chapter eighteen
The alteration of biotic interactions
in fragmented tropical forests
william f. laurance
Smithsonian Tropical Research Institute and National Institute for Amazonian Research (INPA)
Introduction
Tropical rainforests are renowned for their ecological complexity and the seeming
ubiquity of coevolved relationships among species (Janzen 1969; Gilbert
1980). Unfortunately, these forests are being destroyed and fragmented at alarming
rates, to the extent that many tropical protected areas are becoming virtual
islands in a sea of heavily degraded land (Laurance & Bierregaard 1997; DeFries
et al. 2002). The future of tropical biodiversity will be largely determined by
the extent to which natural ecological processes and communities can be maintained
in isolated fragments of forest.
Here I summarize available information on the alteration of biotic interactions,
such as predation and key symbioses like pollination and seed dispersal,
in fragmented tropical forests. This review is necessarily preliminary, given the
great diversity of species and ecological interactions in the tropics and the fact
that the alteration of biotic linkages is among the most poorly understood consequences
of habitat fragmentation.
Initial impacts of fragmentation
Forest fragmentation leads to the reduction and isolation of remnant forest
patches. Although each fragmented landscape is unique, a common pattern is
that most forest fragments in human-dominated landscapes are small, ranging
in size from a hectare or less to a few hundred hectares (Gascon et al. 2000;
Cochrane & Laurance 2002). The biota of such small forest patches are expected
to be vulnerable to edge effects and many other negative consequences of fragmentation
(e.g. Laurance et al. 1998, 2002).
As forest-clearing proceeds, some species that are initially present in a landscape
may be completely extirpated by habitat loss, and others may persist in tiny
numbers in fragments, especially if they are rare or patchily distributed (termed
the sample effect; Wilcox 1980). Forest-dependent species will be demographically
and genetically isolated to varying degrees in fragments, depending on
Biotic Interactions in the Tropics: Their Role in the Maintenance of Species Diversity, ed. D. F. R. P. Burslem,
M. A. Pinard and S. E. Hartley. Published by Cambridge University Press. C Cambridge University Press 2005.
442 W I L L I A M F . L A U R A N C E
Increased forest disturbance
Elevated tree mortality
Invasion of disturbance-adapted butterflies
Leaf-litter ant community composition
Invasion of disturbance-adapted beetles
Leaf-litter invertebrate spp. composition
Leaf-litter abundance & spp. richness
Altered height of greatest foliage density
Lower relative humidity
Faster recruitment of disturbance-adapted trees
Reduced canopy height
Reduced soil moisture
Lower canopy-foliage density
13C in understorey leaves
Altered air temperature
Increased temperature & vapour pressure deficit
Reduced understorey-bird density
Elevated litterfall
PAR penetration to understorey
Lower relative humidity
Number of treefall gaps
13C in understorey air
Higher understorey-foliage density
Altered seedling growth
Invasion of disturbance-adapted plants
Leaf relative water contents
Soil moisture content
Vapor pressure deficit
Leaf conductance
Phosphorus content of falling leaves
Invasion of disturbance-adapted plants
Increased recruitment of Cecropia
Fungal fruiting body density
Edgeparameter
0 100 200 300 400 500
Edge-penetration distance (m)
Figure 18.1 Penetration distances of different edge effects into Amazonian forest
remnants (after Laurance et al. 2002). The distances shown here are typical; landscape
factors such as edge age and the type of modified vegetation adjoining fragments can
also have a significant influence on the intensity and penetration of edge effects.
their intrinsic characteristics (such as their mobility and tolerance of degraded
habitats; Diamond et al. 1987; Laurance 1991) as well as attributes of the particular
landscape (such as the degree to which the fragment is isolated from
other forests and the types of degraded lands surrounding the fragment). When
isolated, small populations (< 100 individuals) are vulnerable to local extinction
from random demographic and genetic effects, frequently exacerbated by
environmental variations and local catastrophes (Leigh 1981; Shafer 1981; Lande
1988).
