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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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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Conceptual approaches

               The theory of island bio geography (MacArthur and Wilson 1967) had a seminal influence in stimulating ecological and conservation interest in fragmented landscapes. This simple, elegant model highlighted the relationship between the number of species on an island and the island’s area and isolation. It predicted that species richness on an island represents a dynamic balance between the rate of colonization of new species to the island and the rate of extinction of species already present. It was quickly perceived that habitat isolates, such as forest fragments, could also be considered as “islands” in a “sea” of developed land and that this theory provided a quantitative approach for studying their biota. This stimulated many studies in which species richness in fragments was related to the area and isolation of the fragment, the primary factors in island bio geographic theory. The development of landscape ecology contributed new ways of thinking about habitat fragments and landscape change. The concept of patches and connecting corridors set within a matrix (i.e. the background ecosystem or land-use type) became an influential paradigm (Forman and Godron 1986). It recognized the importance of the spatial context of fragments. 

               The environment surrounding fragments is greatly modified during landscape changes associated with fragmentation. Thus, in contrast to islands, fragments and their biota are strongly influenced by physical and biological processes in the wider landscape, and the isolation of fragments depends not only on their distance from a similar habitat but also on their position in the landscape, the types of surrounding land-uses and how they influence the movements of organisms (Saunders et al. 1991; Ricketts 2001). The influence of physical processes and disturbance regimes on fragments means that following habitat destruction and fragmentation, habitat modification also occurs. Mcintyre and Hobbs (1999) incorporated this complexity into a conceptual model by outlining four stages along a trajectory of landscape change. 

 These were:

(i) Intact landscapes, in which most original vegetation remains with little or no modification.

(ii) Variegated landscapes, dominated by the original vegetation, but with marked gradients of habitat modification. 
(iii) Fragmented landscapes, in which fragments are a minor component in a landscape dominated by other land uses.
(iv) Relict landscapes with little (<10%) cover of original vegetation, set within highly modified surroundings. 

            This framework emphasizes the dynamics of landscape change. Different stages along the trajectory pose different kinds of challenges for conservation management. Many species are not confined solely to fragments, but also occur in other land uses in modified landscapes. In Nicaragua, for example, riparian forests, secondary forests, forest fallows, live fences, and pastures with dispersed trees each support diverse assemblages of birds, bats, dung beetles and butterflies (Harvey et al. 2006). To these species, the landscape represents a mosaic of land uses of differing quality, rather than a contrast between “habitat” and “non-habitat”. Recognizing landscapes as mosaics emphasizes the need to appreciate all types of elements in
the landscape. 

              This perspective is particularly relevant in regions where cultural habitats, derived from centuries of human land-use, have important conservation values. Different species have different ecological attributes, such as their scale of movement, life history stages, longevity, and what constitutes habitat. These each influence how a species “perceives” a landscape, as well as its ability to survive in a modified landscape. Consequently, the same landscape may be perceived by different taxa as having a different structure and different suitability, and quite differently from the way that humans describe the landscape. A “species-centered” view of a landscape can be obtained by mapping contours of habitat suitability for any given species (Fischer et al. 2004).

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Tropical and subtropical forests

               A second way to assess habitat loss is by contrastingmajor  biomes or ecosystem types. Today, tropical rainforests (also termed tropical moist and humid forests) are receiving the greatest attention, because they are being destroyed so rapidly and because they are the most biologically diverse of all terrestrial biomes. Of the roughly 16 million km2 of tropical rainforest that originally existed worldwide, less than 9 million km2 remains today (Whitmore 1997; MEA 2005). The current rate of rainforest loss is debated, with different estimates ranging from around 60 000 km2 (Achard et al. 2002) to 130 000 km2 per year (FAO 2000). Regardless of which estimate one adheres to, rates of rainforest loss are alarmingly high. Rates of rainforest destruction vary considerably among geographic regions. Of the world’s three major tropical regions, Southeast Asian forests are disappearing most rapidly in relative terms , while the African and New World tropics have somewhat lower rates of percent- annual forest loss (Sodhi et al. 2004). Such averages, however, disguise important smaller scale variation. 

               In the New World tropics, for example, the Caribbean, Meso American, and Andean regions are all suffering severe rainforest loss, but the relative deforestation rate for the region as a whole is buffered by the vastness of the Amazon. Likewise, in tropical Africa, forest loss is severe in West Africa, montane areas of East Africa, and Madagascar, but substantial forest still survives in the Congo Basin (Laurance 1999). Other tropical and subtropical biomes have suffered even more heavily than rainforests. Tropical dry forests (also known as monsoonal or deciduous forests) have been severely reduced, in part because they are easier to clear and burn than rainforests. For instance, along Central America’s Pacific coast, much less than 1% of the original dry forest survives. Losses of dry forest have been nearly as severe in Madagascar and parts of Southeast Asia (Laurance 1999; Mayaux et al. 2005).
 

                Mangrove forests, salt-tolerant ecosystems that grow in tropical and subtropical intertidal zones, have also been seriously reduced. Based on countries for which data exist, more than a third of all mangroves were lost in the last few decades of the 20th century (MEA 2005). From 1990 to 2000, over 1% of all mangrove forests were lost annually, with rates of loss especially high in Southeast Asia (Mayaux et al. 2005). Such losses are alarming given the high primary productivity of mangroves, their key role as spawning and rearing areas for economically important fish and shrimp species, and their importance for sheltering coastal areas from destructive storms and tsunamis (Danielsen et al. 2005).

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Geography of habitat loss

               Some regions of the Earth are far more affected by habitat destruction than others. Among the most imperiled are the so-called “biodiversity hotspots”, which contain high species diversity, many locally endemic species (those whose entire geographic range is confined to a small area), and which have lost at least 70% of their native vegetation (Myers et al. 2000). Many hotspots are in the tropics. The Atlantic forests of Brazil and rainforests of West Africa, both of which have been severely reduced and degraded, are examples of biodiversity hotspots. Despite encompassing just a small fraction (<2%) of the Earth’s land surface, hotspots may sustain over half of the world’s terrestrial species (Myers et al. 2000). Many islands have also suffered heavy habitat loss. For instance, most of the original natural habitat has already been lost in Japan, New Zealand, Madagascar, the Philippines, and Java (WRI 2003). 

               Other islands, such as Borneo, Sumatra, and New Guinea, still retain some original habitat but are losing it at alarming rates (Curran et al. 2004; MacKinnon 2006). Most areas of high human population density have suffered heavy habitat destruction. Such areas include much of Europe, eastern North America, South and Southeast Asia, the Middle East, West Africa, Central America, and the Caribbean region, among others. Most of the biodiversity hotspots occur in areas with high human density and many still have rapid population growth (Cincotta et al. 2000). 

