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2 What is Biodiversity? Biodiversity includes not only the world's species with their unique evolutionary histories, but also genetic variability within and among populations of species and the distribution of species across local habitats, ecosystems, landscapes, and whole continents or oceans. Understanding what constitutes and defines biodiversity is essential for managers and policy-makers who must attempt to incorporate its values into their land- and water-management plans. It is only when we understand all the interacting scientific dimensions of biodiversity outlined in this chapter that we can appreciate the levels of information that must be considered. Biodiversity-management options are inevitably constrained by a combination of biological and sociopolitical realities. In this chapter, we present our biological understanding of biodiversity, which provides the basis for further chapters 3 and 4, which consider the "uses" and "value" of biodiversity. The word biodiversity is used in many ways. Economists and ecologists, ranchers and gardeners, mayors and miners all view biodiversity from different perspectives. When people discuss biodiversity, they often use it as a surrogate for "wild places" or "abundance of species" or even ''large, furry mammals". Yet from the viewpoint of those engaged in biodiversity-related sciences—such as population biology, ecology, systematics, evolution, and genetics—biodiversity has a specific meaning: "the variety and variability of biological organisms" (Keystone Center 1991; Noss and Cooperrider 1994; Wilson and Peter 1988). The Convention on Biological Diversity similarly defines biodiversity as the "variability among living organisms from all sources''. Those definitions are so broad that they can be clearly understood only by considering particular levels of biological organization—genes, species, communities, ecosystems, and even our planet.
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Each level requires different methods of analysis, different modes of understanding, and, ultimately, different approaches to management. For managers, it is not just a matter of counting species or individuals. Managers must consider the role of biodiversity in the functioning of ecosystems and the effects of management and use of resources on ecosystem processes (chapter 3). George Evelyn Hutchison (1965), one of the founders of modern ecology, wrote about the "evolutionary play in the ecological theater". This multilayered drama generates, sustains, shapes, and sometimes even diminishes biodiversity. Charles Darwin's reflections on species diversity underlay one of the most far-reaching theories in the history of ideas: the theory of evolution by natural selection. His travels from England to the strikingly different landscapes of the New World left him awestruck and inspired. Whatever constitutes biodiversity, Darwin recognized that Brazil had a lot of it and certainly more than he left behind in an English midwinter. No modern biologist would disagree. Like Darwin, we often equate biodiversity with the number and novelty of the species present. Species, Populations, and Genes There is genetic diversity within species. If each species were reduced to one small population of genetically similar individuals, we would lose much biodiversity. As we move across a region, the species change, even if the numbers of species in different places might not; a forest and an adjacent grassland might contain almost entirely different assemblages of species, for instance. Moreover, the ecosystem processes in a grassland differ from those in the forest nearby. A population consists of individuals of the same species that live in the same place and interact in various ways, including interbreeding. Populations of the same species living in different places can exchange members, but they often are genetically differentiated to some degree and the further they are separated from each other, the more distinctive they become. Metapopulations are groups of spatially separated populations that occur in patches of habitat across a landscape. Populations can become locally extinct in different habitat patches across a landscape; they infrequently exchange members, and when they do, the passage between local populations is generally hazardous and entails movement across inhospitable habitat. Local populations that make up a metapopulation experience extinction, and habitat left open is recolonized at some finite probability by other local populations within the metapopulation. The genetic variability among individuals within a species can result from gene recombination or mutation, genetic polymorphism (the presence of different forms of the same gene), isolation of gene pools, local selection pressures, habitat (environmental) complexity, landscape mosaics, and environmental gradients. Specific genetic combinations in populations result from natural selection acting
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on individuals in response to biotic and abiotic environments and from random, nonselective fixation of genes. New developments in the study of molecular evolution and modern laboratory techniques make it possible to determine the degree or closeness of relationships within and between populations (Avise 1994, 1995; Hillis and others 1996). Molecular data and traditional anatomical information permit us to deduce phylogenies—the branching patterns of genealogical lineages and ancestry of sets of species (Hillis and others 1996). Genetic Diversity and Adaptation Much genetic variation is detectable only biochemically, but some is evident as variation in anatomy, physiology, behavior, and life-history characteristics—phenotypes—of individuals in a population. Genetic variation is the basis of local adaptations and of common phenomenon of gradual change in phenotype along a geographic transect where the environment changes. Genetic variation is also the basis of coevolution, whereby species evolve adaptations in response to each other's adaptations. There are many examples of adaptive evolution within species. Across the extensive continuous range of the common mussel off the eastern coast of North America, despite its enormous reproductive output and high rates of genetic exchange, populations are genetically differentiated over surprisingly small distances—from a few meters to several kilometers (Koehn and Hilbish 1995). The common yarrow, a composite plant from California, is able to live over a great range of habitats, from the high Sierra Nevada to the Pacific Coast, and shows distinctive, genetically determined forms in different habitats (Clausen and others 1958). Drosophila flies show extensive variation in genome organization according to habitat, elevation, regional geography, and seasonality (Dobzhansky 1970). Effective environmental management includes considerations of genetic variation. For example, salmon stocks in different rivers in the same region exhibit differences in genetic makeup. These are the result of independent evolution of distinct stocks, each of which has adapted to local conditions. The differences seen reflect the histories of the stocks, some resulting from local selection pressures and others from the accumulation of random changes associated with the degree of isolation and population size. Genetic diversity provides an economic basis for protecting and conserving biodiversity (McNeely and others 1990; Oldfield 1984; Potter and others 1993; Reid and Miller 1989; Reid and others 1993; WRI/IUCN/UNEP 1992). For example, Douglas fir trees grow abundantly across the western United States. Their success is due to their diversity despite their similar appearance (Rudolph 1990). Coastal and interior populations show genetic differences in cold hardi-
