Pimm (1988) at the same time identified some flaws when relating the island model theory to certain fragments, since some species were shown to adapt differently and some were even shown to flourish in disturbed habitats. Leaf-cutter ants for example increased many times over and also a number of frog species once thought to be dependent on the forest were shown to be thriving in nearby farm ponds. Interactions with the ‘outside world’ such as these are not catered for by the island theory and show how some species can adapt to forest fragmentation.
Janzen (1983) had earlier identified the inward proliferation of some species in disturbed habitats noting …“as areas of conserved pristine forest are reduced in size they are increasingly susceptible to significant immigration of plants and animals”. This denotes that in some circumstances since natural species are losing the ability to regenerate, fragments are more vulnerable to invasion of plants and animals from an anthropogenic background.
Janzen (1983) provided evidence for this when looking at succession in the place of a fallen tree in a pristine fragment of Amazonia. On returning after 3.5 growing seasons the number of plants growing where the tree had fallen had increased considerably despite shade cast by the forest canopy, suggesting there had been a significant invasion of plants from an anthropogenic background. He suggested that to look at fragments from a conservation point of view it would be better to surround them with species-poor vegetation of non invasive species which have a low food value, for example cotton fields, so as to reduce the chances of species invasion.
Wind has been shown to infiltrate up to 800 metres from the forest edge and studies have shown that wind exposure at forest margins can result in a sudden increase in tree damage and mortality (see figure 1) leading to changes in structure of the vegetation. The ‘BDFFP’ suggested the impacts of these edge effects are far more significant than any other effects on fragmented patches; these include high levels of deterioration of the forest structure. Laurance et. al. (1997) investigating forest patches found that only a decade after fragmentation, considerable edge effects were taking their toll on the new forest boundary areas and that approximately 10% of plant biomass disappeared after two years of isolation (see figure 2). He also found that these tree losses were influencing the reduction of biomass in small fragments.
Figure 1: Amazonian rainforest fragments exhibit sharply elevated rates of tree
mortality and damage near fragment margins.
(From Laurance [on-line] 2001)
Figure 2: Collapse of biomass in three small (ca. 1.4 hectare) rainforest
fragments in the central Amazon.
(From Laurance [on-line] 2001)
Terborgh (1992) examined the biological mechanisms of species loss in habitat fragments and recommended the processes could be put into two groups, direct and indirect effects of fragmentation. A reduction in population sizes of many species as a result of a reduction in habitat area was seen as a direct impact since many species requiring large home ranges have found themselves without adequate living space. For example, a single jaguar requires 30 to 50 km2 to live contentedly. Thus an area the size of Barro Colorado Island, Panama (16 km2) in which jaguars once roamed freely, can no longer sustain a population of jaguars. Table 1 demonstrates such spatial scales applicable to a number of Neptropical organisms required for the completion of their life cycles. Intrusion of edge effects (discussed earlier) was also seen as an inevitable direct effect of fragmentation, where interior forest is suddenly exposed to dry winds, light and possibly fire, changing species composition and structure. The intrusion may well affect both plant and animal communities.
Table 1: Spatial scales required for Neotropical organisms
(From Terborgh, 1992)
The main indirect effect of fragmentation was seen as the population increase of certain species at the expenditure of their predators. For example, the loss of large predators may have a destabilising effect on populations of seed predators which in turn may affect the composition of tree species in the forest. For example, with the decline in population of the jaguar on Barro Colorado Island large rodents have flourished. Proliferation of the large rodents which favour eating large seeds has in turn led to a reduction in large trees and shrubs bearing these seeds. There has since been an increase in smaller flora, no longer limited by their height or shade. This beggars the question, if the predator was successfully reintroduced would the situation be reverted? Terborgh (1992) argues where top predators have been removed it may be unfeasible to reintroduce them since …“once they have gone it may be impossible to reverse the chain reaction leading to loss of diversity”.
Several important questions have therefore been raised concerning the impacts of fragmentation on the plants and animals within them. Whilst studies such as those conducted by the BDFFP have improved our understanding of what happens to fragments it is still a relatively new problem and therefore difficult to predict future outcomes or produce solutions. The general consensus appears to be, the smaller the fragment the less natural it is likely to become. Ideally it should equal or exceed the size required by the largest predator.
“The second most recent common cause of recent extinctions, after habitat destruction, is overhunting.” (Reid, 1992)
The impact of hunting within the humid tropics has been mostly overlooked, despite influencing the tropical forests for much longer than anthropogenic fragmentation, perhaps since wildlife has always served as a major source of food for many indigenous populations and therefore taken for granted.
Hunting can be put into two broad categories, that of subsistence and commercial, the latter involving trade in animals and their products. Subsistence hunting, most prolific within tropical forests, usually involves hunting large numbers of animals and will often serve as a major food source for native populations.
Because of hunting many animal and plant species have now been reduced to the extent that they no longer perform their ‘ecological function’ and in some instances animals have been hunted to such a low number, that even though they still exist within the community they no longer interact notably with other species. In the long term unsustainable hunting unchecked will cause a decrease in density, species richness and biomass.
