are highly varied. For example, in Senegal, GMO technologies are used to produce tilapia that are better adapted to local ecosystems (Eknath et al., 2007). GMO technologies are being used to develop new and cheaper medicines (Concha et al., 2017), and to combat important diseases: trials in Burkina Faso shows that fungi genetically engineered to produce spider toxins caused a 99% collapse in malaria-carrying Anopheles mosquito populations within 45 days (Lovett et al., 2019). GMOs can even be used for conservation purposes, like developing new methods to combat invasive species (Esvelt et al., 2014), creating more effective bioenergy sources (Beer et al., 2009), and making vulnerable species more resistant to climate change (Piaggio et al., 2017). Some scientists even hope to combat plastic pollution by creating a genetically modified bacterium able to consume plastic waste (Austin et al., 2018).
The use of GMOs is not a new phenomenon. Selective breeding, hybridisation, and other forms of artificial selection—techniques that have been used for much of human history—all result in different forms of genetically modified crops and animal species. However, technological advances in genetic engineering have enabled scientists to transfer genes from and between different taxa that have not previously been used in selective breeding programs (i.e. viruses, bacteria, insects, fungi, and shellfish). GMO technologies that transfer genetic material between wholly disparate taxa has led to concern about the unknown and unintended consequences of such “crossovers”. Some people are also concerned that GMOs that escape from captivity or cultivation (e.g. Gilbert, 2010) could hybridise with closely-related wild species, endangering native wildlife while resulting in new, aggressive weeds and virulent diseases. Additionally, the use of GMO crops could potentially harm non-crop species (e.g. insects, birds, and soil organisms) that live in, on, or near the GMO crops. Concerns have also been raised about the potential effects to people eating GMO foods, leading some governments to regulate GMO research and commercial applications differently than traditional agriculture. However, after decades of research, it appears that GMO food is safe to eat (Blancke, 2015).
GMOs offer benefits such as increased production and reduced pesticide use, but there are concerns about unknown and unintended consequences like hybridisation and invasiveness.
Clearly, GMOs offer a wide range of opportunities which could directly and indirectly benefit biodiversity conservation. However, the benefits of GMOs must be examined and weighed against the potential risks on a case by case basis. It is probably wise to proceed cautiously, to study GMOs thoroughly, and to monitor their impacts on ecosystems and human health in the areas where they are used. These investigations could involve workshops where experts come together to perform environmental impact assessments (EIA) (Section 12.2.2). The potential but unknown impacts of GMOs can also be mitigated by limiting the ability of these organisms to spread or reproduce (Muir and Howard, 2004). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/07%3A_Pollution_Overharvesting_Invasive_Species_and_Disease/7.03%3A_Persecution.txt |
Parasites and diseases have always been an important natural factor in regulating the ecology of wildlife, especially in wild populations that have become unsustainably large. Today, however, human activities are facilitating increased spread and transmission of parasites (Box 7.4) and other pathogens, sometimes even creating conditions for epidemics to develop (Figure 7.13). Consequently, parasites and diseases have become a major threat to wildlife, including those already suffering under low population sizes and densities.
Box 7.4 Promoting African and Global Honeybee Health
Vincent Dietemann
Agroscope, Swiss Bee Research Center,
Schwarzenburgstrasse, Switzerland.
[email protected]
Modern society imposes increased pressure on animals and plants to secure the food needed for a growing human population. Honeybees especially contribute to crop productivity in a crucial way thanks to their pollination services. Unfortunately, the number of bee colony losses has surged in recent years in several regions of the world (Goulson et al., 2015), worrying scientists, politicians, and the public. Some regions in Africa, however, have maintained healthy domesticated and wild honeybee colonies (Pirk et al., 2016), which continue to pollinate flowers and enhance the production of many fruit and vegetable crops.
The power of comparisons
Comparative studies have always been important to biological research. Studying a model organism or system under different conditions allows scientists to identify how these organisms or systems react and adapt to their environment. This can even be done at the continental scale and could help us understand the effect of human pressure exerted on honeybees. In general, beekeeping in North America and Europe has been widely industrialized, involving large-scale operations and modern technology, whereas in Africa beekeeping has remained small-scale and mostly low-tech. This gives us an opportunity to determine the effects of beekeeping management and trade on honeybee health.
Different contexts
The varroa mite (Varroa destructor) originally parasitised the eastern honeybee (Apis cerana). In the wake of global honeybee trade, this parasite has invaded most regions of the world that are home to the western honeybee (Apis mellifera) and resulted in major colony losses (Figure 7.E). Although eastern and western honeybees are closely-related species, the western honeybee did not coevolve with this parasite and, thus, has few natural defences against it, with colonies dying within a few years after infestation. Consequently, only those colonies treated against the parasite by beekeepers can survive, and most wild honeybee populations have been decimated. However, there have been exceptions: colonies of the western honeybee in the southern parts of Africa are resistant to the parasite, and large wild populations remain. Several international teams have now turned their attention to resistant African honeybee populations to understand the basis of their survival (Strauss et al., 2016). Researchers hope to use this knowledge to promote the breeding of surviving colonies in currently susceptible populations both within and outside of Africa.
Africa also differs strongly in land use and crop management techniques that are likely to influence honeybees’ nutrition and health. In many areas, small-scale farming prevails, with a lower use of pesticides than in other areas of the world. Understanding how pesticide and other chemical use, as well as how nectar and pollen variety and quality, impact these pollinators will be key to their survival. Therefore, the rest of the world may learn how to maintain healthy honeybees from Africa. Africa, in return, might benefit from global efforts to maintain sustainable pollination services and promote food security.
One way in which humans elevate the impact of parasites and diseases on wildlife is by exposing native species to harmful organisms that they have never previously encountered, and thus have no evolved coping mechanisms. For example, population declines and extirpations of about 200 frog species across the world, including in Africa (Tarrant et al., 2013; Hirschfeld et al., 2016), is due, in part, to a disease caused by the chytrid fungus (Batrachochytrium dendrobatidis). This disease, known as chytridiomycosis (Figure 7.14), affects a frog’s ability to absorb water and electrolytes through the skin (Alroy, 2015). It likely originated in the Korean Peninsula (O’Hanlon et al., 2018), and spread across the world through trade with African clawed frogs (Xenopus laevis, LC) (Weldon et al., 2004). As of yet, there is no cure for this disease, and it continues to be seen as one of the biggest threats currently facing the world’s amphibians.
Disease transmissions can also occur when humans and their pets or livestock interact with wildlife (Cumming and Cumming, 2015). For example, during the early 1990s about 25% of lions in Tanzania’s Serengeti National Park were killed by canine distemper virus which they contracted from domestic dogs living near the park (Kissui and Packer, 2004). Because of the many biological similarities between apes and humans, gorillas (Gorilla spp.), chimpanzees (Pan troglodytes, EN) and bonobos (P. paniscus, EN) are particularly vulnerable to anthroponotic diseases, such as measles, influenza, and pneumonia which can be transferred from humans to animals. But even chytridiomycosis (discussed above) can become an anthroponotic disease, transferred from frog to frog by a careless biologist that handles a healthy frog after a sick one without taking precautions against transmission. Some diseases (e.g. Ebola; flu; and tuberculosis) can be anthroponotic and zoonotic (transferred from animals to humans). While the impact of Ebola on humans in Africa is well-known, it is worth noting that gorillas suffer 90% mortality when exposed to Ebola, compared to 50% mortality in humans. In fact, it was an Ebola outbreak in 2004 that caused the western lowland gorilla (Gorilla gorilla gorilla, CR) to be classified as highly threatened by the IUCN (Genton et al., 2012).
Human activities often facilitate the emergence and spread of infectious diseases, which threaten wildlife, domestic species, and humans alike.
Humans also indirectly facilitate the transmission and spread of parasites and pathogens. While there are some exceptions (notably social insects), transmission and infection rates are typically low for wildlife living in large, complex ecosystems because they have space to move away from disease-carrying droppings, saliva, old skin, and other sources of infection. However, these natural buffers against pathogens and parasites are removed when humans confine those organisms to small areas (such as small fenced reserves) or keep them in crowded conditions. In addition to forcing those organisms to remain in close contact with potential sources of infection, crowded conditions lead to deterioration of habitat quality and food availability. Both these factors increase the organisms’ stress levels and reduce their body conditions which, in turn, lowers their resistance to parasites and diseases (reviewed in Gottdenker et al., 2014).
Human-induced extirpations indirectly facilitate the transmission and spread of parasites and pathogens, even to humans. Such is the case with schistosomiasis (also known as bilharzia), a zoonotic disease carried by a few freshwater snail species. In the 1980s, health care professionals observed an increased incidence of human schistosomiasis around Lake Malawi after overfishing depleted snail-eating fish populations, followed by decreased incidence of schistosomiasis as fish populations recovered in the 1990s (Stauffer et al., 2006). A similar situation occurred in East Africa, where the elimination of apex predators resulted in increased olive baboon (Papio anubis, LC) populations, which not only worsened crop raiding, but also increased parasite infection rates among local peoples (Brashares et al., 2010).
Parasites and diseases also threaten captive wildlife populations, including those kept at zoos and other ex situ conservation facilities (Section 11.5). Because of the proximity in which different species are kept, captive conditions may allow for easier spread of diseases. An added complication with captive populations is that some individuals may function as disease reservoirs. These individuals generally appear healthy because they are fairly resistant to the disease they carry, yet they are able to infect other susceptible individuals. Disease reservoirs frequently limit opportunities for translocation of captive populations (Section 11.2), even when dealing with threatened species. For example, well-meaning people often bring raggedy-looking yet healthy penguins in moult to rehabilitation centres, hoping the penguins will be released once “better”. Yet, those animals might never be released back in the wild to avoid the risk of transmitting diseases to wild penguin populations (Brossy et al., 1999).
The impacts of diseases are bound to become more important in the future of conservation biology, especially as growing human populations and increased competition for space increase the need for single-species management and ex situ conservation (Chapter 11). Disease management should therefore always be taken very seriously, and appropriate steps taken to avoid disease transmissions. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/07%3A_Pollution_Overharvesting_Invasive_Species_and_Disease/7.05%3A_Parasites_and_Diseases.txt |
1. Many threats to biodiversity do not lead to immediate and/or direct mortality, but instead have sublethal impacts that compromise organisms’ fitness over time. Responses to these silent, insidious, and easily-overlooked threats are often delayed, especially when the negative effects are felt only years after exposure.
2. Environmental pollution leaves ecosystems uninhabitable for native wildlife, and cause sickness and death in wildlife and people. Common causes of pollution include pesticides, heavy metals, plastic, fossil fuels, fertilisers, light, heat, and noise, leading to pollution of water, groundwater, air, and soil.
3. Overharvesting is becoming an increasingly damaging threat to biodiversity because people have better access to previously unexploited areas and are adopting increasingly efficient methods for harvesting wildlife products. Persecution, which has its roots in human-wildlife conflict, is becoming an important threat because a growing human population is increasingly encroaching on the shrinking remaining natural habitats.
4. Invasive species outcompete local species and change the structure and composition of their native ecosystems. Human activity is responsible for these invasions, by accidentally or deliberately moving wildlife to new regions of the world. Some invasive species require a great amount of effort and resources to manage.
5. Disease transmission and spread increase when wildlife is confined to small areas and/or crowded conditions. Diseases may also be transmitted between wildlife, domesticated species, and even humans. Managing for diseases is also important in zoos and other ex situ facilities, because diseases spreading from one individual to another can prevent those individuals from being released into the wild.
7.07: Topics for Discussion
1. Which forms of environmental pollution are most prominent in the region where you live? Which natural ecosystems are impacted most severely by this pollution? How are humans affected? What do you think can be done to reduce or even eliminate this pollution?
2. Consider all the fishing, hunting, trapping, collecting, logging, and other wildlife harvesting activities in your region. Which activities are well managed, and which are not? Why is it so difficult to regulate these activities, when so many people know that overharvesting would eventually harm local economies, their families, and their livelihoods? What measures do you think can be put in place to control overharvesting in your region, or at least reduce its impact?
3. Briefly describe what biological control is, and what its benefits are. What are the risks involved in biological control? How can these risks be predicted and avoided?
4. Can you name a few diseases that can be transmitted from people to wildlife, and from wildlife to people? Which species are involved in each of these diseases? How does the transmission of each of these diseases impact the conservation management of species involved?
7.08: Suggested Readings
Benítez-López, A., L. Santini, A.M. Schipper, et al. 2019. Intact but empty forests? Patterns of hunting-induced mammal defaunation in the tropics. PLoS ONE 17: e3000247. https://doi.org/10.1371/journal.pbio.3000247 African mammals are undergoing catastrophic declines due to unregulated hunting.
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Species have evolved and disappeared since the very first species (thought to be microorganisms living in hydrothermal vents) made an appearance on Earth. Some species outcompeted others for access to limiting resources; some were driven to extinction by dangerous pathogens; some just found it hard to survive in constantly evolving ecosystems. While many extinction events have been rather limited in scope and, hence, caused only one or a few extinctions at a time, there have been instances where perturbations were so impactful that they drove very large numbers of species to extinction over a short period of time. There have been five such past mass extinction events—periods marked by the sudden and dramatic loss of a large percentage of species (Figure 8.1). But these mass extinctions have also been followed by periods that favoured increased rates of speciation, during which new species evolved to fill the niches left empty by the extinctions.
Nature’s ability to balance extinctions with speciation was greatly disturbed around 300,000 years ago, when Homo sapiens made their appearance on Earth. Since then, humans have gradually increased their dominance on the natural world, leading to large-scale restructuring and destruction of biological communities. Human modifications of Earth’s climatic, biological, and geochemical environments accelerated greatly during the rise of agriculture (12,000–15,000 years ago) and again during the Industrial Revolution (1760–1840), when fossil fuel usage and urbanisation became the norm. Now, many scientists recognize today’s new and distinct human-dominated geological epoch, the Anthropocene (Waters et al., 2015). One notable feature of the Anthropocene is that species extinctions are increasing at such rapid rates that many conservation biologists now recognize that we are also witnessing the beginnings of Earth’s sixth extinction episode (Barnosky et al., 2011; Ceballos et al., 2017). However, unlike previously, this extinction episode is caused by human activities rather than natural events.
8.01: What is Extinction
The term “extinct” has several nuances in conservation biology, and its meaning can vary somewhat depending on the context:
Over ninety-nine percent of recent extinctions have been caused by human activities.
• A species is globally extinct when no individuals of that species remains alive anywhere in the world. The bluebuck (Hippotragus leucophaeus, EX) has been globally Extinct since the last individual was shot around 1800 (Kerley et al., 2009).
• Four (possibly seven) species of cycad (Encephalartos spp.)—ancient seed plants that were dominant in the age of the dinosaurs—are currently considered extinct in the wild; in other words, they exist only in cultivation; in captivity; or another human-managed situation (IUCN, 2019).
• A species is locally extinct, also called extirpated, when it is extinct in a part of its historic range but can still be found elsewhere in the world. Cheetahs (Acinonyx jubatus, VU) once roamed throughout much of Africa, but are now extirpated in over 90% of their historical range (Durant et al., 2017).
• A species is ecologically extinct (also called functionally extinct) if it persists at such low numbers that its role in an ecosystem is negligible. Africa’s vultures are ecologically extinct over much of their range and thus unable to remove diseased carcasses from the environment, posing both an ecological and socio-economic hazard (see Box 4.4).
8.02: Rates of Extinction
If extinction and speciation are natural processes, an obvious question follows: “Why should we care about the loss of biodiversity?” The answer concerns not individual species extinctions as much as the increasing rate of these extinctions (Figure 8.2). While a species can be wiped off Earth over a relatively short period of time, speciation typically occurs slowly as the genetic makeup of a population shifts over thousands of years. Unfortunately, we are currently losing species 1,000 times faster than natural background extinction rates (for mammals estimated to be 1.8 extinctions per 10,000 species per 100 years, Barnosky et al., 2011), and future rates may be 10,000 times higher that background rates (de Vos et al., 2015). Because over 99% of current species extinctions have been linked to human activity rather than natural processes (Pimm et al., 2014), observations on past extinctions and subsequent speciation may not apply to the present. Moreover, unlike before, humans now share the planet with the species we are wiping out. These losses mean that we are also losing the benefits we gain from nature (Chapter 4) at unprecedented rates. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.00%3A_Prelude_to_Extinction_is_Forever.txt |
While the term “extinction” is relatively easy to define (Section 8.1), determining whether a species is indeed extinct is a more difficult task. One of the most important questions conservation biologists grapple with is deciding how long to wait after the last observation before declaring a species extinct. Answering this question is particularly complicated when considering cryptic and shy species that are difficult to survey, sparsely distributed animals that are hard to find, or plants that are difficult to identify when not in flower.
To complicate matters, over the last few decades, biologists and their colleagues have rediscovered several species that were once thought to be extinct. These rediscovered species are often called Lazarus species, in reference to their apparent return to life. Recent examples include Burundi’s Bururi long-fingered frog (Cardioglossa cyaneospila. NT) rediscovered after a 60-year absence (Blackburn et al., 2016), a Tanzanian coral tree, (Erythrina schliebenii, CR) originally known from only one specimen collected from a deforested region in the 1930s (Clarke et al., 2011), and the coelacanth (Latimeria chalumnae, CR), a fish that was once thought to be extinct for millions of years (Balon et al., 1988). To avoid declaring more extant species as extinct, there is currently a practice of only declaring a species extinct after several decades of intensive searching and “there is no reasonable doubt that the last individual has died” (IUCN, 2012). Consequently, species, such as the black-spotted false shieldback katydid (Aroegas nigroornatus, CR) and the Ethiopian sedge, Cyperus chionocephalus, Critically Endangered, last seen in 1916 (Bazelet and Naskrecki, 2014) and 1836 (Contu, 2013) respectively, have not yet been declared extinct, even though the last individual may have died a long time ago. Similarly, as many as 15 African orchid species—a group that includes some of the most beautiful and specialised plants on Earth, some of which have not seen since 1890—are currently considered Critically Endangered but may actually be extinct (IUCN, 2019).
The rediscovery of species once thought to be extinct should not necessarily be considered a sign of conservation progress. In many cases, Lazarus species were simply overlooked because they were extremely rare and restricted to isolated locations. Such is the case for two forest birds from the island nation of São Tomé and Príncipe, namely the São Tomé grosbeak (Crithagra concolor, CR) and Newton’s fiscal (Lanius newtoni, CR). The grosbeak, the world’s largest canary (50% larger than the second largest canary), was for a long time known only from three specimens collected in 1888–1890; it was thus considered extinct, until its rediscovery over 100 years later, in 1991 (BirdLife International, 2018a). The fiscal shares a remarkably similar history: it was previously known only from records in 1888 and 1928, until its rediscovery in 1990 (BirdLife International, 2018b). Despite these rediscoveries, both species persist as very small (< 250 individuals) populations that are at risk from extinction due to ongoing habitat loss and the impact of invasive predators.
Because extinctions may not always happen immediately after a disturbance, conservation biologists must also consider the lag time between destructive human activities and eventual extinctions. This is illustrated in a study from Kenya’s Kakamega Forest, which found that only half of the species that will eventually go extinct due to habitat loss do so in the first 50 years following habitat fragmentation (Figure 8.3). Long-lived plants can have particularly long extinction lag times, sometimes of several centuries. For example, populations of the Saint Helena olive (Nesiota elliptica, EX) fell below viable levels in the mid-1800s, but the last individual died only in 2003, when the species was officially declared extinct (Cronk, 2016). Species that are doomed to eventual extinction are considered committed to extinction (also called functionally extinct), while the total number of species committed to extinction is referred to as an area’s extinction debt. In one study, researchers used the island biogeography theory to estimate that the average extinction debt for African forest primates was over 30%—that is, more than 30% of forest primates are predicted to go extinct because of habitat destruction and other human activities that have already happened (Cowlishaw, 1999).
On a more positive note, extinction debts may also provide hope for conservation biologists, as the lingering presence of seriously imperilled species affords opportunities to prevent impending extinctions. Conservation biologists are currently illustrating how this can be done by preventing the extinction of three species of pale-coated, desert-adapted ungulates that were formerly common and widespread across the Sahel-Sahara region, namely the scimitar-horned oryx (Oryx dammah, EW), dama gazelle (Nanger dama, CR), and addax (Addax nasomaculatus, CR) (Durant et al., 2014; Brito et al., 2018; IUCN, 2019). The oryx once numbered around one million individuals, with herds of 10,000 seen as recently as 1936. But a population collapse soon followed: by 1985 only 500 oryx survived, and by 2000 it was declared Extinct in the Wild. The addax, relatively common as recently as in the 1970s, also experienced precipitous declines; today fewer than 30 individuals remain in the wild. Similarly, the once-common dama gazelle’s current global population numbers fewer than 250 individuals, fragmented among five subpopulations in Chad, Mali, and Niger. Conservationists noted initial declines already in the 1960s and 1970s, when wild individuals of all three species were caught to initiate captive breeding programs. Luckily, all three species responded well to these programs, and captive populations have grown so strong that reintroduction programs (Section 11.2) have been initiated for the addax (in 1985, in Tunisia), dama gazelle (in 2015, in Morocco), and oryx (in 2016, in Chad). With several reintroductions seemingly successful, there is hope that viable populations of these iconic species may one day again roam free in their previous strongholds. This will only happen if we can reverse or mitigate the threats that causes their population collapses in the first place, namely uncontrolled and illegal hunting, as well as disturbances associated with agriculture, oil exploration, and inconsiderate drilling of wells for groundwater extraction. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.03%3A_When_is_a_Species_Extinct.txt |
Many people hold on to the romanticised belief that historical human societies lived in harmony with nature. Accumulated evidence however indicates that this is not true; early humans have caused extensive ecosystem changes and species extinctions since Homo sapiens appeared on Earth about 300,000 years ago. In fact, even before the arrival of humans, our ancestors had made a mark, by driving species to extinction as early as during the Pleistocene period, which started about 2.5 million years ago (Box 8.1). The impact of early humans was particularly devastating to the wildlife of North America, South America, and Australia, which saw the demise of nearly all their large ( 100 kg) mammals, most famously megaherbivores such as the mammoths (Mammuthus spp.). The Pleistocene extinctions were somewhat less devastating to wildlife in Africa, Europe, and Asia, possibly because large mammals on these continents evolved with human predators, allowing them to develop appropriate defence/escape mechanisms. Nevertheless, Africa’s wildlife did not completely escape the Pleistocene extinctions, as increasingly-sophisticated human activities during that time ensured the demise of as many as 28 large mammal groups, which included Africa’s sabre-toothed cats (Barbourofelidae), nearly all the elephant relatives (Proboscidae), as well as giant hartebeests (Megalotragus spp.), giant buffaloes (Pelorovis spp.), giant hyenas (Pachycrocuta spp.), and giant giraffes (Sivatherium spp.).
Early humans have caused extensive ecosystem changes and species extinctions even before Homo sapiens appeared on Earth about 300,000 years ago.
Box 8.1 Pleistocene Extinctions: Climate Change, Hominin Predation, or Both?
David H.M. Cumming1,2
1FitzPatrick Institute of African Ornithology,
University of Cape Town, South Africa.
2Tropical Resource Ecology Programme, University of Zimbabwe,
Harare, Zimbabwe.
[email protected]
Many scientists consider the present rapid loss of biodiversity to be the start of the 6th mass extinction following five previous extinction episodes (see Figure 8.1), each of which led to large-scale restructuring of Earth’s biodiversity. The 5th global mass extinction took place 65 million years ago (Ma) when a massive meteorite collided with the Earth, and resulted in the extinction of all non-avian dinosaurs, and much else besides. This 5th mass extinction event also marked the transition from the Cretaceous to the Tertiary epoch (65 to 2.5 Ma). The Tertiary epoch was followed by the Quaternary, which includes the Pleistocene period (2.5 million to ~ 12 thousand years ago) and more recently the Holocene period—marked by the development of agriculture, and the subsequent domination of the Earth’s resources by Homo sapiens. The Pleistocene is known for a mini mass-extinction of sorts, which saw the demise of species such as mammoths and sabre-tooth cats. However, unlike previous comprehensive mass extinctions, the Pleistocene was characterized by the extinction of mostly large mammals and very large island birds.
Attempts by scientists to explain these Pleistocene extinctions have been characterized by two centuries of controversy over whether they were caused by climate change or by predatory hominins—the evolutionary line of primates that gave rise to modern humans. The four main hypotheses advanced to account for the loss of Pleistocene fauna are: (i) climate change with minimal if any hominin influence (e.g. Faith et al., 2018); (ii) climate change together with some hominin influence (e.g. Barnosky et al., 2004); (iii) selective hominin predation aided by climate change (e.g. Bartlett et al., 2015), and (iv) hominin predation helped by other large predators without the influence of climate change (e.g. Janzen, 1983, Ripple and Van Valkenburgh, 2010).
The very close relationship between the dispersal of hominins out of Africa, the timing of their arrival elsewhere in the world, and the subsequent extinction of large mammals and birds, provided the primary (if challenged) evidence for human agency in non-African Late- Pleistocene extinctions (e.g. Surovell et al., 2005; Johnson, 2009; Ripple and Van Valkenburgh, 2010). As Haynes (2018) has remarked, “…the proponents of climate change as the only cause of the Late Pleistocene extinctions have not clearly explained how or why so many of the extinct megafaunal genera had survived numerous earlier climate changes.”
Similarly, Faith et al. (2018) have stated that the failure of Pleistocene megaherbivores to adapt to the emergence of C4 grasses was the primary driver of their extinction. However, this claim ignores evidence that many extinct herbivore genera and species previously survived changes in diet over time (Ripple and Van Valkenburgh, 2010), that the diets of particular species were known to vary with location (Ferranec, 2004), and that many large species, which are typically highly mobile generalists, would have had little trouble adapting their ranges and diets to changing climates.
Research and debate on Pleistocene extinctions have tended to focus on the demise of non-African large mammals in the Late Pleistocene, which coincided with the period when hominins (Homo erectus and later also H. sapiens) dispersed across the globe starting at about 2 Ma. In Africa, however, earliest hominins appeared some 7 Ma. It didn’t take long for these early African hominins to develop the skills necessary to manipulate the environment to their advantage. Setting the pace were the Australopithecines, who used stone tools to butcher mammalian carcasses between 4–3 Ma. The Australopithecines and the rest of a diverse group of large predators were joined by H. erectus at the beginning of the Pleistocene, about 2 Ma. This new, qualitatively unique, hunter was able to hunt collaboratively in bands, and was anatomically adapted to throw projectiles forcefully and accurately at large prey (Lombardo and Deaner, 2018). A large brain also placed high nutritional and energetic demands on H. erectus, that could best be met by obtaining meat and bone marrow from proboscidians, the elephant relatives (Surovell et al., 2005; Boschian et al., 2019). Early African hominins were thus well adapted to hunt large prey and contribute to the demise of wildlife, particularly megaherbivores (those over 1,000 kg), through the Early and Middle Pleistocene (Figure 8.A). Other large carnivores at the time may very well have helped hominins drive many Pleistocene herbivores to extinction (Janzen, 1983, Ripple and Van Valkenburgh, 2010; Van Valkenburgh et al., 2016). But this combination may ironically also have led to demise of many of the large Pleistocene predators, through co-extinctions, after their main prey base disappeared (Werdelin and Lewis, 2013).
It thus seems likely that the emergence of a novel and increasingly effective predator during the Early Pleistocene, rather than climate change, was the ultimate factor that tipped the balance against the iconic species that disappeared soon after hominins appeared on Earth. It is worth noting that there is a clear relationship between body size and extinction risk (Figure 8.B), the result of large animals’ relatively long generation lengths, long gestation periods, long periods of caring for young, and an abundance of meat presented to eager hunters. Consequently, even a small increase in mortality may very well result in a large animal’s annual mortality exceeding its generational mortality, the end result being extinction. This relationship also partly explains why present-day elephant populations are unable to withstand poaching in many parts of Africa (Box 7.2).
The lessons from the Pleistocene extinctions are relevant also today. As explained above, accumulated evidence indicates that early humans have caused extensive ecosystem changes and species extinctions extending over more than a million years. Homo sapiens had emerged in Africa about 300,000 years ago (Callaway, 2017). As early humans mastered the use of fire, poison-tipped spears and arrows, pitfall traps, and a host of additional hunting techniques, this unique apex predator proceeded to influence the structure and composition of African (and global) landscapes, and the plant and vertebrate assemblages of the continent (Smith et al., 2019). For the last two million years our ancestors have set in motion a series of trophic cascades that continue to this day and are resulting in increasing loss of diversity of the flora and fauna of the African continent and the rest of the world.
While early extinctions were generally isolated and selective, extinction rates increased rapidly after the rise of agriculture, and especially after European settlers started colonising Africa from the 17th century onward. By no coincidence, the area where most of the extinctions and extirpations during colonialism occurred was on the southwestern tip of Africa, the location of the earliest intensive European settlements on the continent. For example, by 1700, hunting caused the extirpation of every single land animal over 50 kg within 200 km of Cape Town (Rebelo, 1992). As hunters moved further afield in search of targets, Africa saw its first post-colonial large mammal extinctions, namely the bluebuck (Figure 8.4), quagga (Equus quagga quagga, EX), and Cape warthog (Phacochoerus aethiopicus aethiopicus, EX).
Following the demise of many of the Cape Floristic Province’s large animals, humans have driven African species to extinction at an increasing pace. Today, at least 84 Sub-Saharan African species have been confirmed Extinct (Figure 8.5), nine species are Extinct in the Wild, and as many as 202 species are considered possibly Extinct (IUCN, 2019). Among the extinct species are two wildflowers (Acalypha dikuluwensis, EX; Basananthe cupricola, EX) wiped out by mining activities in the DRC; and from Seychelles, an endemic parakeet (Psittacula wardi, EX) that was hunted to extinction. Among the species that persist only in captivity is the Kihansi spray toad (Nectophrynoides asperginis, EW), whose population crashed from more than 20,000 individuals in June 2003 to only five individuals in January 2004 after the establishment of a hydropower plant in eastern Tanzania (Channing et al., 2006). While some species that are Extinct in the Wild may be released back into the wild at some point in future, the four (or possibly seven) cycad species that persist only in captivity will probably not be reintroduced due to ongoing concerns about poaching by plant collectors (Okubamichael et al., 2016).
While most of Africa’s extinctions—at least until now—were isolated events involving one or two species at a time, the region also provides one of the best-studied examples of a recent man-made mass-extinction event. In the mid-1950s, the Uganda Game and Fisheries Department introduced the predatory Nile perch (Lates niloticus) to Lake Victoria to bolster the local fishery industry (Pringle, 2005). An ecological and economic disaster followed, pushing the entire ecosystem to the brink of collapse. First, the local people continued to prefer smaller endemic cichlids—which they could preserve by drying in the sun—to the perch with its oily flesh. This allowed the predatory perch’s population to grow unchecked which, in turn, reduced over 500 endemic cichlid species’ populations by 80% in just a few years (Witte et al., 1992). As the cichlid populations crashed, some local people started consuming perch for protein; however, they preferred smoking the perch over wood fires. To obtain firewood and charcoal, trees were logged around the lake, which in turn increased eutrophication, as well as erosion and siltation. Despite this array of emerging ecological threats, the local fishery continued to harvest the rapidly declining cichlid population. Consequently, as many as 200 cichlid species may have been driven to extinction in the decade following the perch introduction (Goldschmidt et al., 1993).
Judging by the number of extirpations over the last few decades, Africa will undoubtedly see more species pushed to extinction in the coming decades. Of particular concern is West and Southern Africa, which have lost over 75% of its large mammal populations over the past few decades; losses across Sub-Saharan Africa as a whole generally amount to over 50% (Ceballos et al., 2017). Some species will hopefully be spared this fate with the help of people and organizations fighting for their continued survival (Box 8.2). Many species will not be so lucky. The world’s last western black rhinoceros (Diceros bicornis longipes, EX) died in Cameroon in 2011; the northern white rhinoceros (Ceratotherium simum cottoni, CR) may follow this fate within the next few years (see Box 11.4). Lions (Panthera leo, EN) have been extirpated from as many as 16 African nations (Bauer et al., 2015), while cheetahs (Figure 8.6) occur in less than 9% of their historic range (Durant et al., 2017).
Box 8.2 Swimming Dangerously Close to Extinction: Population Crash in Lesotho’s Endemic Maloti Minnow
Jeremy Shelton
Freshwater Research Centre (FRCSA),
Kommetji, South Africa.
[email protected]
Lesotho’s iconic Maloti minnow (Pseudobarbus quathlambae, EN) (Figure 8.C) is a small, stream-dwelling cyprinid, and is the only freshwater fish species endemic to the country. Historically, the species was widespread, but its distribution has become increasingly restricted and fragmented in recent times due to interactions with non-native fishes and habitat degradation (Skelton et al., 2001), leading to it being classified as Endangered by the IUCN (Chakona and Kubheka, 2018). Genetic research has revealed that what was previously considered a single widespread species comprises two genetically distinct lineages: a “Mohale lineage” found in the Mohale catchment, and an “Eastern lineage” which includes populations in five catchments east of the Mohale catchment (Skelton et al., 2001). The genetic divergence between the two Maloti minnow lines is a result of a long period of geographic isolation and warrants that they be conserved as separate evolutionary significant units (ESU). Furthermore, the Mohale lineage, which comprises 77% of the species’ known distribution, is of critical importance for continued survival of the species (Skelton et al., 2001).
Past surveys (e.g. Steyn et al., 1996) have revealed that the Maloti minnow was the only fish species inhabiting the rivers flowing into the Mohale Reservoir. Situated 4 km below the Reservoir, the 20 m high Semongkoaneng waterfall has historically prevented larger fish species from moving upstream into the upper catchment. Following the filling of the Mohale Reservoir in 2003, an inter-basin transfer (IBT) tunnel linking it to Katse Reservoir was opened. Biologists working in the catchment subsequently expressed concern that non-native fishes might colonise the Mohale Reservoir via the IBT tunnel, and from there invade the influent rivers (Rall and Sephaka, 2008). Because the Maloti minnow evolved in the absence of large-bodied fishes, it would not have had an opportunity to evolve adaptations to cope with competition from and predation by larger species and may, therefore, be particularly sensitive to the arrival of other fish.
In 2006, the smallmouth yellowfish (Labeobarbus aeneus, LC), a larger, more aggressive cyprinid, was recorded in Mohale Reservoir (Rall and Sephaka, 2008), suggesting that it had dispersed from Katse Reservoir through the IBT tunnel. By 2013, it had spread into the major influent rivers in that system and coinciding with this was a virtual disappearance of the Maloti minnow from this former stronghold for the Mohale lineage. To illustrate this, surveys in previous decades described healthy populations of several thousand fish (e.g. Steyn et al., 1996), while only five individuals were recorded from the same sites in 2013 (Shelton et al., 2017).
Interestingly, the saving grace for this lineage may have originated from the same source that landed them in this predicament in the first place: human intervention. Prompted by the opening of the Kaste-Mohale IBT, a small team of passionate conservation scientists translocated several Maloti minnows to sections of stream above tall waterfalls, upstream of their natural distribution range (Rall and Sephaka, 2008). These sections, they knew, would be unreachable by larger species swimming upstream from Mohale Reservoir. This assisted colonisation approach has been viewed as controversial, but it may also have saved a tiny minnow from almost certain extinction in the wild. The prospect of losing a charismatic species like the Maloti minnow showcases how projects like the Lesotho Highlands Water Project can easily damage sensitive ecosystems that were not consider in development plans. In order to save the Maloti minnow from extinction, the next step will be to assess the success of translocation efforts and develop a rescue plan for the species. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.04%3A_History_of_Extinctions_in_Sub-Saharan_Africa.txt |
The IUCN has formalised the evaluation of threatened species using an internationally accepted standard of conservation categories describing a taxon’s risk of extinction.
An important task for conservation biologists is to identify and prioritise those species in greatest danger of extinction. Accomplishing this task requires biologists to collect and review all the information we have on each species. To facilitate this major undertaking, the IUCN has formalised the evaluation and reporting of threatened species assessments using an internationally accepted standard of conservation categories to reflect a taxon’s risk of extinction. These nine categories (Figure 8.7), known as Red List Assessments (IUCN, 2017), are:
• Extinct (EX). These species are no longer known to exist. As of mid-2019, the IUCN has listed 84 Sub-Saharan African species as Extinct.
• Extinct in the Wild (EW). These species exist only in cultivation, in captivity, or other human-managed situations. As of mid-2019, the IUCN has listed nine Sub-Saharan African species as Extinct in the Wild.
• Critically Endangered (CR). These species have an extremely high risk of going extinct in the wild. As of mid-2019, the IUCN has listed 880 Sub-Saharan African species as Critically Endangered. Also included in this category are the 202 Sub-Saharan African species that the IUCN considered possibly Extinct.
• Endangered (EN). These species have a very high risk of extinction in the wild. As of mid-2019, the IUCN has listed 1,600 Sub-Saharan African species as Endangered.
• Vulnerable (VU). These species have a moderately high risk of extinction in the wild. As of mid-2019, the IUCN has listed 2,153 Sub-Saharan African species as Vulnerable.
• Near Threatened (NT). These species are close to qualifying for a threatened category but are not currently considered threatened. As of mid-2019, the IUCN has listed 1,034 Sub-Saharan African species as Near Threatened.
• Data Deficient (DD). Inadequate information exists to determine the risk of extinction for these species. As of mid-2019, the IUCN has listed 2,441 Sub-Saharan African species as Data Deficient.
• Least Concern (LC). These species are not considered Near Threatened or threatened. (Widespread and abundant species are included in this category.) As of mid-2019, the IUCN has listed 11,776 Sub-Saharan African species as Least Concern.
• Not Evaluated (NE). Species that have not yet been evaluated. Most species fall in this category.
These categories, and the Red List Criteria (Table 8.1) used to classify each species, are broadly based on population viability analysis (Section 9.2), and consider population size, population trends, and habitat availability. Species that are Extinct in the Wild, Critically Endangered, Endangered, and Vulnerable categories are officially considered “threatened with extinction”. The advantage of this system is that it provides a standard protocol by which decisions can be reviewed and evaluated according to widely accepted yet flexible criteria. Consequently, species, subspecies, varieties, populations, and subpopulations can be assessed on a global or regional level, all under a unified set of standards. The resultant threat status assessment forms the basis of Red Data Books and Red Lists: detailed lists of threatened wildlife by group and/or by region compiled by the IUCN and its affiliate organizations. All global (and many regional) Red List assessments are freely available at http://www.iucnredlist.org, with feedback links provided from which anyone can alert the IUCN if they find errors or have suggestions for improvements.
Table 8.1 The IUCN’s Red List criteria for evaluating a taxon’s threat status. A species that meet any one of criteria A–E could be classified as Critically Endangered.
Red List criteria A–E
Summary criteria used to evaluate a taxon as Critically Endangereda
A. Population size declining
The population size has declined by 90% (or more) over last 10 years or 3 generations (whichever is longer).
B. Geographical range declining
The species is restricted to < 100 km2 and it occurs at a single location and its distribution range is observed/expected to decline.
C. Small and declining populations
There are less than 250 mature individuals left and population has declined by 25% (or more) over last 3 years or 1 generation (whichever is longer).
D. Small populations
There are less than 50 mature individuals left.
E. Population viability analysis
There is a 50% (or greater) chance of extinction within 10 years or 3 generations.
a Additional criteria for Critically Endangered, as well as criteria for Endangered and Vulnerable listings can be found at http://www.iucnredlist.org.
While nearly 20,000 Sub-Saharan African species have been evaluated as of mid-2019 (IUCN, 2019), these assessments only cover a small proportion of the region’s overall biodiversity. Consider for example that as of mid-2019, just over 4,900 Sub-Saharan Africa’s plants have been listed on the IUCN Red List website. Yet, the Cape Floristic Region alone hosts over 6,200 endemic plant species. The assessment gaps are even more conspicuous for lesser-known taxa; for example, only seven species of bryophytes (a group of non-vascular plants that includes mosses) have been assessed as of mid-2019; some readers of this textbook will have more bryophyte species in their gardens. The reasons for such assessment gaps are many, but most boil down to manpower and funding limitations, which restrict our ability to obtain the data needed for comprehensive assessments. It is thus important to understand that the lack of information on these and other poorly known groups does not mean there is no threat. For example, as of mid-2019, no African abalone (Haliotis spp.) have been assessed, even though these highly valued molluscs are some of Africa’s most heavily exploited (and heavily poached) marine organisms (Minnaar et al., 2018). A lack of information about a species’ threats and populations trends is thus a good argument that more studies are needed, sometimes urgently. Similarly, continued monitoring of species thought to be common is also important, as it can shed light on how new threats may emerge or escalate over time.
Course-filter assessments
To fill Red List species assessment gaps, conservation biologists are increasingly relying on broader metrics, or coarse-filter assessments, to identify groups of species that are threatened with extinction. One such method, which reduces the need to evaluate every individual species, is to identify ecosystems that are threatened. This premise rests on the assumption that any threatened ecosystem will contain many threatened species. Hence, protecting and restoring threatened ecosystems will simultaneously allow many populations living in those ecosystems to recover. To facilitate this type of coarse-filter assessment, the IUCN recently established a Red List of Ecosystems (RLE, http://iucnrle.org). The RLE assesses ecosystem status against five criteria: (1) distribution declines, (2) distribution restrictions, (3) environmental degradation, (4) disruption of ecological processes and interactions, and (5) quantitative estimates for risk of ecosystem collapse (Keith et al., 2013). While the ecosystem assessment protocol was only recently developed—only three African ecosystems have been assessed as of mid-2019—its holistic strategy promises a more comprehensive accounting of local biodiversity which could be more informative than an accumulation of single species assessments. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.05%3A_Which_Species_are_at_Risk_of_Extinction.txt |
While a great number of factors may make a species vulnerable to extinction, conservation biologists have observed that species most vulnerable to extinction generally fall under one of six main groups:
• Species with small populations: Some species have very small populations, consisting of just a few individuals. Such small populations are highly vulnerable to random variations in demography or environmental conditions, and to the loss of genetic diversity—all factors that increase the risk of extinction (Section 8.7). Species whose population sizes naturally fluctuate between large and small populations also fall in this category, as they are at an increased risk of extinction during the small population phases of those fluctuations.
• Species with declining populations: Trends in population sizes tend to persist, so populations that are declining in abundance face a high risk of extinction (Caughley, 1994) unless conservation managers identify and address the causes of decline. Species impacted by the threats discussed in Chapters 5–7 generally also have declining populations.
• Species with restricted distribution ranges: Some species, such as those that are restricted to oceanic islands; mountains peaks; or isolated lakes, can be found only in a limited geographic range. A major disturbance, such as a cyclone/hurricane or drought, could easily affect that entire species’ range, potentially driving the species to extinction.
• Species with only one or a few populations: A sufficiently large disturbance—such as a wildfire, storm, or disease outbreak—can wipe out a single population of a species. For a species with only one population, that means its extinction, while the loss of even a single population leaves species with only a few populations more vulnerable to the next disturbance. Species in this category (few populations) overlap with those in the previous category (restricted distribution ranges) because species with few populations tend to have restricted ranges.
• Species that are exploited by people: Overharvesting can easily reduce a population to the point of extinction (Section 7.2). Even if overharvesting is stopped just before the point of extinction, it may still have reduced a population to a size where it becomes susceptible to one or more of the three additional pressures faced by small populations (Section 8.7).
• Species with critical symbiotic relationships: Species that are members of obligate symbiotic relationships (where one species cannot survive without another) will go extinct if its host disappears. For instance, larvae of the rhinoceros stomach botfly (Gyrostigma rhinocerontis) mature in the stomach lining of African rhinoceros, and no other species (Barraclough, 2006). Thus, if the host species (the rhinoceros) were to go extinct, so would the botfly, Africa’s largest fly species. This phenomenon in which one species’ extinction leads to the extinction of other is called a coextinction (Koh et al., 2004), while a series of linked coextinctions is called an extinction cascade (Section 4.2.1).
The following characteristics are also linked with extinction, although the links are not as strong as is the case with the previous six categories:
• Animal species with large body sizes: Large animals generally require large ranges and more food, have lower rates of reproduction, and have smaller population sizes relative to smaller animals. Often, they are harvested by humans for material benefits (see Box 8.1). Consequently, within groups of related species, the largest are generally also the most vulnerable to extinction—that is, a larger species of carnivore, ungulate, or whale is more likely to go extinct than a smaller carnivore, ungulate, or whale.
• Species that require a large home range: Individuals or social groups of some species must forage over wide areas to fulfil their needs. When portions of their range are being degraded or fragmented, the remaining area will eventually be too small to support a viable population.
• Species that are poor dispersers: Moving to more suitable habitat is a common survival response following altered environmental conditions. But species with poor dispersal abilities may be doomed to extinction if they are unable to move to more suitable areas elsewhere (see e.g. discussion on range-shift gaps, Section 6.3.5).
• Seasonal migrants: A migratory species depends on intact ecosystems at two or more locations to complete its life cycle (see Box 5.3). If those ecosystems, either at stop-over sites along migration routes and/or at migratory endpoints, are damaged, the species may be at risk of extinction.
• Species with low genetic diversity: Because genetic diversity (Section 3.2) enables species to adapt to changing environmental conditions, species with low genetic diversity are more vulnerable to extinction because they have less ability to adapt to new diseases, new predators, or recent changes in their ecosystems.
• Species that evolved in stable ecosystems: Species that evolved in relatively stable environments (e.g. tropical ecosystems) are often threatened with extinction because under stabile conditions, a species is unlikely to retain the ability to adapt to environmental changes such as altered microclimates.
• Species with specialised requirements: Specialist species are often threatened with extinction because they are unable to adapt to altered ecosystems.
• Group-living species: A range of factors leaves group-living species at risk of extinction. For example, a herd of ungulates, a flock of birds at their night-time roost, or a school of fish can be harvested in its entirely by people using highly effective techniques. Even if some individuals remain, the harvesting may still leave the population below a critical threshold needed for effective foraging, mating, or territorial defence. This link between population size/density and individual fitness is termed the Allee effect (Section 8.7.2).
• Species that have had no prior contact with people: Species that encounter people for the first time are ecologically naïve—they lack avoidance strategies that promote survival during these encounters. Ecologically naïve species thus have a higher chance of extinction than species that have already survived human contact.
• Species closely related to species that recently went extinct: Groups of closely-related taxa, where some members are threatened or already extinct, often share characteristics that elevate their threat of extinction. Groups of related taxa that include many threatened species include apes, cranes, sea turtles, and cycads.
• Species that live on islands: Island species generally exhibit many of the characteristics mentioned above. In addition, the mere fact that an island is surrounded by ocean means that species that are unable to swim or fly have nowhere to go when they need to escape a threat. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.06%3A_Characteristics_of_Threatened_Species.txt |
While some small populations have persisted against the odds, sufficiently large populations are generally needed to prevent eventual extinction (Halley et al., 2016, see also Section 9.2). Small populations—which include species that have always had small populations and previously large populations that have been reduced to a few individuals—face three additional inherent and unavoidable pressures beyond the threats discussed in Chapters 5–7. These three additional pressures are: (1) loss of genetic diversity; (2) demographic stochasticity; and (3) environmental stochasticity and natural catastrophes. We will now examine how each of these pressures can lead a small population to eventual extinction. Much of this discussion is based on a ground-breaking manuscript by New Zealand ecologist Graeme Caughley, which discusses at length the threats faced by small and declining wildlife populations (Caughley, 1994).
Loss of genetic diversity
Small populations are at risk of losing genetic variation much faster than large populations.
Species with high genetic diversity are generally more able to adapt to and reproduce under new conditions such as those brought by environmental changes (Section 3.2). These adaptations can occur at both individual and population levels. For example, under climate change, some genes may allow some populations to adapt their ranges faster or better tolerate warmer and wetter environments, while phenotypic plasticity—the ability of one gene to express itself differently under different conditions—may allow certain individuals to better adapt to a changing environment. One species that displays remarkable phenotypic plasticity is the crystalline iceplant (Mesembryanthemum crystallinum); by regulating its photosynthetic pathways, an individual plant can adjust its water needs based on the amount of salt and moisture available in the environment (Tallman et al., 1997). Such flexibility may explain why this species, native to southwestern Africa, North Africa, and Europe, has been a successful invader in environments as diverse as those in South America, North America, and Australia.
While populations with many individuals usually also have high levels of genetic diversity, small populations regularly suffer from low levels of genetic diversity. This low genetic diversity not only leaves those populations unable to adapt to changing conditions, but also makes them more susceptible to a variety of deleterious genetic effects (Caughley, 1994). Each of these effects leads to even greater loss of fitness and genetic diversity, hence even larger population declines, and eventually extinction. In the next sections, we discuss further why these deleterious genetic effects are so harmful to small populations.
Genetic drift
In wildlife populations, there are always some alleles that are relatively common, and others that are relatively rare. The relative abundance of any of these alleles may however change from one generation to another purely by chance. While common alleles generally tend to stay common, rare alleles have a high chance of being randomly lost in subsequent generations. Consider how each parent only passes on half of their genetic code to each offspring; this means that the ability of a rare allele to persist is dependent on how many individuals carry it, which individuals produce offspring, and how many offspring those individuals produce. Another important factor is population size (Figure 8.8): in any small population, only a limited number of individuals can carry any single allele, so the smaller the population, the higher the likelihood that alleles are lost to the next generation. This loss of alleles is called genetic drift.
While genetic drift equates to a loss of genetic diversity, there are some cases where populations show no obvious ill effects. Such may have been the case for female elephants in South Africa’s Addo Elephant National Park. Hunting once nearly killed off this entire population; by the time they were adequately protected in 1931, only 11 animals remained, eight of which were female. Of those eight females, at least four were tuskless, while only two, maybe three, females carried both tusks. Over the next decades, Addo’s female elephants have shown increasing degrees of tusklessness; by 2002, only 2% of females had tusks (by comparison, 96–98% of elephant females are normally expected to develop tusks, Maron, 2018). One can therefore postulate that the allele responsible for the tusk development in female elephants became rare, and that the progressive loss of tusked females is a sign of genetic drift (Whitehouse, 2002). While Addo’s female elephants do not show any known limitations from being tuskless, the loss of alleles can also be devastating to the population suffering from genetic drift if, for example, the lost allele(s) coded for traits that would have allowed a species to adapt to a changing environmental condition.
It is important to note that genetic drift is distinct from natural selection. That is, genetic drift involves random changes in the frequency of alleles, whereas natural selection involves changes in traits in response to sexual selection or specific environmental conditions. For example, reduced tusk size in some heavily-hunted elephants in Africa (e.g. Chiyo et al., 2015) is a selective pressure in response to hunting that favour large tusks—this is distinct from Addo’s female elephants that have lost their tusks even in the absence of selective hunting pressure.
Mating among closely related individuals, which occurs in small populations, often results in lower reproductive success and weaker offspring.
Inbreeding depression
In large populations, a variety of instinctive mechanisms are in place to promote heterosis, which occur when offspring have a level of genetic variation that improves their individual evolutionary fitness. Some species are predisposed to disperse from their place of birth to prevent sibling–sibling or parent–offspring mating, while others are restrained from mating with close relatives through sensory cues such as individual odours. Many plants have morphological and physiological traits that facilitate cross-pollination and reduce self-pollination.
However, in small populations with few unrelated mates, the urge to breed might be stronger than the mechanisms that promote heterosis. Under these conditions, rather than forgoing reproduction, breeding among closely-related individuals (or inbreeding) can occur. This breeding among close relatives might result in inbreeding depression, which can occur when closely-related parents give their offspring two copies of a deleterious allele. Individuals suffering from inbreeding depression typically have fewer offspring or have offspring that are weak or fail to reproduce. Such is the case for some mountain gorillas (Gorilla beringei beringei, EN): genetic studies have shown how birth defects in several small populations can be attributed to inbreeding depression (Xue et al., 2015). Inbreeding depression has also been identified as the reason why some small lion populations are more susceptible to diseases (Trinkel et al., 2011). Inbreeding depression can result in a vicious cycle for declining population sizes, where such declines can lead to even more inbreeding depression, and eventually extinction (see Section 8.7.4).
Outbreeding depression
Large populations have many ecological, behavioral, and physiological mechanisms that prevent hybridisation, the production of offspring among genetically distant taxa, whether they be individuals of different species, or individuals of the same species but with different adaptations (the latter being intraspecific hybridisation). As with inbreeding depression, these mechanisms may fail in small populations, leading to outbreeding depression (Frankham et al., 2011). Because offspring that result from outbreeding depression have traits that are intermediate to their parents, they may not be adapted to either of the parents’ ecosystems. For example, one study found that plants suffering from outbreeding depression have weakened defences against herbivory (Leimu and Fischer, 2010). Outbreeding depression may also lead to a breakdown in physiological and biochemical compatibility between would-be parents—hybrid sterility is a well-known consequence of this breakdown. Consequently, species and populations suffering from outbreeding depression often show similar symptoms to inbreeding depression, including lower fitness, weakness, and high rates of mortality.
The opposite of outbreeding depression is hybrid vigour. Under these conditions, the hybrid offspring can be quite strong in an evolutionary sense; they may even outcompete their parent species. Such is the case with the South African endemic black wildebeest (Connochaetes gnou, LC); having recovered from near-extinction, poorly planned translocations are now threatening this species, which readily hybridises with the widespread common wildebeest (Connochaetes taurinus, LC) in areas of contact (Grobler et al., 2011).
Population bottlenecks
In some taxa, such as butterflies, annual plants, and amphibians, population size varies dramatically from generation to generation. During some years, populations can be so large that they appear to face little risk of extinction. However, abundant years can be misleading when followed by successive years of low abundance. Generally, in a population that undergoes extreme size fluctuations, the population size required to ensure continued persistence (i.e., the minimum viable population (MVP), Section 9.2) is in effect much nearer the lowest than the highest number of individuals in any given year. However, during years with low abundance, a phenomenon known as a population bottleneck may occur—that is, the small population size may lead to the loss of rare alleles from one generation to the next. Population bottlenecks may lead to more inbreeding depression which, in turn, reduces reproductive success (Heber and Briskie, 2010) and increases vulnerability to diseases (Dalton et al., 2016). Low genetic diversity in great white sharks (Carcharodon carcharias, VU) living in South Africa’s Indian Ocean is thought to be the result of a population bottleneck (Andreotti et al., 2015).
Populations founded by only a few individuals by definition start off with low genetic diversity, having lasting effects in the population through time.
