ONE OF THE BITS of conventional wisdom about islands most of us accept implicitly is that island vegetation is relatively defenseless against introduced herbivores (Carlquist 1974, Bowen and van Vuren 1997). Scores of anecdotal accounts of denudation of islands by goats, rabbits, pigs, and other introduced herbivores lie behind this conventional wisdom. The reports are so numerous and consistent that one cannot doubt their collective veracity (Coblentz 1978, Courchamp et al. 1999). But the simplistic conclusion to be drawn from these anecdotes—that island floras typically evolve reduced defenses against herbivores—may be understating a more complex and interesting reality.
A less often remarked upon generality is that essentially all islands support herbivores, be they insects, crustaceans, lizards, tortoises, birds, or even mammals. We are thus presented with a paradox: if most islands support native herbivores, then why are island floras so vulnerable to introduced herbivores, especially mammals?
At least two reasons come to mind. There very well may be more. The principal herbivores of remote islands are arthropods, but arthropod herbivores may be mismatched with respect to food plants since plants and arthropods are likely to colonize independently (Janzen 1973a, 1975). Plants generally arrive as seeds transported via wind or in the guts of birds or bats, whereas arthropods can be carried on the wind or in the plumage of birds, or rafted in driftwood. Thus colonizing arthropod herbivores will rarely find their preferred host plants on a given island and will consequently either fail to survive or be obliged to subsist on less preferred plant species on which larvae will develop slowly and in reduced numbers. Mismatching of plants and herbivores could result in reduced herbivore pressure and evolved relaxation of defenses.
There is some support for this idea. Back in the 1970s, two investigations independently reported that sweep net samples of arthropods from Caribbean islands contained conspicuously fewer species and individuals than samples from equivalent sites on the Netoropical mainland (Allen et al. 1973, Janzen 1973b). In keeping with this observation, it was noted shortly afterward that the bird communities of several Antillean islands are consistently deficient in the specialized insectivores that dominate the avifaunas of the mainland (Terborgh and Faaborg 1980). More than 85% of the individual birds captured in standard mist-netted samples at low-elevation sites on either the South or North American mainland were strict insectivores, whereas fewer than 20% of those captured in the Antilles were. The remaining 80% of the Antillean birds were omnivores, nectarivores, frugivores, and granivores, species living at lower trophic levels whose livelihoods were derived in part or in full from plants. This result pointed to something distinctive and fundamental about the organization of island avifaunas, but to my knowledge, no one has pursued it further.
A second reason island floras may be relatively lacking in antiherbivore defenses is that many of the nonarthropod herbivores of islands are terrestrial and therefore unable to access arboreal foliage (Carlquist 1965). One can point to the land iguanas and tortoises of the Galapagos, the flightless geese of Hawaii and other Pacific islands (James and Burney 1997, Steadman 2006), the land crabs of many midoceanic islands, and the out-sized chuckwallas of the Sea of Cortez. In such a setting, a plant has only to grow to a meter or so to escape all but arthropod herbivores. The latter are likely to be controlled by predators—birds, lizards, spiders, and the like (Spiller and Schoener 1990). Reduced herbivory should translate rapidly into reduced investment in antiherbivore defenses, given that tannins and other antiherbivore compounds can constitute up to 35% of the dry weight of foliage (Coley et al. 1985). Thus, before we are tempted to draw broad generalizations about reduced antiherbivore defenses in island vegetation, it would be wise to investigate the specific context of the island(s) in question.
In pursuing this further, it would be helpful to refer to a theoretical framework. There is, in fact, a theory that can allow us to make predictions about levels of herbivory on islands, although the theory was not constructed with islands in mind. Proposed in 1981 by Oksanen, Fretwell, and others, it was termed “the exploitation ecosystems hypothesis” (EEH). A refined statement of it appears in Oksanen and Oksanen (2000). The theory, like most useful theories in ecology, is quite simple in outline. In essence, it follows Hairston, Smith, and Slobodkin (1960) in assuming three trophic regimes in terrestrial ecosystems (figure 5.1). The key variable is productivity. At the lowest productivity levels, barely above zero, there are only producers. Such type-I ecosystems are found only in the most extreme deserts and in the high Arctic or Antarctic (examples in Oksanen et al. 1981, Oksanen and Oksanen 2000). We would expect plants living under such circumstances to allocate relatively little of their meager resources to herbivore defenses (Blossey and Nötzold 1995).
Figure 5.1. Exploitation Ecosystem Hypothesis. Trophic levels are added in stepwise fashion as ecosystem productivity increases (from Oksanen and Oksanen 2000).
At slightly higher productivity levels, the amount of energy transformed by the ecosystem becomes sufficient to support a consumer trophic level. We shall call these type-II ecosystems. Since some arthropods can subsist on quantities of resources that are almost invisible to humans, we would expect arthropods to enter at lower productivity levels than vertebrates. The EEH presumes that, as productivity increases beyond the herbivory threshold, herbivory increases apace, maintaining the plant biomass at a roughly constant level.
At some point (again, probably sooner for arthropods than for vertebrates) productivity crosses a second threshold, and a third trophic level—predators—enters the picture in type-III ecosystems. With still further increases in productivity, predators are presumed to maintain consumers at more or less constant levels, just as the consumers maintained the plant biomass at nearly constant levels in type-II ecosystems. This being so, edible (nonwoody) plant biomass increases with further gains in productivity up to a maximum determined by the physical environment. The EEH thus incorporates both bottom-up and top-down forcing.
Now, what has this to do with islands? It has a lot to do with islands if we make a simple substitution of parameters. The most informative variable of island biogeography—island size—is an excellent surrogate for productivity (other factors, climate, soils, etc., being equal). The substitution of area for productivity was pioneered by Schoener (1989) and is known as the productivity-space hypothesis. Biogeographical arguments can also link island area to the length of food chains (Holt 1996). Applying this logic, the smallest islands should support only producers, somewhat larger islands should support producers and consumers, and so forth.
Our focus for the remainder of this inquiry will be type-II islands, those supporting producers and consumers, but not predators of a dominant herbivore. I shall consider type-II islands originating in two distinct ways, via contraction and via colonization, and show that their herbivore communities display some convergent properties independent of the taxa involved. We shall also see that type-II ecosystems are unlike any we ever encounter in our normal travels. Natural type-II ecosystems have become extremely rare and one has to go, quite literally, to the ends of the earth to find them, at least in the tropics.