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ness, response to moisture stress, and timing of bud bursts. There are also genetic differences between populations only 3–10 km apart that are exposed to different microclimates on north-facing and south-facing slopes. Moreover, genetic variability results in the continued production of diverse phenotypes, some of which are more able than others to resist attacks by western spruce budworm, an important pest for this species. Commercial nurseries make use of local variation in reforestation programs. Current adaptations are important, but genetic diversity is also critical for the future resilience and persistence of natural systems. Variation is important to maintain a population's ability to respond to changing environmental conditions, whether natural or anthropogenic. A notable example is the rapid adaptive evolution of plants that have colonized mine tailings that are polluted by heavy metals in Great Britain (Antonovics and others 1971). This represents natural evolutionary potential, which can be particularly important in the face of rapid global change. For managers of biodiversity, there are practical implications in the observations that some species have many locally distinct populations but others show little geographic variation and that some species have no close relatives but others occur in genera that include hundreds of species. Biologists have recognized that current taxonomy (the classification of organisms on the basis of the evolution of species from their ancestors) is sometimes inadequate for identifying appropriate units for conservation. They have recommended counting "evolutionarily significant units" (ESUs) (Moritz 1994; NRC 1996), historically isolated parts of species that, in addition to representing divergence and diversification in the past, can have direct evolutionary potential. Focusing on ESUs has the goal of ensuring that evolutionary heritage is recognized and protected. Measures of Diversity One of the decisions that managers face is how to assess biodiversity. How do we know whether biodiversity has changed? Scientists use different methods to assess biodiversity. Biodiversity among areas can be compared with statistical indexes of species diversity (Magurran 1988; Pielou 1975). Most indices combine two different metrics: the total number of species and the relative abundances of all species (evenness) in a sample. Such indexes have been criticized on the grounds that similar values of an index might reflect quite different sample compositions. A given index value could reflect a high species richness (a large number of species, many of them rare) or could be attributed to many fewer but commoner species (for example, high relative abundance of many species). The simplest measure of diversity, the number of species in a given area, is called within-area diversity or, technically, alpha diversity. Ecologists generally
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call this measure species richness; they imply no economic value by using rich or its opposite, poor. Only their presence (not their abundance) is taken into consideration in counting the number of species in an area. Species counts are the most visible and most widely known measures of biological diversity. Tourists visit Costa Rica in part because its forests are so rich in bird species and its marine reefs are so rich in corals and fishes (see Costa Rica case study below). The biodiversity of the Camp Pendleton region in southern California includes 345 vertebrate species, a high level that constitutes a large percentage of all terrestrial vertebrates in that richly diverse state (chapter 1). The preeminence of the species as the central unit of biodiversity is explicit in the Convention on Biological Diversity (Heywood 1995; UNEP 1992) and the UN Environment Program's Explanatory Guide to the Convention (Glowka and others 1994). Although simple species-per-area statistics are useful, there are caveats: Species counts are rarely complete. Counts depend in complex ways on the area surveyed and how the survey was conducted. Counts of individual species might need to be weighted by their abundances, percentage covered, or mass. Surrogate measures of biodiversity, such as the numbers of genera (the taxonomic category directly above the species in the Linnean hierarchy) or even higher taxa (such as families), have been used. These can be effective when taxonomy accurately reflects underlying relationships and includes the descendants of a common ancestor, but systematists recommend that such approaches be treated with care and considered to be only interim stages in the development of a deeper understanding of biodiversity. If phylogenetic analyses are available, it can be useful to estimate the number of lineages present to take into consideration uneven species representation. For example, 20 species of lizards might represent only three main lineages in one area, but 15 in another. Such information might be used to identify a focus of active evolutionary diversification in the first case and the survival of ancient lineages in the second. Such tentative conclusions gain force if additional instances of coexisting taxa are found. Case Study: Costa Rica Costa Rica, a small country with high biodiversity, has been pioneering new conservation methods and creating a new biodiversity-conservation ethic by recognizing that its wildlands and biodiversity are among its most important economic assets. The nongovernment Instituto National de Biodiversidad (INBio) has been established to inventory the biodiversity in the country; identify new
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uses of biodiversity for education, industry, and science; and contribute information (and financial resources earned from the economic development of biodiversity) to the conservation of biodiversity in established conservation areas. For decades, Costa Rica has also been the focus of considerable ecological research that has helped to create the knowledge base for the conservation and use of biodiversity. The country's protected areas and wildlands contribute important economic benefits to the nation, including water and electricity, tourism, and scientific research. A growing economic contribution of wildlands involves trade in biochemical and genetic resources for use in the pharmaceutical and biotechnology industries (discussed further in chapter 3). Costa Rica has long been a source of material used for natural-product screening by industry, but samples were typically sold at relatively low prices with no provision for royalties. INBio is now involved in biodiversity prospecting in an effort to earn greater revenues that can be channeled back into conservation. A typical contract between INBio and a pharmaceutical firm might involve an agreement to provide 1,000 samples with a 3% royalty if a discovery is made. In view of the small likelihood of finding a commercial product, the time required for product development, and the lifetime of the patent on the final product, the net present value of such an agreement would amount to roughly $50,000–$500,000, comparable with the traditional spot-fee up-front payments for samples of $100–$200 each) (Reid and others 1993). Pearce and Moran (1994) calculate that the value of tropical forest land for medicinal plants ranges from $0.01 to $21 per hectare. Thus, although the use of land as a source of medicinal plants might not justify conservation in its own right, medicinal values can add to the overall stream of economic benefits from the protection of wildlands. The combined annual economic contribution of wildlands to Costa Rica's gross domestic product (GDP) for watershed protection, ecotourism, scientific research, and biodiversity prospecting is in the range of $87–200 million. Most of the contribution is attributable to tourism and protection of the electricity-generating capacity of the country. Costa Rica's wildlands might eventually contribute even more than this sum to the economy because of their carbon-sequestration potential. Under the Framework Convention on Climate Change, countries might eventually benefit economically from steps that they take to reduce rates of carbon emissions from deforestation. Costa Rica has already received "carbon offset" grants from US companies in return for actions to protect these areas from deforestation, and at $5–$10 per ton of carbon the value of the country's standing forest might be substantial. The current economic contribution of Costa Rica's wildlands can be compared with agriculture's contribution of $864 million in 1989 (GCR 1990) from twice the land area size of the country's protected areas. Even though the average contribution to GDP per unit area of agriculture exceeds that of wildlands, maintenance of wildlands is often the more economically attractive option for the specific soil and ecological conditions in
