Hunting and plant community dynamics

        Although the direct impacts of defaunation driven by overhunting can be predicted to some degree, higher-order indirect effects on community structure remain poorly understood since Redford’s (1992) seminal paper and may have profound, long-term consequences for the persistence of other taxa, and the structure, productivity and resilience of terrestrial ecosystems (Cunningham et al. 2009). Severe population declines or extirpation of the world’s megafauna may result in dramatic changes to ecosystems, some of which have already been empirically demonstrated, while others have yet to be documented or remain inexact. Large vertebrates often have a profound impact on food webs and community dynamics through mobile-linkage mutualisms, seed predation, and seedling and sapling herbivory. Plant communities in tropical forests depleted of their megafauna may experience pollinationbottlenecks, reduced seed dispersal, monodominance of seedling cohorts, altered patterns of seedling recruitment, other shifts in the relative abundance of species, and various forms of functional compensation (Cordeiro and Howe 2003; Peres and Roosmalen 2003; Wang et al. 2007; Terborgh et al. 2008; Chapter 3). On the other hand, the net effects of large mammal defaunation depends on how the balance of interactions are affected by population declines in both mutualists (e.g. highquality seed dispersers) and herbivores (e.g. seed predators) (Wright 2003). For example, significant changes in population densities in wild pigs (Suidae) and several other ungulates and rodents, which are active seed predators, may have a major effect on seed and seedling survival and forest regeneration (Curran and Webb 2000). 

               Tropical forest floras are most dependent on large-vertebrate dispersers, with as many as 97% of all tree, woody liana and epiphyte species bearing fruits and seeds that are morphologically adapted to endozoochorous (passing through the gut of an animal) dispersal (Peres and Roosmalen 2003). Successful seedling recruitment in many flowering plants depends on seed dispersal services provided by large-bodied frugivores (Howe and Smallwood 1982), while virtually all seeds falling underneath the parent’s canopy succumb to density-dependent mortality—caused by fungal attack, other pathogens, and vertebrate and invertebrate seed predators (see review in Carson et al. 2008). A growing number of phytodemographic studies have examined the effects of large-vertebrate removal. Studies examining seedling recruitment under different levels of hunting pressure (or disperser abundance) reveal very different outcomes. At the community level, seedling density in overhunted forests can be indistinguishable, greater, or less than that in the undisturbed forests (Dirzo and Miranda 1991; Chapman and Onderdonk 1998; Wright et al. 2000), but the consequences of increased hunting pressure to plant regeneration depends on the patterns of depletion across different prey species. In persistently hunted Amazonian forests, where large-bodied primates are driven to local extinction or severely reduced in numbers (Peres and Palacios 2007), the probability of effective dispersal of largeseeded endozoochorous plants can decline by over 60% compared to non-hunted forests (Peres and Roosmalen 2003). Consequently, plant species with seeds dispersed by vulnerable game species are less abundant where hunters are active, whereas species with seeds dispersed by abiotic means or by small frugivores ignored by hunters are more abundant in the seedling and sapling layers (Nuñez-Iturri and Howe 2007; Wright et al. 2007; Terborgh et al. 2008). 

                However, the importance of dispersal- limitation in the absence of large frugivores depends on the degree to which their seed dispersal services are redundant to any given plant species (Peres and Roosmalen 2003). Furthermore, local extinction events in large-bodied species are rarely compensated by smaller species in terms of their population density, biomass, diet, and seed handling outcomes (Peres and Dolman 2000). Large vertebrates targeted by hunters often have a disproportionate impact on community structure and operate as “ecosystem engineers” (Jones et al. 1994; Wright and Jones 2006), either performing a key landscaping role in terms of structural habitat disturbance, or as mega-herbivores that maintain the structure and relative abundance of plant communities. For example, elephants exert a major role in modifying vegetation structure and composition as herbivores, seed dispersers, and agents of mortality for many small trees (Cristoffer and Peres 2003). Two similar forests with or without elephants show different succession and regeneration pathways, as shown by long-term studies in Uganda (Sheil and Salim 2004). Overharvesting of several other species holding a keystone landscaping role can lead to pervasive changes in the structure and function of ecosystems. For example, the decimation of North American beaver populations by pelt hunters following the arrival of Europeans profoundly altered the hydrology, channel geomorphology, biogeochemical pathways and community productivity of riparian habitats (Naiman et al. 1986). Mammal overhunting triggers at least two additional potential cascades: the secondary extirpation of dependent taxa and the subsequent decline of ecological processes mediated by associated species. For instance, overhunting can severely disrupt key ecosystem processes including nutrient recycling and secondary seed dispersal exerted by relatively intact assemblages of dung beetles (Coleoptera: Scarabaeinae) and other coprophagous invertebrates that depend on large mammals for adult and larval food resources (Nichols et al. 2009).

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Tropical forest disturbance

                Timber extraction in tropical forests is widely variable in terms of species selectivity, but even highly selective logging can trigger major ecological changes in the understory light environment, forest microclimate, and dynamics of plant regeneration. Even reduced-impact logging (RIL) operations can generate enough forest disturbance, through elevated canopy gap fracture, to greatly augment forest understory desiccation, dry fuel loads, and fuel continuity, thereby breaching the forest flammability threshold in seasonally-dry forests. During severe dry seasons, often aggravated by increasingly frequent continental- scale climatic events, extensive ground fires initiated by either natural or anthropogenic sources of ignition can result in a dramatically reduced biomass and biodiversity value of previously unburnt tropical forests (Barlow and Peres 2004, 2008). Despite these undesirable effects, large-scale commercial logging that is unsustainable at either the population or ecosystem level continues unchecked in many tropical forest frontiers (Curran et al. 2004; Asner et al. 2005). Yet surface fires aggravated by logging disturbance represent one of the most powerful mechanisms of functional and compositional impoverishment of remaining areas of tropical forests (Cochrane 2003), and arguably the most important climatemediated phase shift in the structure of tropical ecosystems

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Cascading effects of overexploitation on ecosystems

         
              All extractive systems in which the over harvested resource is one or more biological populations, can lead to pervasive trophic cascades and other unintended ecosystem-level consequences to non-target species. Most hunting, fishing, and collecting activities affect not only the primary target species, but also species that are taken accidentally or opportunistically. Furthermore, exploitation often causes physical damage to the environment, and has ramifications for other species through cascading interactions and changes in food webs. In addition, overexploitation may severely erode the ecological role of resource populations in natural communities. In other words, overexploited populations need not be entirely extirpated before they become ecologically extinct. In communities that are “half-empty” (Redford and Feinsinger 2001), populations may be reduced to sufficiently low numbers so that, although still present in the community, they no longer interact significantly with other species (Estes et al. 1989). Communities with reduced levels of species interactions may become pale shadows of their former selves. 