The first case I shall present involves a type-II ecosystem created by the contraction of a type-III ecosystem to an area (i.e., productivity level) too small to support predators of vertebrates and some invertebrates. In the case in question, the area contraction took place when the Caroní Valley in Venezuela was flooded in 1986 by the huge (4,300 km2) Guri hydroelectric impoundment (Morales and Gorzula 1986). Flooding fragmented the formerly continuous dry forest of the mainland, creating hundreds of islands ranging from tiny specks of <<1 ha to > 760 ha.
Our first surveys of some of these islands in 1990 indicated that three-quarters or more of all vertebrates present on the nearby mainland had already disappeared from islands of < 12 ha, leaving strongly imbalanced animal communities. Some functional groups were underrepresented (e.g., pollinators, seed dispersers) whereas others were entirely absent (predators of vertebrates). Nearly all persistent species exhibited hyper-abundance, that is, their local population densities on islands were elevated far above their densities on the mainland (Terborgh et al. 1997a,b). Persistent hyperabundant groups included birds, some lizards and amphibians, spiders, small rodents, and several generalist herbivores: red-footed tortoise (Geochelone carbonaria), common iguana (Iguana iguana), red howler monkey (Alouatta seniculus), and leaf-cutter ants (Atta spp., Acromyrmex spp.) (Terborgh et al. 1997b, Lambert et al. 2003, Rao et al. 2001, Aponte et al. 2003, Orihuela et al. 2005).
Since many of our results from the Lago Guri island system have been published elsewhere, I shall provide only a brief summary here, focusing particularly on herbivory. We studied herbivory indirectly via assessments of plant demography at sites supporting high, medium, and low densities of generalist herbivores. Herbivore abundance varied inversely with island size so that “small” islands (below 1.5 ha) supported the highest herbivore densities, “medium” islands (between 3 and 12 ha) supported intermediate densities, and “large” landmasses (88 and 190 ha, mainland) supported low densities. To assess the effects of herbivore density on plant demography, we followed the fates of 3030 small saplings (≥ 1 m tall and < 1 cm diameter at breast height [dbh]), 3997 large saplings (≥ 1 cm, < 10 cm dbh), and 4771 adult trees (≥ 10 cm dbh) for 5 years at 12 sites (table 5.1).
The mortality of small and large saplings was elevated on both small and medium islands, but the differences were not always statistically significant. Far more pronounced were the decreases of recruitment into both stem size classes. Recruitment into the adult tree class (≥ 10 cm dbh) did not differ in relation to landmass size. In sum, demographic effects associated with hyperbundant herbivores were greater for recruitment than mortality and restricted to small stem size classes.
Given that common iguanas and red howler monkeys confine most or all of their feeding activities to the canopy, and that tortoises were not found on small islands, leaf-cutter ants emerged as the herbivore most likely responsible for the low recruitment rate of saplings (Lopez and Terborgh 2007). We obtained further evidence implicating leaf-cutter ants and perhaps other arthropods by setting out tree seedlings under fine wire mesh cages. Seedling survival was high under cages, even at sites supporting Atta densities 100 times greater than observed on the mainland (Lopez and Terborgh 2007). In some cases, uncaged seedlings were defoliated during first night of exposure, whereas seedlings survived up to 3 years under cages (figure 5.2).
TABLE 5.1
Demography of Small and Large Saplings on Small, Medium, and Large Landmasses at Lago Guri, Venezuela, 1997–2002
We found that hyperabundant leaf-cutter ants were relatively unselective in their choice of foliage compared to ants living in widely separated colonies on large landmasses (Rao et al. 2001). Similar observations were made on red howler monkeys (Orihuela et al. 2005). The observation of decreased selectivity under hyperabundance carries important implications.
First, it shows that plant defenses conferring low preference status under “normal” circumstances act in a conditional fashion, being effective only at low herbivore densities. We found that most plant species become vulnerable at high herbivore densities, as indicated by the fact that mortality of saplings exceeded recruitment in nearly every species present on small and medium islands. Relaxed defenses in response to insularity was not a factor in this situation because all plants stranded on Guri islands carried genotypes evolved under mainland conditions. “Edge effects” and exposure to prevailing winds had no discernible effect on the mortality or recruitment of any size class of stems (Terborgh et al. 2006).
Second, the facultative ability of leaf-cutter ants, howler monkeys, and presumably other generalist herbivores to subsist on species of foliage that are ordinarily rejected allows their numbers to increase as much as an order of magnitude above those considered “normal.” Thus the “carrying capacity” for generalist herbivores released from top-down control is many times greater than normal density, at least as a transient condition (Beschta and Ripple 2008).
Figure 5.2. Top: Dry forest understory of a large landmass control site at Lago Guri, Venezuela. Bottom: Understory of a small island supporting a hyperdense population of leaf-cutter ants.
Figure 5.3. Two dimensional NMDS ordination of stems ≥1, <10 cm dbh found in 225 m2 sampling plots located on Atta colonies (squares) and away from Atta colonies (diamonds) on medium islands in Lago Guri, Venezuela. The two sets of points are distinct by multiresponse permutation, p=0.001.
Third, community-wide suppression of plant recruitment by hyper-abundant herbivores leads to collapse of the characteristic dry forest vegetation of the Caroní Valley and its replacement by an entirely novel plant community never before documented.
We were not able to quantify the plant species composition of the vegetation that would emerge under steady-state type-II conditions because transformation of the vegetation of the islands we studied was still in mid-course when the project ended in 2003. We did, however, obtain some hints of what might be in store by inventorying saplings growing on top of five Atta colonies on four medium islands (figure 5.3). The figure shows a nonmetric multidimensional scaling ordination of stems ≥ 1 cm and < 10 cm dbh growing in 225 m2 plots centered on Atta colonies and at sites beyond the foraging radius of existing Atta colonies on the same islands. In each case, points representing Atta colony samples fall near the periphery of the ordination space and far from the corresponding off-Atta-colony samples, indicating marked compositional divergence. Just how marked the divergence was can be judged by a pair of examples. The 3 most abundant species growing on Atta colonies on the island of Ambar, representing 258 out of 419 stems (62%), were not represented in 302 stems from 2 off-colony sites on the same island. Conversely, none of the 3 most abundant species in off-colony samples was contained in the 90 stems growing on an Atta colony on the island of Panorama. Interestingly, there was no consistent direction of divergence of the various Atta colony samples in ordination space, in keeping with the fact that different plants tended to dominate at different sites.
