Of the three types of human impacts on forests mentioned above, the only one for which a consistent effort at assessment has been made at regional or broader scales is the change in forest area. Since the 1980s FAO has provided statistical data on forest cover (FAO 1989, 1995), based on forest inventory, national reporting and high resolution remote sensing. Since the mid-1990s a geographic overview of the current distribution of forest cover has been available (WCMC 1996, 1997). Thus it is possible to provide statistical and mapped data about the amount of forest cover remaining, and to a lesser degree, about the amounts of particular types of forests and trends in the amount of cover. The greater the level of detail about forest types available within such data sets, the more relevant to issues of the conservation of forest biodiversity they will become.
The changes in forest configuration that accompany changes in land use and forest area can have substantial effects on the capacity of forest ecosystems to maintain their original biodiversity. As forest ecosystems are divided into smaller patches, there are numerous effects on their biota, and the responses may vary substantially among species and among forest types. The extensive literature on the effects of forest fragmentation suggests that the effects can be broken down into three major groups: area effects, edge effects and isolation effects. What follows is a brief summary of characteristic components of these effects.
When large forest blocks are broken into smaller ones, not all species are included in the remaining patches, simply because of sampling effects (Wilcox 1980). This is especially true for rare species and for non-mobile organisms, such as trees and many invertebrates, which may be sparsely or patchily distributed within the forest. Large animals and top carnivores are well known to require large areas of habitat. These species are especially vulnerable to the reduction in habitat area caused by forest fragmentation, and they may disappear entirely from forest patches because food or other resources are inadequate to support them (Rylands & Keroughlian 1988, Soulé et al., 1979, Schaller & Cranshaw 1980, Newmark 1987, Laurance et al.,. 1997b). Even smaller species are affected by the size of forest patches; amphibian species richness increased logarithmically with patch size in forest remnants in Madagascar (Vallan 2000). The disappearance of some species from forest fragments can profoundly affect the forest itself, as shown by the effects on tree communities of the disappearance of seed-eating rodents from forest islands in Gatun Lake in Panama (Putz et al., 1990). Other species persist, but in smaller populations, which may encompass less genetic diversity and lead over time to the vulnerability of those species to other ecological changes such as disease. Rare species and those that normally occur at low population densities are especially vulnerable to these effects (Laurance et al., 1997b). Smaller forest patches may also include less environmental variability and therefore fewer microhabitats than more extensive forest areas. This can contribute to the loss of individual species and may cause a reduction in the total species richness per area of forest (Scariot 1999). Fragmentation of forest cover may also alter the nature and proportional impact of natural disturbance regimes and regeneration processes (Laurance et al.,. 1997a, Viana et al.,. 1997).
Another important effect of forest fragmentation results from the creation of interfaces with non-forest environments. These interfaces are associated with environmental gradients resulting from the exposure of the forest edge to drying winds and increased solar radiation (Kapos 1989, Camargo & Kapos 1995, Kapos et al.,. 1997). The physical gradients affect ecological processes including canopy gap formation (Kapos et al.,. 1997), biomass and nutrient cycling (Laurance et al.,. 1997a, 1998a,b, Sizer et al.,. 2000), regeneration (Benitez-Malvido 1998, Sizer & Tanner 1999) and predation (Keyser et al.,. 1998) that can profoundly affect native species. Invasive species, both native and non-native, are often favoured by an increased incidence of forest edges within the landscape, so that substantive changes in species composition have been documented in forest fragments (Brown & Hutchings 1997, Laurance et al.,. 1997b, Lynam 1997, Malcolm 1997, Viana et al.,. 1997). Although some ‘edge effects’ have historically been regarded positively, principally because many game species make use of forest edges, they are generally regarded as detrimental to most native forest species. The magnitude of edge effects within forest fragments can be strongly affected by the land-cover characteristics in the surrounding landscape, the matrix harshness (Murcia 1995, Laurance et al.,. 1997b), and they are also dynamic, frequently increasing in magnitude and extent over time (Gascon et al.,. 2000). Connections among habitat fragments are an important means of reducing genetic isolation and providing additional foraging and refuge areas (Saunders et al.,. 1991)
The other major group of effects of forest fragmentation results from the separation of the forest fragments from each other and from larger blocks of forest. This isolation reduces the movement of species that are reluctant or unable to cross non-forest areas and for those that depend on such species for dispersal. Reduced movement and dispersal also increases the chance of local extinction of individual species as a supply of colonisers or propagules is lacking. Isolation of fragments may also reduce the genetic neighbourhood of some trees, reducing the breadth of the local gene pool for cross fertilisation (Nason et al.,. 1997).
Responses to all of these effects vary among species, but a body of empirical evidence is accumulating that facilitates predictions about the likely effects of fragmentation on any particular forest ecosystem. Also from such evidence, a series of empirical generalisations concerning the spatial configuration of habitat with respect to biodiversity preservation can be paraphrased as follows (Noss & Cooperrider 1994; Thomas et al., 1990).
i. Habitat that is more widely distributed across its original range is more likely to persist than habitat confined to small parts.
ii. Large blocks of habitat are superior to small blocks of habitat.
iii. Blocks of habitat close together are better than blocks far apart.
iv. Habitat in contiguous blocks is better than fragmented habitat.
v. Interconnected blocks of habitat are better than isolated blocks.
The long-term maintenance of forest integrity depends on promoting these characteristics in landscapes. The more that landscapes retain these characteristics, the less is their vulnerability to human-induced change. These generalisations provide a useful basis for assessment and communication of information about forest status in this respect.
Forest structure and composition and their implications for biodiversity are difficult to evaluate at broad geographic scales and may vary widely depending on (among other factors) the kinds and intensity of human activity and local ecological conditions. Therefore, assessing the amount of human activity may be a useful proxy for evaluating its impacts on biodiversity. It is well-documented, for example, that logging increases the probability of recurrent fire in Amazonian rain forests (Uhl et al.,. 1991, Nepstad et al., 1998), and that this in turn will lead to long term changes in species composition (Cochrane & Schulze 1998, 1999). Logging also affects animal community composition and ecological relationships (e.g. Johns 1996, Lambert 1992, Ochoa 2000). Hunting activity near settlements substantially reduces the abundance of mammal species (Muchaal & Ngandjui 1999) and the construction of roads facilitates both logging and hunting as well as land conversion and colonisation.