Fitzgibbon et. al. (1995) looked at the impacts of subsistence hunting in the Arabuko-Sokoke Forest in Kenya. Mammal populations in this region provide an important source of protein and income for local people. In 1991 their harvested biomass exceeded 350 kg/km2, consisting mainly of bushpigs, aardvarks and primates. This amount led Fitzgibbon et. al. (1995) to question the sustainability of the practices involved.
Figure 3: Four-toed elephant shrew
Trapping occurred for the most part toward the edge of the forest, within 1-2 km, since it was rarely worth setting traps further away. Whilst trapping was seen to substantially reduce density of prey on the edges, a relatively small percentage of the population was affected since many could find refuge within the forest. For example trapping of 4-toed elephant shrew (see figure 3) was shown to reduce populations by 40%. This might seem substantial; however the edges contain just 4% of the shrew population and therefore the species is plentiful enough to restore its population. Consequently this type of hunting was seen to be sustainable.
(From Jonathon kingdom-Animal Art
[On-line] 2001)
In contrast, Yellow baboons and Syke’s monkeys (larger primates) hunted throughout the forest were shown to be over-harvested, since they have no refuge. The process was therefore seen to be unsustainable. Fitzgibbon et. al (1995) suggested …“one way to reduce excessive harvesting of primates would be to stop people from hunting primates while allowing trapping to continue around the forest edge”. However, an increasing human population, coupled with reported differences in animal abundance and population dispersion, led Fitzgibbon et. al (1995) to conclude uncontrolled harvesting is unlikely to be sustainable in the future.
The situation within the Arabuko-Sokoke Forest highlights the direct impacts of hunting, however one must as with fragmentation account for indirect impacts too. This is where the hunting and fragmentation issue begin to interconnect since fragmentation may in certain circumstances occur as an indirect result of excessive hunting. The implications of fragmentation and hunting are further complicated when addressing the number of plant and animal populations expected within tropical forests. Howe (1984) suggested …. “animal-mediated dispersal is certain to be critical for the demographic recruitment of many or most tropical species”. Difficulties with this theory are demonstrated when excessive hunting of top predators boost growth of fruit bearing plants (seed dispersers) and so an increase in fruit bearing trees and bushes. On the other hand if seed dispersal is stopped, top predators will lose an important food source and the fruit trees and bushes lose their distribution ability.
Myers (1996) reflecting upon long term repercussions of both fragmentation and hunting suggested a…“loss of species may turn out to be less significant in the long term than the reduction of evolutions capacity to generate new species”. Bearing in mind fragmentation of tropical forests is continuing, perhaps even accelerating and indigenous populations are continuing in some instances to hunt unsustainably the future appears bleak. Education of indigenous populations and of deforesters appears a necessity if forest fragments are allowed to be restored to their natural state or if hunting is to remain sustainable in the long-term.
The complexity of the tropical forests makes it impossible to anticipate when and what species will disappear and which factors (fragmentation/hunting) will contribute the most. Due to the nature of the problem and the difficulty in conducting effective studies it is therefore difficult to make any concise predictions. It is however clear the consequences of forest fragmentation and hunting on wildlife are a direct result of the mutual dependencies of humans on the natural world, and that their impacts are not that different in some respects. Whilst direct impacts individually target either plants or animals, there is potential for indirect impacts to overlap and provoke the same outcomes. Difficulties faced by plants and animals are potentially short term (in evolutionary terms) if man chooses to change his ways, however if man continues the evolutionary consequences within the tropics may plague all concerned for many generations to come.
References
Bierregaard, R.O., Lovejoy, T.E., Kapos, V., Santos, A., and Hutchings, R. 1992. The biological dynamics of tropical rainforest fragments. Bioscience
Fitzgibbon, C.D. et. al. 1995. Subsistence hunting in Arabuko-Sokoke forest and its effects on mammal populations. Conservation Biology 9: 859-866
Howe, H.F. 1984. Implications of seed dispersal by animals for tropical reserve management. Bioogical. Conservation. 30: 261-281.
Janzen, D.H. 1983. No park is an island: increase in interference from outside as park size decreases. Oikos 41: 402-410.
Jonathon kingdom. Jonathon kingdom-Animal Art.
[On-line] Accessed 5 December 2001
Laurance, W.F. et. al. 1997. Biomass collapse in Amazonian forest fragments. Science 278: 1117-1118.
Laurance, W.F. 2000. Dynamics and biomass of Amazonian forest fragments. [On-line] Accessed 5 December 2001.
Myers, N. 1996. The biodiversity crisis and the future of evolution. The Environmentalist 16: 37-47.
Pimm, S.L. 1998. The forest fragment classic. Nature 393: 23-24
Reid, W.V. 1992. Conserving life's diversity: can the extinction crisis be stopped? Environmental Science and Technology 26(6): 1090-1095.
Terborgh, J. 1992. Maintenance of diversity in tropical forests. Biotropica 24(2b): 283-292