              Although difficult to measure, severe declines in large vertebrate populations may result in multi-trophic cascades that may profoundly alter the structure of marine ecosystems such as kelp forests, coral reefs and estuaries (Jackson et al. 2001), and analogous processes may occur in many terrestrial ecosystems. Plant reproduction in endemic island floras can be severely affected by population declines in flying foxes (pteropodid fruit bats) that serve as strong mutualists as pollinators and seed dispersers (Cox et al. 1991). In some Pacific archipelagos, several species may become functionally extinct, ceasing to effectively disperse large seeds long before becoming rare (McConkey and Drake 2006). A key agenda for future research will involve understanding the non-linearities between functional responses to the numeric abundance of strong interactors reduced by exploitation pressure and the quality of ecological services that depleted populations can perform. For example, what is the critical density of any given exploited population below which it can no longer fulfill its community-wide ecological role? 

                 In this section I concentrate on poorly known interaction cascades in tropical forest and marine environments, and discuss a few examples of how apparently innocuous extractive activities targeted to one or a few species can drastically affect the structure and functioning of these terrestrial and aquatic ecosystems.

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Overexploitation in aquatic ecosystems

              Marine biodiversity loss, largely through overfishing, is increasingly impairing the capacity of the world’s oceans to provide food, maintain water quality, and recover from perturbations (Worm et al. 2006). Yet marine fisheries provide employment and income for 0.2 billion people around the world, and fishing is the mainstay of the economy of many coastal regions; 41 million people worked as fishers or fish farmers in 2004, operating 1.3 million decked vessels and 2.7 million open boats (FAO 2007). An estimated 14 million metric tons of fuel was consumed by the fish-catching sector at a cost equivalent to US$22 billion, or ~25% of the total revenue of the sector. In 2004, reported catches from marine and inland capture fisheries were 85.8 million and 9.2 million tons, respectively, which was worth US$84.9 billion at first sale. Freshwater catches taken every year for food have declined recently but on average 500 000 tons are taken from the Mekong river in South-East Asia; 210 000 tons are taken from the Zaire river in Africa; and 210 000 tons of fish are taken from the Amazon river in South America. Seafood consumption is still high and rising in the First World and has doubled in China within the last decade. Fish contributes to, or exceeds 50% of the total animal protein consumption in many countries and regions, such as Bangladesh, Cambodia, Congo, Indonesia, Japan or the Brazilian Amazon. Overall, fish provides more than 2.8 billion people with ~20% or more of their average per capita intake of animal protein. 

             The oscillation of good and bad years in marine fisheries can also modulate the protein demand from terrestrial wildlife populations (Brashares et al. 2004). The share of fish in total world animal protein supply amounted to 16% in 2001 (FAO 2004). These ‘official’ landing statistics tend to severely underestimate catches and total values due to the enormous unrecorded contribution of subsistence fisheries consumed locally. Although the world’s oceans are vast, most seascapes are relatively low-productivity, and 80% of the global catch comes from only ~20% of the area. Approximately 68% of the world’s catch comes from the Pacific and northeast Atlantic. At current harvest rates, most of the economically important marine fisheries worldwide have either collapsed or are expected to collapse. Current impacts of overexploitation and its consequences are no longer locally nested, since 52% of marine stocks monitored by the FAO in 2005 were fully exploited at their maximum sustainable level and 24% were overexploited or depleted, such that their current biomass is much lower than the level that would maximize their sustained yield (FAO 2007). The remaining onequarter of the stocks were either underexploited or moderately exploited and could perhaps produce more. The Brazilian sardine (Sardinella brasiliensis) is a classic case of an overexploited marine fishery. In the 1970s hey-day of this industry, 200 000 tons were captured in southeast Brazil alone every year, but landings suddenly plummeted to <20 000 tons by 2001. Despite new fishing regulations introduced following its collapse, it is unclear whether southern Atlantic sardine stocks have shown any sign of recovery.

             With the possible exception of herring and related species that mature early in life and are fished with highly selective equipment, many gadids (e.g. cod, haddock) and other non-clupeids (e.g. flatfishes) have experienced little, if any, recovery in as much as 15 years after 45–99% reductions in reproductive biomass (Hutchings 2000). Worse still, an analysis of 147 populations of 39 wild fish species concluded that historically overexploited species, such as North Sea herring, became more prone to extreme year-on-year variation in numbers, rendering them vulnerable to economic or demographic extinction (Minto et al. 2008). Marine fisheries are an underperforming global asset—yields could be much greater if they were properly managed. The difference between the potential and actual net economic benefits from marine fisheries is in the order of US$50 billion per year—equivalent to over half thevalue of the global seafood trade (World Bank 2008). The cumulative economic loss to the global economy over the last three decades is estimated to be approximately US$2 trillion, and in many countries fishing operations are buoyed up by subsidies, so that the global fishery economy to the point of landing is already in deficit. Commercial fishing activities disproportionately threaten large-bodied marine and freshwater species (Olden et al. 2007). This results in fishermen fishing down the food chain, targeting ever-smaller pelagic fish as they can no longer capture top predatory fish. This is symptomatic of the now widely known process of ‘fishing down marine food webs’. Such sequential

size-graded exploitation systems also take place in multi-species assemblages hunted in tropical forests (Jerozolimski and Peres et al 2003). 

               In the seas, overexploitation threatens the persistence of ecologically significant populations of many large marine vertebrates, including sharks, tunas and sea turtles. Regional scale populations of large sharks worldwide have declined by 90% or more, and rapid declines of >75% of the coastal and oceanic Northwest Atlantic populations of scalloped hammerhead, white, and thresher sharks have occurred in the past 15 years (Baum et al. 2003; Myers and Worm 2003; Myers et al. 2007). Much of this activity is profligate and often driven by the surging global demand for shark fins. For example, in 1997 line fishermen captured 186 000 sharks in southern Brazil alone, of which 83% were killed and discarded in open waters following the removal of the most lucrative body parts (C.M. Vooren, pers. comm.). Of the large-bodied coastal species affected by this trade, several have virtually disappeared from shallow waters (e.g. greynurse sharks, Carcharias taurus). Official figures show that 131 tons of shark fins, corresponding to US $2.4 million, were exported from Brazil to Asia in 2007. Finally, we know rather little about ongoing extinction processes caused by harvesting. For example, from a compilation of 133 local, regional and global extinctions of marine fish populations, Dulvy et al. (2003) uncovered that exploitation was the main cause of extinctions (55% of all populations), but these were only reported after a median 53-year lag following their real-time disappearance. Some 80% of all extinctions were only discovered through historical comparisons; e.g. the near-extinction of large skates on both sides of the Atlantic was only brought to the world’s attention several decades after the declines have occurred.