New populations founded by only a few individuals are vulnerable to a special type of population bottleneck, the founder effect. The founding individuals of a new population by definition start off with low genetic diversity, much less than the original population that the founders left behind. This low genetic diversity puts the new population at risk of further genetic diversity declines, which have lasting effects through time. This situation can occur naturally when only a small number of individuals disperse to establish a new population or when founder individuals come from a small population that already suffered from low genetic diversity. Being mindful of these concerns is especially important for translocation (Section 11.2) or captive breeding (Section 11.5) projects. For example, to prevent extinction of the world’s smallest gazelle, the Speke’s gazelle (Gazella spekei, EN), a captive population of this species, almost entirely restricted to Somalia, was established in the USA. The founder population for this captive breeding project consisted of only one male and three females, leading to severe levels of inbreeding depression and high mortality rates in offspring (Kalinowski et al., 2000). Understanding the importance of managing for genetic diversity can help avoid these and other challenges that can threaten the success of translocation projects.
Demographic stochasticity
Demographic stochasticity (also known as demographic variation) refers to random variations in a population’s demographic traits (e.g. sex ratios, birth rates, death rates), the cumulative effect of variation in individual organisms’ fitness. In any natural population, some individuals will produce fewer offspring than average, while others will produce more than average; some individuals will produce no offspring at all. Similarly, some individuals die younger than average, while others live longer than average. For populations that are sufficiently large, average birth and death rates provide relatively stable descriptions of key aspects of that population’s demography. However, when a population’s size decreases to below a certain threshold, variations in fitness of a small number of individuals can have a large impact on the overall populations’ demographic parameters, causing population size and other characters to fluctuate up or down unpredictably (Schleuning and Matthies, 2009). Consider, for example, an isolated population of crocodiles with only a few females. As with many other reptiles, offspring sex ratios of crocodiles are determined by the environmental temperature during incubation (Hutton 1987). If, by chance, the population experiences two years of high temperatures, which favour male offspring, and the few females die by chance, the all-male population may be doomed for extinction unless some female crocodiles immigrate from elsewhere.
The social systems of group-living animals can easily be disrupted when their population size or density falls below a critical level.
Small population sizes or low densities can also disrupt social interactions among individuals—especially interactions that affect reproduction—which can cause populations to become demographically unstable. This situation, referred to as the Allee effect, can result in further declines in population size, population density, and population growth rate. Obligate cooperative breeders, such as African wild dogs (Lycaon pictus, EN), are especially vulnerable to the Allee effect (Courchamp et al., 2000) since they need a certain number of individuals to protect their territories and obtain enough food for their offspring (Figure 8.9). Allee effects might also prevent impact group-living species that are not cooperative breeders—recalling the “safety in numbers” mantra, Allee effects seem to prevent the recovery of locally-rare sable antelope (Hippotragus niger, LC) populations in South Africa’s Kruger National Park, as reduced herd sizes increases their exposure to predation (Owen-Smith et al., 2012). But even solitary species that live at low densities are susceptible to Allee effects, since they may find it hard to locate mates once the population density drops below a certain level.
Environmental stochasticity and catastrophes
Environmental stochasticity, the unpredictable variation in environmental conditions, can cause dramatic population size fluctuations over time, and hence, substantially increase the risk of extinction. Consider, for example, how the development rate of many insects is strongly temperature-dependent (e.g. Rebaudo and Rabhi, 2018). In an average or warm year, young insects that hatch on time and feed well may result in ecologically fit adults that produce many young, whereas unusually cold years might reduce hatching success and larval activity, which could also reduce adult fitness (Gibert et al., 2001). So, highly unfavourable conditions in any one year can cause dramatic population declines, or even push a species to extinction if conditions persist over successive years across its range.
Even though a small population may appear to be stable or increasing, an environmental catastrophe can severely reduce population size or even cause extirpation or extinction.
The increased risk of extinction from environmental stochasticity also applies to natural catastrophes that can occur at unpredictable intervals (e.g. droughts, storms, earthquakes, and fires). Range-restricted species are particularly vulnerable to this kind of threat. For example, the biodiversity living in and around several African crater lakes are vulnerable to a rather unique natural phenomenon called “lake burping”. Volcanic chambers underneath some of these lakes are rich in CO2. Small amounts of CO2 may sometimes (or constantly, in some cases) seep up through the lake bed into the surrounding water. Because these lakes are thermally stratified—layers of cold, dense water settle near the bottom while warm, less dense water floats near the top—the CO2-saturated water remains near the bottom of the lake. However, when there is a geologic disturbance, such as a landslide or earthquake, massive amounts of CO2 may suddenly be released, first saturating the warmer water at higher levels with CO2 (killing fish and other oxygen-dependent species in the process), before displacing the breathable surface air in and around the lake. In 1986, one such CO2 eruption killed 1,800 people and 3,500 heads of livestock near Cameroon’s Lake Nyos (Krajick, 2003). Some scientists fear that increased deforestation (which may trigger erosion and landslides) and hydraulic fracturing (which may trigger earthquakes, Section 7.1.1) could trigger similar events at other crater lakes in the region.
Environmental stochasticity tends to increase the probability of extinction more than does demographic stochasticity. As discussed, this is especially true for small populations and range-restricted species.
The extinction vortex
As populations decline in size, they become increasingly vulnerable to the combined impacts from the loss of genetic diversity, inbreeding depression, Allee effects, environmental stochasticity, and demographic stochasticity. All these factors tend to lower reproduction, increase mortality rates, and reduce population size even more, in turn driving populations to extinction at increasingly faster rates over time (Fagan and Holmes, 2006). Conservationists sometimes compare this phenomenon to a vortex, spiralling inward, moving faster (or declining faster in the case of a population) as it gets closer to the centre. At the centre of this extinction vortex (Gilpin and Soulé, 1986) is oblivion—the extinction of the species (Figure 8.10).
The demise of the bluebuck—the first large mammal of Africa to face this fate after European colonisation—may have been the result of an extinction vortex. When European colonists first arrived in South Africa, this ungulate already persisted as a single, small population of an estimated 370 individuals (effective population size at 100 individuals) and a highly restricted (4,300km2) distribution. Considering this small and restricted population’s vulnerable to deleterious genetic factors and demographic stochasticity, a recent study showed that this species was probably caught in an extinction vortex by the time the first colonist shot the first bluebuck (Kerley et al., 2009). This species would thus likely have gone extinct even in the absence of hunting and habitat loss, which only hastened its departure.
Is there any hope for small populations?
Despite the odds and the many threats facing Africa’s wildlife, many species that were once on the brink of extinction have clawed their way back from the abyss towards stable, and sometimes even growing populations. Prime examples include the Pemba flying fox (Pteropus voeltzkowi, VU); considered Critically Endangered in 1996, conservation education programs raised awareness of this unique bat, which now has considered Vulnerable, having recovered to more than 28,000 individuals (Entwistle and Juma, 2016). Similarly, because of habitat destruction and introduced predators, the future of the Seychelles magpie-robin (Copsychus sechellarum, EN) looked rather bleak in 1970, when only 16 individuals remained, all on one island. Today, thanks to habitat restoration efforts, supplemental feeding, invasive species eradication, provisioning of nest boxes, and a translocation program, there are more than 280 Seychelles magpie-robins scattered across five islands (Burt et al., 2016). Another remarkable conservation success story involves the rescue of the southern white rhinoceros (Ceratotherium simum simum, NT), which was reduced to about 20 individuals in a single protected area in the late 1880s. Dedicated conservation efforts since then have seen this iconic species recover to more than 20,000 individuals, with individuals introduced and reintroduced all over Africa and zoos throughout the world. None of these species would have been alive today if it wasn’t for intensive multi-year efforts by dedicated conservation biologists to pull them out of their individual extinction vortices.
Bringing species with small populations back from the edge of extinction requires dedication, careful planning, and significant amounts of resources.
Bringing species with small populations back from the edge of extinction requires dedication, careful planning, and significant amounts of resources. It also requires careful population management to mitigate the negative impacts of founder effects and both demographic and environmental stochasticity (Box 8.4; see also Chapter 11). As these examples show, it can be done. But, given the challenges, it should always be a priority to prevent a species from declining to very low numbers in the first place.
Box 8.3 Fenced Reserves Conserving Cheetahs and African Wild Dogs in South Africa
Kelly Marnewick1,2
1Carnivore Conservation Programme,
Endangered Wildlife Trust,
Johannesburg, South Africa.
2Current address: Department of Nature Conservation,
Tshwane University of Technology,
Pretoria, South Africa.
[email protected]
South Africa is one of the few countries in Africa where numbers of many large carnivore species are stable and, in some cases, increasing. Much of this success can be attributed to the managed metapopulation approach, which involves the reintroduction and subsequent translocation and management of populations in geographically isolated fenced reserves, between which natural dispersal is highly unlikely. As of 2016, more than 300 cheetahs are being managed in 51 reserves encompassing 10,995 km2 (mean: 195 km2 range: 20–1,000 km2) and nearly 250 African wild dogs in 11 reserves encompassing 5,086 km2 (mean: 216 km2 range: 19–1,000 km2). The reserves are situated across the country within a variety of land tenure systems including state and provincial protected areas and privately owned and community-run game reserves. Most reserves derive income primarily from ecotourism.
Each reserve forms part of the national network. Animals are moved between reserves to maintain the genetic integrity and demographic balance of individual subpopulations, but also to minimise direct management in the long term. Translocations are planned to mimic natural processes as far as possible but, due to the intricacies involved in managing animals between several reserves, this is not always possible. For wild dogs, small groups of unrelated adult males and females are artificially bonded to form packs, which mimics natural pack formation in the wild. For cheetahs, sub-adults are removed once they disperse from their maternal range. The animals are generally immobilised in the field and transported awake in crates on vehicles to their new reserves. Soft releases (Section 11.2.1) are preferred: these involve the animals being kept in temporary holding bomas of approximately 1 ha in size for about three months. The formation of artificial social groups is also done during this period. Intensive post-release monitoring is done at intervals reliant on reserve resources, but daily monitoring is recommended. The success rate of reintroductions has been high and, for wild dogs, has been strongly linked to the social cohesion of released groups (Marneweck et al., 2019), and the integrity of perimeter fences (Gusset et al., 2008).
This highly collaborative process involves multiple stakeholders, including conservation NGOs, provincial government conservation departments, private reserve owners and managers, researchers, local communities, and tourists. Effective and responsible population management tools help to prevent local populations growing too large or too small, and best practice guidelines ensure the ethical handling and management of animals. Individual reserves are responsible for providing infrastructure and other requirements including managing sustainable prey populations, perimeter fences, bomas and post release monitoring, as well as ensuring that a management plan is in place and adhered to. In many cases, students or volunteer organizations conduct post-release monitoring. National, high-level management is coordinated by the Endangered Wildlife Trust (EWT) and is funded through donations from corporations, individual philanthropists, conservation trusts, and foundations.
The managed metapopulation approach to carnivore conservation has increased the number and distribution of both cheetahs and African wild dogs in South Africa and built technical capacity in the country for metapopulation management (Davies-Mostert and Gusset, 2013), which has also been applied to species, such as lions, elephants, and black rhinoceros (Diceros bicornis, CR). Opportunities abound in other countries to use lessons learned in South Africa for the recolonisation of other areas where large mammals have been locally or regionally extirpated. Additionally, projected human population expansion, and the habitat fragmentation that comes with it, means that this approach is likely to become an indispensable tool in maintaining the viability of populations in disconnected landscapes. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.07%3A_Problems_of_Small_Populations.txt |
One of the more interesting conservation debates to have emerged in recent years involve efforts to reverse extinction. This field, known as de-extinction or resurrection biology aims to revive extinct species, and eventually to reintroduce viable populations to their original locations (Seddon, 2017). One possible method, called “breeding back”, aims to produce individuals genetically similar to an extinct species by selective breeding of extant species that carry genetic material of their extinct relatives. This is the main method currently being used to revive the aurochs (Bos primigenius, EX), the ancestor of today’s domestic cattle (Stokstad, 2015). Other “breeding back” projects place less emphasis on genetics and more on morphology, by selectively breeding individuals with certain traits to produce individuals that visually appear similar to the extinct species. Such is the case at The Quagga Project, where selectively breeding of plains zebras (Equus quagga, NT) with quagga-like characteristics (reduced striping and brown hues) are resulting in animals (Figure 8.11) that look increasingly like extinct quaggas (Harley et al., 2009).
The second popular method used for de-extinction is cloning. This involves the transfer of viable genetic material from an extinct species to the eggs (or embryo) of a closely related surrogate mother, who will hopefully give birth to an individual of the extinct species. Cloning has been used in selective breeding of livestock for many years, and plans are also currently underway to use cloning to prevent the extinction of highly threatened species such as the northern white rhinoceros (see Box 11.4). Despite the promise that cloning offers for reviving extant and recently extinct species, cloning species that went extinct many years ago has been more challenging. So far, attempts to clone Spain’s Pyrenean ibex (Capra pyrenaica pyrenaica, EX) and Australia’s gastric-brooding frog (Rheobatrachus silus, EX) have produced individuals that lived for only a few minutes (Ogden, 2014).
Despite the progress made, de-extinction is one of the most controversial and polarising debates to emerge among conservation biologists in recent years. Proponents of de-extinction hope that the early work described above paves the way for the resurrection of extinct species once the threats that drove them to extinction have been managed. Many resurrection biologists have even started establishing banks where genetic material of threatened species is cryopreserved for future use. They also hope that their work will inspire more people to be interested in science in general and especially in conservation. Protected areas with extinct species may even draw tourists that can fund conservation projects, while reintroductions of once-extinct species can revive lost ecosystem services.
Bringing a species back from extinction is, however, highly controversial for several reasons. First, there is the argument that the limited funds available for conservation are better spent on species currently facing extinction rather than on projects with possibly insurmountable technical challenges. Others argue that there is no point in spending millions of dollars to bring back an extinct species if we cannot even solve the extinctions drivers that caused the demise of the extinct species in the first place. Conservationists also wonder how the issues facing small populations will be managed, especially early in the process. Many believe that these small compromised populations will simply occupy valuable space in zoos and protected areas that can be better used for protecting extant species. There are also major misgivings about whether the resurrected species will fill the same ecosystem function as before since they may behave differently; in fact, some worry that unpredictable behaviors may introduce new harmful threats to ecosystems. Many conservationists are also worried that the public’s concern for species currently threatened may fade if there is a perception that we can simply revive the species after the last individual died. Lastly, ethical questions are frequently raised about humans essentially trying to “play God” with these “vanity projects”, and the possibility that this entire field will undermine one of the most important foundations of conservation biology—that we need to act now because extinction is forever. Clearly there are some advantages, but also disadvantages, to de-extinction. Most importantly, much research still needs to happen for this to be a viable idea. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.08%3A__Is_De-extinction_a_Solution.txt |
1. The rates of species extinction are currently 1,000 times greater than natural background levels; this may soon increase to 10,000 times. Over 99% of modern extinctions can be attributed to human activity. These extinctions are leading to the rapid loss of ecosystem services.
2. The IUCN has developed quantitative criteria that assign species to nine conservation categories based on their risk of extinction. Species that are Extinct in the Wild (EW), Critically Endangered (CR), Endangered (EN), and Vulnerable (VU) are officially considered “threatened with extinction”. The five other categories are Extinct (EX), Near Threatened (NT), Least Concern (LC), Data Deficient (DD), and Not Evaluated (NE).
3. Species with the following characteristics are particularly vulnerable to extinction: species with small populations, species with declining populations, species with restricted distribution ranges, species with one or only a few small populations, species that are exploited by people, and species with critical symbiotic relationships.
4. Small populations are at high risk of extinction because they are vulnerable to several deleterious genetic factors, as well as demographic and environmental stochasticity. Such populations often require intensive management to prevent them from becoming a victim of an extinction vortex.
5. De-extinction as a scientific field aims to revive extinct species using methods such as selective breeding and cloning. But this practice is controversial, and not practical with current technology.
8.10: Topics for Discussion
1. Why should you, or anyone else, be concerned if a species becomes extirpated (also known as locally extinct)? How does this concern compare to when a species becomes globally extinct?
2. Use the IUCN Red List (http://www.iucnredlist.org) to identify one species in your country that is currently threatened with extinction. How might this species be affected by the various challenges facing small populations? Consider genetic, physiological, behavioral, and ecological factors, as appropriate.
3. Think of an imaginary animal that was recently discovered. Imagine this animal is also threatened with extinction. Name three characteristics that make your imaginary animal vulnerable to its threats. Now discuss some steps that can be implemented to ensure that your animal will continue to survive.
4. A herd of 80 rhinoceros have been moved from South Africa to Australia to “save the species” because “there is no safe place in Africa for rhinos today” (see Hayward et al., 2017). What do you think of this plan? Do you think the project will be successful? What are the main opportunities and challenges? Once you’re done answering the question read Lundgren et al. (2017) and decide if you still feel the same.
8.11: Suggested Readings
Bonebrake, T.C., F. Guo, C. Dingle, et al. 2019. Integrating proximal and horizon threats to biodiversity for conservation. Trends in Ecology and Evolution 34: in press. https://doi.org/10.1016/j.tree.2019.04.001 Coordination is critical when combatting multiple threats.
Caughley, G. 1994. Directions in conservation biology. Journal of Animal Ecology 63: 215–24. https://doi.org/10.2307/5542 The classic overview of the challenges faced by small and declining populations.
Ceballos, G., P.R. Ehrlich, and R. Dirzo. 2017. Biological annihilation via the ongoing sixth mass extinction signaled by vertebrate population losses and declines. Proceedings of the National Academy of Sciences 114: E6089–E6096. https://doi.org/10.1073/pnas.1704949114 Extinctions by the numbers.
IBPES. 2019. Nature’s dangerous decline ‘unprecedented’: Species extinction rates accelerating. IBPES media release. https://www.ipbes.net/news/Media-Release-Global-Assessment We are in the midst of a biodiversity crisis.
IUCN. 2012. IUCN Red List Categories and Criteria: v. 3.1 (Gland: IUCN). https://portals.iucn.org/library/sites/library/files/documents/RL-2001-001-2nd.pdf A summary of the IUCN Red List classifications.
Jones, H.P. 2010. Seabird islands take mere decades to recover following rat eradication. Ecological Applications 20: 2075–80. https://doi.org/10.1890/10-0118.1 Control of invasive species can be an effective method to achieve recovery of threatened species.
Koh, L.P., R.R. Dunn, N.S. Sodhi, et al. 2004. Species coextinctions and the biodiversity crisis. Science 305: 1632–34. https://doi.org/10.1126/science.1101101 Thousands of species face extinction due to the decoupling of important symbiotic relationships.
McClenachan, L., A.B. Cooper, K.E. Carpenter, et al. 2012. Extinction risk and bottlenecks in the conservation of charismatic marine species. Conservation Letters 5: 73–80. https://doi.org/10.1111/j.1755-263X.2011.00206.x Most marine species have not been evaluated for their risk of extinction.
Stearns, B.P., and S.C. Stearns. 2010. Still watching, from the edge of extinction. BioScience 60: 141–46. https://doi.org/10.1525/bio.2010.60.2.8 Many threatened species increasingly rely on human actions to avoid extinction.
Tranquilli, S., M. Abedi-Lartey, F. Amsini, et al. 2012. Lack of conservation effort rapidly increases African great ape extinction risk. Conservation Letters 5: 48–55. https://doi.org/10.1111/j.1755-263X.2011.00211.x Neglecting species in conservation need increases extinction risk. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/08%3A_Extinction_is_Forever/8.09%3A_Summary.txt |
Andreotti, S., S. Heyden, R. Henriques, et al. 2016. New insights into the evolutionary history of white sharks, Carcharodon carcharias. Journal of Biogeography 43: 328–39. https://doi.org/10.1111/jbi.12641
Balon, E.K., M.N. Bruton, and H. Fricke. 1988. A fiftieth anniversary reflection on the living coelacanth, Latimeria chalumnae: Some new interpretations of its natural history and conservation status. Environmental Biology of Fishes 23: 241–80. https://doi.org/10.1007/BF00005238
Barnosky, A.D., N. Matzke, S. Tomiya, et al. 2011. Has the Earth’s sixth mass extinction already arrived? Nature 471: 51–57. https://doi.org/10.1038/nature09678
Barnosky, A.D., P.L. Koch, R.S. Feranec, et al. 2004. Assessing the causes of late Pleistocene extinctions on the continents. Science 306: 70–75. https://doi.org/10.1126/science.1101476
Barraclough, D.A. 2006. Bushels of bots. Natural History 115: 18–21.
Bartlett, L.J., D.R. Williams, G.W. Prescott, et al. 2015. Robustness despite uncertainty: Regional climate data reveal the dominant role of humans in explaining global extinctions of Late Quaternary megafauna. Ecography 39: 152–61. https://doi.org/10.1111/ecog.01566
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Even without human influences, the size of any wildlife population may be stable, increasing, decreasing, or even fluctuating. These population changes, combined with occasional natural perturbations, can and have driven some species and populations to extinction. Such natural extinction events generally occur at local scales, and are interspersed by long periods of little change, so that overall ecosystem stability is not compromised. Moreover, as explained by the intermediate disturbance hypothesis (e.g. Bongers et al., 2009), localised disturbances and subsequent local extinctions play an important role in maintaining regional biodiversity, as they increase opportunities for a greater variety of species to live in an area (Figure 9.1), at least until succession drives them out again. Some species that colonise the empty niches left by extinctions or extirpations may even evolve to become new species over time.
9.01: Monitoring Population Size
The primary aim of population monitoring is to detect changes in the environment, population size, and species distribution over time. Such monitoring efforts frequently focus on a particular area or a population of concern, but it can also target more common but sensitive species, such as butterflies and macroinvertebrates, which can function as indicator species to assess ecosystem condition (Section 4.2.6). The great number of methods (which are all types of surveys) used to monitor populations usually fall into one of three different categories: biodiversity inventories, population censuses, and demographic studies.
Biodiversity inventories
The primary aim of population monitoring is to detect changes in the environment, population size, and species’ distributions over time.
A biodiversity inventory is an attempt to document which species are present in some defined locality. Such an effort can focus on one specific taxa (e.g. a bird survey) or several taxa, on a small area (e.g. a city park) or large area (e.g. a large national park), over a short period of time (e.g. a few hours) or long period of time (e.g. several years, Box 9.1). There are many methods to compile a biodiversity inventory, ranging from uncomplicated to highly organised, performed by a single person or a large team of experts. Some of the most popular methods for biodiversity inventories include site visits by professional naturalists and questionnaires distributed among local people. To tap into the knowledge and eagerness of amateur naturalists, conservation biologists are also increasingly compiling biodiversity inventories using citizen science surveys (see Box 15.3). Rapid biodiversity assessments (RAP) are sometimes used to compile an inventory under tight deadlines to answer urgent questions and inform urgent decisions. A bioblitz is a special type of biodiversity inventory during which experts on a range of taxa come together to record all the living species within a designated area over a brief period (usually over 24 hours).
Box 9.1 The Role of Biodiversity Inventories in the Management of Gorongosa National Park
Marc Stalmans1 and Piotr Naskrecki2
1 Scientific Services, Gorongosa National Park,
Mozambique.
2 E.O. Wilson Biodiversity Laboratory, Gorongosa National Park,
Mozambique.
The 4,000 km2 Gorongosa National Park in central Mozambique was proclaimed in 1960 to protect one of the highest densities of large herbivores at the southern end of Africa’s Great Rift Valley (Tinley, 1977). National Parks are often victims of war and political instability and Gorongosa National Park is no exception. It suffered grievously during the protracted period of civil war from the early 1970s to early 1990s. During this time the park lost 90–99% of its elephants, common hippopotamuses (Hippopotamus amphibious, VU), African buffalo (Syncerus caffer, NT), plains zebras, and common wildebeest (Connochaetes taurinus, LC) through poaching by warring parties and hunters from nearby cities and rural communities.
Since then, restoration efforts that started in earnest in 2004 have brought about a spectacular recovery of several affected large mammal populations (Bouley et al., 2018; Stalmans et al., 2019). But conservation management cannot only focus on these flagship species. Considering, amongst others, the impact of climate change and the importance of agriculture (with its associated pests and pollinators) to rural communities, it is vital that conservationists understand the breadth of biodiversity and its relationship to ecosystem functioning. To accomplish this, a program of systematic biodiversity surveys is currently being undertaken in Gorongosa. Each year, a group of international and national specialists team up with park technicians and rangers to conduct a three-week long bioblitz in a subsection of the park. These surveys also serve as training opportunities to prepare young Mozambican scientists to apply modern biological survey methods and technologies.
By the beginning of 2019, a total of nearly 5,900 species represented by some 44,000 observations have been entered into the park’s biodiversity database. Based on these data, initial projections suggest that Gorongosa protects 37,500–76,500 different species. Vertebrates are likely to number 850–1,000 species, while plants are estimated to number 2,000–3,000 species. Single orders of insects far exceed those numbers; for example, there may be 3,000–5,000 species of wasps, and 4,000–6,000 species of moths of which 15–25% may be new to science. Local ecosystems are also particularly rich. For the surveys around the Bunga inselbergs (Figure 9.A) in 2015, at least 580 species of butterflies and moths were collected, most of them never before recorded from the Park. Forty species of katydids were recorded, with two species of significance. A large population of Debrona cervina, a large arboreal katydid was discovered, until now known only from two type specimens collected in 1890. Also collected was the predaceous katydid Peringueyella macrocephala, Mozambique’s largest katydid, previously known only from a handful of specimens collected between 1850 and 1965. About 100 species of grasshoppers were recorded, including two species new to science. Additionally, about 30 species of mantids were recorded, including Rhomboderella thorectes, a species previously known from the single holotype collected in the early 1900s. It is expected that the full inventory of the park’s biodiversity will span a period of 20 years.
Biodiversity inventories play an important role in the management of Gorongosa and other national parks, especially in long-term conservation planning. Amongst others, baseline data obtained from our surveys will be used in future to measure overall biodiversity responses to large mammal population changes, evolving patterns of land use around the park, and the impact of climate change. Future biodiversity surveys will also target little known areas adjacent to the park to provide information required for corridor planning.
While biodiversity inventories seldom offer the kinds of detailed data required to predict likelihood of a species’ persistence, they have several uses in conservation. First, a biodiversity inventory can be a comparatively inexpensive and straightforward method to broadly monitor an area’s species and populations. Biodiversity inventories conducted over a wide area can also help determine the distribution of a species, while a comparison with follow-up inventories can highlight distribution changes (which often correspond to population changes). This was well illustrated in a study that used repeated citizen scientist surveys to investigate how songbird distributions have changed across South Africa, Lesotho, and eSwatini between 1987 and 2013 (Péron and Altwegg, 2015).
Population censuses
A population census (also called a count) uses a repeatable sampling protocol to estimate the abundance or density of a population or species which, in turn, can tell us whether a population is doing well or not. When a species is easy to detect, relatively sedentary, and the sampling area is small, a comprehensive census of all individuals may be possible. However, comprehensive censuses are generally very difficult, if not impossible, to conduct when implemented on large or highly mobile populations, or over large areas. In these cases, it may be better to restrict the census to a more manageable subsection of the population, by dividing the area of interest into sampling units, and randomly censusing only some of the units. Population estimates that capture only a fraction of the overall population can then serve as an index for broader trends, or it can be used to estimate the total population size through extrapolation, if the researcher knows which fraction of the population or area was counted.
Some popular methods for censusing subsections of wildlife populations are, sampling plots, distance sampling, and mark-recapture surveys. Sampling plots are popular in studies focussing on plants and invertebrates, allowing biologists to systematically count each individual observed in a small area (Figure 9.2). Birds and mammals are often censused using distance sampling, during which all observed individuals on predetermined transects or from points are tallied. The number of individuals observed in the count area can then be extrapolated to obtain population size (or density) estimates for individual (or multiple) species observed across the entire area of interest. Aerial censuses are often used to conduct distance sampling transects over large and open areas, while point counts and walked line-transects are more popular for small areas or closed-canopy ecosystems (White and Edward, 2000). Mark-recapture surveys, mark-resight surveys, and sight-resight surveys are popular for species that are easy to catch, trap, or individually recognized. In this case, captured (and thus counted) individuals would be marked for future identification, after which the total population in an area is estimated by accounting for the proportion of marked and unmarked individuals seen on subsequent visits. The marking of animals can be done with a variety of procedures, including using highly visible tags, paint approved for animal use, or unique marks on the animal itself. One creative study used tourists’ photographs to generate a mark-recapture dataset, which was used to estimate the size of cheetah and African wild dog (Lycaon pictus, EN) populations in South Africa’s Kruger National Park (Marnewick et al., 2014). Like inventories, population censuses can sometimes also lead to unexpected yet important findings: the first comprehensive population survey of sea turtles breeding on Africa’s Atlantic coast recently alerted marine biologists to the fact that Gabon hosts several globally important rookeries (Box 9.2).
Box 9.2 Sea Turtle Conservation along Africa’s Atlantic Coast
Angela Formia
Wildlife Conservation Society,
Global Conservation Program,
New York, NY, USA.
[email protected]
Virtually all the characteristics of sea turtles’ life histories make them difficult to study and conserve. They are long-lived, slow growing, migratory, and almost entirely ocean-dwelling. Although they return to their natal beaches to reproduce, these are usually thousands of kilometres from their developmental and adult foraging grounds. In addition, sea turtle habitat often overlaps with areas of high human use such as developed coastlines and intensive fisheries. Describing population ranges and assessing interaction with human threats is thus critical to their survival.
Over recent decades, we have learnt much about sea turtles along the coastline of Africa (Figure 9.B) thanks to extensive research efforts. For instance, we know that these coasts host globally important populations of green turtles (Chelonia mydas, EN) in Mauritania, Guinea Bissau, Equatorial Guinea and Republic of the Congo; loggerheads (Caretta caretta, VU) on Cabo Verde; hawksbills (Eretmochelys imbricata, CR), on São Tomé and Principe; leatherbacks (Dermochelys coriacea, VU) in Equatorial Guinea and Gabon; and olive ridleys (Lepidochelys olivacea, VU) in Gabon and Angola.
One of Africa’s most remarkable sea turtle populations is Gabon’s leatherback rookery, the biggest in the world with as many as 15,000 to 41,000 nesting females (Witt et al., 2009). Gabon also hosts the largest olive ridley rookery in the Atlantic (Metcalfe et al., 2015), and foraging grounds for green and hawksbill turtles. Until the late 1990s, virtually nothing was known about these populations, other than the fact that eggs and adults were frequently collected for human consumption. Since then, a multi-pronged approach has been adopted to describe and protect Gabon’s sea turtles. Intensive coastal monitoring has allowed scientists to assess spatio-temporal trends in nesting frequency and abundance, and levels of nest-site fidelity and reproductive success. Using techniques, such as satellite telemetry, flipper tagging, oceanic modelling, and dispersal simulations, and genetic and isotopic analyses, researchers have been able to map sea turtle behavior at sea, in Gabon’s coastal waters, and during post-nesting migrations to foraging grounds off South America and South Africa (i.e. Formia et al., 2006, Maxwell et al., 2011, Witt et al., 2011, Pikesley et al., 2018).
Building upon this knowledge, measures have been established to quantify and reduce the impact of threats to Gabon’s sea turtles. In 2002, the Gabonese government created a system of national parks and protected areas encompassing approximately 80% of Gabon’s sea turtle nests; in 2017, a new network of 20 marine protected areas (MPA) was officially created, covering 26% of Gabon’s territorial waters (Parker, 2017). Laws enacted in 2011 prohibit all hunting, capture, and commercialisation of sea turtles. Trained observers on-board industrial fishing vessels quantify sea turtle bycatch from bottom trawling and tuna seiners and reduce mortality by treating and releasing captured turtles. In addition, the Gabonese government now requires that all shrimp trawlers use turtle excluder devices (TED), aluminium grids sewn into the nets allowing sea turtles and other large bycatch to escape, while conserving shrimp catch; similar devices are being developed for fish trawlers. Ongoing efforts are shifting traditional turtle hunting and other destructive practices toward more sustainable fisheries. Turtle-watching ecotourism also represents a growing potential to increase awareness and incentivize conservation efforts.
Nevertheless, African sea turtle conservation remains a formidable challenge. Although the economic context is changing rapidly, impoverished coastal villagers in many countries continue to collect turtles and eggs for local consumption or market sale, and many wealthier urbanites continue to consider them delicacies. These problems are often compounded by corruption, political instability, inadequate law enforcement, and development priorities focused on destructive exploitation. With funding deficits, combating these challenges sometimes seems like a losing battle, but public attitudes are slowly shifting. Even in remote beach villages, the idea that a turtle alive is worth more than dead is no longer such a bizarre concept.
Demographic studies
Demographic studies monitor individuals of different ages and sizes over time (Figure 9.3) to obtain a more comprehensive dataset than would be produced by population censuses. Most demographic studies use the same methods that what would be used for a population census; however, in addition to counting and marking, individuals would also be aged, measured for size and body condition, and sexed, when possible. The best demographic studies involve collecting these data from the same individuals over time, which is easiest when working with sedentary species (e.g. plants), populations in an enclosed space (e.g. in a small fenced reserve), animals that are fairly resident and/or habituated to human presence, or individuals carrying biologging devices (Kays et al., 2015). This may not always be possible, in which case biologists may obtain data from different individuals during each field session, to serve as an index for larger population trends.
The data obtained from demographic studies are often used in combination with mathematical modelling to guide and refine conservation strategies. For example, researchers frequently compare the age structure (i.e. the percentage of juveniles, reproductively active adults, and older post-reproductive-age adults) of a declining population to that of a stable population to identify causes of decline, and the population parameters that are most sensitive to disturbances. This information can then be used to predict population sizes at different points in the future, and how those populations may respond to different management scenarios. The aim of many demographic studies is to predict, and identify strategies to reduce, extinction risk (see Section 9.2).
Recent progress in collecting survey data
Conservation activities are regularly impeded by insufficient information. This is especially true in tropical regions of the world, where most threatened species lack demographic data, and some species lack reliable data altogether. Faced with these gaps, biologists have started using several innovative methods to fill data gaps. Prominent examples include using market surveys (e.g. Kümpel et al., 2010, Ingram et al., 2015) and interviews with local people (e.g. Edwards and Plagányi, 2008) to obtain much-needed baseline survey data. It is important to note that such datasets, obtained second-hand rather than directly, can be unreliable and biased, especially if data are collected from harvesters unwilling to report on their own illegal activities. It is thus important to combine potentially unreliable datasets with reliable datasets, or obtain independent verification, before using such data to make important decisions. One such example comes from West Africa, where researchers wanted to quantify extinction risk for the Nigeria-Cameroon chimpanzee (Pan troglodytes ellioti, EN). Here, biologists related unreliable market survey data to two reliable datasets—orphan intake rate at wildlife sanctuaries and the number of young in wild groups—to estimate that the region’s chimpanzees might be extinct within the next 20 years because hunting was two to 13 times higher than the population could sustain (Hughes et al., 2011).
Camera traps, hair snares, and faecal samples all provide non-invasive sampling techniques to obtain baseline data needed for conservation assessments.
Collecting genetic material on elusive and rarely-seen animals with non-invasive techniques such as hair snares and faecal sampling are also becoming increasingly popular means of collecting survey data. Researchers in Gabon did just that, by using genetic material obtained from dung to estimate the population size, gender ratio, age distribution, breeding status, relatedness, and dispersal patterns of the region’s forest elephant (Loxodonta cyclotis) population (Eggert et al., 2013). These non-invasive techniques reduce the need for researchers to be in the field, thereby reducing both the researchers’ exposure to dangerous conditions and disturbances to the populations they are trying to monitor.
Camera traps represent another non-invasive survey technique whose popularity has greatly increased in recent years. These special cameras, often placed at supplemental food or next to wildlife paths, are activated automatically when an animal passes into the area covered by the camera’s motion sensors (Figure 9.4). This photographic record of movement can then be used to obtain biodiversity inventories, population size estimations, or even to compile demographic datasets (Steenweg et al., 2017). Creative researchers at South Africa’s Robben Island even successfully combined camera trapping with human facial recognition technology—more generally associated with law enforcement—to automate monitoring of individual African penguins (Spheniscus demersus, EN) (Sherley et al., 2010). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/09%3A_Applied_Population_Biology/9.00%3A_Prelude_to_Applied_Population_Biology.txt |
Biologists often use the positive relationship between population size and likelihood of persistence (Section 8.7) to predict the probability that a population may go extinct at some point in the future. One of the most popular tools for making such predictions is population viability analysis (PVA). A PVA can be thought of as a type of extinction risk assessment; it uses demographic data and mathematical methods to predict at what point in the future a population or species is likely to perish. In addition, by considering a species’ resource requirements and the availability of limiting resources, biologists can use the results of a PVA to identify a species’ most vulnerable life stages, and to estimate how management techniques may influence population size and extinction risk. In this way, PVAs can guide conservation decisions by highlighting the need to, for example, modify harvesting regulations, perform translocations (Section 11.2), or provide and protect a greater amount of suitable habitat. Even the IUCN’s Red List Criteria (Section 8.5) uses PVA as a criterion to help prioritise conservation targets: populations with low extinction risk may not require immediate attention, while those approaching extinction thresholds will gain higher priority.
A word of warning
The purpose of this chapter is to provide a brief introduction on the usefulness of quantitative population biology methods such as PVA in conservation. While the methods for studying population sizes, fluctuations, and demographics are very powerful, they are also highly technical, and require specialist knowledge of mathematical procedures. Erroneous predictions from using incorrect methods, violating assumptions, and/or using inadequate data would run counter to well-intentioned objectives; and so, the increased popularity and use of PVA by insufficiently trained conservation scientists is of serious concern. For many people, the methods highlighted in this chapter are best learned by studying under the supervision of an expert, to better understand each model’s assumptions and the newest developments in the field. For people with advanced mathematical skills who might want to study more on their own, texts such as Quantitative Conservation Biology (Morris and Doak, 2002) and Bayesian Methods for Ecology (McCarthy, 2007) may help. The development of user-friendly software packages, such as VORTEX and RAMAS (reviewed in Brook et al., 2000), have also expanded the PVA user-group in recent years. Nonetheless, when obtaining results—even from seasoned demographic modellers—it is important to remember that we cannot account for all future possibilities. Interpreting the results of a PVA, as any other model predicting the future, requires a great deal of caution and a healthy dose of common sense.
Probability of extinction
Population viability analysis (PVA) uses demographic data and mathematical methods to predict if a population or species is likely to persist or perish.
The main purpose of a PVA is to estimate the viability (or time to extinction) of a species or population from observed population sizes and growth rates. Consider a population with 100 individuals that loses 50% of its individuals each year. A simple model will suggest that this population will lose 50 individuals the first year, 25 individuals the second year, and so on, until no individuals are left in the seventh year. The probability of extinction for this population is thus 100%, and the time to extinction is seven years. But how do we deal with the more realistic complex variations in population sizes we see in nature?
A more realistic PVA begins by constructing a mathematical model representing the population of interest using data obtained from a demographic study, which may include the current age (or size) structure of the population, average birth rates, and average survival rates of each age class. This dataset would be organised in a format suitable for PVA modelling using a database package, and then analysed using the methods of matrix algebra. Because results from this initial model have only one outcome—a population that is either stable or growing/declining at a fixed rate—it is called a deterministic model. Typically, deterministic models are then tailored to include a variety of independent environmental parameters, such as food availability, storm frequency, or invasive competitors. Variability can also be added into the model by allowing some or all the model elements (e.g. survival rate or habitat availability) to vary within their observed ranges of values. Catastrophic events, such as a fire that kills a large proportion of the population, can also be programd to occur at random points in time. Hundreds or even thousands of simulations of this complex model can then be run to determine changes in population size over time, the probability of population extinction within a certain period, and the median time to extinction. Because of the variability built into this more complex model, each iteration’s output will vary from the next; for that reason, it is called a stochastic model. The choice of models and the parameters included depend on the goals of the analysis and the management options under consideration.
Minimum viable population
When a PVA shows that a population has a relatively high risk of extinction, a logical next step would be to determine what could be done to prevent the extinction from happening. In general, protecting larger populations reduces extinction probability (Figure 9.5). To understand exactly how large is large enough, a PVA can also be used to estimate a minimum viable population (MVP). As the name implies, an MVP is the smallest number of individuals necessary for a population to have a chance of long-term persistence, despite the potential effects of demographic, environmental, and genetic stochasticity, and natural catastrophes faced by small populations (Section 8.7). This is well illustrated in the influential paper by Shaffer (1981), who compared setting MVP targets to planning for floods; engineers cannot rely on the average annual rainfall when designing flood control systems near rivers and wetlands. Instead, they must design systems that can also handle extreme rainfall and flooding events. These extreme events may occur rarely, perhaps once every 50 years, but they will likely occur during the lifetime of a flood control system. Similarly, to maximise the long-term protection of a threatened species, we must take actions that protect them in both average and extreme years characterized by catastrophic events such as cyclones/hurricanes, forest fires, and disease epidemics (Anderson et al., 2017). This is especially true considering future climate change scenarios, where every year may be uncharacteristically harsh, in effect a 50-year event.
A minimum viable population (MVP) is an estimate of the smallest number of individuals necessary for a population to have a good chance of long-term persistence.
Several studies have attempted to come up with a “universal” MVP value that could ensure that a population of any species has a reasonable chance of persistence. The estimates vary greatly. For example, a universal MVP estimate from the 1980s, the “50/500 rule”, suggested that at least 50 individuals are necessary to prevent inbreeding, and 500 to prevent genetic drift (Frankham et al., 2014). While this 50/500 rule is currently used to guide the IUCN Red List Criteria for small populations (see e.g. Table 8.1, Criteria D), more recent studies suggested that this estimate is much too low. For example, one study that considered over 1,000 species calculated that 1,377 individuals must be protected to ensure the survival of the population and species (Brooke et al., 2006). Another study argued that 4,169 adults needed to be protected (Traill et al., 2007), while a third study identified 7,316 adults as the universal MVP (Reed et al., 2003). The reason why these estimates are highly variable is because MVPs are context specific, with the results varying greatly by species, location, and degree of threat (Flather et al., 2011). For some species, it might be necessary to protect large numbers of individuals—maybe thousands or tens of thousands for invertebrates and annual plants with that can experience large population size fluctuations. For other species, such as those that are long-lived and reproduce regularly, protecting only a few hundred individuals may suffice. Unfortunately, many threatened species have population sizes much smaller than any of these recommended minimums. For example, half of the 23 surveyed elephant populations remaining in West Africa have fewer than 200 individuals (Bouché et al., 2011), a number considered to be vastly inadequate for their long-term survival, especially in the absence of strong conservation management.
While a universal MVP value will probably never be agreed upon, species and location specific MVP estimates have great value for guiding conservation efforts. For example, it can suggest the minimum number of individuals that need to be released to improve chances of reintroduction success (Section 11.2). MVP estimates can also be combined with a species’ home range requirements to determine a minimum dynamic area (MDA), which is the smallest area of suitable habitat required to sustain the MVP. The use of MVP and MDA, and factors influencing it, were well illustrated in several studies on South Africa’s fragmented cheetah (Acinonyx jubatus, VU) population. Here, researchers originally found that translocations every 1–5 years can greatly improve the likelihood of persistence for 20 subpopulations with at least 10 cheetahs each, or for 10 subpopulations with at least 15 cheetahs each (Lindsey et al., 2009). But a follow-up study then showed that these results were context specific, and highly dependent on the presence of other predators that compete for the same prey (Lindsey et al., 2011). With no competitors, a minimum dynamic area of 200 km2 would be sufficient to support 10 cheetahs. However, a reserve of at least 700 km2 would be needed for 15 cheetahs and 15 lions, and even more if other competitor carnivores are present.
Because of the close relationship between population viability and habitat availability, these two factors are often considered together in population and habitat viability assessments (PHVA).
Because of the close relationship between population viability and habitat availability, these two factors are often considered together in population and habitat viability assessments (PHVA, http://www.cpsg.org/our-approach/workshop-processes/phva-workshop-process). Such an assessment was recently performed for Sierra Leone’s western chimpanzees (Pan troglodytes verus, CR), where 53 conservation partners came together to develop a recovery plan for this highly threatened species (Carlsen et al., 2012). As an illustration of how conservation projects can bring people from different walks of life together, the participants for this PHVA came from universities, government, NGOs, and the private sector, and included Paramount Chiefs, representatives from the UN and Sierra Leone’s government ministries, and well as experts in tourism, communications, population modelling, and wildlife rehabilitation.
Effective population size
One of the most important considerations when estimating MVPs is deciding which individuals to include in the calculations. Because population viability depends greatly on a population’s ability to produce young to increase in size (or at least balance out mortality), it makes sense that reproductive status is important in MVP estimations. For that reason, biologists often calculate MVPs using the effective population size (Ne), an estimate of how many individuals or pairs in a population are actively breeding. Consider, for example, a school of 1,000 dolphins; it might have 990 immature individuals and only 10 mature dolphins (five males and five females) that are actively breeding. Even though the full population consists of 1,000 dolphins, the effective population size is only 10—just the mature breeding dolphins.
A population’s effective size is often much smaller than the total population size because not all individuals are capable of breeding at any one time.
It is worth noting that the effective population size may sometimes be even smaller than the number of individuals capable of breeding at any one time. Factors that cause such a scenario include unequal sex ratios, variation in reproductive output, or an inability to find mates. Health status may also play a role; for example, many long-lived seabird species will forego breeding in years where adults did not attain a necessarily healthy body condition, or years when food is scarce (Crawford et al., 2008). Such reduced effective population sizes can lead to drastic population declines, especially when unsuitable conditions persist over consecutive years.
It is also important to remember that the individuals included in effective population size are not the only ones deserving conservation attention. For example, while young animals may not immediately contribute to population growth and stability, they remain a conservation priority for their potential to contribute to population viability in future. Protecting non-reproductive individuals is also important to avoid having cooperative breeders such as African wild dogs (Lycaon pictus, EN) succumb to Allee effects (Section 8.7.2). Individuals that forego reproduction because of poor body condition (e.g. malnourished individuals) can easily become reproductive, and contribute to population viability, if their stressors are mitigated. Lastly, for many species (e.g. many plants, fungi, bacteria, and protists), many (sometimes all) individuals may be dormant for long periods in the soil as seeds, spores, tubers, or other structures. While these dormant individuals may not be part of the effective breeding population, they still contribute to population viability in the long term.
Maximum sustainable yield
A population’s maximum sustainable yield provide an estimate of the greatest number of individuals that can be harvested without detriment to the population.
An important but under-utilised benefit of PVAs is the ability to help conservation managers estimate sustainable harvest rates for wildlife populations at risk of overharvesting (Milner-Gulland and Rowcliffe, 2007). Many threatened species can withstand some level of harvesting, so long as harvest rates are lower than recruitment rates. To estimate the sustainable level of harvesting, biologists may use PVA to estimate a population’s maximum sustainable yield—the greatest number of individuals that can be harvested without detriment to the population (Box 9.3). When estimating maximum sustainable yields for overharvested taxa, it is important to consider not only the total population size (or effective population size), but also harvesting biases produced by harvester preferences and techniques. For example, when estimating hunting quotas, hunter preferring larger animals (Lindsey et al., 2013; Barthold et al., 2016) and animal behavior (Caro et al., 2009) can significantly influence model output. Also, in fisheries management, it is important to consider the outsized role older and larger fish play in recruitment rates, or the indirect damage fishing does the environment or to juvenile individuals (De Leo and Micheli, 2015). Lastly, it is important to consider how harvesting right at maximum sustainable yield levels may leave those populations less buffered to future disturbances (Cumming and Cumming, 2015)—it is thus advisable to maintain harvest quotas well below maximum levels.
Box 9.3 Sustainably Harvesting Fruit Bats Through Better Understanding of Life Histories
David T. S. Hayman
Molecular Epidemiology and Public Health Laboratory,
Hopkirk Research Institute, Massey University,
Palmerston North, New Zealand.
[email protected]
Full-time hunters are employed to shoot them […] Nothing is known about the natural factors that encourage or repress population growth in the straw-coloured fruit bat, or on what age category these factors have maximum effects. […] No laws, customs or taboos protect the straw-coloured fruit bat from exploitation […].
Funmilayo, 1978
Funmilayo’s comments from the 1970s still ring true about the problems currently facing straw-coloured fruit bats (Eidolon helvum, NT) throughout their African distribution range. The species is hunted widely in West and Central Africa (Figure 9.C), with scientists estimating that over 128,000 and up to 306,000 individuals are killed annually in Ghana (Kamins et al., 2011) and Côte D’Ivoire (Niamien et al., 2015). In Ghana, the population “is hunted far beyond maximum sustainable yield”, concluded an initial study that predicted maximum sustainable yield based on population sizes and estimated intrinsic rate of increase (Kamins et al., 2011). This overhunting has been going on for a long time. Notes from a 1909 field trip to DRC reported them hunted in the “hundreds” (Allen et al., 1917).
Despite these reports of intensive harvesting, straw-coloured fruit bats remain abundant with colonies comprised of several million individuals often reported. The species is also highly mobile, migratory, and panmictic, breeding freely across its continental distribution (Peel et al., 2017). These life history traits make it difficult to determine how hunting is impacting the population because presence and size of colonies are highly variable in space and time (Hayman and Peel, 2016).
What can the natural history of the species tell us about their possible vulnerability to the pressures inherent of being hunted? Straw-coloured fruit bats exhibit classic life history traits of long-lived species. These bats invest time and energy into single, well-developed pups that they nurse and carry, as Funmilayo noted, “until they are capable of independent existence, which gives [the pup] a high chance of survival”. This investment in individual young and single annual breeding events means that straw-coloured fruit bats are susceptible to over-harvesting.
Good information on birth and death rates are required to accurately assess the impact of harvesting. Hayman et al. (2012) estimated birth rate and survival probability parameters in a single colony of up to 1 million straw-coloured fruit bats that roost in trees in Accra, Ghana, demonstrating the feasibility of obtaining such information. Histological examination of tooth growth layers allowed age estimation and life-table analyses to estimate an annual survival probability for juveniles of 43% and adults of 83%. Mark-recapture data using radio-collar telemetry and multi-state models to address confounding emigration estimated lower annual adult survival probability, c. 63%. True survival probabilities likely exist between these estimates, as follow up studies from four further locations suggest (Hayman and Peel, 2016), because permanent emigration may underestimate capture–recapture estimates and population decline may bias life table estimates. Birth rates for the species are high (0.96 young per female per year). Improved estimation of these key parameters will allow for critical analyses of harvest sustainability of straw-coloured fruit bat populations in future.
Sensitivity analysis
A particularly useful feature of PVA—and models in general—is that model parameters can be individually evaluated to better understand the implications of different management strategies. This is usually accomplished with a sensitivity analysis, a method that determines which parameter or combination of parameters has the biggest influence on population viability. Obviously, parameters that greatly influence population viability should become the focus of conservation efforts, whereas parameters that have a minimal effect can be given less attention. Some of the most popular model parameters to investigate are demographic parameters by age class, which can identify which life stages are most sensitive to conservation management. Such a sensitivity analysis might reveal that slight changes in adult mortality rates greatly affect population viability, whereas relatively large changes in juvenile recruitment rates have a minimal impact. Crouse et al. (1987) obtained such a result in their classic study on loggerhead turtles (Caretta caretta VU) living off the USA’s Atlantic coast. At the time, great effort was invested in improving hatching success and ensuring that hatchlings to reach the sea. However, Crouse’s study showed that, even if 100% egg and hatchling success was achieved, sea turtles will remain threatened unless adult survival were also improved. Results from this study subsequently played a significant role in initiating global efforts to reduce sea turtle bycatch during fisheries operations (see e.g. Fennessy and Isaksen, 2007; Ayinla et al., 2011). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/09%3A_Applied_Population_Biology/9.02%3A_Estimating_Extinction_Risk.txt |
Lack of adequate data
Population biologists often require several years of survey data to distinguish long-term population trends from “model noise”—short-term population fluctuations caused by weather and other unpredictable events (Figure 9.6). For that reason, general guidelines suggest that at a minimum, six (Morris and Doak, 2002) to 10 (McCarthy et al., 2003) years’ worth of population data are required before a PVA is attempted.
Population biologists often require several years of survey data to distinguish long-term population trends from short-term population fluctuations.
In recent years, considerable effort has been invested in collating, summarising, and making available demographic datasets. One example is the Demographic Species Knowledge Index (Conde et al., 2019) meant to summarise demographic data obtained from ex situ conservation facilities (Section 11.5). Nevertheless, most African species continue to lack multi-year datasets, while many threatened species lack reliable survey data altogether. Because the enormous task of filling these data gaps is impractical, there is a need to be strategic as to which populations to consider for PVA purposes. For example, it does not make sense to conduct a PVA on each species in a threatened ecosystem when a few carefully selected indicator species will suffice to monitor ecosystem health (McGeoch et al., 2002). Other priorities for PVA efforts include (1) species harvested by humans, (2) species most sensitive to ecosystem changes, (3) species with the greatest uncertainty regarding viability, and (4) species that are the focus of current management efforts (Wilson et al. 2015).
But even in the absence of reliable and complete datasets, PVAs can still be useful. For example, sensitivity analysis can inform future data collection efforts, particularly to fill gaps that lead to high levels of uncertainty, or to verify data accuracy for particularly sensitive parameters.
Data reliability
While strategically filling data gaps should be a priority, it should not come at the expense of data quality and reliability. Many—perhaps most—population monitoring programs are poorly designed (Buckland and Johnston, 2017), leading to biased data, poor survey precision, and misleading results. Poorly designed surveys not only waste valuable time and resources, but the erroneous results also seriously hamper conservation efforts.
To overcome these shortcomings, there are five criteria that a well-designed monitoring program should satisfy (Buckland and Johnston, 2017). First, survey sites should represent the region or species of interest. Second, a sufficiently large number of monitoring sites should be chosen. Third, monitoring programs should be set up that every target species—whether common or rare—is adequately counted. Fourth, species selected for monitoring should represent the community of interest, rather than charismatic species that are easily detected. Fifth, multiple surveys need to be conducted over time to detect long-term population trends. Given resource constraints, some compromises in survey design may at times be required. It may also be worth considering the use of citizen scientists and new technologies such as camera traps (Section 9.1.4) to improve data collection efficiency and to provide back-up evidence of reported species for follow-up expert review, if needed.
Model reliability
While PVAs can provide reasonably accurate predictions when based on reliable data (Brook et al., 2000; McCarthy et al., 2003), many conservationists continue to be sceptical of PVA results and their ability to predict future population changes over time (Crone et al., 2013). Part of the reason is our inability to accurately account for unanticipated future events, such as unusual weather events or the arrival of a new invasive species. There are also mechanistic challenges to PVA modelling, including their sensitivity to model assumptions and slight changes in model parameters i.e. slight changes in model input generate vastly different results. For this reason, some biologists have started to discourage the use of PVAs in conservation management, especially when faced with inadequate data (Ellner et al., 2002).
While this scepticism is important and model interrogation should always be welcomed (both aspects usually lead to model improvements), PVA will continue to play a crucial role in conservation in the foreseeable future. It is however important for biologists using PVA to be familiar with the challenges associated with model reliability, as well as the assumptions and limitations of each PVA model. It always helps to begin any PVA model with a clear understanding of the ecology of the target population, the threats it faces, and its demographic characteristics, which in turn enables the modellers to better evaluate model results. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/09%3A_Applied_Population_Biology/9.03%3A_Challenges_to_PVA_Implementation.txt |
1. Read the manuscript by Pfab and Witkowski, (2000), which is a PVA study that is relatively easy to understand. Can you identify some strengths and weaknesses of this PVA? Which assumptions did this study make? What parameters were used? Are there any other model parameters you think could have been useful?