Plants able to survive and even increase at Atta colony sites included both common and rare elements of the local dry forest vegetation. The five colony sites supported from 90 to 275 saplings of 14 to 38 species, a majority of which can be presumed to be survivors from precolony times rather than newly established individuals (table 5.2). Each site was dominated by a small number of species, from 1 to 5, that made up 50% or more of the stems. The great majority of species were represented by only 1, 2, or 3 stems at each site. The collection of dominant species is taxonomically diverse, yet most of them were exceptional in possessing coriaceous evergreen leaves, an uncommon feature in the semideciduous dry forest vegetation of the Caroní Valley. Another characteristic that may have deterred Atta herbivory, found in two legumes (Acacia sp., Calliandra laxa), was the possession of compound leaves with finely divided leaflets that were individually much smaller than the usual load carried by Atta workers.
Another noteworthy feature of the results is that the lists of species that dominated on each island show little overlap. Here we appear to have a good example of what Hurtt and Pacala (1995) have termed “winner by default.” Any given island will carry only a sample of the regional floristic diversity and a given site within an island will offer an even more limited diversity. Thus, the “best competitor” in the regional species pool will not always be on hand to “win” in a given situation and other species will succeed instead. In an open competition run over many generations in the presence of hyperabundant herbivores, the winners might be further pared down to an even smaller group of species than we observed on the four islands.
The species listed in table 5.2 appear to be the vanguard of a drastically altered vegetation adapted to a type-II world of hyperabundant herbivores. One can anticipate that most of the less common species still surviving on Atta colonies at the time of our census will eventually die out, leaving only the most resistant species. One can further anticipate that a huge loss in plant diversity will accompany the winnowing process. Speculating even further, one could anticipate that a type-II world at equilibrium would be characterized by a low diversity of highly defended plant species and, accordingly, reduced densities of herbivores.
TABLE 5.2
Numbers of the Five Most Abundant Sapling Species Found in Five 225 m2 Plots Centered on Atta Colonies on Four Medium Islands in Lago Guri, Venezuela
Is this merely wild speculation, or can we find real-world examples of equilibrial type-II ecosystems with which to test the idea? The answer is yes, though well-documented examples are few. Before humans transformed the ecology of the world’s islands, the oceans undoubtedly contained hundreds or perhaps thousands of islands supporting type-II ecosystems. Many islands of the Pacific and the Indonesian archipelago would have qualified, as would many of the Philippines and West Indies. But human conquest of the world’s islands was accompanied by habitat destruction, introductions of domestic and commensal animals, and consequent extinctions that have forever altered the ecology of the vast majority of the world’s islands. Introduced rats, rabbits, cats and other human commensals have fundamentally disrupted the ecology of even remote subantarctic islands like Macquarie, Kerguelen, Crozet and the Tristan da Cunha group (Courchamp et al. 1999). But fortunately, a few extremely isolated islands have survived more or less intact, and it is to these we must go to find the answer to our question.
In pondering this issue, and pursuing it in the literature, I found three cases that are supported by sound natural history data. Two are isolated islands in the Indian Ocean: Christmas Island and the Aldabra Atoll, and the third is East Plana Cay in the Bahamas. Each of these islands supports a generalist herbivore in the absence of predators, and in each case, the herbivore belongs to a different taxonomic class or phylum. On Christmas Island the herbivore is a land crab, Becarcoidea natalis; on Aldabara it is a tortoise, Geochelone gigantea; and on East Plana Cay, it is a mammal, the Bahamian hutia, Geocapromys ingrahami (table 5.3).
In all three cases, the herbivores maintain population densities and biomasses greatly exceeding those of equivalent herbivores in the presence of predators (Coe et al. 1976, Iverson 1982). We shall see that these three cases, disparate as they are in geography and taxonomy, have much in common with each other and with the case of the Lago Guri islands already considered.
All three islands are small, isolated from other islands and remote from the mainland, suggesting low turnover (MacArthur and Wilson 1967). We can thus safely presume that the type-II ecosystems they support are ancient and that their extraordinary herbivores and the plants upon which they subsist have been evolving together for millennia. Research conducted on each of the three islands offers distinct insights into the nature and operation of type-II ecosystems.
TABLE 5.3
Generalist Herbivores of Three Remote Oceanic Islands: Their Population Densities and Biomasses
Christmas Island lies 360 km south of Java in the Indian Ocean and supports only one macroherbivore, the red crab, Becarcoidea natalis. The crabs, weighing up to 500 g, live in burrows on the forest floor at densities estimated at 1.3/m2 (Green 1997). The crabs consume leaf litter and any other edible plant parts that fall to the ground. Crabs as a dominant herbivore are not unusual. Related species occupy scores of islands in the Pacific Ocean and the mangrove zone of tropical shorelines around the world (Sherman 2002).
The crabs of Christmas Island have recently come under threat, but in a way that initiated a fortuitous experiment. In a tragic but typical inadvertency, the notoriously destructive yellow crazy ant, Anoplolepis gracilipes, arrived on Christmas Island over 70 years ago. For decades it remained at low density until 1989, when huge, multiqueened, “supercolonies” were noticed. Since then, the ant has been spreading in a front across the island with worker densities reaching thousands/m2 (O’Dowd et al. 2003). Crabs have no defense against the ants and are killed by them so that ant-occupied zones have become crabless. The slow spread of the ant across the island allowed investigators to compare tracts of forest with and without crabs.
