There are those drivers that result from changes in atmospheric properties (e.g., CO2 enrichment, N deposition, climate change), those that arise from direct manipulation of land (e.g., land use change, intensification), and those that involve shifts in organism presence or abundance (e.g., extinctions, invasions, outbreaks).
Of the human-induced changes in atmospheric composition, the enhancement of CO2 concentration is arguably the most pervasive. Given the high levels of CO2 in the soil, atmospheric CO2 enrichment is unlikely to affect the soil biota directly, and effects of enrichment on the decomposer community are likely in the first instance to be plant driven. This can occur at the level of both the individual plant and the plant community. At the whole-plant level, CO2 enrichment can promote soil organisms through increasing resource quantity, for example, by promoting NPP (Korner & Arnone 1992) and rhizodeposition (Paterson et al. 1996). Further, CO2 enrichment can alter litter quality, frequently enhancing litter C:N ratios, although both positive and negative effects of CO2 enrichment on decomposability of litters have been reported (Franck et al. 1997; Couteaux et al. 1999). Elevation of CO2 concentrations can also alter the role that soil fauna play in decomposition, since these organisms are more important in catalyzing breakdown of litter of poorer quality (Couteaux et al. 1991). At larger spatial scales, CO2 enrichment is also likely to alter plant community composition, frequently favoring plant species or functional types with faster growth and higher litter quality over slower growing plants (Collatz et al. 1998; Herbert et al. 1999). This is, in general, likely to favor decomposer activity.
Effects of CO2 enrichment on the composition of the plant community directly alters the community of root pathogens and root herbivores, due to changed plant-soil feedback (Bever 1994). Trade-offs between plant growth rate and plant defense against herbivores and pathogens (van der Meijden et al. 1988) result in more specialist root herbivores and root pathogens following CO2 enrichment. However, effects of root herbivores and pathogens are also influenced by factors other than resource availability and resource quality—such as the recognition of plant roots by pathogens, herbivores, and their natural antagonists—and by the ability of plants to culture and enumerate the antagonists of their enemies (van der Putten 2003). Litter with a high C:N ratio may not favor microbial decomposition, but it could still enhance specific microbial antagonists (Hoitink & Boehm 1999). Therefore, exposure of plant communities to CO2 enrichment may lead to changed root pathogen and herbivore communities, reduced or more specialized activity of these root organisms, and to changed host recognition or altered exposure to the natural enemies of the herbivores and pathogens belowground.
Atmospheric N deposition also has positive effects on those soil biota that depend on plant derived resource quality through promoting NPP, and unlike CO2 enrichment it often causes plants to produce litter with lower C:N ratios. Further, N deposition can have important belowground effects by altering the composition of the plant community, usually by favoring plant species that are adapted to fertile situations and that produce high-quality litter (Aerts & Berendse 1988). However, direct effects of N deposition on soil microbes can either promote or inhibit soil processes; chronic N deposition can either promote litter decay through enhancing microbial cellulolytic activity, or suppress it by inhibiting ligninolytic enzyme activity (Carreiro et al. 2000). Effects of N
deposition on the root herbivore and pathogen community will be mediated primarily through the plant community, but changes in the physico-chemical soil properties may also exert direct effects on these soil organisms.
Global warming and associated climate change events resulting from atmospheric CO2 enrichment can influence soil organisms directly, although the main effects of climate change on soil biota are again likely to be indirect and driven by plant responses. At the whole-plant level, increased temperature promotes the kinetics of nutrient mineralization, plant nutrient uptake, and NPP (Nadelhoffer 1992) that should in turn promote decomposer organisms. At the level of the plant community, elevated temperature has the capacity to promote plant functional types with either superior (Pastor & Post 1988; Starfield & Chapin 1996) or poorer litter quality (Hattersley 1983; Harte & Shaw 1995). In this light, the implications of climate-driven vegetation change for the decomposer subsystem are likely to be context specific. Global warming may also disrupt natural communities because of different dispersal and migration capacities of individual species (Warren et al. 2001). Migration of plant species without their natural root pathogens and herbivores may lead to enhanced abundance in the new territories, due to the escape from specific natural enemies and the presence of relatively aspecific mutualistic symbionts, for example, mycorrhizal fungi (Klironomos 2002).