                Human populations are often densest in coastal areas, many of which have experienced considerable losses of both terrestrial habitats and nearby coral reefs. Among others, coastal zones in Asia, northern South America, the Caribbean, Europe, and eastern North America have all suffered severe habitat loss (MEA 2005). Finally, habitat destruction can occur swiftly in areas with limited human densities but rapidly expanding agriculture. Large expanses of the Amazon, for example, are currently being cleared for large-scale cattle ranching and industrial soy farming, despite having low population densities (Laurance et al. 2001). Likewise, in some relatively sparsely populated areas of Southeast Asia, such as Borneo, Sumatra, and New Guinea, forests are being rapidly felled to establish oilpalm
or rubber plantations (MacKinnon 2006; Laurance 2007; Koh and Wilcove 2008). Older agricultural frontiers, such as those in Europe, eastern China, the Indian Subcontinent, and eastern and midwestern North America, often have very little native vegetation remaining.

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Valuing Ecosystem Services

              Ecosystems and their constituent species provide an endless stream of products, functions, and services that keep our world running and make our existence possible. To many, even the thought of putting a price tag on services like photosynthesis, purification of water, and pollination of food crops may seem like hubris, as these are truly priceless services without which not only humans, but most of life would perish.A distinguished economist put it best in response to a seminar at the USA Federal Trade Commission, where the speaker downplayed the impact of global warming by saying agriculture and forestry “accounted for only three percent of the US gross national product”. The economist’s response was: “What does this genius think we’re going to eat?” (Naylor and Ehrlich 1997). Nevertheless, in our financially-driven world, we need to quantify the trade - offs involved in land use scenarios that maximize biodiversity conservation and ecosystem services versus scenarios that maximize profit from a single commodity. Without such assessments, special interests representing single objectives dominate the debate and sideline the integration of ecosystem services into the decision-making process (Nelson et al. 2009). 

               Valuing ecosystem services is not an end in itself, but is the first step towards integrating these services into public decision-making and ensuring the continuity of ecosystems that provide the services (Goulder and Kennedy 1997; National Research Council 2005; Daily et al. 2009). Historically, ecosystem services have been mostly thought of as free public goods, an approach which has too frequently led to the “tragedy of the commons” where vital ecosystem goods like clean water have been degraded and consumed to extinction (Daily 1997). Too often, ecosystem services have been valued, if at all, based on “marginal utility” (Brauman and Daily 2008). When the service (like cleanwater) is abundant, themarginal utility of one additional unit can be as low as zero. However, as the service becomes more scarce, the marginal utility of each additional unit becomes increasingly valuable (Goulder and Kennedy 1997). Using the marginal value for a service when it is abundant drastically underestimates the value of the service as it becomes scarcer. As Benjamin Franklin wryly observed, “When the well’s dry,we know the worth of water.” As the societal importance of ecosystem services becomes increasingly appreciated, there has been a growing realization that successful application of this concept requires a skilful combination of biological, physical, and social sciences, as well as the creation of new programs and institutions. 

              The scientific community needs to help develop the necessary quantitative tools to calculate the value of ecosystem services and to present them to the decision makers (Daily et al. 2009). A promising example is the InVEST (Integrated Valuation of Ecosystem Services and Tradeoffs) system (Daily et al. 2009; Nelson 2009) developed by the Natural Capital Project (www.naturalcapital.org). However, good tools are valuable only if they are used. A more difficult goal is convincing the private and public sectors to incorporate ecosystem services into their decision - making processes (Daily et al. 2009). Nevertheless, with the socio-economic impacts and human costs of environmental catastrophes, such as Hurricane Katrina, getting bigger and more visible, and with climate change and related carbon sequestration schemes having reached a prominent place in the public consciousness, the value of these services and the necessity of maintaining them has become increasingly mainstream. Recent market-based approaches such as payments for Costa Rican ecosystem services, wetland mitigation banks, and the Chicago Climate Exchange have proven useful in the valuation of ecosystem services (Brauman and Daily 2008). Even though the planet’s ecosystems, the biodiversity they harbor, and the services they collectively provide are truly priceless, market-based and other quantitative approaches for valuing ecosystem services will raise the profile of nature’s services in the public consciousness, integrate these services into decision-making, and help ensure the continuity of ecosystem contributions to the healthy functioning of our planet and its residents.

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Nature’s Cures versus Emerging Diseases

            While many people know about how plants prevent erosion, protect water supplies, and “clean the air”, how bees pollinate plants or how owls reduce rodent activity, many lesser-known organisms not only have crucial ecological roles, but also produce unique chemicals and pharmaceuticals that can literally save people’s lives. Thousands of plant species are used medically by traditional, indigenous communities worldwide. These peoples’ ethno botanical knowledge has led to the patenting, by pharmaceutical companies, of more than a quarter of all medicines (Posey 1999), although the indigenous communities rarely benefit from these patents (Mgbeoji 2006). Furthermore, the eroding of traditions worldwide, increasing emigration from traditional, rural communities to urban areas, and disappearing cultures and languages mean that the priceless ethno botanical knowledge of many cultures is rapidly disappearing in parallel with the impending extinctions of many medicinal plants due to habitat loss and over harvesting (Millennium Ecosystem Assessment 2005a). Some of the rainforest areas that are being deforested fastest, like the island of Borneo, harbor plant species that produce active anti-HIV (Human Immunodeficiency Virus) agents (Chung 1996; Jassim and Naji 2003). 

             Doubtlessly, thousands more useful and vital plant compounds await discovery in the forests of the world, particularly in the bio diverse tropics (Laurance 1999; Sodhi et al. 2007)however, without an effective strategy that integrates community based habitat conservation, rewarding of local ethno botanical knowledge, and scientific research on these compounds, many species, the local knowledge of them, and the priceless cures they offer will disappear before scientists discover them.

              As with many of nature’s services, there is a flip side to the medicinal benefits of biodiversity, namely, emerging diseases ( Jones et al. 2008). The planet’s organisms also include countless diseases, many of which are making the transition to humans as people increasingly invade the habitats of the hosts of these diseases and consume the hosts themselves. Three quarters of human diseases are thought to have their origins in domestic or wild animals and new diseases are emerging as humans increase their presence in formerly wilderness areas (Daily and Ehrlich 1996; Foley et al. 2005). Some of the deadliest diseases, such as monkey pox, malaria, HIV and Ebola, are thought to have initially crossed from central African primates to the people who hunted, butchered, and consumed them (Hahn et al. 2000; Wolfe et al. 2005; Rich et al. 2009). Some diseases emerge in ways that show the difficulty of predicting the consequences of disturbing ecosystems. The extensive smoke from the massive 1997–1998 forest fires in South east Asia is thought to have led to the fruiting failure of many forest trees, forcing frugivorous bats to switch to fruit trees in pig farms. The bats, which host the Nipah virus, likely passed it to the pigs, from which the virus made the jump to people (Chivian 2002). Another classic example from Southeast Asia is the Severe Acute Respiratory Syndrome (SARS). 