Removal of the island’s dominant herbivore has resulted in a stunning transformation of the vegetation (O’Dowd et al. 2003: figure 5.4). All three trophic levels present on the island have been affected: consumers, producers, and decomposers. In the natural state of the island, crabs consumed most plant matter falling from the canopy: leaves, flowers, and fruits (Green et al. 1999). Seedlings of many species are also consumed (O’Dowd and Lake 1990, Green et al. 1997). Crab foraging thus maintains the forest floor in a condition strikingly reminiscent of that of small Lago Guri islands, bare of leaf litter and most regenerating plants (compare figures 5.2 and 5.4). Extirpation or exclusion of the crab released seeds and seedlings from predation, whereupon the understory quickly became crowded with tree saplings (Green et al. 1997). Seedling diversity jumped from 6 to 22 species per 80 m2 (O’Dowd and Lake 2003). Leaf litter that had previously been consumed by crabs now lay on the forest floor to decompose slowly, as in mainland forests. Portions of Christmas Island that have been invaded by the ant are undergoing a catastrophic shift in vegetation, perhaps as profound as the one we documented on islands in Lago Guri, with the distinction that the change is in response to a release from herbivore pressure rather than the opposite.
Figure 5.4. Understory of forest on Christmas Island, Indian Ocean: Top: Natural state with red crabs. Bottom: Without red crabs after invasion of the yellow crazy ant (Anoplolepis gracilipes) (from O’Dowd et al. 2003, p. 815).
The Aldabra Atoll supports the Aldabra giant tortoise, one of three surviving members of a once-extensive radiation in the western Indian Ocean of up to eight species of tortoises (Gerlach 2004, 2005). Approximately 150,000 tortoises weighting up to 250 kg each occupy the 155 km2 Aldabra Atoll. The atoll consists of several discrete islands, some of which lack surface water and, consequently, tortoises. Occupied portions of the island support tortoise densities of up to 2,700 per km2 (Coe et al. 1979; table 5.3).
The principal islands of the western Indian Ocean, Madagascar, Mauritius, Reunion, and Rodrigues, all harbored giant tortoises that were quickly exterminated, along with the elephant bird, dodo, solitaire, and other species, after humans discovered the islands. Nevertheless, the legacy of the extinct tortoises lives on in the native vegetation as indicated by the presence of many plant species possessing the unusual trait of heterophylly (figure 5.5).
The juvenile leaves of these plants are mostly small and grasslike, not at all resembling the adult leaves. Recently, a team of researchers conducted leaf choice experiments with captive Aldabra tortoises. The tortoises overwhelmingly selected adult over juvenile leaves (figure 5.5) despite greater natural accessibility of the latter (Eskildsen et al. 2004). Moreover, they showed that the transition from juvenile to adult leaf morphology takes place at a height equivalent to the reach of a foraging tortoise (figure 5.6).
Figure 5.5. Heterophylly in some plants of the Mascarene Islands (Mauritius Reunion, and Rodrigues) western Indian Ocean. (from Eskildsen et al. (2004). Juvenile leaves are on the left: a) Diospyros egrettarum, b) Tarenna borbonica, c) Eugenia lucida, d) Cassine orientalis, e) Turraea casimiriana, f) Maytenus pyria, g) Gastonia mauritiana.
The last of the three cases concerns the hutias of East Plana Cay. The Bahamian hutia was thought possibly to be extinct until Garrett Clough confirmed its presence in 1966 on East Plana Cay, a 450 ha island lying to the windward of other Bahamian islands (Clough 1969). Perhaps its small size and windward position served to protect it from invasion by rats (Rattus spp.), for humans, rats, cats, dogs, etc., had long since exterminated the hutia populations of all other Bahamian islands.
The vegetation of East Plana Cay is low, shrubby, and relatively undiverse. The diet of hutias is comprised principally of the foliage, and doubtless other parts, of six common plant species belonging to the following genera: Strumpfia, Conocarpus, Foresteria, Phyllanthus, Croton, and Tournefortia. These include members of families, e.g., Boraginaceae, Combretaceae, Euphorbiaceae, that produce potent antiherbivore defenses, so one can surmise that the vegetation of East Plana Cay is comprised of a selection of the most resistant species from the Bahamian flora (Clough 1972).
Figure 5.6. A. Proportions of adult (black bars) versus juvenile (open bars) leaves of seven heterophyllous plant speces eaten by Aldabra tortoises. B. Vertical ranges of juvenile (white bars) and adult (black bars) foliage of seven heterophyllous plant species. Checkered bars indicate foliage showing transitional morphology. Numbers below the bars refer to sample sizes. The horizontal line represents the browse line for Aldabra tortoises (from Eskildsen et al. 2004).
Persistence of the hutia on only one small island made it highly vulnerable to extinction, prompting Clough and others to establish an additional population by releasing 11 hutias (6 males and 5 females) on Little Wax Cay (24° 53′ N, 76° 47′ W), a small island in the Exuma group, some 300 km to the northwest of East Plana Cay (Campbell et al. 1991). That was in 1973. Twelve years later, in 1985, another investigator estimated the number of hutias on Little Wax Cay at 1200. Four years after that, a third party led by David Campbell returned to the island in April, 1989, to conduct vegetation analysis (Campbell et al. 1991).
Even as one approached Little Wax Cay from the sea, it is obvious that the vegetation of the cay had been massively perturbed. Large areas of the island were bald, without closed, living canopy, in sharp contrast to neighboring cays, which do not have hutias. Many of the trees and shrubs were recently killed and remained as gaunt skeletons, which had not yet decomposed. Closer examination of the cay revealed that large areas were paved with hutia fecal pellets. (Campbell et al. 1991, p. 538)
Campbell et al. go on to state that they found no evidence of seven plant species documented by Russell in a 1958 survey of Little Wax Cay undertaken prior to the introduction of hutias. They conclude that “as the edible plants of Little Wax Cay are being destroyed by hutias, the vegetation of the Cay is likely to become dominated by toxic plants, and it is inevitable that the population of hutias on the Cay will soon begin to fall” (Campbell et al 1991).
The results of Campbell et al. clearly indicate that the vegetation of Little Wax Cay was lacking in defenses against herbivory prior to the introduction of hutias. Whether hutias had ever previously been on the island is not known, but they had presumably been absent for at least 100 years prior to the introduction, allowing time for the vegetation to adjust to type-I conditions. Similar uncertainty applies to the history of East Plana Cay, as well. The Bahamas once supported a large owl that might have controlled hutias, but the owl has been extinct for several thousand years since the Bahamas were colonized by humans (Steadman et al. 2007).