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much of the country. With growing evidence of the economic importance of wildlands and protected areas—and growing understanding that the most economically valuable use of many of these regions is as wildlands—the country is now seeking to enhance the economic benefits from the natural systems and to establish mechanisms whereby the benefits can serve as incentives for conservation. Species Counts Might Not Be Representative For many groups of species (such as nematode worms, mites, and aquatic fungi), what we know for certain is only that we do not know many, perhaps even most of the species. Consequently, we can measure species numbers for some but not all of the species in an area. Only for a few regions do we have even partial inventories of the species present. For example, an inventory of fungi, lichens, bryophytes, vascular plants, mollusks, arthropods, amphibians, mammals, fishes, and birds has been done for the Pacific Northwest, where controversy rages over the old-growth forests, but the effort is incomplete because of variation in our knowledge of different groups. We know all the birds and mammals, but our knowledge of insects and fungi is far from complete. Even that example is exceptional because of the large number of species and groups that were inventoried. Measures of species numbers are usually just counts of easily observed or identified species. Costa Rican forests are rich in birds, but whether the forests are relatively rich in other species—fungi, for example—is unknown. Areas rich in one group of species are often rich in another, but not always. Remote islands might have many bird species but few or no mammals or amphibians. Species Counts and the Area Counted Larger areas contain more species than smaller areas; the United States has more species (of everything) than does the state of Tennessee. However, the number of species does not increase in simple proportion to the area. An area, that is half the size of another area, might have 85% of the larger area's species. Consequently, we cannot use the number of species per unit area as a biodiversity measure without understanding the biological context. Thus, equally perhaps, when we conclude that Costa Rica or riparian habitats within western US rangelands are diverse (see the western rangelands case study in chapter 1), we mean that they are richer in species than other areas of similar size. Endemism and Diversity Across Space As we move across a region, the species composition might change greatly even though the species numbers might not. This change in species in a region is an important measure of diversity in its own right. We call the difference in
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composition between-area diversity, or beta diversity. The various ways of measuring such diversity all arise from this insight: two adjacent environments might both contain 10 species, but the number they share could range from 0 (when beta diversity is highest) to 10 (when beta diversity is lowest—zero). Endemics are the species that are prevalent in or peculiar to an area. The greater the fraction of endemics areas hold, the greater the between-area diversity as we cross its boundary. Endemism and between-area diversity are also related to the typical size of a species range. The smaller the typical range, the more quickly one moves from an area with one set of species to an area with another set. Areas differ greatly in endemism. All the forests of eastern North America hold 160 species of birds, and the tropical forests of Hawaii once held about 130 species. Hawaii's breeding species were fewer than eastern North America's but were all endemic and had small ranges. Fewer than 25 of eastern North America's birds are endemic (Pimm and Askins 1995). The distinctiveness of an area's flora and fauna leads to several concerns of managers: why some areas with few species contribute greatly to biodiversity, why endemic species contribute to much to biodiversity, and why some species in some places contribute nothing to regional or global biodiversity. An area's endemic species dominate discussions of protecting biodiversity because it is the loss of these species that causes a global loss of species diversity. Usually, endangered species are endemics with small ranges (Collar and others 1994; Pimm and others 1995). Few endangered species are rare over very large areas. However, many species with formerly wide geographical distributions—such as the grizzly bear, mountain lion, leopard, bald eagle, and peregrine falcon—have become endangered because of severe habitat loss, persecution, and widespread use of pesticides. Thus, concerns about biodiversity at Camp Pendleton (see case study in chapter 1) focus on the several species, such as the California gnatcatcher, now found only or almost only there. The concern about endemics means that there can be conflicts between measures of local versus regional diversity. For example, across much of the eastern United States, the fragmentation of once-continuous tracts of forest has led to a local increase in species via the invasion of widespread open-area species, such as cowbirds, bobwhite quail, and white-tailed deer. Forest managers and wildlife managers (Dasmann 1964; Giles 1978; Leopold 1933), once viewed the creation of openings in continuous forest as important for increasing game-species productivity and non-game-species diversity (biodiversity). Earlier editions of the wildlife-managers handbook published by the Wildlife Society (Giles 1969; Mosby 1960, 1963; Schemnitz 1980) considered the development of forest edges to be an important management tool. Those recommendations have been deleted in the latest edition (Bookhout 1994). A forest edge might have a higher number of species per unit area, but these are generally common and widespread species. The creation of forest edges and fragmentation by logging eliminates the continuous habitat required by forest interior species, such as ovenbirds, worm-eating
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warblers, and waterthrushes. Therefore, although edges can increase local diversity, it constitutes as a loss to regional and global biodiversity. It has long been known that birds could be classified as forest interior or edge species (for example, Kendeigh 1944), but it has only recently been appreciated that edges create biological and physical environments that can be detrimental to some species. Birds have been particularly well-studied in this context, and edge habitats have been shown to contribute to increased nest predation and cowbird brood parasitism (for example, Böhning-Gaese and others 1993; Robinson and others 1995; Terborgh 1989). Humanity has both deliberately and accidentally introduced species worldwide. Obviously, introductions of nonindigenous species—that is, plants, animals, and microorganisms in areas outside their natural geographical ranges (OTA 1993) can add nothing to global biodiversity. Replacing a region's endemic species with species that are more widespread can increase biodiversity locally, but it also reduces between-area diversity by homogenizing global flora and fauna. The mediterranean regions are an excellent example: introduced grasses and forbs can increase diversity at the local level, but they generally reduce biodiversity in western rangelands (see the case study in chapter 1). Introduced species can also be seriously harmful. Some introduced trees have reduced large areas of the Everglades nearly to single-species stands and have correspondingly endangered native species (see the Everglades case study in chapter 3). According to the 1993 report of the Office of Technology Assessment (p 5), approximately 15% of the nonindigenous species in the United States cause severe harm. High-impact species—such as the zebra mussel, gypsy moth, of leafy spurge (Euphorbia esula) (weed)—occur throughout the country. Almost every part of the United States confronts at least one highly damaging nonindigenous species today. They affect many national interests: agriculture, industry, human health, and the protection of natural areas. . . . Harmful nonindigenous species cost millions to perhaps billions of dollars annually. Introduced species, of course, can be beneficial. Very few of the foods we grow are endemic to the United States. Virtually all other crops are nonindigenous. Many introduced species—such as Kentucky bluegrass and wisteria—come to be perceived by some people as occurring naturally in a region. Some people believe that other nonindigenous species, such as the green crab in waters near Martha's Vineyard, detract from the integrity of the environment. Regardless of these examples of beneficial effects of introduced nonindigenous species, the adverse effects of nonindigenous species on endemics have resulted in their being one of the leading causes of global extinctions (Nott and others 1996; Pimm and others 1995). According to Norse (1993), the other causes are overexploitation, physical alteration of habitat (including habitat destruction and degradation), pollution, and global atmospheric changes.