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Non-timber forest products

              Non-timber forest products (NTFPs) are biological resources other than timber which are extracted from either natural or managed forests (Peters 1994). Examples of exploited plant products include fruits, nuts, oil seeds, latex, resins, gums, medicinal plants, spices, dyes, ornamental plants, and raw materials such as firewood, Desmoncus climbing palms, bamboo and rattan. The socio-economic importance of NTFP harvest to indigenous peoples cannot be underestimated. Many ethnobotanical studies have catalogued the wide variety of useful plants (or plant parts) harvested by different aboriginal groups throughout the tropics. For example, the Waimiri-Atroari Indians of central Amazonia make use of 79% of the tree species occurring in a single 1 ha terra firme forest plot (Milliken et al. 1992), and 1748 of the ~8000 angiosperm species in the Himalayan region spanning eight Asian countries are used medicinally and many more for other purposes (Samant et al. 1998). Exploitation of NTFPs often involves partial or entire removal of individuals from the population, but the extraction method and whether vital parts are removed usually determine the mortality level in the exploited population. Traditional NTFP extractive practices are often hailed as desirable, low-impact economic activities in tropical forests compared to alternative forms of land use involving structural disturbance such as selective logging and shifting agriculture (Peters et al. 1989). As such, NTFP exploitation is usually assumed to be sustainable and a promising compromise between biodiversity conservation and economic development under varying degrees of market integration. The implicit assumption is that traditional methods of NTFP exploitation have little or no impact on forest ecosystems and tend to be sustainable because they have been practiced over many generations. However, virtually any form of NTFP exploitation in tropical forests has an ecological impact.

             The spatial extent and magnitude of this impact dependson the accessibility of the resource stock, the floristic composition of the forest, the nature and intensity of harvesting, and the particular species or plant part under exploitation. Yet few studies have quantitatively assessed the demographic viability of plant populations sourcing NTFPs. One exception are Brazil nuts (Bertholletia excelsa, Lecythidaceae) which comprise the most important wild seed extractive industry supporting millions of Amazonian forest dwellers for either subsistence or income. This wild seed crop is firmly established in export markets, has a history of 200 years of commercial exploitation, and comprises one of the most valuable non-timber extractive industries in tropical forests anywhere. Yet the persistent collection of B. excelsa seeds has severely undermined the patterns of seedling recruitment of Brazil nut trees. This has drastically affected the age structure of many natural populations to the point where persistently overexploited stands have succumbed to a process of senescence and demographic collapse, threatening this cornerston of the Amazonian extractive economy (Peres et al. 2003). A boom in the use of homeopathic remedies sustained by over collecting therapeutic and aromatic plants is threatening at least 150 species of European wild flowers and plants and driving many populations to extinction (Traffic 1998). Commercial exploitation of the Pau-Rosa or rosewood tree (Aniba rosaeodora, Lauraceae), which contains linalol, a key ingredient in luxury perfumes, involves a one-off destructive harvesting technique that almost invariably kills the tree. This species has consequently been extirpated from virtually its entire range in Brazilian Amazonia (Mitja and Lescure 2000). Channel 5 and other perfumes made with Pau-Rosa fragrance gained wide market demand decades ago, but
the number of processing plants in Brazil fell from 103 in 1966 to fewer than 20 in 1986, due to the dwindling resource base. Yet French perfume connoisseurs have been reluctant to accept replacing the natural Pau-Rosa fragrance with
synthetic substitutes, and the last remaining populations of Pau-Rosa remain threatened. 

              The same could be argued for a number of NTFPs for which the harvest by destructive practices involves a lethal injury to whole reproductive individuals. What then is the impact of NTFP extraction on the dynamics of natural populations? How does the impact vary with the life history of plants and animals harvested? Are current extraction rates truly sustainable? These are key questions in terms of the demographic sustainability of different NTFP offtakes, which will ultimately depend on the ability of the resource population to recruit new seedlings either continuously or in sporadic pulses while being subjected to a repeated history of exploitation. Unguarded enthusiasm for the role of NTFP exploitation in rural development partly stems from unrealistic economic studies reporting high market values. For example, Peters et al. (1989) reported that the net-value of fruit and latex extraction in the Peruvian Amazon was US$6330/ ha. This is in sharp contrast with a Mesoamerican study that quantified the local value of foods, construction materials, and medicines extracted from the forest by 32 indigenous Indian households (Godoy et al. 2000). The combined value of consumption and sale of forest goods ranged from US$18 to US$24 ha 1 yr 1, at the lower end of previous estimates (US$49 - US$1 089 ha 1 yr 1). NTFP extraction thus cannot be seen as a panacea for rural development and in many studies the potential value of NTFPs is exaggerated by unrealistic assumptions of high discount rates, unlimited market demands, availability of transportation facilities and absence of product substitution.

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Tropical forest vertebrates

             Humans have been hunting wildlife in tropical forests for over 100 000 years, but the extent of consumption has greatly increased over the last few decades. Tropical forest species are hunted for local consumption or sales in distant markets as food, trophies, medicines and pets. Exploitation of wild meat by forest dwellers has increased due to changes in hunting technology, scarcity of alternative protein, larger numbers of consumers, and greater access infrastructure. Recent estimates of the annual wild meat harvest are 23 500 tons in Sarawak (Bennett 2002), up to 164 692 tons in the Brazilian Amazon (Peres 2000), and up to 3.4 million tons in Central Africa (Fa and Peres 2001).
Hunting rates are already unsustainably high across vast tracts of tropical forests, averaging sixfold the maximum sustainable harvest in Central Africa (Fa et al. 2001). Consumption is both by rural and urban communities, who are often at the end of long supply chains that extend into many remote areas (Milner-Gulland et al. 2003). 

               The rapid acceleration in tropical forest defaunation due to unsustainable hunting initially occurred in Asia (Corlett 2007), is now sweeping through Africa, and is likely to move into the remotest parts of the neotropics (Peres and Lake 2003), reflecting human demographics in different continents. Hunting for either subsistence or commerce can profoundly affect the structure of tropical forest vertebrate assemblages, as revealed by both village- based kill-profiles (Jerozolimski and Peres 2003; Fa et al. 2005) and wildlife surveys in hunted and unhunted forests. This can be seen in the residual game abundance at forest sites subjected to varying degrees of hunting pressure, where overhunting often results in faunal biomass collapses, mainly through declines and local extinctions of large-bodied species (Bodmer 1995; Peres 2000). Peres and Palacios (2007) provide the first systematic estimates of the impact of hunting on the abundances of a comprehensive set of 30 reptile, bird, and mammal species across 101 forest sites scattered widely throughout the Amazon Basin and Guianan Shield. Considering the 12 most harvest sensitive species, mean aggregate population biomass was reduced almost eleven-fold from 979.8 kg/km2 in unhunted sites to only 89.2 kg/km2 in heavily hunted sites. 