Plants of type-II insular ecosystems do carry anti-herbivore defenses—but only against native herbivores. Defenses found in the vegetation of type-II islands are various, depending on the accessibility of propagules and/or foliage to native herbivores. On Christmas Island, where terrestrial crabs are the herbivore, defenses are expressed at the propagule (seed and seedling) stage (Green et al. 1997); on Aldabra and other islands of the Western Indian Ocean, where tortoises were the principal herbivore, it is at the stage of juvenile leaves; and on East Plana Cay, where a mammal capable of climbing is the selective agent, conventional chemical defenses are expressed in mature foliage (Campbell et al. 1991). Given that native herbivores of type-II islands are often earthbound, like crabs and tortoises, they might select for height-limited defenses that would prove ineffective against introduced mammals like goats or cattle. Height-limited defenses are also found in African acacias, though the height at which thorns cease to be produced is the height of a giraffe (Archibald and Bond 2003).
Herbivore densities in type-II ecosystems are consistently high multiples of those observed in type-III systems on continental mainlands. This was true both for the secondary type-II systems of Lago Guri islands and the three primary type-II systems described just above. Hyperabundant herbivores thus appear to be characteristic of type-II systems. Transitions from type-II to type-I or from type-III to type-II ecosystems may entail what Scheffer et al. (2001) have termed “catastrophic regime shifts” involving major changes in plant species composition.
The intense herbivore pressure that prevails in type-II systems could be expected to drive plant-herbivore arms races. To this point there is little evidence, though consistently high herbivore densities suggest that the herbivores “win.” Plant investment in antiherbivore defenses necessarily entails trade-offs with growth and reproduction and must therefore be self-limiting (Coley et al. 1985). Animals subsisting on heavily defended plant material may themselves experience decrements in growth and reproductive performance, but such decrements may not be strongly disadvantageous in the context of predator-free type-II islands. In the language of foraging ecology, the herbivores of type-II systems become energy maximizers instead of time minimizers (MacArthur and Pianka 1966).
Any plants that were fully resistant to a resident herbivore could take over an island like Aldabra or East Plana Cay and shut out the herbivores, but that does not appear to happen. Plant diversity on type-II islands appears to be low, but it is far from zero. Hyperabundant herbivores thus fail to eliminate plant diversity and persist on type-II islands, presumably for millennia. This could be understood if selection favored herbivore genotypes that could tolerate the defenses of the most common plant species. Such frequency-dependent selection would prevent monopolization of the vegetation by any one plant species and would help stabilize plant diversity, though perhaps at a low level compared to type-III systems.
The evolution of plant defenses is usually considered in relation to the feeding preferences of herbivores, but defenses can also serve as a currency of interspecific competition between plants (Blossey and Nötzold 1995). Fast-growing, weakly defended plants should predominate under low herbivory, such as in type-I systems. Where predators regulate herbivore densities, herbivore pressure is likely to fluctuate in both space and time, establishing a regime of lottery competition (Chesson and Warner 1981). Plants sharing a common herbivore could display reciprocal demography, just as do prey species sharing a common predator (Holt 1977). Thus, a regime of low, patchy herbivory (type III) could be expected to maintain higher overall levels of plant diversity than one without herbivory (type I) or continuously high herbivory (type II). In the absence of herbivory, interspecific competition between plant species would limit diversity, whereas under intense herbivory, only species with strong defenses could persist (Lubchenco 1978). An analogy to the intermediate disturbance hypothesis seems apt here (Connell 1978, Molino and Sabatier 2001). If herbivore pressure proves to be a strong regulator of plant diversity on islands, then the presence/absence of generalist herbivores could act as a major biotic filter for plant species composition superimposed on the traditional geographic filters of area, isolation, and elevation.
How does the EEH intersect with classical island biogeography? Perhaps the intersection is broader than we currently imagine. Productivity and herbivory have not been major issues in island biogeography. Investigators have most often focused on the number of species of birds or lizards or, less commonly, other groups, such as bats, ants, and beetles. Inspired by MacArthur and Wilson (1967), investigators have overwhelmingly fixated on the physical parameters of area, isolation, and elevation, while remaining largely blind to the potential of interisland variation in biotic conditions to contribute to explanations of biogeographic patterns. An outstanding exception to this statement is found in the prescient work of Schoener and his colleagues (see their chapter in this volume).
Development of a more holistic view of island biogeography, one that takes into account both physical and biotic variables, has been hindered by the lack of a biotic complement to the MacArthur-Wilson theory. Here I suggest that the EEH, and modifications thereof, can provide the missing biotic complement. I’m not suggesting that the EEH, or anything like it, can substitute for MacArthur-Wilson. The success of MacArthur-Wilson is outstanding and beyond debate. What I am suggesting is that the biotic conditions of an island can, and undoubtedly do, contribute to explaining such biogeographic features as the presence or absence of individual species and the species richness of a particular taxon.
To support this contention, I offer four highly abbreviated examples. (1) MacArthur himself was puzzled by a phenomenon he termed “density overcompensation” (MacArthur et al. 1972). The term refers to the oft-repeated finding of greater total bird densities on islands than in similar habitat on the corresponding mainland, notwithstanding greater species diversity on the latter. We observed density overcompensation in birds on Lago Guri islands and obtained evidence pointing to bottom-up (productivity) effects associated with the presence of howler monkeys at hyperabundant densities and a concomitant acceleration of nutrient cycling (Feeley and Terborgh 2005, 2006, 2008). Top-down effects (reduced predation) could also help to explain density overcompensation. (2) Diamond’s (1975) famous “checkerboard” distributions represent a biotic mechanism (competitive exclusion) that operates to regulate the presence/absence of individual species on particular islands (see Simberloff and Collins, this volume). (3) Schoener and Spiller (1996) have shown that spider diversity on tiny Bahamian islets is strongly regulated from the top down by the presence or absence of the lizard Anolis sagrei, an important predator of spiders. (4) Exogenous inputs, such as nutrients withdrawn from the sea and transported to seabird nesting islands as fish and manure, can transform the vegetation of entire islands in a bottom-up effect (Croll et al. 2005).