In addition to physical isolation, an important driver of ecological change in
forest fragments is edge effects (Fig. 18.1) – diverse physical and biotic alterations
associated with the abrupt, artificial margins of habitat fragments (Janzen 1986;
A L T E R E D I N T E R A C T I O N S I N F R A G M E N T E D F O R E S T S 443
Lovejoy et al. 1986; Murcia 1995; Laurance et al. 2002). Fragmented tropical forests
often have sharply elevated tree mortality, a proliferation of disturbance-adapted
trees and vines (Viana et al. 1997; Laurance et al. 1998), and hotter, drier conditions
near forest boundaries (Kapos 1989). Most edge effects penetrate anywhere
from a few metres to a few hundred metres into fragments, although some
phenomena (such as surface fires and certain invading species) can penetrate as
far as 2–10 km into forests (Laurance 2000). As a result of such changes, some
forest-dependent species decline near fragment margins whereas disturbanceadapted
and generalist species increase, sometimes dramatically (Lovejoy et al.
1986; Restrepo et al. 1999; Laurance et al. 2002).
Fragmented forests are not merely reduced and isolated; they also face additional
threats from hunting, logging, surface fires, pollution, and climatic and
atmospheric change (e.g. Peres 2001; Weathers et al. 2001; Cochrane & Laurance
2002; Milner-Gulland et al. 2003; Laurance 2004). The impacts of such ancillary
threats on fragmented populations are poorly understood, in part because they
vary greatly in intensity among different landscapes and regions, complicating
efforts to develop simple generalizations. However, because hunting is nearly
ubiquitous and can have important impacts on tropical vertebrates and their
ecological roles in fragmented forests (Cullen et al. 2000; Peres 2001), its effects
will be considered briefly here.
Many species are highly vulnerable in fragmented and hunted tropical landscapes,
including large predators and herbivores, many understorey birds, some
primates, a variety of insects and various plant species (Lovejoy et al. 1986;
Didham et al. 1998; Restrepo et al. 1999; Laurance et al. 2002). Clearly, the diverse
changes in species composition, species abundances, forest dynamics and physical
environment alter some – and possibly many – biotic interactions in tropical
forest fragments. As discussed below, what is far more poorly understood is
the degree to which these altered interactions accelerate local species extinctions
or disrupt key ecological processes in fragmented forests, especially over
longer time scales ranging from centuries to millennia (e.g. Kellman et al. 1996;
Corlett & Turner 1997).
Pollination
Most tropical plants are pollinated by insects, birds or bats, and rely on pollinators
to maintain genetic diversity through outcrossing (Bawa 1990; Ricklets
& Renner 1994; Momose et al. 1998). Recent analyses confirm that inbreeding
depression in small, isolated populations of plants can reduce seed set, germination,
survival and resistance to stress; and that populations with reduced
genetic variation can experience reduced growth and increased extinction rates
(cf. Keller and Waller 2002 and references therein).
Three factors largely determine the vulnerability of a plant species to the
loss of its pollinators in fragmented forests. The first is the degree to which the
plant is reliant on pollinators for reproduction. Many tropical plants are obligate
444 W I L L I A M F . L A U R A N C E
outbreeders, being either dioecious or self-incompatible, or possess mixed mating
systems that are highly outcrossed through various mechanisms that select
for non-self pollen or outcrossed seed (Bawa & Opler 1975; Renner & Ricklefs
1995; Kenta et al. 2002). Many species are also rare, with effective populations
often encompassing hundreds to many thousands of hectares (Murawski &
Hamrick 1991; Chase et al. 1996; Stacy et al. 1996; Nason et al. 1998), and they frequently
have clumped distributions, and thus can be vulnerable to population
isolation and an erosion of genetic diversity. Hence, for most tropical plants,
animal pollinators are critical.
The second factor is the degree to which the plant depends on specific pollinator
species. Plants that rely on just one or a few pollinator species may be
more vulnerable to habitat disturbance than are those can utilize a suit of different
pollinators (but see Aizen et al. 2002). Only a relatively small number
of plant species, such as orchids and figs, have obligate relationships with a
single bee or wasp pollinator (Bawa 1990). A restricted suite of pollinators is
also common on faunally depauperate islands, where a number of plants rely
on specific birds (Cox 1983) or flying foxes (Cox et al. 1991; Rainey et al. 1995)
for pollination and seed dispersal. The majority of other plant species, however,
have a higher degree of redundancy in their pollinators (e.g. Horvitz &
Schemske 1990). Notably, most tropical trees rely on insects and are pollination
generalists (Ashton 1969; Bawa & Opler 1975; Stacy et al. 1996; Momose et al.