              In KilumIjim, Cameroon, most large mammals, including elephants, buffalo, bushbuck, chimpanzees, leopards, and lions, have been lost as a result of hunting (Maisels et al. 2001). In Vietnam, 12 large vertebrate species have become virtually extinct over the last five decades primarily due to hunting (Bennett and Rao 2002). Pangolins and several other forest vertebrate species are facing regionalscale extinction throughout their range across southern Asia [Corlett 2007, TRAFFIC (The Wildlife Trade Monitoring Network) 2008], largely as a result of trade, and over half of all Asian freshwater turtle species are considered Endangered due to over-harvesting (IUCN 2007). In sum, game harvest studies throughout the tropics have shown that most unregulated, commercial hunting for wild meat is unsustainable (Robinson and Bennett 2000; Nasi et al. 2008), and that even subsistence hunting driven by local demand can severely threaten many medium to large-bodied vertebrate populations, with potentially far-reaching consequences to other species. However, persistent harvesting of multi-species prey assemblages can often lead to post-depletion equilibrium conditions in which slow-breeding, vulnerable taxa are eliminated and gradually replaced by fast-breeding robust taxa that are resilient to typical offtakes. For example, hunting in West African forests could now be defined as sustainable from the viewpoint of
urban bushmeat markets in which primarily rodents and small antelopes are currently traded, following a series of historical extinctions of vulnerable prey such as primates and large ungulates (Cowlishaw et al. 2005).

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Overexploitation in tropical forests

             Tropical deforestation is driven primarily by frontier expansion of subsistence agriculture and large development programs involving resettlement, agriculture, and infrastructure. However, animal and plant population declines are typically pre-empted by hunting and logging activity well before the coup de grâce of deforestation is delivered. It is estimated that between 5 and 7 million hectares of tropical forests are logged annually, approximately 68-79% of the area that was completely deforested each year between 1990 and 2005 [FAO (Food and Agriculture Organization of the United Nations) 2007]. Tropical forests account for ~25% of the global industrial wood production worth US$400 billion or ~2% of the global gross domestic product [WCFSD (World Commission on Forests and Sustainable Development) 1998]. Much of this logging activity opens up new frontiers to wildlife and non-timber resource exploitation, and catalyses the transition into a landscape dominated by slash-and burn and large-scale agriculture. Few studies have examined the impacts of selective logging on commercially valuable timber species and comparisons among studies are limited because they often fail to employ comparable methods that are adequately reported. The best case studies come from the most valuable timber species that have already declined in much of their natural ranges.

            For instance, the highly selective, but low intensity logging of broadleaf mahogany (Swietenia macrophylla), the most valuable widely traded Neotropical timber tree, is driven by the extraordinarily high prices in international markets, which makes it lucrative for loggers to open-up even remote wilderness areas at high transportation costs. Mechanized extraction of mahogany and other prime timber species impacts the forest by creating canopy gaps and imparting much collateral damage due to logging roads and skid trails (Grogan et al. 2008). Mahogany and other high-value tropical timber species worldwide share several traits that predispose them to commercial extirpation: excellent pliable wood of exceptional beauty; natural distributions in forests experiencing rapid conversion rates; low-density populations (often <1 tree/ha); and life histories generally characterized as non-pioneer late secondary, with fast growth rates, abiotic seed dispersal, and low-density seedlings requiring canopy disturbance for optimal seedling regeneration in the understory (Swaine and Whitmore 1988; Sodhi et al. 2008). One of the major obstacles to implementing a sustainable forestry sector in tropical countries is the lack of financial incentives for producers to limit offtakes to sustainable levels and invest in regeneration. Economic logic often dictates that trees should be felled whenever their rate of volume

increment drops below the prevailing interest rate (Pearce 1990). Postponing harvest beyond this point would incur an opportunity cost because profits from logging could be invested at a higher rate elsewhere.

            This partly explains why many slow-growing timber species from tropical forests and savannahs are harvested unsustainably (e.g. East African Blackwood (Dalbergia melanoxylon) in the Miombo woodlands of Tanzania; Ball 2004). This is particularly the case where land tenure systems are unstable, and where there are no disincentives against ‘hit-and-run’ operations that mine the resource capital at one site and move on to undepleted areas elsewhere. This is clearly shown in a mahogany study in Bolivia where the smallest trees felled are ~40 cm in diameter, well below the legal minimum size (Gullison 1998). At this size, trees are increasing in volume at about 4% per year, whereas real mahogany price increases have averaged at only 1%, so that a 40-cm mahogany tree increases in value at about 5% annually, slowing down as the tree becomes larger. In contrast, real interest rates in Bolivia and other tropical countries are often >10%, creating a strong economic incentive to liquidate all trees of any value regardless of resource ownership. Tropical deforestation is driven primarily by frontier expansion of subsistence agriculture and large development programs involving resettlement, agriculture, and infrastructure. However, animal and plant population declines are typically pre-empted by hunting and logging activity well before the coup de grâce of deforestation is delivered.

            It is estimated that between 5 and 7 million hectares of tropical forests are logged annually, approximately 68-79% of the area that was completely deforested each year between 1990 and 2005 [FAO (Food and Agriculture Organization of the United Nations) 2007]. Tropical forests account for ~25% of the global industrial wood production worth US$400 billion or ~2% of the global gross domestic product [WCFSD (World Commission on Forests and Sustainable Development) 1998]. Much of this logging activity opens up new frontiers to wildlife and non-timber resource exploitation, and catalyses the transition into a landscape dominated by slash-andburn and large-scale agriculture. Few studies have examined the impacts of selective logging on commercially valuable timber species and comparisons among studies are limited because they often fail to employ comparable methods that are adequately reported. The best case studies come from the most valuable timber species that have already declined in much of their natural ranges. For instance, the highly selective, but low intensity logging of broadleaf mahogany (Swietenia macrophylla), the most valuable widely traded Neotropical timber tree, is driven by the extraordinarily high prices in international markets, which makes it lucrative for loggers to open-up even remote wilderness areas at high transportation costs. Mechanized extraction of mahogany and other prime timber species impacts the forest by creating canopy gaps and imparting much collateral damage due to logging roads and skid trails (Grogan et al.

2008). Mahogany and other high-value tropical timber species worldwide share several traits that predispose them to commercial extirpation: excellent pliable wood of exceptional beauty; natural distributions in forests experiencing rapid conversion rates; low-density populations

(often <1 tree/ha); and life histories generally characterized as non-pioneer late secondary, with fast growth rates, abiotic seed dispersal, and low-density seedlings requiring canopy disturbance for optimal seedling regeneration in the understory (Swaine and Whitmore 1988; Sodhi et al. 2008). One of the major obstacles to implementing a sustainable forestry sector in tropical countries is the lack of financial incentives for producers to limit offtakes to sustainable levels and invest in regeneration.