It is likely that one could find many more examples to add to these if one searched the literature. Suffice it to say that biotic interactions of various kinds, including bottom-up and top-down effects, can contribute to a more complete understanding of island biology.
These speculations lead us to reconsider the nature of island vegetation in relation to the exploitation ecosystem hypothesis. The smallest islands should support type-I ecosystems. The relevant range of island areas has not been determined, but the presence of crabs and/or reptilian herbivores on islands of less than 1 km2 suggests that most tropical type-I islands must be tiny (Burness et al. 2001). Even mammals can persist on some very small islands. East Plana Cay is only 4.5 km2 and Little Swan Island, which supported an endemic hutia until domestic cats were released onto it in the 1960s, is only 2.5 km2 (Morgan and Woods 1986). Islands supporting type-II ecosystems were probably once numerous in the world’s oceans in all but the most remote (and perhaps high-latitude) locations. Plant species native to such islands must have carried defenses against resident herbivores, but, as practically all such islands are now inhabited by man and his commensals, the ecosystems of extremely few survive intact. Predators enter the picture on much larger islands where they maintain herbivores at the low densities typical of type-III ecosystems (Burness et al. 2001).
Finally, the world’s largest islands (e.g., Madagascar, New Guinea, New Zealand) once carried complete ecosystems, replete with top carnivores and megaherbivores (here defined operationally as herbivorous animals large enough to escape predation as adults; Burness et al. 2001). Megaherbivores, like the hyperabundant herbivores of type-II ecosystems, are capable of overriding all but the most assertive antiherbivore defenses, so we could expect that relatively undefended plant species would be relegated to fugitive status as ephemerals or gap colonists, or confined to rock faces or other inaccessible sites, as is the case of a number of highly endangered plants of the Hawaiian archipelago (Carlquist 1970).
Megaherbivores have roamed the continental landmasses of the earth since the early Mesozoic, with only a temporary hiatus after the end-Cretaceous extinctions. As recently as the late Pleistocene, proboscidians (elephants) of several genera were found on all continents except Australia and Antarctica. Judging from the known distribution of elephants in Africa today, proboscidians were ubiquitous generalists, ranging essentially everywhere between the extremes of rainfall, temperature, and elevation gradients. Even now, African elephants occur from the edge of the Sahara to the Cape of Good Hope, from the Indian Ocean to the Atlantic, and from the lowlands of the Congo Basin to above timberline on Mt. Kilimanjaro and Mt. Kenya (Coe 1967, Owen-Smith 1988). The ubiquity of proboscidians in Africa, and their former presence elsewhere in the world, including the high Arctic, underscores the extreme implausibility of climate change as the factor responsible for the disappearance of proboscidians and other megafauna from all parts of the world except Africa and southern Asia (Barnosky et al. 2004).
Unfortunately, the EEH does not consider megaherbivores, an oversight that exemplifies the shifting baseline of our anthropocentric society. Nevertheless, the EEH can be extended quite simply by adding a type-IV regime to accommodate megaherbivores, but there remain some questions about the range of productivity levels that would support type-II, -III, and -IV ecosystems.
It stands to reason that, if type-IV ecosystems once occupied all but the most extreme situations within continents, type-III ecosystems would have occupied very limited areas. Indeed, given the prehistoric ubiquity of megaherbivores and their island counterparts, such as the elephant bird, giant tortoises, and moas, it is reasonable to wonder whether Type-III ecosystems ever existed other than on islands. Today, elephants are found in areas of extremely low productivity in the Namibian desert where rainfall is less than 100 mm/yr (Viljoen 1989). Referring back to figure 5.1, that would place the threshold to type-IV ecosystems at the far left of the diagram at a level of productivity around 0.1 kg/m2yr-1.
We can thus surmise that type-IV ecosystems occupied more than 90% of the unglaciated, nondesert habitat of the planet since the Mesozoic (extinction crises and their aftermaths excepted). Type-I, -II, and -III ecosystems would have been relegated primarily to islands where water barriers filtered the colonization of large vertebrates (Holt 1996). The type-II and -III ecosystems that now occupy most of the more-or-less “natural” habitat remaining on the continents are therefore of recent anthropogenic origin.
To summarize, I propose that the four ecosystem states, I, II, III, and IV, comprise a trophic cascade in herbivory (table 5.4). As in more conventional top-down trophic cascades, successive states are characterized by alternating, high (types II and IV) and low (types I and III) levels of herbivory (Paine 1980, Scheffer et al. 2001). Plant defenses should adapt to herbivore pressure through natural selection, induced responses, and/or species selection based on constitutive properties. Plant diversity should be low in the absence of herbivory (type I; pure bottom-up forcing) and in the presence of hyperabundant herbivores or megafauna (types II and IV; strong top-down forcing); it should be high in the presence of predators that cause a moderate level of herbivory to fluctuate in space and time (type III; mixed top-down and bottom-up forcing).
TABLE 5.4
The Trophic Cascade in Herbivory
I grant that some of this is unabashed speculation, but everything I propose can be supported or refuted by appropriate empirical tests. Those desiring to conduct such tests should not delay. Already, more than 90% of the earth’s ice-free terrain has been fundamentally altered. Continental areas were generally type-IV until human-mediated overkill liquidated megaherbivores nearly everywhere. Now, type-IV ecosystems remain only in small and shrinking portions of Africa and southern/southeastern Asia. The remainder of continental earth has relaxed to type-III conditions (lacking megaherbivores but retaining large carnivores such as wolves and jaguars) or type-II conditions (large carnivores eliminated and native herbivores replaced by livestock; Valone et al. 2001). The implications for conservation of this trophic downgrading of the earth’s ecosystems are largely unexplored. The best chances for finding examples of type-I, -II, and -III ecosystems that have arisen naturally and are still undegraded must remain among the world’s islands. Sadly, very few islands remain anywhere that have not undergone anthropogenic shifts in state. Documenting the ecology of these last remaining intact islands before alien species arrive and transform them should be a research goal of the highest priority.
