Convention on biological diversity
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II.Species diversitySpecies, or alpha diversity, is a measure of the number of species of all or various taxa per unit area, properly referred to as species richness. Another measure is gamma diversity, which reports the sum of alpha diversities across a broader region. Knowledge of the diversity of plant and animal species found in forests and their distribution is incomplete. While some groups, such as mammals and birds are often documented reasonably fully, other taxa such as invertebrates and microbes remain virtually unknown (FAO, 1999). The UNEP Global Biodiversity Assessment (UNEP, 1995) suggested that of the estimated 13.6 million species of organisms worldwide (including aquatic and oceanic species), less than 1.8 million have been described. The majority of species are insects (8 million), fungi (1.5 million) and bacteria (1 million) and the vast majority of these are undescribed. Even among terrestrial plants, many species are incompletely known in some regions. For example, only 30-40% of 15,000 botanical specimens held in the Royal Forest Herbarium could be identified during the Flora of Thailand Project (OEPP, 1997). For the poorly known groups of organisms, the estimated number of total species often entails considerable, rather tenuous extrapolation. For example, the commonly quoted estimate of global fungus diversity of 1.5 million species (Hawksworth, 1991) has relied on extrapolation from studies in temperate zones. However, a recent study of fungi growing on the tropical palm, Licuala, suggests that even 1.5 million may be a conservative estimate (Fröhlich and Hyde, 1999). Further, even when a species is identified, considerable work remains to determine its range, habitat requirements and relative priority for conservation. It is well known that primary tropical forests, particularly tropical wet forests with their high per unit area productivity, support much higher species richness than either temperate or boreal forests (UNEP 1995). Latitudinal gradients in species richness have often been reported (e.g., Stevens 1989, Ricketts et al., 1999). However, the fundamental processes that influence forest biological diversity are more or less common to all forest biomes. Species richness in forests is affected by several factors operating at local and regional scales, including long-term history (e.g., glacial events), ecological space available for niches, forest productivity (which is ultimately related to climate, soils and available water), extent of forest cover, forest landscape heterogeneity, species interactions and, for vertebrates, body size (Holling, 1992; Ritchie and Olff, 1999). At large scales, Thompson (2000) showed a direct linear relationship between bird, mammal and salamander/reptile species richness and net primary productivity in boreal and temperate forests in Canada. Helle and Niemi (1996) reported that numbers and species richness of birds increased with stand age-class in boreal forests, although the proportion of Neotropical migrant species declined with successional age in the Palaearctic region. Some longitudinal variation also exists in the species richness of mammalian herbivores of the boreal region (Danell et al., 1994). At local scales in primary forests, ecological niche space may be very high because of the complex vertical and horizontal structures of this forest type. Patches of forest have differing suitability as habitat for a given species depending on their size, presence of required structures, plant species composition, age and isolation from other patches; these factors influence patch quality. Some patches may be occupied by a given species for extended periods, while others may be unoccupied or occupied infrequently. Quality of a patch to a particular species may be affected by its isolation from other forests because the species is incapable of migrating across the hostile surrounding areas (Lovejoy et al., 1986). Hence, not all available habitat is necessarily occupied. For this reason, landscape structure and connectivity among forest patch size and continuity are extremely important in maintaining biological diversity (Hanski, 1999a, b). Species/unit area curves offer a quantified prediction that the larger the area of forest, to some upper size limit, the greater the number of species that can occupy the forest (Preston, 1962). The asymptote of the curve depends to a large extent on beta diversity of the landscape. A high beta diversity will result in high regional diversity, but with low local richness or alpha diversity (Cornell and Lawton, 1992). The relationship between species richness and forest area is valuable for predicting numbers of species in a sample area and possible changes as a result of forest management. The same relationship also predicts that most species in small patches will be generalists, and that specialist species, with their stricter habitat requirements, will primarily exist in self-sustaining populations in large continuous forest areas. This is particularly true of animal species with large body sizes and/or large