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Landscapes as Biodiversity In many discussions, biodiversity refers to "a diversity of landscapes". We consider a region that contains both grassland and forest to be more diverse than one that contains only grassland. It is the mixture of ponderosa pine savannas, grasslands, wetlands, and riparian woodlands that gives Boulder, Colorado, its diverse environment (see the Boulder case study in chapter 3). Camp Pendleton has not only a large number of threatened and endangered species, but also a diversity of marine, estuarine, riparian, and terrestrial "habitat types" (see the Camp Pendleton case study in chapter 1). Costa Rica (see the case study this chapter) has many more life zones than an area of comparable size (for example, West Virginia) in eastern North America. Terms like grassland and forest denote different associations of species. Grassland and forest edge have high between-area diversity. Boundaries between associations often correspond to obvious physical differences in the environment and differences in ecological processes, such as nutrient cycling. The use of landscape terms to describe biodiversity raises three questions for managers to deal with: How do we classify landscapes? How do landscapes differ with respect to ecosystems and ecosystem processes? What linkages exist between ecosystem processes across diverse landscapes? The term association of species is deliberately vague. Ecologists apply it to areas (with the sets of species they contain) that range from a few square meters to continents. On the largest scale, we refer to tundra, coniferous forest, deciduous forest, grassland, savanna, desert, tropical rain forest and so on. Ecologists call these major regions biomes. On smaller scales, Noss and Peters (1995) classify and identify the endangered "ecosystems" of the United States. By ecosystem they mean distinct assemblages of plants and animals. For example, naming an ecosystem "Florida scrub" is a statement of the likelihood that we will find a set of plant and animals widely across this ecosystem. In addition, the species typically will be different from those in other ecosystems. On an even smaller scale, we have finer divisions of environments variously called habitats, associations, communities, and biotopes. When sufficient data are available, formal statistical procedures enable a manager to recognize a biome, landscape, ecosystem, habitat, biotope, or other ecological association. The procedures group smaller areas into larger divisions according to the principle that species are similar within and different between those divisions (Hengeveld 1990; Pielou 1975). For most of the cases, the recog-
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nition of divisions is informal, often with reference to an expert or general guide; such informality does not deny the utility of the divisions. In the 1880s, C.H. Merriam, one of the great natural historians of the West, characterized the biodiversity of northern Arizona and mapped it into seven ''life zones'' on the basis of altitudinal bands of temperature and the appearance of the vegetation. Merriam's classification preceded formal vegetation surveys and statistical analyses. Nonetheless, his classification retains its utility as a broad guide to where to find plants and animals and where the boundaries between their distributions will likely lie. On a much finer scale, the conspicuous zonation of intertidal rocky shores provided the initial motivation for studying near-shore marine ecosystems (Gilsen 1930). Grassland and forest clearly do more than refer to the similarity of species within and the differences between associations. A grassland, like any other environment, has its own typical set of ecological processes, and these might be different qualitatively and quantitatively from those in the nearby forest. The plants in grasslands, for example, might be adapted to frequent fires; indeed, without fires, trees might invade and forest take over. In contrast, the dominant ecological processes in a lake might be related to the nature of the nutrient effects inputs from surrounding areas. Sometimes the threats to biodiversity are the human impacts on natural ecosystem processes, such as changes in the hydrology and fire regimes of the Everglades (see the case study in chapter 3). Biodiversity on the landscape scale involves more than the mosaics that differ in composition (such as forest versus grassland). It also includes the connections and dynamics between and among patches and their implications for the functioning of ecosystems (Turner and Gardner 1991). Connections can occur through the flow of water, energy, materials, or organisms. For example, water moves through upland to riparian and wetland areas and then to streams, carrying with it dissolved nutrients. The accumulation of water in wetland or riparian areas leads to soil saturation, decomposition by anaerobic pathways, dominance by different plant and microbial species, and substantial effects on the chemistry of streams. Nitrogen fertilizer that is leached from upland agricultural systems can be taken up and retained by riparian plants or denitrified to nitrogen gas in soils (Hedin and others 1998; Peterjohn and Correll 1984, 1986). In either case, the maintenance of landscape diversity controls the overall exchange of nutrients between terrestrial and aquatic ecosystems. On a coarser scale, the seasonal movement of migratory birds between tropical and temperate ecosystems connects these otherwise independent biomes (see Everglades case study, chapter 3). This flow of organisms requires that managers in each region consider the dynamics of the other region in their analyses. By its very nature, biotic exchange over long distances implies a lack of independence.
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taxon of monogenean trematode parasites, and the mountain beaver is home to the world's largest flea, itself a long-branch taxon. The redwood and sequoia trees are sister species that form a long-branch taxon. They are distantly related to the dawn redwood of China, which is extinct in the wild but was preserved in Chinese monasteries and is itself a long-branch taxon. Biologically Based Ranking and Rating Methods Biologists assess the importance of conserving biodiversity in various ways. Some are based on conserving species, others on maintaining community or ecosystem functions. From the perspective of the field of biological systematics, species do not all have equal value when it comes to biodiversity maintenance and conservation. Several approaches have been used to assign such value. One can use a generalized hierarchical approach, working along a genealogical to phylogenetic continuum from genetically distinct sister populations to groups at various taxonomic levels. Populations of a species that vary geographically in degree of genetic distinctiveness would have greater value than populations of a species that are genetically more uniform. Similarly, with respect to a given protected species, a related species that is more distinct genetically would have greater value than one that is only slightly different. That kind of ranking can be used in a phylogenetic ranking of taxa; species that are phylogenetically increasingly remote would have increasing value because the goal is to maintain the greatest amount of biological diversity. The method can be made precise when sufficient information on relationships is available (Faith 1994). With such a scheme, long-branch taxa have the greatest value. That scheme can be combined with habitat, community, ecosystem, and geographic (bioregion) approaches. A habitat or region that has several long-branch taxa is more valuable for biodiversity maintenance than one that has none or only one. In contrast, one might choose to focus on a region that is relatively poor in long-branch taxa because many factors go into valuation, and pragmatic concerns or special interest in a species might force decision-making. When this happens, it is wise policy to identify a rationale underlying the decision. Another consideration in biodiversity maintenance is the geographic distribution of a species. In general, species that are widely distributed require less attention than species that are narrowly distributed, although that widely distributed species that have low population density might be of more concern than an endemic that is well protected and in good demographic condition. The components of biodiversity are hierarchical and intricately linked. For example, the genetic variability within a species is related to their continued adaptation and evolution in the face of biological, physical, and chemical changes. A variety of species in an ecosystem might increase productivity and stability. The pattern of ecosystems on the landscape influences energy flow, nutrient