I wish to express my deep gratitude to Lauri Oksanen for his friendship and for the inspiration his ideas have given me, for they have opened my eyes to a diversity of island ecosystems I had never previously imagined. I am grateful to Luis Balbas and to EDELCA (Electrificación del Caroní) for long-standing support of the Lago Guri project. Financial support from the MacArthur Foundation and National Science Foundation is gratefully acknowledged (DEB-9707281, DEB-0108107). I also thank two reviewers for insightful and helpful comments.
Allen, J. D., W. Barnthouse, R. A. Prestbye, and D. R. Strong. 1973. On foliage arthropod communities of Puerto Rican second growth vegetation. Ecology 54:628–37.
Aponte, C., G. R. Barreto, and J. Terborgh. 2003. Consequences of habitat fragmentation on age structure and life history in a tortoise population. Biotropica 35:550–55.
Archibald, S., and W. J. Bond. 2003. Growing tall vs. growing wide: tree architecture and allometry of Acacia karroo in forest, savanna, and arid environments. Oikos 102:3–14.
Barnosky, A. D., P. L. Koch, R. S. Feranec, S. L. Wing, and A. B. Shabel. 2004. Assessing the causes of Late Pleistocene extinctions on the continents. Science 306:70–75.
Beschta, R. L., and W. J. Ripple. 2008. Wolves, trophic cascades, and rivers in the Olympic National Park, USA. Ecohydrology 1:118–30.
Blossey, B., and R. Nötzold. 1995. Evolution of increased competitive ability in invasive nonindigenous plants: A hypothesis. Journal of Ecology 83:887–89.
Bowen, L., and D. van Vuren. 1997. Insular endemic plants lack defenses against herbivores. Conservation Biology 11:1249–54.
Burness, G. P., J. M. Diamond, and T. Flannery. 2001. Dinosaurs, dragons, and dwarfs: the evolution of maximal body size. Proceedings of the National Academy of Sciences U.S.A. 98:14518–23.
Campbell, D. G., K. S. Lowell, and M. E. Lightbourn. 1991. The effect of introduced hutias (Geocapromys ingrahami) on the woody vegetation of Little Wax Cay, Bahamas. Conservation Biology 5:536–41.
Carlquist, S. 1965. Island Life. Garden City, NY: The Natural History Press.
———. 1970. Hawaii: A Natural History. Garden City, NY: The Natural History Press.
———. 1974. Island Biology. New York: Columbia University Press.
Chesson, P. L., and R. R. Warner. 1981. Environmental variability promotes coexistence in lottery competitive systems. American Naturalist 117:123–43.
Clough, G. C. 1969. The Bahamian hutia: A rodent refound. Oryx 10:106–9.
———. 1972. Biology of the Bahamian hutia, Geocapromys ingrahami. Journal of Mammalogy 55:670–72.
Coblentz, B. E. 1978. The effects of feral goats (Capra hircus) on island ecosystems. Biological Conservation 13:279–86.
Coe, M. J. 1967. The Ecology of the Alpine Zone of Mount Kenya. Monographiae Biologicae No. XVII. The Hague: W. Junk.
Coe, M. J., D. Bourn, and I. R. Swingland. 1979. The biomass, production and carrying capacity of giant tortoises on Aldabra. Philosophical Transactions of the Royal Society of London, Series B 286:163–76.
Coe, M. J., D. H. Cummings, and J. Phillipson. 1976. Biomass and production of African large herbivores in relation to rainfall and primary production. Oecologia 22:341–54.
Coley, P. D., J. P. Bryant, and F. S. Chapin III. 1985. Resource availability and plant anti-herbivore defense. Science 230:895–99.
Connell, J. H. 1978. Diversity in tropical rain forests and coral reefs. Science 199:1302–10.
Courchamp, F., M. Langlais, and G. Sugihara. 1999. Control of rabbits to protect island birds from cat predation. Biological Conservation 89:219–25.
Croll, D. A., J. L. Maron, J. A. Estes, E. M. Danner, and G. V. Byrd. 2005. Introduced predators transform subarctic islands from grassland to tundra. Science 307:1959–61.
Diamond, J. M. 1975. Assembly of species communities. In Ecology and Evolution of Communities, ed. M. L. Cody and J. M. Diamond, 342–444. Cambridge, MA: Harvard University Press.
Eskildsen, L. I., J. M. Olesen, and C. G. Jones. 2004. Feeding response of the Aldabra giant tortoise (Geochelone gigantea) to island plants showing heterophylly. Journal of Biogeography 31:1785–90.
Feeley, K. J., and J. W. Terborgh. 2005. The effects of herbivore density on soil nutrients and tree growth in tropical forest fragments. Ecology 86:116–24.
———. 2006. Habitat fragmentation and effects of herbivore (howler monkey) abundances on bird species richness. Ecology 87:144–50.
———. 2008. Trophic drivers of species loss from forest fragments. Animal Conservation 11:366–68.
Gerlach, J. 2004. Giant Tortoises of the Indian Ocean. Frankfurt am Main: Edition Chimaira.
———. 2005. Interpreting morphological and molecular data on Indian Ocean giant tortoises. In African Biodiversity: Molecules, Organisms, Ecosystems, ed. B. A. Huber, B. J. Sinclair, and K-H. Lamp, 213–19. New York: Springer.
Green, P. T. 1997. Red crabs in rain forest on Christmas Island, Indian Ocean: Activity patterns, density and biomass. Journal of Tropical Ecology 13:17–38.
Green, P. T., P. S. Lake, and D. J. O’Dowd. 1999. Monopolization of litter processing by a dominant land crab on a tropical oceanic island. Oecologia 199: 435–44.
Green, P. T., D. J. O’Dowd, and P. S. Lake. 1997. Control of seedling recruitment by land crabs in rain forest on a remote oceanic island. Ecology 78:2474–86.
Hairston, N. G., F. E. Smith, and L. B. Slobodkin. 1960. Community structure, population control, and competition. American Naturalist 94:421–44.
Holt, R. D. 1977. Predation, apparent competition and the structure of prey communities. Theoretical Population Biology 12:197–229.
———. 1996. Food webs in space: An island biogeographic perspective. In Food Webs: Integration of Pattern and Process, ed. G. A. Polis and K. O. Winemiller, 313–23. New York: Chapman and Hall.