2. For this exercise, you are going to construct a simple PVA for a threatened frog species on a sheet of paper. This frog formerly occupied an expansive lowland forest, which over time was disturbed and degraded. A recent survey was able to find only ten frogs (five males and five females), all in one small, isolated forest patch that can accommodate up to 20 frogs. In the spring, males and females form mating pairs; each pair typically produce zero, one, two, three, or four, offspring that survive to breeding age the following year (to create this demographic dataset, flip four coins for each mated pair; the number of heads is the number of offspring). The sex of the offspring is assigned at random (flip a coin for each young animal, with heads for males and tails for females. Individuals not mated because of uneven sex ratios do not breed. After the breeding season, all the adult frogs die. (A) Run five different population simulations for five generations each, and chart population size over time. What percentage of populations would go extinct within the 10 generations? (B) Perform a sensitivity analysis by making the frogs’ living conditions more severe. For example, lower the number of frogs found during the survey to six, or impose 50% mortality on offspring every year due to introduced rats. (C) Perform another sensitivity analysis by making the frogs’ living conditions more accommodating. For example, examine the impact of supplying extra food to the frogs, which would allow more offspring to be produced each year. Examine the results of all your different models to determine which factor is most important to address to ensure the frog species does not go extinct.
9.05: Suggested Readings
Anderson, S.C., T.A. Branch, A.B. Cooper, et al. 2017. Black-swan events in animal populations. Proceedings of the National Academy of Sciences 114: 3252–57. https://doi.org/10.1073/pnas.1611525114 Ignoring extreme weather events may severely underestimate extinction risk.
Buckland, S.T., and A. Johnston. 2017. Monitoring the biodiversity of regions: Key principles and possible pitfalls. Biological Conservation 214: 23–34. https://doi.org/10.1016/j.biocon.2017.07.034 Five principles for reliable surveys.
Danielsen, F., N.D. Burgess, P.M. Jensen, et al. 2010. Environmental monitoring: The scale and speed of implementation varies according to the degree of peoples’ involvement. Journal of Applied Ecology 47: 1166–68. https://doi.org/10.1111/j.1365-2664.2010.01874.x Involving the local community in environmental monitoring increases the speed of conservation actions.
Guschanski, K., L. Vigilant, A. McNeilage, et al. 2009. Counting elusive animals: Comparing field and genetic census of the entire mountain gorilla population of Bwindi Impenetrable National Park, Uganda. Biological Conservation 142: 290–300. https://doi.org/10.1016/j.biocon.2008.10.024 Genetic techniques are providing new opportunities for studying populations.
Jewell, A. 2013. Effect of monitoring technique on quality of conservation science. Conservation Biology 27: 501–08. https://doi.org/10.1111/cobi.12066 The methods that researchers use to tag and monitor species can affect and sometimes even harm the species being studied.
Mascia, M.B., S. Pailler, M.L. Thieme, et al. 2014. Commonalities and complementarities among approaches to conservation monitoring and evaluation. Biological Conservation 169: 258–67. https://doi.org/10.1016/j.biocon.2013.11.017 Describes different objectives in monitoring, with a goal of developing standard procedures for evaluating projects.
Sebastián-González, E., J.A. Sánchez-Zapata, F. Botella, et al. 2011. Linking cost efficiency evaluation with population viability analysis to prioritize wetland bird conservation actions. Biological Conservation 144: 2354–61. http://doi.org/10.1016/j.biocon.2011.06.015 Different management approaches are evaluated for their cost effectiveness on bird populations in Spain.
One of the following two texts:
Bibby, C., M. Jones, and S. Marsden. 1998. Expedition Field Techniques: Bird Surveys (London: Royal Geographic Society). http://www.conservationleadershipprogram.org/media/2014/09/Bird_Surveying_Manual.pdf Methods for conducting biological surveys. Written on birds, but applicable to other taxa.
White, L., and A. Edwards. 2000. Conservation Research in the African Rain Forests: A Technical Handbook (New York: WCS). http://apes.eva.mpg.de/eng/pdf/documentation/WhiteEdwards2000 Methods for obtaining data on animals and their environment. Written for forest work, but also applicable in other ecosystems. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/09%3A_Applied_Population_Biology/9.04%3A_Topics_for_Discussion.txt |
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Anderson, S.C., T.A. Branch, A.B. Cooper, et al. 2017. Black-swan events in animal populations. Proceedings of the National Academy of Sciences 114: 3252–57. https://doi.org/10.1073/pnas.1611525114
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Bouley P., M. Poulos, R. Branco, et al. 2018. Post-war recovery of the African lion in response to large-scale ecosystem restoration. Biological Conservation 227: 233–42. https://doi.org/10.1016/j.biocon.2018.08.024
Brook B.W., J.J. O’Grady, A.P. Chapman, et al. 2000. Predictive accuracy of population viability analysis in conservation biology. Nature 329: 512–19. https://doi.org/10.1038/35006050
Brook, B., M. Burgman, and R. Frankham. 2000. Differences and congruencies between PVA packages: The importance of sex ratio for predictions of extinction risk. Conservation Ecology 4: 6.
Brook, B.W., L.W. Traill, and C.J.A. Bradshaw. 2006. Minimum viable population sizes and global extinction risk are unrelated. Ecology Letters 9: 375–82. https://doi.org/10.1111/j.1461-0248.2006.00883.x
Buckland, S.T., and A. Johnston. 2017. Monitoring the biodiversity of regions: Key principles and possible pitfalls. Biological Conservation 214: 23–34. https://doi.org/10.1016/j.biocon.2017.07.034
Carlsen, F., K. Leus, K. Traylor-Holzer, et al. 2012. Western chimpanzee population and habitat viability assessment for Sierra Leone: Final report. IUCN/SSC CBSG—Europe (Copenhagen: CBSG Europe). http://www.cpsg.org/sites/cbsg.org/files/documents/Sierra%20Leone%20Chimpanzee%20PHVA%20Final%20Report.pdf
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Crouse, D.T., L.B. Crowder, and H. Caswell. 1987. A stagebased population model for loggerhead sea turtles and implications for conservation. Ecology 68: 1412–23. https://doi.org/10.2307/1939225
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De Leo, G.A., and F. Micheli. 2015. The good, the bad and the ugly of marine reserves for fishery yields. Philosophical Transactions of the Royal Society B 370: 20140276. https://doi.org/10.1098/rstb.2014.0276
Eggert, L.S., R. Buij, M.E. Lee, et al. 2014. Using genetic profiles of African forest elephants to infer population structure, movements, and habitat use in a conservation and development landscape in Gabon. Conservation Biology 28: 107–18. https://doi.org/10.1111/cobi.12161
Ellner, S.P., J. Fieberg, D. Ludwig, et al. 2002. Precision of population viability analysis. Conservation Biology 16: 258–61. https://doi.org/10.1046/j.1523-1739.2002.00553.x
Fennessy, S.T., and B. Isaksen. 2007. Can bycatch reduction devices be implemented successfully on prawn trawlers in the Western Indian Ocean? African Journal of Marine Science 29: 453–63. https://doi.org/10.2989/AJMS.2007.29.3.12.342
Flather, C.H., G.D. Hayward, S.R. Beissinger, et al. 2011. Minimum viable populations: Is there a ‘magic number’ for conservation practitioners? Trends in Ecology and Ecolution 26: 307–16. https://doi.org/10.1016/j.tree.2011.03.001
Formia, A., B.J. Godley, J.F. Dontaine, et al. 2006. Mitochondrial DNA diversity and phylogeography in West and Central African green turtles (Chelonia mydas). Conservation Genetics 7: 353–69. https://doi.org/10.1007/s10592-005-9047-z
Frankham, R., C.J.A. Bradshaw, and B.W. Brook. 2014. Genetics in conservation management: Revised recommendations for the 50/500 rules, Red List criteria and population viability analyses. Biological Conservation 170: 56–63. https://doi.org/10.1016/j.biocon.2013.12.036
Funmilayo, O. 1978. Fruit bats for meat: are too many taken? Oryx 14: 377–78. https://doi.org/10.1017/S0030605300016008
Hayman, D.T.S., and A.J. Peel. 2016. Can survival analyses detect hunting pressure in a highly connected species? Lessons from straw-coloured fruit bats. Biological Conservation 200: 131–39. https://doi.org/10.1016/j.biocon.2016.06.003
Hayman, D.T.S., R. McCrea, O. Restif, et al. 2012. Demography of straw-colored fruit bats in Ghana. Journal of Mammalogy 93: 1393–404. https://doi.org/10.1644/11-MAMM-A-270.1
Hughes, N., N. Rosen, N. Gretsky, et al. 2011. Will the Nigeria-Cameroon chimpanzee go extinct? Models derived from intake rates of ape sanctuaries. In: Primates of Gashaka, ed. by V. Sommer and C. Ross (New York: Springer). http://doi.org/10.1007/978-1-4419-7403-7
Ingram, D.J., L. Coad, B. Collen, et al. 2015. Indicators for wild animal offtake: Methods and case study for African mammals and birds. Ecology and Society 20: 40. http://doi.org/10.5751/ES-07823-200340
Kamins, A.O., O. Restif, Y. Ntiamoa-Baidu, et al. 2011. Uncovering the fruit bat bushmeat commodity chain and the true extent of fruit bat hunting in Ghana, West Africa. Biological Conservation 144: 3000–08. https://doi.org/10.1016/j.biocon.2011.09.003
Kays, R., M.C. Crofoot, W. Jetz, et al. 2015. Terrestrial animal tracking as an eye on life and planet. Science 348: aaa2478. http://doi.org/10.1126/science.aaa2478
Kümpel, N.F., E.J. Milner-Gulland, G. Cowlishaw, et al. 2010. Assessing sustainability at multiple scales in a rotational bushmeat hunting system. Conservation Biology 24: 861–71. https://doi.org/10.1111/j.1523-1739.2010.01505.x
Lindsey, P., C.J. Tambling, R. Brummer, et al. 2011. Minimum prey and area requirements of the vulnerable cheetah Acinonyx jubatus: Implications for reintroduction and management of the species in South Africa. Oryx 45: 587–99. https://doi.org/10.1017/S003060531000150X
Lindsey, P., K. Marnewick, H. Davies-Mostert, et al. 2009. Cheetah (Acinonyx jubatus) population habitat viability assessment workshop report (Johannesburg: IUCN CBSG and EWT). http://www.cbsg.org/sites/cbsg.org/files/documents/South%20African%20Cheetah%20PHVA%202009.pdf
Lindsey, P.A., G.A. Balme, P. Funston, et al. 2013. The trophy hunting of African lions: Scale, current management practices and factors undermining sustainability. PLoS ONE 8: e73808. https://doi.org/10.1371/journal.pone.0073808
Marnewick, K., S.M. Ferreira, S. Grange, et al. 2014. Evaluating the status of and African wild dogs Lycaon pictus and cheetahs Acinonyx jubatus through tourist-based photographic surveys in the Kruger National Park. PloS ONE 9: e86265. https://doi.org/10.1371/journal.pone.0086265
Maxwell, S.M., G.A. Breed, B.A. Nickel, et al. 2011. Using satellite tracking to optimize protection of long-lived marine species: Olive ridley sea turtle conservation in Central Africa. PloS ONE 6: e19905. https://doi.org/10.1371/journal.pone.0019905
McCarthy, M.A. 2007. Bayesian Methods for Ecology (Cambridge: Cambridge University Press). https://doi.org/10.1017/CBO9780511802454
McCarthy, M.A., S.J. Andelman, and H.P. Possingham. 2003. Reliability of relative predictions in population viability analysis. Conservation Biology 17: 982–89. https://doi.org/10.1046/j.1523-1739.2003.01570.x
McGeoch, M.A., B.J. van Rensburg, and A. Botes. 2002. The verification and application of bioindicators: A case study of dung beetles in a savanna ecosystem. Journal of Applied Ecology 39: 661–72. https://doi.org/10.1046/j.1365-2664.2002.00743.x
Metcalfe, K., P.D. Agamboué, E. Augowet, et al. 2015. Going the extra mile: Ground-based monitoring of olive ridley turtles reveals Gabon hosts the largest rookery in the Atlantic. Biological Conservation 190: 14–22. https://doi.org/10.1016/j.biocon.2015.05.008
Milner-Gulland, E.J., and M.J. Rowcliffe. 2007. Conservation and Sustainable Use: A Handbook of Techniques (Oxford: Oxford University Press).
Morris, W.F., and D.F. Doak. 2002. Quantitative Conservation Biology (Sunderland: Sinauer).
Niamien, M., J. Coffi, B. Kadjo, et al. 2015. Initial data on poaching of Eidolon helvum (Kerr, 1792) near-threatened species in Côte D’ivoire, West Africa. European Journal of Scientific Research 35: 219–27.
Ogutu, J.O., H.-P. Piepho, M.Y. Said, et al. 2016. Extreme wildlife declines and concurrent increase in livestock numbers in Kenya: What are the causes? PLoS ONE 11: e0163249. https://doi.org/10.1371/journal.pone.0163249
Parker, L. 2017. New ocean reserve, largest in Africa, protects whales and turtles. National Geographic. http://on.natgeo.com/2samx3a
Peel, A.J., J.L.N. Wood, K.S. Baker, et al. 2017. How does Africa’s most hunted bat vary across the continent? Population traits of the straw-coloured fruit bat (Eidolon helvum) and its interactions with humans. Acta Chiropterologica 19: 77–92. https://doi.org/10.3161/15081109ACC2017.19.1.006
Péron, G., and R. Altwegg. 2015. Twenty-five years of change in southern African passerine diversity: Nonclimatic factors of change. Global Change Biology 21: 3347–55. https://doi.org/10.1111/gcb.12909
Pfab, M.F., and E.T.F. Witkowski. 2000. A simple population viability analysis of the critically endangered Euphorbia clivicola R.A. Dyer under four management scenarios. Biological Conservation 96: 263–70. http://doi.org/10.1016/S0006-3207(00)00088-4
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9.07: Summary
1. Protecting a threatened species requires a firm grasp of its population biology. Long-term monitoring using biodiversity inventories, population censuses, and demographic studies can reveal temporal changes in population size and distribution and help to distinguish short-term fluctuations from long-term decline.
2. Biologists are increasingly relying on innovative methods to track wildlife populations and demographics. Among the most popular are market surveys, hair snares and faecal sampling, while photos taken by tourists and camera traps have also been used to obtain population-level data.
3. Population viability analysis (PVA) uses demographic, genetic, and environmental data to predict changes in population sizes and extinction risk over time. Sensitivity analysis can be used to guide conservation action by estimating how different management actions will affect a population’s extinction probability.
4. Minimum viable population estimates can be used to determine how many individuals are needed to reduce the threat of extinction, while maximum sustainable harvest estimates can be used to set harvest limits on species threatened by overharvesting.
5. Many surveys are poorly designed, leading to biased data, poor survey precision, and misleading results, which hamper our ability to halt biodiversity losses. To overcome these challenges, surveys should be representative, sufficiently large, and conducted repeatedly over time. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/09%3A_Applied_Population_Biology/9.06%3A_Bibliography.txt |
Habitat loss (and its associated degradation) is currently the most important threat facing Africa’s wildlife (Figure 10.1). When an ecosystem is destroyed or degraded, its ability to sustain wildlife is compromised, and the individuals that depend on that ecosystem for survival either need to adapt or move elsewhere, or they will die. Conversely, preventing ecosystem degradation and destruction is one of the single most important actions we can take to protect biodiversity. In the process, we also improve our own well-being, given that natural ecosystems are our first line of defence against natural disasters, and provide us with food, clean water, and other ecosystem services.
Broadly speaking, ecosystem conservation involves three different activities: (1) monitoring ecosystems, (2) maintaining ecosystems, and (3) restoring damaged ecosystems. While many books have been written on each of these three activities, the broad overview this chapter provides will hopefully enable readers to gain a basic understanding of the tools and methods used in ecosystem conservation.
10.01: Ecosystem Monitoring
A complex and adaptive ecosystem in which all the chemical, physical, and biological components, functions, and processes are intact and functioning normally is considered a healthy ecosystem (but see Cumming and Cumming (2015), for a discussion on this value-based term). In contrast, disturbing any of an ecosystem’s components, functions, and/or processes will, by definition, alter them to some degree (Table 10.1). In many cases, ecosystems that have been exposed to certain forms and levels of disturbances remain healthy because there is redundancy in the roles performed by different ecosystem components (Section 4.2.1). This ability of an ecosystem to withstand certain forms and levels of disturbances is referred to as ecosystem stability. Ecosystem stability could be the result of one or both of two qualities: resistance and resilience. Resistance is the ability of an ecosystem to retain the same characteristic communities and natural cycles throughout and after a disturbance event, while resilience is the ability of an ecosystem to rapidly recover or adapt after a disturbance event. For example, if the number of native aquatic insect species decline after non-native fishes are introduced to previously fish-free ponds, the pond’s ecosystem has low resistance. But if the native insect community recovers rapidly after the non-native fishes were removed, the ecosystem is resilient.
Table 10.1 Three ways how humans have changed the natural world.
Natural function
Changes attributed to human activities
Land surface
As much as half of the world’s ice-free land surface has been transformed to cater to people’s need for natural resources. Much of these changes are driven by agricultural activities.
Nitrogen cycle
Human activities release massive amounts of nitrogen into natural ecosystems on a daily basis. Much of this occurs through the use of nitrogen fertilisers, burning fossil fuels, and cultivating nitrogen-fixing crops.
Atmospheric carbon cycle
Scientists estimate that humans would have doubled levels of carbon dioxide in the Earth’s atmosphere by the middle of this century. This is primarily the result of fossil fuel use and deforestation.
Wildlife populations
Between 1970 and 2014, Sub-Saharan Africa have lost three quarters of its freshwater vertebrates; the rate of these declines shows no sign of reducing.
Pollutants
Pollution from human activities have become so omnipresent that it is hard to escape its impacts. Microplastics have been found in drinking water and the food we eat (Chapter 7).
Sources: MEA, 2005; Kulkarni et al., 2008; http://www.livingplanetindex.org
Many forms of ecosystem disturbances are easy to observe. Consequently, monitoring these visible forms of disturbances—such as the outright destruction of a forest or plastic pollution on a beach—focuses less on detection and more on developing systematic survey protocols (Section 9.1) that can provide information on whether a disturbance is spreading and increasing in intensity, or whether conservation action is successful in containing the threat. However, some disturbances are subtler, unobtrusive, and thus difficult to detect; examples include pesticide drift and agricultural runoff (Section 7.1). Adopting a “wait and see” approach to detecting these invisible forms of disturbances can be particularly damaging, since that approach generally ends at a point where the harm will either be impossible to reverse or will require significantly more resources and time than would have been the case if the problem was addressed earlier. In this way, there are many similarities between monitoring ecosystem health and human health—some ailments are easier to diagnose than others, but we avoid the worst-case scenarios by screening regularly for diseases and treating the threatening ones promptly.
Perhaps the most popular method conservation biologists use to monitor ecosystem health is known as biomonitoring. By monitoring the abundance and/or fitness of sensitive species (Box 10.1), biologists can sometimes detect ecosystem degradation before it becomes apparent to the human eye or escalates to a point where it starts impacting human lives (Bornman and Bouwman, 2012). Monitoring environmental indicators such as macroinvertebrates (Figure 10.2) is particularly popular when examining the ecological condition of aquatic ecosystems; mayflies, caddisflies, and stoneflies—specialists of undisturbed streams—are often replaced by flies and midges in polluted and disturbed environments. Sometimes however, when plants or animals are not easily monitored, certain aspects of those species can still be monitored. One example is monitoring total plant biomass as a proxy for soil nutrients or intensity of herbivory. Another option is to perform a bioassay, during which a sensitive organism (typically water fleas or plankton) is released into a potentially contaminated environment to see if death or declining health occurs.
By monitoring the abundance and/or fitness of sensitive species, biologists can detect threats to biodiversity before it becomes apparent to the human eye.
Box 10.1 Using Insects to Monitor Environmental Health
Rosina Kyerematen
Department of Animal Biology and Conservation Science,
University of Ghana,
Legon, Ghana.
[email protected]
Insects are important to nearly every terrestrial food web in the world and serve a multitude of different purposes: some insects are responsible for pollination of plants while others are scavengers that clean up dead plant and animal material. In some cases, our understanding of an insect species’ ecological role can make it suitable as an indicator of environmental health. Biomonitoring looks at the presence and abundance of organisms within their natural communities to assess the impact of environmental disturbances; this knowledge can then be used to guide ecosystem management. An indicator taxon is one whose impact can be specifically and precisely measured; its abundance serves as a measure of the overall health of an ecosystem. Understanding how the presence and abundance of indicator species, and the relative abundance of tolerant and intolerant species, reflects the relative health of an environment can allow for rapid surveys of impaired ecosystems to assess trends as well as to track changes following remediation and restoration efforts.
Over the past few years, biomonitoring with insects as indicator taxa has become increasingly popular in Ghana. Butterflies are especially popular because they show varying relative sensitivities to environmental change; the abundance of certain butterfly species can for example be used to study the impact of habitat loss, fragmentation, and climate change (Kyerematen et al., 2018). The presence and abundance of butterflies more characteristic of open and disturbed ecosystems (Figure 10.A) can, for example, be used as an indicator of forest degradation.
Aquatic insects, particularly benthic macroinvertebrates, are also useful bioindicators. Freshwater resources, such as lakes and rivers, provide water for drinking and washing to local people, and a home for economically important taxa, such as fish and shellfish. Protecting these water sources is therefore important for safeguarding people’s health and livelihoods. The presence, absence, and diversity of certain benthic macroinvertebrates, even at the order level, can provide valuable information about whether a waterbody is being degraded or not (Kyerematen et al., 2014; Nnoli et al., 2019). A recent study showed that dragonfly and damselfly diversities and populations along the coastal Densu River in Ghana vary widely depending on the physical condition of the river and surrounding area (Acquah-Lamptey et al., 2013).
With their high diversity and varying tolerances for ecosystem conditions, insects are extraordinarily suited as ecological indicators in environmental monitoring. Each insect species is also part of a wider biological community with important ecological roles. If lost, not only will an abundance of other life be affected, such a loss may also hint at a looming crisis facing people living in those compromised ecosystems.
At times, conservation biologists may need to measure the physical environment to assess environmental health. This approach is particularly common when tracking pollution, for example by monitoring for changes in biochemical indicators. For example, measuring total phosphorus, nitrogen, and dissolved oxygen load in streams and other surface water can help scientists track eutrophication (Section 7.1.1). Measuring these and other biochemical indicators is usually accomplished directly via chemical analysis of environmental samples, such as soil and water. Sometimes however, biochemicals indicators are tracked indirectly via biological samples obtained from plants and animals. Because they bio-accumulate heavy metals and other pollutants, filter feeders such as clams and mussels (e.g. Bodin et al., 2013) are particularly useful in this regard as they can be used to detect very low concentrations of harmful chemicals in the environment.
Monitoring ecosystems with geospatial analysis
A persistent challenge facing biologists who monitor ecosystems—and other aspects of biodiversity—is achieving consistency across space and time. Consider a survey of a sensitive bird community to track ecosystem change; not only will different observers have varying levels of experience, but they will almost certainly see different birds during an early morning census compared to one later in the afternoon due to differences in biology and behavior between species. These factors introduce error into monitoring data, which in turn can mask the effect a biologist tries to measure (Buckland and Johnston, 2017). Laboratory scientists’ control for these confounding factors by making multiple measurements under strictly controlled conditions. But for conservation biologists working outside in the wind and rain, repeated observations under similar conditions can be near impossible.
Geospatial analysis offers a variety of tools that allow biologists to overcome some of these traditional field monitoring challenges. These tools use geographic information systems (GIS) computer software packages to store, display, and manipulate a wide variety of data representing the natural environment, biodiversity, and human land-use patterns as they relate to one another on Earth’s surface. GIS thereby allows biologists to easily visualise and analyse spatial relationships between mapped data, which may include aspects such as vegetation types, climate, soils, topography, geology, water availability, species distributions, existing protected areas, human settlements, and human resource use (Figure 10.3). Understanding such relationships helps conservation biologists to prioritise their actions, for example by identifying areas where data are lacking, where an environmental change requires further investigation, or where gaps in regional protected areas network exist.
Remote sensing is a special branch of geospatial analysis directed at obtaining ecosystem data without making physical contact (i.e. boots on the ground) with the observation site. Before the turn of the 20th century, the most popular form of remote sensing was aerial photography from airplanes. These aerial photographs facilitated geographers’ ability to draw maps of landscape features, including human infrastructure and natural vegetation patterns. Remote sensing opportunities greatly expanded from 1960 onward, with the launch of the National Aeronautical Space Administration’s (NASA) first Earth observation satellites, to take photographs of Earth from space for weather forecasting. Subsequent satellite programs expanded their scope to also collect additional data of Earth’s surface and atmosphere. While much of this data would have been useful to conservation, early satellite data products were very expensive and thus largely out of reach of the larger conservation community. This all changed in 2008 when NASA started distributing their Earth observation products for free to the public, heralding an era in which remote sensing became a standard tool in the conservation field.
Remote sensing offers a variety of tools that allow biologists to monitor biodiversity beyond the abilities of traditional field monitoring techniques.
Today, hundreds of Earth observation satellites circle the planet, offering near real-time access to unbiased and consistent environmental datasets of nearly all terrestrial surfaces, oceanic surfaces and floor depths, and the atmosphere, all from the comfort of a computer connected to the internet (Wilson et al., 2013). Scientists use these products in ecosystem monitoring efforts, including monitoring water quality (Dube et al. 2015), forest loss (Laporte et al., 2007), coral reef health (McClanahan et al., 2011), desertification (Symeonakis et al., 2004), and fire regimes (Archbald et al., 2010). Linking the information obtained by Earth observation satellites to biological information collected on the ground has proved invaluable in monitoring species’ threat statuses (Di Marco et al., 2014), ecosystem connectivity (Wegmann et al., 2014), and habitat suitability (Torres et al., 2010), as well as understanding how biodiversity responds to environmental changes (Box 10.2).
Box 10.2 Remote Sensing and Spatial Analysis for African Conservation
Barend F. N. Erasmus
Global Change Institute (GCI), University of the Witwatersrand,
Johannesburg, South Africa.
[email protected]
Remote sensing is the art and science of observing objects or landscapes from a distance, without being in direct contact with the environment. Although you can think of wildlife photography as a type of remote sensing, the term usually refers to aerial photography (Figure 10.B), or images taken via satellite. For both cases, there is a trade-off between the area covered by each photo (the footprint), how much digital storage space is available, and how often a satellite takes a picture of the same area. For satellites that take pictures of the earth systematically, along a pre-defined path, cloud cover determines how often you can get a usable image. Recent improvements in technology, now allow for the deployment of constellations of satellites that have a collective point-and-track capability to observe an area almost continuosly.
Our eyes are sensitive to the colours red, green, and blue, together called the visible part of the electromagnetic spectrum. This is actually a very small part of the entire spectrum, and scientists have found, for example, that vegetation seen as uniformly green shows a lot of variation in the infrared part of the spectrum. For this reason, many satellites carry cameras that can “see” infrared light, and by proxy, measure vegetation health, biomass, and sometimes even structure. This capability, together with frequent revisit rates, allows for a unique view of how African landscapes change, whether through habitat loss, seasonal changes, or drought. Analyses of 13 years of remotely sensed data show that the most arid parts of southwest Botswana now experience typical summer vegetation conditions later in the year (Dubovyk et al., 2015), so herbivores must cope with a much longer dry season now than they experienced in 2000.
African landscapes are always changing, and sometimes in unpredictable ways—localised thunderstorms at the beginning of the rainy season can rapidly change a dry dustbowl landscape into green grazing. These “wet footprints” can cover areas as small as 1 km × 3 km; in contrast, a large frontal weather system can cover dozens or even hundreds of kilometres. An ungulate looking for green grazing not too far from drinking water, while trying to avoid predators and hunters, needs the ability to detect such green patches, and the strength and knowledge to move there. Obtaining this knowledge from an ungulate’s perspective of the landscape is no small task. Hopcraft et al. (2014) show how common wildebeest and plains zebra in the Serengeti have different migration strategies: wildebeest move to green grass as quickly as they can, with little effort to avoid predators, whereas zebra gauge predation risk and forage quality concurrent to their trek.
Our ability to understand ungulate movements and other ecosystem patterns has been greatly enhanced over the last few years, as animal-tracking technology (where GPS positions are logged and stored using radio-frequency tracking, mobile phone networks, satellite systems, or any combination) and the resulting analyses became more sophisticated. It is now possible to distinguish between locations where animals often spend a little time (for example, a preferred shady tree for resting during the heat of the day), or the same location where they infrequently spend a lot of time (for example, once during an oestrus cycle, a lactating lioness will spend a lot of time at her den, with cubs). It is also possible to “see” when such an animal changes “mode” of movement. For example, researchers have found that springbok (Antidorcas marsupialis, LC) rams around the Etosha pan in Namibia show sedentary behavior (small movements around a specific area) when grazing is good, which transitions to searching behavior during dry seasons (long, relatively fast movements in a straight line) as they move to areas that had water or grass during previous drought periods (Lyons et al., 2013).
If we combine this animal tracking capability with regular remotely sensed images of vegetation quality, then we can start to answer questions about why certain animals move to certain areas. It also highlights the fact that it is very seldom good enough to put a fence around an area and call it preserved. Due to the changing nature of Africa’s savannahs, animals need to be able to move to areas with water or forage, often outside reserves, when they become available. In the face of a changing climate and changing rainfall patterns, this ability to move long distance to reach vital resources remains one of the best adaptation responses that African ungulates may have to cope with climate change. However, fences, roads, and farmland may block these migrations across the landscape, putting those populations at risk of extinction (Section 5.1.1).
African conservation areas need to make provision for these animal movements, or risk conservation areas without sustainable animal populations. This presents a problem—how do we investigate options for large animals to move across a transformed rural landscape and minimise human-wildlife conflict while, on the other hand, still providing access to green grass or drinking water at a specific protected area? Both resources change in location and time of year, and only regular, detailed remote sensing of vegetation, combined with detailed animal movement studies, will provide the necessary picture in time and space.
The popularity and utility of these products have preceded and facilitated the expansion of other remote sensing applications in ecosystem monitoring. Among the most popular are radar products (Figure 10.4), which have become the standard method for obtaining elevation and other terrain data (NASA, 2009, 2013), as well as estimates of carbon stocks (Carreiras et al., 2012). Advances have also been made in using hyper-spectral imagery to monitor soil properties (Mashimbye et al., 2012) and even individual trees (Naidoo et al., 2012). LiDAR has enabled biologists to map three-dimensional vegetation, which can be used to explain animal movements (Loarie et al., 2013; Davies et al., 2016) and measure carbon stocks and forest loss (Burton et al., 2017).
Conservation biologists are often faced with the shifting baseline syndrome, where the reference points they use to measure their progress may be vastly different from earlier states.
Despite the opportunities presented by remote sensing, it is critical to remember that it is not a substitute for traditional field monitoring methods. Most importantly, remote sensing applications cannot be considered reliable without verification using field data (see Burton et al., 2017). Most remote sensing products are also relatively new, which does not allow for enough opportunities to compare across time. In the absence of historical remotely sensed data, geospatial analysts usually choose the best reference site currently available; this may expose those analysts to shifting baseline syndrome, because the chosen reference site may be vastly different from earlier states the scientists are actually interested in studying (Bunce et al., 2008; Papworth et al., 2009). Remote sensing is, therefore, not a cure-all for ecosystem monitoring challenges; it is simply a powerful tool to supplement traditional field-based monitoring. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/10%3A_Conserving_Ecosystems/10.00%3A_Prelude_to_Conserving_Ecosystems.txt |
Even when monitoring data show that an ecosystem is healthy, management is often required to maintain those desired conditions. That is because very few ecosystems are completely free of human influences. For example, rivers and streams carry pollutants far beyond the point of contamination (Section 7.1) and roads acting as firebreaks suppress natural fire regimes. Today, even the most isolated patches of habitats may not be completely protected from the influences of global processes such as climate change. To maintain complex and adaptive ecosystems, conservation biology is guided by four complementary management principles: (1) maintain ecosystem processes, (2) minimise external threats, (3) be adaptive, and (4) be minimally intrusive.
Maintaining critical ecosystem processes
Ecologists generally divide ecosystem processes into four disparate yet interdependent categories: water cycling, nutrient cycling (which include the carbon and nitrogen cycle), energy flow, and community dynamics. The linkages between these processes create feedback loops, where changes in one factor may be amplified elsewhere. Maintaining ecosystem processes is thus very important because small, seemingly small, changes can have major impacts on biological communities.
The water cycle
The water cycle refers to the distribution of water through an ecosystem, and includes the absorption and distribution of water vapour, rainwater, and surface water in lakes, rivers, and oceans. Since much of the water cycle happens out of sight and is generally associated with large-scale phenomena, such as weather patterns and anthropogenic climate change, land managers sometimes fail to recognize how local factors influence the water cycle. This is a grave mistake; many deadly ecological disasters (e.g. desertification, flooding, and landslides) can be attributed to disturbances to the water cycle at the local scale.
Outside of ensuring sustainable use of water resources, maintaining vegetation cover arguably plays the most important role in preserving the water cycle at local scales. Plants and their roots enable soil to store and release water, and make these water reserves available for soil organisms, which in turn aid in decomposition of dead plants and animals (Section 4.2.2). In contrast, a loss of vegetation cover increases surface runoff, which leads to deteriorating soil conditions through nutrient leaching and erosion of fertile topsoil. An increasing number of studies have also shown how forest loss can change a region’s climate by reducing rainfall which in turn exacerbates drought conditions (Lawrence and Vandecar, 2015). For example, forest clearing for agriculture has reduced rainfall by 50% over much of West Africa (Garcia-Carreras and Parker, 2011). Many forest restoration programs thus focus on reversing these losses.
Complex and adaptive ecosystems provide more opportunities for people to benefit from ecosystem services than uniform plantations with single species
When restoring degraded forests and other ecosystems to repair the water cycle and other ecosystem services, it is important to remember that complex ecosystems with locally-adapted plants are generally the most effective in maintaining the water cycle and other ecosystem services (Burton et al., 2017). There have been cases across Africa where well-intended restoration efforts used fast-growing timber species such as gum (Eucalyptus spp.) and pine (Pinus spp.). While these single-crop plantations may superficially resemble a forest in structure and may even provide some of the same ecosystem services as native plants, some of these fast-growing exotic plants also bring significant environmental harm and negative externalities passed onto local people (van Wilgen and Richardson, 2014). Of particular concern is their role in disrupting local water cycles (Section 7.4.2), ironically the very aspect these forest restoration efforts aim to rehabilitate. The choice of species used for restoration should this be carefully considered to avoid unintended consequences later on.
The nutrient cycle
The nutrient cycle involves the cycling of essential nutrients such as carbon, nitrogen, sulphur, and phosphorus through the ecosystem. Like the water cycle, natural vegetation cover plays an important role in maintaining the nutrient cycle. That is because plant roots slow water runoff which, in turn, help soil to retain nutrients dissolved in water. Plants also form a major component of above-ground and below-ground biomass. When dead plant biomass is decomposed along with animal waste products, nutrients previously absorbed through plant roots are released back into the soil and water, where they can once again be absorbed by living plants and other consumers.
Unfortunately, vegetation cover, decomposition, and fire dynamics (discussed below) alone cannot ensure a healthy nutrient cycle. Much of Africa is nutrient impoverished because soil nutrients are lost quicker than they are replaced. One of the main causes is unsustainable agricultural practices (Sanchez, 2010), such as farming on sandy soils and in tropical forests. These areas are nutrient-poor, so crop yields are typically low. Because these areas are prone to leaching, a large proportion of synthetic fertilisers added to supplement impoverished soils leaches into groundwater or washes into nearby streams and lakes, threatening water supplies by causing harmful algae blooms and eutrophication (Section 7.1.1). To compensate, the failing farmers may resort to even more unsustainable land conversions (Wallenfang et al., 2015). Careful management of the nutrient cycle is thus critical for both biodiversity conservation and socio-economic well-being, particularly given its importance to food security (Drechsel et al., 2001). To achieve this, there is an urgent need to adopt more sustainable land management practices (Chapter 14).
Ecosystem processes are linked into multiple feedback loops, so changes in one factor are amplified elsewhere.
The energy cycle
Energy flow—a crucial component of ecosystem productivity (Section 4.2.2)—refers to the capture and storage of solar energy by primary producers (photosynthetic plants, algae, and some bacteria), and the distribution of that energy to consumers, detritivores, and decomposers. Although solar energy can appear as an unlimited resource in many ecosystems, the energy available to consumers (i.e. herbivores and carnivores) is limited because only about 10% of the energy obtained at one trophic level is passed on to the next (Figure 10.5). Being at the top of the food chain, apex predators are in a particularly vulnerable position because seemingly small disruptions at lower trophic levels will have a cumulative impact on the energy available to them. Such disruptions may include reduced prey populations (e.g. overharvesting of herbivores, Section 7.2) or foraging disruptions (e.g. a predator needing to walk further to find prey). Research from Southern Africa’s Kalahari Desert has showed that such disruptions, which are amplified by predators’ high-energy lifestyles (it takes a lot of energy to bring down a large ungulate!), may put apex predators such as cheetahs on a downward spiral of energy deficits (Scantlebury et al., 2014). While such impacts may not always lead to direct mortality, these insidious, subtle, and easily-overlooked sublethal impacts compromise individuals’ ability to reproduce, with extinction being the end result. To avoid such a scenario, maintaining energy flow generally involves maintaining complex, species-rich ecosystems so that consumers have ample opportunities to fulfil their energy needs for finding prey, growth, reproduction, and other activities.
Community dynamics
In ecosystem conservation, maintaining viable populations of different interacting species is as important as maintaining important ecosystem processes, such as ecosystem productivity and ecological succession. This focus often falls on maintaining populations that form part of important mutualistic relationships such as pollination and seed dispersal (Section 4.2.5), predator-prey interactions, and even healthy levels of competitive and parasitic interactions (which allow more species to persist). Of interest is the preservation of keystone species and ecosystem engineers, which has an outsized effect on community dynamics (Section 4.2.1). As illustrated in this, and other, chapters, disrupting community dynamics through pollution, overharvesting, or any other threat facing biodiversity (Chapter 5–7), generally leads to impoverished natural communities. Impoverished communities may in turn provide opportunities for invasive species to colonise an area, further perpetuating biodiversity losses. In Chapter 11, we discuss further how populations and species can be maintained.
Fire Dynamics
Although fire is generally not considered one of the four fundamental ecosystem processes, it plays such an important role in African biodiversity management, including maintaining the four fundamental ecosystem processes, that it deserves its own discussion. African farmers understand the importance of setting fire to keep cropland and grazing pastures productive; burning existing vegetation releases carbon and other essential nutrients beneficial for plant growth into the environment. Similarly, fire also plays a critical role in the flow of energy, community dynamics, and overall maintenance of fire-dependent ecosystems, such as grasslands, savannahs, and Mediterranean communities. Suitably low intensity fires seldom kill living plants; rather, they encourage seed germination and seedling growth by reducing dead material that may crowd new growth, by exposing bare mineral soil (the substrate required for many seeds to germinate), and by releasing vital nutrients into the soil. This periodic removal of dead material also prevents fuel load accumulation, thereby preventing future fires from becoming destructive. In contrast, without fire, fire-dependent ecosystems will slowly transform into unproductive scrublands suffocated by encroaching woody vegetation (Smit and Prins, 2015). Then, when wildfires do occur (e.g. through human negligence or lightning) the resultant accumulated fuel loads increase the intensity and heat of fires, creating very dangerous and difficult to control scenarios.
Obviously, given the potentially destructive force of fire, land managers who use fire as a management tool must consider many aspects before setting a prescribed burn, also known as a controlled burn. Foremost, to prevent a fire from becoming destructive to natural communities and nearby human developments, burning must be done in a well-planned manner with careful consideration given to the area’s ecology, weather forecasts, and fire-readiness of the site (Goldammer and de Ronde, 2004; Kelly and Brotons, 2017). It is also recommended that prior to burning (or any other conspicuous management operation for that matter), land managers develop a public outreach plan to explain to local people the importance of fire in ecosystems management, and the steps taken to keep them and their properties safe. To further improve community relations and education, South Africa’s Working on Fire program (Figure 10.6) provides scholarships, fire training, and employment opportunities to local youths.
Fire management plans that match natural fire regimes produce the best results for effective ecosystem management. Land managers accomplish this by ensuring that their burn plans mimic the local area’s natural fire season, fire frequency, and flame intensity, while also accounting for management goals and local ecological factors such as rainfall and geology (see e.g. van Wilgen et al., 2010, 2014). The size of each burn area must also be considered. Best practices suggest not burning the entirety of a community at a time; rather, burning only portions of an area allows for more habitat heterogeneity, provides opportunities for non-burrowing animals to take refuge in unburned areas, and maximises ecosystem diversity. Bringing all of these aspects together, scientists working in Tanzania’s Ngorongoro Conservation Area determined that the area would respond best if land managers burn up to 20% of their grasslands annually or biannually (Estes et al., 2006). Similarly, field experiments in certain South African grasslands have found that plant diversity is highest when burns occur every second year, in winter or autumn (Uys et al., 2004). Other ecosystems, such as the Cape Floristic Region’s fynbos, may need to burn only once every decade (Kraaij et al., 2013). However, because of the high density of houses in some fynbos ecosystems, the periodic fires needed for locally-adapted vegetation to persist are often extinguished because of the threat to human settlement (van Wilgen et al., 2012).
While fire plays an important role in many African landscapes, it is important to note that overly frequent fires can be a threat even to fire-dependent communities. For example, habitat degradation resulting from too many fires in quick succession can leave a natural community vulnerable to invasions by harmful species (Masocha et al., 2011). Overly frequent fires can also prevent seedling recruitment by directly killing vulnerable young plants, and by depleting the seed bank because seedlings do not have sufficient time to mature and set seed.
Fire-sensitive ecosystems (e.g. tropical forests, high mountains, and peat bogs) must also be managed carefully to avoid fire disturbance, which can lead to habitat loss and edge effects (Chapter 5). One way to accomplish this is to educate farmers living adjacent to fire-sensitive ecosystems on how to safely manage their land with fire. Conservationists also need to be considerate when managing fire-dependent ecosystems adjacent to fire-sensitive ecosystems, as is the case with the unique patches of fire-dependent savannahs—remnants from the last Ice Age 15,000 years ago—that are surrounded by forest within Gabon’s Lobé National Park (Jeffery et al., 2014). Careful fire management, led by good science, is bound to become increasingly important in the future, given that wildfires are expected to become more frequent and more intense under climate change (Pricope and Binford, 2012).
Minimising external threats
Human activity cannot and does not need to be eliminated from nature; in fact, the structure and diversity of many of today’s natural landscapes—and to which today’s wildlife are adapted to—are in part the result of past human activities (e.g. Garcin et al., 2018). Today, there are over 7 billion people on Earth, so our impacts are more pervasive than for the majority of history. There is, thus, an urgent need to utilise natural resources in such a way that future generations will also benefit from the ecosystem services that previous generations have left us. This requires a concerted effort from every sector in human society to minimise those threats we impose on the ecosystems around us. This includes preventing pollution, large-scale human disturbances, overharvesting, and habitat destruction (Chapter 5–7).
Major strides have been made in recent years towards achieving these goals. Governments are updating laws to safeguard the environment, industries are refining recycling and waste disposal methods, new techniques are being developed to remove pollutants from the environment, and individual citizens are becoming more aware of their individually small but collectively significant impacts on the environment. We should be proud of the progress being made and continue to strive for improvements. But one external threat that requires greater attention and understanding is invasive species.
Controlling invasive species
Invasive species degrade and destroy natural ecosystems by outcompeting native species, disturbing ecosystem processes, and altering the physical environment (Section 7.4). Limiting these harmful impacts can be particularly challenging since exotic species that establish themselves in a new area can build up such large numbers, become so widely dispersed, and be so thoroughly integrated into ecosystems (i.e. naturalised) that eradicating them entirely would be extraordinarily difficult and expensive, or as in the case of tickberry (Lantana camara) perhaps even impossible (Bhagwat et al., 2012). This is not only a problem facing conservation biology, but also agriculture, where invasive species often spread from one farm to another, forestry, where invasive species are spread between saw mills and along logging routes, and fisheries, where native resources are outcompeted, sometimes up-ending an entire local industry. The impact of invasive species on farming communities is particularly severe—they lose tens of billions of dollars each year while trying to combat deteriorating grazing lands, reduced crop yields, and escalating pest control expenses. One study from South Africa calculated that invasive plants result in financial losses of US \$646 million each year—this figure would have been US \$5 billion if invasive species control measures already implemented were absent (de Lange and van Wilgen, 2010). Consequently, a range of stakeholders have invested considerable resources in combatting invasive species.
The most important step in preventing biological invasions is to prevent the initial establishment of problem species. This requires educating people about the dangers posed by invasive species.
Because invasive species are often very hard to eradicate once established (Figure 10.7), the foremost step in avoiding invasive species’ harmful impacts is to avoid opportunities for new invasions (Section 7.4.1). This requires raising awareness across all levels of society about the dangers posed by invasive species, both to the natural world and to agricultural and natural resources systems. There is also a need for citizens, scientists, and industry to monitor for potential and known invasive species, and promptly implement intensive control efforts to stop establishment and spread. The Global Register of Introduced and Invasive Species (http://www.griis.org) is a free, online searchable source to facilitate these tasks, by providing information about the impact and control of invasive species. Governments can also partake in efforts to control invasive species. While most African countries screen agricultural imports for pests, countries, such as Australia and New Zealand, take this task particularly seriously, with trained officials screening each visitor (and returning residents) and package for hitchhiking species before they cross those countries’ borders. Lastly, it would require increased dialogue between conservation biologists and land managers to make a careful and thorough assessment prior to the deliberate introduction of a new species, even if thought of as beneficial.
Despite best practices, not all invasions can be prevented, and for those that do occur, an early detection and rapid response strategy offers the best chance to limit harm. This usually involves raising awareness of potential invasive species to ensure biologists and other stakeholders will recognize a new invasion, efforts to screen for such species on a regular basis, and implementing direct attack approaches, such as using herbicides, pesticides, or mechanical control once detected. While addressing a new invasion as soon as possible, it is also important to consider and contain the risks each direct attack approach carries. For example, herbicides and pesticides carry a risk of killing non-target native species via pesticide drift (Section 7.1), while mechanical control may cause disturbances such as trampling, undue soil disturbance, and even pollution.
Controlling invasive species that have become established will require substantially more resources and manpower to combat than those detected early on. Even so, it is still worth initiating control measures as early as possible since harm and resource needs will only escalate over time. To overcome these impediments to invasive species control, the South African government established an exemplary model, the Working for Water program, in 1995. Working for Water has three goals—poverty relief, water conservation, and invasive species control—which it accomplishes by combining invasive species control with job creation and social upliftment (van Wilgen and Wannenburgh, 2016). Specifically, the program hires and trains unemployed people to eradicate water-thirsty invasive shrubs and trees across South Africa; some of the removed plants are subsequently sold as firewood at local markets at a profit for the participants. In its first 20 years, the program has created over 227,000 person years of employment and treated over 28,000 km2 of invasive species (Figure 10.8). Hopefully programs, such as these, will inspire more governments to act to restore degraded ecosystems to their previously healthy state.
Another method to manage established pests is biological control, also called biocontrol. Biocontrol typically relies on one or more natural enemies from an invasive species’ original range to control the pest in its introduced range (Section 4.2.7). One of the main benefits of biocontrol is that it ensures cost-effective, long-term, area-wide control of an invasive species, beyond the capabilities of chemical pesticides and mechanical control. Biocontrol also allows for opportunities to control invasive species that are hard to manage with chemical pesticides and mechanical control (at least without significant additional harm to the environment), such as submerged aquatic weeds (Coetzee et al., 2011). Third, biocontrol agents are highly host-specific, thereby eliminating the impact that chemical pesticides and mechanical control have on non-target organisms. Lastly, an effective biocontrol agent ideally eliminates the need for chemical pesticides and mechanical control, thereby reducing threats such as pesticide pollution (Section 7.1) and ecosystem degradation (Section 5.1).
Biocontrol does have some drawbacks. Primary among them is the significant upfront investment required, as candidate species first need to be found, and then extensively tested for host specificity and potential interactions with native wildlife before being released. Biocontrol also requires careful monitoring after release to determine effectiveness as well as to carefully check for impacts on non-target native species. This monitoring needs to be conducted over the long term, because biocontrol agents typically require several years before they establish self-sufficient colonies in the wild and might only then show signs of unintended impacts. Because alternative methods for controlling invasive species can also kill biocontrol agents (causing conflicting results and wasted resources), additional coordination is required before applying biocontrol and alternative pest management strategies simultaneously in the same area. Lastly, there is no guarantee that a biocontrol agent will be effective. For example, the tickberry continues to thrive despite the release of over 40 biocontrol agents (Zalucki et al., 2007). But when successful, the long-term savings from these upfront investments are generally well worth it. One study in Benin found that biocontrol of water hyacinth required a US \$2 million upfront investment, but the resultant water quality improvements increased local incomes by US \$84 million per year (de Groote et al., 2003). Another study from South Africa estimated a net gain of 50–3,500 times the investment, depending on the specific biocontrol agent used (de Lange and van Wilgen, 2010).
Several very successful biological control programs have been implemented in Africa over the past century. Most famous is the rescue of the cassava crop (see Box 4.3). Another successful biocontrol program, implemented across much of the continent, has reduced Kariba weed (Salvinia molesta) by over 95% within just a few years (e.g. Mbati and Neuenschwander, 2005; Diop and Hill, 2009; Martin et al, 2018). South Africa has been particularly active in the research and introduction of biocontrol agents. From 1913, when South Africa started controlling invasive cacti with hemipterans, to 2017, a total of 93 biocontrol agents have been released for the control of 59 invasive species (Zachariades et al., 2017).
While the most popular biocontrol agents generally involve insects, disease-causing pathogens can also be used for biocontrol. Feline panleukopenia virus (also known as feline distemper) was highly effective in managing a feral cat population that caused the extirpation of seabirds on Sub-Antarctic Marion Island; some birds are now even returning as breeders (Bester et al., 2002). In an effort to reduce the use of chemical pesticides on food crops, efforts are also currently underway to find fungal pathogens to control introduced pests impacting African crops, including pea leafminers (Liriomyza huidobrensis), originally from South America (Akutse et al., 2013), and banana weevils (Cosmopolites sordidus), originally from Southeast Asia (Akello et al., 2008).
The most effective pest control programs use an integrated pest management (IPM) approach that relies on using multiple control methods either simultaneously or in succession.
Some of the most effective pest control programs use an integrated pest management (IPM) approach that relies on using multiple pest control methods described above either simultaneously or in succession (van Wyk and van Wilgen, 2002). Strategic planning to coordinate best practices can also help offset some of the costs of invasive species control (Rahlao et al., 2010) and ensure that important pest sources are not missed (van Wilgen et al., 2007). When considering the best method to control an invasive species, it may also help to consider how our own actions inadvertently encourage invasive species. For example, an over-reliance on synthetic fertiliser has been shown to cause eutrophication (Chislock et al., 2013) and encourage growth of aquatic invasive plants (Coetzee and Hill, 2012; Bownes et al., 2012).
Even though the impacts of invasive and other exotic species are generally considered negative, they do occasionally provide some benefits. For example, Australian pines (Casuarina equisetifolia) have been planted widely throughout Africa for timber, charcoal, and to stabilise eroding lands. Some people harvest invasive species to eat or to sell; examples include water hyacinth (Figure 10.9), prickly pears (Opuntia spp.), and Mediterranean mussels (Mytilus galloprovincialis). The latter has also become an important food source for the African black oystercatcher (Haematopus moquini, NT) (Kohler et al., 2009). Similarly, nearly 18 species of diurnal raptors, including four globally threatened species, nest and roost—sometimes in colonies numbering thousands of individuals—in stands of invasive Australian gum (Eucalyptus spp.) trees (Allan et al., 1997; Jenkins, 2005). Many birds also favour fruits produced by invasive plants, such as the tickberry and syringa (Melia azedarach); this however allows those plants to spread even further. Because they reproduce so fast, water hyacinth has been investigated as a resource for producing bioenergy (Güereña et al., 2015). A ground spider (Prodida stella, CR), endemic to the Seychelles and threatened by sea level rise (Gerlach, 2014), was first described from specimens collected on Australian pine, and may thus persist on trees further inland as their original range is submerged by rising sea levels. Nevertheless, assessments generally show that costs incurred from invasive species outweigh the benefits (Mwangi and Swallow, 2008). It is thus important to consider whether native species can fulfil the same functions in those cases where benefits of invasive species are touted.
Adaptive management
In decades past, ecosystem management in Africa was generally conducted following a laissez-faire (i.e. hands-off) approach where natural processes were allowed to follow their own course, with interventions only implemented when subjectively deemed absolutely necessary. While this passive management style may have worked in an era when ecosystems were less fragmented by roads and fences, and the impact of pollution and invasive species transported along rivers and streams were limited. But maintaining healthy ecosystems (especially small ones) through limited action is becoming increasingly difficult in today’s human-dominated world. Leaving concerns unattended may seem fine in the short term, but such neglect can create problems that are very difficult to contain later on. For this reason, it is increasingly necessary to actively manage ecosystems not only to achieve conservation goals, but even to just avoid degradation of current conditions.
A major challenge of active (and reactive) conservation management is that nature consists of thousands of interacting components and feedback loops. Because it is impossible to fully understand how these different components interact, management strategies are typically implemented with an incomplete understanding of how interventions may impact broader ecosystem processes.
Management actions that draw from experimentation and prior experiences often deliver the desired results. However, despite good intentions and careful consideration, some interventions may later give rise to a cascade of unintended consequences that run counter to overall conservation goals. South Africa’s Kruger National Park provides a good example. Conservation managers here once thought they could better protect the local wildlife by fencing the Park and constructing artificial waterholes at regular intervals, a policy that was implemented primarily in the 1960s and 1970s (Smith et al., 2007; Venter et al., 2008; van Wilgen and Biggs, 2011). But instead of providing conservation benefits, the fences blocked dispersal routes, which caused grazing herbivores to be increasingly sedentary around waterholes, leading to overgrazing. The increased availability of surface water also allowed elephant populations to increase to a point where they became a threat to other taxa, particularly large fruit-bearing trees such as baobabs (Adansonia digitata). The construction of artificial waterholes in arid sections of the Park saw an expansion of plains zebra (Equus quagga, NT) and common wildebeest (Connochaetes taurinus, LC) distributions, which in turn also attracted more lions into these arid areas. This increasing grazing competition and predation pressure caused populations of four locally rare antelopes to fall by 73–88% between 1986 and 2006, prompting fears of imminent extirpations of these iconic species in one of Africa’s premier protected areas.