1998).
The third factor is the impact of habitat fragmentation on the pollinators. At
present, only a few generalizations can be made about the responses to fragmentation
of pollinators, as species respond to habitat fragmentation in individualistic
ways depending on their tolerance of edge and degraded habitats, population
density, trophic position and other factors (Terborgh 1974; Laurance 1991;
Didham et al. 1998). Among Amazonian butterflies, for example, a number of
forest-interior species (especially in the Ithomiinae) decline in forest fragments,
whereas light-loving generalist species increase (Brown & Hutchings 1997). Of 12
euglossine-bee species studied by Powell & Powell (1987), four declined in abundance
in recently isolated Amazonian forest fragments but other euglossines
increased in the modified habitats surrounding the fragments (see also Becker
et al. 1991). Larger pollinators, such as many hummingbirds (Aizen & Feinsinger
1994a; Stouffer & Bierregaard 1995; Aldrich & Hamrick 1998) and nectar- and
fruit-eating bats (Kalko 1998; Law & Lean 1999; Sampaio 2000), appear relatively
resilient to habitat fragmentation and can often traverse the degraded lands
surrounding fragments (Steffan-Dewenter & Tscharntke 1999). When subjected
to both fragmentation and hunting, however, flying foxes have often declined
precipitously, especially on oceanic islands (Cox et al. 1991; Rainey et al. 1995).
Even when pollinators do not decline markedly in fragmented forests, changes
in the distribution of resources can alter pollinator foraging behaviour with
A L T E R E D I N T E R A C T I O N S I N F R A G M E N T E D F O R E S T S 445
respect to flower fidelity and foraging distance, which affects the quality and
quantity of pollen arriving at the plant (Aizen & Feinsinger 1994a; Aldrich &
Hamrick 1998; Ghazoul & McLeish 2001).
To date, only a few studies have provided unequivocal evidence of reduced
plant reproduction in response to fragmentation-induced changes in pollinators.
One of the most important studies is that of Aizen and Feinsinger (1994a),
who worked not in tropical rainforest but in dry subtropical forest in the Argentinian
Chaco. Among 16 plant species they studied (most of which were mainly
bee-pollinated), 13 had reduced seed production in forest fragments. Moreover,
among three self-compatible species, plants in fragments appeared to receive
most pollen from themselves or from closely related neighbours, suggesting
that pollinators were unable to transport pollen between different fragments
(Aizen & Feinsinger 1994a).
Other evidence of negative fragmentation effects is more anecdotal. In the
central Amazon, Dick (2001a) observed an absence of small beetles and stingless
bees on mass-flowering trees in cattle pastures, and suggested that plants that
are pollinated exclusively by small beetles, such as those in the nutmeg family
(Myristicaceae), could be especially vulnerable to fragmentation. In Costa Rican
forest fragments, Ghazoul and McLeish (2001) found no decrease in the abundance
of Trigona bee pollinators for the widespread tree Anacardium excelsum,
but observed that pollination success and seed production was positively correlated
with fragment size. They proposed that Trigona bees rarely move among
fragments, reducing genetic variability, fertilization and seed set among trees
in small fragments. Aldrich and Hamrick (1998) found that reproduction of the
bat-dispersed tree Symphonia globulifera in forest fragments in Costa Rica was
dominated by a few large, isolated trees in pastures that had high rates of selfing,
suggesting that fragmentation might eventually cause a genetic bottleneck
in the population. Finally, Murren (2002, 2003) found that epiphytic orchids
(Catasetum viridiflavum) on small man-made islands in Lake Gatun, Panama had
lower seed set than did plants in nearby mainland areas, suggesting that pollination
by euglossine bees was reduced on islands, although there was no evidence
of reduced gene flow among island populations.
A collapse of plant–pollinator mutualisms has also been observed on oceanic
islands, which are somewhat analogous to habitat fragments in being isolated
and having relatively depauperate biotas. In Hawaii, the loss or severe decline of
native sphingid-moth pollinators has been implicated in the very low fruit set
of endemic Brighamia (Campanulaceae). Moreover, Hawaiian plants like Stenogyne
kanehoana (Lamiaceae) have not set seeds for many years because their honeycreeper
pollinators have been extirpated (Renner 1998). The decimation of flying
foxes on many Pacific islands has also led to reduced reproduction in endemic
plants that rely on these large bats for pollination (Cox et al. 1991; Rainey et al.