             Economic logic often dictates that trees should be felled whenever their rate of volume increment drops below the prevailing interest rate (Pearce 1990). Postponing harvest beyond this point would incur an opportunity cost because profits from logging could be invested at a higher rate elsewhere. This partly explains why many slow-growing timber species from tropical forests and savannahs are harvested unsustainably (e.g. East African Blackwood (Dalbergia melanoxylon) in the Miombo woodlands of Tanzania; Ball 2004). This is particularly the case where land tenure systems are unstable, and where there are no disincentives against ‘hit-and-run’ operations that mine the resource capital at one site and move on to undepleted areas elsewhere. This is clearly shown in a mahogany study in Bolivia where the smallest trees felled are ~40 cm in diameter, well below the legal minimum size (Gullison 1998). At this size, trees are increasing in volume at about 4% per year, whereas real mahogany price increases have averaged at only 1%, so that a 40-cm mahogany tree increases in value at about 5% annually, slowing down as the tree becomes larger. In contrast, real interest rates in Bolivia and other tropical countries are often >10%, creating a strong economic incentive to liquidate all trees of any value regardless of resource ownership.

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A brief history of exploitation

                Our rapacious appetite for both renewable and non - renewable resources has grown exponentially from our humble beginnings—when early humans exerted an ecological footprint no larger than that of other large omnivorous mammals— to currently one of the main driving forces in reorganizing the structure of many ecosystems. Humans have subsisted on wild plants and animals since the earliest primordial times, and most contemporary aboriginal societies remain primarily extractive in their daily quest for food, medicines, fiber and other biotic sources of raw materials to produce a wide range of utilitarian and ornamental artifacts. Modern hunter-gatherers and semi-subsistence farmers in tropical ecosystems, at varying stages of transition to an agricultural economy, still exploit a large number of plant and animal populations. By definition, exploited species extant today have been able to co-exist with some background level of exploitation. However, paleontological evidence suggests that prehistoric peoples have been driving prey populations to extinction long before the emergence of recorded history. 

                 The late Paleolithic archaeology of big-game hunters in several parts of the world shows the sequential collapse of their majestic lifestyle. Flint spearheads manufactured by western European Cro - Magnons became gradually smaller as they shifted down to ever smaller kills, ranging in size from mammoths to rabbits (Martin 1984). Human colonization into previously people-free islands and continents has often coincided with a rapid wave of extinction events resulting from the sudden arrival of novel consumers. Mass extinction events of large-bodied vertebrates in Europe, parts of Asia, North and South America, Madagascar, and several archipelagos have all been attributed to post-Pleistocene human overkill (Martin and Wright 1967; Steadman 1995; McKinney 1997; Alroy 2001). These are relatively well corroborated in the (sub)fossil record but many more obscure target species extirpated by archaic hunters will remain undetected. In more recent times, exploitation-induced extinction events have also been common as European settlers wielding superior technology greatly expanded their territorial frontiers and introduced market and sport hunting. One example is the decimation of the vast North American buffalo (bison; Bison bison) herds. In the 1850s, tens of millions of these ungulates roamed the Great Plains in herds exceeding those ever known for any other mega herbivore, but by the century’s close, the bison was all but extinct. 

                 Another example is the extirpation of mono dominant stands of Pau-Brasil legume trees (Caesalpinia echinata, Leguminosae- Mimosoidae) from eastern Brazil, a source of red dye and hardwood that gave Brazil its name. These were once extremely abundant and formed dense clusters along 3000 km of coastal Atlantic forest. This species sustained the first trade cycle between the new Portuguese colony and European markets and was relentlessly exploited from 1500 to 1875 when it finally became economically extinct (Dean 1996). Today, specimens of Pau-Brasil trees are largely confined to herbaria, arboreta and a few private collections. The aftershock of modern human arrival is still being felt in many previously inaccessible tropical forest frontiers, such as those in parts of Amazonia, where greater numbers of hunters wielding fire arms are emptying vast areas of its harvest sensitive mega fauna (Peres and Lake 2003). In many modern societies, the exploitative value of wildlife populations for either subsistence or commercial purposes has been gradually replaced by recreational values including both consumptive and non-consumptive uses. In 1990, over 20 million hunters in the United States spent over half a billion days afield in pursuit of wild game, and hunting licenses finance vast conservation areas in North America. 

                In 2006, ~ 87.5 million US residents spent ~US$122.3 billion in wildlife-related recreational activities, including ~US$76.6 billion spent on fishing and/or hunting by 33.9 million people (US Census Bureau 2006). Some 10% of this total is spent hunting white-tailed deer alone (Conover 1997). Consumptive uses of wildlife habitat are therefore instrumental in either financing or justifying much of the conservation acreage available in the 21st century from game reserves in Africa, Australia and North America to extractive reserves in Amazonia, to the reindeer rangelands of Scandinavia and the saiga steppes of Mongolia. Strong cultural or social factors regulating resource choice often affect which species are taken. For example, while people prefer to hunt large bodied mammals in tropical forests, feeding taboos and restrictions can switch “on or off” depending on levels of game depletion (Ross 1978) as predicted by foraging theory. This is consistent with the process of de-tabooing speciesthat were once tabooed, as the case of brocket deer among the Siona-Secoya (Hames and Vickers 1982). However, several studies suggest that cultural factors breakdown and play a lesser role when large-bodied game species become scarce, thereby forcing discriminate harvesters to become less selective (Jerozolimski and Peres 2003).

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Overexploitation

                In an increasingly human-dominated world, where most of us seem oblivious to the liquidation of Earth’s natural resource capital, exploitation of biological populations has become one of the most important threats to the persistence of global biodiversity. Many regional economies, if not entire civilizations, have been built on free-for-all extractive industries, and history is littered with examples of boom - and - bust economic cycles following the emergence, escalation and rapid collapse of unsustainable industries fuelled by raw renewable resources. The economies of many modern nation -states still depend heavily on primary extractive industries, such as fisheries and logging, and this includes countries spanning nearly the entire spectrum of per capita Gross National Product (GNP), such as Iceland and Cameroon. Human exploitation of biological commodities involves resource extraction from the land, freshwater bodies or oceans, so that wild animals, plants or their products are used for a wide variety of purposes ranging from food to fuel, shelter, fiber, construction materials, household and garden
items, pets, medicines, and cosmetics. 

               Overexploitation occurs when the harvest rate of any given population exceeds its natural replacement rate, either through reproduction alone in closed populations or through both reproduction and immigration from other populations. Many species are relatively insensitive to harvesting, remaining abundant under relatively high rates of offtake, whereas others can be driven to local extinction by even the lightest levels of offtake. Fishing, hunting, grazing, and logging are classic consumer-resource interactions and in natural systems such interactions tend to come into equilibrium with the intrinsic productivity of a given habitat and the rates at which resources are harvested. Furthermore, efficiency of exploitation by consumers and the highly variable intrinsic resilience to exploitation by resource populations may have often evolved over long periods. Central to these differences are species traits such as the population density (or stock size), the per capita growth rate of the population, spatial diffusion from other less harvested populations, and the direction and degree to which this growth responds to harvesting through either positive or negative density dependence. 