Despite good intentions, conservation activities may give rise to unintended consequences that run counter to overall conservation goals. Adaptive management can reduce further harm.
While it is hard to escape the looming threat of unintended consequences harming biodiversity, their impacts can be mitigated. One of the first steps involves setting management goals and objectives through close working relationships with a wide variety of stakeholders, including local people (Box 10.3) and scientists who can provide additional insights into the local social-ecological context within which land managers operate. Also important is developing a monitoring protocol that enables land managers to assess whether conservation goals are being met. When monitoring shows that management actions are ineffective or detrimental, it is absolutely critical to be willing and able to integrate this improved understanding into revised management strategies. This process, whereby new knowledge gained through repeated cycles of learning is used to revise and refine conservation strategies and management goals, is known as adaptive management (Venter et al., 2008; van Wilgen and Biggs, 2011). Rather than punishing management errors, adaptive management embraces the idea that we live in a complex world, and that management strategies frequently need be revised to account for new knowledge, shifting priorities, and even evolving societal values. Some of the best adaptive management plans explicitly mandate regular (e.g. every 5 years) reviews, where management strategies, goals, and objectives are formally reviewed and updated.
Box 10.3 Environmental Governance in the Serengeti Ecosystem
Alex Wilbard Kisingo
College of African Wildlife Management,
Mweka, Tanzania.
[email protected]
The greater Serengeti ecosystem of Tanzania (Figure 10.C) represents one of the most biologically productive ecoregions in the world. In recognition of its biological importance, much of this ecosystem is safeguarded under a mosaic of government, private, and co-managed protected areas, as well as community conserved areas (Kisingo, 2013a). Among these protected areas is the world-famous Serengeti National Park, as well as the Ngorongoro Conservation Area.
The various governance models that dictated human actions in the colonial, post-colonial, and contemporary eras have had a profound effect on the ecosystem’s conservation outcomes (Polasky et al., 2008; Sinclair et al., 2008). During most of the colonial period and immediately after Tanzania gained independence in 1961, management of the ecosystem was in the hands of government, which followed a fortress conservation approach (i.e. “fences and fines”). This included evicting traditional peoples from their ancestral lands, restricting the use of wildlife and other natural resources, and imposing heavy penalties on those on the wrong side of wildlife laws. This approach severed the link between people and nature, creating an atmosphere of hostility and resistance to conservation initiatives amongst the local communities. The results have been increased subsistence and commercial poaching, human encroachment into wildlife habitats, and a general lack of community support for conservation (Kisingo 2013b). As a result, Tanzania has seen its wildlife populations reduced and migratory corridors blocked, while invasive plant species are spreading and human-conservation conflicts are intensifying.
In contrast to this counter-productive approach, involving local people in decision making and implementation can enhance attainment of conservation and social outcomes. Such an involvement can occur through training and employment in conservation enterprises, developing social infrastructure through conservation-related financing, and improving a democratic governance space through increased awareness and capacity building (Kisingo 2013a). This socio-economic transformation, where local communities experience the benefits of conservation activities first-hand, can even create a positive feedback loop where local people not only reduce their dependence on protected wildlife resources for survival, but also become inspired to initiate their own grassroots conservation initiatives.
In 1989, nearly 30 years after Tanzania gained independence, some attempts were made to empower local people in the governance of the Serengeti ecosystem. The first major step in this regard involved developing the Serengeti Regional Conservation Strategy (SRCP), which encouraged the creation of integrated conservation and development projects (ICDPs, see Section 14.3) as a means for local people to gain direct benefit from conservation (Kisingo, 2013a). This was followed by the establishment of the Ngorongoro Pastoral Council in 2000, to include pastoralist communities living in the Ngorongoro Conservation Area (NCA) in decision making and protected areas management. Then, in 2003, the Ikona and Makao Community Wildlife Management Areas (WMAs) were established as a first step towards the co-management of wildlife in the Serengeti ecosystem.
Unfortunately, despite good intentions, these initiatives only had a limited impact in empowering local communities. An important reason for this failure was a general incompatibility between local policies and national legislations, notably a legal framework that favoured a top-down approach to decision making over the wishes and values of local communities. In the NCA, communities were more involved in the provisioning of scholarships and socio-economic assistance, in effect surrendering decision making power to government authorities. Local communities thus either never had a real seat at the table during the development and implementation of management strategies, or lost what little they had over time (Kisingo, 2013a). It should thus not come as a surprise that management of the Serengeti ecosystem generally presents mixed results in terms of conservation and social outcomes. Although wildlife populations are still better off in government protected areas when compared to WMAs, neglecting these community conserved areas (which function as critical wildlife corridors) undoubtedly also negatively impacts wildlife in government protected areas.
To be more effective at meeting conservation goals, there is an urgent need to adapt and reengineer ecosystem governance structures in the greater Serengeti ecosystem. Involving more stakeholders, particularly local people, would be an important first step. This involvement should be honest and transparent. Instead of calling one or two community representatives to a workshop to secure a rubber-stamp, start by gathering input from a wide variety of stakeholders and make an effort afterwards to show how that input was considered in final plans. In this way, even those who do not get what they want will at least know that they were heard, and hopefully understand why their desires were not met. There is also an urgent need for capacity building, to develop conservation management expertise among members of local communities, rather than the current over-reliance on outside experts who do not understand local dynamics. This could start small: community members can, for example, be involved in anti-poaching efforts through the use of Village Game Scouts (VGS), a strategy that reduces operational costs in tandem with boosting local incomes. The most promising and eager scouts can then be provided with additional professional development opportunities. Empowering local communities in this way, and enabling them to benefit from conservation activities, will provide opportunities for a wider variety of people, beyond just foreign tourists, to understand how conservation actions can support and benefit livelihoods. This will result in social outcomes that are desired by local communities, and in return, conservation authorities who then gain local support for future actions.
Being minimally intrusive
While adaptive management often necessitates active management, it does not embrace a philosophy of intervention for its own sake. If managers try to maintain their land by intervening in every aspect in nature, conservation will become inhibitively expensive, especially as the costs of managing unintended consequences escalate. Rather, effective adaptive management plans generally embrace a philosophy of “management by exception” (Venter et al. 2008), whereby interventions are implemented only when certain pre-determined thresholds of concern are exceeded, or when monitored trends suggest that those thresholds will soon be exceeded. (For examples of such thresholds, see van Wilgen and Biggs, 2011).
Thinking back to the example from Kruger National Park, discussed above, when it became clear the four species of antelopes were on their way to extirpation, a research program was initiated to identify the causes for these population declines. Based on this research, park managers decided that the best way forward was to close the majority of the artificial waterpoints, and move some antelopes to a predator-proof enclosure where they can safely breed (van Wilgen and Biggs, 2011). To reverse the damage caused fences, park managers also prioritised re-establishing free movement of animals by removing these fences between the park and some adjacent properties (Venter et al., 2008). While lion and zebra populations subsequently moved out of the Park’s arid region as hoped, and elephant populations stabilised, numbers of the four antelope species remained stubbornly low. This prompted park managers to evaluate whether the resources invested in these locally rare but globally common ungulates could be better spent on globally rare species occurring in the park, such as black rhinoceros (Diceros bicornis, CR). In other words, they adapted their threshold of concern from focussing on locally rare species to prioritising globally rare species (van Wilgen and Biggs, 2011). This decision remains controversial, even among park managers concerned about the potential loss of genetic diversity within the four ungulates. But adaptive management allows for shifting priorities; park managers may very well revisit their decisions if and when the park’s rhinoceros are secured, and funds are freed up for other activities. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/10%3A_Conserving_Ecosystems/10.02%3A_Maintaining_Complex_and_Adaptive_Ecosystems.txt |
Ecosystems are regularly disturbed by natural phenomena such as floods from cyclones/hurricanes, or wildfires started by lightning. Nevertheless, natural disturbances typically lead to succession and a return to ecological conditions that can sustain high levels of biodiversity. In contrast, ecosystems that humans have damaged or destroyed through activities such as unsustainable agriculture and deforestation, overgrazing, or pollution tend to lose their ability to rebound without human intervention.
Ecological restoration aims to restore damaged ecosystems to a point where their ecosystem functions and species composition resemble their original or near-original state.
Ecological restoration is the practice of restoring damaged ecosystems to a point where their ecosystem functions and species composition resemble their original or near-original state. Restoration ecology, in turn, is the scientific study of restoring damaged ecosystems, communities, and populations. Ecological restoration is often the best method for providing for the long-term use of degraded sites, whether considered from the perspectives of ecology, social benefit, or even economic benefit. Consequently, and unsurprisingly, this practice and the science originally developed in response to attempts to restore economically valuable ecosystem functions: creating wetlands to prevent flooding, reclaiming mining sites to prevent pollution and soil erosion, revegetating overgrazed rangelands to increase grass production, and planting trees on cleared areas to improve agroforestry.
Best practices in ecological restoration have undergone major advances in recent decades. In the past, restoration methods mostly aimed for quick economic benefit, which resulted in simplified ecosystems that either failed to establish or degraded after a short time. To avoid such costly mistakes, restoration plans of today increasingly aim for the permanent re-establishment of healthy ecosystems that could support sustainable industries such as ecotourism, wildlife management, carbon sequestration, and low-level grazing by livestock. Ecological restoration often also makes economic sense; a study from South Africa found that every US \$1 invested in restoring ecosystem services would generate US \$8.30 for the local economy (de Wit et al., 2008).
Many grassroots conservation groups are at the forefront of initiatives that use ecosystem restoration to help make the connection between healthy ecosystems and socio-economic well-being. One prime example is the Green Belt Movement, a Kenyan initiative led by rural women to combat deforestation and restore degraded forests. They do this by helping rural women work together to grow and plant trees. Since its founding in 1977, the organization has overseen the planting of over 51 million trees, which has helped restore forests on Mount Kenya, the Aberdares, and the Mau Complex. The planted trees have prevented erosion, stored rainwater, and provided firewood, timber, and food. In addition, over 30,000 women have been trained in sustainable trades such as forestry, beekeeping, and food processing.
Ecological restoration approaches
There are four main approaches to ecological restoration (Figure 10.10):
• Natural regeneration. Degraded areas, such as abandoned fields or logged areas, are allowed to naturally reseed and return to grasslands or forests. Land managers often choose this approach when active restoration is too expensive, when earlier restoration attempts have failed, or when experience has shown that the ecosystem is resilient and can recover on its own (e.g. Crouzeilles et al., 2017).
• Rehabilitation. Land managers improve conditions of a degraded ecosystem by transitioning it to another, different ecosystem type. For example, land managers could rehabilitate a degraded forest by transitioning it to a tree plantation. Rehabilitation could involve replacing just a few species or many species.
• Partial restoration. Land managers restore some ecosystem functions and some of the species that were dominant or characteristic of the ecosystem. For example, as a part of a grassland restoration, land managers might initially replant a few key species that are hardy and contribute to ecosystem functioning; they could delay restoration of rare species until later phases.
• Complete restoration. Land managers restore an area to benchmark ecosystem structure, mix of species, and ecosystem functioning. Complete restoration usually requires an active program to modify the site, reintroduce native species, and eliminate or reduce the factors that were degrading the ecosystem.
Before a restoration project is initiated, and the type of approach is decided upon, land managers must consider how quickly the ecosystem can recover, resource needs and availability, the availability of locally adapted taxa, and the work that might be required to allow the restored community to persist over the long term. Examples of specific considerations include how to prepare soils, how to handle translocated organisms, when and how much fertiliser and water to add, and how to prevent invasions by unwanted species (Galatowitsch and Richardson, 2005; Zabbey and Tanee, 2016). It is also important to remember that ecosystems generally fail to recover if the factor that caused them to become degraded in the first place is not removed or reduced. For instance, efforts to reverse desertification (Section 5.3.4) would require a reduction of grazing pressure and unsustainable agricultural practices.
To measure restoration success, biologists often aim to restore degraded areas to conditions (ecosystem functions or species composition) comparable to a chosen benchmark or reference site. Reference sites provide practical targets for restoration and can be used to quantitatively assess of the success of a restoration project. Comparing restoration progress against a reference site also allows land managers to intervene or adjust their methods if restoration goals are not being met. This approach, in which land managers monitor conditions and adjust their protocols as and when needed, is known as adaptive restoration. (For a general discussion on adaptive management, see Section 10.2.3.)
Major restoration targets
Many human-altered ecosystems in Africa have proven to be good candidates for ecological restoration. These include tropical rainforests, wetlands, rangelands, and coral reefs. In addition, restoration projects in urban areas (Box 10.4) have become popular in recent years in part due to the enhanced quality of life for people living in the area.
Box 10.4 Sustainable Forest Restoration Using Natural Vegetation
Samuel Kiboi
School of Biological Sciences, University of Nairobi,
Nairobi, Kenya.
[email protected]
Deforestation is one of the main driving forces of biodiversity loss in Africa. Many rural and urban communities rely on wood biomass for energy in the form of either charcoal or firewood. This means that they must continuously source for the firewood or charcoal by harvesting living or dead trees. In many cases, the available energy source is live trees on farmlands which are planted as border trees, or random remnants of pre-existing vegetation within the farm. In some instances, farmers who have land in less densely populated areas have portions of forested areas or woodlots which are under continuous disturbance from wood harvesting. This is more common in rangelands or areas that have lower agricultural productivity. In other areas, such as urban settlements bordering forests, such as Kibera in Nairobi Kenya, there has been extensive harvesting of firewood and sometimes selectively for medicinal purposes or wood carving (Furukawa et. al., 2011).
Given the known benefits of intact forests, including improving food security and climate change mitigation, there are currently several efforts aimed at increasing Kenya’s forest cover both in protected and unprotected areas. The Kenya Forest Service has always been at the forefront of restoration in protected areas, particularly in gazetted forest areas. Despite the general enthusiasm to increase forest cover, many structural and informational challenges remain. Most reforestation programs classify seedlings as either “exotic” or “indigenous”, but do not consider which species are best suited to local conditions. In addition, despite the numerous reforestation programs initiated by individuals, government entities, and corporations, there is generally minimal follow-up maintenance after planting, which can jeopardise an entire project. The first three years after planting are especially crucial for proper seedling establishment and require intensive management, including weeding, mulching, and protection from herbivores. Perhaps the biggest challenge to the sustainability of these reforestation initiatives is the slow growth rate of many valuable indigenous trees that does not meet short-term harvest demands while also allowing for longer-term forest regeneration.
At the University of Nairobi, successful urban forest islands with potential natural vegetation have been established using the “Miyawaki method” (Miyawaki, 2004). This method uses native trees to restore indigenous forests at timelines shorter than if natural regeneration was allowed to take its course (Figure 10.D). To create an urban green space on the university property, we selected 16 native tree species using a vegetation science study of remnant forests around Nairobi. Within 16 months, many of the trees had established well, with the best performing species, Ehretia cymosa, growing to over 2 m (Kiboi et. al., 2014). This study illustrates the importance of selecting locally adapted species in forest restoration initiatives.
Sustainable restoration practices can alleviate the short-term pressure from restored ecosystems while they mature to a self-sustaining structure. Not only should locally adapted species be promoted, but also native species that can be continuously coppiced, where new shoots rapidly replace harvested branches and portions of branches. Exotic species, such as Australian gum, pine, and mesquite (Prosopis juliflora) often display these characteristics, but those species are often invasive with detrimental effects on native ecosystems and communities. Fortunately, many African plant species are also good candidates for sustainable restoration initiatives, including camphor bush (Tarchonanthus camphoratus), sickle-leaved false-thorn (Albizia harveyi), silver clusterleaf (Terminalia sericea), and weeping wattle (Peltophorum africanum) (Kennedy, 1998; Kaschula et al., 2005). Although coppices may have more variable increases in biomass compared to initial planted stands, it is a sustainable way of biomass management especially in areas that experience high demand for harvestable wood. In addition, planting native trees and shrubs in farmlands typically provides beneficial ecosystem services through increasing the abundance and diversity of native insectivorous birds and pollinators of crops. In these various ways, the right management practices can lead to benefits for local people, biodiversity, and sustainable conservation practices.
Tropical forests: Tropical forests cover less than 10% of Earth’s land surface; yet, they contain more than half of all terrestrial species (Cortlett and Primack, 2011). When these forests are lost, we lose substantial biodiversity and ecosystem services. For this reason, tropical forest restoration initiatives in Africa and elsewhere have received much attention in recent years. Towards the end of this chapter we will discuss a major global effort focussed on restoring degraded tropical forests, known as REDD+.
Africa has already lost over 40% of its wetlands through human activity, with current loss rates among the highest in the world.
Wetlands: Africa has already lost over 40% of its wetlands through human activity, with current loss rates among the highest in the world (Davidson, 2014). Because of the recognized importance of wetlands in providing flood control and other ecosystem services (Section 5.5.3), damaged wetlands are frequently targeted in restoration efforts. Wetlands are defined by their hydrology; therefore, wetland restoration projects often the focus on restoring a site’s original hydrology. One such example comes from South Africa, where authorities (with support from the World Bank) have been working on restoring Africa’s largest estuarine lake at iSimangaliso Wetland Park—a multistep process that involves restoration of the estuary’s hydrology, controlling invasive plants around the wetland, and improving farming practices in the surrounding area (Whitfield et al., 2013). Wetland restoration can also occur through activities like dam removals (Section 11.3.2) or replacing exotic vegetation that deplete groundwater with native vegetation to promote groundwater retention (Sirami et al., 2013). Importantly, true wetland restorations are notoriously difficult to accomplish. It can be relatively easy to replant a wetland to look as it previously looked, but to restore the foundational hydrology often requires sophisticated engineering. In many cases, partial wetland rehabilitation is the best that can be achieved.
Mangrove swamps (Figure 10.11) provide nursery grounds for many economically important fisheries, protect coastal communities against powerful storms, and prevent saltwater from intruding into freshwater systems (van Bochove et al., 2014). They are also among the world’s most important carbon sinks, storing four times more carbon per hectare than other types of tropical forests (Donato et al., 2011). Yet, over 35% of the world’s mangrove swamps have already been degraded by agriculture, urban expansion, pollution, and commercial shellfish farming (MEA, 2005; Giri et al., 2011). To regain these lost services, several communities are now restoring their mangroves, while also adopting more sustainable practices to reduce damage to these important habitats (Feka et al., 2009). One of Africa’s most ambitious mangrove restoration projects have been initiated in Senegal, where more than 300,000 local citizens planted more than 150 million mangrove trees across 140 km2 between 2006 and 2013 (Cormier-Salem and Panfili, 2016). Mangrove (as any other) restoration projects do need to be planned carefully to ensure success. For example, it is important to choose ecologically-appropriate species to plant, rather than the fastest growing species that promises quick (but not necessarily optimal) results. Studies from Eritrea have also shown how fertiliser runoff caused by wave action could reduce lead to project failure (Sato et al., 2005). Another concern is that mangroves are often exploited, restored, and managed as forests, while the primary determinants of their function and structure—hydrology, soils, and nutrients—are neglected (Lewis, 2005; Gopal, 2013). Recent work showed that natural regeneration of mangrove swamps may produce more diverse, resilient, and productive ecosystems compared to planting efforts (Wetlands International, 2016). These issues will need to be addressed to ensure the long-term sustainability of mangrove restoration efforts.
Seasonal drylands: Through extensive land mismanagement (primarily overgrazing and unsustainable agriculture), a large portion of Africa’s seasonal drylands are undergoing desertification, the conversion of once-productive land to desolate man-made deserts—large dry unproductive dust bowls with no vegetation. The degradation of these lands has crippled agriculture, obliterated natural biological communities, and displaced millions of people. While many drylands seem to regenerate naturally when pressures associated with land mismanagement are removed at an early stage, extended periods of mismanagement hamper recovery by leading to a loss of natural seed banks, nutrients, and microsites that allow for seedling establishment.
Somalia is home to one of the world’s most effective desertification reversal programs. Since the early 1990s, when Somalia’s national government collapsed, Somalis have been tormented by warlords and civil war. The lack of effective governance also saw the rise of an unregulated charcoal trade; groves of thorn trees hundreds of years old were set ablaze, before the so-called “black gold” was exported to Arabia. The resultant wildfires and removal of trees caused an erosion crisis, turning grazing lands that once supported a diverse pastoralist community into unproductive wastelands. The resulting famine, exacerbated by droughts, caused even more Somalis to turn to a life of crime, piracy, and terrorism in a desperate effort to support their families. To reverse this decline, the humanitarian NGO Adeso successfully persuaded a regional government to create and enforce a ban on charcoal exports. Adeso also started educating local people about the links between the environment and their own lives, and introduced sustainable alternatives to the charcoal trade, such as promoting the use of solar cookers to reduce the need for charcoal fuel. To reverse desertification and prevent further erosion, Adeso showed local communities how to construct small and simple rock dams; the dams also provide a microenvironment suitable for thorn tree seeds to germinate. Adeso has been so successful in these ventures that they subsequently expanded their work to Kenya and South Sudan.
Coral Reefs: Coral reefs are one of the world’s most important marine ecosystems, both ecologically and economically. They provide food to local communities, support ecotourism industries, and protect coasts by reducing wave energy by as much as 97% (Ferrario et al., 2014). Yet, coral reefs are also one of the most threatened marine ecosystems, impacted heavily by overharvesting, pollution, sedimentation, and climate change. Nevertheless, restoring coral reefs is well worth it; a meta-analysis found that it is nearly 20 times cheaper to restore coral reefs than to construct artificial systems for coastal protection (Ferrario et al., 2014). As such, several initiatives are now in progress to restore coral reefs, ranging from transplanting corals and boosting sea urchin populations for seaweed control to creating artificial reefs that can act as substrate for coral settlements (Lindahl, 2003; Edwards and Gomez, 2007).
The future of ecological restoration
Research in restoration ecology has grown rapidly in recent years. Many reviews (e.g. Suding, 2011) and books (e.g. Falk et al., 2016) have recently been published on the topic. The Society for Ecological Restoration (SER) was established in 1988 to support the field, and two scientific journals (Restoration Ecology and Ecological Restoration) publish hundreds of papers each year on the topic, in addition to the papers published in other ecological and conservation journals. The growth in research provides scientists and land managers more studies and evidence to inform planning and improvement of restoration projects.
A recent development in the field involves biodiversity offsets (ten Kate et al., 2004; MacFarlane et al., 2016). A system generally used by developers, biodiversity offsets aim to achieve no net loss of biodiversity during economic development; some projects even aim for a net overall biodiversity gain. Developers accomplish this by compensating for the ecosystem damage (or loss of threatened species populations, Kormos et al., [2014]) that may be incurred during a development project. This compensation usually follows one or more of three main strategies: (1) reducing the extent of damage at the development site, (2) restoring or protecting natural communities at a different “receptor site” as compensation for what is being lost, and (3) enhancement of the remaining natural communities after development.
While biodiversity offsets (and other restoration initiatives in general) sound good in theory, it is important to remember that the most effective biodiversity conservation strategy remains protecting and managing intact ecosystems. Studies and practical experience have shown that ecological restoration efforts often fail to recreate key characteristics of their reference sites, including species composition or ecosystem functioning, even after years of effort and investment. It is also important to remember than some African ecosystems regenerate very slowly—tropical forests require more than 100 years to develop (Bonnell et al., 2011)—so even effective restorations may take decades to provide the full range of benefits. In cases where biodiversity offsets are pursued, it is critical to ensure that these initiatives indeed offer true conservation gains by mitigating the various associated risks (Coralie et al., 2015; Gordon et al., 2015; Maron et al., 2016).
Because restoring damaged environments takes considerable time and resources, preserving intact ecosystems should be prioritised. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/10%3A_Conserving_Ecosystems/10.03%3A_Restoring_Damaged_Ecosystems.txt |
Complex natural ecosystems play an important role in mitigating the destructive effects of climate change. Prominently, living plants sequester greenhouse gases from the atmosphere (Zarin et al., 2015); in contrast, their loss due to habitat loss increases greenhouse gas emissions (Section 6.1). Studies have also shown how ecosystems with high complexity (Betts et al. 2018) and species diversity (Mokany et al., 2014; Isbell et al., 2015) are better buffered against climate change. Lastly, by maintaining and restoring carbon-sequestering ecosystems, we also provide opportunities for climate-sensitive to persist despite the threat of climate change pace (Section 11.4).
One of the foremost initiatives aimed at combatting climate change through ecosystem conservation and restoration is known as the Reducing Emissions from Deforestation and Forest Degradation (REDD+, http://www.un-redd.org) program. Set up by the UN, REDD+ provides financial incentives to local communities and landowners that make conservation of carbon-sequestering ecosystems worth more than destroying them. Funding for REDD+ is obtained through carbon trading programs, in which individuals and organizations looking to offset their emissions buy carbon credits. The funds obtained through REDD+ are then invested in initiatives that promote ecological restoration and reduce local dependence of intact ecosystems by creating alternative income streams such as sustainable crop, timber, honey, milk, and meat production.
REDD+ provides financial incentives to local communities and landowners by making conservation of carbon-sequestering ecosystems worth more than destroying them.
The original aim of REDD+ was to safeguard primary old-growth forests, but the diversity of goals set out by REDD+ also include improving ecosystem connectivity, protecting threatened species, and preventing further loss and degradation of carbon-sequestering ecosystems other than forests. Africa has been a major beneficiary of this program. Since its inception in 2007, REDD+ projects have been funded in 27 Sub-Saharan African countries (http://www.reddprojectsdatabase.org), affording protection for over 1.6 million ha of forest (Panfil and Harvey, 2016), and providing opportunities for thousands of Africans who would not have had access to these funds otherwise.
Yet the future of these opportunities is not secure. Concerns exist regarding the effectiveness of REDD+ programs, much of which is based on problematic implementation, long-term funding security, lack of monitoring, and lack of concrete conservation goals (Phelps et al., 2011; Panfil and Harvey. 2015; Fletcher et al., 2016). For example, there are concerns that REDD+ programs can develop into a form of perverse subsidies, such as when native vegetation is cleared to establish plantations (Figure 10.12) with trees that have a high risk of becoming invasive (Lindenmayer et al., 2012). Similarly, there are concerns about the strong emphasis on forests, possibly at the expense of other important ecosystems and ecosystem services (Bond, 2016). Conservation biologists continue to be hopeful that REDD+, as with the range of approaches described in this chapter, will provide opportunities for land managers and scientists to successfully protect and restore biodiversity now and into the future.
10.05: Summary
1. Ecosystem conservation and management involves three different activities: (1) monitoring ecosystem components, (2) maintaining healthy ecosystems, and (3) restoring damaged ecosystems.
2. An ecosystem in which all the chemical, physical, and biological components and processes are functioning normally is considered healthy. Ecosystems that remain healthy through disturbance are resistant, while ecosystems that rapidly recover after disturbance are resilient.
3. Ecosystems can be monitored using direct observation, environmental or biochemical indicators, and remote sensing analysis. It is important that any monitoring method be consistent and repeatable across space and time.
4. To maintain ecosystems that can support diverse ecological communities, conservationists are guided by three complementary management principles: (1) maintain critical ecosystem processes (water cycling, nutrient cycling, energy flow, community dynamics), (2) minimise external threats, and (3) be adaptive yet minimally intrusive.
5. Ecological restoration is the practice of restoring damaged ecosystems to an agreed-upon benchmark. This can be accomplished via rehabilitation, partial restoration, complete restoration, or taking no action. The strategy followed will depend on each project’s goals and resource availability.
10.06: Topics for Discussion
1. Read the articles by Bunce et al. (2008) and/or Papworth et al. (2009) about shifting baselines. Then think of a natural ecosystem in your region; it could even be an ecosystem in a protected area. How do you think that ecosystem looked 50 years ago? What about 100 years ago? And 1,000 years ago? At what time do you think the ecosystem was able to support the most diverse ecological community? What would you do to restore (or maintain) the ecosystem to this state? Would such a restoration project impact some species negatively? Is that a problem?
2. Consider all the aquatic communities in your region (ponds, marshes, streams, rivers, lakes, estuaries, coastal waters, etc.). Who is responsible for managing these ecosystems, and how do they balance the need for protecting biodiversity with the needs of society for natural resources? What additional conservation projects would you implement to help protect those ecosystems in the coming decades?
3. Imagine that the last population of a threatened bird species (which draws birdwatchers to the area) lives along a river nearby. This river also has numerous endemic species of fish, shellfish, and insects. A foreign company recently obtained permission to dam the river for hydropower generation. Beyond the impact of flooding, the dam will also cause various forms of pollution which will destroy the threatened birds’ food source and nesting area. Upon writing about the challenge in the local newspaper, you receive US \$1 million from an anonymous donor to save the bird. The company is willing to forego the development in exchange for the US \$1 million. It will cost an additional \$750,000 to implement an effective ecological restoration program that can reverse the threatened birds’ population declines. Is it better to buy out the company and not devote additional resources to ecological restoration and researching the bird? Or would you rather spend the money on finding alternative ways to protect the bird and the other endemic species? Explain your answers.
10.07: Suggested Readings
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One of the following two manuscripts:
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There are many examples in this textbook illustrating how species have been saved from the brink of extinction. For some, the solution was simple: halt the threats that caused their populations to decline. In other cases, more drastic steps were required, like moving the last remaining individuals into captivity until the threats have been reversed. Many species that persist with low population sizes would likely not have survived without human intervention (Figure 11.1).
In each of these success stories, the most important first steps involved determining the ecological needs of the species at risk and understanding the factors that made that species vulnerable to extinction (Section 8.5). This chapter reviews some of the most important concepts for understanding and managing those needs and risks. The concepts reviewed in this chapter include methods to study species and populations, actions that can be taken to increase population sizes, and strategies that can help maintain evolutionary processes such as genetic exchange. This chapter also considers how to manage for climate change and discuss the importance of ex situ conservation strategies.
11.01: Studying Species and Populations
To save a species from extinction, it is vital to have a firm grasp on the species’ distinctive characters, in other words its natural history. To obtain this natural history information, 10 important factors need to be considered:
To save a species from extinction, it is vital to have a firm grasp on the species’ distinctive characters, in other words its natural history.
• Population biology: How many individuals are there in the population? How many males, females, juveniles, breeding adults, and individuals past breeding age are there? What is the species’ life expectancy? How have these aspects changed over time? (see also Chapter 9)
• Habitat: In what kind of environment can the species be found? How do these ecosystems change over time and space? Does the species have a complex life history that requires multiple habitats (e.g. frogs that live on land generally need water for breeding)? What factors are important to maintain suitable habitat?
• Distribution: Where in the world can the species of concern be found? At what rate is its distribution increasing/decreasing? What factors drive these increases/decreases?
• Morphology: What are the defining traits, or range of traits, of the species’ appearance? How do the species’ unique morphological characteristics help it survive? Are there closely-related species that appear similar (i.e. cryptic species) and with which it can be misidentified?
• Limiting resources: What types of resources does the species need to survive? Are any of these resources in short supply? Does the distribution of these important resources change over time and space?
• Physiology: Are there any special requirements the species’ physical and biochemical processes need for it to grow, survive, and reproduce? What are the conditions under which meeting these requirements is especially hard?
• Behaviour: How do individuals act or behave (Box 11.1)? Is the species sedentary, nomadic, or migratory? Do individuals group together, disperse at random throughout landscapes, or space themselves out at regular distances? How do these behaviors help it survive?
• Genetics: How much do genes vary within the species? How are the species’ genetics linked to its morphology, physiology, and behavior? Are there local genetic adaptations? Is the genetic variation in key traits sufficient to allow the species to adapt to environmental changes? Are there any deleterious genetic concerns? (Section 8.7.1)
• Biological interactions: In what ways do individuals of the species interact with each other and with other species? Which of these interactions are critical for survival? Are there any competitors, predators, parasites, or diseases affecting the species?
• Interactions with humans: How sensitive is the species to human activity? Do humans use the species in any way? Is the species sustainably harvested? Is the species associated with human-wildlife conflict (Section 14.4)?
Box 11.1 The Overlooked Role of Behavioural Ecology in the Conservation of African Mammals
Adrian M Shrader
Mammal Research Institute,
Department of Zoology and Entomology,
University of Pretoria, South Africa.
[email protected]
When considering the management and conservation of wild animals, topics linked to population and community ecology (e.g. carrying capacity, Hayward et al., 2007a) often come to mind. This is not surprising, as these disciplines consider broad patterns of population dynamics (e.g. birth rates and mortality rates), which are key to achieving management and conservation goals. While this information is necessary, in many instances, it fails to explain the mechanisms behind the patterns observed and answer key questions. For example, why do species prefer specific habitats? Why do some herbivores adjust their home ranges with the seasons? To answer these sorts of questions, we need to understand an animal’s behavioral ecology.
Take for example the challenge of understanding the impacts that elephants cause within protected areas. A standard way to assess these impacts is to record which tree species are damaged and how many trees are affected (e.g. broken branches, bark stripping) (Boundja and Midgley, 2010). While this provides information on the trees most vulnerable to elephant damage, it does not explain why elephants are damaging the trees. Is it because the trees are a key part of the elephants’ diet, or are these trees just abundant across the landscape and in the way of a moving herd? To answer these questions, we turn to behavioral ecology. By observing foraging elephants, or by walking down their feeding paths after they have left, we can determine the animals’ diet, and generate an acceptability index (number eaten ÷ number available) of each tree species (Shrader et al., 2012). These data allow us to better understand the reasons behind elephant damage.
Other situations where behavioral ecology can help include reintroductions, population management, and human-animal conflict mitigation. For example, in South Africa, oribi are locally threatened by habitat loss and poaching. One conservation strategy is to relocate individuals away from known threats. Oribi are grassland specialists (Figure 11.A) that require both short and tall grasslands—therefore, release sites require a mosaic of these habitats. Moreover, within grasslands oribi perceive woodland patches to be dangerous, and tend to avoid feeding within 15 m of them (Stears and Shrader, 2015). If we do not consider how oribi utilise their environment, our estimate of available habitat at a release site may be greater than the area utilised. This mistake could reduce relocation success.
With regards to population management, behavioral ecology is central to the conservation of southern white rhinoceros (Ceratotherium simum simum, NT) in the Hluhluwe-iMfolozi Park, South Africa. Within the park, the management policy incorporates space use and social ecology of the rhinos to facilitate population regulation (i.e. dispersal). To do this, the population can grow in the central core of the park. When rhino numbers get too high in the core, individuals naturally disperse into surrounding low-density areas, at which point they are captured by wildlife officers and transported to other areas. Thus, rhino behavior itself is used to indicate when there are too many individuals within the fenced park (Linklater and Shrader, 2017).
Finally, behavioral ecology has helped reduce human-elephant conflict through the understanding that elephants are afraid of bees and will avoid feeding close to them. To capitalise on this fear, fences that incorporate beehives were designed and constructed around agricultural fields in northern Kenya, which helped reduce crop damage from raiding elephants. Of 32 raids recorded in the area, only one was at a farm with a beehive fence (King et al., 2011). These examples showcase how behavioral ecology can support, expand, and strengthen management and conservation of wildlife. These same principles can be applied to protect a wide range of animals across Africa, and elsewhere.
Understanding the natural history of a species directly informs conservation strategies. For example, if we know where a species occurs and what its habitat needs are, we are in a better position to prioritise which areas need to be protected or how ecosystems need to be restored. Similarly, if we know that an important food resource is missing, perhaps during a drought or due to human activities, conservationists could provide supplemental feeding until the limiting resource has recovered (Figure 11.2). Depending on the species in question, some factors play a more important role than others. For example, managing a disease outbreak may play a more important role in the conservation of a widespread migratory bird (that can spread diseases to other species), while managing for genetic diversity may play a more prominent role in the conservation of a small population of fishes restricted to only one lake. For many widespread species, different factors affect different subpopulations. In such cases each subpopulation might need to be managed as its own evolutionary significant unit (ESU; see e.g. Dubach et al., 2013) to retain unique local adaptations and genetic markers.
Obtaining natural history data
Conservationists rely on several resources and techniques to obtain natural history information. Initial steps often involve reviewing published and unpublished literature to understand what is known (and not known) about a species. Literature reviews do have some drawbacks: they can take a long time, may uncover contradictory information, and may lack critical information relevant to a local area or specific population. For this reason, and especially when decisions need to be made under tight schedules, conservation biologists may need to speed up their initial species review by sourcing natural history information from subject matter experts who are familiar with the species or ecosystem of concern.
Conservation biologists are increasingly recognizing the importance of traditional ecological knowledge—detailed insights that rural people have on the species around them.
Conservation biologists are also increasingly recognizing the importance of traditional ecological knowledge (TEK)—detailed insights that rural people have on the ecology, behavior, and distribution of the species around where they live (Shackeroff and Campbell, 2007; Brook and McLachlan, 2008). For example, while termites are often considered a pest by people living in urban settings, scientists are increasingly relying on TEK to understand the important contributions of termites to food security to human health, as well as to learn about ecological sustainable methods for their control when needed (Sileshi et al., 2009).
While literature reviews, expert opinions, and traditional ecological knowledge are important first steps to collect natural history information, the most reliable method remains fieldwork, where multiple individuals from the population of concern in the area of interest are observed repeatedly over time. Indeed, most of natural history information we have today was obtained during detailed notetaking by naturalists—biologists who dedicate much of their time to better understand the natural world—in the field.
Unfortunately, there are still major gaps in our understanding of the living world. Consequently, a very large number of threatened species, including better-known groups (e.g. reptiles, Tolley et al., 2016), lack the kinds of data necessary to ensure that we can give them the best chance of survival. Filling these gaps is also becoming harder since it is costly and sometimes logistically impossible (or dangerous) for naturalists to spend an extended period in the field. There is also a trade-off in the breadth and depth of data collection possible: the more area one covers, the less detailed the data; conversely, when one collects more detailed data, the scope of the study is reined in for logistical constraints. Further, there is also a limit to the number of organisms any one individual observer can study at any one time.
Recent technological advances have greatly increased our ability to overcome the logistical constraints that impede conservation fieldwork. One of the most useful developments involves the miniaturisation (and reduced costs) of animal-borne biologging devices, such as radio telemetry and GPS tags (Kays et al., 2015). Previously reserved for projects with large grants that focused on large animals, the big clunky devices of a few decades ago have made way for devices small enough to fit comfortably on animals as small as beetles and frogs. Some biologging devices are now also solar-powered and transmit data through Earth observation satellites in real time, allowing researchers to track the behaviors of several organisms at a time from the comfort of their offices. Even better, some tracking technologies also collect environmental data and movement data simultaneously, allowing us to better understand how wildlife responds to changing environmental conditions. These new and sophisticated datasets can then be used to better understand threats to species (e.g. Scantlebury et al., 2014; Childress et al., 2016) and inform management of protected areas (e.g. Maxwell et al., 2011).
Species distribution modelling (SDM), also known as environmental niche modelling, is becoming increasingly popular for determining a species’ distribution and habitat needs. SDMs overlay species location data, obtained during field work or using biologging devices, onto a selection of relevant environmental variables (e.g. forest cover, elevation, soil type) using GIS software, after which special modelling algorithms estimate the species’ ecological niche and distribution (Figure 11.3, see also Figure 10.3). This information enables conservation biologists to identify previously unknown habitat patches (which may represent undiscovered and unprotected populations) or empty habitats (which may be used in translocations, see Section 11.2). The appeal of SDMs lies in the availability of user-friendly software packages that can use very limited datasets. For example, one study from West Africa successfully combined market survey data and SDM to determine the potential for sustainable extraction of 12 medicinal plant species (van Andel et al., 2015). Another study used SDMs to develop a holistic picture of diversity and endemism patterns of nearly all 250 African bat species (Herkt et al., 2016). While distribution modelling offers very useful conservation tools, it is important to learn about the different techniques under the guidance of an expert to avoid making costly mistakes (McPherson et al., 2006; Pearson et al., 2006).
Species distribution modelling, also known as environmental niche modelling, is becoming increasingly popular for determining a species’ distribution and habitat needs.
Experimentation offers powerful methods to better understand competing theories and hypotheses, and to gain insight into how specific management actions may influence population dynamics. Experimentation is often associated with controlled environments such as laboratories; however, this is often impossible and sometimes even unethical to perform laboratory experiments on threatened species. Instead, conservation researchers may opt for natural experiments, which allows for the target species or population to be studied in its natural ecosystem.
A chronosequence study is a special type of natural experiment that overcome the long-term commitment some studies require to attain meaningful results. Also called space-for-time experiments, chronosequence studies allow us to infer long-term trends over a short study period using study systems that share similar qualities but are differently aged. Chronosequence studies are particularly popular when studying ecological restoration projects (Section 10.3) since some ecological processes often require many decades to develop (Bonnell et al., 2011). In one such example, conservation biologists needed only three summers worth of vegetation surveys to show that some species recolonise coastal dune forests in the Maputaland-Podoland-Albany Biodiversity Hotspot only after 100 years since disturbance (Wassenaar et al., 2005).
Sometimes, despite their best efforts, biologists may still fail (or may not have enough time) to obtain much needed natural history information during a critical period. To overcome such a challenge, biologists have, at times, used natural history information of a substitute species (which is different from surrogate species, Section 13.3.5) to fill data gaps for a rare species (Caro et al., 2005). An example of this application comes from the USA where researchers used behavioral observations of a common butterfly to predict dispersal of another closely related butterfly that was too rare to properly study (Hudgens et al., 2012). It is important to note that using information from substitute species does have serious limits (Henry et al., 2019). For example, considering that different populations of a single species may have very different environmental needs and adaptations, using data from a different species may be even less useful. Care must therefore be taken when using data from substitute species with proper acknowledgement of the assumptions and uncertainty this approach adds to one’s research. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/11%3A_Preventing_Extinctions/11.00%3A_Prelude_to_Preventing_Extinctions.txt |
Because the probability of extinction increases rapidly for small populations (Section 8.7), conservation biologists often invest considerable energy into increasing the size of small and declining populations. Often, these projects involve improving the extent and quality of suitable habitat (Chapter 10) or mitigating threats such as overharvesting (Chapter 12). When appropriate, conservation biologists may sometimes resort to translocations—moving individuals from sites where they are threatened (e.g. unprotected lands or a paper park) or overabundant (e.g. a well-managed protected area or ex situ conservation facility) to sites where they can offer a larger contribution to conservation efforts.
Understanding a species’ ecological needs is critically important for translocations, because it influences the choice of release site and type of preparations needed.
Conservation biologists generally recognize four basic translocation approaches:
• Restocking (also called augmentation) occurs when wildlife managers increase the size and genetic diversity of existing populations, by releasing individuals that have been raised in captivity or that have been obtained from other wild populations.
• Reintroduction occurs when wildlife managers release individuals into areas where they occurred in the past but not at present. The areas must be ecologically suitable and the factors that caused the extirpation must have been reduced or eliminated for a reintroduction to be successful.
• Introduction involves creating new populations by moving individuals to suitable areas outside that species’ historical range. Introductions are usually considered when reintroductions are impossible because the species’ historical range has been degraded too severely or because persistent threats will lead to reintroduction failure.
• Assisted colonisation (also called assisted migration) is a special class of introduction where biologists “assist” species with poor dispersal capabilities to adapt their ranges in response to environmental changes. It is anticipated that this strategy will become an important conservation tool in preventing extinctions where climate change outpaces the speed of natural migration.
Important considerations for translocations
Section 11.1 broadly discussed the importance of understanding the ecological and other natural history needs when protecting threatened species. Understanding a species’ ecological needs is equally, if not more, important for translocations, because it influences the choice of release site and type of preparations needed (Figure 11.4). Complementing the 10 factors mentioned in Section 11.1, the next section briefly introduces some of the most important considerations during translocations.
Determining need and feasibility
Perhaps the most important factor to consider before starting a translocation is to determine whether it is necessary. Translocations carry risks, not only for the target population to be moved, but also the individuals left behind and for the recipient ecosystem. These risks expose translocation projects to a high risk of failure, particularly if preparations are inadequate and essential resources (e.g. funding, trained staff) are in short supply. Translocations also demand considerable resources—resources that can at times be better spent mitigating the threats the target population face. While these considerations may seem obvious, a recent review found that most translocations projects are initiated without proper cost-benefit analyses (Pérez et al., 2012). To improve translocation practices, conservationists seriously considering a translocation project are encouraged to review the 10 criteria outlined in Pérez et al. (2012), some of which also overlap with the considerations mentioned below.
Support from local stakeholders
It is also important to consider, at an early stage, how the public will view the translocation project. Some people may feel the resources used in a translocation are better invested elsewhere; others dislike translocations because they view it as a threat to their livelihood—this is especially true when carnivores are involved (Gusset et al., 2008a). Because of these and other potential conflicts and emotions, it is crucial that translocation projects (like any conservation activity) obtain the support from local stakeholders at an early stage. It is helpful to be transparent from the outset and to explain the project’s goals, as well as the benefits the local community may gain (e.g. attract more tourists, restore a degraded ecosystem service). Good public outreach also provides opportunities to address the public’s concerns and misconceptions about the project and about biodiversity conservation in general.
Identifying suitable habitat
It goes without saying that the probability for success is greatly improved when the translocated individuals are released in good quality habitat. This is particularly true for species with poor dispersal capabilities, such as plants that reproduce through vegetative propagation: the plants could die in an environment that is too sunny, shady, wet, or dry. While this point may seem obvious, many translocations fail because individuals are released in inferior habitats (Armstrong and Seddon, 2007). One of the reasons for this potential habitat mismatch is because wildlife may perceive the environment differently than humans, so a site that may look good to the human eye may lack one or more overlooked limiting resource. Refugee species—species forced to live in suboptimal habitat due to threats present in their preferred habitat (e.g. Ali et al., 2017)—also present a challenge to biologists who may unwittingly view inferior habitat as optimal and base conservation decisions on essentially bad information. The same challenge presents itself at ecological traps—unsuitable environments that an organism mistakenly perceives as optimal habitat (e.g. Sherley et al., 2017). These are some of the most important reasons why biologists need to be cautions when using species distribution models (SDM) when identifying areas suitable for translocations. To mitigate costly translocation failures, it is advisable that releases start small, and have multiple phases, to assess how released individuals respond to their new environment. Conducting experimental and adaptive releases can also reduce uncertainty by evaluating different release scenarios (Menges et al., 2016).
Species forced to live in suboptimal habitat due to threats present in their preferred habitat may lead biologists to unwittingly view inferior habitat as optimal.
It is also important to ensure that any habitat identified as suitable is free from threats such as pollution and invasive species that may lead to declining health or even death for released individuals. A project in the Cape Floristic Region in South Africa provides a good example of how alert conservation biologists mitigated a threat that could have caused a translocation failure. The Clanwilliam sandfish (Labeo seeberi, EN) was once widespread in the region’s Olifants-Doring River system. However, recent surveys indicated that the species had gone extinct in the Olifants River. Although biologists did find some juvenile fish in the Doring River and some of its tributaries, they also noticed that invasive fish predated on most of those juveniles before they reached adulthood (Jordaan et al., 2017). These ill-fated individuals were thus dispersing from the last remaining reproductive subpopulation persisting in the headwaters of one single Doring tributary to other parts of the river, which acted as a population sink. To prevent the species’ extinction, biologists initiated a habitat restoration plan involving restoring natural stream flow regimens and eradicating predatory invasive fish in the headwaters of a second Doring tributary. They then installed barriers that prevented invasive fish from reaching the restored area before translocating 338 juvenile fish (Figure 11.5) there. With this habitat restoration plan, the biologists hope to establish a second viable population, and to improve the juveniles’ chances of surviving to adulthood before they disperse back to areas where the invasive fishes occur (Jordaan et al., 2017).
Considering genetics and behavior
Translocation projects also need to consider the genetic makeup, social organization, and behavior of a species that is being released. It is preferable to use individuals from the same genetic stock as individuals that already occur (or have occurred) in the release area to avoid outbreeding depression and to capture local adaptations (Sections 8.7.1). Such efforts simultaneously also contribute to conservation of genetic diversity, as opposed to the pollution thereof if individuals from different genetic stock are mixed.
Group-living species, particularly those vulnerable to Allee effects (Section 8.7.2), need to be released in sufficient numbers so they can maintain their natural social organization and behavior. For species that need to be released in groups, it is preferable to release socially integrated animals rather than individuals unfamiliar with each other (Gusset et al., 2008b). Releasing groups of animals does have its own set of challenges. For example, social groups abruptly released from captivity may disperse explosively, possibly leading to project failure. This happened with African buffalo (Syncerus caffer, NT) herds translocated to South Africa’s Addo Elephant National Park which fragmented into smaller groups after release, making them more vulnerable to lion (Panthera leo, VU) predation (Tambling et al., 2013). Fortunately, in this case, the buffaloes underwent several behavioral modifications over time, which eventually allowed their numbers to stabilise (Box 11.2). This contrasts with failed rock hyrax (Procavia capensis, LC) reintroductions in South Africa, where group disintegration post release exposed the animals to unsustainable predation levels (Wimberger et al., 2009).
Box 11.2 Large Predator Reintroductions: A Balancing Act
Craig J. Tambling
Department of Zoology and Entomology, University of Fort Hare,
Alice, South Africa.
[email protected]
Large predator numbers are declining, and African carnivores are no exception (Ripple et al. 2014). How to conserve African carnivores are a hotly debated topic now, with “fortress” type conservation areas considered the most viable option by many (Packer et al., 2013). In South Africa, this conservation model is the norm, and many small protected areas are now translocating large carnivores for ecotourism. However, these large carnivore translocations have repercussions for resident prey species. Understanding the ecological and biodiversity consequences of these translocations is thus important for the management of these small protected areas (Tambling et al., 2014).
In 2003, lions and spotted hyenas (Crocuta crocuta, LC) were reintroduced into the Addo Elephant National Park Main Camp Section after being absent from the area for over 100 years. Post-release monitoring of the six reintroduced lions indicated that at least 50% of their diet in the first two years following reintroduction was African buffalo. This was especially concerning to South African National Parks as this resident buffalo population contributes substantially to game auction sales each year, with the money raised being used to expand the national park system in South Africa (SANParks, 2009).
Following high predation rates of buffalo by lion and a 2007 buffalo census suggesting low juvenile recruitment, the coexistence of lion and buffalo in Addo was questioned. These concerns lead to a detailed assessment of buffalo behavior and demographics between 2008 and 2011 (Tambling et al., 2012), which showed that by 2008–2009, juvenile buffalo recruitment (Figure 11.B) had rebounded to levels reminiscent of those prior to the lion reintroduction. Direct observations of the buffalo population showed drastic behavioral alteration following the high initial predation rates by lions. These behavioral changes included: (1) increased breeding herd sizes, (2) a reduction in nocturnal movement, and (3) greater use of open habitats at night and early morning when lions are hunting. These behavioral adjustments enabled the active defence of the breeding herds, reducing successful predation by lions and ensuring an increase in buffalo recruitment. Although this study suggests that prey populations are capable of behavioral adjustments to reduce predation, this is not always the case, with some species unable to respond, leading to precipitous declines in prey populations such as eland (Tragelaphus oryx, LC) (Leaver, 2014).
The ultimate aim of most translocation projects is to establish populations that are self-sustaining, free from inbreeding, and interactive participants of their communities and ecosystems.
Case studies of predator-prey interactions following large predator reintroductions highlight the management challenges faced by small reserves where ecotourism, biodiversity, and financial goals each need to be met. Due to the small size of these “fortress” reserves, a local overabundance of predators can have severe ecological effects on prey populations. However, in many reserves, the high demand for large predators for ecotourism often results in costly reactive, rather than scientifically sound proactive, management. There is, however, a growing body of research on the proactive management of large carnivores, where wildlife managers aim to replicate ecological processes (i.e. lion inter-birth intervals) to limit management interventions required to control large predator numbers (Ferreira and Hofmeyr, 2014). In small reserves, lion inter-birth intervals are shorter than in large ecosystems, and so lengthening the inter-birth intervals to that observed in large ecosystems can reduce lion population growth rate in these small reserves (Miller et al., 2015). Understanding predator-prey interactions is important regardless of the conservation model employed to protect these large charismatic species.
How many individuals to release
The ultimate aim of translocation projects is to establish ecologically relevant populations, meaning populations that are self-sustaining, free from inbreeding, and an interactive participant of its community and ecosystem. The probability of achieving this goal increases as more individuals are being released. Because translocation projects typically do not have an unlimited supply of individuals to release, wildlife managers often rely on quantitative models (Section 9.2) to estimate the minimum number of individuals that should be released and how many times releases should occur. For example, a population viability analysis (PVA) on western lowland gorillas (Gorilla gorilla gorilla, CR) reintroduced to Gabon and the Republic of the Congo showed that the probability of persistence of an apparently established population could be increased significantly if more individuals were released (King et al., 2014).
The ability to establish new populations through translocations does not reduce the need to protect threatened species still in their natural habitats.
While releasing more individuals certainly improves the likelihood of establishing a self-sustaining population, it is also important to determine how many individuals the target community can sustain. In other words, the release area should contain enough suitable habitat to support the territories of all the released individuals. To determine how many individuals can be sustained, wildlife managers may calculate the release area’s carrying capacity—an estimate of the maximum number of individuals an ecosystem can support. The carrying capacity concept has its roots in the livestock trade, where farmers wanted to maximise the number of animals on their land without risking overgrazing. While the concept has gained popularity in conservation biology in recent decades, calculating the carrying capacity for wildlife is very complex because of all the multi-faceted interactions that characterize healthy ecosystems. For example, the carrying capacity for a wild population can depend on factors such as food, water, shelter, soil nutrients, and sunlight availability, as well as more species-specific natural history factors such as habitat quality, home range, sex ratios (Tambling et al., 2014), and interactions with other species (Lindsey et al., 2011).
Over the past few decades, through trial and error, adaptive management (Section 10.2.3), and the collection of vast amounts of demographic data, scientists have made significant progress in calculating carrying capacity for wildlife populations. Perhaps the most progress has been made in calculating carrying capacities for large ungulates, by monitoring vegetation biomass, which in turn is affected by soil nutrients and rainfall (Fritz and Duncan, 1994). Much progress has also been made in calculating carrying capacities of predators by monitoring prey densities (Hayward et al., 2007a). For most populations, however, carrying capacity isn’t explicitly calculated, but implicitly estimated based on intuition. Refining existing carrying capacity models and developing new methods for other taxa remain an active area of research that will hopefully reduce conservation biologists’ over-reliance on intuition in future years. But even in the absence of carrying capacity calculations, wildlife managers can track a population’s health and overall fitness. When the health of a particularly successful population or its environment starts declining, a root cause may be that too many individuals have been released, or the population is being sustained above carrying capacity.