home range requirements, such as most carnivores. However, as forest patches become smaller and the distance between them becomes larger with treeless areas in between, populations of smaller-bodied species, such as invertebrates, also become affected. Further, most species that have narrow habitat tolerances are area-sensitive. Thresholds exist in the reduction of continuous forest into patches beyond which species extinctions occur in a rapid and non-linear manner that is often unpredictable (Andren, 1994; Haila, 1999; Hanski, 2000). Andren (1994) indicated that when forests are reduced to 10-30% of the original cover, species’ declines become abruptly non-linear. Keystone speciesKeystone species are those whose removal from an ecosystem would have a disproportionately large impact on processes in that system (Power et al., 1996). Both plants and animals can be keystone species. For example, in Peru, twelve species of figs and palms maintain all frugivorous animals for three months each year (Terbourgh, 1986). There have, indeed, been instances of frugivores declining in association with loss of their food trees in logged tropical forests (Frumhoff, 1995). It is the degree of interaction, or the linkages, between a particular keystone species and other species within the system that is important. Keystone species do not always occupy an elevated trophic status (Power et al., 1996), and it is likely that most keystone species have not been identified as such because many are soil invertebrates, pollinators, mutualists or pathogenic organisms (Krebs, 1985; Power et al., 1996). Jones et al. (1994) used the term ‘ecosystem engineers’ (or keystone modifiers, after Mills et al. [1993]) to refer to species that create, modify or maintain certain environments. These are keystone species because they play important functional roles in ecosystems by creating habitats required by other species. Therefore, the removal of the engineering species from an environment results in species impoverishment beyond loss of the species itself. In North America, beavers (Castor canadensis) are a keystone species because they modify several hectares of forest by creating impoundments through damming streams. These impoundments are then used by many aquatic and semi-aquatic species. Ants may be other keystone species as they are responsible for seed and spore dispersal of certain plants (including many fungi), soil mixing and aeration and decomposition of wood. Ants can also have specific competitive effects on the distribution of other ant species and ground-dwelling arthropods, certain canopy arthropods and some vertebrates, including woodpeckers and ant eaters in tropical forest systems (Puntilla et al., 1994; Elmes, 1992; Holldobler and Wilson, 1990). Some woodpecker species may also be considered keystone modifiers because they excavate cavities in large trees that are subsequently essential for successful reproduction of many other species. If the habitat requirements of woodpeckers are provided through careful planning, then habitat will be created for secondary cavity-using species including forest bats, cavity-nesting passerines, cavity-nesting owls, certain wasp species and tree squirrels (Thompson and Angelstam, 1999). Species that concentrate their populations annuallyAn important aspect of the life history of many animal species is their movement to restricted areas that meet their needs for some portion of the year. During periods when a species congregates at high density in localized areas, a large proportion of local populations, or even the entire population, is particularly susceptible to disturbance or habitat loss. In forest ecosystems, there are numerous examples of species that move into localized high-density areas, often in winter. For instance, many species of amphibians congregate to breed in vernal pools in forests or forests margins. An example of a migratory species that concentrates is the monarch butterfly (Danaus plexippus) of North America (Malcolm and Zalucki, 1993), which breeds over much of the United States and Canada and overwinters in small areas in Mexico. A component of ecosystem management is therefore to safeguard the important or key habitats associated with species that concentrate and such species are arguably a special case for protection to maintain biological diversity (Thompson and Angelstam, 1999). Endangered speciesAcross all biomes, the 2000 IUCN Red List of Threatened Species lists over 11,000 species threatened8 with extinction (although this is based on a complete sample of only mammals, birds and coniferous trees, and very incomplete assessments of other taxonomic groups). While this is less than one percent of the world’s described species, it includes 24% of all mammal species and 12% of all bird species. While plants have not yet been evaluated systematically, in an earlier work Walter and Gillett (1998) found that at least 12.5% of flowering plants were threatened. The 2000 Red List also estimates that 25% of reptiles, 20% of amphibians and 30% of fish species are threatened. It has been estimated that current species extinction rates are between 100 and 1,000 