            So far having killed 774 people, the SARS coronavirus has been recently discovered in wild animals like the masked palm civet (Paguma larvata) and raccoon dog (Nyctereuteus procyonoides) that are frequently consumed by people in the region (Guan et al. 2003). SARS like corona viruses have been discovered in bats (Li et al. 2005) and the virus was probably passed to civets and other animals as they ate fruits partially eaten and dropped by those bats (Jamie H. Jones, personal communication). 

              It is probable that SARS made the final jump to people through such animals bought for food in wildlife markets. The recent emergence of the deadly avian influenza strain H5N1 provides another good example. Even though there are known to be at least 144 strains of avian flu, only a few strains kill people. However, some of the deadliest pandemics have been among these strains, including H1N1, H2N, and H3N2 (Cox and Subbarao 2000). H5N1, the cause of the recent bird flu panic, has a 50% fatality rate and may cause another human pandemic. At low host densities, viruses that become too deadly, fail to spread. It is likely that raising domestic birds in increasingly higher densities led to
the evolution of higher virulence in H5N1, as it became easier for the virus to jump to another host before it killed its original host. There is also a possibility that increased invasion of wilderness areas by people led to the jump of H5N1 from wild birds to domestic birds, but that is yet to be proven. Malaria, recently shown to have jumped from chimpanzees to humans (Rich et al. 2009), is perhaps the best example of a resurging disease that increases as a result of tropical deforestation (Singer and Castro 2001; Foley et al. 2005; Yasuoka and Levins 2007). 

               Pearson (2003) calculated that every 1% increase in deforestation in the Amazon leads to an 8% increase in the population of the malaria vector mosquito (Anopheles darlingi). In addition, some immigrants colonizing deforested areas brought new sources of malaria (Moran 1988) whereas other immigrants come from malaria-free areas and thus become ideal hosts with no immunity (Aiken and Leigh 1992). Collectively, the conditions leading to and resulting from tropical deforestation, combined with climate change, human migration, agricultural intensification, and animal trafficking create the perfect storm for the emergence of new diseases as well as the resurgence of old ones. In the face of rapid global change, ecologically intact and relatively stable communities may be our best weapon against the emergence of new diseases.

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Mobile Links

             “Mobile links” are animal species that provide critical ecosystem services and increase ecosystem resilience by connecting habitats and ecosystems as they move between them (Gilbert 1980; Lundberg and Moberg 2003;). Mobile links are crucial for maintaining ecosystem function, memory, and resilience (Nystrm and Folke 2001). The three main types of mobile links: genetic, process, and resource links (Lundberg and Moberg 2003), encompass many fundamental ecosystem services (Sekercioglu 2006a, 2006b). Pollinating nectarivores and seed dispersing frugivores are genetic links that carry genetic material from an individual plant to another plant or to a habitat suitable for regeneration, respectively. Trophic process links are grazers, such as antelopes, and predators, such as lions, bats, and birds of prey that influence the populations of plant, invertebrate, and vertebrate prey. Scavengers, such as vultures, are crucial process links that hasten the decomposition of potentially disease-carrying carcasses (Houston 1994). Predators often provide natural pest control (Holmes et al. 1979). Many animals, such as fish-eating birds that nest in colonies, are resource links that transport nutrients in their droppings and often contribute significant resources to nutrient-deprived ecosystems (Anderson and Polis 1999). Some organisms like woodpeckers or beavers act as physical process linkers or “ecosystem engineers” (Jones et al. 1994). By building dams and flooding large areas, beavers engineer ecosystems, create new wetlands, and lead to major changes in species composition. In addition to consuming insects (trophic linkers), many woodpeckers also engineer their environment and build nest holes later used by a variety of other species (Daily et al. 1993). Through mobile links, distant ecosystems and habitats are linked to and influence one another (Lundberg and Moberg 2003). 

               The long-distance migrations of many species, such as African antelopes, songbirds, waterfowl, and gray whales (Eschrichtius robustus) are particularly important examples of critical mobile links. However, many major migrations are disappearing (Wilcove 2008) and nearly two hundred migratory bird species are threatened or near threatened with extinction (Sekercioglu 2007). Dispersing seeds is among the most important functions of mobile links. Vertebrates are the main seed vectors for flowering plants (Regal 1977; Tiffney and Mazer 1995), particularly woody species (Howe and Smallwood 1982; Levey et al. 1994; Jordano 2000). This is especially true in the tropics where bird seed dispersal may have led to the emergence of flowering plant dominance (Regal 1977; Tiffney and Mazer 1995). Seed dispersal is thought to benefit plants in three major ways (Howe and Smallwood 1982):
 

• Escape from density-dependent mortality caused by pathogens, seed predators, competitors, and herbivores (Janzen-Connell escape hypothesis).

• Chance colonization of favorable but unpredictable sites via wide dissemination of seeds. 

• Directed dispersal to specific sites that are particularly favorable for establishment and survival.

               Although most seeds are dispersed over short distances, long-distance dispersal is crucial (Cain et al. 2000), especially over geological time scales during which some plant species have been calculated to achieve colonization distances 20 times higher than would be possible without vertebrate seed dispersers (Cain et al. 2000). Seed dispersers play critical roles in the regeneration and restoration of disturbed and degraded ecosystems (Wunderle 1997;), including newly-formed volcanic soils (Nishi and Tsuyuzaki 2004). Plant reproduction is particularly pollination limited in the tropics relative to the temperate zone (Vamosi et al. 2006) due to the tropics greater biodiversity, and up to 98% of tropical rainforest trees are pollinated by animals (Bawa 1990). Pollination is a critical ecosystem function for the continued persistence of the most biodiverse terrestrial habitats on Earth. Nabhan and Buchmann (1997) estimated that more than 1200 vertebrate and about 289 000 invertebrate species are involved in pollinating over 90% of flowering plant species (angiosperms) and 95% of food crops. Bees, which pollinate about two thirds of the world’s flowering plant species and three quarters of food crops (Nabhan and Buchmann 1997), are the most important group of pollinators. In California alone, their services are estimated to be worth $4.2 billion (Brauman and Daily 2008). However, bee numbers worldwide are declining (Nabhan and Buchmann 1997). In addition to the ubiquitous European honeybee (Apis mellifera), native bee species that depend on natural habitats also provide valuable services to farmers, exemplified by Costa Rican forest bees whose activities increase coffee yield by 20% near forest fragments (Ricketts et al. 2004). 

              Some plant species mostly depend on a single (Parra et al. 1993) or a few(Rathcke 2000) pollinator species. Plants are more likely to be pollinator-limited than disperser limited (Kelly et al. 2004) and a survey of pollination experiments for 186 species showed that about half were pollinator-limited (Burd 1994). Compared to seed dispersal, pollination is more demanding due to the faster ripening rates and shorter lives of flowers (Kelly et al. 2004). Seed disperser and pollinator limitation are often more important in island ecosystems with fewer species, tighter linkages, and higher vulnerability to disturbance and introduced species. Island plant species are more vulnerable to the extinctions of their pollinators sincemany island plants have lost their ability to self-pollinate and have become completely dependent on endemic pollinators (Cox and Elmqvist 2000). Pollination limitation due to the reduced species richness of pollinators on islands likeNew Zealand and Madagascar (Farwig et al. 2004) can significantly reduce fruit sets and probably decrease the reproductive success of dioecious plant species. Predators are important trophic process links and can control the populations of pest species. Formillennia, agricultural pests have been competing with people for the food and fiber plants that feed and clothe humanity. Pests, particularly herbivorous insects, consume 25–50% of humanity’s crops every year (Pimentel et al. 1989). In the US alone, despite the US$25 billion spent on pesticides annually (Naylor and Ehrlich 1997), pests destroy 37% of the potential crop yield (Pimentel et al. 1997). 