Preparing individuals for release
Translocation projects using individuals obtained from the wild are generally much more successful than those using captive-bred individuals, given that wild individuals are already adapted to a life where they must fend for themselves. Nevertheless, some projects may have to use captive-bred individuals, particularly when the target species is extinct in the wild, or when individuals were brought to an ex situ conservation facility because it is easier to breed them under human care in controlled conditions. In such cases, a great amount of effort may be required to prepare the captive-bred individuals for releases.
A great amount of effort may be required to prepare captive-bred individuals for translocation because they may have lost adaptations required for survival and reproduction in the wild.
A major drawback when using captive-bred individuals is that they may have lost the important adaptations required for survival and successful reproduction in the wild. Pre-release training, which varies according to the species, can sometimes overcome this drawback. For predators, it may involve providing low risk prey, such as chickens and domestic rabbits in holding facilities until their hunting skills are better developed (Houser et al., 2011). For plants propagated indoors, it may involve hardening them off by placing them outside for increasingly longer periods to gradually introduce them to sun, wind, and temperature changes during the day. To help young birds disassociate humans from food, human trainers sometimes use puppets or wear costumes (Figure 11.6) during feeding time to mimic the appearance and behavior of wild individuals (Valutis and Marzluff, 1999). Another method, which may promote behavioral enrichment, involves cross-fostering, in which unrelated parents helps raise the offspring of a threatened species. In carnivore conservation, this technique has shown much promise to augment litter size and encourage gene flow using orphaned African wild dog (Lycaon pictus, EN) pups (McNutt et al., 2008). Interspecific cross-fostering has also been used in bird conservation, where biologists use common species to incubate eggs abandoned by threatened species (e.g. Powell and Cuthbert, 1993). However, cross-fostering using different species may lead to a new set of problems, like behavioral changes and hybridisation, if the young subsequently associate with the wrong species. A great amount of care and research are thus needed before such strategies are attempted.
Whether using captive-bred or wild individuals for translocations, individuals may have to be fed, sheltered, trained, or otherwise cared for after release to give them time to become more familiar with their new surroundings. This approach, known as soft release, involves keeping the released individuals in pre-release holding facilities for a period; it may also include some form of assistance after release to increase opportunities for success. Soft releases also provide an opportunity to introduce captive-bred organisms to wild individuals of the same species that can act as “instructors” for survival in the new environment, or for unfamiliar individuals to bond into cohesive units (Gusset et al., 2006).
The alternative to soft release is a hard release—an abrupt release of individuals from captivity without assistance such as food supplementation. While hard releases are popular (because they are relatively easy to perform), it is a risky strategy that faces a high risk to failure (Brown et al., 2007; Wimberger et al., 2009). Hard releases can however be appropriate under the right conditions (Hayward et al., 2007b). For example, hard releases are often use in head-starting programs (Figure 11.7) for reptiles and amphibians (Scheele et al., 2014), where conservation biologists collect wild individuals and raise them past their most vulnerable life stages before releasing them again where they were collected.
Post-release monitoring
A translocation project does not end after the last individual was released. Rather, ongoing monitoring should be implemented to determine whether a translocation was successful, what degree of success was achieved, whether adaptive management is needed, whether additional releases should be conducted, or whether the project should be aborted. A well-designed monitoring plan can also highlight the consequences of translocation on the broader ecosystem, such as the impact that predators introduced to a new area may have on prey populations (Box 11.2) and competing species (Groom et al., 2017). Because some responses in translocated populations can be rather subtle and take many years to show or subside, post-release monitoring should ideally be a long-term endeavour. For example, by monitoring seemingly successful elephant reintroductions across five protected areas in South Africa, researchers found that stress hormones in released animals continued to decline 24 years post release (Jachowski et al., 2013). Long-term monitoring will also help wildlife managers better understand the ultimate fate of the released individuals. Many apparently successful translocations fail because the released individuals die after several years without ever reproducing. Highlighting the importance of post-release monitoring, one study from South Africa found that 70% of captive-bred oribi (Oribia oribi, LC) died within two months of release, mostly due to predation (Grey-Ross et al., 2009). Another study found that reintroduced cheetahs were all killed within a year of release (Houser et al., 2011). These were expensive lessons, but post-release monitoring ensured that the reason for failures are known and can be addressed ahead of future releases.
Helping other translocation projects
Strategies used in successful translocation projects were nearly always informed by releases conducted by other wildlife managers who circulated their experiences to the wider conservation community. It is important to pay this effort forward; new translocation projects should make every effort to track and publish their results to inform others. While it is always easier to present the results of successful projects, publishing the lessons from failed projects is also important (Wimberger et al., 2010; Godefroid et al., 2011). Equally important is the publication of project costs, to enable wildlife managers to better determine under which conditions translocations represent a cost-effective conservation strategy. For example, a large African wild dog reintroduction program in South Africa achieved their initial goal of establishing nine self-sustaining packs much more quickly than expected—five years rather than 10—yet reintroducing all these populations cost 20 times more than if the funds were used to enhance protection of existing packs within protected areas (Lindsey et al., 2005). With more information available, future conservationists would hopefully be able to have better guidelines to maximise cost-effectiveness and the likelihood of project success. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/11%3A_Preventing_Extinctions/11.02%3A_Saving_Species_Through_Translocations.txt |
Some ecosystems are transient in nature—their character is temporary and will change because of disturbance and succession. Consequently, species that occupy those transient habitats are bound to be naturally extirpated at one time or another. Consider, for example, a small population of wildflowers occurring in a river’s floodplain; at some stage, there is going to be a flood that will wash away those flowers. But the flooding also disperses seeds downstream, allowing for new wildflower populations to establish in suitable habitat elsewhere. These shifting populations linked by movements between them are better characterized as a metapopulation (a “population of populations”) (Figure 11.8) consisting of several subpopulations. For some metapopulations, every subpopulation is transient: their distribution changes dramatically with each generation. Other metapopulations involve relatively permanent subpopulations with only a few individuals dispersing each generation. Some metapopulations consist of one or more source populations whose sizes are stable or increasing, and several sink populations whose sizes fluctuate depending on environmental conditions. Some sink subpopulations may undergo such dramatic fluctuations that they would be extirpated in unfavourable years were it not for population rescue by immigrants from source populations.
A metapopulation (a “population of populations”) consists of several subpopulations linked by movements of individuals between them.
Habitat fragmentation threatens metapopulation dynamics by reducing opportunities for dispersal across the landscape (Chapter 5). When there is too little movement of individuals between habitat fragments, the dwindling subpopulations within those fragments are at risk of extirpation or even extinction (Section 8.7). In contrast, well-connected subpopulations maintain themselves by colonising empty niches, exchanging genetic material, and adapting to changing environments. Dispersal also maintains critical ecosystem processes, such as pollination and seed dispersal (Section 4.2.5). Consequently, conservation biologists have invested significant resources in recent years to maintain and restore wildlife movements within fragmented ecosystems.
Connectivity in terrestrial ecosystems
Maintaining and restoring ecosystem connectivity—the ability of ecosystems to facilitate the dispersal of individuals between different areas—involves maintaining and restoring wildlife movements that are (at risk of being) impeded by human activities. The most popular method to maintain (or restore) connectivity in a fragmented landscape is to maintain (or restore) habitat linkages, also called wildlife corridors, habitat corridors, dispersal corridors, or movement corridors. All these terms refer to continuous tracts of suitable habitat with little to no dispersal barriers that connect otherwise isolated habitat patches and populations.
Maintaining and restoring ecosystem connectivity is an important strategy for conserving wildlife whose movements are impeded by human activities.
Some of the most prominent efforts to restore habitat linkages involve habitat restoration. For example, plans are currently underway to use forest regeneration to reconnect nine forest fragments in Tanzania’s East Usambara Mountain; if successful, this project would establish the largest contiguous forest block (over 3,000 km2) in the Eastern Arc Mountain Biodiversity Hotspot (Newmark, 2008). The positive impact of this project is expected to be immense. It has been estimated that the restoration of just 80 km2 of forest would stave off the first fragmentation-induced extinctions by over 2,000 years, compared to an estimated seven years until the first extinction if these forest fragments were to remain unconnected (Newmark et al., 2017).
Connectivity is important in every ecosystem on Earth. However, given the linear characteristic of riparian zones along rivers and stream—and hence a larger proportional impact of edge effects (Section 5.1.2)—we might consider connectivity in these spatially restricted systems to be particularly important (Figure 11.9). Protecting and restoring riparian zones as habitat linkages resonates with a variety of people because these areas provide a range of important ecosystem services, including flood control and water purification (Section 4.2.4). Conservationists can tap into this energy by lobbying for laws that prohibit activities such as logging, housing, and industrial developments within a certain distance from a river or stream. By protecting ecosystem services associated with riparian zones, these laws simultaneously also maintain wildlife refuges (Monadjem and Reside, 2008), source populations (Vosse et al., 2008), and habitat linkages (Bentrup et al., 2012; McLennan and Plumptre, 2012). In contrast, inadequate protection of riparian ecosystems not only compromises connectivity, but also negatively affect species not overtly dependent on these buffer areas. For example, research from Southeast Asia has shown that losing riparian ecosystems in an otherwise palm oil dominated landscape reduced stream quality, which in turn reduced local fish diversity by up to 36% (Giam et al., 2015). In contrast, protecting riparian zones were found to increase palm oil yields (Horton et al., 2018). With so many riparian areas currently being degraded and destroyed, there is an urgent need for stronger riparian protection laws (Chapter 12), and for more effective enforcement of those laws.
Restoring connectivity may also involve removing or otherwise mitigating human constructs that block wildlife dispersal. This is a major aim of TFCAs, which aim to restore dispersal between protected areas (Jones et al., 2012) by removing fences and other human constructs while still maintaining sustainable land tenures (Andersson et al., 2013). These efforts, accomplished through partnerships with local communities, are re-establishing historical mass migration routes, which in turn will hopefully also boost those areas’ ecotourism potential (Box 11.3). Efforts to revive extinct mass migrations also seem to be paying off! For example, in Botswana, the removal of veterinary fences—meant to prevent spread of diseases from wildlife to livestock, but also cutting off the world’s second largest wildebeest migration—have seen several hundred plains zebras (Equus quagga, NT) returning to old migration routes within four years (Bartlam-Brooks et al., 2011).
Box 11.3 Transfrontier Conservation Areas: Managing Biodiversity Across International Boundaries
Simon M. Munthali
Kavango-Zambezi Transfrontier Conservation Area Programme,
Kasane, Botswana.
http://www.kavangozambezi.org
TFCAs are components of a larger ecosystems that straddles the border between two or more countries, encompassing one or more protected areas as well as multiple-resource areas used by communities and private landholders. They are also managed for sustainable use of natural resources (Singh, 1998). The concept recognizes that borders are political rather than ecological (Dallimer and Strange, 2015), and aims to ensure that key ecological processes continue to function where political borders have divided ecosystems, river basins, or wildlife corridors (Cumming, 1999).
TFCAs are widely being established in Africa. One of these is the 520,000 km2 Kavango-Zambezi TFCA (KAZA)—a conservation and development initiative of Angola, Botswana, Namibia, Zambia, and Zimbabwe.
The benefits of the KAZA include:
• Re-establishment of the seasonal wildlife migration routes and connectivity among the many protected areas (national parks, community conservancies, and wildlife and forest reserves) within the region (Figure 11.C). The primary wildlife focus is the savannah elephant (Loxodonta africana), whose population of about 250,000 is predominantly concentrated in Chobe National Park (Botswana), Hwange National Park (Zimbabwe), and Bwabwata National Park (Namibia). Elephants need unimpeded movement to protected areas where population densities are much lower, such as Luengue-Luiana and Mavinga National Parks (Angola), and Sioma Ngwezi and Kafue National Parks (Zambia). This movement would reduce pressure on the ecosystems that are currently overpopulated and enable elephants and other species to better coexist—especially grazing herbivores that depend on the same habitats as the elephant.
• Expanding the wildlife-based economy, primarily ecotourism, into agricultural marginal areas (with predominantly Kalahari sand soils), through community-private partnerships. Through these partnerships, local communities would benefit from employment and business opportunities in ecotourism activities.
• Opportunities for local communities to participate in decision-making, and influencing policies and legislation related to natural management such as coordination of the fishing closed season between Namibia and Zambia during the fish breeding season (December–March) in the Zambezi River.
• Formation of alliances among different stakeholders (governments, private sector, NGOs, and local communities) to maximise skills and resources in promoting sustainable land use, conserving biodiversity and alleviating poverty.
Despite these benefits, there are obstacles to progress in attaining the benefits of the KAZA. Notable among these are social and political factors, such as increasing human population density, increasing cultivation of land, and expanding human settlements in wildlife corridors. Many of these factors trigger human-wildlife conflicts and poaching both for local consumption of bushmeat and for the illegal sale of elephant ivory. To mitigate these threats, the following strategies are being implemented:
• A Master Integrated Development Plan for the KAZA has been developed, which provides initial zoning. Its key feature is spatially allocating land into various uses (human settlement, agriculture, and protected wildlife areas, including wildlife dispersal corridors). The Master Integrated Development Plan also assists in creating awareness about the value of the wildlife corridors, which traverse communal areas.
• Promotion of conservation agriculture as a tool for improving land stewardship, intensification of agriculture, and improving crop yields per unit area of land, and therefore decreasing the likelihood of cutting down forested areas in and around wildlife corridors to plant new agricultural fields. Currently, within the KAZA, conservation agriculture is being piloted in Angola, Namibia, and Zambia. Conservation agriculture is crop production that strives to achieve acceptable profits together with high and sustained production levels while concurrently conserving the environment.
• Promotion of community-private partnerships in ecotourism development. Over the past four years, Ngoma safari lodge (Botswana), and Machenje sport fishing lodge (Zambia) have been developed specifically in support of securing wildlife corridors. They also provide incentives to the local communities for adopting wildlife conservation as a supplement to their land use practices. These lodges are in addition to the numerous existing tourist resorts in the KAZA.
• A law enforcement and anti-poaching strategy for the KAZA is being developed to coordinate transboundary law enforcement surveillance and fines to prevent poaching of protected wildlife. In addition, KAZA partner countries are integrating other security agencies, such as the military, police, immigration, and customs officials, to prevent the illegal export of wildlife products such as elephant ivory and bushmeat out of the KAZA.
• Reducing human-wildlife conflicts (Section 14.4) through improved land use planning, solar-powered electrified fencing encircling clusters of village fields and facilities and use of chilli-pepper-based olfactory repellents to deter elephants from entering crop fields.
The KAZA has made considerable progress to date in coordinating conservation efforts among the wildlife agencies and national parks across five countries in Southern Africa. The principal success has been measures to allow the continued migration of elephants along existing migration routes across international borders. The challenges ahead—from inadequate funding for wildlife patrolling and anti-poaching activities to increasing populations of rural people outside the protected areas and across migration routes—remain significant.
Section 5.1.1 discussed how inconsiderate fence placements threaten wildlife, while the paragraph above explained how removing fences can improve connectivity. Ironically, and illustrating the difficulties conservationists face when dealing with conflicting demands, strategically placed fences can sometimes also be used as a conservation tool. For example, researchers working on a fragmented lion population in Botswana found that the most effective way to improve this population’s viability was through strategic placement of fences to direct dispersal between protected areas (Cushman et al., 2016). Strategically placed predator-proof fences may at times also be required to avoid human-wildlife conflict (Packer et al., 2013, but see Creel et al., 2013), and to facilitate the recovery of threatened species, as is the case for Africa’s rarest antelope, the hirola (Beatragus hunter CR) (Ng’weno et al., 2017). The final word here is that management must remain responsive to both positive and negative impacts of tools, such as fences, rather than relegating them to bins, such as good or bad. (See also Dupuis-Desormeaux et al., [2018] for the use of fence-gaps and exclusionary fences to mitigate some negative fence impacts.)
Protecting and restoring stepping stone habitats can maintain connectivity in areas where it is impractical to establish or restore continuous habitat linkages.
At times, when it is impractical to establish or restore continuous habitat linkages, biologists may opt to protect and restore stepping stone habitats (Figure 11.10). As the name implies, stepping stone habitats are a special type of habitat linkage that facilitate dispersal along a patchwork of isolated habitat patches within a matrix of unsuitable or inhospitable habitat. Stepping stones thereby divide long dispersal events through a long stretch of inhospitable terrain up into shorter, and thus more manageable, sections. Stepping stone habitats are particularly important for migratory species that rest and refuel at stop-over sites between the end-points of their migratory route (Runge et al., 2015)—each stop-over site can be viewed as a stepping stone habitat. Prominent examples of stepping stone habitats that deserve protection include sacred forests which can act as stop-over sites for migratory forest birds; wetlands and estuaries (see Box 5.3), which can act as stop-over sites by migratory waterbirds; and small forest reserves, which can act as stepping stones between a network of other protected areas (Riggio and Caro, 2017).
Connectivity in freshwater ecosystems
Dams have always played an important role in hydropower generation and securing a year-round supply of water for farms, industries, and cities. Unfortunately, recent evidence suggests that reservoirs may create more problems than they solve (Section 5.3.2). Of concern is their contribution to greenhouse gases (Deemer et al., 2016), as well as their role in blocking dispersal of aquatic organisms. To counter these negative impacts, governments across the world are decommissioning and removing dams and other types of artificial water impoundments. For instance, over the past 30 years more than 1,174 dams were removed in the USA; the 72 dams removed in 2016 alone restored more than 3,000 km of streams (Thomas-Blate, 2016). Similar efforts are also underway in Europe (http://www.ecrr.org), where river restoration efforts have been initiated at over 1,100 locations across 31 countries. Unfortunately, not only are efforts to restore freshwater connectivity lagging across Africa; in many cases, even more rivers are currently being dammed (Winemiller et al., 2016).
While dams play an important role in hydropower generation and securing a year-round supply of water, recent evidence suggests that they create many environmental problems, including blocking species dispersal.
Connectivity in marine ecosystems
Ecosystem connectivity is also important in marine ecosystems. Many marine organisms, including economically important species, breed and feed in different areas at different times of the year, and use established dispersal routes to move between those areas. It is thus important to protect these dispersal routes so we can maintain these marine ecosystems and ecosystem services.
Maintaining movement dynamics in marine seascapes involves protecting and restoring marine corridors, estuarine linkages, and coastal habitat linkages.
There are three main strategies to maintain and restore movement dynamics of marine seascapes. First, marine corridors—zones used by whales and other marine species to move between feeding and breeding grounds—should be protected. Marine biologists in several countries successfully reduced collisions between whales and ocean-faring vessels with minor adjustments to shipping lanes that previously crossed marine corridors (Silber et al., 2012). Second, estuarine linkages should be protected, and restored where needed. For example, biologists in South Africa restored the natural flow regime of the St Lucia Estuary, Africa’s largest estuarine lake, by removing dredge spoil in the estuary mouth (Nunes et al., 2018). Third, coastal habitat linkages—beaches and littoral shallows used by wildlife for dispersal, breeding, and feeding—need to be maintained. Studies from South Africa have highlighted how poor protection of connectivity pathways between coastal habitats can compromise these areas’ high levels of species richness and endemism (von der Heyden, 2009; Harris et al., 2014).
Mimicking connectivity
In the absence of habitat linkages, wildlife managers may be able to mimic dispersal dynamics by sporadically translocating a few individuals between subpopulations. Managing populations in this way may be a good alternative in cases where areas earmarked for translocations are too small to sustain a single viable population. Such is the case in South Africa, where conservation biologists occasionally move threatened predators within a small and fragmented protected areas network, where none of the areas are large enough to host a viable population on their own (see Box 8.3). Managing isolated and small populations so intensively nearly always requires sound underlying principles and extensive quantitative analyses (Chapter 9) for guidance.
Management considerations in connectivity conservation
While intuitively appealing, there are a few potential drawbacks to connectivity that conservation planners should consider when planning to establish new habitat linkages (reviewed in Haddad et al., 2014). Prominently, connecting historically isolated populations may lead to outbreeding depression, for example when populations with different local adaptations are connected. Habitat linkages may also act as bottlenecks that expose dispersing animals to greater risks of predation and enable pests and diseases to spread easier. Care must be taken to ensure that wildlife do indeed perceive the landscape “connected”; a habitat linkage that may look good to the human eye may in fact be perceived as inhospitable habitat to wildlife (Newmark, 2008). A recent study from the Americas has shown that the habitat quality of a single stepping stone habitat can determine whether a migration is successful or not (Gómez et al., 2017).
Although the benefits for reconnecting fragmented landscapes generally outweigh the drawbacks, it is important to carefully plan to avoid those drawbacks.
Although the benefits for connecting landscapes for conservation generally outweigh the drawbacks (Haddad et al., 2014), it is important to carefully plan to avoid those drawbacks. Genetic studies can be useful in both determining connectivity among populations (von der Heyden, 2009; Godley et al., 2010) and help researchers detecting potential deleterious factors, such as outbreeding depression (Figure 11.11, see also Frankham et al., 2011; Ralls et al., 2018). Modelling approaches that combine a target species’ movement limitations with radio tracking technologies (e.g. Godley et al., 2010) or remotely sensed environmental variables (e.g. Wegmann et al., 2014) could help to estimate whether a landscape is indeed connected. Much effort has also been invested in finding the optimal width of habitat corridors. For example, one study in lowland forests suggested that corridors that are 30–40 m wide might be adequate for migration of most species while corridors that are 200 m wide will be adequate for all species (Laurance and Laurance, 1999). This is useful guidance, but ecosystems vary, as do target species (Wilson et al., 2010; Pryke and Samways, 2012) and, thus, some corridors may need to be even wider. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/11%3A_Preventing_Extinctions/11.03%3A_Managing_and_Facilitating_Movement_Dynamics.txt |
Earth’s temperature is well on its way to exceed the 2°C increase cap set by global authorities in 2016 (Paris Agreement, Section 12.2.1). Many species that need to adapt to these changes are unable to do so, either because of their limited dispersal capabilities or because of human-induced habitat fragmentation (Section 6.3.5). Others that can disperse may risk decoupling of important symbiotic relationships, as the species involved may not disperse at the same speed, or the same distance (Section 6.3.2). While slowing habitat loss could slow the overall impacts of climate change (Section 10.4), preventing the extinction of many climate-sensitive species will require a range of pro-active conservation management strategies that allow species to adapt at their own pace as and when needed.
Preventing the extinction of climate-sensitive species will require a range of pro-active conservation strategies that allow those species to adapt at their own pace as and when needed.
One of the most important strategies for protecting climate-sensitive species is to identify and protect their likely future habitats. This task of predicting where suitable habitats may be found in future is generally accomplished by identifying and projecting a species’ climatic niche (or bioclimatic envelope) using species distribution models (SDM, Pearson and Dawson, 2003). Section 11.1.1 described how SDM use location data overlaid onto environmental variables to estimate a species’ environmental niche, and how this information can then be used to predict where else a species may occur in a landscape. A similar strategy is followed when predicting a species’ future climate-adapted range. Here, location data are overlaid onto present-day climate variables (e.g. average temperature and rainfall) to define the species’ climatic niche; these niche limits are then projected onto the landscape of interest using future climate scenarios (Section 6.2). Much effort has also been made in recent years to incorporate aspects, such as physiology (Kearney and Porter, 2009) and biological interactions (e.g. Araújo and Luoto, 2007), in predicting future ranges.
Once future ranges have been identified, the next task is to recognize and protect/restore critical dispersal pathways (Section 11.3). While a general strategy of increasing ecosystem-wide connectivity will certainty also benefit climate-sensitive species, conservationists could specifically target climate adaption, by maintaining and restoring climate corridors—dispersal pathways between the current and future ranges (Mawdsley et al., 2009). Several efforts (e.g. Williams et al., 2005; Phillips et al., 2008; Ayebare et al., 2013) are currently underway to establish and protect species-specific and community-specific climate corridors, as predicted using advanced distribution modelling techniques. These and other studies have shown that likely climate corridors often include north–south river valleys, ridges, and coastlines to facilitate poleward distribution shifts, while habitat linkages that cross gradients of elevation, rainfall, and soil types will help climate adaptation across more complex landscapes.
Species with dispersal limitations and specialised interactions may not always benefit from increased connectivity. Instead, those species may rely on climate refuges—areas that are resilient to climate change and thus able to continue to support climate-sensitive communities in future. Africa offers two good examples that illustrate how climate refuges can be identified. The first study, on South African birds, identified climate refuges as areas where temperatures seldom rise above the threshold known to negatively impact a specific species’ fitness (Cunningham et al., 2013). The second study, on northern Mozambique’s coral reefs (McClanahan and Muthiga, 2017), identified two kinds of climate refuges: (a) areas where temperatures never reached a point where it would kill the corals, and (b) areas situated in deeper and cooler water but with the full spectrum of light, which allowed corals to thrive while avoiding heat stress. Both these studies highlight why protecting and restoring complex natural ecosystems (see also Betts et al. 2018) is so important for climate change mitigation.
Climate-sensitive species that are dispersal-limited may not benefit from increased connectivity. Instead, they will rely on climate refuges—areas that are resilient to climate change.
Assisted colonisation is an alternative conservation strategy to save species with dispersal limitations and specialised interactions. Also called assisted migration, assisted colonisation involves the pro-active translocation of climate-sensitive species from their present ranges to their future ranges. Sometimes, even species able to self-disperse may require assisted colonisation. For example, African penguins (Spheniscus demersus, EN) are currently undergoing population declines because of climate change-induced shifts in fish populations on which they depend for food (Sherley et al., 2017). To re-establish this important biological interaction, conservationists are currently using assisted colonisation to establish two new penguin colonies further east from existing colonies (Birdlife South Africa, 2019), in an area where fish populations have remained healthy (Figure 11.12).
As with any translocation project, introducing climate-sensitive species to new areas carries significant risks, including decoupling them from critical limiting resources and symbiotic relationships. It is thus imperative to start small, by translocating only a few well-monitored individuals. If monitoring shows that the initial releases were successful, one can then plan for further releases over time. Because this strategy is still new, it is also important to disseminate your experiences to the broader conservation community, for example by presenting results at conferences or in scientific journals. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/11%3A_Preventing_Extinctions/11.04%3A_Managing_Species_Sensitive_to_Climate_Change.txt |
The best strategy for protecting biodiversity over the long term is to protect existing wild populations in their natural ecosystems. This strategy, known as on-site, or in situ conservation, not only protect entire ecological communities—including thousands of species and their interactions—but also natural processes and ecosystem services. However, if the last populations of a threatened species are too small to remain viable, if they continue to decline despite conservation efforts, or if their threats do not subside, then in situ conservation may prove ineffective. In such cases, sometimes the only option left to prevent an imminent extinction is to capture those last remaining individuals and transfer them to a facility where they can be cared for under artificial, human-controlled conditions. This strategy is known as off-site, or ex situ conservation, and may involve individuals that were collected in the wild, orphaned, confiscated, or displaced and have nowhere else to go. Thanks to ex situ efforts, several African plants and animals that are extinct in the wild continue to survive in zoos, aquaria, and botanical gardens. Examples include four to seven species of ancient cycads (Encephalartos spp.) from Southern Africa, and the pygmy Rwandan water lily (Nymphaea thermarum, EW), which is the world’s smallest water lily (IUCN, 2019).
For species facing imminent extinction, sometimes the only option left may be to capture the remaining individuals and transfer them to captivity.
Ex situ and in situ conservation are complementary strategies (Figure 11.13; see also Conde et al., 2011). For example, many ex situ conservation programs aim to raise enough healthy individuals to support translocation projects when appropriate habitats are available. Ex situ conservation efforts were instrumental in preventing the extinction of the live-bearing Kihansi spray toad (Nectophrynoides asperginis, EW). Populations of this Tanzanian endemic declined precipitously following the establishment of a hydroelectric dam, which caused the toad’s waterfall spray-zone habitat to dry up. The species was subsequently declared Extinct in the Wild in 2009. Tanzanian conservationists, however, demonstrated good foresight by inviting zoos from the USA to collect adults for a captive breeding effort even before the dam was built. This effort is now yielding positive results: after a decade of captive breeding, the erection of an artificial sprinkler system for habitat restoration, and experimental releases (Vandvik et al., 2014), nearly 10,000 toads were released to their former range in May 2018 (Anon, 2018).
Safeguarding a well-represented sample of the world’s biodiversity play only a small role in ex situ conservation efforts. Maintaining self-sustaining wildlife populations under human care not only reduce the need to collect individuals for research from the wild; it also allows researchers to study aspects such as physiology, genetics, and demographics of threatened species (Conde et al., 2019) using methods that might not be possible without animals in captivity. These studies can then provide knowledge and experience to help protect both ex situ and in situ populations. For example, the establishment of the Demographic Species Knowledge Index (Conde et al., 2019), summarise demographic data obtained from ex situ conservation facilities, play a crucial role in filling gaps in datasets for population viability analyses (Section 9.2) Ex situ facilities also play a critical role in captive breeding, head-starting, public outreach, education, and fundraising for in situ conservation. Many ex situ facilities have also become directly involved—and sometimes even taking leading roles—in field conservation efforts (Wilson et al., 2019). Lastly, many ex situ facilities directly connect conservation to social and economic progress through off-site education, employment, and implementation of a range of different community development activities (Ferrie et al., 2013).
Recent efforts to increase knowledge transfer among ex situ facilities has greatly enhanced their contribution to overall conservation efforts. Facilitated by organizations such as the IUCN’s Conservation Planning Specialist Group (CPSG), ex situ facilities now regularly share information on best practices for care and handling of species in human care, including aspects such as nutritional requirements, optimal housing conditions, and veterinary techniques to anaesthetize, immobilise, and reduce stress for animals when they are being moved or during medical treatments (see http://www.cpsg.org). Much of this information is stored in a central database called the Zoological Information Management System (ZIMS). Maintained by Species360, ZIMS keeps track of animal husbandry, medical, and breeding information on over 6.8 million animals belonging to more than 21,000 species for over 1,000 member institutions in 90 countries. Ex situ facilities that maintain these records and comply with operations standards in animal welfare, conservation, education, and research can also apply to become an accredited institution with the Pan-African Association for Zoos and Aquaria (PAAZA), or its parent organization, the World Association of Zoos and Aquariums (WAZA). As of mid-2019, four Sub-Saharan African ex situ facilities were accredited by WAZA, and 19 by PAAZA.
Types of ex situ facilities
Many types of facilities help to preserve ex situ populations. Here we describe some of the most common, including zoos and aquaria for animals, and botanical gardens and seed banks for plants.
Ex situ conservation facilities compliment field conservation efforts through captive breeding, public outreach, education, knowledge generation, and fundraising.
Zoos around the world currently contribute to the conservation of nearly 7,000 species of terrestrial vertebrates (mammals, birds, reptiles, and amphibians) by caring for more than 500,000 individual animals. They do not do this alone; they often work with government agencies, universities, and a variety of other organizations who use zoo animals for research, education, and other conservation activities. While zoos traditionally focussed on displaying charismatic animals that draw visitors, many zoos are now also investing in the conservation of small threatened vertebrates, as well as invertebrates, such as butterflies, beetles, dragonflies, spiders, and molluscs (many of which are also cheaper to maintain). South Africa’s National Zoological Gardens, which houses more than 9,000 individual animals belonging to 705 species, is Africa’s largest zoo by variety of captive species and individuals. The zoo also hosts a variety of daily school programs meant to inspire kids to a career in conservation; these include holiday courses, a zoo club, and guided tours at night.
Aquaria are the aquatic version of zoos, specialised in caring, displaying, and conserving marine and freshwater biodiversity, such as fishes, corals, molluscs, and crustaceans (Figure 11.14). One such institution is South Africa’s uShaka Marine World, the world’s fifth largest aquarium and home to more than 390 marine species—most from the Western Indian Ocean—held in 11 million litres of seawater. Most organisms currently in aquaria have been obtained from the wild, but conservationists are constantly refining techniques to breed more species in captivity to limit wild collecting. Recent and dramatic increases in aquaculture, which currently accounts for roughly a third of fish and shellfish production globally, have made ex situ conservation of aquatic species even more important. The hope is that these ex situ populations will help maintain genetic stocks and act as insurances against disease outbreaks introduced by domestic fish, molluscs, and crustaceans.
Botanical gardens (and arboretums, which specialise on trees and other woody plants) are dedicated to the collection, cultivation and educational curation of living plant species. Botanical gardens across the world house more than 6 million living plants, representing over 80,000 species—approximately 25% of the world’s vascular flora (Wyse Jackson, 2001). The world’s oldest and largest botanical garden—the Royal Botanic Gardens in London, UK—maintains over 28,000 plant taxa, nearly 10% of plant taxa in the world. In Sub-Saharan Africa, there are at least 153 botanical gardens in 33 countries, which range from small community-organised centres to world-famous conservation hubs, such as South Africa’s Kirstenbosch Botanical Garden. Like zoos and aquaria, botanical gardens play a critical role in conservation efforts through public outreach and education. For example, Ghana’s Aburi Botanical Garden established a model medicinal plant garden where the public can gain first-hand knowledge on how to combine conservation, cultivation, and sustainable use of medicinal plants (Gillett et al. 2002).
A few botanical gardens and research institutes have developed collections of seeds, known as seed banks, which take advantage of the fact that seeds of most plants can survive for long periods when stored in cold, dry conditions. The seeds deposited in seed banks may be obtained from the wild, or from cultivated specimens. When gathering material from the wild, botanists generally target populations from across a species’ geographical and habitat ranges so their collections can capture as much of each species’ genetic diversity as possible. In this way, seed banks play a crucial role not only in conservation of plant species richness, but also genetic diversity. Seed banks may even be the only means some plant species are protected. Because many seeds of each species are usually collected, seed banks also provide a convenient opportunity for translocation projects. That is because safeguarded seed collections can be used to propagate not just large numbers of seedlings but, in some cases, custom-developed genetic mixtures to maximise local adaptations. The world’s largest and most diverse seed bank is the Millennium Seed Bank, UK. At the end of 2018, the Millennium Seed Bank catalogued over 2.25 billion seeds from over 39,000 species; its billionth seed, from an African bamboo, was deposited in April 2007. In addition to safeguarding a portion of plant diversity, the Millennium Seed Bank has also benefitted countries, such as Botswana, Burkina Faso, and Mali through the redistribution of banked seeds to aid ecological restoration efforts.
Seed banks contribute to conservation of genetic diversity of plants by collecting material across target species’ geographical and habitat ranges.
Challenges facing ex situ facilities
While the contribution of ex situ conservation facilities to overall biodiversity conservation strategies is significant (Conde et al., 2011), there are some drawbacks that need to be considered. For example, due to the limited number of individuals that can be maintained under human care, especially for larger animals, there is an increased risk that captive populations may suffer from threats facing small populations, such as inbreeding depression and demographic stochasticity (Section 8.7). There is also a concern that ex situ conservation can contribute to hybridisation concerns, for example if different cryptic species are accidentally managed as a single species. To avoid these threats, many ex-situ facilities manage their captive populations jointly as a single interbreeding metapopulation. They do this through studbooks which track the origin, pedigree, and demographic history of each individual in participating facilities. By maintaining and referring to these studbooks, ex situ conservation facilities can make informed decisions regarding transfer and breeding recommendations. The establishment of a European studbook for African dwarf crocodiles (Osteolaemus spp.) even addressed concerns about potential hybridisation between cryptic species (Schmidt et al., 2015).
Ex-situ facilities often manage captive populations as a single metapopulation using studbooks to track the origin and demographic history of breeding individuals.
Funding also remains an obstacle, given that ex situ facilities typically require large, long-term, funding commitments, in comparison to many in situ conservation activities. One consequence of funding limitations is that ex situ facilities mostly focus on showy or charismatic species that attract visitors, so small and less charismatic species are not always afforded equal protection (Brooks et al., 2009). Many ex situ facilities are also more inclined to house non-threatened species that are easier and less costly to care for, rather than threatened species with specialised needs (Table 11.1). For example, despite the fear of looming mass amphibian extinctions due to a disease caused by the chytrid fungus (Batrachochytrium dendrobatidis) (Alroy, 2015), 75% of ex situ amphibian collections consist of non-threatened species, with only 6.2% of all threatened amphibians afforded ex situ protection (Dawson et al., 2016). Neglecting threatened species in ex situ conservation efforts also creates a feedback loop, by maintaining a limited understanding on how to care for the species most in need.
Table 11.1 Number and percentages of terrestrial vertebrate species from Sub-Saharan Africa currently maintained in the world’s ex situ facilities. Values in parenthesis represent percentage of all speciesa, threatened speciesb, and CITES-listed speciesc for each taxon class, respectively.
Mammals
Birds
Reptiles
Amphibians
Total
Worldwidea
659 (55%)
1,470 (65%)
197 (27%)
44 (5%)
2,370 (47%)
Africa
110 (9%)
234 (10%)
34 (4%)
6 (1%)
384 (8%)
Asia
136 (11%)
327 (14%)
22 (3%)
2 (0%)
487 (10%)
Oceania
37 (3%)
61 (3%)
6 (1%)
1 (0%)
105 (2%)
Europe
191 (12%)
465 (20%)
73 (10%)
19 (2%)
748 (15%)
North America
145 (12%)
311 (14%)
53 (7%)
14 (2%)
523 (10%)
South America
40 (3%)
72 (3%)
197 (27%)
44 (5%)
353 (7%)
Threatened speciesb
45 (23%)
42 (20%)
22 (21%)
8 (4%)
117 (16%)
Extinct in the Wild
1 (100%)
0 (0%)
0 (0%)
1 (100%)
2 (100%)
Critically Endangered
7 (26%)
4 (19%)
5 (25%)
3 (5%)
19 (15%)
Endangered
13 (16%)
12 (15%)
2 (5%)
3 (3%)
30 (10%)
Vulnerable
24 (27%)
26 (23%)
15 (33%)
1 (2%)
66 (23%)
CITES-listed speciesc
95 (50%)
121 (62%)
45 (25%)
1 (6%)
262 (45%)
Appendix I species
30 (58%)
4 (44%)
8 (80%)
1 (6%)
43 (49%)
Appendix II species
58 (44%)
112 (62%)
37 (22%)
0 (0%)
207 (43%)
Appendix III species
7 (100%)
5 (100%)
0 (0%)
0 (0%)
12 (100%)
Source: https://zims.species360.org, current as of April-2019. Compiled by Johanna Staerk (Species360).
Fortunately, ex situ facilities have responded to these concerns by developing several innovative mechanisms that enables them to contribute more to the conservation of threatened species. For example, ex situ facilities all agree that attracting more visitors attracts more funding. To attract more visitors, zoos and aquaria are increasingly keeping animals in enclosures that are representative of their natural environments; this keeps the animals heathier and providing more opportunities to exhibit natural behaviors which, in turn, leave visitors more satisfied. Some zoos and aquaria have also established special displays where visitors can feed, touch, or otherwise interact with animals. Many ex situ facilities have also started inviting local artists to display sculptures and other artwork, which adds to the experience for visitors and attracting people that might not otherwise have visited. A rather unusual—but very successful—attempt to increase foot traffic comes from the USA, where the California Academy of Sciences hosts dance parties with laser shows, food, and drinks every Thursday night (http://www.calacademy.org/nightlife), which visitors can enjoy while visiting the Academy’s aquarium and other conservation exhibits.
While the contribution of ex situ facilities to species conservation is significant, many rare species are ill-suited for ex situ efforts. Some species simply do not adapt or reproduce in captivity, while others that do relatively well in captivity experience behavioral and physiological changes or acquire diseases (Brossy et al., 1999) that prevent releases in the wild. Even so, the conservation biologists working at ex situ facilities constantly try to find ways to overcome these challenges. For example, staff at ex situ facilities sometimes use assisted reproductive techniques such as artificial incubation of bird and reptile eggs, or artificial insemination (Box 11.4) to overcome reproductive challenges (e.g. if individuals cannot mate because they are in different locations). Others use cryopreservation and genome resource banks for the long-term storage of embryos, eggs, sperm, or purified DNA, at least until those tissues can be used to increase a species’ genetic diversity, or perhaps even to resurrect an extinct species (see de-extinction, Section 8.8). However, many ex situ conservation techniques are difficult and expensive to implement. When possible, it is almost always preferable to preserve species in situ where they can be self-sustaining, free from inbreeding, and an interactive participant of their community and ecosystem.
Box 11.4 Saving the Northern White Rhinoceros with Assisted Reproduction Technologies
Morné de la Rey
Embryo Plus,
Brits, South Africa.
http://www.embryoplus.com
A few centuries ago, Earth’s wildernesses enabled animals to roam and breed relatively freely. Today, sprawling cities, agriculture, and fences not only restrict animals’ ability to forage, but also limit reproduction between differing gene pools. These stresses create smaller and more isolated populations which are being edged toward extinction.
There are several landscape-scale conservation initiatives to counteract these imbalances. But some species and populations are so rare that they depend on intensive management to remain viable. Assisted reproductive techniques (ART) provide promise for helping such species. Over the past 30 years, ARTs have greatly enhanced how the livestock industry preserve, improve, and proliferate genetic stock. Now, efforts are also underway to use ARTs to ensure the preservation of biodiversity.
The many types of ARTs
ARTs include a wide array of medical procedures to address infertility, and to make reproduction possible between individuals unable to do so naturally (e.g. animals in different protected areas). In this way, biologists can ensure genetic exchange while eliminating the risks inherent in translocation such as spread of diseases, adaptation to new environments, and disruption of group dynamics.
ARTs have various levels from relatively simple to very complex. The most basic technique is artificial insemination. A major advantage of this technique is that it can multiply male genetic contributions by inseminating more females than would be possible in nature. Much progress has also been made in improving viability of cryopreserved semen to overcome challenges with timing of female reproductive cycles and other logistical constraints.
As for multiplying female genetic contributions, methods involve embryo transfer and in vitro fertilisation (IVF). With multiple ovulation embryo transfer (MOET) egg fertilisation occurs naturally; with IVF, it occurs in a laboratory incubator. In both cases, an embryo is eventually transferred to a surrogate mother which will carry it until birth. Scientists are currently working on improving viability of stored germplasm, so that embryos can be cryopreserved until a suitable surrogate mother is ready.
The third technique is nuclear transfer, also known as cloning. This very delicate procedure involves replacing the haploid DNA of an unfertilised egg with diploid DNA of another; cells are then cultured, after which the embryo is transferred to a surrogate mother.
Using ARTs to save the northern white rhino
Once widespread across Central Africa, poaching has pushed the northern white rhinoceros (Ceratotherium simum cottoni, EW) to the brink of extinction. Today, only two females remain, both in a semi-captive setting at Ol Pejeta Conservancy, Kenya. Incapable of natural reproduction, this species is committed to extinction without drastic intervention.
A cutting-edge initiative is currently underway to use ARTs to save this iconic species. While the project’s exact trajectory is still being developed, likely steps include optimising procedures for harvesting, maturing, and fertilising eggs, followed by embryo transfer into surrogate southern white rhinoceros. Some preliminary successes have also been achieved to generate stem cells from skin biopsies (Ben-Nun et al., 2011), which could be used in cloning. Genetic material (tissue samples and semen) of several northern white rhinos has been cryopreserved at various places around the world. However, there is a limited amount of sperm available (there are no males left), and so artificial insemination and IVF with northern white rhinoceros depends on embryo transfer successes. Many partners have been assembled to pool resources and ideas in support of this initiative, including Ol Pejeta Conservancy, Embryo Plus, Fauna & Flora International (FFI), Back to Africa, Dvur Karlove Zoo, Leibniz-IZW, Avantea, San Diego Zoo, and Kenya Wildlife Service.
Refining ARTs on other species
Before ARTs are implemented on the near-extinct rhinoceros, it is advisable to optimise procedures on another species. A logical choice would be the closely related southern white rhinoceros (C. simum simum, NT). However, the southern subspecies is also threatened, so we should look for more common mammals first.
Veterinarians at Embryo Plus routinely perform ARTs on domestic cattle, so efforts are currently focussed on building from this experience to work with wild bovines (Figure 11.D). For example, Embryo Plus recently produced the world’s first African buffalo (Syncerus caffer, NT) through IVF; the healthy calf named Pumelelo (meaning success in isiZulu) was born in June 2016. Embryo Plus has also produced several western Zambian sables (Hippotragus niger kirkii) from southern sable (H. niger niger) surrogates using embryo transfer. Plans are also underway to investigate the viability of using eland (Tautragus oryx, LC) and domestic horses (Equus ferus) as surrogate mothers for mountain bongo (T. eurycerus isaaci, CR) and Grevy’s zebra (E. grevyi, EN), respectively.
From dream to reality
The long-term objective of the northern white rhinoceros project is to establish a viable breeding herd which can be reintroduced into secure habitats. But much work remains for this dream to become reality. While there was one successful attempt in producing a healthy bongo calf by transferring an embryo to an eland mother (Woolf, 1986), inter-species embryo transfer remains challenging. Due to a rhinoceros’ size, we also need to ensure ART procedures can be performed safely without placing undue stress on the patient. Lastly, because each species’ embryos have different requirements in the laboratory, extensive research is necessary before ARTs can be attempted on a new species.
Although ARTs in wildlife management is still in its infancy, we are confident that early breakthroughs hold promise for the survival of the northern white rhinoceros, as well as other threatened species that may one day benefit from these procedures. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/11%3A_Preventing_Extinctions/11.05%3A_Ex_Situ_Conservation_Strategies.txt |
Most of today’s species-centric conservation initiatives are biased towards species that are showy, charismatic, or economically important. Consequently, conservation efforts for the vast majority of taxa are neglected, particularly in Africa where conservation funding is often more limited than elsewhere. One well-known example is known as plant blindness, the perception that animals take precedence above plants in conservation efforts. This isn’t just a case of hurt feelings among botanists: there are likely significantly more plant than animal species that should be considered as threatened (see Table 2.1); however, thorough threat assessments are hampered because, as a group, plants receive significantly less funding compared to animals (Negron-Ortiz, 2014). One explanation for this disparity is that plants are often seen as the backdrop of the environment rather than the critical foundation (as primary producers) of every food web on Earth. While showy plant species may indeed have highly visible roles in maintaining the environment and regional economies, neglected species may play an equally—sometimes even more—important role in maintaining ecosystems and ecosystem services (Schleuning et al., 2016).
Fortunately, the number of professional and amateur societies interested in protecting neglected taxa, such as reptiles, amphibians, invertebrates, fungi, and plants are rising. Some groups of experts are also organised into Specialist Groups (https://www.iucn.org/ssc-groups) by the IUCN. These societies and expert groups highlight the plight of neglected taxa and are willing to provide in-house expertise on best practices for protecting those species.
11.07: Summary
1. A species may be threatened by a combination of many factors, all of which must be addressed in a comprehensive conservation plan that considers its natural history.
2. New populations of threatened species can be established in the wild using either captive-raised or wild-caught individuals. Animals used in translocation projects sometimes require special care and behavioral training before release as well as care and monitoring after release.
3. Maintaining and facilitating movement dynamics is very important for protecting wildlife in their natural ecosystems. To do this, connectivity must be preserved by ensuring that habitat linkages such as wildlife corridors and stepping stone habitats that are intact, functional, and free from human-made obstacles.
4. Preventing biodiversity losses under climate change requires ecosystem preservation, maintaining and restoring climate corridors and refugia, and assisted colonisation for species unable to adapt their ranges quick enough.
5. Some species that are in danger of going extinct in the wild can be maintained in zoos, aquaria, botanical gardens, and seed banks; this strategy is known as ex situ conservation. Ex situ conservation contributes to field conservation through research, skills development, public outreach, conservation education, fundraising, captive breeding, and head-starting.
11.08: Topics for Discussion
1. How do you judge whether a reintroduction project is successful? Develop simple and then increasingly detailed criteria to evaluate a project’s success.
2. Tying concepts from different chapters together, what are the biggest challenges standing in the way of conserving Africa’s migratory birds?
3. Use the advanced search functions on the IUCN Red List website (https://www.iucnredlist.org) to pick one species occurring in your country that is threatened by climate change. Referring to Chapter 6, how does climate change threaten this species? What strategies can be used to prevent this species’ extinction?
4. What roles do ex situ facilities play in the conservation of threatened species in Africa? Discuss two or three different roles. Do you think there are certain aspects in conservation that they can make a larger contribution to than is currently the case?
5. Find two or three examples of wild or semi-wild populations of African species maintained on other continents? Are those species threatened in their natural distribution ranges? Does maintaining populations of African species on other continents represent a successful conservation strategy? Explain your answer.
11.09: Suggested Readings
Hoffmann, M., J.W. Duckworth, K. Holmes, et al. 2015. The difference conservation makes to extinction risk of the world’s ungulates. Conservation Biology 29: 1303–13. https://doi.org/10.1111/cobi.12519 Conservation action does make a difference, as illustrated by ungulate conservation.
IUCN/SSC. 2013. Guidelines for Reintroductions and Other Conservation Translocations (Gland: IUCN/SSC). https://portals.iucn.org/library/efiles/documents/2013-009.pdf Guidelines for species reintroductions and translocations.
Mawdsley, J.R., R. O’Malley, and D.S. Ojima. 2009. A review of climate-change adaptation strategies for wildlife management and biodiversity conservation. Conservation Biology 23: 1080–89. https://doi.org/10.1111/j.1523-1739.2009.01264.x A review of methods to combat climate change.
Menges, E.S., S.A. Smith, and C.W. Weekley. 2016. Adaptive introductions: How multiple experiments and comparisons to wild populations provide insights into requirements for long-term introduction success of an endangered shrub. Plant Diversity 38: 238–46. https://doi.org/10.1016/j.pld.2016.09.004 Experimentation can reduce uncertainty in population translocation.
Miller, B., W. Conway, R.P. Reading, et al. 2004. Evaluating the conservation mission of zoos, aquariums, botanical gardens, and natural history museums. Conservation Biology 18: 86–93. https://doi.org/10.1111/j.1523-1739.2004.00181.x Eight tough questions directed at ex situ facilities.
Nogués-Bravo, D., D. Simberloff, C. Rahbek, et al. 2016. Rewilding is the new Pandora’s box in conservation. Current Biology 26: R87–R91. https://doi.org/10.1016/j.cub.2015.12.044 Are there limits to wildlife translocation dreams?
Pérez, I., J.D. Anadón, M. Díaz, et al. 2012. What is wrong with current translocations? A review and a decisionmaking proposal. Frontiers in Ecology and the Environment 10: 494–501. https://doi.org/10.1890/110175 Important considerations for translocations.
Sherley, R.B., K. Ludynia, B.M. Dyer, et al. 2017. Metapopulation tracking juvenile penguins reveals an ecosystem-wide ecological trap. Current Biology 27: 563–68. https://doi.org/10.1016/j.cub.2016.12.054 Detecting and dealing with ecological traps requires holistic thinking. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/11%3A_Preventing_Extinctions/11.06%3A_Thoughts_on_Neglected_Taxa.txt |
Ali, A.H., A.T. Ford, J.S. Evans, et al. 2017. Resource selection and landscape change reveal mechanisms suppressing population recovery for the world’s most endangered antelope. Journal of Applied Ecology 54: 1720–29. https://doi.org/10.1111/1365-2664.12856
Alroy, J. 2015. Current extinction rates of reptiles and amphibians. Proceedings of the National Academy of Sciences 112: 13003–08. https://doi.org/10.1073/pnas.1508681112
Andersson, J., M. de Garine-Wichatitsky, D. Cumming, et al. 2013. Transfrontier Conservation Areas: People Living on the Edge (New York: Routledge).
Anon. 2018. 9,873 Kihansi spray toads return to their environment. IPP Media. https://go.shr.lc/2B6N7wB
Araújo, M.B., and M. Luoto. 2007. The importance of biotic interactions for modelling species distributions under climate change. Global Ecology and Biogeography 16: 743–53. https://doi.org/10.1111/j.1466-8238.2007.00359.x
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Wilson, J.W., R. Bergl, L.J. Minter, et al. 2019. The African elephant Loxodonta spp. conservation programs of North Carolina Zoo: Two decades of using emerging technologies to advance in situ conservation efforts. International Zoo Yearbook 53: in press. https://doi.org/10.1111/izy.12216
Wilson, J.W., R.L. Stirnemann, Z.S. Schaikh, et al. 2010. The response of small mammals to natural and human-altered edges associated with Afromontane forests of South Africa. Forest Ecology and Management 259: 926–31. http://doi.org/10.1016/j.foreco.2009.11.032
Wimberger, K., C.T. Downs, and M.R. Perrin. 2009. Two unsuccessful reintroduction attempts of rock hyraxes (Procavia capensis) into a reserve in the KwaZulu-Natal Province, South Africa. South African Journal of Wildlife Research 39: 192–201. https://doi.org/10.3957/056.039.0213
Wimberger, K., C.T. Downs, and R.S. Boyes. 2010. A survey of wildlife rehabilitation in South Africa: Is there a need for improved management? Animal Welfare 19: 481–99.
Winemiller, K.O., P.B. McIntyre, L. Castello, et al. 2016. Balancing hydropower and biodiversity in the Amazon, Congo, and Mekong. Science 351: 128–29. http://doi.org/10.1126/science.aac7082
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The negative impact of human activities on the natural environment is apparent wherever you look. Some impacts are an unavoidable consequence of human activities; vast resources are currently invested in finding ways to mitigate those impacts. Other impacts, often entirely preventable, are rooted in greed. Consider how the worst polluters are corporations that prioritise profits over environmental and human health. Similarly, many threatened species continue to be illegally exploited in an unsustainable manner; in the worst cases, the profits from poaching are funding human-rights atrocities and organised criminal networks. Because society pays the price for environmental crimes—which generally benefit only a few people—there is broad interest in preventing environmental abuse, and to punish the perpetrators.
Because society pays the price for environmental crimes—which benefit only a few people—there is broad interest in preventing environmental abuse, and to punish the perpetrators.
Environmental crimes are generally divided into two categories: wildlife crimes—the illegal exploitation of biodiversity (including but not restricted to wildlife trafficking and biopiracy), and pollution crimes—the illegal trade and disposal of waste and hazardous substances. As with other crimes, environmental crimes are generally defined by legislative action, when governments pass environmental laws and regulations that restrict certain kinds of activities. The effectiveness of these laws and regulations in protecting the environment relies on three main factors: (1) identifying conservation priorities, (2) establishing regulations that addresses those needs, and (3) enforcing environmental laws and regulations.
12.01: Identifying Legislative Priorities
Humans have always depended on the environment to fulfil their most basic needs. Before the Industrial Revolution, fulfilling those basic needs generally occurred at sustainable levels. Over the last few centuries, however, exponential human population growth and rates of resource extraction have put enormous pressure on the environment. Today, many wildlife populations and ecosystems are unable to cope with these pressures. Increased globalisation has exacerbated many of these problems. For example, with most Asian rhinoceros and pangolin populations on the brink of extinction (IUCN, 2019), Asian traders are increasingly filling their orders for elephant (Figure 12.1), rhinoceros, and pangolin body parts from African suppliers (Biggs et al., 2013; Wasser et al., 2015; Heinrich et al., 2016).