               For example, many long-lived and slow -growing organisms are particularly vulnerable to the additive mortality resulting from even the lightest offtake, especially if these traits are combined with low dispersal rates that can inhibit population diffusion from adjacent unharvested source areas, should these be available. These species are often threatened by over hunting in many terrestrial ecosystems, unsustainable logging in tropical forest regions, cactus “rustling” in deserts, overfishing in marine and freshwater ecosystems, or many other forms of unsustainable extraction. For example, overhunting is the most serious threat to large vertebrates in tropical forests (Cunningham et al. 2009), and overexploitation, accidental mortality and persecution caused by humans threatens approximately one-fifth (19%) of all tropical forest vertebrate species for which the cause of decline has been documented [IUCN (International Union for Conservation of Nature) 2007]. Overexploitation is the most important cause of freshwater turtle extinctions (IUCN 2007) and the third-most important for freshwater fish extinctions, behind the effects of habitat loss and introduced species (Harrison and Stiassny 1999). 

               Thus, while population declines driven by habitatloss and degradation quite rightly receive a great deal of attention from conservation biologists (MEA 2006), we must also contend with the specter of the ‘empty’ or ‘half-empty’ forests, savannahs, wetlands, rivers, and seas, even if the physical habitat structure of a given ecosystem remains otherwise unaltered by other anthropogenic processes that degrade habitat quality. Overexploitation also threatens frogs: with Indonesia the main exporter of frog legs for markets in France and the US (Warkentin et al. 2009). Up to one billion wild frogs are estimated to be harvested every year for human consumption (Warkentin et al. 2009). I begin this chapter with a consideration of why people exploit natural populations, including the historical impacts of exploitation on wild plants and animals. This is followed by a review of effects of exploitation in terrestrial and aquatic biomes. Throughout the chapter, I focus on tropical forests and marine ecosystems because many plant and animal species in these realms have succumbed to some of the most severe and least understood overexploitation-related threats to population viability of contemporary times. I then explore impacts of exploitation on both target and non-target species, as well as cascading effects on the ecosystem. This leads to a reflection at the end of this chapter of resource management considerations in the real-world, and the clashes of culture between those concerned with either the theoretical underpinnings or effective policy solutions addressing the predicament of species imperiled by overexploitation.

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Processes that affect community structure

              Interactions between species, such as predation, competition, parasitism, and an array of mutualisms, have a profound influence on the structure of communities. The loss of a species or a change in its abundance, particularly for species that interact with many others, can have a marked effect on ecological processes throughout fragmented
landscapes. Changes to predator-prey relationships, for example, have been revealed by studies of the level of predationon birds’ nests in fragmented landscapes (Wilcove 1985). An increase in the amount of forest edge, a direct consequence of fragmentation, increases the opportunity for generalist predators associated with edges or modified land-uses to prey on birds that nest in forest fragments. In Sweden, elevated levels of nest predation (on artificial eggs in experimental nests) were recorded in agricultural land and at forest edges compared with the interior of forests (Andrén and Angelstam 1988). Approximately 45% of nests at the forest edge were preyed upon compared with less than 10% at distances >200 m into the forest. At the landscape scale, nest predation occurred at a greater rate in agricultural and fragmented forest landscapes than in largely forested landscapes (Andrén 1992). The relative abundance of different corvid species, the main nest predators, varied in relation to landscape composition. 

               The hooded crow (Corvus corone cornix) occurred in greatest abundance in heavily cleared landscapes and was primarily responsible for the greater predation pressure recorded at forest edges. Many mutualisms involve interactions between plants and animals, such as occurs in the pollination of flowering plants by invertebrates, birds or mammals. A change in the occurrence or abundance of animal vectors, as a consequence of fragmentation, can disrupt this process. For many plant species, habitat fragmentation has a negative effect on reproductive success, measured in terms of seed or fruit production, although the relative impact varies among species (Aguilar et al. 2006). Plants that are self-incompatible (i.e. that depend on pollen transfer from another plant) are more susceptible to reduced reproductive success than are self-compatible species. This difference is consistent with an expectation that pollination by animals will be less effective in small and isolated fragments. However, pollinators are a diverse group and they respond to fragmentation in a variety of ways (Hobbs and Yates 2003). Changes in ecological processes in fragments and through out fragmented landscapes are complex and poorly understood. 

                Disrupted interactions between species may have flow-on effects to many other species at other trophic levels. However, the kinds of changes to species interactions and ecological processes vary between ecosystems and regions because they depend on the particular sets of species that occur. In parts of North America, nest parasitism by the brown-headed cowbird (Molothrus ater) has a marked effect on bird communities in fragments (Brittingham and Temple 1983); while in eastern Australia, bird communities in small fragments may be greatly affected by aggressive competition from the noisy miner (Manorina melanocephala) (Grey et al. 1997). Both of these examples are idiosyncratic to their region. They illustrate the difficulty of generalizing the effects of habitat fragmentation, and highlight the importance of understanding the consequences of landscape change in relation to the environment, context and biota of a particular region.

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Patterns of community structure in fragmented landscapes

             For many taxa birds, butterflies, rodents, reptiles, vascular plants, and more species richness in habitat fragments is positively correlated with fragment size. This is widely known as the species-area relationship. Thus, when habitats are fragmented into smaller pieces, species are lost; and the likely extent of this loss can be predicted from the species-area relationship. Further, species richness in a fragment typically is less than in an area of similar size within continuous habitat, evidence that the fragmentation process itself is a cause of local extinction. However, the species-area relationship does not reveal which particular species will be lost. Three explanations given for the species-area relationship (Connor and McCoy 1979) are that small areas: (i) have a lower diversity of habitats;
(ii) support smaller population sizes and therefore fewer  species can maintain viable populations;and (iii) represent a smaller sample of the original habitat and so by chance are likely to have fewer species than a larger sample. While it is difficult to distinguish between these mechanisms, the message is clear: when habitats are fragmented into smaller pieces, species are lost.