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cycling, and population movements. The value of agricultural or forest productivity is undeniable, but its intrinsic relationship to soil microbial processes, hydrological and atmospheric cycles, pollinators, and pest predators is largely unknown and unappreciated by most sectors of society. Chapter 3 discusses the values of biological components in detail. Given that funds for conservation are limited, how should they best be allocated to ensure the most efficient conservation of biological diversity? That question confronts decision-makers in institutions as varied as government departments responsible for protected areas and nongovernment organizations, such as The Nature Conservancy. Typically, the answer involves setting priorities for habitat or ecosystem conservation; and this, in turn, requires assigning relative values to the areas under consideration for protection. Although ultimate decisions of which habitats or ecosystems will be protected might be influenced by considerations of the cost of protecting various sites or assessments of the likely threat to a site in the absence of protection, the initial ranking of sites should be based on biological criteria. No approach to priority-setting can serve all biodiversity-conservation objectives. For example, one logical goal of conservation would be to conserve both the greatest diversity of species and the greatest diversity of natural habitats. Consider two hypothetical ecosystems, one with 1,000 endemic species and one with 10. If sufficient money were available to protect two 1,000-hectare sites, where should they be. Locating both in the species-rich site would protect far more species but would sacrifice the protection of unique habitats. Placing one conservation site in each ecosystem would protect the diversity of ecosystems but with a tremendous loss of species diversity. There is no scientifically based means of comparing the value of a "unit" of habitat protection with a "unit" of species protection, so there is no single solution to the problem. Biological value is assessed with reference to five basic criteria: Richness, the number of species or habitats in a given area. A region with more species or habitats per unit area is given higher value than a region with fewer. Thus, tropical forests, with their high number of species, are often seen as having higher conservation priority than adjacent tropical dry forests which are slightly less rich in species. Endemism, the narrowness of the distribution of the species in an area. A region with many endemic species is given higher value than a region with fewer. Thus, Madagascar, some 80% of whose plant species are found nowhere else, has higher conservation priority than a region with a lower proportion of endemic species. Rarity of species or habitats in a region. A region with rare species or habitats is given higher value than a region with abundant ones. Thus, wetlands in arid regions are given higher value than wetlands in temperate regions. Ecosystem services, the importance of the natural habitat, or resident
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single species capable of influencing ecosystem function (see chapter 3) for various services of importance to humans. Thus, a forested watershed that is the source of public water is seen as having higher conservation value than one that is not. Protected status and representation, the relative protection of the species or ecosystem that already exists. Protection of an ecosystem that is not yet represented in a system of protected areas is given higher value than one that is. Some examples of the use of biological ranking methods are discussed below. Rare Species and Habitats The Nature Conservancy's method for ranking "elements of natural diversity" is the best-known example of a valuation approach that is based primarily on the rarity of and threat to species and biological communities. The conservancy obtains information about the known or estimated numbers of subpopulations, the estimated numbers of individuals, the narrowness of ranges and habitats, trends in population and habitat, threats, and fragility, and then it assigns a rank of 1–5 (with 1 representing extreme vulnerability) (Johnson 1995). It then focuses its habitat-acquisition efforts on areas that have more rare and imperiled species. In addition, a variety of quantitative tools permit a population's status or viability to be assessed or a habitat's ecological importance to be determined. Box 2-1 classifies and lists some of these techniques as a quick guide for a manager seeking widely available literature relevant to some local and pressing situation. Representative Biological Communities The 1982 World Conservation Union Bali Action Plan (McNeely and Miller 1984) called for the establishment of a worldwide network of national parks and protected areas covering all terrestrial biogeographic regions, and it set a target of protecting at least 10% of each bioregion. The union later conducted a series of systematic regional reviews to identify gaps in protected-area coverage, with emphasis on ensuring representative coverage of protected areas. Other international efforts, such as the UNESCO Man and the Biosphere Program, also have chosen to emphasize representative coverage of protected areas in their conservation priority-setting. By the late 1980s, about 15 of some 227 biogeographic provinces still had no protected areas, and 30 had five or fewer areas that encompassed less than 1,000 km2 (Reid and Miller 1989).
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BOX 2-1 Quantitative Tools to Assess Biological Importance Category Reference Population Status and Viability Analyses Minimum viable populations General Plants Animals Gilpen and Soulé 1986; Goodman 1987; Harris and others 1987; Soulé 1987 Ballou and others 1995; Boyce 1992; Ruggiero and others 1994; Shaffer 1980; Soulé 1987 Menges 1990, 1998; Schemske and others 1994 Groom and Pascual 1998; Lamberson and others 1992; Reed and others 1988 Landscape Design Issues Metapopulations Ecosystem fragmentation Habitat corridors (connectivity) Population sources and sinks Gilpin and Hanski 1991; Hastings and Harrison 1994; Hanski and Gilpin 1997; McCullough 1996; Tilman and Kareiva 1997 Andrén and Anglestam 1988; Delcourt and Delcourt 1992; Forman 1995; Harris 1984; Lynch and Whigham 1984; Murcia 1995; Saunders and others 1991; Schwartz 1997; Shafer 1990; Robinson and others 1995; Turner and Gardner 1991; Wahlberg and others 1996; Wilcox and Murphy 1985; Yahner and Scott 1988 Adams and Dove 1989; Beier and Noss 1998; Forman 1995; Forman and Gordon 1986; Hudson 1991; Mackintosh 1989; Simberloff and others 1992 Donovan and others 1995a, b; Howe and Davis 1991; Pulliam 1988, 1996; Pulliam and Danielson 1991; Trine 1998 Species Introductions Nonindigenous species Harmful species Brothers and Spingam 1992; Drake and others 1989; Mooney and Drake 1986; Parker and Reichard 1998; Ruesink and others 1995 OTA 1993 Reserve Locations Rare species and biodiversity hot spots Siting decisions Bedward and others 1992; Forey and others 1994; Gaston 1994; Groombridge 1992; Johnson 1995; Myers 1980, 1988, 1990; Prendergast and others 1993; Reid 1998; Wilson 1992 Andelman and others in press