Identifying which species and ecosystems need to be prioritised for legislative action can be confusing, and sometimes even seems in conflict with more readily available information at hand. For example, many hunters believe that the animals they target persist in healthy numbers despite claims to the contrary from conservation biologists. In other areas, logging companies claim they operate sustainably, yet tropical forests continue to shrink. In the face of conflicting information, it is critical for conservation biologists to rely on consistent, repeatable, and transparent methods to identify those populations, species, and ecosystems that may need (additional) regulatory protections.
Currently, the most popular method to identify legislative priorities is to use the IUCN’s Red List criteria, developed to reflect a taxon’s risk of extinction (Section 8.5). Following these criteria (which can be applied on a global or local scale), species that are considered Extinct in the Wild, Critically Endangered, Endangered, and Vulnerable are officially considered “threatened with extinction” and would thus receive higher priority than species that are Near Threatened or Least Concerned.
Although coarse filter approaches, which focus on groups of species and threatened ecosystems (Section 8.5.1), have been a catalyst for many international treaties and protected areas, legal mechanisms at the national and regional level do not always allow for its use. Through lobbying and education, these legislative branches will hopefully improve their receptiveness for coarse filter approaches in setting future legislative agendas. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/12%3A_Biodiversity_and_the_Law/12.00%3A_Prelude_to_Biodiversity_and_the_Law.txt |
When conservation priorities have been identified, there are several options available to preserve biodiversity. One option could involve the establishment of protected areas where ecological restoration (Section 10.3) and species conservation projects (Chapter 11) can be carried out. Conservation biologists could also start an environmental education program (Section 15.5) that would help people live more sustainably on unprotected lands (Chapter 14). Under certain conditions, however, especially when control and protection measures fail, restrictions or outright bans of some human activities may be necessary (Keeley and Scoones, 2014). The most effective restrictions and bans involve legislative actions that also establish mechanisms to enforce environmental laws and regulations, and mechanisms that reduce consumer demand (Challender and MacMillan, 2014).
Environmental laws and regulations are implemented at three different levels: international treaties, national laws, and local laws. While the scope of each of these levels differs, they are intricately connected with one another. International treaties influence national laws, but also depend on their enforcement to succeed, while national laws are guided by local needs as well as customary laws that have been in place for generations. Ideally speaking, international and national laws set minimum benchmarks, which regional and local governments adopt and enforce. Local and regional laws may sometimes set stricter standards in areas where the environment is more sensitive, more damaged, or more important for human well-being. Local and national legislatures may also choose to ignore broader legislation, through non-cooperation and non-enforcement. But this is not advisable as it may lead to further environmental deterioration, loss of funding, and even trade embargoes and sanctions (Section 12.4.4) that could harm local economies.
International agreements
International agreements provide frameworks that allow countries to work together to protect biodiversity (Sands and Peel, 2012). These international agreements, called treaties or conventions, are needed for five important reasons: (1) many species migrate and disperse across administrative borders, (2) ecosystems do not follow administrative boundaries, (3) pollution spreads by air and water across regions and around the globe, (4) many biological products are traded internationally, and (5) some environmental problems (e.g. climate change and pollution) require global cooperation and coordination. To pass international treaties, agreements are negotiated at international conferences under the authority of international bodies such as the UN, UNEP, or IUCN and come into force when they are ratified by an agreed-upon number of countries. These treaties are then implemented at the local level when signatory countries pass national laws to enforce them.
International agreements provide frameworks that allow countries to work together to protect biodiversity.
One of the most important international environmental treaties is the Convention on Biological Diversity (CBD, https://www.cbd.int). The CBD formulated and signed following the UN Earth Summit (also called Rio Summit) held in Rio de Janeiro, Brazil in 1992, has played a major role in raising awareness of the value of biodiversity to humanity. At this meeting, representatives from 178 countries formulated and eventually signed the CBD, obligating signatory countries to protect biodiversity through careful management of nature for the benefit of humans. The CBD was expanded in 2010 to also include recommendations for the protection of IUCN Red Listed species and ecosystems, as part of the Aichi Biodiversity Targets (Table 12.1).
Table 12.1 The UN, with governments across the world, have agreed to work on five strategic goals and 20 specific targets (collectively known as Aichi Biodiversity Targets) to halt the loss of biodiversity and protect and restore what remains.
CBD strategic goal
Aichi Target
A.
Address underlying causes of biodiversity losses
1. Improve awareness of biodiversity values
2. Integrate biodiversity values into development
3. Eliminate perverse subsidies; incentivise sustainability
4. Implement plans for sustainable consumption and production
B.
Reduce pressures on biodiversity
5. Reduce the rate of habitat loss by at least 50%
6. Ensure sustainable use of marine resources
7. Ensure sustainable agriculture, aquaculture, and forestry
8. Reduce pollution to non-detrimental levels
9. Identify and control priority invasive species
10. Reduce pressures on climate-sensitive ecosystems
C.
Safeguard ecosystems, species, and genetic diversity
11. Increase coverage of well-managed protected areas
12. Prevent the extinction of threatened species
13. Prevent genetic erosion of biodiversity
D.
Enable more people to enjoy the benefits of biodiversity
14. Restore and safeguard ecosystems and essential services
15. Restore and enhance resilience of degraded ecosystems
16. Ensure fair and equitable sharing of ecosystem services
E.
Implement participatory biodiversity strategies
17. Implement participatory national biodiversity strategies
18. Respect and conserve traditional knowledge
19. Improve, share, and apply biodiversity knowledge
20. Mobilise resources to address Aichi Targets
There are also several international agreements seeking the direct protection of targeted threatened species. One of the most important treaties of this nature is CITES (Convention on International Trade in Endangered Species of Wild Fauna and Flora, https://cites.org), agreed upon in 1973 in Washington, DC. This treaty, ratified by 175 countries, establishes lists (known as Appendices) of species for which member nations agree to ban, restrict, control, and monitor international trade. Over 35,000 species of plants and animals appear on these appendices, many also listed as threatened by the IUCN. With a few exceptions, the international trade of wild-caught specimens on Appendix I is prohibited; trade in Appendix II species is strictly regulated to ensure sustainability, while trade in Appendix III species require a certification of origin. Once member countries pass local laws to comply with CITES, police, customs inspectors, wildlife officers, and governmental agents appointed for that purpose can arrest individuals possessing or trading in products from the listed species. The World Conservation Monitoring Centre (WCMC), which operates within UNEP, is tasked with managing the CITES database and monitoring whether member countries are enforcing recommendations.
The Convention on the Conservation of Migratory Species of Wild Animals (sometimes shortened to Bonn Convention, http://www.cms.int) is another important treaty that seeks the protection of specifically targeted species. The Bonn Convention came into force in 1983, and has over 120 Parties, including 37 from Sub-Saharan Africa. As with CITES, the Bonn Convention categorises species under Appendices. Species on Appendix I are threatened with extinction; “Range States” to Appendix I species are obliged to afford those species’ strict protections. Appendix II lists species whose populations would significantly benefit from international cooperation. Three important agreements that involve Sub-Saharan species have been concluded under the Bonn Convention: (1) the African-Eurasian Waterbird Agreement (http://www.unep-aewa.org), which, amongst others, things bans the use of lead shot around aquatic ecosystems; (2) the Gorilla Agreement (http://www.cms.int/gorilla), which binds Parties to protect gorillas in their habitats; and the Agreement on the Conservation of Albatrosses and Petrels (https://acap.aq), which coordinates international efforts to mitigate known threats to seabirds.
Several international agreements seek the protection of important ecosystems. Perhaps the most prominent is Convention Concerning the Protection of the World’s Cultural and Natural Heritage (http://whc.unesco.org), which protects natural (and cultural) areas of international significance. As of mid-2019, UNESCO (the organization managing the list of World Heritage Sites) recognized 35 natural World Heritage Sites in Sub-Saharan Africa; this includes some of the world’s most famous conservation areas, such as Serengeti National Park in Tanzania, Bwindi Impenetrable Park in Uganda, and the Aldabra Atoll of the Seychelles. In addition, five World Heritage Sites are recognized for their natural and cultural significance: this includes Gabon’s Ecosystem and Relict Cultural Landscape of Lopé-Okanda, Tanzania’s Ngorongoro Conservation Area, and the Maloti-Drakensberg Park—a transboundary site composed of South Africa’s Drakensberg National Park and Lesotho’s Sehlathebe National Park (Figure 12.2).
Another important treaty that seems ecosystem protection is the Ramsar Convention on Wetlands (http://www.ramsar.org), which recognizes the ecological, scientific, economic, cultural, and recreational value of freshwater, estuarine, and coastal marine ecosystems. All but three Sub-Saharan African countries have signed the Ramsar Convention; this binds each member country to conserve and sustainably utilise its wetlands (particularly those that support migratory waterfowl), and to officially declare at least one internationally significant wetland as protected. As of mid-2019, 252 Sub-Saharan African wetlands, covering over 1 million km2, were declared internationally significant under Ramsar guidelines. South Africa and Burkina Faso have the most Ramsar Wetlands (23 and 22, respectively), while the Republic of the Congo has the largest area (138,138 km2) designated. The world’s largest Ramsar wetland, the DRC’s Ngiri-Tumba-Maidombe, is 65,696 km2 in size (over twice the size of Lesotho!).
International treaties are particularly important to the marine environment, since about two-thirds of the world’s oceans (50% of the planet) fall outside any country’s jurisdiction.
International treaties are particularly important to the marine environment, since about two-thirds of the world’s oceans (50% of the planet) are considered international waters—that is, being outside any country’s exclusive economic zone (EEZ), all states have the freedom to fish, travel, do research, etc. in these areas. Three examples of international agreements protecting such marine ecosystems are (1) the 1972 Convention on the Prevention of Marine Pollution by Dumping of Wastes and Other Matter (http://www.imo.org/en/OurWork/Environment/LCLP) which regulates pollutants into the marine environment, (2) the 1982 Convention on the Law of the Sea (http://www.un.org/Depts/los) which establishes guidelines for management of marine natural resources, and (3) the 2009 Agreement on Port State Measures (http://www.fao.org/port-state-measures) which sanctions monitoring for illegal, unreported, and unregulated fishing at shipping ports.
In addition to being party to these and other global treaties, several African countries are also members of agreements that address regional environmental concerns. Among the most prominent is the 2003 Revised African Convention on the Conservation of Nature and Natural Resources (or Maputo Convention). The most progressive reforms of the Maputo Convention include the recognition that nature is a finite resource, that the needs of future generations and traditional peoples must be considered, and that the harmful impacts of civil strife on the environment must be mitigated.
The 2016 Paris Agreement, which deals with greenhouse gas emission reduction and climate change mitigation, serves to illustrate the difficult political negotiations (Figure 12.3) involved in the adoption of an international treaty. Although the negative effects of climate change have been known for several decades (Section 6.1), until recently there has been a distinct lack of action to curb global greenhouse gas emissions. For example, as an early call to action on reducing greenhouse gas emissions, representatives from 154 countries signed the UN Framework Convention on Climate Change (UNFCCC) at the Earth Summit in May 1992. In the following years, negotiations during annual UNFCCC conferences (formally known as “Conference of the Parties”, or COPs) led in the Kyoto Protocol, adopted in Japan in 1997, which marked the first attempt to set legally binding emission reduction targets. Despite broad appeal among its 192 parties, the Kyoto Protocol faced an uphill battle from the start because the USA (the world’s biggest greenhouse gas emitter at the time) refused to ratify it, and China (which recently overtook the USA as the biggest emitter) was exempted from compliance. While this has left the Kyoto Protocol largely a failure, it provided important lessons that contributed to the successful passing of the Paris Agreement (http://unfccc.int), which was negotiated and adopted through consensus by 195 countries (this time including the USA and China) in December 2015. The Paris Agreement went into effect on 4 November 2016 after the minimum 55 countries ratified it, marking a breakthrough in the decades-long battle to curb global greenhouse gas emissions. By mid-2019, all but one country in the world (the non-signatory being the Holy See, who as UNFCCC observer nation that cannot sign but strongly support the Agreement) have signed and/or ratified the Agreement. Most relevant to African member states are the mechanisms set up to provide developing countries with large amount of aid for climate change mitigation and adaption, much of which involves ecosystem conservation (see REDD+, Sections 15.3).
While it is still too early to judge the effectiveness of the Paris Agreement, the 1987 Montreal Protocol on Substances that Deplete the Ozone Layer (http://ozone.unep.org) illustrates how international cooperation can be effective in preventing environmental disasters. In the 1970s, scientists discovered that a range of chemicals (primarily chlorofluorocarbons, or CFCs) commonly used in agriculture, energy production, and even common household items (such as refrigerators and aerosol spray canisters) were depleting the atmospheric ozone layer. The ozone layer is critical for human life; by cutting the amount of harmful ultraviolet radiation from the sun that reaches the Earth’s surface, protection from the ozone layer reduces skin cancer, cataracts, and crop damage. In response to this threat, the Montreal Protocol aimed to phase out those substances that were responsible for ozone depletion. Since then, the ozone layer has steadily recovered; current projections suggest that the ozone layer will return to 1980 levels in the second half of the 21st century. Towards the end of his tenure as Secretary General of the UN (1997–2006), Ghana’s Kofi Annan declared, “Perhaps the single most successful international agreement to date has been the Montreal Protocol”. The Montreal Protocol’s success is directly due to this widespread adoption and implementation.
National and local laws
Traditional African societies have long recognized that preserving the environment is important for human well-being. Consequently, many African cultures had mechanisms in place before the arrival of European colonists that allowed these historical societies to exploit communal resources on a long-term, sustainable basis. These mechanisms included mystical beliefs, local customs, and cultural taboos that ensured the protection of wildlife and land with cultural and spiritual significance. While sacred forests are prominent examples, not all sacred sites are/were forested. For example, the sandy beaches on Guinea-Bissau’s Poilão Island was also regarded as sacred by the people of the Bijagós Archipelago, ensuring the protection of one of the world’s most important green turtle (Chelonia mydas, EN) nesting sites (Catry et al., 2002). These mechanisms, referred to as customary laws, also limited access to certain territories and imposed restrictions on harvesting methods, harvest times, and types of individuals that may be harvested. Strict sanctions for violations ensured that customary laws were generally followed, often through self-policing. In some ways, this traditional approach to natural resource management was not so different from certain wildlife management systems in Europe and elsewhere at the time—or even from today’s more formal law systems—which place restrictions on how we utilise nature. While some customary laws continue to regulate activities in certain regions of Africa (e.g. Walters et al., 2015), in many areas, they were lost when European authorities replaced traditional authorities during colonisation.
Today, an increasing number of international treaties and environmental organizations are achieving their conservation goals by promoting respect for and inclusion of the cultural and spiritual values that traditional peoples attach to the environment. This includes the CBD, UN, IUCN, and African Union, all promoting the integration of traditional ecological knowledge (TEK) in conservation activities and regulations (Mauro and Hardison, 2000). A growing number of national governments are also institutionalising these efforts by passing laws recognizing traditional rights, providing traditional peoples with land titles, and declaring areas of spiritual and cultural significance as protected. Conservation scientists are also increasingly relying on TEK to better understand ecological networks (Sileshi et al., 2009; Gómez-Baggethun et al., 2013), to ensure sustainable utilisation of natural resources (Mbata et al., 2002; Terer et al., 2012), and to secure the continued survival of severely threatened species such as the Cross River gorilla (Gorilla gorilla diehli, CR), of which fewer than 300 individuals remain (Etiendem et al., 2011).
While governments are becoming increasingly respectful of customary laws and traditional lifestyles, in many areas the customs inherent to them have fallen by the wayside under increased industrialisation, urbanisation, and globalisation. An increasing number of traders of traditional products are also using more effective collection and harvesting techniques, thereby pushing many species to extinction (Section 7.2). To fill these regulatory voids and to ensure sustainable utilisation of natural resources, statuary (passed by legislatures); regulatory (passed by regulatory agencies; and case (passed by judicial bodies) laws are playing an important role in protecting Africa’s natural heritage.
Laws that protect the environment (and which can be passed by local or national branches of government) can generally be divided into three categories:
• Natural resource management laws, which define the limits of fair and sustainable use of land, water, minerals, and biodiversity.
• Pollution laws, which regulate dumping of waste and other harmful substances into the environment.
• Tax incentives, which encourage environmentally responsible behaviors.
Environmental laws that address natural resource use are well known because they impact the activities of the public and some businesses. These include hunting, trapping, and fishing regulations that limit the size and number of animal and plant products that can be collected, and the equipment that can be used for harvesting. Such regulations are typically enforced through licencing requirements, harvest reporting, and law enforcement patrols. Authorities may also set up mechanisms to restrict the sale, transport, and killing of sensitive species, including restricting the sale of firearms and ammunition.
Many people have also been exposed to restrictions that control the ways in which land is used to protect biodiversity. For example, uncontrolled fires may severely damage natural communities, so practices (such as building campfires) that contribute to accidental fires are often rigidly controlled. In some areas, vehicles and even foot traffic may be restricted to protect ecosystems and resources that are sensitive to disturbance, such as bird and turtle nesting areas on beaches, or sources of drinking water. One of the most popular methods of restricting activities in sensitive ecosystems and around sensitive resources is to pass laws that establish protected areas (Chapter 13).
Commercial operations are also subject to laws that govern natural resource use. Zoning laws, for example, prevent development of sensitive areas, such as riparian forests, beaches, wetlands, and floodplains. In areas where development is permitted, national laws typically require environmental impact assessments (EIAs, see Dana et al. 2012; Biamah et al. 2013) prior to development (Figure 12.4). Construction sites are surveyed during these assessments to ensure that damage is not done to threatened species or sensitive ecosystems. For major regional and national projects such as dams, mines, oil extraction, and highway construction, environmental impact statements must often be prepared that describe a project’s potential damage, and mediatory actions taken.
For industries that exploit threatened species and ecosystems, certification of a product’s origin is increasingly being used as a mechanism to ensure that wild populations are not depleted by illegal collections (Poole and Shepherd, 2016). These certifications may state that environmental regulations, sustainable practices, and socially responsible methods have been followed, or that products were farmed, captive reared, or horticulturally derived rather than collected in the wild. To offset the damage caused by deforestation, various governments have recently made a concerted effort to minimise threats to their forests, including announcing timber export bans and moratoriums on commercial logging. Further afield, the USA, European Union, and Australia have also started placing bans on imported timber that was illegally harvested, some of which was sourced in Africa. Such bans are very effective in reducing the market value of unsustainably sourced products, while also increasing the market share for responsible businesses.
In recognizing the immense harm invasive species inflict on the environment (Section 7.4), some countries have also enacted laws aimed at combatting invasive species. One example is South Africa, where over 500 current and potential invasive species are classified under three categories (http://www.invasives.org.za): Category 1 (destroy immediately, may not be owned), Category 2 (kept only with permit, no trade), and Category 3 (no trade, no breeding, but no need to remove) (Zengeya et al., 2017). Category 2 includes popular pets, such as mallards (Anas platyrhynchos, LC), that can hybridise with native waterfowl, as well as plants, such as gum trees (Eucalyptus spp.) that reduce local water availability (Section 7.4.2). Complementing this effort, the city of Cape Town’s local government launched a competition (http://www.capetowninvasives.org.za) (with prizes) during the first half of 2017 for people who report the location for any of 28 priority invasive species.
Laws that regulate waste management and prevent pollution (Section 7.1) deal with aspects such as air emissions, sewage treatment, hazardous waste, solid waste, and wastewater dumping. In the unfortunate event that pollution ends up in the environment, such laws may also sanction contaminant clean-up. The primary aim of most pollution laws is to protect human health, property, and natural resources such as drinking water, forests, and commercial and sport fisheries. At the same time, they also protect biological communities that would otherwise be destroyed by pollution. For example, air pollution that exacerbates respiratory disease (in humans and animals) also damages commercial forests. Similarly, drinking water pollution which sickens people also kills aquatic species, such as turtles, amphibians, and fish. These examples once again show how intricately human health and economic well-being are linked to the health of the environment.
Tax incentives can encourage environmentally responsible behaviors and reward individuals who contribute to biodiversity conservation.
Most laws meant to protect biodiversity are restrictive in nature, but some regulations take a different tact by rewarding individuals who contribute to biodiversity conservation. Although under-utilised in Africa, perhaps the most popular regulatory reward mechanisms are subsidies and tax incentives. For example, several industrialized countries provide subsidies and tax rebates for citizens and industries that install sustainable energy alternatives, such as solar panels, acquire greener transport options, such as hybrid and electric vehicles, and invest in green infrastructure, such as green roofs and permeable surfaces (Section 14.2). South Africa took its first step of this kind in 2016, when BirdLife South Africa’s Fiscal Benefits Project influenced the introduction of a new tax incentive into national legislation that rewards citizens for making conservation commitments on their land (Stevens, 2017). This tax incentive allows landowners to pay reduced taxes based on the value of their land they have formally declared and manage as a protected area. (For a detailed financial analysis of a similar incentive in Canada, see Schuster et al., 2017). By financially rewarding responsible citizens, national governments can put a smile on their citizens’ faces, while also saving money over the long term given that it is often cheaper to protect intact ecosystem services than restoring damaged ecosystems. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/12%3A_Biodiversity_and_the_Law/12.02%3A_Environmental_Laws_and_Policies.txt |
A single unlawful act—whether negligent or on purpose, by one single person or business—can harm countless ecosystems over a very wide geographic area. Such harm may persist for long periods of time (years, decades, and longer), and impact the lives of thousands of people. For that reason, mutual respect dictates that people and corporations alike abide by the environmental laws and regulations governing their activities. Unfortunately, while most people and businesses comply with environmental laws, it seems that there will always be those who take more than their fair share, corrupt government officials who facilitate smuggling, and greedy corporations that ignores the laws or searches for ways around them for profit. Consequently, there is a constantly need for government structures to evaluate whether environmental laws and regulations are enforced, whether violators are prosecuted, and whether amendments or new laws are needed.
Mutual respect dictates that people and corporations alike abide by the environmental laws and regulations.
Environmental laws can be enforced in several ways. In general, the system works when offences are investigated, and violators are apprehended by law enforcement officers, such as the police (Figure 12.5). Vigilant citizens can also play a role by reporting offences to authorities; financial rewards are increasingly being offered as an incentive for citizens to report environmental crimes. Some districts, environmental agencies, and protected areas may also employ dedicated environmental compliance officers, such as game rangers and anti-poaching units, to monitor human activities that may negatively impact biodiversity. Sometimes, the mere presence of environmental compliance officers is enough to deter illicit activities. With adequate enthusiasm, training, support, and equipment, these teams can have a positive impact on an area’s biodiversity and its people in a short period of time.
When caught, violators are usually punished by being charged fines and/or civil damages, and/or being sentenced to time in prison. Serve severe penalties can act as strong deterrent to those who consider engaging in environmental crimes. For example, Zimbabwe recently sentenced a rhinoceros poacher to 35 years in prison (Rademeyer, 2016), Nigeria sanctioned 26 mining companies for not complying with environmental laws (NAN, 2015), and Cameroon fined two ivory traffickers US \$500,000 plus five years in prison (WWF, 2017). While such severe fines are usually reserved for major offences involving charismatic species, a South African court recently signalled that all biodiversity matters by sentencing a Spanish couple in possession of illegally collected plants to 12 years in prison, in addition to a US \$150,000 fine (Steyn, 2015). As always, it is critically important that every violator is treated equally under the law, whether the violator is the owner of a company that dumps noxious chemicals into a river, a corrupt government official who facilitates smuggling of illegal wildlife products, or an individual caught hunting illegally in a protected area.
New technologies in environmental law enforcement
While environmental law enforcement is the single best predictor of conservation success across Africa (Hilborn et al., 2006; Tranquilli et al., 2012), catching and prosecuting perpetrators can be a difficult and dangerous task. Over the last few years, an increasing number of law enforcement officers have died while protecting the environment (WWF, 2016). Journalists reporting on environmental crimes are also increasingly persecuted, kidnapped, and even murdered (RSF, 2015). Well-organised environmental crime syndicates linked to drug smuggling, terrorism, and other human-rights abuses use increasingly sophisticated tools and tactics to evade detection. Moreover, armed poachers frequently outnumber law enforcement officials. Consequently, refining old and developing new strategies in environmental law enforcement are increasingly necessary.
In Kenya, molecular methods have helped increase conviction rates for environmental crimes from 43% in 2013 to over 90% in 2016.
One of the most promising developments in wildlife conservation has been the rapid development of molecular and other analytical tools and increased data processing capacity, leading to better detection, tracking, and prosecution environmental crimes. One promising development has been the increased use of genetic analysis to aid law enforcement. For example, DNA barcoding—a genetic analysis method that can identify the species of unknown tissue samples—helped expose illegal trade in five species of cycad (Encephalartos spp.), each of them threatened and listed on CITES Appendix I (Williamson et al., 2016). Elsewhere, biologists have started using stable isotope analysis—a technique that analyses an animal’s diet—to determine the origin (captive-bred or wild-caught) of parrots that are for sale (Alexander et al., 2019), and whether rare cycads were wild-collected before or after the practice was banned (Retief et al., 2014). To fully harness the power of molecular methods, wildlife agencies in South Africa and Kenya have even set up dedicated wildlife crime forensic laboratories (Wasser et al., 2007, 2015), where conservationists work closely with forensic scientists to solve wildlife crimes (Box 12.1). These initiatives have already paid off in Kenya, where molecular methods have helped increase conviction rates for environmental crimes from 43% in 2013 to over 90% in 2016 (ODPP, 2017)!
Box 12.1 Insect Biodiversity Helps Solve African Wildlife Crimes
Martin H. Villet
Southern African Forensic Entomology Research Laboratory,
Department of Zoology and Entomology, Rhodes University,
Grahamstown, South Africa.
[email protected]
Poaching and pollution are crimes akin to murder and poisoning and forensic biologists have a set of vitally important tools to convict perpetrators of such crimes: insect biodiversity.
While police detectives sometimes evaluate insects found at crime scenes to help solve murders, forensic entomologists and anti-poaching investigators can use the biodiversity associated with the decomposition of carcasses to solve poaching crimes. The flies and beetles involved in decomposition are like two hands of a clock, flies ticking along in days and beetles indicating weeks.
At least 14 families of flies, including blow flies (Calliphoridae, Figure 12.A), flesh flies (Sarcophagidae), house flies (Muscidae), cheese skippers (Piophilidae), and soldier flies (Stratiomyidae), breed on carcasses in Africa (Villet, 2017), some of them arriving within an hour of the death of the animal to start the decomposition clock. They lay eggs, which hatch into larvae that eat the carcass and grow at a steady pace. The age of these larvae can be estimated by measuring their size when a carcass is found, providing a minimum time between death and discovery. The larvae eventually mature into pupae that give rise to adult flies; this process can also be calibrated to inform the timeline of evidence.
Over 90 species of beetles from at least 10 families also breed on carcasses (Villet, 2011), and can be used to cross-validate evidence from fly larvae; their longer life cycles provide a record that spans a longer period of weeks. Beetles also arrive in a sequence linked to the decomposition process. This pattern of ecological succession starts with clown beetles (Histeridae) and rove beetles (Staphylinidae) that prey on fly larvae, followed by hide beetles (Dermestidae) and carrion beetles (Silphidae) that feed on the dried tissues left by the fly larvae and, finally, by spider beetles (Anobiidae) and hair beetles (Trogidae) that eat the hair, feathers, scales, skin and cartilage left at the end of decomposition (Villet, 2011). The ecological succession clock covers a month or more, depending on the weather and the characteristics of the carcass.
Insect biodiversity can reveal other forensic details, too (Villet, 2015). For instance, insects that feed on drugged or poisoned animal tissue can bioaccumulate contaminants and provide samples for analysis even after the carcass has become too decomposed to analyse directly. Insects may even indicate the presence of these contaminants through their behavior. Animal remains that have been transported from elsewhere and dumped may harbour insects that indicate the route that was travelled. For example, poached parts of African animals bearing insects from Asia have almost certainly travelled through those areas.
The diversity of insects can also provide evidence of environmental crimes involving pollution, a field called environmental forensics. Lethal levels of pollution will change the structure of insect communities, affecting the most sensitive species first. This insight underlies the certified South African Scoring System for aquatic biomonitoring and related scoring systems developed in other African countries for rating the health of rivers based on the biodiversity of their invertebrate inhabitants (Villet, 2015). Sub-lethal levels of pollutants affect insect reproduction and development, which can be detected in impaired reproduction in adult insects and developmental anomalies in insect larvae (Villet, 2017), including increased asymmetry between the left and right sides of the body (termed fluctuating asymmetry) and peculiar developmental patterns. Such research is called environmental forensic entomotoxicology, and it is an exciting new field linking the study of insect biodiversity, environmental damage, and conservation biology.
Conservationists have also become more mindful of the strategies they use to plan and conduct law enforcement monitoring. For example, park managers in Chad now use sophisticated mapping technologies to plan and monitor vulnerable wildlife as well as anti-poaching patrols (Box 12.2), while conservationists working in Cameroon are using acoustic sensors which identify times and areas of increased poaching activity (Astaras et al., 2017). Biologists in the Albertine Rift in turn use a spatial planning software package called Marxan (http://marxan.org)—generally used to identify the locations of new protected areas—to ensure law enforcement activities are more cost effective (Plumptre et al., 2014). To keep staff out of harm’s way and to cover more ground, environmental agencies have also started using unmanned aerial vehicles (UAVs) for law enforcement monitoring (see Box 15.1).
Conservationists are becoming more mindful of the strategies they use to plan and conduct law enforcement monitoring.
Box 12.2 Protecting Elephants in a Hostile Region
Lorna Labuschagne
Previous Address:
African Parks, Zakouma National Park,
N’Djaména, Chad.
Current address:
Frankfurt Zoological Society, Serengeti Conservation Project,
Arusha, Tanzania.
[email protected]
“Extinction is forever” is a phrase we hear often, but perhaps don’t consider deeply enough. The passenger pigeon (Ectopistes migratorius, EX) in North America is a prime example. Early naturalist accounts describe how this species were once so numerous that flocks blackened the sky, and yet it was possible to kill each one. Africa’s elephants are currently under similarly huge pressure, especially in Central and West Africa. The well documented story of the elephants of Zakouma National Park in Chad is a good example, where an estimated 4,000 elephants lost their lives between 2002 and 2010 to feed the insatiable demand for ivory (https://www.africanparks.org/the-parks/zakouma).
In the past, a densely-packed elephant herd was an effective defence against horsemen with spears, whose hunting method centred on isolating an individual. With armed groups coming from as far afield as Darfur, Sudan, the modus operandi of the poachers on horseback has not changed much over the past 200 years, except that the spear has been replaced with an automatic rifle. With today’s poachers shooting indiscriminately into a tightly packed herd, the result is a devastating massacre. In the past, as many as 60 elephants of all ages were killed in a single attack in Zakouma, with many dying later from festering bullet wounds and small calves ending up lost or orphaned. The trauma of such slaughter on these intelligent animals is hard to imagine and is perhaps best understood by the fact that the Zakouma herds stopped breeding for almost five years. So how does one endeavour to stop such carnage on a free-roaming population, and allow elephants to live a normal life again, especially in an open system where herds range widely?
Each area in Africa is different, and it is important to remember that what works in one area will not always work in another. To address a poaching problem, the situation must be carefully assessed, historical information evaluated, and a “feel” for the threats acquired. It is also important to remember that no anti-poaching team can function without the support staff that keeps them equipped and mobile (mechanics, buyers, bookkeepers, etc.).
Below is a list of key initiatives forming the basis of an efficient protection system for a conservation area (Figure 12.B).
• Finding and Tracking Animals: To protect a species, a good understanding of its movements is needed. Several parks in Africa achieve this by fitting satellite GPS collars on individual elephants in different herds. Animal tracking has in the past been primarily used for research purposes, but today the data is also used to monitor elephant movements and adapt anti-poaching patrols accordingly.
• Communication: It is impossible to stop poaching without good communication—be it by mobile or satellite phone, a radio network, or personal trackers with a messaging function. Where the terrain allows it, a digital VHF radio network should ideally be put in place with linked relay stations and portable radios to ensure communication throughout most of the protected area.
• Central Control Room: A Central Control Room (CCR) (Figure 12.C) is where all anti-poaching activities are coordinated day and night. Ideally the park should work on a predictable but unpredictable anti-poaching system; rangers and their families know when they will be on patrol again (predictable), but the day-to-day deployment is unpredictable and coordinated by the CCR using all information available, such as real-time elephant movements, in their decision making. Where to deploy patrols should ideally be made by at least three people and the command then given to the Patrol Leaders, who are trained and equipped with GPS units. Where ranger posts or Forward Operational Bases (FOBs) are used, the unpredictable component can also include not knowing which ranger post or FOB they will be sent to, or with whom. A good rotational policy among rangers plays an important part in keeping rangers alert and motivated.
• Anti-Poaching Monitoring Technology: Today sophisticated mapping techniques are especially helpful for planning and reporting. Not only that, many are available at no cost on the internet. These tools allow conservation practitioners to monitor, record, and display the movements of animals, anti-poaching patrols, aircraft, and poaching incidents, and plot all of them in different layers on a map or satellite image. This is a key component to monitoring the patrol effort and coverage of an area and should be coordinated by the team in the CCR.
• Accessibility Throughout the Year: Although it can be difficult based on the area or budget, conservationists must be inventive in adapting to changing weather and field conditions throughout the year. Airstrips, for example, should be carefully placed to support rangers year-round; accessibility throughout the year is important for logistics but also for evacuation of ill or wounded rangers if needed.
• Intelligence Gathering: Not much happens in rural Africa without somebody knowing about it. The key is to get that information to your CCR. Cell phones are increasingly common, and you might also consider having a toll-free phone number. In areas without good GSM coverage, another option is to put in place a “Village Radio” system, where digital VHF radios are programd in such a way that private calls can be made, allowing for a radio to be installed in a village and still protect the sensitive communications of a park. Having communication in key villages around a park, which speak with the CCR about any illegal activity that they have picked up in the surrounding communities, helps provide much needed security to local people and an important link with the park management.
As protectors of elephants and other wildlife, park managers must assess the situation, decide what can be done in an area, and try it, but most importantly, park managers must employ an adaptive management strategy (Section 10.2.3) and continue adapting over time to a changing situation. Poachers change their strategies, and therefore so must park managers. Ultimately, the goal is to reduce the number of poaching incidents to allow wildlife populations to recover. This can be a daunting task with pitfalls among successes; always remember, keeping field ranger morale high is a key component to ensuring success.
In recent years, some of the world’s biggest conservation organizations banded together to form the SMART (Spatial Monitoring and Reporting Tool) partnership. The main goal of the partnership is improving protected areas management, particularly environmental law enforcement, by enabling law enforcement officials and biologists to more easily collate and process information collected during monitoring and patrols. The partnership accomplishes this through the development of a freely available and fully customisable software package that includes real-time mapping, basic analysis tools, and automatic report generation abilities (Wilson et al., 2019). These features allow park managers to be more strategic in their conservation work by allowing them to better plan, evaluate, and implement their activities. SMART is rapidly becoming the standard in environmental law enforcement across the developing world, and several national governments in Africa have already adopted SMART as its environmental crime monitoring platform.
Despite this progress, older technologies are still being used very effectively in law enforcement. To name a few examples, conservationists continue to rely on tools, such as passive integrated transponder (PIT) tags (Gibbons and Andrews, 2004) and embedding GPS transmitters (e.g. Christy and Stirton, 2015) to identify and track stolen wildlife products such as rhinoceros horns, elephant tusks, valuable timber, and expensive ornamental plants. Lastly, environmental law enforcement officials continue to rely so heavily on well-trained domestic dogs to detect trafficked wildlife products and apprehend environmental criminals that several organizations now specialise in training dogs for conservation purposes. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/12%3A_Biodiversity_and_the_Law/12.03%3A_Environmental_Law_Enforcement.txt |
Despite all the efforts to protect biodiversity through laws and regulations, the scale of environmental crimes continues to increase year after year. Today, the US \$91–258 billion environmental crime industry is the world’s fourth largest illegal enterprise, after drug smuggling, counterfeiting, and human trafficking (Nellemann et al., 2016). Increased financial support for environmental law enforcement could certainty help: current spending to combat environmental crimes, globally estimated at US \$20–30 million a year, is a mere drop in the bucket compared to the losses incurred from these same crimes, which are 10,000 times greater (Nelleman et al., 2016). There is also a need to prevent environmental crimes before they happen, given that the damages incurred cannot always be undone by punishing the offenders.
Environmental crime is the world’s fourth largest illegal enterprise, after drug smuggling, counterfeiting, and human trafficking.
An important step towards reducing the scale of environmental crimes is to address the ineffectiveness of environmental regulations. That includes addressing the range of tactics that criminals use to facilitate non-compliance (Chapron et al., 2017), but also ensuring there are mechanisms that remove the incentives for people to engage in environmental crimes. In the following section, we look at some of the most prominent challenges that complicate environmental law enforcement.
Lack of capacity
The foremost reason why environmental laws fail is that authorities often lack the capacity for effective monitoring and enforcement. Lack of capacity is a major problem in the marine fisheries industry due to the size of the oceans and the cost of patrolling them. This is particularly prominent in West Africa’s oceanic waters (Figure 12.6), which experience the world’s highest levels of illegal and unregulated fishing (Agnew et al., 2009; Gremillet et al., 2015). Apprehending the environmental criminals operating in these and other areas requires resources and manpower, both in short supply.
Lack of capacity is exacerbated by law enforcement officers that turn a blind eye to actions they deem innocuous, or when prosecutors fail to indict criminals out of fear of reprisals. Regulatory controls may also no longer exist or be enforced in regions that experience substantial political instability, economic hardship, civil unrest, or war (Hanson et al., 2009; Beyers et al., 2011). But even in areas where regulatory controls exist, prosecution can be complex, and thus very hard, especially when the illegal activities cross international boundaries and different legal jurisdictions. Such a breakdown of legal mechanisms often leaves natural resources vulnerable to whoever can exploit them.
Conflicting government priorities
Clashing priorities between different government structures complicate the enforcement of environmental laws. We see this when agencies overseeing mining activities issue inappropriate permits because of pressure for economic development, or because of corrupt agreements between businesses and government officials (Mascia and Pailler, 2011). Another example of mixed priorities occurs when a national government gives permission to extractive companies to exploit protected areas (Section 13.7.3) or communal lands without first consulting and obtaining local input and consent. Such government-sanctioned violations are generally very difficult to prosecute and require an active and caring citizenry to take their governments to task.
One of the most popular methods for citizens to make themselves heard is activist activities, such as public protests. Concerned individuals can also launch petitions on websites such as https://www.change.org, http://www.greenpeace.org, and https://www.avaaz.org. There is even a website for whistle-blowers (https://wildleaks.org) who want to report environmental crimes anonymously. Another positive development is the growing number of successful lawsuits that concerned citizens and environmental justice organizations have brought against their governments for environmental violations (e.g. CER, 2017; Yende, 2017). The Kenyan conservation organization Wildlife Direct has taken this a step further; they are keeping citizens informed about lawsuits involving environmental crimes through a website dedicated to tracking and reporting on such cases (https://wildlifedirect.org/legal-program-3).
Informal economies, traditional activities, and the law
Law enforcement can at times be counterproductive. This is true especially in areas where the separation between informal/unreported and illegal activities is blurred. For example, traditional people who graze their livestock, collect medicinal plants, or hunt and trap animals in protected areas that were established on ancestral land seldom have criminal intent. But because formal law systems seldom account for these informal activities, those people are engaging in illegal activities. Similarly, confusing terminology may also lead to unintended conflict. For example, in some parts of Cameroon, the cultural definition of a hunter describes someone who owns a gun and makes a living from hunting animals (Hofner et al., 2018). In this context, some people consider it within the law to trap animals with snares, or even to make sporadic “hunting” trips into a protected area where hunting is forbidden, given that it is not for commercial purposes.
When dealing with vulnerable people, conservation initiatives must also consider impacts on livelihoods and the potential for sustainable utilisation
When dealing with vulnerable peoples whose livelihoods are threatened, an approach that involves sensitivity and compassion generally offers more effective and enduring resolutions. Many conservation initiatives have not only failed but have also created long-lasting negative attitudes by preventing traditional peoples from sustaining their livelihoods. Before implementing new regulations, governments should carefully consider if they would disrupt livelihoods. If so, it might be wise to consider if some form of sustainable utilisation isn’t possible. For example, while pastoralist activities in Tanzania’s Ngorongoro Conservation Area may lead to wildlife declines (Boone et al., 2002, but see Ogutu et al., 2016), conservation authorities decided on a suitable compromise by allowing Maasai herders to graze their livestock in the area on the condition that they exit daily. (See also Section 13.5.2 for discussion on zoning.)
When banning activities which have become unsustainable, it helps to provide start-up resources and market access to help affected people comply with new restrictions.
Despite best intentions, sustainable utilisation is not always possible, and the actions of people engaging in environmentally detrimental activities are not compatible with conservation goals. In such cases, it is important to implement controls that enable those people to transition toward sustainable activities. Failing that, conservation activities may unintentionally force the affected people to resort to illegal activities such as poaching out of desperation to obtain food and income. When banning previously-allowed activities that have become unsustainable, it helps to provide affected people with start-up resources and market access to help them comply with new restrictions while also meeting broader societal needs. For example, to fulfil income and nutritional needs when bushmeat harvesting is banned, it might be necessary to help hunters transition to farming with animals that reproduce quickly in captivity. Raising poultry can be a good alternative to bushmeat because chickens grow quickly, provide eggs, feed on insect pests, and need little land for maintenance. Farming with locally-adapted wildlife, such as large snails (e.g. Carvalho et al., 2015) and cane rats (Thryonomys swinderianus, LC) (e.g. van Vliet et al., 2016), has also proven to be a profitable and sustainable alternative to the bushmeat trade (for a review on wildlife farming for conservation, see Tensen, 2016). Initially, many people may resist the risks involved in leaving behind familiar activities. It is, therefore, important to explain carefully the reasoning behind those changes (e.g. “bushmeat hunting drives away tourist dollars”, Rogan et al., 2017). It may also be beneficial to enable the affected individuals to travel to areas where they can see first-hand how more sustainable activities can benefit local people.
Trade embargoes and sanctions
The basic premise of CITES is that, so long as participating countries abide by agreed-upon regulations, trade involving species of concern will not be stopped, only monitored. However, when agreements are not met, or compliance falls short, then trade is banned in part or in whole. For example, in early 2016, following failures to comply with international trade regulations, CITES instituted blanket suspensions on trade of all CITES-regulated products against 14 African countries (CITES, 2016). Another example from 2016: after various high-profile environmental crimes that involved CITES-regulated species, the USA afforded protection to lions and elephants under their Endangered Species Act (https://www.fws.gov/endangered). Trophy hunters from the USA now face significant regulatory and logistical barriers which has all but eliminated hunting of these species for wealthy American hunters, threatening a US \$500 million per year industry that supports over 53,000 jobs and protects over 1.4 million km2 of land (SCIF, 2015). In both these examples, businesses operating within the law are unfortunately also impacted.
Many formerly destructive companies are now voluntarily pursuing opportunities to prove, through special certifications, that their products are harvested responsibly and sustainably.
To avoid scenarios such as these, it is much more advantageous for pressure to mount from within non-complying countries and industries before outside pressure takes effect. To that end, many formerly destructive companies are now voluntarily pursuing opportunities to prove, through special certifications, that their products are harvested responsibly and sustainably. Four prominent certification agencies operating in Africa are the Forest Stewardship Council (FSC) which sets guidelines for the responsible management of forests, the Marine Stewardship Council (MSC) which sets standards for sustainable fisheries, the Roundtable on Sustainable Palm Oil (RSPO) which promotes sustainable production of palm oil, and the Rainforest Alliance which promotes sustainable agriculture (For a more complete treatment of sustainability standards, see https://www.isealalliance.org, https://www.evidensia.eco, and https://www.iisd.org/ssi). Given that most certification schemes were established relatively recently, they are not without their flaws, but collaborations with conservation biologists (e.g. Christian et al., 2013; WWF, 2013) play a major role in ensuring continued improvements. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/12%3A_Biodiversity_and_the_Law/12.04%3A_The_Limits_of_Environmental_Laws_and_Regulations.txt |
Challenging problems are often solved when a diverse group of people from different backgrounds and viewpoints come together for mutual benefit. Solving environmental crime is no different: history shows that building trust and respect for fellow human beings and future generations is more powerful than the threat of force. Effective law enforcement efforts, which require multi-level cooperation from international structures down to individual people (Box 12.3), are often characterized by partnerships between wildlife agencies and local people (Biggs et al., 2016). These partnerships may take the form of environmental educational campaigns to incentivise conservation action and sustainable resource use (Abensperg-Traun, 2009). Cooperation between different individuals at the grassroots level is another very effective means to ensure sustainable resource use. These efforts may take many forms, but they begin with individual and group decisions to prevent the destruction of habitats and species to preserve something of perceived economic, cultural, biological, scientific, or recreational value. Through collaboration and cooperation, both from the grassroots level up, and governments down, conservation biologists can achieve their goals, by ensuring free and fair treatment of all citizens regardless of their diverse and sometimes opposing viewpoints on natural resource management.
Box 12.3 Thoughts on Poaching and Illegal Wildlife Trafficking in Sub-Saharan Africa
Tamar Ron
Biodiversity Conservation consultant.
[email protected]
The two main causes of the alarmingly rapid wildlife loss in Africa today are: (1) unsustainable use of land and natural resources, mostly related to decision making that does not prioritise conservation considerations; and (2) overharvesting of wild animals and plants through poaching and illegal logging.
Poaching and illegal logging can be locally driven for subsistence use, resulting from poverty; for lack of other protein, energy, and income sources; and, at times, by the intensifying impact of armed conflict or post-conflict situations. In a country like Angola, for example, where war has significantly diminished wildlife populations, the continuous impact of intensive bushmeat poaching may well lead to the extinction of remnant core populations of species that have initially survived the armed conflict. It is severely impacting even the iconic and endemic giant black sable (Hippotragus niger variani, CR). Bushmeat poaching is typically unselective; it targets mainly large and medium sized mammals, but also smaller mammals, birds, reptiles and freshwater fish. Similarly, illegal logging for wood, charcoal, or slash-and-burn-based cultivation, results in rapid and irreversible biodiversity degradation and loss.
In contrast, commercial poaching is often driven by international trafficking, whether of live animals (Figure 12.D) or animal and plant products, from source countries to destination markets. Illegal traffickers are often well financed, sophisticated, and involved in other forms of serious crime, at times even in terrorism (e.g. Nellemann et al., 2014). Illegal wildlife trade is selective and forms an imminent threat to iconic species with commercial value, such as elephants, rhinoceros, big cats, great apes, pangolins, sea turtles, parrots, and rosewood, to name a few.
While the core causes, nature, and impacts of subsistence and commercial poaching are different, local community members are a centrepiece of both. Their intimate acquaintance with local wildlife and their habitats is vital; therefore, effective wildlife protection is based on their active engagement. After many generations of alienation, local community members must be included in the decision-making process for the sustainable management of natural resources. The Namibian conservancies (Section 14.3) offer a model of success in engaging communities in conservation, by protecting their rights, securing their fair benefits as the resource owners and not merely as workers, and providing them with adequate training.
Wildlife crime is a serious threat to biodiversity and while local influences should be recognized and addressed, these crimes should be treated as a global enforcement priority. Continuous and coordinated national efforts of all relevant sectors, and with global cooperation, are essential to success. Such efforts must include: (1) improving awareness at all levels; (2) adequate legislation and policies; (3) realistically deterring punishments and forfeiture of wildlife crime revenues; (4) strengthening enforcement and intelligence capacities in all source, destination, and transit countries; (5) addressing governance challenges; and (6) trying to eradicate the markets, or at least to reduce the demand for illicit wildlife products. The poaching drivers, international crime syndicates, and middlemen, must be targeted. Enforcement focused mostly at the poachers’ level can never achieve the desired results. Often, they are no less victims than their target species. Further, if the world wants to protect iconic species, it cannot be expected that the burden of their conservation, and human-wildlife-conflict damages, should fall solely on those communities that happen to share their habitat. The global effort and substantive support required should not be viewed as a contribution, but as mutual responsibility for achieving a global goal.
Lastly, we may have to accept that total eradication of wildlife crime may not be achievable. There is no magic remedy, nor a single perpetrator. Nevertheless, integrated efforts to reduce these crimes must be strengthened at all levels. We simply cannot give up on our fellow species. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/12%3A_Biodiversity_and_the_Law/12.05%3A_Conclusion.txt |
1. Environmental laws and regulations are implemented at three different levels: international treaties, national laws, and local laws. Each of these levels is intricately connected: international treaties influence—but also depend on—national laws to succeed, while national laws are guided by local needs and customary laws that have been in place for generations.
2. International treaties and conventions provide frameworks for countries to cooperate on protecting species, ecosystems, and other levels of biodiversity. International agreements are important because: (1) many species migrate and disperse across borders, (2) ecosystems do not follow administrative boundaries, (3) pollution spreads by air and water across regions and around the globe, (4) many biological products are traded internationally, and (5) some environmental problems require global cooperation and coordination.
3. National governments protect biodiversity by regulating natural resource use and preventing pollution. Subsidies and tax incentives can also be used to reward citizens and businesses that engage in environmentally-responsible behaviors.
4. There is a constant need to evaluate environmental laws and regulations to ensure that they are enforced, violators are prosecuted, and new laws and amendments are passed as and when needed. Law enforcement agencies and scientists are constantly looking for new ways to address enforcement shortcomings.
5. When banning environmentally detrimental activities that are not compatible with conservation goals, it is critical to help the affected people to transition toward sustainable activities. Failing that, conservation activities may unintentionally force those people to resort to illegal activities such as poaching out of desperation to obtain food and income.
12.07: Topics for Discussion
1. Identify two or three environmental laws regulating human activities in your country. Discuss how these national laws relate to international laws and local laws.
2. A wide range of international and national laws protect threatened species and ecosystems. Why do species and ecosystems covered by such laws not recover?
3. Trapping birds and small mammals are traditional activities for boys and young men across Africa, and traditional hunting is seen as a competitive sport to sharpen the mind. These activities are also important to meet local income and nutritional needs. How should we deal with the hunting of threatened species outlawed by national governments, but encouraged by local traditions? Should people be allowed to hunt for bushmeat, even if it includes great apes and other rare species, to pay for necessities, such as schooling and medicine?
12.08: Suggested Readings
Biggs, D., R. Cooney, D. Roe, et al. 2016. Developing a theory of change for a communitybased response to illegal wildlife trade. Conservation Biology 31: 5–12. https://doi.org/10.1111/cobi.12796 Four pathways to community-level action against wildlife crimes.
Chapron, G., Y. Epstein, A. Trouwborst, et al. 2017. Bolster legal boundaries to stay within planetary boundaries. Nature Ecology and Evolution 1: 0086. https://doi.org/10.1038/s41559-017-0086 Various tactics are used to make biodiversity laws less effective.
Christian, C., D. Ainley, M. Baileyet al. 2013. A review of formal objections to Marine Stewardship Council fisheries certifications. Biological Conservation 161: 10–17. https://doi.org/10.1016/j.biocon.2013.01.002 Strengths and weaknesses of schemes certifying sustainability.
Christy, B., and B. Stirton. 2015. How killing elephants finances terror in Africa. National Geographic. http://on.natgeo.com/1I5N2aO. A riveting account of two journalists tracking poaching routes.
Etiendem, D., L. Hens, and Z. Pereboom. 2011. Traditional knowledge systems and the conservation of Cross River gorillas: A case study of Bechati, Fossimondi, Besali, Cameroon. Ecology and Society 16: 22. http://doi.org/10.5751/ES-04182-160322 Traditional belief systems can be used to promote conservation.
Harfoot, M., S.A.M. Glaser, D.P. Tittensor, et al. 2018. Unveiling the patterns and trends in 40 years of global trade in CITES-listed wildlife. Biological Conservation 223: 47–57. https://doi.org/10.1016/j.biocon.2018.04.017 Even legal wildlife trade is having a massive impact on biodiversity.
Hilborn, R., P. Arcese, M. Borner, et al. 2006. Effective enforcement in a conservation area. Science 314: 1266. https://doi.org/10.1126/science.1132780 Enforcement of environmental laws really makes a difference.
Plumptre, A.J., R.A. Fuller, A. Rwetsiba, et al. 2014. Efficiently targeting resources to deter illegal activities in protected areas. Journal of Applied Ecology 51: 714–25. https://doi.org/10.1111/1365-2664.12227 Prioritising law enforcement efforts can save money.
Tulloch, A.I.T., N. Auerbach, S. Avery-Gomm, et al. 2018. A decision tree for assessing the risks and benefits of publishing biodiversity data. Nature Ecology and Evolution 2: 1209–17. https://doi.org/10.1038/s41559-018-0608-1 Publishing biodiversity data has risks that should be considered
Wasser, S.K., L. Brown, C. Mailand, et al. 2015. Genetic assignment of large seizures of elephant ivory reveals Africa’s major poaching hotspots. Science 349: 84–87. https://doi.org/10.1126/science.aaa2457 Genetics can be used to identify poaching hotspots. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/12%3A_Biodiversity_and_the_Law/12.06%3A_Summary.txt |
Abensperg-Traun, M. 2009. CITES, sustainable use of wild species and incentive-driven conservation in developing countries, with an emphasis on southern Africa. Biological Conservation 142: 948–63. http://doi.org/10.1016/j.biocon.2008.12.034
Agnew, D.J., J. Pearce, G. Pramod, et al., 2009. Estimating the worldwide extent of illegal fishing. PLoS ONE 4: e4570. https://doi.org/10.1371/journal.pone.0004570
Alexander, J., C.T. Downs, M. Butler, et al. 2019. Stable isotope analyses as a forensic tool to monitor illegally traded African grey parrots. Animal Conservation 22: 134–43. https://doi.org/10.1111/acv.12445
Astaras, C., J.M. Linder, P. Wrege, et al. 2017. Passive acoustic monitoring as a law enforcement tool for Afrotropical rainforests. Frontiers in Ecology and the Environment 15: 233–34. https://doi.org/10.1002/fee.1495
Beyers, R.L., J.A. Hart, A.R.E. Sinclair, et al. 2011. Resource wars and conflict ivory: The impact of civil conflict on elephants in the Democratic Republic of Congo-the case of the Okapi Reserve. PloS ONE 6: e27129. https://doi.org/10.1371/journal.pone.0027129
Biamah, E.K., J. Kiio, and B. Kogo. 2013. Chapter 18–Environmental Impact assessment in Kenya. Developments in Earth Surface Processes 16: 237–64. https://doi.org/10.1016/B978-0-444-59559-1.00018-9
Biggs, D., F. Courchamp, R. Martin, et al. 2013. Legal trade of Africa’s rhino horns. Science 339: 1038–39. http://doi.org/10.1126/science.1229998
Biggs, D., R. Cooney, D. Roe, et al. 2016. Developing a theory of change for a community-based response to illegal wildlife trade. Conservation Biology 31: 5–12. https://doi.org/10.1111/cobi.12796
Boone, R.B., M.B. Coughenour, K.A. Galvin, et al. 2002. Addressing management questions for Ngorongoro Conservation Area, Tanzania, using the SAVANNA modelling system. African Journal of Ecology 40: 138–50. https://doi.org/10.1046/j.1365-2028.2002.00357.x
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With its rich biological diversity, Africa plays a critical role in global conservation efforts. Yet, many of the continent’s most threatened species and ecosystems continue to face an uncertain future. In light of increasing human populations that need an increasing amount of natural resources each year, safeguarding the region’s biodiversity is a major challenge. One of the best ways to meet this challenge is to designate protected areas—regions where human activities are regulated or, at times, even prohibited by law.