Hurtt, G. C., and S. W. Pacala. 1995. The consequences of recruitment limitation: Reconciling chance, history, and competitive differences between plants. Journal of Theoretical Biology 176:1–12.
Iverson, J. B. 1982. Biomass in turtle populations: A neglected subject. Oecologia 55:69–76.
James, H. F., and D. A. Burney. 1997. The diet and ecology of Hawaii’s extinct flightless waterfowl: Evidence from coprolites. Biological Journal of the Linnean Society 62:279–97.
Janzen, D. H. 1973a. Dissolution of mutualism between Cecropia and its Azteca ants. Biotropica 5:15–28.
———. 1973b. Sweep samples of tropical foliage insects: Effects of seasons, vegetation types, elevation, time of day, and insularity. Ecology 54:687–708.
———. 1975. Behavior of Hymenaea courbaril when its predispersal seed predator is absent. Science 189:145–47.
Lambert, T. D., G. H. Adler, C. M. Riveros, L. Lopez, R. Ascanio, and J. Terborgh. 2003. Rodents on tropical land-bridge islands. Journal of Zoology (London) 260:179–87.
Lopez, L., and J. Terborgh. 2007. Seed predation and seedling herbivory as factors in tree recruitment failure on predator-free forested islands. Journal of Tropical Ecology 23:129–37.
Lubchenco, J. 1978. Plant species diversity in a marine intertidal community: Importance of herbivore food preference and algal competitive abilities. American Naturalist 112:23–39.
MacArthur, R. H., J. M. Diamond, and J. R. Karr. 1972. Density compensation in island faunas. Ecology 53:330–42.
MacArthur, R. H., and E. Pianka. 1966. On optimal use of a patchy environment. American Naturalist 100:603–9.
MacArthur, R. H., and E. O. Wilson. 1967. The Theory of Island Biogeography. Princeton, NJ: Princeton University Press.
Molino, J-F., and D. Sabatier. 2001. Tree diversity in tropical rain forests: A validation of the intermediate disturbance hypothesis. Science 294:1702–4.
Morales, L. C., and S. Gorzula. 1986. The interrelations of the Caroní River Basin ecosystems and hydroelectric power projects. Interciencia 11:272–7.
Morgan, G. S., and C. A. Woods. 1986. Extinction and the zoogeography of West Indian land mammals. Biological Journal of the Linnean Society 28: 167–203.
O’Dowd, D. J., P. T. Green, and P. S. Lake. 2003. Invasional “meltdown” on an oceanic island. Ecology Letters 6:812–17.
O’Dowd, D. J., and P. S. Lake. 1990. Red crabs in rainforest, Christmas Island: differential herbivory of seedlings. Oikos 58:289–92.
Oksanen, L., S. D. Fretwell, J. Arruda, and P. Niemelä. 1981. Exploitation ecosystems in gradients of primary productivity. American Naturalist 118: 240–61.
Oksanen, L., and T. Oksanen. 2000. Logic and realism of exploitation ecosystems hypothesis. American Naturalist 155:703–23.
Orihuela Lopez, G., J. Terborgh, and N. Ceballos. 2005. Food selection by a hyperdense population of red howler monkeys (Alouatta seniculus). Journal of Tropical Ecology 21:445–50.
Owen-Smith, R. N. 1988. Megaherbivores: The influence of very large body size on ecology. Cambridge: Cambridge University Press.
Paine, R. T. 1980. Food webs: Linkage, interaction strength, and community structure. Journal of Animal Ecology 49:667–85.
Rao, M., J. Terborgh, and P. Nuñez. 2001. Increased herbivory in forest isolates: Implications for plant community structure and composition. Conservation Biology 15:624–33.
Scheffer, M., S. Carpenter, J. A. Foley, and B. Walker. 2001. Catastrophic shifts in ecosystems. Nature 413:591–96.
Schoener, T. W. 1989. Food webs from the small to the large. Ecology 70: 1559–89.
Schoener, T. W., and D. A. Spiller. 1996. Devastation of prey diversity by experimentally introduced predators in the field. Nature 381:691–94.
Sherman, P. M. 2002. Effects of land crabs on seedling densities and distributions in a mainland neotropical rain forest. Journal of Tropical Ecology 18: 67–89.
Spiller, D. A., and T. W. Schoener. 1990. A terrestrial field experiment showing impact of eliminating top predators on foliage damage. Nature 347:469–72.
Steadman, D. W. 2006. Extinction and biogeography of tropical Pacific birds. Chicago: University of Chicago Press.
Steadman, D. W., R. Franz, G. S. Morgan, N. A. Albury, B. Kakuk, K. Broad, S. E. Franz, K. Tinker, M. P. Pateman, T. A. Lott, D. M. Jarzen, and D. L. Dilcher. 2007. Exceptionally well preserved late Quarternary plant and vertebrate fossils from a blue hole on Abaco, The Bahamas. Proceedings of the National Academy of Science U.S.A. 104:19897–902.
Terborgh, J., and J. Faaborg. 1980. Saturation of bird communities in the West Indies. American Naturalist 116:178–95.
Terborgh, J., K. Feeley, M. Silman, P. Nuñez, and B. Balukjian. 2006. Vegetation dynamics of predator-free land-bridge islands. Journal of Ecology 94:253–63.
Terborgh, J., L. Lopez, and J. Tello. 1997a. Bird communities in transition: The Lago Guri Islands. Ecology 78:1494–501.
Terborgh, J., L. Lopez, J. Tello, D. Yu, and A. R. Bruni. 1997b. Transitory states in relaxing land bridge islands. In Tropical Forest Remnants: Ecology, Management, and Conservation of Fragmented Communities, ed. W. F. Laurance and R. O. Bierregaard, Jr., 256–74. Chicago: University of Chicago Press.
Valone, T. J., M. Meyer, J. H. Brown, and R. M. Chew. 2001. Timescale of perennial grass recovery in desertified arid grasslands following livestock removal. Conservation Biology 16:995–1002.
Viljoen, P. J. 1989. Spatial distribution and movements of elephants (Loxodonta africana) in the northern Namib Desert region of the Kaokoveld, South West Africa/Namibia. Journal of Zoology 219:1–19.