1995).
446 W I L L I A M F . L A U R A N C E
In other cases, however, fragmentation has had little effect on plant reproduction
because native pollinators are unaffected or recover quickly from population
declines. Two species of hummingbird-pollinated plants in the Argentinian
Chaco actually had stable or increased reproduction, evidently because hummingbirds
preferred the flower-rich fragments and readily moved among them
(Aizen & Feinsinger 1994a). After being decimated by a hurricane, populations of
fig wasps recovered quickly (within 5–16 months) because the wasps were able to
disperse at least 60 km from other forests that were unaffected by the hurricane
(Bronstein & Hossaert-McKey 1995). Harrison (2003) inferred that some fig-wasp
species in an isolated forest fragment (4500 ha) in Borneo had arrived by dispersing
from forests at least 30 km away, although wasps that pollinated dioecious
figs were poorly represented and appeared to be much more dispersal-limited
than monoecious-fig wasp species. Nason and Hamrick (1997) found that small
populations of the neotropical tree Spondias mombin in forest fragments actually
had higher rates of pollen immigration than did trees in intact forests, suggesting
that fragmentation did not result in genetic isolation of populations. Indeed,
studies of genetic markers suggest that small fragments and isolated trees in
pastures can act as important stepping stones for many pollinators (Boshier
et al. 1995; Chase et al. 1996; Stacy et al. 1996; Dick 2001b; White et al. 2002; Dick
et al. 2003; Murren 2003). Finally, Murcia (1996) found no effect of fragment size
or edge effects on pollination of 16 plant species in Colombian cloud forests.
In some fragmented landscapes, plant reproduction and gene flow have been
maintained despite severe declines of natural pollinators, when exotic animal
species take over the pollination duties. Isolated canopy trees (Dinizia excelsa) in
Amazonian pastures were genetically ‘rescued’ by abundant Africanized honeybees
(Apis mellifera scutellata), which more than compensated for the decline of
native insect pollinators and contributed to an increase in the production of
outcrossed seeds in pasture trees, relative to those in intact forest (Dick 2001a,
2001b; Dick et al. 2003). Similarly, in the Argentinian Chaco, the proportion of
flower visits by Africanized honeybees increased in forest fragments relative to
intact forest (Aizen & Feinsinger 1994b). In Hawaii, a number of endemic plant
species that formerly relied on honeycreepers for pollination are now being pollinated
by introduced bird species (Renner 1998). Although exotic pollinators
help to maintain plant reproduction in some cases, it is possible that they may
displace native pollinators and thereby contribute to the breakdown of natural
plant–pollinator interactions.
In summary, habitat fragmentation and related disturbances alter many
plant–pollinator interactions, but the effects of these changes on plant reproduction
vary considerably among different studies and landscapes. In some cases,
ecological redundancy of pollinators and exotic pollinators limit the impact of
declining native pollinators, but in many other cases plant fecundity is clearly
being reduced. In general, pollen movement among the most isolated fragments
A L T E R E D I N T E R A C T I O N S I N F R A G M E N T E D F O R E S T S 447
is more likely to be mediated by large pollinators, such as birds, bats and large
bees, than by small insects (Steffan-Dewenter & Tscharntke 1999). Surprisingly,
self-compatible and self-incompatible plants may show little overall difference
in their vulnerable to fragmentation, despite the latter being more reliant on
animal pollinators (Aizen et al. 2002). However, canopy trees, which are typically
pollination generalists, could be less vulnerable to fragmentation than are
understorey plants, which rely on fewer pollinator species (Aizen & Feinsinger
1994a; Momose et al. 1998; Renner 1998). Plant–pollinator relationships are often
asymmetrical, with the pollinator being more specialized on the host plant than
vice versa (as in some solitary bees; Renner 1998), and thus host-specific pollinators
could be vulnerable in fragmented forests if their host plants decline.
Seed dispersal, seed predation and herbivory
Although some tropical plants rely on wind, gravity or water to disperse their
propagules, the bulk of species have animal-dispersed fruits or seeds. Birds and
bats are the main dispersers for smaller-seeded plants, whereas terrestrial and
arboreal mammals disperse larger-seeded species (Levey et al. 1994). A distinction
can be drawn between primary dispersers (such as monkeys) that remove fruit
or seeds directly from tree crowns, and secondary dispersers (such as terrestrial
mammals) that remove fallen fruit and seeds (Wright 2003).