Protecting existing wild populations in their natural ecosystems not only protects ecological communities and interactions, but also natural processes and ecosystem services.
Biodiversity conservation is most effective when we maintain healthy, functioning, and intact ecosystems. Although it is true that many species and populations live outside protected areas, and some wildlife populations (Craigie et al., 2010) and natural communities (Lindsey et al., 2014) are declining even when protected, well-managed protected areas continue to be the most effective method to safeguard biodiversity (Brooks et al., 2009; Ihwagi et al., 2015). Illustrating the point, a global meta-analysis, which included 952 locations across Sub-Saharan Africa, found that wildlife populations are 15% larger and species richness is 11% higher inside protected areas compared to populations directly outside (Gray et al., 2016). Differences may be even starker at individual sites: tea fields on Tanzania’s East Usambara Mountains held only 8% of the bird species present in the adjacent protected forest (Newmark, 2008), while some vultures in Eswatini now exclusively breed in protected areas (Monadjem and Garcelon, 2005). Studies from Tanzania have also shown how wildlife in protected areas are more resilient to climate change (Beale et al., 2013a), because habitat loss and fragmentation occur at four times their respective rates outside protected areas relative to inside them (see also Potapov et al., 2017). Consequently, until such a time that we can live more sustainably on unprotected lands, protected areas will remain an important cornerstone in our efforts to protect biodiversity. But how do we know what or where to protect, how much to protect, or how to effectively manage a protected area?
13.01: Establishing Protected Areas
A protected area is “a clearly defined geographical space (Figure 13.1), recognized, dedicated and managed through legal or other effective means to achieve the long-term conservation of nature with associated ecosystem services and cultural values” (Dudley, 2008). Given this broad definition, it comes as no surprise that governments, organizations, and local communities use a variety of mechanisms to establish protected areas. The most popular of these mechanisms are:
• Government action, which can occur at a national, regional, or local level.
• Community-based initiatives by local people and traditional groups.
• Land purchases and holdings by private individuals and organizations.
• Protected areas established through co-management agreements.
• Development of biological field stations or marine laboratories.
Government protected areas
Government actions are generally considered the most secure form of protection because they involve passage of laws and buy-in from multiple levels of society. Of course, legislation establishing a protected area does not guarantee that the species and ecosystems therein are adequately preserved. Small populations, especially those living in small protected areas, often require active management (Section 8.7.5) to ensure their continued survival. Another concern is that laws protecting national parks and other wildlife sanctuaries are not strictly enforced, leading to so-called paper parks—parks that appear on official government lists, but with wildlife monitoring, law enforcement, and ecosystem management lacking on the ground (Laurance et al., 2012). However, government-sanctioned protected areas do lay a solid foundation for partnerships among governments, international conservation organizations, multinational banks, research institutes, and educational organizations. Such partnerships can bring together funding, training, and scientific and management expertise to maximise the potential value of those protected areas.
Community conserved areas
In many areas, local people already protect biological communities, forests, wildlife, rivers, and coastal waters in the vicinity of their homes. Protection on these community conserved areas is enforced by village elders and councils to ensure the sustainable use of natural resources such as food supplies and drinking water. Natural areas have also been set aside by royal families and churches to provide a space for spiritual activities (see Box 2.1) and sustainable harvesting of medicinal plants (see Box 5.2). Because human activities are highly restricted in these sacred spaces, they provide an important refuge for biodiversity. Today, an increasing number of traditional communities link cultural advocacy directly to conservation through the establishment of protected areas on their lands as a safeguard against developments that would compromise their way of living. Other communities establish protected areas to attract tourists and ensure the protection of special wildlife. One such example is the Iyondji Bonobo Community Reserve in the DRC, which protects bonobos (Pan paniscus, EN), forest elephants (Loxodonta cyclotis), as well as one of the world’s most enigmatic birds, the Congo peafowl (Afropavo congensis, VU) (Dupain et al., 2013).
Traditional communities may link cultural advocacy to conservation by establishing protected areas as a safeguard against developments that would compromise their way of living.
Privately protected areas
Over the last few decades, many African countries have adopted a more Western form of land tenure under private ownership. Wealthy individuals or groups of people have taken advantage of this opportunity by acquiring large tracts of land for ecotourism purposes (de Vos et al., 2019). Because the ecotourism potential of these privately protected areas depends on how well the property is managed (Clements et al., 2016), private landowners often invest considerable effort to maintain and even increase wildlife populations on their land. Privately protected areas have unique advantages over government-protected areas. For example, they have local buy-in from landowners and their employees by design; this is often a significant stumbling block for government-protected areas. Private sites could also employ innovative funding mechanisms that allow them to fast-track land acquisition, perhaps in response to threats such as development. In some areas, privately protected areas may even employ more people, pay better wages, and contribute more to local economies that government protected areas (Sims-Castley et al., 2005). Privately protected areas can, therefore, play a significant role in overall conservation efforts (see Box 2.3), particularly in areas where threatened species (Cousins et al., 2010) and ecosystems (Gallo et al., 2009) are underrepresented in government-protected areas.
Because the ecotourism potential of private protected areas depends on how they are managed, landowners prioritize maintaining and even increasing wildlife populations on their land.
Despite the advantages of privately protected areas, we must also consider the drawbacks. Like many community conserved areas, privately protected areas are not permanently protected by the same mechanisms and oversight as government protected areas are. Ownership and management style can also change at the whim of the landowner, or perhaps the heirs. At times, management practices may be detrimental to the species and ecosystems these privately protected areas claim to protect, for example, through introduction of invasive species and harmful breeding practices (Milner et al., 2007), and by resisting regulatory controls (Cousins et al., 2010). Innovative strategies will thus be required to ensure that these areas do contribute to biodiversity protection, which include education, support, and methods that balance financial gains with conservation goals.
Co-managed protected areas
Local people who support conservation and the protection of their local natural resources are often inspired to take the lead in protecting their local biodiversity. Governments and conservation organizations can assist such initiatives by allowing local people to access specialist expertise and obtain financial assistance to develop conservation and ecotourism infrastructure. These conservation areas, characterized by partnerships between different levels of society that share decision-making responsibilities and consequences of management actions, have been termed co-managed protected areas. Tanzania, where the management of more than two million hectares of forests and woodlands have been transferred to local groups (Blomley et al., 2019), has been particularly active in this regard. One of the biggest strengths of co-management is that, with proper consultation and engagement, it avoids eco-colonialism—the unfortunate practice by some governments and conservation organizations of disregarding the rights and practices of local people during the establishment and management of new conservation areas or environmental laws and regulations.
Contractual parks offer a good model on how to avoid eco-colonialism. These protected areas are established and managed through agreements with private or communal landowners whose land forms part of a protected area (usually a national park). This not only allows a larger area to be protected, but also allows local people to benefit from biodiversity conservation through benefit sharing and job generation initiatives. Contractual parks play an important role, especially in South Africa, where it is used as a tool to meet both conservation goals and restitution of previously dispossessed land. One such example is the |Ai-|Ais/Richtersveld TFCA (Figure 13.2), which protects a huge number of succulent plant species and a variety of desert ecosystems at the border between Namibia and South Africa. Much of this national park is make up of communal lands, with the landowners—the local Nama people—having co-management and benefit-sharing agreements with the South African government (Reid et al., 2004). Incorporating activities of the landowners in sections of the park enriches tourism experiences, such as boating, hiking, and birdwatching, and contributes to preserving the Namas’ cultural identity, pastoral lifestyle, and threatened local languages (Chennels, 1999).
Field stations and marine laboratories
Biological field stations and marine laboratories are a special kind of protected area that provide a dedicated stable space for scientists, students, and even the general public to pursue research projects on all kinds of natural phenomena in an intact environment (Tydecks et al., 2016). By facilitating collaboration and long-term observation, work done at field stations in Africa has led to several fundamental scientific advances, including improved understanding of environmental responses to climate change and acid rain, as well as advances in social development through conservation activities. Today, there are biological field stations in at least 24 Sub-Saharan African countries (Tydecks et al., 2016). Among them are Namibia’s Gobabeb Research and Training Centre which focuses on desert conservation, Kenya’s Mpala Research Centre (Box 13.1) which investigates the potential for wildlife and livestock to coexist, Nigeria’s A.P. Leventis Ornithological Research Institute (see Box 15.4) which focuses on bird conservation, and Uganda’s Makerere University Biological Field Station which has a long, distinguished record of primate research.
Box 13.1 Mpala Research Centre: A Living Laboratory for (More than Just) Scientists
Anchal Padukone and Dino J. Martins
Mpala Research Centre,
Nanyuki, Kenya.
http://www.mpala.org
In the heart of Kenya’s Laikipia district, Mpala Conservancy stretches over 200 km2 of semi-arid savannah, acacia bushland, wooded grassland, rocky escarpments and riverine communities along the Ewaso Nyiro and Ewaso Narok rivers. The area is home to an abundance of wildlife, including all the classic savannah mammals: impala (Aepyceros melampus, LC), Grant’s gazelles (Nanger granti, LC), reticulated giraffe (Giraffa camelopardalis reticulata, EN), leopards (Panthera pardus, VU), lions (P. leo, VU), spotted hyenas (Crocuta crocuta, LC), and some of the largest savannah elephant (Loxodonta africana, VU) and African wild dog (Lycaon pictus, EN) populations in Kenya. There are also a few species typical of the northern regions of the Somali-Maasai centre of endemism, such as Grevy’s zebra (Equus grevyi, EN) and gerenuk (Litocranius walleri, NT). Mpala also functions as a working cattle ranch, with upwards of 2,000 cattle, camels, and sheep that are available for use by researchers.
This “multiple use” landscape and its neighbouring ranches provide exceptional opportunities for researchers to study interactions among humans, their domestic herds and wildlife in an area where they coexist. Since much of East Africa’s wildlife is found in similar areas outside formal protection, such research could provide essential and widely applicable knowledge for conservation efforts. They will be particularly important as conservation managers will increasingly have to balance wildlife and rangeland management needs to remain effective, in a context of human population growth and economic development. The Mpala Research Centre, established here in 1994, attracts hundreds of scientists every year who use this “living laboratory” to pursue projects varying in scope from the population biology of individual species to community-level dynamics and ecosystem functioning (Rubenstein and Rubinoff, 2014).
Having a research station in this area facilitates long-term and large-scale field experiments, including the Kenya Long-term Exclosure Experiment (KLEE). The 18 KLEE plots are designed to keep out different groups of animals: some plots only exclude megaherbivores (e.g. elephants and giraffes); others exclude all large herbivores; still others only exclude domestic cattle, among other combinations. This allows for controlled studies of the effects of different groups of herbivores on the vegetation and on each other. This research reveals that while domestic stock and wild grazers compete for forage during the dry season, the presence of zebras enhances cattle weight gain during the wet season—perhaps because zebras consume dead grass parts, improving forage quality for cattle (Riginos et al., 2012). Other studies at Mpala have also shown that wildlife and livestock can coexist and facilitate each other’s success, given the right approaches in management (Odadi et al., 2011; Ogutu et al. 2016).
Another long-time focus of research at Mpala is the threatened Grevy’s zebra. Dr. Daniel Rubenstein (Princeton University) and his research group examine the influence of environmental features on competitive behavior and reproductive patterns in plains zebra and Grevy’s zebra. In turn, they are interested in how these social processes influence zebra population size. Their findings have the potential to inform management strategies in areas where Grevy’s numbers are too low to be self-sustaining (Rubenstein, 2010).
Involving the non-scientific community, especially those living around conservancies, is crucial to the long-term success of conservation efforts. Recognising this, Mpala has hosted several citizen science initiatives. For example, The Great Grevy’s Rally was a photographic census that relied on inputs from both scientists and members of the public, who travelled to conservancies, such as Mpala, to take pictures of every Grevy’s zebra they could find. Researchers processed these images using the Image Based Ecological Information System (IBEIS, http://ibeis.org) to differentiate individuals using their stripe patterns. This allows them to determine population size and structure, and assess whether zebra numbers are stable, increasing, or decreasing.
Also hosted at Mpala, the Kids Twiga Tally (Kahumbu et al., 2016) was a similar “sight-resight” survey of reticulated giraffes that relied on IBEIS software to distinguish between individuals and determine population structure. Its 70 young participants (Figure 13.A) came from both city schools and nearby pastoralist communities, spanning a range of socio-economic backgrounds. After spending two days taking pictures of giraffes on GPS-enabled cameras, the children returned to their schools having contributed meaningfully to conservation science. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.00%3A_Prelude_to_The_Importance_of_Protected_Areas.txt |
Protected areas vary greatly in how they are managed. For some, particularly those that protect very sensitive and/or recovering wildlife populations and ecosystems, human activity—even activities, such as photography, hiking, or bird watching (which can cause trampling and disturb shy animals)—may at times need to be forbidden except for specially arranged guided tours. For others, extraction of natural resources may be permitted albeit regulated.
To distinguish how protected areas are managed, the IUCN developed six categories to classify protected areas based on how the land is used (Table 13.1). Of these categories, the first five can be defined as true protected areas, because the environment is managed primarily for biological diversity. The sixth category, Protected area with sustainable use of natural resources, refers to extractive reserves that are managed primarily for the sustainable production of natural resources, such as timber and grazing lands. Nevertheless, extractive reserves can play an important role in conservation: (1) they frequently protect much larger areas than do other types of protected areas; (2) they still provide habitat for many species that were present pre-extraction; and (3) they often border and can thus provide a buffer around, and wildlife linkage between, category I–V protected areas.
Managers of extractive reserves must seek balance between the harvest of natural resources and risking environmental degradation from unsustainable practices.
Table 13.1 Description of Categories I–VI of the IUCN’s classification of protected areas.
Category
Description
Ia
Strict nature reserve
Managed strictly for biodiversity conservation. Serves as reference sites for research and monitoring. Human visitations and impacts highly regulated.
Ib
Wilderness area
Generally large and relatively unmodified natural areas without significant human habitations. Managed to preserve the area’s natural character and ecological integrity.
II
National Park
Large natural areas set aside for protection of biodiversity and ecosystem processes. Also managed to support human activities (spiritual, education, scientific, recreation) compatible with biodiversity protection.
III
National monument of feature
Managed to protect a natural feature (e.g. seamount, geological feature, ancient grove) with outstanding cultural and/or natural significance. Can cover a small area, and often have high visitor value.
IV
Habitat/species management area
Protected area dedicated to the protection of a specific species of habitat. May at times required regular and active intervention to ensure primary management goals are met
V
Protected landscape/seascape
An area with a significant natural or cultural value, created by the interaction between people and nature. Managed to safeguard the interactions that sustains the area’s value. Often act as model for sustainability
VI
Managed-resource protected area
Managed primarily for the low-level, non-industrial, sustainable use of natural resources. Generally large, with most of its ecosystems intact.
Source: After Dudley, 2008
It is important to note that not all protected areas are covered under the IUCN’s six-category system. Prominently are RAMSAR wetlands (Section 12.1.2) which are not incorporated under formal protected areas, but still protected under international law. Other examples include locally-managed marine areas and indigenous reserves, some of which are as effectively managed as formal protected areas. The IUCN is currently working on a new classification, called ‘other effective area-based conservation measures’ (OECM; IUCN WCPA, 2018), to officially recognize the contribution of areas falling outside formal protected area networks to biodiversity conservation efforts. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.02%3A_Classification_of_Protected_Areas.txt |
Historically, the boundaries of protected areas was often determined through pragmatic considerations, such as the availability of funds and land, and political influence, rather than ecological considerations. Many conservation areas were thus established on “lands that nobody wants”: marginal areas with little agriculture and development potential, or areas that were too remote to have high commercial value (a trend that continues even today: Venter et al., 2018). Other protected areas were established in locations with charismatic megafauna, so ecosystems without those species remained unprotected. Consequently, some of Africa’s most threatened species and ecosystems remain under-protected (Beresford et al., 2011).
In a crowded world with finite natural resources and limited funding, it is increasingly important to be strategic about where protected areas are established.
In a crowded world with finite natural resources and limited funding, it is becoming increasingly important to be strategic about where protected areas are established. To do this, conservation biologists and policy makers must answer three key questions: (1) What is most important to protect? (2) Where would it be best protected? (3) How could it be most effectively protected? Three criteria can be used to answer the first two of these questions:
• Distinctiveness (or irreplaceability): Ecosystems with species that are distinct in their taxonomy (e.g. ecosystems that contain the only species in a taxonomic group) or geographic distribution (e.g. endemic species), or ecosystems with unique attributes (e.g. scenic landscapes, unusual geological features).
• Endangerment (or vulnerability): Areas that contain concentrations of species threatened with extinction, or ecosystems in danger of being destroyed.
• Utility: Species and ecosystems that people value, including culturally significant species, economically valuable species or ecosystems, or areas that can contribute to combating climate change.
Using these criteria, scientists have developed several broadly complementary methods to prioritise areas for protection. The approaches differ more in what traits they emphasise rather than in their fundamental principles. Thus, although some people may argue about which approach is better, each approach contributes to the protection of biodiversity.
Species approach
Many protected areas are created to protect (e.g. threatened, culturally significant, or keystone) species. Species that provide the motivation to establish a protected area are known as focal species. As a prominent example using the focal species concept, the Alliance for Zero Extinction (http://www.zeroextinction.org) identified 67 priority sites across Sub-Saharan Africa (853 sites globally) that contain the last remaining populations of one or more Endangered or Critically Endangered species. Flagship species, such as gorillas, are a special kind of focal species because they capture public attention, have symbolic value, and are important for ecotourism purposes. Many flagship species and focal species are also umbrella species, because their protection indirectly benefits other species and ecosystem components with which they share their landscape.
Protected areas are often established to protect threatened or charismatic species, unique ecosystems, and or wilderness areas.
Ecosystem approach
There is debate among conservation biologists over whether ecosystems rather than individual species should be the primary target of conservation efforts. Supporters of an ecosystem approach argue that protecting and managing ecosystems can preserve more species and provide more value to people than spending the same amount of money to protect individual species. Focusing on ecosystems also allows for greater flexibility in justifying conservation efforts, because it can be easier to demonstrate the economic value of ecosystems for helping to control floods, filtering water, and providing opportunities for recreation. To that end, the WWF has identified 238 ecoregions across the globe (the “Global 200”)—57 of them in Sub-Saharan Africa—that are most crucial to the biodiversity conservation (Olson et al., 2002). This Global 200 analysis formed the basis of a more recent global assessment that identified 41 at-risk ecoregions—areas of high conservation priorities because they are undergoing high levels of habitat conversion and have low protected areas coverage (Watson et al., 2016). Africa has several at-risk ecoregions, particularly in Angola, South Africa, the DRC, and West Africa’s Sahel region. The IUCN Red List of Ecosystems (RLE, Section 8.5.1) is another example of an ecosystem-focused prioritization for conservation. While the ecosystems approach overcomes several limitations of the species approach, some conservationists argue that focussing on distinct ecosystems may, in itself, be detrimental, and that the scope of conservation should be expanded, for example by also including biogeographic transition zones (van Rensburg et al., 2013).
Wilderness approach
Wilderness areas are large areas where people have had little influence on the environment (relative to other areas), they have few people living in them, and are unlikely places for human development in the short term. These areas are conservation priorities because they may be the only places where animals that require large home ranges can continue to survive in the wild. Further, wildernesses can serve as controls or benchmarks for researchers to measure the effect of human disturbance on nature. The most popular way to identify wilderness areas is to identify areas without roads. While very few roadless areas remain, many of the world’s most important roadless wildernesses, some larger than 10,000 km2, are in Africa (Ibisch et al., 2016). Of concern is that, second to South America, Africa also leads the world in wilderness losses over the past decade (Potapov et al., 2017). It is worth emphasising that even wilderness areas have had a long history of human activity (Roberts et al., 2017). It is not always necessary or even possible to eliminate all human activity from such areas, if those activities do not obstruct conservation goals.
Hotspot approach
Multiple prominent initiatives have prioritised conservation in areas where large concentrations of species can be protected in a relatively small area. Perhaps the most prominent example is the Global Biodiversity Hotspots initiative. Combining a species approach with an ecosystem approach, Global Biodiversity Hotspots are areas with exceptionally high levels of biological diversity and endemism—that is, irreplaceable biodiversity—that are threatened with imminent habitat destruction (Table 13.2). Norman Myers, a British biologist who launched his conservation career as a wildlife photographer in Kenya, originally proposed the Biodiversity Hotspot concept (Myers, 1988). Working with a team of prominent scientists, Myers identified 25 Hotspots (five of them in Sub-Saharan Africa), which contained 44% of all vascular plant species and 35% of all terrestrial vertebrate species on only 1.4% of the Earth’s land surface (Myers et al., 2000). More recently, Conservation International (CI) identified an expanded set of 36 Biodiversity Hotspots (Mittermeier et al., 2005), eight of which are in Sub-Saharan Africa (Figure 13.3). This expanded set of Biodiversity Hotspots covers only 2.3% of Earth’s surface yet contains over 50% of all plant species and over 40% of all terrestrial vertebrate species.
Table 13.2 A natural history comparison of Sub-Saharan Africa’s eight Global Biodiversity Hotspots.
Location
Original extent (× 1,000 km2)
Remaining undisturbed vegetation (%)
Number of species
Plants
Birds
Mammals
Guinean Forests of West Africa
620
15
9,000
917
390
Succulent Karoo
103
29
6,356
225
75
Cape Floristic Region
90
20
9,000
320
127
Maputaland-Pondoland-Albany
274
25
8,100
631
202
Coastal Forests of Eastern Africa
291
10
4,050
633
198
Eastern Afromontane
1,018
11
7,600
1,300
490
Indian Ocean Islandsa
601
10
13,500
503
211
Horn of Africa
1,659
5
5,000
697
220
Source: Mittermeyer et al., 2004; https://www.cepf.net/our-work/biodiversity-hotspots.
a Includes Madagascar and Mascarene islands
While the Global Biodiversity Hotspots highlight some of the most important global conservation priorities, none of these Hotspots are small enough to be contained in a single protected area—in fact, most of these Hotspots identify whole regions, not projects, requiring conservationists to still make decisions for prioritising protection within them. To create actionable priorities from within regional hotspots, several initiatives aim to identify local hotspots of species richness that can be conserved as one protected area of a manageable size. One such approach is the Key Biodiversity Areas (KBA) program (Eken et al., 2004), which identifies conservation priorities using standardised criteria and thresholds that account for concentrations of threatened species and/or globally significant population aggregations. The KBA program functions as an umbrella designation for several taxon-specific approaches, most prominently BirdLife International’s Important Bird and Biodiversity Areas (IBA) program (Fishpool and Evans, 2011). Other KBA programs include PlantLife International’s Important Plant Areas program (e.g. Smith and Smith 2004), as well as the Important Sites for Freshwater Biodiversity program (Darwall et al., 2005). One example from Guinea used KBA criteria and thresholds regarding threatened mammals to provide suggestions for expanding the country’s protected areas network (Brugiere and Kormos, 2009).
Gap analysis approach
Assessing the performance of existing protected areas can be done by spatially comparing their footprint to prioritised conservation areas (as above). Such an assessment offers not only an assessment of existing protected areas performance, but also offers a means to identify conservation gaps—important areas that still need to be protected to meet broader conservation goals. Such assessments, which systematically evaluate whether different aspects of biodiversity are adequately protected, are collectively known as systematic conservation planning assessments (McIntosh et al., 2017). Perhaps the most popular systematic conservation planning method is gap analysis, during which scientists overlay maps of species (or ecosystem) distributions with maps of protected areas to identify species (called gap species, see also Figure 10.3) or ecosystems that are not adequately protected in existing protected areas networks (Box 13.2).
Gap analysis enables conservation planners to identify species or ecosystems that are not adequately protected in existing protected areas networks.
Box 13.2 Identifying Key Sites for Conservation in the Albertine Rift
Andrew J. Plumptre12
1Albertine Rift Program,
Wildlife Conservation Society,
Kampala, Uganda.
2Current Address:
Key Biodiversity Area Secretariat,
c/o BirdLife International,
Cambridge, UK.
[email protected]
The Albertine Rift is one of the richest regions on Earth for vertebrate diversity (Figure 13.B). Spanning about 100 km either side of the international border of the eastern DRC, it includes forests, wetlands and savannahs from eastern DRC and western Uganda, Rwanda, Burundi, and Tanzania, and runs from the northern end of Lake Albert to the southern end of Lake Tanganyika. It contains more than 40% of Africa’s mammals, 52% of Africa’s birds, as well as 19% of its amphibians and plants, in only 1% of the continent’s surface area. It also contains more endemic and globally threatened species than any other ecoregion in Africa (Plumptre et al., 2007). Endemic large charismatic species include the eastern gorilla (Gorilla beringei, CR), golden monkey (Cercopithecus kandti, EN), Congo bay owl (Phodilus prigoginei, EN), and Ruwenzori turaco (Ruwenzorornis johnstoni, LC). The lakes in the Albertine Rift each also contain several hundred unique fish species. Unfortunately, this rich biodiversity also occurs in one of the most densely populated parts of Africa, and the threats to existing protected areas are high.
The Wildlife Conservation Society (WCS) has been working to support the conservation of six key landscapes in the Albertine Rift (ARCOS, 2004), particularly focusing on (a) identifying critical areas for conservation of threatened and endemic species; (b) undertaking research and monitoring of species and key landscapes; and (c) supporting the conservation of critical sites and the creation of new protected areas to conserve large and small mammals, birds, reptiles, amphibians and plants in all protected areas, as well as sites where new protected areas might be established. These surveys have identified critically important areas in eastern DRC, such as the Itombwe and Kabobo Massifs where new species have been identified and some species were rediscovered, having been last seen more than 50 years ago. Working with local communities, the surveys have been used to design the boundaries of newly established protected areas to ensure that they capture as much of the biodiversity as feasible. Once the local people in the area are presented with survey results and options for protection discussed, they often realise the importance of their site and propose more stringent protection measures than conservationists initially thought possible.
Using species distribution models (SDM) of the region’s endemic and globally threatened species, WCS gained an understanding of where these species should occur both now and under future climate change scenarios (Ayebare et al., 2018). Using Marxan software (Possingham et al., 2000), WCS then identified those areas that would conserve all the species of conservation interest at minimum cost (Plumptre et al., 2019). This procedure identified the Itombwe and Kabobo Massifs together with the Sitebi Hills east of Mahale Mountains National Park in western Tanzania as being critical for conservation of species that are currently not adequately protected (Figure 13.C).
These results were used to develop an Albertine Rift Action Plan (Plumptre et al., 2016), together with detailed conservation action plans for the preservation of the six core landscapes and their unique and threatened species, both inside and outside of protected areas, now and into future.
When identifying conservation gaps, it is important to think carefully about the taxa or ecosystem used to make the assessment. Many conservation assessments assume that one well-known species group can act as a biodiversity indicator (also known as a biodiversity surrogate or surrogate species) for lesser-known taxa, so establishing a protected area to protect one gap species will also afford protection to other under-protected taxa. While this is true to some level, several studies have shown that this may not always be the case (Rodrigues and Brooks, 2007; Carwardine et al., 2008; Jones et al., 2016).
Optimization Approach
Prioritisation efforts typically need to consider multiple factors in addition to biodiversity, such as cost-effectiveness, socio-economics, site condition, and potential threats that may impact a proposed protected area. Technical computer software known as “decision support tools” are providing a new way to identify conservation priorities that meet a suite of conservation objectives. One of the most popular packages is Marxan (http://marxan.org), a freely available program that identifies the optimal location for protected areas based on flexible user-defined criteria (Watts et al., 2009). The user-defined criteria can be complex; for example, one can set the model parameters to choose the areas that best protect certain aspects of biodiversity (e.g. protect at least 25% of each vegetation type) while reducing costs and minimising impact on other stakeholders; model input can include measured data, as well as expert input. In one such example, conservation biologists from South Africa, Eswatini, and Mozambique used Marxan to identify potential locations for new protected areas in the Maputaland Centre of Endemism which the three countries share. They found that adding 4,291 km2 to the existing protected areas network could generate US \$18.8 million in revenues while fulfilling their conservation objectives: protecting 44 landcover types, 53 species, and 14 ecological processes (Smith et al., 2008).
Decision support tools help identify conservation priorities that meet a suite of objectives, including cost-effectiveness, socio-economics, and site condition.
Regardless which prioritization approach one follows, it is important to remember that prioritising species and ecosystems in need of protection does not amount to “doing conservation”. Real conservation only happens when a conservation plan that will implement those suggestions is drawn up and put in place. A review of eight different systematic conservation assessments in South Africa provides a good foundation to guide conservation biologists in the process from prioritization to implementation (Knight et al., 2006). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.03%3A_Prioritization_-_What_Should_be_Protected.txt |
As of mid-2019, there were just over 7,500 protected areas covering over 4.5 million km2 of land and ocean surface (UNEP-WCMC, 2019) scattered across Sub-Saharan Africa (Figure 13.4). The country with the largest number of protected areas is South Africa with over 1,500 protected areas, while the country with the largest total area under protection is Tanzania, with over 360,000 km2. While these statistics may seem impressive, seeing these numbers in perspective is important before performance is judged. Currently, one of the most prominent sets of targets used to measure conservation progress is laid out in the international Aichi Biodiversity Targets (https://www.cbd.int/sp/targets). The global conservation area target reads:
By 2020, at least 17 per cent of terrestrial and inland water areas and 10 per cent of coastal and marine areas, … are conserved … and integrated into the wider landscape and seascape.
While Sub-Saharan Africa as a region is well on its way to achieving its goal of protecting 17% of terrestrial areas, the percentage of land protected is very uneven among countries.
The good news is that as a region, Sub-Saharan Africa is well on its way to achieving the Aichi terrestrial target, since just under 17% of the region’s total land and inland water surfaces were protected as of mid-2019 (UNEP-WCMC, 2019). Further good news is that 22 Sub-Saharan African countries have protected more than 17% of their land area, with Seychelles (42%), Republic of the Congo (41%), and Tanzania (38%) leading the way. Sub-Saharan Africa’s protected areas network is also one the best performers globally in affording protection to migratory birds (Runge et al., 2015) and terrestrial megafauna (Lindsey et al., 2017).
Despite this progress, some notable gaps remain. Foremost, the percentage of land protected is very uneven among countries. While a few countries have reached the Aichi protected areas target, there were also 10 countries with less than 5% of their land protected, and an additional six countries which protect less than 10%. Furthermore, the amount of land protected does not necessarily translate to adequate protection for all ecosystems (Watson et al., 2016). For example, despite having the most protected areas, South Africa protects only 8% of its land, well below the Aichi target. Many protected areas also qualify as paper parks (Tranquilli et al., 2012, 2014), with a questionable contribution towards achieving conservation goals.
A neglected system: marine protected areas
When thinking about conservation in Africa, many people’s minds will wander towards images of a charismatic terrestrial mammal, like an elephant, lion, or gorilla. But what about all the whales, dolphins, sea urchins, starfish, nudibranchs, and other wonderful marine creatures? Perhaps due to the outsized influence of Africa’s famous land mammals on the ecotourism sector, Africa’s marine conservation efforts have always lagged behind terrestrial conservation efforts. In total, just over 700,000 km2 (7%) of Sub-Saharan Africa’s marine environment is protected (UNEP-WCMC, 2019). The gaps in marine conservation are even more obvious when one considers that as of mid-2019, only six countries have achieved the 10% Aichi Target, with Gabon (29%) and St. Helena (28%) leading the way. Marine protection is particularly lacking along the Atlantic coast (Klein et al., 2015), where many of 15 coastal countries protect less than 1% of their coastal and oceanic waters. It is also worth keeping in mind that the 10% coverage target (a modest goal that many countries may fail to achieve), may not be enough to achieve key conservation and sustainable development goals (Spalding et al., 2008). For example, to reverse declining commercially important fish populations, it is estimated that as much as 30% of the marine environment may need to be protected (O’Leary et al., 2016).
There is clearly an urgent need to establish more marine protected areas (MPAs), protected areas within oceanic and coastal environments (Box 13.3). There is also an urgent need to scale up law enforcement in the marine environment (Brashares et al., 2004). Increasing our marine protection efforts—which even local communities can initiate (Rocliffe et al., 2014)—is well worth it: it strengthens local fisheries (Kerwath et al., 2009; Lester et al. 2009) and offers educational and recreational opportunities, such as swimming and diving, which in turn generates ecotourism revenue. For example, Africa’ oldest MPA, Tsitsikamma National Park in South Africa (established in 1964), attracts over 170,000 visitors each year (Chadwick et al., 2014); the tourism revenues support numerous jobs and are a major stimulant of the local economy (Oberholzer et al., 2010). This is in stark contrast to the marine environment off West Africa, where unregulated fisheries are putting tremendous strain on local economies amid a lack of ecotourism infrastructure (Agnew et al., 2009; Gremillet et al., 2015).
Box 13.3 Marine Protected Areas in East Africa and the Western Indian Ocean
Abraham J. Miller-Rushing
Acadia National Park, US National Park Service,
Bar Harbor, ME, USA.
How can MPAs in the Western Indian Ocean best enhance the preservation of biodiversity and the economies in this Global Biodiversity Hotspot? The ecosystems of the East African coast and nearby islands are diverse—mangrove forests, river deltas, coastal lagoons, rocky shores, sandy beaches, coral reefs, mud flats, seagrass beds, and open water. These areas are also economically important, with millions of people dependent on these waters’ shrimp, fish, and other natural resources for their livelihoods.
How effective are these MPAs, both in protecting biodiversity and people’s livelihoods? In 2006, an assessment of eight MPAs in Kenya, Tanzania, and Seychelles found several shortcomings, including inadequacies in staffing, funding, stakeholder engagement, and articulation of goals and management practices. Also, there needed to be additional monitoring and research to inform management and policy (Hockings et al., 2006). Despite these faults, the abundance and size of fish increased dramatically in several MPAs within 10 years of implementing fishing restrictions (McClanahan et al., 2007). The size and quality of fish caught in surrounding fishing grounds also increased substantially, probably due to fish dispersing from the MPAs.
Following these successes, the number and management of MPAs in the area have steadily increased and improved, at least, in part, due to cultivating better relationships with local stakeholders. One such example comes from the Quirimbas archipelago, just off the coast of northern Mozambique, where the Quirimbas National Park (over 1,000 km2) is managed through a cooperative effort of 40 villages, the government of Mozambique, and WWF. At the northern end of the Quirimbas archipelago, a few kilometres north of Quirimbas National Park, the Vamizi Conservation Project (Figure 13.D) protects an additional 230 km2 around the islands of Vamizi, Rongui and Macaloe. The Vamizi Project was initiated in 2002 as an innovative community-based management project involving local communities, international NGOs, and a group of individual investors. After protection, fish populations quickly began rebounding and had positive spill-over effects on fish around the reserve (da Silva et al., 2015). The stories of the abundant fish have contributed to a challenge for the project—attracting commercial fishermen from outside the area. To help ensure the financial and scientific sustainability of the project, partners developed a luxury ecotourism site and a research centre on Vamizi Island.
Other protected areas have met variable degrees of success, as conservation managers and communities in the region test different approaches and figure out how best to sustain MPAs in a challenging environment. Different approaches are likely to work in different situations, depending on availability of resources, local stakeholders, and other constraints. As MPAs in the region continue to develop, coordination among countries could improve the value of the MPAs to biodiversity conservation. Already there are examples of multiple pathways to improving and expanding MPAs to protect biodiversity and achieve sustainable fisheries in this region (McClanahan et al., 2016). The future is hopeful. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.04%3A_How_Much_Land_Should_We_Protect.txt |
The unplanned way in which protected areas have historically been established means that their design may at times impede rather than aid their goals. For example, many protected areas are too small to sustain viable populations of the species they are meant to protect. To avoid and mitigate such mistakes, conservation biologists are increasingly exploring methods to design more efficient protected areas networks.
Conservation biologists often start the process of designing protected areas networks by considering “the four Rs”:
• Representation: A network of protected areas should protect as much of the biodiversity (including species, ecosystems, genetic diversity, ecosystem processes, etc.) of a region, country, or subcontinent (depending on the scale of planning) as possible.
• Resiliency: Protected areas should be large enough that they can maintain biodiversity (including species, ecosystems, genetic diversity, etc.) for the foreseeable future, including in the face of climate change and natural disasters such as cyclones/hurricanes and uncontrollable wildfires.
• Redundancy: A network of protected areas should not rely on a single protected area to conserve key aspects of a region’s biodiversity; rather important aspects of biodiversity should be included in multiple protected areas to ensure their long-term existence.
• Reality: Each protected area requires sufficient funding, political will, defensibility, and local buy-in to support biodiversity over the long term.
In addition to the four Rs (which can also be applied to species protection), the following questions can also help guide planning of protected area networks (Figure 13.5):
• How large of an area must be protected and what landscape features must it include to effectively and sustainably protect biodiversity that may not be able to persist outside it?
• Is a single large protected area better, or are multiple smaller reserves more effective?
• What shape should a protected area be?
• When creating multiple protected areas, should conservation managers create them near one another or far apart? Should they be connected in some way, or should they be isolated from one another?
• How should human activities be accommodated? What activities should be allowed?
To prepare readers for discussions with land managers, conservation planners, and policy makers who are in the process of developing new protected areas, the next section introduces some of the most important principles related to protected areas design. It is important to note that this introduction is not meant to serve as a universal set of rules for the design of protected areas. Because every project presents a special and unique set of circumstances, a single set of simplistic or overly general guidelines will not suffice. Also, the principles discussed below have been explored mainly in terms of protecting terrestrial vertebrates, vascular plants, and large invertebrates, so it is still uncertain how they apply to freshwater and marine nature protected areas.
What size should a protected area be?
The design of protected areas, and their size, is often addressed through the lens of the island biogeography model that states that large islands can accommodate more species and larger populations than small islands (Section 5.1). Research on extinction rates of populations (Newmark, 1996; Woodroffe and Ginsberg, 1998) and species richness (Harcourt et al., 2001; Brashares et al., 2001) has shown that protected areas function very much like islands. Specifically, because large protected areas contain greater habitat diversity than small protected areas, larger protected areas can accommodate (a) more species, (b) a larger range of ecosystem processes, and (c) viable populations of large species that range over large areas and live at low densities.
Large protected areas are generally preferred over small ones because they can support a greater variety of ecosystems and larger wildlife populations.
Given the range of costs and benefits of establishing large protected areas, conservation biologists have debated whether creating a single large reserve or several small reserves of the same total area—known as the SLOSS (Single Large Or Several Small) debate—is better. As discussed in Section 5.1.1, habitat fragmentation is currently one of the main drivers of species extinctions; it divides large populations into more vulnerable subpopulations, leads to undesirable edge effects, creates barriers to dispersal, and provides entry points for invasive species. These negative impacts are also of concern for protected areas, especially those that are small and fragmented (leading to larger perimeter:area ratios). For example, fragmentation concentrates elephants (Vanak et al., 2010) and apex predators (Cozzi et al., 2013) in the core of protected areas, greatly limiting the effective protected area for these taxa. However, these same impacts do not alter ungulate foraging (Kiffner et al., 2013), leading, potentially, to overgrazing near reserve borders. Studies have also shown how wildlife experience higher levels of mortality near protected area boundaries (Balme et al., 2010). Ignoring such edge effects could disrupt the long-term conservation value of a protected area, particularly small ones that could effectively function as edge habitat in its entirety. Because one big fragmented reserve has many of the characteristics of several small protected areas, conservation planners should aim to establish properly-placed large protected areas and to keep them as intact as possible. It is thus good practice to restrict and even remove highways, fences, farms, logging operations, and other human activities inside protected areas because of how they fragment habitats and reduce habitat availability overall.
But how do we know when a protected area is big enough? Ultimately, optimal size depends on the area over which important natural processes take place, which varies depending on the ecosystem. In some cases, the functional size may be quite small, such as a desert spring, a mountain bog, or a rocky outcrop. In contrast, the functional size of tropical forests, seasonal drylands, and desert communities are typically quite large, possibly spanning across country borders. Understanding and planning for protecting these different targets thus requires a familiarity with the functioning and ecology of each ecosystem.
When considering the size of a proposed protected area, conservation managers must also consider how well the area can be monitored and defended from threats. In some instances, an entire community may be incorporated into a relatively small protected area that is easy to monitor and defend against pollution, invasive species, and so forth. More often however, only a portion of the target community can be protected. In such cases it is important to consider how secure the conservation target will ultimately be. For example, if an aquatic organism needs protection, clearly the protection of its immediate habitat is critical. However, if a major threat is upstream from its habitat, then protection of the immediate habitat alone will be insufficient. Instead, managers would need to find ways to prevent outside threats from impacting populations inside the protected area. One option could be to discuss the threats and how to mitigate them with surrounding landowners, perhaps by facilitating their adoption of sustainable land-use practices. If the magnitude of the threats cannot be reduced to acceptable levels, a prioritization program might be used to identify critical sub-components of a larger ecosystem that will still accomplish the necessary protection. These kinds of considerations can become very complex and involved. But they are also very important to consider as options, especially when dealing with ecosystems situated between a variety of stakeholders.
Zoning as a solution to conflicting demands
While the general consensus seems to be that larger protected areas are better than smaller ones, establishing a properly-placed large protected area can be an imposing challenge. In a few special cases, large protected areas may be established through cooperation between multiple levels of society. More often, however, conservation biologists are faced with limited resources, and stakeholders can reasonably ask why a large park is required in an area that can otherwise be used to support agriculture or other types of businesses that may provide quick profits and jobs.
One way to deal with such conflicting demands while still achieving the target of protecting a large area is through a method called mixed-use zoning. Mixed-use zoning prioritises the overall conservation objectives for a protected area but also sets aside designated areas where certain regulated human activities are permitted (Box 13.4). In this way, some areas (or zones) may be designated for subsistence agriculture, shade-grown crops, timber production, hunting, ecotourism, or water management. Other areas are designated are dedicated to recovery of threatened species, ecotourism, ecosystem restoration, and scientific research. This is the case at the W-Arly-Pendjari (WAP) Complex, which straddles the border zone between Benin, Burkina Faso, and Niger. The core of the complex consists of three national parks covering 14,948 km2, set aside for strict biodiversity conservation. These national parks are surrounded by as many as 16 additional reserves, partial reserves, and hunting concessions, bringing the total area of protected Sudano-Sahelian savannah to 26,000 km2 (WHC, 2018).
Mixed-use zones sets aside areas for certain regulated human activities within a larger conservation area. This approach helps abate conflicting land use pressure.
Box 13.4 Zoning: Something for Everyone in the Forests of Dzangha-Sangha
Richard Carroll1,2
1World Wildlife Fund,
Washington DC, USA.
2Current address:
The Pimm Group, Nicholas School of the Environment,
Duke University, Durham, NC, USA.
[email protected]
Located in the dense forests of southwestern Central African Republic (CAR), in a wedge between neighbouring Cameroon to the West and the Republic of the Congo to the East, the Dzanga-Sangha Project (DSP) aims to conserve CAR’s last lowland tropical forest by integrating conservation and regional development. The DSP, which formally began in 1988 with the establishment of a cooperative agreement between WWF and the CAR government, is an integrated conservation and development project (ICDP); its activities are focused on protected area management, rural development, tourism, and project administration, as well as sustainable use of natural resources and applied ecological and social research. The focal area of the DSP is the Dzanga-Sangha Complex of Protected Areas (Figure 13.E), an area of 4,589 km2 comprising the Dzanga-Sangha Special Reserve (3,359 km2) and Dzanga Ndoki National Park (1,143 km2). The Complex is home to healthy populations of forest elephant (Figure 13.F), western lowland gorilla (Gorilla gorilla gorilla, CR), chimpanzee, and other wildlife characteristic of the Northwest Congolian Moist Lowland Forest (Carroll, 1992). The forest also shelters the BaAka Pygmies, a hunter-gatherer community whose traditional livelihood is directly linked to the forest and its plant and wildlife resources (Robinson and Remis, 2014).
Many of the WWF-supported programs in Central Africa have sought to create the conditions for traditional peoples, such as Pygmies to maintain their lifestyles, and to adapt to changing social conditions should they choose. In the case of the DSP, two-thirds of the Complex area is classified as a “Special Reserve”, a designation that the CAR government created to accommodate traditional peoples’ use of the forest. While traditional hunting and gathering are broadly allowed in the Special Reserve, national laws specifically prohibit hunting of “integrally protected species”, such as gorillas, chimpanzees and elephants, in the Complex and elsewhere in CAR.
To establish a “safe zone” where wildlife can reproduce away from human pressures (Blom et al., 2004), and to accommodate tourism, one third of the Complex is designated as a national park. Hunting is not allowed in the national park; as compensation, 40% of all tourist receipts go to a village association, which includes BaAka, and 50% pays salaries for local employees of the park and special reserve. In other words, 90% of the dividends earned from tourism activities goes to the local people affected by conservation activities. The local community by and large supports the designation of this no-hunting zone, both to sustain their traditional activities and those of tourists.
Building on the successes in CAR, the DSP is also an active partner in the 36,000 km2 transboundary Sangha Tri-National (STN) initiative. Reflecting the Peace Park concept, the initiative is a multi-national effort to protect a large block of contiguous forests, the heart of which lies at the meeting point of the Congo-CAR-Cameroon boundaries. This initiative includes CAR’s Dzanga-N’Doki National Park, as well as two adjacent national parks: Cameroon’s Lobéké National Park (430 km2) and Republic of the Congo’s Nouabalé-Ndoki National Park (4,190 km2). These three national parks are surrounded by extensive buffer zones that include the Dzanga-Sangha Special Reserve, forests around Lobeke (700 km2) and the peripheral zone in Republic of the Congo with almost 12,000 km2 of logging concessions. STN was declared as the first landscape level World Heritage Site in 2012.
In summary, the Dzanga-Sangha Project is an ambitious, long-term effort of the CAR government, WWF, and other participating partners to save the largest and most biologically diverse tract of forest remaining in the region. Moreover, the evolution of the STN initiative demonstrates the shift from site-focused conservation to a more eco-regional or landscape strategy that incorporates the impact of human activities and the movement of animal populations across international boundaries.
Through its Biosphere Reserves program, UNESCO has pioneered a formal zoning approach that integrates human activities, scientific research, biodiversity conservation, and tourism at a single location (Coetzer et al., 2014). A biosphere reserve is divided into three zones to delineate different levels of human use (Figure 13.6). The core of a typical biosphere reserve is a no-take zone (also called a core zone), strictly protected for biodiversity and ecosystem functioning. Around the core area is a restricted-use buffer zone, where people can conduct traditional, low-impact activities, such as collecting edible plants and small amounts of wood for fuel, and scientists can conduct non-destructive research. Outside of the buffer zone is a transition zone that allows some sustainable development (such as small-scale farming) and some medium-impact natural resource extraction (such as selective logging and fishing). As of mid-2019, there were 73 UNESCO Biosphere Reserves in 26 different Sub-Saharan African nations (http://www.unesco.org/new/en/natural-sciences); new reserves are regularly being added.
Zoning is also proving effective in resolving conflicting demands over marine environments. Like terrestrial biosphere reserves, zoned MPAs consist of core zones where marine organisms can escape and recover from human disturbances, and multiple-use zones where activities such as harvesting of natural resources are permitted. Of course, harvesting fish and other marine species is not the only human activity that needs to be regulated. For example, many marine organisms are sensitive to anthropogenic noise, which interferes with communication and other important behaviors (Shannon et al., 2015). Creating multiple types of multiple-use areas can allow for different intensities of human activities. The is well illustrated at Eritrea’s Sheik Said Marine National Park; here, only approved scientific research is allowed in the restricted zone, low-impact ecotourism activities such as snorkelling and spiritual activities are allowed in the sanctuary zone, while noisy motorboats and limited take are allowed in the habitat protection zone (Habtemariam and Fang, 2016).
Zoned marine protected areas include core zones where marine organisms can escape human disturbance, and multiple-use zones where certain activities are permitted.
While resolving conflicting demands for space, zoning also provides benefits to biodiversity and people. For example, when compared to nearby unprotected sites, zoned MPAs typically have greater total weight of commercially important fish, greater numbers of individual fish, and greater coral reef cover (Lester et al., 2009). Conditions that allow marine organisms within MPAs to thrive, in turn, create opportunities for fish and other sea creatures to spill from the MPA into adjacent unprotected areas where they can be caught by local fishers, with a goal of a more sustainable harvest overall. A study from South Africa evaluated this hypothesis by attaching radio transmitters onto several white stumpnose (Rhabdosargus globiceps VU), an important fish for both commercial and recreational fisheries (Kerwath et al., 2009). This study showed that tagged fish spent 50% of their time outside the MPA, which would make them theoretically available to fishermen half of the time, while fish that did not leave protected waters produced offspring that could later disperse into multiple-use areas.
Despite the clear benefits of zoning, enforcing restrictions remains a major challenge. Even with good public outreach efforts and the threat of fines, harvesters of natural resources may still move toward and sometimes even into restricted zones to access more abundant or accessible natural resources. The resultant overharvesting at the margins of protected areas may prevent wildlife from dispersing into a wider area, which then make it hard for people that abide by the rules to access natural resources. The primary challenge in zoning is thus to find a compromise that the various stakeholders are willing to accept, and that provides opportunities for the long-term sustainable use of natural resources. Once those compromises have been agreed upon, a combination of local involvement, public outreach, clear posting of information signs, and visible enforcement of zoning restrictions can significantly increase the success of a zoning plan.
Connectivity among protected areas
Although large protected areas are preferable to smaller ones, sometimes small protected areas are the only available option, and conservation biologists must figure out how to protect biodiversity in these small areas. This is important in an African context, where most protected areas are very small, and only very few are sufficiently large to truly fulfil biodiversity needs (Table 13.3). To help conservation biologists meet this challenge, there is a growing body of evidence showing that small protected areas, even ones less than a hectare, can in fact be effective at maintaining viable wildlife populations. But how can that be? Does it suggest that small conservation areas are also useful for conservation purposes?
Table 13.3 A size comparison of Sub-Saharan Africa’s 10 largest protected areas.
Name
Location
Size (km2)
Established
Prince Edward Island Marine Protected Area
South Africa
181,230
2013
Termit & Tin Toumma National Nature and Cultural Reserve
Niger
97,000
2012
Ouadi Rimé-Ouadi Achim Faunal Reserve
Chad
77,950
1969
Air and Ténéré Reserves
Niger
77,360
1988
Central Kalahari Game Reserve
Botswana
52,800
1961
Namib-Naukluft National Park
Namibia
49,768
1979
Borana Controlled Hunting Area
Ethiopia
45,366
1973
Selous Game Reserve
Tanzania
44,000
1905
Ngiri-Tumba-Maindombe*
DRC
65,696
2008
Okavango Delta system*
Botswana
55,374
1996
* Ramsar wetlands
One of the main reasons why some wildlife populations can persist in small protected areas is that these areas violate an important assumption—that protected areas are isolated from one another. But we now know that wildlife populations often disperse between protected areas through the surrounding habitat matrix (Pryke et al., 2015). This dispersal maintains both metapopulation dynamics (Section 11.3) and reduces the risk of deleterious genetic effects (Section 8.7.1), allowing a network of small protected areas to effectively function as one large conservation area (Wegmann et al., 2014). In contrast, reserve isolation create population sinks for wildlife meant to be protected (Newmark, 2008). Consequently, re-establishing or maintaining connectivity within protected areas networks, and particularly among small reserves, has become an important strategy for enhancing their conservation value
Landscape connectivity may enable a network of small protected areas to effectively function as one large conservation area.
Many of the strategies used to maintain and restore ecosystem connectivity (Section 11.3) can be applied to protected areas management. However, this can be challenging given that administrative boundaries seldom consider natural ecosystem boundaries (Dallimer and Strange, 2015). Consequently, many ecosystems are artificially divided between different countries, each with its own development needs and management styles. Furthermore, many border barriers meant to restrict movement of people also restrict wildlife movement.
Transfrontier conservation areas enable two or more countries to collaboratively manage a shared ecosystem for mutual benefit.
Bioregional management seeks to conserve such large ecosystems that cross political borders. One way to accomplish this is to establish a transfrontier conservation area (TFCA) (also known as Peace Park or transboundary protected area), in which two or more countries collaboratively manage a shared ecosystem for mutual benefit (Hanks, 2008; see also Box 2.2 and Box 11.3). In addition to pooling scarce resources, this cooperative management style often includes removal of human-made physical barriers such as fences to allow free movement of animals (and sometimes also people, such as pastoralists) within the TFCA (Section 11.3.1). Sub-Saharan Africa first transboundary protected area was created in 1954, with the establishment of W National Park in Benin, Burkina Baso, and Niger, so named because the River Niger is shaped like letter “W” in this area. But it was only after the creation of the Peace Parks Foundation in 1997, and the Kgalagadi Transfrontier Park 2000, on the border between Botswana and South Africa (Anderson et al., 2013), that the concept gained widespread popularity in the region.
What about small isolated reserves?
At times, there will be no other choice than to accept that a small reserve is the only option available to achieve in situ conservation. In those cases, it is certainly better to accept the challenge. For many species, especially plants, a small protected area is the only buffer they have against extinction (Wintle et al., 2019). Biologists in South Africa have also pioneered an initiative to maintain species that require large home ranges in small, isolated protected areas by artificially managing dispersal dynamics (see Box 8.3). Small reserves, especially those located in or near populated areas (see Box 14.2), can also serve as locations for public outreach, conservation education, recreation, and citizen science that can improve public engagement with nature and awareness of conservation issues (Miller and Hobbs, 2002). Lastly, in addition to serving as stepping stones (Section 11.3.1), even small protected areas in urban areas provide various ecosystem services, including mitigating the urban heat island effect and reducing flooding (Feyisa et al., 2014, see also Section 7.1.6). In each of these cases, conservation biologists must creatively consider how to replicate natural processes across a small and/or fragmented protected areas network to ensure that they function on a scale that will maintain the target populations and communities. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.05%3A_Designing_Protected_Areas.txt |
Many people today have a misconception that the job of a conservation manager is done once a protected area is established. This might have been true if nature were “in balance” (a flawed concept in today’s human-dominated world, see e.g. Pimm, 1991). However, reality is very different. In many cases, humans have modified the environment so much that important populations and ecosystem processes cannot be maintained without at least some intervention, even inside protected areas. It is also important to regulate the activities of people who enter protected areas, particularly those who feel that reserves and national parks are shared public spaces that should be open to hunting, fishing, logging, farming, or mining activities. If we ignore these threats by leaving protected areas unmanaged and regulations unenforced, the biodiversity they are supposed to protect will almost certainly be lost over time.