             However, many pests have evolved resistance to the millions of tons of synthetic pesticide sprayed each year (Pimentel and Lehman 1993), largely due to insects’ short generation times and their experience with millions of years of coevolution with plant toxins (Ehrlich and Raven 1964). Consequently, these chemicals poison the environment (Carson 1962), lead to thousands of wildlife fatalities every year, and by killing pests’ natural enemies faster than the pests themselves, often lead to the emergence of new pest populations (Naylor and Ehrlich 1997). As a result, the value of natural pest control has been increasingly recognized world wide, some major successes have been achieved, and natural controls now form a core component of “integrated pest management” (IPM) that aims to restore the natural pest-predator balance in agricultural ecosystems (Naylor and Ehrlich 1997). Species that provide natural pest control range from bacteria and viruses to invertebrate and vertebrate predators feeding on insect and rodent pests (Polis et al. 2000; Perfecto et al. 2004; Sekercioglu 2006b). For example, a review by Holmes (1990) showed that reductions in moth and butterfly populations due to temperate forest birds was mostly between 40–70% at low insect densities, 20–60% at intermediate densities, and 0–10% at high densities. Although birds are not usually thought of as important control agents, avian control of insect herbivores and consequent reductions in plant damage can have important economic value (Mols and Visser 2002). Takekawa and Garton (1984) calculated avian control of western spruce budworm in northern Washington State to be worth at least US$1820/ km2/year. To make Beijing greener for the 2008 Olympics without using chemicals, entomologists reared four billion parasitic wasps to get rid of the defoliating moths in less than three months (Rayner 2008). Collectively, natural enemies of crop pests may save humanity at least US $54 billion per year, not to mention the critical importance of natural controls for food security and human survival (Naylor and Ehrlich 1997). Promoting natural predators and preserving their native habitat patches like hedgerows and forests may increase crop yields, improve food security, and lead to a healthier environment. Often underappreciated are the scavenging and nutrient deposition services of mobile links. Scavengers like vultures rapidly get rid of rotting carcasses, recycle nutrients, and lead other animals to carcasses (Sekercioglu 2006a). Besides their ecological significance, vultures are particularly important in many tropical developing countries where sanitary waste and carcass disposal programs may be limited or non-existent (Prakash et al. 2003) and where vultures contribute to human and ecosystem health by getting rid of refuse (Pomeroy 1975), feces (Negro et al. 2002), and dead animals (Prakash et al. 2003). Mobile links also transport nutrients from one habitat to another. 

              Some important examples are geese transporting terrestrial nutrients towet lands (Post et al. 1998) and seabirds transferring marine productivity to terrestrial ecosystems, especially in coastal areas and unproductive island systems (Sanchez-pinero and Polis 2000). Seabird droppings (guano) are enriched in important plant nutrients such as calcium, magnesium, nitrogen, phosphorous, and potassium (Gillham1956).Murphy (1981) estimated that seabirds around the world transfer 104 to 105 tons of phosphorous from sea to land every year. Guano also provides an important source of fertilizer and income to many people living near seabird colonies. Scavengers and seabirds provide good examples of how the population declines of ecosystem service providers lead to reductions in their services (Hughes et al. 1997). Scavenging and fish eating birds comprise the most threatened avian functional groups, with about 40% and 33%, respectively, of these species being threatened or near threatened with extinction (Sekercioglu et al. 2004). The large declines in the populations of many scavenging and fish-eating species mean that even if none of these species go extinct, their services are declining substantially. Seabird losses can trigger trophic cascades and ecosystem shifts (Croll et al. 2005). Vulture declines can lead to the emergence of public health problems. In India, Gyps vulture populations declined as much as 99% in the 1990s (Prakash et al. 2003). Vultures compete with feral dogs, which often carry rabies. As the vultures declined between 1992 and 2001, the numbers of feral dogs increased 20-fold at a garbage dump in India (Prakash et al. 2003). Most of world’s rabies deaths take place in India (World Health Organization 1998) and feral dogs replacing vultures is likely to aggravate this problem. Mobile links, however, can be double-edged swords and can harm ecosystems and human populations, particularly in concert with human related poor land-use practices, climate change, and introduced species. 

               Invasive plants can spread via native and introduced seed dispersers (Larosa et al. 1985; Cordeiro et al. 2004). Land use change can increase the numbers ofmobile links that damage distant areas, such as when geese overload wetlands with excessive nutrients (Post et al. 1998). Climate change can lead to asynchronies in insect emergence and their predators timing of breeding (Both et al. 2006), and in flowering and their pollinators lifecycles (Harrington et al. 1999). Mobile links are often critical to ecosystem functioning as sources of “external memory” that promote the resilience of ecosystems (Scheffer et al. 2001). More attention needs to be paid to mobile links in ecosystem management and biodiversity conservation (Lundberg and Moberg 2003). This is especially the case for migrating species that face countless challenges during their annual migrations that sometimes cover more than 20 000 kilometers (Wilcove 2008). Some of the characteristics that make mobile links important for ecosystems, such as high mobility and specialized diets, also make them more vulnerable to human impact. Protecting pollinators, seed dispersers, predators, scavengers, nutrient depositors, and other mobile linksmust be a top conservation priority to prevent collapses in ecosystem services provided by these
vital organisms.

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Biodiversity and Ecosystem Function

               The role of biodiversity in providing ecosystem services is actively debated in ecology. The diversity of functional groups (groups of ecologically equivalent species (Naeem and Li 1997)), is as important as species diversity, if not more so (Kremen 2005), and in most services a few dominant species seem to play the major role (Hooper et al. 2005). However, many other species are critical for eco system functioning and provide “insurance” against disturbance, environmental change, and the decline of the dominant species (Tilman 1997; Ricketts et al. 2004; Hobbs et al. 2007). As for many other ecological processes, it was Charles Darwin who first wrote of this, noting that several distinct genera of grasses grown together would produce more plants and more herbage than a single species growing alone (Darwin 1872). Many studies have confirmed that increased biodiversity improves ecosystem functioning in plant communities (Naeem and Li 1997; Tilman 1997). Different plant species capture different resources, leading to greater efficiency and higher productivity (Tilman et al. 1996). Due to the“sampling-competition effect” the presence of more species increases the probability of having a particularly productive species in any given environment (Tilman 1997). Furthermore, different species’ ecologies lead to complementary resource use, where each species grows best under a specific range of environmental conditions, and different species can improve environmental conditions for other species (facilitation effect; Hooper et al. 2005). Consequently, the more complex an ecosystem is, the more biodiversity will increase ecosystem function, as more species are needed to fully exploit the many combinations of environmental variables (Tilman 1997). More biodiverse ecosystems are also likely to be more stable and more efficient due to the presence of more pathways for energy flow and nutrient recycling (Macarthur 1955; Hooper et al. 2005; Vitousek and Hooper 1993; Worm et al. 2006). Greenhouse and field experiments have confirmed that biodiversity does increase ecosystem productivity, while reducing fluctuations in productivity (Naeem et al. 1995; Tilman et al. 1996). 

               Although increased diversity can increase the population fluctuations of individual species, diversity is thought to stabilize overall ecosystem functioning (Chapin et al. 2000; Tilman 1996) and make the ecosystem more resistant to perturbations (Pimm 1984). These hypotheses have been confirmed in field experiments, where speciesrich plots showed less yearly variation in productivity (Tilman 1996) and their productivity during a drought year declined much less than species poor plots (Tilman and Downing 1994). Because more species do better at utilizing and recycling nutrients, in the long-term, species-rich plots are
better at reducing nutrient losses and maintaining soil fertility (Tilman et al. 1996; Vitousek and Hooper 1993). Although it makes intuitive sense that the species that dominate in number and/or biomass are more likely to be important for ecosystem function (Raffaelli 2004; Smith et al. 2004), in some cases, even rare species can have a role, forexample, in increasing resistance to invasion (Lyons and Schwartz 2001). A keystone species is one that has an ecosystem impact that is disproportionately large in relation to its abundance (Hooper et al. 2005; Power et al. 1996;). Species that are not thought of as “typical” keystones can turn out to be so, sometimes in more ways than one (Daily et al. 1993). Even though in many communities only a few species have strong effects, the weak effects of many species can add up to a substantial stabilizing effect and seemingly “weak” effects over broad scales can be strong at the local level (Berlow 1999). Increased species richness can “insure” against sudden change, which is now a global phenomenon (Parmesan and Yohe 2003; Root et al. 2003). Even though a few species may make up most of the biomass of most functional groups, this does not mean that other species are
unnecessary (Walker et al. 1999). 

              Species may act like the rivets in an airplane wing, the loss of each unnoticed until a catastrophic threshold is passed (Ehrlich and Ehrlich 1981b). As humanity’s footprint on the planet increases and formerly stable ecosystems experience constant disruptions in the form of introduced species, pollution , climate change, excessive nutrient loads, fires, and many other perturbations, the insurance value of biodiversity has become increasingly vital over the entire range of habitats and systems, from diverse forest stands sequesterin CO2 better in the long-term (Bolker et al. 1995; Hooper et al. 2005; but see Tallis and Kareiva 2006) to forest-dwelling native bees’ coffee pollination services increasing coffee production in Costa Rica (Ricketts et al. 2004;). With accelerating losses of unique species, humanity, far from hedging its bets, is moving ever closer to the day when we will run out of options on an increasingly unstable planet.

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Soils and Erosion

                 Without forest cover, erosion rates skyrocket, and many countries, especially in the tropics, lose astounding amounts of soil to erosion. Worldwide, 11 million km2 of land (the area of USA and Mexico combined) are affected by high rates of erosion (Millennium Ecosystem Assessment 2005b). Every year about 75 billion tons of soil are thought to be eroded from terrestrial ecosystems, at rates 13–40 times faster than the average rate of soil formation (Pimentel and Kounang 1998). Pimentel et al. (1995) estimated that in the second half of the 20th century about a third of the world’s arable land was lost to erosion. This means losing vital harvests and income (Myers 1997), not to mention losing lives to malnutrition and starvation. Soil is one of the most critical but also most underappreciated and abused elements of natural capital, one that can take a few years to lose and millennia to replace. A soil’s character is determined by six factors: topography, the nature of the parent material, the age of the soil, soil organisms and plants, climate, and human activity (Daily et al. 1997). For example, in the tropics, farming can result in the loss of half the soil nutrients in less than a decade (Bolin and Cook 1983), a loss that can take centuries to restore. In arid areas, the replacement of native deep rooted plants with shallow-rooted crop plants can lead to a rise in the water table, which can bring soil salts to the surface (salinization), cause water logging, and consequently result in crop losses (Lefroy et al. 1993). Soil provides six major ecosystem services (Daily et al. 1997):

• Moderating the hydrologic cycle. 
• Physical support of plants. 
• Retention and delivery of nutrients to plants. 
• Disposal of wastes and dead organic matter. 
• Renewal of soil fertility. 
• Regulation of major element cycles.

                Every year enough rain falls to cover the planet with one meter of water (Shiklomanov 1993), but thanks to soil’s enormous water retention capacity, most of this water is absorbed and gradually released to feed plants, underground aquifers, and rivers.However, intensive cultivation, by lowering soil’s organic matter content, can reduce this capacity, leading to floods, erosion, pollution, and further loss of organic matter (Pimentel et al. 1995). Soil particles usually carry a negative charge, which plays a critical role in delivering nutrient cations (positively-charged ions) like Ca2þ, Kþ, Naþ, NH4þ, and Mg2þ to plants (Daily et al. 1997). To deliver these nutrients without soil would be exceedingly expensive as modern hydroponic (water-based) systems cost more than US$250 000 per ha (Canada’s Office of Urban
Agriculture 2008; Avinash 2008). 

                Soil is also critical in filtering and purifying water by removing contaminants, bacteria, and other impurities (Fujii et al. 2001). Soils harbor an astounding diversity of microorganisms, including thousands of species of protozoa, antibiotic-producing bacteria (which produce streptomycin) and fungi (producing penicillin), as well as myriad invertebrates, worms and algae (Daily et al. 1997). These organisms play fundamental roles in decomposing dead matter, neutralizing deadly pathogens, and recycling waste into valuable nutrients. Just the nitrogen fixed by soil organisms like Rhizobium bacteria amounts to about 100 million metric tons per year (Schlesinger 1991). It would cost at least US$320 billon/year to replace natural nitrogen fertilization with fertilizers (Daily et al. 1997). As the accelerating release of CO2, N2O (Nitrous Oxide), methane and other greenhouse gases increasingly modifies climate (IPCC 2007), the soil’s capacity to store these molecules is becoming even more vital. Per area, soil stores 1.8 times the carbon and 18 times the nitrogen that plants alone can store (Schlesinger 1991). For peatlands, soil carbon storage can be 10 times greater than that stored by the plants growing on it and peatland fires release massive amounts of CO2 into the atmosphere (Page and Rieley 1998). Despite soil’s vital importance, 17% of the Earth’s vegetated land surface (Oldeman 1998) or 23% of all land used for food production
[FAO (Food and Agriculture Organization of the United Nations) 1990] has experienced soil degradation since 1945. Erosion is the best-known example of the disruption of the sedimentary cycle. 

                       Although erosion is responsible for releasing nutrients from bedrock and making them available to plants, excessive wind and water erosion results in the removal of top soil, the loss of valuable nutrients, and desertification. The direct costs of erosion total about US$250 billion per year and the indirect costs (e.g. siltation, obsolescence of dams, water quality declines) approximately $150 billion per year (Pimentel et al. 1995). Sufficient preventive measures would cost only 19% of this total (Pimentel et al. 1995). The loss of vegetative cover increases the erosional impact of rain. In intact forests, most rain water does not hit the ground directly and tree roots hold the soil together against being washed away (Brauman et al. 2007), better than in logged forest or plantations (Myers 1997) where roads can increase erosion rates (Bruijnzeel 2004). The expansion of farming and deforestation have doubled the amount of sediment discharged into the oceans. Coral reefs can experience high mortality after being buried by sediment discharge (Pandolfi et al. 2003; Bruno and Selig 2007). Wind erosion can be particularly severe in desert ecosystems, where even small increases in vegetative cover (Hupy 2004) and reduced tillage practices (Gomes et al. 2003) can lessen wind erosion substantially. Montane areas are especially prone to rapid erosion (Milliman and Syvitski 1992), and revegetation programs are critical in such ecosystems (Vanacker et al. 2007). Interestingly, soil carbon buried in deposits resulting from erosion, can produce carbon sinks that can offset up to 10% of the global fossil fuel emissions of CO2 (Berhe et al. 2007). 

                 However, erosion also lowers soil productivity and reduces the organic carbon returned to soil as plant residue (Gregorich et al. 1998). Increasing soil carbon capacity by 5–15% through soil-friendly tillage practices not only offsets fossil-fuel carbon emissions by a roughly equal amount but also increases crop yields and enhances food security (Lal 2004). An increase of one ton of soil carbon pool in degraded cropland soils may increase crop yield by 20 to 40 kilograms per ha (kg/ha) for wheat, 10 to 20 kg/ha for maize, and 0.5 to 1 kg/ha for cowpeas (Lal 2004).

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Regulation of the Hydrologic Cycle

                One of the most vital and immediate services of ecosystems, particularly of forests, rivers and wetlands, is the provisioning and regulation of water resources. These services provide a vast range of benefits from spiritual to life-saving, illustrated by the classification of hydrologic services into five broad categories by Brauman et al. (2007): improvement of extractive water supply, improvement of in-stream water supply, water damage mitigation, provision of water-related cultural services, and water-associated supporting services. Although 71% of the planet is covered by water, most of this is seawater unfit for drinking or agriculture (Postel et al. 1996). Fresh water not locked away in glaciers and icecaps constitutes 0.77% of the planet’s water (Shiklomanov 1993). To provide sufficient fresh water to meet human needs via industrial desalination (removing the salt from seawater) would cost US$3 000 billion per year (Postel and Carpenter 1997). Quantity, quality, location, and timing of water provision determine the scale and impact of hydrologic services (Brauman et al. 2007). These attributes can make the difference between water as a blessing (e.g. drinking water) or a curse (e.g. floods). Water is constantly redistributed through the hydrologic cycle. Fresh water comes down as precipitation, collects in water bodies or is absorbed by the soil and plants. 

                 Some of the water flows unutilized into the sea or seeps intounderground aquifers where it can remain for millennia unless extracted by people; mining this “fossil” groundwater is often unsustainable and is a serious problem in desert regions like Libya (Millennium Ecosystem Assessment 2005c). The cycle is completed when water vapor is released back into the atmosphere either through evaporation from land and water bodies or by being released
from plants (transpiration) and other organisms. Rising environmental temperatures are expected to increase evaporation and consequent precipitation in some places and raise the likelihood of droughts and fires in other places, both scenarios that would have major consequences for the world’s vegetation (Wright 2005). These changes in turn can lead to further climatic problems, affecting agriculture and communities worldwide. Ecosystems, particularly forests, play major roles in the regulation of the hydrologic cycle and also have the potential to moderate the effects of climate change. Tropical forests act as heat and humidity pumps, transferring heat from the tropics to the temperate zones and releasing water vapor that comes back as rain (Sodhi et al. 2007). Extensive tropical deforestation is expected to lead to higher temperatures, reduced precipitation, and increased frequency of droughts and fires, all ofwhich are likely to reduce tropical forest cover in a positive feedback loop (Sodhi et al. 2007). Forest ecosystems alone are thought to regulate approximately a third of the planet’s watersheds on which nearly five billion people rely (Millennium Ecosystem Assessment 2005c). With increasing human population and consequent water pollution, fresh water is becoming an increasingly precious resource, especially in arid areas like the Middle East, where the scarcity of water is likely to lead to increasing local conflicts in the 21st century (Klare 2001; Selby 2005). 

                  Aquatic ecosystems, in addition to being vital sources of water, fish, waterfowl, reeds, and other resources, also moderate the local climate and can act as buffers for floods, tsunamis, and other water incursions. For example, the flooding following Hurricane Katrina would have done less damage if the coastal wetlands surrounding New Orleans had had their original extent (Day et al. 2007). The impact of the 24 December 2004 tsunami in Southeast Asia would have been reduced if some of the hardest hit areas had not been stripped of their mangrove forests (Dahdouh-guebas et al. 2005; Danielsen et al. 2005). These observations support analytical models in which thirty “waru” trees (Hibiscus tiliaceus) planted along a 100 m by 1 meter band reduced the impact of a tsunami by 90% (Hiraishi and Harada 2003), a solution more effective and cheaper than artificial barriers. Hydrologic regulation by ecosystems begins with the first drop of rain. Vegetation layers, especially trees, intercept raindrops, which gradually descend into the soil, rather than hitting it directly and leading to erosion and floods. By intercepting rainfall and promoting soil development, vegetation can modulate the timing of flows and potentially reduce flooding. Flood mitigation is particularly crucial in tropical areas where downpours can rapidly deposit enormous amounts of water that can lead to increased erosion, floods, and deaths if there is little natural forest to absorb the rainfall (Bradshaw et al. 2007). Studies of some watersheds have shown that native forests reduced flood risks only at small scales, leading some hydrologists to question directly connecting forest cover to flood reduction (Calder and Aylward 2006). However, in the first global-scale empirical demonstration that forests are correlated with flood risk and severity in developing countries, Bradshaw et al. (2007) estimated that a 10% decrease in natural forest area would lead to a flood frequency increase between 4% and 28%, and to a 4–8% increase in total flood duration at the country scale. 

                Compared to natural forests, however, afforestation programs or forest plantations may not reduce floods, or may even increase flood volume due to road construction, soil compaction, and changes in drainage regimes (Calder and Aylward 2006). Non-native plantations can do more harm than good, particularly when they reduce dry season water flows (Scott et al. 2005). Despite covering only 6% of the planet’s surface, tropical forests receive nearly half of the world’s rainfall, which can be as much as 22 500 mm during five months of monsoon season in India (Myers 1997). In Southeast Asia, an intact old-growth dipterocarp forest intercepts at least 35% of the rainfall, while a logged forest intercepts less than 20%, and an oil palm (Elaeis spp.) plantation intercepts only 12% (Ba 1977). As a consequence, primary forest can moderate seasonal extremes in water flow and availability better than more intensive land uses like plantation forestry and agriculture. For example, primary forest in Ivory Coast releases three to five times as much water at the end of the dry season compared to a coffee plantation (Dosso 1981). However, it is difficult to make generalizations about hydrologic response in the tropics. For example, local soil and rainfall patterns can result in a
65-fold variation in tropical natural sedimentation rates (Bruijnzeel 2004). This underlines the importance of site-specific studies in the tropics, but most hydrologic studies of ecosystems have taken place in temperate ecosystems (Brauman et al. 2007).

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Climate and the Biogeochemical Cycles

                 Ecosystem services start at the most fundamental level: the creation of the air we breathe and the supply and distribution of water we drink. Through photosynthesis by bacteria, algae, plankton, and plants, atmospheric oxygen is mostly generated and maintained by ecosystems and their constituent species, allowing humans and innumerable other oxygen-dependent organisms to survive. Oxygen also enables the atmosphere to “clean” itself via the oxidation of
compounds such as carbon monoxide (Sodhi et al. 2007) and another form of oxygen in the ozone layer, protects life from the sun’s carcinogenic, ultraviolet (UV) rays. Global bio geochemical cycles consist of “the transport and transformation of substances in the environment through life, air, sea, land, and ice” (Alexander et al. 1997). 

                 Through these cycles, the planet’s climate, ecosystems, and creaturesare tightly linked. Changes in one component can have drastic effects on another, as exemplified by the effects of deforestation on climatic change (Phat et al. 2004). The hydrologic cycle is one that most immediately affects our lives and it is treated separately below. As carbon-based life forms, every single organism on our planet is a part of the global carbon cycle. This cycle takes place between the four main reservoirs of carbon: carbon dioxide (CO2) in the atmosphere; organic carbon compounds within organisms; dissolved carbon in water bodies; and carbon compounds inside the earth as part of soil, limestone (calcium carbonate), and buried organic matter like coal, natural gas, peat, and petroleum (Alexander et al. 1997). 

                 Plants play a major role in fixing atmospheric CO2 through photosynthesis and most terrestrial carbon storage occurs in forest trees (Falkowski et al. 2000). The global carbon cycle has been disturbed by about 13% compared to the pre-industrial era, as opposed to 100% or more for nitrogen, phosphorous, and sulfur cycles (Falkowski et al. 2000). Given the dominance of carbon in shaping life and in regulating climate, however, this perturbation has already been enough to lead to significant climate change with worse likely to come in the future [IPCC (Intergovernmental Panel on Climate Change) 2007]. Because gases like CO2, methane (CH4), and nitrous oxide (N2O) trap the sun’s heat, especially the long-wave infrared radiation that’s emitted by the warmed planet, the atmosphere creates a natural “greenhouse” (Houghton 2004). 

                Without this greenhouse effect, humans and most other organisms would be unable to survive, as the global mean surface temperature would drop from the current 14 C to –19 C (IPCC 2007). Ironically, the ever-rising consumption of fossil fuels during the industrial age and the resultant increasing emission of greenhouse gases have created the opposite problem, leading to an increase in the magnitude of the greenhouse effect and a consequent rise in global temperatures (IPCC 2007). Since 1750, atmospheric CO2 concentrations have increased by 34% (Millennium Ecosystem Assessment 2005a) and by the end of this century, average global temperature is projected to rise by
1.8 –6.4 C (IPCC 2007). Increasing deforestation and warming both exacerbate the problem as forest ecosystems switch from being major carbon sinks to being carbon sources (Phat et al. 2004; IPCC 2007). If fossil fuel consumption and deforestation continue unabated, global CO2 emissions are expected to be about 2–4 times higher than at present by the year 2100 (IPCC 2007). As climate and life have coevolved for billions of years and interact with each other through various feedback mechanisms (Schneider and Londer 1984), rapid climate change would have major consequences for the planet’s life-support systems. There are now plans under way for developed nations to finance the conservation of tropical forests in the developing world so that these forests can continue to provide the ecosystem service of acting as carbon sinks (Butler 2008). 

                 Changes in ecosystems affect nitrogen, phosphorus, and sulfur cycles as well (Alexander et al. 1997; Millennium Ecosystem Assessment 2005b; Vitousek et al. 1997). Although nitrogen in its gaseous form (N2) makes up 80% of the atmosphere, it is only made available to organisms through nitrogen fixation by cyanobacteria in aquatic systems and on land by bacteria and algae that live in the root nodules of lichens and legumes (Alexander et al. 1997). Eighty million tons of nitrogen every year are fixed artificially by industry to be used as fertilizer (Millennium Ecosystem Assessment 2005b). However, the excessive use of nitrogen fertilizers can lead to nutrient overload, eutrophication, and elimination of oxygen in water bodies. Nitrogen oxides, regularly produced as a result of fossil fuel combustion, are potent greenhouse gases that increase global warming and also lead to smog, breakdown of the ozone layer, and acid rain (Alexander et al. 1997). Similarly, although sulfur is an essential element in proteins, excessive sulfur emissions from human activities lead to sulfuric acid smog and acid rain that harms people and ecosystems alike (Alexander et al. 1997). Phosphorous (P) scarcity limits biological nitrogen fixation (Smith 1992). 

                In many terrestrial ecosystems, where P is scarce, specialized symbiotic fungi (mycorrhizae) facilitate P uptake by plants (Millennium Ecosystem Assessment 2005b). Even though P is among the least naturally available of major nutrients, use of phosphorous in artificial fertilizers and runoff from animal husbandry often also leads to eutrophication in aquatic systems (Millenium Ecosystem Assessment 2005b). The mining of phosphate deposits and their addition to terrestrial ecosystems as fertilizers represents a six fold increase over the natural rate of mobilization of P by the weathering of phosphate rock and by plant activity (Reeburgh 1997). P enters aquatic ecosystems mainly through erosion, but no-till agriculture and the use of hedgerows can substantially reduce the rate of this process (Millenium Ecosystem Assessment 2005a).

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Ecosystem functions and services

                  In our increasingly technological society, people give little thought to how dependent they are on the proper functioning of ecosystems and the crucial services for humanity that flow from them. Ecosystem services are “the conditions and processes through which natural ecosystems, and the species that make them up, sustain and fulfill human life” (Daily 1997); in other words, “the set of ecosystem functions that are useful to humans” (Kremen 2005). Although people have been long aware that natural ecosystems help support human societies, the explicit recognition of “ecosystem services” is relatively recent (Ehrlich and Ehrlich 1981a; Mooney and Ehrlich 1997).
 

                 Since the entire planet is a vast network of integrated ecosystems, ecosystem services range from global to microscopic in scale (Millennium Ecosystem Assessment 2005a). Ecosystems purify the air and water, generate oxygen, and stabilize our climate. Earth would not be fit for our survival if it were not for plants that have created and maintained a suitable atmosphere. Organisms decompose and detoxify detritus, preventing our civilization from being buried under its own waste. Other species help to create the soils on which we grow our food, and recycle the nutrients essential to agriculture. Myriad creatures maintain these soils, play key roles in recycling nutrients, and by so doing help to mitigate erosion and floods. Thousands of animal species pollinate and fertilize plants, protect them from pests, and disperse their seeds. And of course, humans use and trade thousands of plant, animal and microorganism species for food, shelter, medicinal, cultural, aesthetic and many other purposes. Although most people may not know what an ecosystem is, the proper functioning of the world’s ecosystems is critical to human survival, and understanding the basics of ecosystem services is essential. 

                Entire volumes have been written on ecosystem services (National Research Council 2005; Daily 1997), culminating in a formal, in-depth, and global overview by hundreds of scientists: the Millennium Ecosystem Assessment (2005a). It is virtually impossible to list all the ecosystem services let alone the natural products that people directly consume, so this discussion presents a brief introduction to ecosystem function and an overview of critical ecosystem services.

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Where is biodiversity?

                Just as biodiversity has varied markedly through time, so it also varies across space. Indeed, one can think of it as forming a richly textured land and seascape, with peaks (hotspots) and troughs (cold spots), and extensive plains in between. Even locally, and just for particular groups, the numbers of species can be impressive, with for example c.900 species of fungal fruiting bodies recorded from 13 plots totaling just 14.7 ha (hectare) near Vienna, Austria (Straatsma and Krisai- Greilhuber 2003), 173 species of lichens on a single tree in Papua New Guinea (Aptroot 1997), 814 species of trees from a 50 ha study plot in Peninsular Malaysia (Manokaran et al. 1992), 850 species of invertebrates estimated to occur at a sandy beach site in the North Sea (Armonies and Reise 2000), 245 resident species of birds recorded holding territories on a 97 ha plot in Peru (Terborgh et al. 1990), and >200 species of mammals occurring at some sites in the Amazonian rain forest (Voss and Emmons 1996). 

                 Although it remains the case that for no even moderately sized area do we have a comprehensive inventory of all of the species that are present (microorganisms typically remain insufficiently documented even in otherwise well studied areas), knowledge of the basic patterns has been developing rapidly. Although long constrained to data on higher vertebrates, the breadth of organisms for which information is available has been growing, with much recent work particularly attempting to determine whether microorganisms show the same geographic patterns as do other groups.

Land and water

                 The oceans cover 340.1 million km2 (67%), the land 170.3 million km2 (33%), and fresh waters (lakes and rivers) 1.5 million km2 (0.3%; with another 16 million km2 under ice and permanent snow, and 2.6 million km2 as wetlands, soil water and permafrost) of the Earth’s surface. It would therefore seem reasonable to predict that the oceans would be most biodiverse, followed by the land and then freshwaters. In terms of numbers of higher taxa, there is indeed some evidence that marine systems are especially diverse. For example, of the 96 phyla recognized by Margulis and Schwartz (1998), about 69 have marine representatives, 55 have terrestrial ones, and 60 have freshwater representatives. However, of the species described to date only about 15% are marine and 6% are freshwater. The fact that life began in the sea seems likely to have played an important role in explaining why there are larger numbers of higher taxa in marine systems than in terrestrial ones. 

                 The heterogeneity and fragmentation of the land masses (particularly that associated with the breakup of the “supercontinent” of Gondwana from 180 Ma) is important in explaining why there are more species in terrestrial systems than in marine ones. Finally, the extreme fragmentation and isolation of freshwater bodies seems key to why these are so diverse for their area.

Biogeographic realms and ecoregions

                 Of the terrestrial realms, the Neotropics is generally regarded as overall being the most biodiverse, followed by the Afrotropics and Indo Malaya, although the precise ranking of these tropical regions depends on the way in which organismal diversity is measured. For example, for species the richest realm is the Neotropics for amphibians, reptiles, birds and mammals, but for families it is the Afrotropics for amphibians and mammals, the Neotropics for reptiles, and the Indo-Malayan for birds (MEA 2005). In parts, these differences reflect variation in the histories of the realms (especially mountain uplift and climate changes) and the interaction with the emergence and spread of the groups, albeit perhapscomplicated by issues of geographic consistency in the definition of higher taxonomic groupings.
 

              The Western Indo-Pacific and Central Indo-Pacific realms have been argued to be a center for the evolutionary radiation of many groups, and are thought to be perhaps the global hotspot of marine species richness and endemism (Briggs 1999; Roberts et al. 2002). With a shelf area of 6 570 000 km2, which is considered to be a significant influence, it has more than 6000 species of molluscs, 800 species of echinoderms, 500 species of hermatypic (reef forming) corals, and 4000 species of fish (Briggs 1999). At the scale of terrestrial ecoregions, the most speciose for amphibians and reptiles are in the Neotropics, for birds in Indo-Malaya, Neotropics and Afrotropics, and for mammals in the Neotropics, Indo-Malaya, Nearctic, and Afrotropics. Amongst the freshwater ecoregions, those with globally high richness of freshwater fish include the Brahmaputra, Ganges, and Yangtze basins in Asia, and large portions of the Mekong, Chao Phraya, and Sitang and Irrawaddy; the lower Guinea in Africa; and the Paraná and Orinoco in South America (Abell
et al. 2008).

Altitude and Depth

                  Variations in depth in marine systems and altitude in terrestrial ones are small relative to the areal coverage of these systems. The oceans average c.3.8 km in depth but reach down to 10.9 km (Challenger Deep), and land averages 0.84 km in elevation and reaches up to 8.85 km (Mt. Everest). Nonetheless, there are profound changes in organismal diversity both with depth and altitude. This is in large part because of the environmental differences (but also the effects of area and isolation), with some of those changes in depth or altitude of a few hundred meters being similar to those experienced over latitudinal distances of several hundred kilometers (e.g. temperature). In both terrestrial and marine (pelagic and benthic) systems, species richness across a wide variety of taxonomic groups has been found
progressively to decrease with distance from sea level (above or below) and to show a pronounced hump-shaped pattern in which it first increases and then declines (Angel 1994; Rahbek 1995; Bryant et al. 2008). The latter pattern tends to become more apparent when the effects of variation in area have been accounted for, and is probably the more general, although in either case richness tends to be lowest at the most extreme elevations or depths. Microbial assemblages can be found at considerable depths (in some instances up to a few kilometers) below the terrestrial land surface and the seafloor, often exhibiting unusual metabolic capabilities (White et al. 1998; D’Hondt et al. 2004). Knowledge of these assemblages remains, however, extremely poor, given the physical challenges of sampling and of doing so without contamination from other sources.

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