               Factors other than area, such as the spatial and temporal isolation of fragments, land management or habitat quality may also be significant predictors of the richness of communities in fragments. In Tanzania, for example, the number of forest under story bird species in forest fragments (from 0.1 to 30 ha in size) was strongly related to fragment size, as predicted by the species-area relationship (Newmark 1991). After taking fragment size into account, further variation in species richness was explained by the isolation distance of each fragment from a large source area of forest. Species show differential vulnerability to fragmentation. Frequently, species with more specialized ecological requirements are those lost from communities in fragments. In several tropical regions, birds that follow trails of army ants and feed on insects flushed by the ants include specialized ant-following species and others that forage opportunistically in this way. In rainforest in Kenya, comparisons of flocks of ant-following birds between a main forest and forest fragments revealed marked differences (Peters et al. 2008). The species richness and number of individuals in ant-following flocks were lower in fragments, and the composition of flocks more variable in small fragments and degraded forest, than in the main forest. This was a consequence of a strong decline in abundance of five species of specialized ant-followers in fragments, where as the many opportunistic followers (51 species) were little affected by fragmentation (Peters et al. 2008). The way in which fragments are managed is a particularly important influence on the composition of plant communities. In eastern Australia, for example, grassy woodlands dominated by white box (Eucalyptus albens) formerly covered several million hectares, but now occur as small fragments surrounded by cropland or agricultural pastures. 

               The species richness of native understory plants increases with fragment size, as expected, but tree clearing and grazing by domestic stock are also strong influences (Prober and Thiele 1995). The history of stock grazing has the strongest influence on the floristic composition in woodland fragments: grazed sites have a greater invasion by weeds and a more depauperate native flora. The composition of animal communities in fragments commonly shows systematic changes in relation to fragment size. Species-poor communities in small fragments usually support a subset of the species present in larger, richer fragments. That is, there is a relatively predictable change in composition with species “dropping out” in an ordered sequence in successively smaller fragments (Patterson and Atmar 1986). Typically, rare and less common species occur in larger fragments, whereas those present in smaller fragments are mainly widespread and common. This kind of “nested subset” patternhas been widely observed: for example, in butterfly communities in fragments of lowland rainforest in Borneo (Benedick et al. 2006). At the landscape level, species richness has frequently been correlated with heterogeneity in the landscape. This relationship is particularly relevant in regions, such as Europe, where human land-use has contributed to cultural habitats that complement fragmented natural or semi-natural habitats. In the Madrid region of Spain, the overall richness of assemblages of birds, amphibians, reptiles and butterflies in 100 km2 landscapes is strongly correlated with the number of different land-uses in the landscape (Atauri and De Lucio 2001). However, where the focus is on the community associated with a particular habitat type (e.g. rainforest butterflies) rather than the entire assemblage of that taxon, the strongest influence on richness is the total amount of habitat in the landscape. For example, the richness of wood land dependent birds in fragmented landscapes in southern Australia was most strongly influenced by the total extent of wooded cover in each 100 km2 landscape, with a marked threshold around 10% cover below which species richness declined rapidly (Radford et al. 2005).

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Metapopulations and the conservation of subdivided populations

                              Small populations are vulnerable to local extinction, but a species has a greater likelihood of persistence where there are a number of local populations interconnected by occasional movements of individuals among them. Such a set of subdivided populations is often termed a “meta population” (Hanski 1999). Two main kinds of meta population have been described. A mainland-island model is where a large mainland population (such as a conservation reserve) provides a source of emigrants that disperse to nearby small populations. 

               The mainland population has a low likelihood of extinction, where as the small populations become extinct relatively frequently. Emigration from the mainland supplements the small populations, introduces new genetic material and allows recolonization should local extinction occur. A second kind of meta population is where the set of interacting populations are relatively similar in size and all have a likelihood of experiencing extinction. Although colonization and extinction may occur regularly, the overall population persists through time. The silver-spotted skipper (Hesperia comma), a rare butterfly in the UK, appears to function as a meta population (Hill et al. 1996). In 1982, butterflies occupied 48 of 69 patches of suitable grassland on the North Downs, Surrey. Over the next 9 years, 12 patches were colonized and seven populations went extinct. Those more susceptible to extinction were small isolated populations, whereas the patches more likely to be colonized were relatively large and close to other large occupied patches. 

              The conservation management of patchily distributed species is likely to be more effective by taking a metapopulation approach than by focusing on individual populations. However, “real world” populations differ from theoretical models. Factors such as the quality of habitat patches and the nature of the land mosaic through which movements occur are seldom considered in theoretical models, which emphasize spatial attributes (patch area, isolation). For example, in a meta population of the Bay checker spot butterfly (Euphydryas editha bayensis) in California, USA, populations in topographically heterogeneous fragments were less likely to go extinct than those that were in topographically uniform ones. The heterogeneity provided some areas of suitable topo climate each year over a wide range of local climates (Ehrlich and Hanski 2004). There also is much variation in the structure of subdivided populations depending on the frequency of movements between them. At one end of a gradient is a dysfunctional metapopulation where little or no movement occurs; while at the other extreme, movements are so frequent that it is essentially a single patchy population.

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Processes that affect species in fragmented landscapes

                The size of any population is determined by the balance between four parameters: births, deaths, immigration, and emigration. Population size is increased by births and immigration of individuals, while deaths and emigration of individuals reduce population size. In fragmented landscapes, these population parameters are influenced by several categories of processes.    

Deterministic processes

                Many factors that affect populations in fragmented landscapes are relatively predictable in their effect. These factors are not necessarily a direct consequence of habitat fragmentation, but arise from land uses typically associated with subdivision. Populations may decline due to deaths of individuals from the use of pesticides, insecticides or other chemicals; hunting by humans; harvesting and removal of plants; and construction of roads with ensuing road kills of animals. For example, in Amazonian forests, subsistence hunting by people compounds the effects of forest fragmentation for large vertebrates such as the lowland tapir (Tapir terrestris) and white-lipped peccary (Tayassu pecari), and contributes to their local extinction (Peres 2001). Commonly, populations are also affected by factors such as logging, grazing by domestic stock, or altered disturbance regimes that modify the quality of habitats and affect population growth. For example, in Kibale National Park, an isolated forest in Uganda, logging has resulted in long-term reduction in the density of groups of the blue monkey (Cercopithecus mitza) in heavily logged areas: in contrast, populations of black and white colobus (Colobus guereza) are higher in regrowth forests than in unlogged forest (Chapman et al. 2000). Deterministic processes are particularly important influences on the status of plant species in fragments (Hobbs and Yates 2003).