Protected areas management should ideally be guided by a carefully-designed management plan assembled and regularly reviewed by a team of experts.
Every single protected area on Earth requires some form of management to be effective. Ideally, a protected area’s management is guided by a carefully-designed management plan assembled and reviewed by a team of experts (Henschel et al., 2014). While the details of each protected area’s management plan will be different, important aspects to address include monitoring and maintaining complex and adaptive ecosystems (Chapter 10), managing threatened species (Chapter 11), and providing resources, training, and memorable experiences to local people and visitors (discussed below). Management plans should also address which activities are prohibited (e.g. hunting or campfires) which activities are encouraged (e.g. wildlife photography, citizen science projects), and how rules and regulations will be enforced (Chapter 12). Lastly, the best management plans have a system in place to ensure that goals and activities are regularly reviewed and updated to account for new knowledge and experiences, and changing priorities.
In some protected areas, particularly small ones, it may be necessary to artificially maintain conditions that enable local wildlife to persist. One such example is the maintenance of natural fire regimes in fire-adapted ecosystems (Section 10.2.1). Another example is the temporary (or sometimes permanent) supply of limiting resources, such as exposed mineral licks, carcasses for scavengers, and nest boxes for bats and birds. Conservation managers might also establish artificial water sources or plant native fruit trees to support local (or translocated) wildlife. When taking such steps, it is important to strike a balance between establishing protected areas free from human influence and creating semi-natural areas in which plants and animals become so dependent on people that their persistence is not sustainable over the long term.
Management actions are generally implemented without completely understanding how the action will influence local ecosystem processes and wildlife populations. In light of this uncertainty, and despite good intentions, it should come as no surprise that some management actions may not achieve conservation goals. Some management actions may even later show to have unintended consequences that harm local biodiversity. While some actions are easy to reverse, some may put conservation managers on a cycle of reactionary management that is hard to escape. For that reason, it is important to carefully consider both the benefits and drawbacks of a management action before implementation. It is also important to be ready and willing to adapt management protocols as and when needed (see adaptive management, Section 10.2.3).
The importance of monitoring
An important aspect of a protected area management plan involves setting up a well-designed, long-term monitoring plan to assess whether conservation goals are being met. The exact types of information gathered will depend on the goals and objectives of each protected area, but can include tracking threatened wildlife populations, monitoring ecosystem health, or evaluating whether a threat is increasing or decreasing. These assessments may involve a wildlife survey (Section 9.1), taking regular measurements of various ecosystem indicators (Section 10.1), and/or conducting regular law enforcement monitoring (Section 12.3). In recognition of the linkages between the wellbeing of people and success of conservation (Oberholzer et al., 2010; Oldekop et al., 2016; Hauenstein et al., 2019), many conservation biologists are now also combining biodiversity monitoring with monitoring local peoples’ well-being.
A protected area management plan should include a long-term monitoring plan to assess whether conservation goals are being met.
Monitoring may highlight uncomfortable realities for conservation managers. An example could be management actions that prove to harm biodiversity (discussed above). Another uncomfortable reality is when one species needs to be prioritised over another. This is the case on protected islands off Southern Africa’s west coast, where biologists have resorted to selectively culling Cape fur seals (Actocephalus pusillus, LC) that predate on three species of threatened seabirds; in one case, this predation led to the abandonment of an entire seabird breeding colony (Makhado et al., 2009). Even more problematic is when one threatened species causes significant harm to another. This is the case in Uganda’s Kibale National Park, where chimpanzees kill as much as 12% of the area’s Ugandan red colobus monkeys (Procolobus tephrosceles, EN) each year (Watts and Mitani, 2002; Lwanga et al., 2011). It is however important to not confuse sustainable levels of predation with real threats that can lead to extinction. For example, in Ethiopia, the big-headed African mole rat (Tachyoryctes microcephalus, EN) is the favoured prey of the similarly-threatened Ethiopian wolf (Canis simensis, EN). However, rather than predation by the wolves, habitat loss from agriculture and overgrazing is the most important threat to the survival of the mole rat (Lavrenchenko and Kennerley, 2016), as well as the wolf (Marino and Sillero-Zubiri, 2011).
The control of any wildlife population, even invasive species in protected areas, can become very emotional for the public. It may even give rise to animal rights advocacy groups that oppose or even impede conservation. Such is the case in South Africa, where a well-organised group of local citizens opposed the eradication of invasive Himalayan tahrs (Hemitragus jemlahicus, NT), relatives of goats, which threatened imperilled Fynbos plants in a World Heritage Site (Gaertner et al., 2016). To avoid unnecessary conflict with such citizen groups, which can quickly turn into a public relations nightmare, it is important to consider whether drastic management actions are necessary. If so, it is wise to involve and educate the public for the need of such actions at an early stage.
Because monitoring can be resource-intensive, it is important to ensure the scale and methods of monitoring are appropriate for management needs. For small reserves, tracking only a few ecosystem components during periodic site visits might be sufficient. In contrast, for large or remote protected areas, geospatial analysis with environmental data obtained through remote sensing methods (Section 10.1.1) may be a more feasible way to monitor legal and illegal human impacts, such as logging (Figure 13.7), shifting cultivation, hunting, and mining. Many protected areas are also increasingly reliant on local people, researchers, tourists, and other groups of people to contribute to monitoring, particularly through citizen science projects (Section 15.4.1).
The importance of working with local people
The future of a protected area almost always depends on the degree of support, neglect, or hostility it receives from people who may be living inside the protected area, or in the surrounding area. Local people are unlikely to support conservation areas where there is a history of mistrust or disagreement between them and conservation authorities, or where park managers have not worked with and/or discussed conservation goals with them. This is particularly true when local people have been displaced by conservation actions (Cross, 2015; Baker et al., 2012) or are victims of human-wildlife conflict (Section 14.4). Such victims will understandably be angry and frustrated and may even reject conservation regulations altogether. Escalating cycles of hostility due to enforcement efforts can even lead to outright violence, during which protected areas staff, residents, and tourists can be threatened, hurt, or even killed.
The future of a protected area depends on the degree of support, neglect, or hostility it receives from people who live inside the protected area, or in the surrounding area.
To avoid such an ugly scenario, a central part of any protected area’s management plan must be a policy to ensure that local communities value, and benefit from, conservation activities. The ultimate goal of such a policy should not only be to ensure that local people become strong supporters of conservation efforts, but that they later also willingly contribute to them. At a very basic level, this can be accomplished by developing a range of ecotourism opportunities, particularly those that encourage participation in citizen science projects (Section 15.4.1), and those that afford opportunities where the goals and benefits of a protected area can be explained to local people. South Africa’s SANParks does this by encouraging school visits and accommodating a variety of income groups through a multi-tiered fee system (Beale et al., 2013b). When conservation displaces local people or limits activities previously allowed, it might also be worth investigating whether there is room to practice traditional activities in a sustainable way. Such is the case in South Africa, where the regional conservation authority Ezemvelo KZN Wildlife allows local people to sustainably harvest plant resources, such as thatching grass and medicinal plants, from protected areas they manage (Beale et al., 2013b; see also Section 13.5.2).
The next level of involvement includes benefit sharing. This often takes the form of compensatory payments for people who have lost assets due to conservation actions (Hall et al., 2014; see also Section 14.4). Some park managers also provide educational and employment support to local communities. One example comes from Botswana’s Okavango Delta region, where employment opportunities generated through ecotourism ventures at Moremi Game Reserve greatly improved relationships between local communities and park managers (Mbaiwa and Strongza, 2011; see also Section 14.3). African Parks, who manages 15 national parks across 10 African countries, have made local involvement (Figure 13.8) and community development a core part of their mission, which they accomplish by constructing schools, facilitating entrepreneurship, and funding healthcare services. The third level of involvement involves co-management partnerships, where local people directly participate in park management and planning (discussed in Section 13.1.4).
The importance of accommodating visitors
Developing plans that accommodate outside visitors is also an important aspect of protected areas management. Tourists are some of the most important outside visitors to attract. Their spending stimulates local economies, and provides funds for salaries, maintenance, and other conservation initiatives (Ferraro and Hanauer, 2014). When tourism activities are combined with citizen science projects (Section 15.4.1), visitors can also contribute to monitoring, further expanding the capacity of protected areas staff. Accommodating university students and other researchers is also important, as they could provide valuable information to park managers and training to staff at a steeply discounted price, compared to work by expensive outside consultants who may not always understand local dynamics.
While ecotourism provides opportunities for employment, income, and monitoring, it is important to manage the multiple threats introduced by visitors.
While visitors provide significant benefits, it is important to monitor harmful elements they may knowingly or unknowingly introduce (Buckley et al., 2016). For example, visitors may introduce invasive species (Spear et al., 2013; Foxcroft et al., 2019) or induce behavioral changes in the animals they observe (Geffroy et al., 2015). Visitors may also directly damage protected ecosystems: frequent boating and diving among reefs can degrade reef communities when divers’ flippers, boat hulls, and anchors crush fragile corals. Visitors may even kill wildlife directly when they trample wildflowers, disrupt nesting birds, collide into animals that are crossing roads, or spread diseases to wildlife (Ryan and Walsh, 2011). When visitor activities are restricted, especially previously-allowed activities, park managers need to be able to explain reasons for the current policies and ensure that reasonable alternatives are available. For example, if the number of tourists visiting a special wildlife spot must be restricted to prevent damage to a site, the tourists could be offered the chance to visit a different site or participate in another activity.
The IUCN Green List of Protected Areas
A challenge that park managers frequently face is objectively determining how well their protected areas are managed. While profit margins, tourist numbers, species diversity, and population indices offer some form of evaluation criteria, it is not a fool-proof system: some well-managed protected areas are not very accessible to tourists, while carelessly increasing species richness or wildlife populations will likely have detrimental consequences. Tools such as the Management Effectiveness Tracking Tool (Stolton et al., 2007), Spatial Monitoring and Reporting Tool (Moreto, 2015), and Rapid Assessment and Prioritisation of Protected Areas Management (Ervin, 2003) have helped park managers assess whether the goals of their management plans were being achieved. But with no global standard of best practices against which protected areas are objectively assessed, park managers are mostly left to evaluate success based on their own subjective criteria and goals.
To fill this gap, the IUCN recently established the Green List of Protected Areas (http://www.iucn.org/greenlist) which aims to increase the number of protected areas that are effectively and fairly managed (Figure 13.9). Nominated protected areas will be evaluated against a set of standards which attest to management structures that can achieve long-term positive impacts on biodiversity and people. This list of standards, adapted to reflect local contexts within which evaluated protected areas operate, is divided into four higher level components: (1) good governance, (2) sound design and planning, (3) effective management, and (4) successful conservation outcomes (Figure 13.10). There are even plans, through a “Fair Finance” initiative, to reward protected areas that receive Green List status by making resources available to further strengthen their accomplishments.
The Green List has only recently been established; hence, not many protected areas have been evaluated by the time this book was written. Sub-Saharan Africa’s first Green List sites were Kenya’s Lewa Wildlife Conservancy and Ol Pejeta Conservancy, both which formed part of the 2014 initial trial period. Both sites were re-certified in 2018, when Kenya’s Ol Kinyei Conservancy was also added to the Green List. Hopefully many more sites will follow suit in the near future. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.06%3A_Managing_Protected_Areas.txt |
The biggest challenges that park managers will face in the coming decades stem from a growing human population. When key natural resources, such as firewood and bushmeat, become harder to find, conflict is inevitable as more people look for new lands where they can fulfil their needs. As more people encroach into protected areas, so too will habitat loss, pollution, invasive species, and diseases. Despite conservationists’ best efforts to build collaborations with nearby communities, park managers need to anticipate that this ever-greater demand for space and natural resources will add additional challenges to their work plans. Below, we discuss three challenges that will likely continue to pose threats in future, and for which there are not always easy solutions.
Funding limitations
To enable protected areas to achieve their full potential, there must be adequate funding to support a team of well equipped, properly trained, and motivated staff (James et al. 2001; Gill et al., 2017). There is also a need for buildings, vehicles, communications equipment, and other appropriate infrastructure and resources to enable the staff to fulfil their duties, and for tourists to have a memorable time. The cost of these resources can quickly add up; for example, researchers estimated that more than \$1 billion is needed each year to manage Africa’s protected areas that include lion populations (Lindsey et al., 2018). Yet, Africa’s protected areas are frequently understaffed, lack basic equipment, and face funding shortages (Tranquilli et al., 2014; Watson et al., 2014). Without the means to travel, communicate, and protect themselves, even motivated staff may find themselves stuck at their duty stations, unaware of what is happening elsewhere in their park. Some of these challenges can be solved with an adequate ecotourism plan, which can be facilitated from the grassroots level up or government level down. A growing number of funding mechanisms, including private and international donors, have also started to fill funding gaps (Section 15.3) which, in turn, has allowed more NGOs to assist in conservation areas management (Tranquilli et al., 2012; Lindsey et al., 2014). Above all, a carefully assembled management and monitoring plan, which is adequately funded, is key to the success of protected areas.
Planning for climate change
Because protected areas are fixed in space and time, many species that are currently protected will adjust their ranges beyond the borders of existing protected areas due to climate change. One study from South Africa found that 62% of bird species will lose some degree of protection over the next few decades, with five species losing at least 85% of their protected ranges (Coetzee et al., 2009). Studies in West Africa yielded remarkably similar results, where 63% of amphibians, 63% of mammals, and 55% of bird species face decreased protection due to changing climate (Baker et al., 2015). The situation is even worse for taxa with too little protection as it is. For example, suitable habitat for only 5% of African bat species is currently protected; due to climate change, it will further decrease by 2050 (Smith et al., 2016).
To ensure the future protection of species vulnerable to climate change, we must incorporate species’ predicted distribution ranges into the planning of protected areas networks. For species that disperse easily, this requires protecting gaps in their current and future ranges (Hole et al., 2011), as well as protecting, maintaining, and restoring potential dispersal pathways (Section 11.3). For poor dispersers, conservationists could start experimenting with assisted colonisations, or identify and protect their climate refugia (Section 11.4). For many species, however, establishing protected areas in their future ranges will be nearly impossible simply because no land is available. These species will greatly depend on conservation efforts outside protected areas, which we will discuss in Chapter 14.
Facing degazettement
It may be reasonable to assume that protected areas (especially government protected areas, established by law) afford permanent protection to biodiversity on those lands. Unfortunately, that is not the case—between 1950 and 2017, at least 227 different protected areas in Sub-Saharan Africa lost (partially or fully) lost their legal protected status (WWF and CI, 2016), in a process formally known as protected area downgrading, downsizing, and degazettement (PADDD, http://www.padddtracker.org). There are a variety of reasons behind PADDDs. For example, some protected areas have been PADDDed because of environmental degradation caused by conflicting land uses, including illegal logging, illegal agriculture, and land invasions; in such cases governments (in consultation with conservation managers) may determine that the resources needed for land rehabilitation are better spent protecting other sites (Fuller et al., 2010). Others have been PADDDed because incorrect procedures were followed during establishment—in such cases, it might be prudent to carefully consider if a compromise could be reached that combines the goals of conservation and development (Section 14.3). However, the vast majority of African PADDDs are enacted because of more sinister motives, such as to undercut conservation restrictions (Mascia and Pailler, 2011). For example, when examining each threat individually, data from WWF and CI (2016) suggest that mining pressure was the leading cause of previous African PADDDs. Considering that nearly 30% of African protected areas are still earmarked for oil and gas exploration (Leach et al., 2016), the threat from mining will likely also continue in the foreseeable future (Durán et al., 2013; Edwards et al., 2013).
Mining pressure is currently the leading cause for downgrading and degazettement of African protected areas.
Most conservationists consider the PADDD process a bad precedent that should be avoided unless necessary. While there are legitimate reasons behind some PADDDs (Fuller et al., 2010), few are enacted with conservation goals in mind. In many cases, government officials remove the protected status of lands without even consulting conservation scientists and park managers. Such decisions are particularly frustrating when important areas that protect threatened species and ecosystems are affected. Combatting the continuing threat of PADDDs will depend on national and international conservation organizations partnering with vigilant citizens who take ownership of their natural treasures. Until citizenry can trust that government officials have the interests of their natural heritage at heart, protected areas PADDDs will remain a highly controversial topic. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/13%3A_The_Importance_of_Protected_Areas/13.07%3A_Challenges_for_Protected_Areas.txt |
1. Establishing protected areas is the most effective method for safeguarding biodiversity. Seventeen percent of Sub-Saharan Africa’s land surface is included in over 7,500 protected areas, with new reserves and parks regularly designated. In contrast, only 7% of the region’s marine and coastal environments are protected, with protection highly uneven among countries.
2. Government agencies and conservation organizations set priorities for establishing new protected areas based on the relative distinctiveness, endangerment, and utility of a species or ecosystems. Many protected areas are established to preserve species of special significance, unique ecosystems, wilderness areas, and concentrations of threatened species. Gap analysis is used to identify elements of biodiversity not accommodated in existing protected area networks.
3. While protected areas have previously been designed haphazardly, conservation biologists are developing guidelines for designing more effective protected areas. As general guidelines, protected areas should be large whenever possible, and should not be fragmented. Conservation planners should also aim to create linked networks of conservation areas to encourage wildlife dispersal.
4. Protected areas must be actively managed to maintain biodiversity. Monitoring provides much needed information to evaluate whether management activities are achieving their intended objectives or need to be adapted.
5. Managing interactions with local people and visitors is critical to the success of protected areas and should be part of a management plan. To obtain and maintain local support, managements plans should consider benefit sharing and co-management partnerships.
13.09: Topics for Discussion
1. Obtain a map of your region’s protected areas (e.g. nature reserves and national parks) and multiple-use managed areas (e.g. hunting and logging concessions). (https://protectedplanet.net is a good source.) If you could designate an additional protected area, where would it be? What shape would your protected area be? What would the management goals be for these additional areas? Explain all your answers.
2. Think about a protected area that you have visited. What is the main goal of this protected area? Do you think park management is succeeding in the goal? What are they doing particularly well? What could they do to manage the protected area better?
3. Think of a protected area near you that safeguards an aquatic environment, such as a beach, estuary, or lake. What unique challenges do you think the people managing that protected area face that managers of a terrestrial protected area do not face?
4. Many countries are developing protected areas that cross international borders. What are the main goals of these parks? Are they achieving their goals? What are the main challenges?
5. How can national parks continue to function optimally in countries where the central governments have largely ceased to function, and where corruption is rampant?
Suggested Readings
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Fox, H.E., M.B. Mascia, X. Basurto, et al. 2012. Reexamining the science of marine protected areas: Linking knowledge to action. Conservation Letters 5: 1–10. https://doi.org/10.1111/j.1755-263X.2011.00207.x Biological, social, and policy issues are all important in effective marine protected areas.
Gross, J.E., S. Woodley, L.A. Welling, et al. 2016. Adapting to Climate Change: Guidance for Protected Area Managers and Planners (Gland: IUCN). https://portals.iucn.org/library/node/46685 Best practices guidelines for managing protected areas under climate change.
Ihwagi, F.W., T. Wang, G. Wittemyer, et al. 2015. Using poaching levels and elephant distribution to assess the conservation efficacy of private, communal and government land in northern Kenya. PLoS ONE 10: e0139079. https://doi.org/10.1371/journal.pone.0139079 A comparison of different types of protected areas in protecting biodiversity
Knight, A.T., A. Driver, R.M. Cowling, et al. 2006. Designing systematic conservation assessments that promote effective implementation: Best practice from South Africa. Conservation Biology 20: 739–50. https://doi.org/10.1111/j.1523-1739.2006.00452.x An overview of eight systematic conservation plans from design to implementation.
Mascia, M.B., and S. Pallier. 2011. Protected areas downgrading, downsizing, and degazettement (PADDD) and its conservation implications. Conservation Letters 4: 9–20. https://doi.org/10.1111/j.1755-263X.2010.00147.x In many areas of the world, the official protection of national parks and other conservation areas is being withdrawn.
Mbaiwa, J.E., and A.L. Stronza. 2011. Changes in resident attitudes towards tourism development and conservation in the Okavango Delta, Botswana. Journal of Environmental Management 92: 1950–59. https://doi.org/10.1016/j.jenvman.2011.03.009 Local people being positive about conservation.
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Well-managed protected areas are essential tools for securing intact ecosystems and the biodiversity they sustain. However, it will not be sufficient to rely solely on protected areas to preserve biodiversity. Harming ecosystems such as rivers and streams on unprotected lands has repeatedly been shown to decrease biodiversity also within protected areas (Colvin et al., 2011; Woodborne et al., 2012). Furthermore, many species only occur on unprotected lands (Beresford et al. 2011), and some species even fare better outside protected areas (Murgatroyd et al., 2016). Other species need to move out of protected areas to access important seasonal resources: about two-thirds of Kenya’s large animals regularly move from protected areas into unprotected rangelands in search of food and water (Young et al., 2005; Western et al. 2009a). Lastly, restoring damaged areas and maintaining intact ecosystems inside and outside of protected areas provides ecosystem services such as water and air purification. And so, while we must continue to pursue protected areas for the many benefits they do offer, we must not forget about the value of the spaces between protected areas. In this chapter, we will explore how efforts on unprotected lands can complement conservation efforts in protected areas.
14.01: Human-Dominated Landscapes
In every country on Earth, significant portions of unprotected lands still harbour some of their original biota (Figure 14.1). Consider, for example, remote regions that are considered “wilderness” by governments and the general public. Most of these areas are inhabited by low-density human societies that practice a traditional way of life. With relatively little outside influence from modern technology, these traditional peoples are often dependent on—and thus highly concerned with—the health of their environment. More importantly, traditional peoples have been an integral part of their environments for thousands of years. The present mixture and relative densities of wildlife in these “wildernesses” thus reflect the historical activities (e.g. fishing, hunting, fire management, land clearing, and planting of useful plant crops) of the people living in those areas (Roberts et al., 2017). These activities do not degrade the environment if human population densities remain low and natural resources are harvested sustainably. To regulate these activities, most traditional societies have an established system of rights to natural resources, known as customary laws, which an increasing number of governments recognize (Section 12.2.2). Conservation biologists should follow this example: rather than being considered a threat to the “pristine” environments in which they live, traditional peoples should be seen as important partners in conservation efforts because protecting their lifestyles also ensures the protection of biodiversity (Box 14.1).
Traditional peoples are important partners in conservation efforts because protecting their lifestyles also ensures the protection of biodiversity.
Box 14.1 Traditional People and Conservation: Turning the Page
Abraham J. Miller-Rushing1 and John W. Wilson
1Acadia National Park, US National Park Service,
Bar Harbor, ME, USA.
How can we balance the conservation of biodiversity with the rights of traditional peoples? Conservation has a mixed history in this regard (Brockington et al., 2006). In many cases, conservation projects and traditional peoples have supported each other, but there are also many examples where vulnerable peoples have been abused and dispossessed of their ancestral lands. Here we briefly discuss three such examples from different regions of Africa: the “Pygmies” of Central Africa’s forests, the “Bushmen” (also known as San, or First People) of the Kalahari Desert, and the Maasai of East Africa’s savannahs.
Relocating local people from protected areas, either before or after establishment, is a relatively common strategy across the world. Amongst other reasons, these relocations are thought of as a strategy to reduce stress on wildlife from hunting and other forms of resource extraction which, in turn, enables park managers to reach their conservation and tourism goals quicker. However, in many cases relocated people are not adequately supported as they transitioned to new lifestyles, leading to increased poverty, declining health, and loss of identity. When vulnerable traditional peoples are relocated, humanity also loses a rich cultural heritage, including local knowledge about native wildlife and traditional medicines.
The Baka, a “Pygmy” tribe that lives in southern Cameroon, have greatly suffered from poorly executed conservation activities. This traditional hunter-gatherer community often identify themselves as “forest peoples” for their strong ties to the forests of Central Africa, which defines their history, culture, and livelihoods. Their ties are now under strain; over the past few decades, several hunting concessions and protected areas were established on ancestral Baka land without proper consultation. In addition to losing their ancestral land, there have also been multiple instances where law enforcement patrols arrested, abused, tortured, and even killed Baka community members that were suspected of illegal hunting (FPP, 2016). Regardless of the back story, it is inherently unjust for hunting to be legal for rich foreigners on ancestral lands where traditional peoples are now prohibited from doing so. This is all even truer when the people who forgo their traditional activities in the name of conservation do not receive compensation for their losses in return.
The second example involves the “Bushmen” (also known as San, or First People). For tens of thousands of years, the Bushmen have lived a nomadic or semi-nomadic hunting and gathering lifestyle in the Kalahari Desert of Southern Africa. Over the last few centuries however, the Bushmen have struggled to maintain their lifestyles as larger African tribes, and later European colonists have staked claims on their land. In 1961, the government of Botswana established the Central Kalahari Game Reserve to protect wildlife and dedicate a place for the Bushmen to practice their traditional lifestyles. However, the government also drilled wells, and established a school and health post which unintentionally encouraged the Bushmen to adopt a more sedentary lifestyle in which they grew crops and raised livestock in the reserve. In 1985, the government of Botswana decided that the Bushmen’s lifestyles were no longer compatible with the goals of the reserve. After a process that most observers, including the UN (Anaya, 2010), considered inadequate and unethical, between 1997 and 2001 the Bushmen were banned from hunting within the reserve, and forced to relocate to settlements outside its boundaries.
In 2006, the High Court of Botswana ruled that the relocations were illegal, and that the government had to allow the Bushmen to return to the reserve. The government, however, allowed only a very limited number of people to return, prohibited traditional hunting, and took other actions that failed to meet their human rights obligations (Shaheed, 2016). The government defended these actions by saying that they were protecting biodiversity, and that the hunting pressure by Bushmen amounted to poaching of protected wildlife. To outside observers, it seemed that keeping them away from a lucrative diamond mine, established in the reserve in 2014, was the stronger motivation (Haines, 2016).
The Maasai are also very familiar with land conflict in the name of conservation. They occupy an area of 160,000 km2 of seasonal lands in the Great Rift Valley in southern Kenya and northern Tanzania. Here they live a semi-nomadic, pastoralist lifestyle, and rely almost entirely on livestock for food and income. In the early 1900s, British colonists began taking lands away from the Maasai to create ranches and later also protected areas. This included the iconic Maasai Mara National Reserve and Serengeti National Park on either side of the Kenya-Tanzania border. Together these parks protect one of the last great mammal migrations in the world, attracting tourists from far and wide.
Maasai pastoralists who live adjacent to these and other protected areas regularly come in conflict with wildlife, particularly lions, which pose a threat to the herders and their livestock. The Maasai have a tradition of hunting lions, particularly after livestock predation, but also as a traditional rite of passage to manhood. Killing lions is now illegal in Kenya as the region’s lion population is dwindling from the impacts of habitat loss and too many lion hunters. As an alternative to retaliatory killings, the Kenyan government financially compensates herders for predation losses; however, the government has not always been consistent in this compensation (Goldman et al., 2010). At times, corrupt officials have also exploited the Maasai by taking or misusing their share of tourism revenues (although corruption seems to be declining and compensation and revenue sharing have improved in recent years). Also, in Tanzania, tens of thousands of Maasai were evicted from their ancestral land in the Serengeti wildlife corridor in 2013. And yet, instead of strict protection, the land in question was earmarked for a foreign-owned hunting concession (Smith, 2014). (The Tanzanian government has since reversed this land grab and fired those behind it.)
Conservation conflicts with traditional peoples are unnecessary and avoidable. Research—done by conservation biologists—shows that traditional peoples have a relatively minor impact on the environment, especially when compared to that of commercial hunters and exploited migrant labourers who have little incentive to live sustainability (Thibault and Blaney, 2003; Poulsen et al., 2009; Fa et al., 2016). Conservation biologists and traditional peoples thus face similar concerns: habitat loss and fragmentation, human encroachment, commercial hunting and other capitalist ventures, and armed militias. Such is the case even in 2019, in southern Cameroon, where the clearing of the Meyomessala forests for Chinese-owned rubber plantations not only threatens many threatened and iconic species, but also the livelihoods of the local Baka community (Sixtus, 2008). As such, working together will benefit both conservation biologists and traditional peoples.
Fortunately, there are several models illustrating the compatibility of conservation and traditional lifestyles. One example comes from northern Kenya, where the community-run Northern Rangeland Trust sustainably manages 44,000 km2 of land. Here, core areas are set aside for wildlife, while buffer areas are subjected to a rotational livestock grazing framework. This model benefits both livestock and wildlife by preventing overgrazing, increasing habitat heterogeneity, and maximising overall biodiversity, thereby allowing 480,000 people (from 15 ethnic groups) living in the area to enjoy peace, prosperity, and healthy ecosystems. Livestock and wildlife populations—which show limited competition under the right conditions (Kartzinel et al., 2015)—are also thriving, and at times they may even benefit one another (Odadi et al., 2011). A major accomplishment here is that the population size of Africa’s rarest antelope, the hirola, doubled in only three and a half years. The area now boasts a growing tourism industry, which boosts employment, and provides revenue that is spent on health, education, and other societal needs.
Some challenges remain, most notably human-wildlife conflict, and overcoming a history of mistrust. Droughts and political instability have also contributed to tensions between conservationists and traditional peoples in some areas. Even so, conservation biologists have an obligation to follow best practices and ensure that human rights are not violated through conservation activities (Borrini-Feyerabend and Hill, 2015). There is also a need to abide by international laws and to recognize the rights of people who want to live a traditional lifestyle; this includes respecting their right to free, prior, and informed consent (FPIC, see UN-REDD, 2013) before mutually-beneficial conservation actions are implemented. With more African countries adopting a western form of land tenure that enables private land ownership, empowering traditional peoples by helping them to obtain legal title (the right to land ownership that is recognized by the government) can also help establish trust and locally managed protected areas where resources are harvested sustainably (Rai and Bawa, 2013). Recent studies have shown that deforestation rates decline on such legally designated traditional lands, particularly in places undergoing rapid land-use changes (Nolte et al., 2013). Gaining trust and working with traditional peoples is not always easy. But only under those circumstances can we have any hope of establishing a sustainable conservation model that will weather the test of the times.
Communal lands in East Africa that are dedicated to pastoralism provide a good example illustrating the compatibility between traditional peoples and conservation efforts (McGahey et al., 2007). In contrast to modern livestock farming systems that maintain livestock in a restricted area, pastoralism involves regularly herding livestock to new areas in search of fresh pasture and water. Pastoralists sometimes also use fire to enhance land productivity, which, together with intense but short-term grazing, maintains native grassland and savannah ecosystems by preventing encroachment of woody plants (Section 10.2.1). By keeping their grazing lands suitable for livestock, pastoralists also maintain those areas in a state that is suitable for native biodiversity and the natural resources they need for survival. In fact, levels of biodiversity on well-managed pastoral land may rival (Msuha et al., 2012) or even exceed (ILRI, 2006) that of adjacent protected areas where such activities are excluded.
Biodiversity conservation and human welfare operate hand-in-hand on unprotected lands, but with a fragile balance that requires continued maintenance.
However, when the rules governing pastoral systems break down due to agricultural developments and fences that impede pastoralist movements, the resultant breakdown of traditional grazing systems lead to overgrazing, negatively impacting people, livestock, and wildlife (Western et al., 2009b; Groom and Western, 2013). These are the kinds of threats facing the hirola (Beatragus hunteri, CR), also known as the Hunter’s antelope, which shares much of its range with Somali pastoralists in north-eastern Kenya. With fire suppression, elephant extirpation, and a breakdown of traditional grazing systems, trees are encroaching on this facultative grazer’s last remaining strongholds; the loss of grasslands is also harm cattle production in the region (Ali et al., 2017). On pastoralist lands, biodiversity conservation and human livelihoods truly operate hand in hand, but with a fragile balance that requires continued maintenance.
People who live in rural areas and sell natural resources that they extract from healthy ecosystems also play an important role in conservation by engaging in sustainable natural resource management. Consider all the unprotected estuaries and marine areas that support commercial fisheries for a moment. When fisheries are managed in a sustainable way, not only does this benefit the people that depend on these commercially important species, but other native species can also thrive. Such a win-win outcome was illustrated in Cameroon’s mangrove swamps, where the unsustainable harvesting of mangrove trees for smoking fish was mitigated with fuel-efficient stoves. These new stoves reduced both the amount of time and firewood needed for smoking fish, benefiting the mangroves as well as the fishers’ profit margins (Feka et al., 2009). Fuel-efficient stoves form an important role in reducing firewood harvesting pressure in Cameroon and many other African countries.
People living in urban centres can also contribute to conservation efforts by raising environmental awareness among fellow citizens, participating in activism, lobbying, and fundraising activities, and generating knowledge through citizen science projects (see Box 15.3). They can also help reduce the multiple pressures that their cities exert on the surrounding environment. Among the most exciting recent developments have been the development and installation of green infrastructure, such as urban forests, green roofs, urban wetlands, permeable sidewalks, urban farms, and rain gardens (Figure 14.2). Not only does green infrastructure save money by reducing energy consumption and pollution clean-up costs, it also reduces overall maintenance (Odefey et al., 2012) and improves overall well-being (Demuzere et al., 2014). Consequently, green infrastructure is increasingly being integrated in urban planning across the world, including North America (EPA, 2018), Europe (Natural England, 2009), Asia (e.g. Kennedy et al., 2016), and in South Africa (Culwick et al., 2016).
Many urbanites are also eager to work with government agencies and conservation NGOs to make their cities more biodiversity friendly by restoring urban waterways and wetlands (Box 14.2), and replanting abandoned industrial sites and other damaged urban areas with native vegetation that can support pollinators, birds, and other wildlife. Such efforts foster neighbourhood pride, create a sense of community, and provide a sense of satisfaction to people who like to be close to nature. These restored areas, and other urban green spaces, can also serve to highlight the links between human well-being and nature, which may make those city dwellers who remain on the side-lines of conservation more receptive to the more challenging aspects of conservation, such as prescribed fire and invasive species management (Gaertner et al., 2016). Establishing and maintaining areas to protect biodiversity where people live and work, termed reconciliation ecology (Rosenzweig, 2003), will increase in importance as Africa’s urban centres continue to expand over the next decades (Seto et al., 2011).
Box 14.2 Importance of Protected Areas in Cities: Insights from the City of Cape Town
Pippin M. L. Anderson
Department of Environmental and Geographical Science,
University of Cape Town, South Africa.
[email protected]
With urbanisation rates of 3% per year, Sub-Saharan Africa’s sprawling cities are predicted to increase in area by seven times over coming decades (Anderson et al., 2014). Much of this growth will be associated with weak decentralised governance and limited resources for environmental management. Just under half of all Africans live below the poverty line, and African city-dwellers tend to make greater use of natural resources than citizens of other continents (Anderson et al., 2014). Meanwhile, biodiversity and conservation concerns tend to be eclipsed by, or be independent of, other pressures such as poverty, unemployment, and access to food, water, and housing. Despite the challenges of promoting biodiversity conservation under these conditions, sprawling cities will benefit greatly from conserving some green space within their boundaries.
The City of Cape Town in South Africa has exceptional biodiversity; the region around the city hosts some 9,000 plant species on just 90,000 km2 of land. But as urban areas spread, natural areas are being developed to meet housing and economic needs. Elsewhere, native ecosystems are being suppressed by agricultural transformation, invasive species, annual floods and coastal erosion are intensifying, and altered fire regimes are hindering natural processes. These alterations to the natural landscape have significant economic consequences due to eroded ecosystem services (O’Farrell et al., 2012) conservatively valued at US \$150–450 million per annum (de Wit et al., 2009). Similarly, urbanisation also reduces human well-being through loss of education and recreation opportunities afforded by green spaces (Goodness and Anderson, 2014).
The Edith Stephens Nature Reserve, a small (0.39 km2) protected area in the City of Cape Town, illustrates the important role of green space in an urban environment (Figure 14.A). Despite the small size and isolated nature of the reserve, the site provides several benefits to the surrounding community. Originally established to protect a wetland that holds the threatened Cape quillwort (Isoetes capensis, EN), the reserve offers nature-watching opportunities to the public via a bird hide overlooking the wetland, and a quiet neutral space for socialising that is not linked to gang or political territories. The reserve also has an environmental education centre that regularly offers children’s workshops and holiday programs, as well as a teacher-training program. Even so, the site faces considerable urban pressures, including heavy foot traffic across the reserve, nutrient run-off from the adjacent agricultural land and urban area, and informal activities, such as brick cleaning on the property’s boundary.
The reserve was included in the Cape Flats Nature project initiated in 2002, which sought to link urban nature conservation with social justice in the city’s most economically marginal areas (Katzschner, 2013). The idea was to establish conservation practices that integrate ecological sustainability with community empowerment and social well-being. Even though the Cape Flats Nature project ended in 2010, the principles of the project live on through ongoing efforts by the reserve’s small staff and local volunteers recruited from adjacent communities and government public works programs. The reserve has developed good relations with its neighbours, and the staff claim that they enjoy lower incidences of crime than many other protected areas in the city. The Edith Stephens Nature Reserve exemplifies the kind of hybrid conservation practice that is required in an urban African setting where social and biodiversity requirements need to be balanced, especially in small areas with significant development pressures.
The impact of agriculture
Habitat loss from agricultural expansion is arguably the biggest current challenge to biodiversity conservation in Africa (Balmford et al., 2012; Laurance et al. 2013; Maxwell et al., 2016). At the root of this problem is the need to supply food and other resources to a growing human population. Exacerbating the situation, much of Sub-Saharan Africa’s arable land has already been degraded to such a degree that it cannot sustain viable food production anymore (Drechsel et al., 2001). Most of these losses are not due to natural factors, but to poor land management practices, such as overgrazing, continual ploughing of fields, and heavy use of fertilisers. These practices release nitrogen, carbon, and oxygen into the atmosphere and compromise the soil’s ability to hold water, leading to erosion, soil salinisation, desertification, and even climate change (Vågen et al., 2005). This not only lead to collapsing ecosystem services, but also increased competition for space as even more land must be converted for agriculture, and to accommodate people and their activities. Such land conflicts are only going to become worse with climate change (Zabel et al., 2014).
In light of the seemingly irreconcilable conflict between agriculture and conservation, some conservation biologists have suggested that the only way in which we can secure a future for biodiversity is through a land sparing approach, in which agricultural investments are focussed on intensifying practices on land already dedicated to farming and no more. One of the main drawbacks of such a high-yield approach is that the impact of intensive agricultural practices degrades natural ecosystems even far from the immediate area, for example through nutrient and pesticide pollution (Section 7.1). For that reason, others support a land sharing approach that promotes biodiversity-friendly agricultural practices, even if that means agricultural lands continue to expand. One of the main drawbacks of this land-sharing approach is that it still alters ecosystem composition, which would threaten species that need large territories and habitats, in addition to leading to more human-wildlife conflict (Section 14.4).
While the land-sparing versus land-sharing frameworks make for good intellectual debate, the reality is that both are undesirable scenarios when carried out as opposite extremes. There is no denying that agriculture is important—it’s the primary source of livelihood for millions of Africans, and critical for food security. But because there is a finite amount of land available for food production, we have no choice but to develop methods that will allow greater yields on existing agricultural lands without depleting the soil or damaging more ecosystems. In other words, we need to adopt a hybrid approach where some lands are dedicated to large protected areas where human activities are restricted, some lands are dedicated to wildlife-friendly agro-ecosystems, and some lands are used for intensive food production (Fischer et al., 2014; Law et al., 2017). Much of Africa has employed a similar structure historically and even today, where large areas that are relatively untouched or regenerating are interspersed with low-intensity traditional agricultural systems that continue to support a range of native species (Şekercioğlu, 2012). Traditional farming systems offer many strategies showing how natural ecosystem services can be used to improve yields, including the use of biocontrol and crop diversification to keep pests and diseases at bay, and planting of nitrogen-fixing legumes to improve soil fertility. This is in stark contrast to intensive modern agricultural practices dedicated to single crop specialisation; these impoverished ecosystems cannot maintain themselves but rather rely on continuous use of fertilisers and pesticides to remain productive.
Biodiversity-friendly farming practices can also produce economic benefits; and so, many government programs are now promoting sustainable agricultural intensification.
With the increased realisation that biodiversity-friendly farming practices can also produce economic benefits, many government programs have begun to promote and subsidise the adoption of sustainable agricultural intensification (http://www.fao.org/ag/ca/AfricaTrainingManual.html; see also Pretty et al., 2011; Garnett et al., 2013). Also known as conservation agriculture, this farming approach blends traditional agricultural practices with improved and locally adapted crops, as well as integrated crop and pest management strategies (Figure 14.3) to boost yields on existing farmland while creating cost and labour savings; it may even reduce the amount of land under cultivation (Stevenson et al., 2013). Some of these strategies include minimal tilling, crop rotation, intercropping, and terracing (to prevent agricultural runoff which, in turn, prevents erosion). Soil nutrient levels are enhanced through fertiliser microdosing, and by planting legumes, encouraging decomposition by termites, and using crop residues as mulch before composting it directly into the soil. Crop yields are further improved by maintaining windbreaks such as riparian buffer zones, which have the added benefit of enhancing the diversity of seed dispersers, biocontrol agents, and pollinators. In one study, fertiliser microdosing increased sorghum and millet yields and, thus, incomes for 25,000 smallholder farmers in Mali, Burkina Faso, and Niger by 50–130% (Tabo et al., 2011). Another region-wide study found that conservation agricultural techniques could increase average crop yields by nearly 400 kg per hectare (Corbeels et al., 2014). It is important to remember that conservation agriculture is a departure from both traditional and intensive monocrop farming techniques. Adequately training farmers in best practices and new techniques is therefore crucial to program success (Gatare et al., 2013).
To take advantage of the multiple benefits to be gained from integrated, biodiversity-friendly farming techniques, entire industries have started adopting such practices. Among the most prominent are cacao and coffee, where many growers now produce their crops under native shade trees (Box 14.3). While Africa has had a long history of shade-grown cacao and coffee production, recent decades have seen many farmers transitioning towards intensive farming in full sun, which allows for easier mechanisation. But crops grown in full sun are generally also of lower quality and more susceptible to pest outbreaks (Kellerman et al., 2008; Bisseleua et al., 2009; Tscharntke et al., 2011). In contrast, shade-grown cacao and coffee benefits both the farmers and biodiversity: a study from Ethiopia found that shade coffee farms had over double the number of bird species in comparison with nearby forest sites (Buechley et al., 2015). Shade farming could also be a strategy for farmers trying to cope with increasing temperatures due to climate change (Blaser et al., 2018). Considering the large global markets for cacao and coffee, reverting to traditional growing methods here would have a large positive impact on the environment simply due to the economy of scale.
Box 14.3 Preserving Biodiversity Through Shaded Agroforestry
Hervé D. Bisseleua
World Agroforestry Centre (ICRAF),
Nairobi, Kenya.
[email protected]
Chocolate is one of the most universal treats in the world, but could your sweet tooth be increasing biodiversity loss? The chocolate tree (Theobroma cacao) is traditionally grown in areas with dense and diverse canopies of shade trees, home to an abundant variety of plants and animals (Figure 14.B). The chocolate industry is strongly dependent on small-scale agriculture, but also highly vulnerable to pest and disease outbreaks, and climate change. These production challenges combined with increasing global demand for chocolate has increased economic and social pressures to achieve higher yields within a shorter timeframe. Higher yields could be achieved through reduced shade tree management and increased use of chemical pesticides and fertiliser. But these techniques lead to deforestation, biodiversity loss, and loss of ecosystem functioning. Higher yield techniques in the short term are also not sustainable over the long term: work in Cameroon and elsewhere showed that the promotion of high-yielding hybrid cacao varieties under direct full sun have contributed to more frequent outbreaks of pests and diseases (Kellerman et al., 2008; Bisseleua et al., 2009; Tscharntke et al., 2011).
To achieve more sustainable cacao production, and to benefit from ecosystem services, such as enhanced biological control of pests and diseases, and increased soil fertility, West Africa’s cacao farmers are now gradually returning to agroforestry practices that embrace increased shade tree diversity. Farmers adopting these techniques are already reaping benefits. For instance, in Ghana and Cameroon, cacao yields from shaded cacao agroforestry systems are 12–23% higher compared to full sun systems (Bisseleua et al., 2009; Asare and Raebild, 2016). In eastern Côte d’Ivoire, the use of leguminous trees as shade in rehabilitated cacao plantations is also reported to increase the survival rate and yield of cacao trees (Smith Dumont et al., 2014). Cacao grown in shade may produce for 60–100 years, whereas production may only last for 20 years without shade (Obiri et al., 2007). In addition to environmental services, diversified shade trees may also provide additional income opportunities, such as timber and firewood production, medicine, local spices, and fruit, from native shade trees such as the njangsang tree (Ricinodendron heudelotii) and bush mango (Irvingia gabonensis) (Smith Dumont et al., 2014). Importantly, in all these multi-strata systems, a higher density and diversity of shade trees means higher densities and diversity of pollinators and biological pest control agents such as ants and social wasps, which in turn increase cacao yields even more (Bisseleua et al., 2017).
In conclusion, better land management practices, such as allowing a diversity of shade trees to grow among the cacao crop, increases both biodiversity and revenue for farmers. Tropical agroforestry is thus a promising approach to reconcile biodiversity conservation and economic development. Educating farmers on shaded agroforestry systems and creating complimentary economic incentives and policies would help farmers adapt to better management practices faster, ultimately allowing agroforestry systems to contribute more to biodiversity conservation. A guilt-free sweet tooth, indeed!
Sustainable agricultural practices are much needed in the conservation portfolio of Africa, where farmers have been slow to adopt conservation agriculture. Two main challenges are noted. First, spreading ideas and innovations across different agricultural sectors has been challenging, and much work remains to implement them. Second, due to incredible ecosystem diversity across the region, specific practices are not equally suitable everywhere; there is thus need for more research into flexible practices that can be modified to meet local growing conditions (Giller et al., 2009). Strengthening coordination of agricultural research and cooperation at local and regional scales (Gonthier et al., 2014), as well as better land-use allocation (Law et al., 2015) may solve some food production challenges. Adapting biodiversity-friendly certification schemes to consider local dynamics might also encourage more African farmers to adopt biodiversity-friendly techniques (Gove et al. 2008; Buechley et al., 2015). Not only would biodiversity benefit, but these practices would enable farmers across Africa to receive higher prices for their crops and recover a large portion of the land lost to land degradation each year without the need for more land conversion.
The impact of logging, mining, and other extractive industries
Conservationists are also dependent on extractive industries. Rather than criticise, it is more productive to partner with and influence these industries to contribute to conservation efforts.
As with intensive agriculture, high-impact resource extraction industries have not traditionally been compatible with conservation needs. These include mining, oil and gas extraction, dredging, quarrying, and logging, which have often been associated with complete ecosystem destruction. While it is easy to criticise these industries for their impact on nature, it is important for conservationists to remember that we are all dependent on those industries in some way or other, even to perform our conservation activities. Rather than criticise, it is more productive to partner with and influence with industries to contribute to conservation efforts.
There are many examples illustrating how partnerships between conservation biologists and extractive industries can benefit conservation. One of the best examples comes from the timber industry, which has the potential to greatly increase forest conservation opportunities (Clark et al., 2009). Traditionally known for leaving unsightly clear-cuts behind them, research has shown that improved logging techniques facilitate quicker ecosystem recovery after harvesting, which in turn also benefits biodiversity. For example, while the response of wildlife to logging differs depending on the harvesting method and forest type (Ofori-Boateng et al., 2013), primates (Stokes et al., 2010; Morgan et al., 2018), amphibians (Ofori-Boateng et al. 2013), and birds (Şekercioğlu, 2002) can all tolerate responsible-done light-touch logging techniques (but see also Bicknell et al., 2013; Gatti et al., 2015).
Guided by this research, some sectors of the timber industry have been keen to adopt more sustainable logging techniques (Figure 14.4) that focus on reducing damage to the soil, stream banks, and remaining trees, while removing just the largest trees. This approach reduces soil disturbance, erosion, waste, and carbon emissions. The Forest Stewardship Council (FSC) and other similar organizations are setting certification standards for sustainable logging, which enable certified logging operations to sell their products at higher prices on world markets. In addition to impact reduction within logged areas, certification schemes typically also require timber companies to avoid logging high conservation value forests, which is a good strategy for protecting ecosystem services and biodiversity in general. Some agroforestry companies also allow local people to cultivate rare medicinal and aromatic plants in the shaded areas on their concessions which reduces harvesting pressure on wild populations (Rao et al., 2004). Lastly, a key element for wildlife management in logged forests is to stop hunters, fishers, trappers, and plant collectors from entering the impacted area after timber harvest by closing unused logging roads. (For more discussion on the impact of the logging road on biodiversity, see Laurance et al., 2014 and Benítez-López et al., 2017.) (For an example in fisheries, see Box 7.2)
Many African mining companies have also become active partners in conservation. These partnerships include contributions like providing funding for conservation activities, participating in biodiversity offset programs (Section 10.3.3), and subsidising conservation agriculture efforts. South Africa offers several examples illustrating how extractive industries can develop productive conservation partnerships. Two South African diamond trading companies, De Beers Group and E. Oppenheimer and Son, converted 2,500 km2 of their properties earmarked for diamond mining and exploration into protected areas used for ecotourism and environmental research. Also, in South Africa, the petrochemical company SASOL supports a wide range of environmental programs, including sponsoring natural history field guides, anti-poaching programs, and threatened species recovery projects.
Despite these and other examples of progress, many challenges associated with extractive industries remain unsolved. For example, pollution from these industries continues to threaten Africa’s environment, and many extractive companies remain unfriendly towards conservation activities. Many of these challenges stem from efforts to cut costs by ignoring legal and regulatory requirements, particularly when it involves stakeholders from outside Africa who prioritise short-term profits over long-term sustainability. Tardy monitoring and enforcement of regulatory requirements (sometimes driven by corruption) also remain as obstacles (Linder and Palkovitz, 2016). The best solutions for overcoming these challenges involve continued diplomacy to establish new (and strengthen existing) partnerships with such industries. Conservation biologists also need to educate industry workers and the general public to be on the lookout for environmental violations and inadequate law enforcement, which needs to be reported and addressed before more damage is done. Above all, it is important to remind legislators and other members of society that environmental damage can harm, sometimes irreversibly, our own ability to have fulfilling lives. | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/14%3A_Conservation_on_Unprotected_Lands/14.00%3A_Prelude_to_Conservation_on_Unprotected_Lands.txt |
Infrastructure development poses a significant and escalating challenge to biodiversity conservation efforts. Dams and fences impede wildlife dispersal and migrations (Section 5.1.1), power distribution lines and high-rise buildings pose a collision hazard to birds and bats (Rushworth et al., 2014; Frick et al., 2017), and city expansions compete with biodiversity for space. Expanding road networks are particularly harmful because roads open new areas for deforestation, urban sprawl, agricultural expansion, and unsustainable hunting (Laurence et al., 2006; Benítez-López et al., 2017). A recent review found that 75% of Sub-Saharan Africa’s development corridors—large-scale infrastructure developments meant to stimulate economic growth—would cut through sparsely-populated and low-quality agricultural, range, and forest lands (Laurance et al., 2015; Sloan et al., 2016). These developments not only threaten the wildlife living within those areas (Benítez-López et al., 2017), but also the carbon-sequestering potential of large swathes of tropical forests (Laurance et al., 2015). Most alarming, the corridors’ planners have shown very little regard for existing biodiversity conservation efforts, given that the proposed transportation network cuts directly through 408 existing protected areas, which includes 69 national parks, biosphere reserves, World Heritage Sites, and Ramsar wetlands (Sloan et al., 2016). In contrast, only five of the 33 planned and active development corridors would cross areas of low conservation priority and with promising agricultural potential (Laurance et al., 2015).
Improving existing infrastructure in disturbed and populated areas makes more economic sense than creating new developments in wilderness areas.
Conservation versus development is not a zero-sum game. Rather, biodiversity conservation improves our own well-being by enabling us to obtain the necessary resources to support our livelihoods and our industries’ profit margins (Chapter 4). One way to maintain these benefits while also promoting economic development (Section 15.1) is to focus on improving existing infrastructure in disturbed and populated areas, rather than creating new developments that bisect marginal lands, protected areas, and wilderness areas. Developing marginal lands and wildernesses seldom makes sense, not only because these areas are sparsely populated, but also because many are low-nutrient environments that would never support sustainable agriculture (Balmford et al., 2001; Laurance et al., 2015).
When new developments are necessary, there are usually opportunities to balance diverging interests. For example, to offset the large land footprint of renewable energy, new wind farms and transmission cables could be directed to already degraded land. In many cases, the compromise might even contribute more to socio-economic developmental goals than the original plans. This was well illustrated in Tanzania, where a proposed road development would have disrupted the famous Mara-Serengeti migration route for large mammals, with potentially dire consequences to the area’s ecotourism industry (Dobson et al., 2010; Holdo et al., 2011). To avoid such an impact, scientists used computer models to identify an alternative route that would not only minimise disturbance, but also achieve greater socio-economic development (Hopcraft et al., 2015). Studies, such as these, have provided important foundations for similar work to mitigate the impact of fences on wildlife (Durant et al., 2015), and by making minor adjustments to shipping lanes to reduce collisions between whales and ocean-faring vessels (Silber et al., 2012).
There are also opportunities to make existing infrastructure more wildlife friendly. Of interest is the maintenance of connectivity despite the presences of potential barriers such as fences and roads. For example, strategically placed fence-gaps and exclusionary fences, as well as tunnels placed under fences can be used to facilitate continued dispersal of selected species in fenced areas (Dupuis-Desormeaux et al., 2018). Similarly, warning signs (Figure 14.5), overpasses (e.g. Ford et al., 2009) and underpasses (e.g. Dell’Amore 2012) along paved roads can keep motorists safe from collisions with large animals. One study from Canada found that strategically placed wildlife crossings could reduce vehicle collisions involving large mammals by 96% (Ford et al., 2009), also reducing the chance of human injuries and damage to vehicles (Huijser et al., 2009). While this field of research is still relatively new, much headway has been made in making wildlife crossings cost-effective (https://arc-solutions.org) and determining their optimal placement (Bastille-Rousseau et al. 2018). | textbooks/bio/Ecology/Conservation_Biology_in_Sub-Saharan_Africa_(Wilson_and_Primack)/14%3A_Conservation_on_Unprotected_Lands/14.02%3A_Smart_Development_Outside_Conservation_Areas.txt |
Subsets and Splits