times higher than the natural background rate and before humans began to cause extinctions (Pimm et al., 1995; Lawton and May, 1995). It is difficult to provide comprehensive estimates of how many forest species are threatened since there is no globally accepted habitat classification system. More specific data are available for individual tree species and some forest habitat types. For example, 900 threatened bird species rely on tropical rainforests, with 42% of these occurring almost exclusively in lowland rain forest and 35% occurring in montane rain forest. Among mammals, 33% of those threatened occur in lowland rain forest and 22% in montane forest, though it is not clear what proportion of these mammals are totally dependant on forests for their survival (Hilton-Taylor, 2000). The World List of Threatened Trees (Oldfield et al., 1998) documents over 7,300 species facing extinction (Table 5). Based on gross estimates of the total number of tree species in the world, this represents about 9% of the world’s tree flora. Although not specifically targeting endangered or threatened tree species because of their endangered condition, the list of priority species established by the FAO Panel of Experts on Forest Gene Resources (FAO, 2000c) also suggests that around 9% (48 species out of 524 tree species recorded) of the world’s most important trees are threatened, at species or population level. Table 5: Globally Threatened Tree Species9
Species and communities as indicators of forest change and habitat lossIndicator species are generally readily recognizable and are usually discrete biological entities whose predicament and/or conservation needs easily capture the imagination of the general public and government agencies alike. It is worth noting that tree species diversity (provided that this can be assessed) may prove to be an adequate surrogate for overall species diversity in most forest ecosystems. Trees as indicators may be particularly relevant for those forest ecosystems where many animal and fungal species are obligate associates of particular tree species (e.g. Campbell, 1989). Comparatively good data are available on the distribution of many tree species (Anon., 1996; Van Bueren and Duivendoorn, 1996). Indicators may be used to suggest the effects of change within a system at particular scales, or to indicate population trends that result from altered ecological processes. Most often indicators are chosen as ‘umbrella species’, defined as indicator species that correlate strongly with the presence of other species, (i.e. their presence indicates, with very high probability, the presence of several to many other species). All umbrella species are indicators, but not all indicators are necessarily umbrella species. The use of indicator species monitoring is an integral component of a programme of ecosystem management. In similar forest patches, not all subsets of species are necessarily the same, and across landscapes, habitats are not uniform. These variables can lead to a wide disparity in species richness measures across samples from a landscape. Community assemblage may depend to a large degree on history and stochasticity, as well as availability of suitable habitats, environmental gradients and the supply of species through immigration. Numerous processes affect a species’ capacity to occupy a forest stand, including competition for resources and demographic factors. In studying the effects of change in forested systems, the concept of alternative states in the same forest types must be considered. In other words, the presence or absence of a species in an area may be simply a manifestation of the same community in a different state, with no change in its functional capacity. Further, under the metapopulation theory, at any given time a patch of habitat may or may not be occupied by a certain species (Levins, 1992). Therefore, from a forest management perspective, the lack of a particular species in an area is not necessarily cause for concern under a properly replicated census design. On the other hand, absence of that particular species may indeed suggest serious problems. Efforts to map species diversity have been limited, mainly to regional levels, and have not usually included a separate analysis of forest species. Because of the difficulty and vast resources needed to make an inventory of the total number of species in an area, certain better-known genera or families are often used as surrogates for the presence of other species or to indicate functioning ecosystems. However, species may respond individually to change, as predicted by the unique niche hypothesis, suggesting that changes in numbers of one species do not necessarily correlate to changes in others [Reference needed]. Care must, therefore, be taken in the selection of indicator species, as well as in interpretation of census results. The characteristics of useful biological diversity indicator species are discussed in Caro and O’Doherty (1998). Hunter (1990) suggested that indicators were species that should be monitored to reveal the presence/absence of certain specific conditions or structures (fine filters) that