Direct use of land by humans for production of food and fiber arguably has the greatest impact on terrestrial ecosystems of all global change phenomena. At the broad scale, forest, grassland, and arable systems differ tremendously in their functional composition of vegetation as well as disturbance regimes. This in turn has important implications for the composition, abundances, and activities of the soil organisms present; generally cropping systems contain lower levels of many components of soil fauna, microbial biomass, and organic matter than do comparable areas under forest or grassland (Lavelle 1994; Wardle 2002). Conversion of land to agriculture often has adverse effects on the performance of the decomposer subsystem, and artificial inputs are therefore required to substitute for services provided by soil biota. Agricultural intensification also influences the decomposer subsystem (De Ruiter et al. 1993; Giller et al. 1997). For example, agricultural tillage favored bacterial-based energy channels over fungal-based channels (Hendrix et al. 1986), promoted small-bodied soil animals relative to large-bodied ones (Wardle 1995), and altered the relative contribution of different subsets of the soil biota to litter decomposition (Beare et al. 1992). Comparable effects appear to result from intensification of forest management (Blair & Crossley 1988; Sohlenius 1996).
Another major land use change is extensification, or even the complete abandonment of production on arable land and grassland to conserve, or restore, former biodiversity (van der Putten et al. 2000). However, there has been little work on the restoration of diversity belowground. Soil communities respond more slowly to changes imposed by land abandonment than communities above ground (Korthals et al. 2001), and responses are often idiosyncratic (Hedlund et al. 2003). Therefore, besides dispersal limitations of many organisms to restoration areas (Bakker & Berendse 1999), slow development rates of the belowground community may be an important controlling factor of ecosystem services and goods provided by these newly developing restored ecosystems.
Alien species are most likely to alter community and ecosystem properties when they show large functional differences to the native species of the community being invaded. The functional attributes of most alien species do not differ greatly from those of native biota (Thompson 1995), but differences, when they do exist, can cause profound implications for both the aboveground and belowground components of the ecosystem. For example, invasion of the N-fixing shrub Myrica faya into Hawaiian montane forests lacking nitrogen fixing plants has important effects on ecosystem N inputs (Vitousek & Walker 1989). Introduction of deer and goats into New Zealand rainforests, which lack native browsing mammals, has caused large shifts in the soil food web composition and diversity, and in ecosystem C sequestration (Wardle et al. 2001). Invasion of European earthworms into those North American forests that lack a native earthworm fauna has been shown to alter soil microbes, fauna, and supply of plant-available nutrients from the soil (Hendrix & Bohlen 2002). Human-induced extinctions of organisms may also be of functional importance, but only in instances in which the lost species plays an important functional role. Historical examples include probable large-scale alteration of soil processes following vegetation change caused by extinctions of dominant megaherbivore species, for example in Siberia (Zimov et al. 1995) and Australia (Flannery 1994).
Soil organisms and processes are capable of showing a variety of responses to drivers of global change, and the nature of these responses is likely to be context specific. Different global change drivers do not operate independently of one another; a given ecosystem is likely to be affected by several drivers operating simultaneously. Interactions between different global change drivers (e.g., between CO2 enrichment and N deposition [Lloyd 1999]or between invasive plants and CO2 enrichment [Smith et al. 2000]) may have important, though largely unrealized, implications for the decomposer subsystem and for the interactions between plants, root pathogens, and symbiotic mutu-alists (Richardson et al. 2000). This will in turn affect those ecosystem services driven by soil organisms. Importantly, soil biota do not function in isolation, and ecosystems are driven by feedbacks between the aboveground and belowground biota (van der Putten et al. 2001; Wardle 2002; Bardgett & Wardle 2003).
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