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Species Richness and Endemism A number of priority-setting systems focus on the protection of areas that are particularly rich in species or that have many endemics. For example, Myers (1989) identified 10 "hot spots" that deserved conservation emphasis—tropical forest areas with high species richness and relatively high endemism that also faced exceptional degrees of threat from human activities. The list was expanded to include eight additional regions—four in the humid tropics and four in Mediterranean-type habitats (Myers 1990). Myers estimates the number of plant species in a region and the percentage of those species that are endemic, evaluates the threat of habitat loss for the region, and then ranks highest regions with large numbers of threatened endemics on relatively small areas. Birdlife International has followed a similar approach, identifying regions that have relatively high numbers of bird species with restricted ranges (less than 50,000 km2). In all, 221 "endemic bird areas" have been identified and are being emphasized as a focus for conservation action (Johnson 1995). The Maintenance of Biodiversity and Ecosystem Services The preceding sections have focused on the varied definitions of biodiversity, how it is related to landscape-scale patterns, its genetic basis, and its evolutionary origin and significance. Those descriptive accounts identify what biodiversity is; they do not address how it is maintained or influenced by specific interactions or what the role of species—individually or collectively—might be in ecosystem functioning. These are developed more fully in chapter 3. The role of ecological interactions in influencing whether species can coexist locally has been recognized at least since the time of Darwin (1859), who showed that a clipped grassy plot harbored more species than an undisturbed one. Since then, an extensive literature has developed the theme that various interactions can influence the genetic structure and morphological appearance of local populations (Tollrian and Harvell 1999), the probability of species coexistence (Paine 1969), and the biological structure and function of entire freshwater assemblages (Brooks and Dodson 1965; Carpenter and Kitchell 1993; Werner 1986). Probably all known taxa, ranging from pathogens to (especially) humans, are involved in this interactive natural world. The dynamic relationships and their immediate and long-term consequences obviously influence the determination and evaluation of species diversity patterns. The locally resident species also affect considerations of ecosystem function. For instance, these are increasingly factored into conservation priority-setting, particularly with regard to the protection of water quality. Many countries have forest policies that require the protection of forested buffers along rivers and streams to reduce siltation and protect the rivers from changes in water temperature. In some cases, protected areas have been estab-
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lished in watersheds specifically to protect the quality of the water supply of an urban area or to protect dams from siltation. The management implications of changes on local species composition and therefore probably richness and of its capacity to alter ecosystem function are developed in chapter 3 in the Everglades case study and the section Ecosystem Services, and in chapter 6 in the Lake Washington case study. Nature is complex and highly interactive: management decisions increasingly consider the totality of the biological matrix; no species lives in isolation, and changes in one are certain to affect the ecological and evolutionary continuity and the performance of others and of the assemblage in which they are imbedded. Summary Biodiversity includes not only the world's species with their unique evolutionary histories, but also genetic variability within and among populations of species and the distribution of species across local habitats, ecosystems, landscapes, and whole continents or oceans. Because biodiversity is such a broad concept, methods for its quantification are necessarily broad. In this chapter, we have attempted to define the components of biodiversity and to describe some of the ways to measure them. In the following chapters, case studies illustrate management decisions driven by various concepts of what biodiversity is or does. For instance, aesthetic considerations were influential in the preservation of open spaces in Boulder, CO (chapter 3), whereas water quality issues motivated the restoration of Lake Washington (chapter 6). The Everglades case study describes a major federal project in which biodiversity itself and habitat restoration were the primary considerations (chapter 3). Given such variation in mission, managers must consider both the maintenance of viable local populations of species of interest and the maintenance of biodiversity on larger scales, which is essential for the functioning of ecosystems. This chapter has addressed the many components of biodiversity that managers need to consider; the next chapter extends our understanding of how people value the components of biodiversity. Throughout the report, case studies illustrate management decisions that were based on the varied biodiversity components. References Adams LW, Dove LE. 1989. Wildlife reserves and corridors in the urban environment: a guide to ecological landscape planning and resource conservation. Columbia MD: National Inst Urban Wildlife. Andelman S, Fagan W, Davis F, Pressey RL. In press. Tools for conservation planning in an uncertain world. BioScience. Andrén H, Angelstam P. 1988. Elevated predation rates as an edge effect in habitat islands: experimental evidence. Ecol 69:544–7.
OCR for page 38
Antonovics J, Bradshaw AD. Turner RG. 1971. Heavy metal tolerance in plants. Adv Ecol Res 7:1–85. Avise JC. 1994. Molecular markers, natural history and evolution. New York NY: Chapman & Hall. 511 p. Avise JC. 1995. Mitochondrial DNA polymorphism and a connection between genetics and demography of relevance to conservation. Cons Biol 9:686–90. Böhning-Gaese K, Taper ML, Brown JH. 1993. Are declines in North American insectivorous songbirds due to causes on the breeding range? Cons Biol 7:76–86. Ballou JD, Gilpin M, Foose TJ (eds). 1995. Population management for survival and recovery: analytical methods and strategies in small population conservation. New York NY: Columbia Univ Pr. 375 p. Bedward M, Pressey RL, Keith DA. 1992. A new approach for selecting fully representative reserve networks: addressing efficiency, reserve design, and land suitability with an iterative analysis . Biol Cons 62:115–25. Beier P, Noss RF. 1998. Do habitat corridors provide connectivity? Cons Biol 12: 1241–52. Bookhout TA (ed). 1996. Research and management techniques for wildlife and habitats, 5th ed. Bethesda MD: The Wildlife Society. 740p. Boyce M. 1992. Population viability analysis. Ann Rev Ecol Syst 23:481–506. Brooks JL, Dodson SI. 1965. Predation, body size and composition of plankton. Science 150:28–35. Carpenter SR, Kitchell JF. 1993. The trophic cascade in lakes. Cambridge UK: Cambridge Univ Pr. Brothers TS, Spingarn A. 1992. Forest fragmentation and alien plant invasion of central Indiana old-growth forests. Cons Biol 6:91–100. Clausen J, Keck DD, Hiesey WM. 1958. Experimental studies on the nature of species. III. Carnegie Inst Washington Publ No 581. Collar NJ, Crosby MJ, Stattersfield AJ. 1994. Birds to watch 2. Cambridge UK: BirdLife Intl. Darwin C. 1859. The origin of species: by means of natural selection or the preservation of favored races in the struggle for life. London UK: J Murray. Dasmann, R.F. 1964. Wildlife biology. New York NY: J Wiley. Delcourt PA, Delcourt HR. 1992. Ecotone dynamics in space and time. In: Hansen AJ, di Castri F (eds). Landscape boundaries: consequences for biotic diversity and ecological flows. New York NY: Springer-Verlag. p 19–54. Dobzhansky T. 1970. Genetics of the evolutionary process. New York NY: Columbia Pr. Donovan TM, Lamberson RH, Kimber A, Thompson III FR, Faaborg J. 1995a. Modeling the effects of habitat fragmentation on source and sink demography of neotropical migrant birds. Cons Biol 9:1396–1407. Donovan TM, Thompson FR III, Faaborg J, Probst JR. 1995b. Reproductive success of migratory birds in habitat sources and sinks. Cons Biol 9:1380–95. Drake JA, Mooney HA, di Castri F, Groves RH, Kruger FJ, Rejmanek M, Williamson M (eds). 1989. Biological invasions: a global perspective. New York NY: J Wiley. Faith D. 1994. Phylogenetic pattern and the quantification of organismal biodiversity. Philos Trans Roy Soc London B 345:45–58. Forey PL, Humphries CJ, Vane-Wright RI (eds). 1994. Systematics and conservation evaluation. Oxford UK: Clarendon Pr. Forman RTT. 1995. Land mosaics: the ecology of landscapes and regions. New York NY: Cambridge Univ Pr. Forman RTT, Godron M. 1986. Landscape ecology. New York NY: J Wiley. Gaston KJ. 1994. Rarity. New York NY: Chapman & Hall. 205 p. GCR [Gobierno de Costa Rica]. 1990. Estudio Nacional de Biodiversidad. Ministerio de Recursos Naturales, Energia y Minas. San Jose, Costa Rica. Giles RH. 1969. Wildlife management techniques, 3rd ed. Washington DC: The Wildlife Soc. Giles RH. 1978. Wildlife management. San Francisco CA: WH Freeman.