were missed with coarse filter (or forest type) management. Indicators may be employed to assess change in a long list of forest aspects including processes and structures from sites to landscapes, for example, landscape heterogeneity, or site diversity. There is a substantial body of literature on community responses as indicators of pollution effects (e.g., Cairns, 1985). If forest processes have been altered, then the structure of animal communities will likely also have been affected (Holling, 1992). It often makes sense to examine species assemblages as indicators, rather than try to monitor individual species, in part because of a narrow focus in following individual species and also because of the general lack of autecological information (Kremen, 1992; Dufrêne and Legendre, 1997). For many taxa of animals or plants, all (or most) species in an area may be sampled simultaneously. Discarding the broader data set in favour of a few chosen species is not cost-effective when the tools are available to use all the data. For example, theoretically all songbirds can be recorded in singing male point counts, all salamanders may be equally vulnerable to pitfall trapping or shelter-board surveys, and most small mammals species are readily captured in traps. Therefore, many of these data sets lend themselves to community analyses. When choosing the communities or species assemblages to monitor, their sensitivity to forest change at multiple scales relevant to the problem should be considered. Westoby (1998) noted that it is important to choose the appropriate scale for assessing local species richness and, because the analysis is statistical, an adequate number of points with which to perform regressions must be obtained. Ordination of communities can aid in decisions about what communities might perform well as indicators, or the properties of a particular group of species as indicators at a given scale. The pattern of co-variation over time among species within a community may indicate impacts of effects. Population viability analysisThe presence of a species does not necessarily indicate that the population is sufficiently large to ensure long-term survival of the species. Further, even large population size does not guarantee population survival (Mangel and Tier, 1994), hence populations of species in more than one location are important to long-term persistence. Any monitoring programme for rare species should be concerned with population viability analysis (PVA) and also modelling to determine minimum viable populations (MVP) for key or rare species. The outputs from such modelling suggest critical population sizes needed to maintain genetic diversity. PVA modelling considers three main points: genetic factors, population demography and ecological factors (Boyce, 1992). Among the important ecological factors are the occurrence of ecological thresholds such as those caused by fragmentation and Allee effects (inability to locate mates) due to habitat loss and change in habitat structure. More recently, Hanski et al. (1996) have developed the term minimum viable metapopulation, to describe the common situation of limited emigration from source areas, such as protected areas, to surrounding habitats. They have suggested that small numbers of a local population have a high probability of extinction due to stochastic events. However, such modelling requires certain minimal data, which are most often not available, including means and variances of vital population demographic statistics and how these may differ in various habitat patches, dispersal rates, dispersal distances and success (Doak and Mills, 1994). Thus, unless considerable research is undertaken, a detailed understanding of the likelihood of survivorship in a population will not be obtained and yet this may be crucial to management decisions. Selection of elements for an inventory and to monitor, and monitoring methodsAn inventory is a compilation of data, often with maps, on the aspects of biological diversity chosen to monitor, while monitoring refers to the collection of data over time. Forest biological diversity is an issue of scale. Therefore, compiling an inventory and monitoring of biological diversity is a hierarchical procedure, requiring various techniques and technologies. Depending on technical capacity and state of knowledge, aspects to monitor, ranked from least to most demanding, include: forest types (area), forest ecosystems (area, age classes), landscape structure (patch sizes, associated selected variables), species (indicators, endangered species, key species, useful species) and genetic diversity. A starting point for any monitoring programme is an understanding of the distribution of the element to be monitored. A serious limitation to detecting change in global forest conditions is the lack of capacity in many countries to perform even large-scale monitoring of forest types and landscape structure and, more generally, an incomplete knowledge of forest species and forest ecosystems. Classification systems for each level of biological diversity must exist as a pre-condition to compilation of an inventory and monitoring |