Many seed-dispersing animals are also seed predators, although different
species vary greatly in the degree to which they consume seeds; for example,
peccaries and white-tailed deer are almost exclusively predators of large seeds,
whereas tapirs act primarily as large-seed dispersers and agoutis are major seed
predators but aid plant recruitment by caching and burying seeds (Brewer et al.
1997). Insects (notably bruchid beetles) are usually seed predators (Janzen 1970).
Seed dispersal can provide a critical escape from severe density-dependent seed
predation, seedling herbivory and pathogen attack near parent plants (Janzen
1970; Connell 1971), and there is a strong indication that some tree species
would be unable to recruit without mammalian dispersers (Augspurger 1984;
Chapman & Chapman 1995).
How does forest fragmentation affect dispersal and predation of plant
propagules? At least on small man-made islands, these processes appear to
be profoundly disrupted. On tiny (< 1 ha) islands created in the late 1980s in
Venezuela, predators of vertebrates have disappeared and some of their prey,
including rodents, howler monkeys, iguanas and leaf-cutter ants, have increased
10- to 100-fold in density. As a consequence, seedlings and saplings of canopy
trees are severely reduced, providing evidence of a trophic cascade that is ramifying
throughout the animal and plant communities (Terborgh et al. 2001). On
equally small (< 2 ha) but older (c. 80-year-old) islands in Lake Gatun, Panama,
most mammals have disappeared but rats (Proechimys semispinosus) have increased
sharply in abundance, and predation on large seeds is sharply elevated relative
448 W I L L I A M F . L A U R A N C E
to mainland forest (Asquith et al. 1997). On Lake Gatun islands that are too small
to support any mammals, tree communities are dominated by a few large-seeded
species (Putz et al. 1990; Leigh et al. 1993) that may out-compete smaller-seeded
trees when key predators of large seeds, such as agouti and paca, are absent
(Terborgh 1992).
It is unclear, however, whether such trends are typical of ‘real’ forest fragments,
which are surrounded by a matrix of modified land and subjected to
varying levels of hunting pressure. Fragmented forests in Los Tuxtlas, Mexico,
have experienced sharply reduced grazing pressure and increased plant regeneration,
leading to dense monospecific carpets of seedlings on the forest floor
(Dirzo & Miranda 1991). As discussed by Asquith et al. (1997), this may arise from
defaunation of larger (> 1 kg) predatory and herbivorous mammals by hunters
and, perhaps more importantly, from suppression of small-rodent populations
by abundant vipers (Bothrops spp.). The net effect is that plant recruitment at
Los Tuxtlas is very different from that observed on small islands.
Plant recruitment is often altered in fragmented forests. In Costa Rican fragments,
Guariguata et al. (2002) found higher seed predation by small rodents,
and reduced dispersal, for one canopy tree species (Dipterix panamensis) but not
for a second species (Carapa guianensis). Tabarelli et al. (1999) found that small
(< 15 ha) forest fragments of Brazilian Atlantic forest isolated for over 50 years
had more abiotically dispersed trees and shrubs, and fewer animal-dispersed
trees, than did larger (370–7900 ha) fragments, probably because of large-scale
declines of frugivores in small fragments. In Tanzania, a number of tree species
in small (< 9 ha) forest fragments show reduced recruitment, apparently because
of sharp declines in frugivorous birds and primates in the fragments (Cordeiro &
Howe 2001, 2003).
In at least two studies, fragmentation reduced plant recruitment by promoting
an influx of seed predators from surrounding degraded habitats. In a national
park in Indonesian Borneo, large influxes of vertebrate seed predators from
adjoining degraded lands led to a dramatic collapse in recruitment of dipterocarp
canopy trees during a major masting event (Curran et al. 1999). Similar
patterns were observed at Pasoh Forest Reserve in Peninsular Malaysia, where
pigs (Sus scrofa) achieved abnormally high population densities by feeding on
oil palms surrounding the reserve and then severely damaged the understorey
vegetation (Ickes et al. 2001) and dipterocarp recruitment within the reserve.
Moreover, when the surrounding oil palms were felled in 2001, pig populations
crashed and dipterocarp recruitment in the reserve recovered following a general
flowering event the following year (S. J. Wright, personal communication).
Some evidence suggests that large-seeded plants may be more vulnerable than
small-seeded species in fragmented landscapes because the large vertebrates on
which they rely for dispersal, such as ungulates and primates, are often sharply
A L T E R E D I N T E R A C T I O N S I N F R A G M E N T E D F O R E S T S 449
reduced in abundance by the combined effects of hunting and fragmentation
(Fuentes 2000; Peres 2001). In central Panama, hunting and fragmentation had
negative effects on long-distance recruitment of the large-seeded palm Attalea
butyraceae (Wright & Duber 2001). Seed predation and dispersal of seeds away
from presumptive parent trees were reduced on small islands and in areas subjected
to heavy hunting pressure, relative to lightly hunted and intact-forest sites.
Seedling densities were highest around parent trees on islands and in heavily
poached sites, suggesting that human disruption of mammal communities may
eventually reduce local tree diversity by disrupting natural density-dependent
seed predation (Wright & Duber 2001). Small-seeded plants can also be vulnerable
if their key dispersers respond negatively to fragmentation, as occurs in
some species dispersed by birds (e.g. Cordeiro & Howe 2001, 2003) or by dung
beetles (e.g. Chapman et al. 2003).
Forest fragmentation can also affect rates of leaf herbivory, although this phenomenon
has been relatively little studied in the tropics. As discussed above,
Terborgh et al. (2001) found intense herbivory on seedlings and saplings on small
man-made islands in Venezuela, following population explosions of leaf-cutting
ants and other herbivores on the islands. Forest fragments in the Amazon and
Brazilian Atlantic region also show atypically high densities of leaf-cutting ants,
especially near fragment margins, which could increase mortality and turnover
of young trees (Vasconcelos & Cherrett 1995; Laurance 2003). Herbivorous insects
generally increase near the margins of Amazonian forest fragments, possibly
in response to increased leaf flush there (Lovejoy et al. 1986; Didham 1997).
Herbivory from mammals (mainly rodents) on seedlings and saplings was evidently
greater on small islands than on large islands and mainland areas in
Panama (Asquith et al. 1997). With the exception of Benitez-Malvido et al. (1999),
who found little difference in insect damage to leaves between fragmented and
intact Amazonian forests, most studies have found increased herbivory in forest
fragments. The decline of many predators in fragments (see below), and the concomitant
increases in herbivorous vertebrates and insects, may partially explain
this pattern.
It must be emphasized that the effects of habitat fragmentation on plant
recruitment and survival are often context specific, depending on the nature
of the fragmented landscape, its constituent fauna and the relative impact of
hunting on predatory and herbivorous mammals. As proposed by Wright (2003),
the indirect effects of hunting on plant communities could depend on whether
large carnivores are being hunted (such as for the fur trade) or whether hunting
is principally for food, in which case agoutis, deer and other herbivorous mammals
are mainly taken. In the former case, abundances of larger-bodied seed
predators and herbivores may actually increase, whereas in the latter case they
are likely to decline.
450 W I L L I A M F . L A U R A N C E
Predation
Large predators such as jaguars, tigers and some raptors are more vulnerable
to local extinction in fragmented forests than are species at lower trophic
levels, because predators require large home ranges and are frequently persecuted
by humans. The disappearance of large predators may have important
ecological impacts in fragmented landscapes, especially if top-down regulation
of prey species is common (Terborgh 1992). The phenomenon of ‘mesopredator
release’ was coined to describe population explosions of medium-sized omnivores
(e.g. agoutis, coati mundis, opossums, raccoons, coyotes) in fragmented
habitats where large predators have disappeared (Soul´e et al. 1988; Terborgh
1992; Crooks & Soul´e 1999).
Mesopredator release has been documented in fragmented temperate habitats
(Soul´e et al. 1988; Crooks & Soul´e 1999) but its importance in the tropics is
uncertain. The near extirpation of jaguars, pumas and harpy eagles on Barro
Colorado Island, Panama might have led to unusually high densities of mediumsized
mammals such as agoutis, coati mundis, sloths and howler monkeys (Glanz
1982; Terborgh 1992), which could account for elevated nest predation on the
island (Loiselle & Hoppes 1983; Sieving 1992). It is unclear, however, whether
medium-sized mammals are actually hyper-abundant on Barro Colorado Island
or merely appear to be so, as they are habituated to humans and therefore
conspicuous to researchers on the island (Wright et al. 1994). In tropical Australia,
atypically high abundances of native rodents in fragmented rainforests may
result from the loss of larger rainforest-dependent predators, although edgerelated
forest disturbances could also favour rodents (Laurance 1994, 1997).