OCR for page 39
Gilpin M, Hanski I (eds). 1991. Metapopulation dynamics: empirical and theoretical investigations. New York NY: Academic Pr. 336 p. Gilpin ME, Soulé ME. 1986. Minimum viable populations: processes of species extinction. In: Soulé ME (ed). Conservation biology: the science of scarcity and diversity. Sunderland MA: Sinauer. p 19–34. Gilsen T. 1930. Epibioses of the Gullmar Fiord I. A study in marine sociology, 1877–1927. Kristinberg Zool Sta 3:1–375. Glowka L, Burhenne-Gulmin F, Synge H. 1994. A guide to the Convention on Biological Diversity. Gland Switzerland: World Conservation Union. Goodman D. 1987. The demography of chance extinction. In: Soulé ME (ed). Viable populations for conservation. New York NY: Cambridge Univ Pr. p 11–34. Groom MJ, Pascual MA. 1998. The analysis of population persistence: an outlook on the practice of viability analysis. In: Fiedler PL, Kareiva PM (eds). Conservation biology, 2nd ed. New York NY: Chapman & Hall. p 4–27. Groombridge B (ed). 1992. Global biodiversity: status of the earth's living resources. Compiled by World Conservation Monitoring Centre. New York NY: Chapman & Hall. 585 p. Hanski IA, Gilpin ME (eds). 1997. Metapopulation biology: ecology, genetics, and evolution. New York NY: Academic Pr. 512 p. Harris LD. 1984. The fragmented forest: island biogeography theory and the preservation of biotic diversity. Chicago IL: Univ Chicago Pr. 211 p. Harris RB, Maguire LA, Shaffer ML. 1987. Sample sizes for minimum viable population estimation . Cons Biol 1:72–5. Hastings A, Harrison S. 1994. Metapopulation dynamics and genetics. Ann Rev Ecol Syst 25:167–88. Hedin LO, von Fischer JC, Ostrom NE, Kennedy BP, Brown MG, Robertson GP. 1998. Thermodynamic constraints on nitrogen transformations and other biogeochemical processes at soilstream interfaces. Ecology 79:684–703. Hengeveld R. 1990. Dynamic biogeography. Cambridge UK: Cambridge Univ Pr. Heywood VH. 1995. Global biodiversity assessment. New York NY: Columbia Univ Pr. 1140 p. Hillis DM, Moritz C, Mable BK. 1996. Molecular systematics, second edition. Sunderland MA: Sinauer. Howe RW, Davis GJ. 1991. The demographic significance of "sink" populations. Biol Cons 57:39–255. Hudson WE. 1991. Landscape linkages and biodiversity. Washington DC: Island Pr. Hutchinson GE. 1965. The ecological theater and the evolutionary play. New Haven CT: Yale Univ Pr. Johnson N. 1995. Biodiversity in the balance: approaches to setting geographic conservation priorities. Washington DC: Biodiversity Support Prog. 115 p. Kendeigh SC. 1944. Measurement of bird populations. Ecol Monog 14:67–106. Keystone Center. 1991. Final consensus report of the Keystone policy dialogue on biological diversity on Federal lands. Keystone CO: The Keystone Center. Koehn RK, Hilbish TJ. 1995. The adaptive importance of genetic variation. In: exploring evolutionary biology: readings from American Scientist. Slatkin M (ed). Sunderland MA: Sinauer. p 182–9 Lamberson R, McKelvey R, Noon BR, Voss C. 1992. A dynamic analysis of northern spotted owl viability in a fragmented forest landscape. Cons Biol 6:505–12. Leopold AS. 1933. Game management. New York: Charles Scribner. Lynch JF, Whigham DF. 1984. Effects of forest fragmentation on breeding bird communities in Maryland, USA. Biol Cons 28:287–324. Mackintosh G (ed). 1989. Preserving communities and corridors. Washington DC: Defenders of Wildlife.
OCR for page 40
Magurran AE. 1988. Ecological diversity and its measurement. Princeton NJ: Princeton Univ Pr. McCullough DR (ed). 1996. Metapopulations and wildlife conservation. Washington DC: Island Pr. 429 p. McNeely JA, Miller KR. 1984. National parks, conservation, and development. Washington DC: Smithsonian Inst Pr. McNeely JA, Miller KR, Reid WV, Mittermeier RA, Werner TB. 1990. Conserving the world's biological diversity. Available from: IUCN, World Resources Inst, Conservation International, World Wildlife Fund-US, World Bank. Menges ES. 1990. Population viability analysis for an endangered plant. Cons Biol 4:52–62. Menges ES. 1998. Evaluating extinction risks in plant populations. In: Fiedler PL, Kareiva PM (eds). Conservation biology, 2nd ed. New York NY: Chapman & Hall. p 49–65. Mooney HA, Drake JA (eds). 1986. Ecology of biological invasions of North America and Hawaii. New York NY: Springer-Verlag. Morrone JJ, Katinas L, Criscio JV. 1996. On temperate areas, basal clades and biodiversity conservation. Oryx 30(3):187–94 Moritz C. 1994. Defining ''evolutionarily significant units'' for conservation. Trends Ecol Evol 9:373–5. Mosby HS (ed). 1960. Manual of game investigational techniques. Bethesda MD: The Wildlife Soc. Mosby HS (ed). 1963. Wildlife investigational techniques, 2nd ed. Bethesda MD: The Wildlife Soc Murcia C. 1995. Edge effects in fragmented forests: implications for conservation. Trends Ecol Evol 10:58–62. Myers N. 1980. Conversion of tropical moist forests. Washington DC: National Res Coun, National Acad of Sci. 205 p. Myers N. 1988. Threatened biotas: "hotspots" in tropical forests. Environmentalist 8:187–208. Myers N. 1990. The biodiversity challenge: expanded "hotspots" analysis. Environmentalist 10:243–56. Norse EA. 1993. Global marine biological diversity. Washington, DC: Island Pr. Noss RF, Peters RL. 1995. Endangered ecosystems: a status report on America's vanishing habitat and wildlife. Washington DC: Defenders of Wildlife. 132 p. Noss RF, Cooperrider A. 1994. Saving nature's legacy: protecting and restoring biodiversity. Washington DC: Island Pr. Nott MP, Rogers E, Pimm S. 1995. Modern extinctions in the kilo-death range. Curr Biol 5(1):14–7 NRC [National Research Council]. 1996. Science and the endangered species act . Washington DC: National Acad Pr. Oldfield ML. 1984. The value of conserving genetic resources. Sunderland MA: Sinauer. OTA [Office of Technology Assessment]. 1993. Harmful nonindigenous species in the United States. US Congress, Office of Technology Assessment, OTA-F-565. Washington DC: GPO. Paine RT. 1969. A note on tropic complexity and community stability. Amer Nat 103:91–3. Parker IM, Reichard SH. 1998. Critical issues in invasion biology for conservation science. In: Fiedler PL, Kareiva PM (eds). Conservation biology, 2nd ed. New York NY: Chapman & Hall. p 283–305. Pearce DW, Moran D. 1994. The economic value of biological diversity. London UK: Earthscan. Peterjohn WT, Correll DL. 1984. Nutrient dynamics in an agricultural watershed: observations on the role of riparian forest. Ecology 65:1466–75. Peterjohn WT, Correll DL. 1986. The effect of riparian forest on the volume and chemical composition of baseflow in an agricultural watershed . In: Correll DL (ed). Watershed research perspectives. Washington DC: Smithsonian Inst Pr. p 244–62. Pielou EC. 1975. Ecological diversity. New York NY: J Wiley. Pimm SL, Askins RA. 1995. Forest losses predict bird extinction in eastern North America. Proc Natl Acad Sci USA 92:9343–7.