Isolation

                 Isolation of populations is a fundamental consequence of habitat fragmentation: it affects local populations by restricting immigration and emigration. Isolation is influenced not only by the distance between habitats but also by the effects of human land-use on the ability of organisms to move (or for seeds and spores to be dispersed) through the landscape. Highways, railway lines, and water channels impose barriers to movement, while extensive croplands or urban development create hostile environments for many organisms to move through. Species differ in sensitivity to isolation depending on their type of movement, scale of movement, whether they are nocturnal or diurnal, and their response to landscape change. Populations of one species may be highly isolated, while in the same landscape individuals of another species can move freely.Isolation affects several types of movements, including: (i) regular movements of individuals between parts of the landscape to obtain different requirements (food, shelter, breeding sites); (ii) seasonal or migratory movements of species at regional, continental or inter-continental scales; and (iii) dispersal movements (immigration, emigration) between fragments, which may supplement population numbers, increase the exchange of genes, or assist recolonization if a local population has disappeared. In Western Australia, dispersal movements of the blue breasted fairy-wren (Malurus pulcherrimus) are affected by the isolation of fragments (Brooker and Brooker 2002). There is greater mortality of individuals during dispersal in poorly connected areas than in well-connected areas, with this difference in survival during dispersal being a key factor determining the persistence of the species in local areas. For many organisms, detrimental effects of isolation are reduced, at least in part, by habitat components that enhance connectivity in the landscape (Saunders and Hobbs 1991; Bennett 1999). These include continuous “corridors” or “stepping stones” of habitat that assist movements (Haddad et al. 2003), or human land-uses (such as coffee-plantations, scattered trees in pasture) that may be relatively benign environments for many species (Daily et al. 2003). In tropical regions, one of the strongest influences on the persistence of species in forest fragments is their ability to live in, or move through, modified “countryside” habitats (Gascon et al. 1999; Sekercioglu et al. 2002).

Stochastic processes

                When populations become small and isolated, they become vulnerable to a number of stochastic (or chance) processes that may pose little threat to larger populations. Stochastic processes include
the following.

• Stochastic variation in demographic parameters such as birth rate, death rate and the sex ratio of offspring. 

• Loss of genetic variation, which may occur due to inbreeding, genetic drift, or a founder effect from a small initial population size. A decline in genetic diversity may make a population more vulnerable to recessive lethal alleles or to changing environmental conditions. 

• Fluctuations in the environment, such as variation in rainfall and food sources, which affect birth and death rates in populations. 

•  Small isolated populations are particularly vulnerable to catastrophic events such as flood, fire, drought or hurricanes. A wildfire, for example, may eliminate a small local population where as in extensive habitats some individuals survive and provide a source for recolonization.

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Patterns of species occurrence in fragmented landscapes

                Many studies have described the occurrence of species in fragments of different sizes, shapes, composition, land-use and context in the landscape. For species that primarily depend on fragmented habitat, particularly animals, fragment size is a key influence on the likelihood of occurrence. As fragment size decreases, the frequency of occurrence declines and the species may be absent from many small fragments. Such absences may be because the fragment is smaller than the minimum area required for a single individual or breeding unit, or for a self-sustaining population. Some species persist in fragmented landscapes by incorporating multiple fragments in their territory or daily foraging movements. In England, the tawny owl (Strix aluco) occupies territories of about 26 ha (hectares) in large deciduous woods, but individuals also persist in highly fragmented areas by including several small woods in their territory (Redpath 1995). 

                There is a cost, however: individuals using multiple woods have lower breeding success and there is a higher turnover of territories between years. Species that require different kinds of habitats to meet regular needs (e.g. for foraging and breeding) can be greatly disadvantaged if these habitats become isolated. Individuals may then experience difficulty in moving between different parts of the landscape to obtain their required resources. Amphibians that move between a breeding pond and other habitat, such as overwintering sites in forest, are an example. Other attributes (in addition to fragment size) that influence the occurrence of species include the type and quality of habitat, fragment shape, land use adjacent to the fragment, and the extent to which the wider landscape isolates populations. In the Iberian region of Spain, for example, the relative abundance of the Eurasian badger (Meles meles) in large forest fragments is significantly influenced by habitat quality and forest cover in the wider landscape (Virgos 2001). 

                 In areas with less than 20% forest cover, badger abundance in forests was most influenced by isolation (i.e. distance to a potential source area >10 000 ha), whereas in areas with 20–50% cover, badgers were most influenced by the quality of habitat in the forest fragments. A key issue for conservation is the relative importance of habitat loss versus habitat fragmentation (Fahrig 2003). That is, what is the relative importance of how much habitat remains in the landscape versus how fragmented it is? Studies of forest birds in landscapes in Canada and Australia suggest that habitat loss and habitat fragmentation are both significant influences, although habitat loss generally is a stronger influence for a greater proportion of species (Trczinski et al. 1999; Radford and Bennett 2007). Importantly, species respond to landscape pattern in different ways. In southern Australia, the main influence for the eastern yellow robin (Eopsaltria australis) was the total amount of wooded cover in the landscape; for the grey shrike-thrush (Colluricincla harmonica) it was wooded cover together with its configuration (favoring aggregated habitat); while for the musk lorikeet (Glossopsitta concinna) the influential factor was not wooded cover, but the configuration of habitat and diversity of vegetation types (Radford and Bennett
2007).

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Changes to ecosystem processes

              Removal of large tracts of native vegetation changes physical processes, such as those relating to solar radiation and the fluxes of wind and water (Saunders et al. 1991). The greatest impact on fragments occurs at their boundaries; small remnants and those with complex shapes experience the strongest “edge effects”. For example, the micro climate at a forest edge adjacent to cleared land differs from that of the forest interior in attributes such as incident light, humidity, ground and air temperature, and wind speed. In turn, these physical changes affect biological processes such as litter decomposition and nutrient cycling, and the structure and composition of vegetation. Changes to biophysical processes from land use in the surrounding environment, such as the use of fertilizers on farmland, alterations to drainage patterns and water flows, and the presence of exotic plants and animals, also have spill-over effects in fragments. 

               Many native vegetation communities are resistant to invasion by exotic plant species unless they are disturbed. Grazing by domestic stock and altered nutrient levels can facilitate the invasion of exotic species of plants, which markedly alters the vegetation in fragments (Hobbs and Yates 2003) and habitats for animals. The intensity of edge effects in fragments and the distance over which they act varies between processes and between ecosystems. In tropical forests in the Brazilian Amazon, for example, changes in soil moisture content, vapor pressure deficit, and the number of tree fall gaps extend about 50 m into the forest, whereas the invasion of disturbance-adapted butterflies and beetles and elevated tree mortality extend 200 m or more from the forest edge (Laurance 2008). 

               In most situations, changes at edges are generally detrimental to conservation values because they modify formerly intact habitats. However, in some circumstances edges are deliberately managed to achieve specific outcomes. Manipulation of edges is used to enhance the abundance of game species such as deer, pheasants and grouse. In England, open linear “rides” in woods may be actively managed to increase incident light and early succession habitat for butterflies and other wildlife (Ferris-Kaan 1995). Changes to biophysical processes frequently have profound effects for entire landscapes. 

              In highly fragmented landscapes in which most fragments are small or have linear shapes, there may be little interior habitat that is buffered from edge effects. Changes that occur to individualfragments accumulate across the landscape. Changes to biophysical processes such as hydrological regimes can also affect entire landscapes. In the Western Australian wheat belt, massive loss of native vegetation has resulted in a rise in the level of groundwater, bringing stored salt (NaCl) to the surface where it accumulates and reduces agricultural productivity and transforms native vegetation (Hobbs 1993).

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