Another potential impact of forest fragmentation is that generalist omnivores,
seed predators and brood parasites may flood into fragments from surrounding
degraded lands (Curran et al. 1999; Ickes et al. 2001), causing negative impacts on
nesting birds, other small vertebrates and plant communities, especially near
forest edges. This phenomenon has been demonstrated in many fragmented
temperate and boreal forests (e.g. Gates & Gysel 1978; Wilcove 1985; Paton 1994),
but relatively few such studies have been conducted in the tropics. Experimental
studies provided no evidence of increased nest or seed predation near forest
edges in central Amazonia (Meyer 1999; Bruna 2002) and tropical Australia (Laurance
et al. 1993; Laurance & Grant 1994; Harrington et al. 1997), although Lindell
(2000) found increased predation on quail eggs near forest edges in Costa Rica.
It is clear, nevertheless, that important changes in predator and insectivore
communities do occur in fragmented forests. In the central Amazon, predation
on understorey invertebrates may be reduced because insectivorous bats, understorey
birds and army ants all decline sharply (Laurance et al. 2002). In tropical
Australia, larger rainforest-dependent predators, such as spotted-tailed quolls
(Fig. 18.2), rufous owls and larger pythons, decline sharply in forest fragments
and are replaced by smaller, generalist predators like barn owls, red-bellied
A L T E R E D I N T E R A C T I O N S I N F R A G M E N T E D F O R E S T S 451
Figure 18.2 The decline in fragmented landscapes of forest-dependent predators, such as
this spotted-tailed quoll (Dasyurus maculatus) from tropical Australia, may lead to
population explosions of omnivorous mammals that in turn prey on nesting birds,
invertebrates, seeds and other rainforest species (photo by S. E. Williams).
blacksnakes and white-tailed rats (Laurance 1997). As discussed above, at least on
small man-made islands, striking trophic cascades have been observed following
the disappearance of large predators (Terborgh et al. 1997, 2001). Such wholesale
changes in predator communities must alter predation pressure on different
groups of vertebrates and invertebrates.
Other ecological interactions
Fragmentation could potentially alter a great diversity of ecological interactions.
For example, army ants require extensive areas for foraging and disappear from
small forest fragments in the Amazon and East Africa; in turn, obligate antfollowing
birds, which prey on insects that flee the marauding ants, also frequently
disappear (Harper 1989; Bierregaard et al. 1992; Peters & Kramer 2003).
The disappearance of collared peccaries (Tayassu pecari) from Amazonian forest
fragments leads to declines of several frog species that require the muddy pools
created by wallowing peccaries, for reproduction (Zimmerman & Bierregaard
1986). The decline of dung beetles in fragmented forests could have large effects
on the phoretic mites that rely on these beetles for dispersal (Klein 1989). Altered
452 W I L L I A M F . L A U R A N C E
levels of insect herbivory in fragments could potentially affect the vulnerability
of rainforest plants to infection by fungal pathogens (Benitez-Malvido et al.
1999).
It is apparent that these and a great many other ecological interactions are
being altered by habitat fragmentation. For some phenomena, such as population
explosions of leaf-cutting ants along forest-fragment margins (Vasconcelos &
Cherrett 1995; Laurance 2003) that could negatively affect plant recruitment,
we have some inkling of their potential impacts on forest ecosystems. For many
others – such as the potential effects of fragmentation on disease and parasite
dynamics, and on ecological interactions that occur belowground or in the forest
canopy – very little is known. In the future, researchers should continue
to document the remarkably varied impacts of habitat fragmentation on biotic
interactions, but we must also attempt to determine the effects of these alterations
on species survival and key ecological processes, and their generality
across a range of fragmented landscapes.
Acknowledgements
I thank Rhett Harrison, Christopher Dick, Joseph Wright, Helene Muller-Landau,
Heraldo Vasconcelos and three anonymous referees for helpful comments on
drafts of the manuscript.
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