OCR for page 41
Pimm SL, Russell GJ, Gittleman JL, Brooks TM. 1995. The future of biodiversity. Science 269:347–50. Pitman III WC, Cande S, LaBrecque J, Pindell J. 1993. Fragmentation of Gondwana: the separation of Africa from South America. In: Goldblatt P (ed). Biological relationships between Africa and South America. New Haven CT: Yale Univ Pr. p 15–34. Potter CS, Cohen JI, Janczewski D (eds). 1993. Perspectives on biodiversity: case studies of genetic resource conservation and development. Washington DC: AAAS Pr. Prendergast JR, Quinn RM, Lawton JH, Eversham BC, Gibbons DW. 1993. Rare species, the coincidence of diversity hotspots and conservation strategies. Nature 365:335–7. Pulliam HR. 1988. Sources, sinks, and population regulation. Amer Nat 132:652–61. Pulliam HR. 1996. Sources and sinks: Empirical evidence and population consequences. In: Rhodes Jr OE, Chesser RK, Smith MH (eds). Population dynamics in ecological space and time. Chicago IL: Univ Chicago Pr. p 45–69. Pulliam HR, Danielson BJ. 1991. Sources, sinks, and habitat selection: A landscape perspective on population dynamics. Amer Nat 137:S50–66. Reed JM, Doerr PD, Walters JR. 1988. Minimum viable population size of the red-cockaded woodpecker. J Wildlife Mgmt 52:385–91. Reid WV. 1998. Reid WV, Laird SA, Meyer CA, Gamez R, Sittenfeld A, Janzen DH, Gollin MA, Juma C. 1993. Biodiversity prospecting: using genetic resources for sustainable development. Washington DC: World Resources Inst. Reid WV, Miller KR. 1989. Keeping options alive: the scientific basis for conserving biodiversity. Washington DC: World Resources Inst. Robinson SK, Thompson FR III, Donovan TM, Whitehead DR, Faaborg J. 1995. Regional forest fragmentation and the nesting success of migratory birds. Science 267:1987–90. Rudolph SG. 1990. Ancient forests as genetic reserves for forestry. In: Norse EA (ed). Ancient forests of the Pacific Northwest. Washington DC: Island Pr. p 129–32 Ruesink JL, Parker IM, Groom MJ, Kareiva PJ. 1995. Reducing the risks of nonindigenous species introductions. BioScience 45: 465–77. Ruggiero LF, Hayward GD, Squires JR. 1994. Viability analysis in biological evaluations: concepts of population viability analysis, biological population and scale. Cons Biol 8:364–72. Salthe SN. 1972. Evolutionary biology. New York NY: Holt, Rinehart and Winston. Salthe SN. 1985. Evolving hierarchical systems: their structure and representation. New York NY: Columbia Univ Pr. Saunders DA, Hobbs RJ, Margules CR. 1991. Biological consequences of ecosystem fragmentation: a review. Cons Biol 5:18–32. Schemnitz SD (ed). 1980. Wildlife management techniques manual, 4th ed. Washington DC: The Wildlife Soc. Schemske DW, Husband BC, Ruckelshaus MH, Goodwillie C, Parker IM, Bishop JG. 1994. Evaluating approaches to the conservation of rare and endangered plants. Ecology 75:584–606. Schwartz MW (ed). 1997. Conservation in highly fragmented landscapes. New York NY: Chapman & Hall. 436 p. Shafer CL. 1990. Nature reserves: island theory and conservation practice. Washington DC: Smithsonian Inst Pr. Simberloff DS, Farr JA, Cox J, Mehlman DW. 1992. Movement corridors: conservation bargains or poor investments? Cons Biol 6:493–504. Soulé ME (ed). 1987. Viable populations for conservation. New York NY: Cambridge Univ Pr. Terborgh, J. 1989. Where have all the birds gone? Princeton NJ: Princeton Univ Pr. Tilman D, Kareiva P (eds). 1997. Spatial ecology: the role of space in population dynamics and interspecific interactions. Princeton NJ: Princeton Univ Pr.
OCR for page 42
Tollrian R, Harvell CD. 1999. The ecology and evolution of inducible defenses. Princeton NJ: Princeton Univ Pr. Trine CL. 1998. Wood thrush population sinks and implications for the scale of regional conservation strategies. Cons Biol 12:576–85. Turner MG, Gardner RH (eds). 1991. Quantitative methods in landscape ecology. New York NY: Springer-Verlag. UNEP [UNEP]. 1992. Convention on biological diversity. Nairobi Kenya: UNEP. Wahlberg N, Moilanen A, Hanski I. 1996. Predicting the occurrence of endangered species in fragmented landscapes. Science 273:1536–8. Werner EE. 1986. Species interactions in freshwater fish communities. In: Diamond J, Case TJ (eds). Community ecology. New York NY: Harper and Row. p 344–58. Wiley EO. 1981. Phylogenetics: the theory and practice of phylogenetic systematics. New York NY: Wiley/Interscience. Wilcox BA, Murphy DD. 1985. Conservation strategy: the effects of fragmentation on extinction. Amer Nat 125:879–87. Wilson EO. 1992. The diversity of life. Cambridge MA: Harvard Univ Pr. 424 p. Wilson EO, Peter EM (eds). 1988. Biodiversity. Washington DC: National Acad Pr. WRI/IUCN/UNEP. 1992. Global biodiversity strategy: guidelines for action to save, study, and use earth's biotic wealth sustainability and equitably . Available from: World Resources Inst, IUCN, UNEP. Yahner RH, Scott DP. 1988. Effects of forest fragmentation on depredation of artificial nests. J Wildlife Mgmt 52:158–61.
Representative terms from entire chapter: