Soil biodiversity controlled by stress and perturbation
It is well supported that the local species diversity of a given group of organisms is mainly driven by stress (resource availability, temperature, pH, etc.) and perturbations (fires, grazing, cultivation, etc.) (Grime, 1973; Huston, 1994; Huston & Deangelis, 1994; Rosenzweig, 1995; Wardle, 2002). Along gradients of stress or perturbation, species richness theoretically follows a humpbacked pattern of variation. In ecosystems characterized by harsh environmental conditions (strongly stressed or perturbed), a reduction in diversity may occur because only a small number of species may have the requisite traits to survive under such constraints (Wardle, 2002). Lower levels of diversity under more productive or less disturbed conditions may also occur as a result of a higher intensity of interspecific competition (Grime, 1973; Huston, 1994; Huston & Deangelis, 1994), a decrease in the heterogeneity of resource spatial distribution (Tilman, 1982) or as a consequence of mechanisms not related to biotic interactions (Abrams, 1995).
While this peaked response of biodiversity to productivity/disturbance gradients has long been presented as ubiquitous in nature (Huston, 1994; Rosenzweig, 1995), a cursory examination of the available literature reveals that it may actually not be as prevalent as commonly believed (Mackey & Currie, 2000, 2001; Mittelbach et al., 2001). Rather, positive monotonic, negative monotonic, and unimodal relationships all appear, and non-significant relationships are also described in a large number of studies. This suggests that the commonly accepted humpbacked diversity–productivity/disturbance relationship may not be the rule (Mackey & Currie, 2001; Mittelbach et al., 2001). In fact, the shape of the biodiversity response to adversity gradients may vary according to, for example, the importance of competition in driving community assembly, the nature of disturbances (e.g. frequency, intensity, predictability, natural versus anthropogenic), the dispersal capacity of the focal taxon, the size of the species pool adapted to highly productive habitats, the efficiency of sampling in detecting rare species or the scale at which the study is conducted (Mackey & Currie, 2000, 2001; Mittelbach et al., 2001; Pärtel et al., 2007).
The results currently available for soil biota are rather inconsistent (Bardgett et al., 2005b; Wardle, 2006). As an example, the diversity of microbial communities (Derry et al., 1999; Degens et al., 2000) (Fig. 9a), chilopods or woodlice (Paoletti, 1988; cited in Wardle, 2006) seems to increase monotonically along a gradient of resource availability. In a synthesis of results from nine different studies, Wardle (2006) found little if any support for decomposer diversity decreasing at the most favourable extremity of stress gradients. He concluded that many soil taxa are not strongly regulated by competition, and that increasing resource availability is not likely to result in competitive exclusion. According to Abrams (1995), such a pattern may also reflect other mechanisms, such as an increase with resource availability in the abundance of rare species, the availability of rare resources, the occurrence of combinations of resources or conditions necessary for specialist species, or in the overall resource volume that allow the coexistence of a greater number of species. Additionally, the strong heterogeneity of the soil environment and the large number of environmental factors affected by this heterogeneity probably also account for the lack of unimodal diversity–productivity relationship observed for most soil organisms.
Not all soil taxa present monotonic responses to productivity gradients, and humpbacked patterns have also been described for some of them. Dahmouche (2007) (after Bouché, 1972), for instance, found that species richness of earthworm assemblages decreased along a gradient of carbon availability (Fig. 9b), although this variation may also be explained by some environmental constraints that are likely to covary with soil carbon levels (e.g. hydromorphy). In this group, however, the systematic limitation of community richness to 10–12 species, whatever the level of available trophic resource (Lavelle et al., 1995; Decaëns et al., 2008b), as well as the well-known cases of competitive exclusion described in some grassland ecosystems (Chauvel et al., 1999; Decaëns et al., 2004) support the idea of strong competitive pressure in the more productive soils. Competition within earthworm communities may be enhanced by the high adaptability and ecological plasticity between juveniles and adults, which lead to a rapid saturation of the ecological space and a relative uniformity in species richness among communities (Decaëns et al., 2008b). Mycorrhizal fungi is another group for which competition is known to limit diversity when nitrogen availability increases (Egerton-Warburton & Allen, 2000; Jonsson et al., 2000; Egerton-Warburton et al., 2007).
Variations in soil biodiversity along disturbance gradients have been extensively illustrated in the scientific literature, mainly through studies on the impact of soil cultivation (Wardle, 2002; Bardgett et al., 2005b). They usually provide little support for the existence of a humpbacked pattern, but rather suggest a negative and monotonic response of most groups to anthropogenic perturbations induced by soil management practices (Wardle, 1995, 2002; Bardgett et al., 2005b). This is the case for microbial communities (Degens et al., 2000), nematodes (Bloemers et al., 1997), springtails (Chauvat et al., 2007), earthworms (Fragoso et al., 1997; Decaëns & Jiménez, 2002; Decaëns et al., 2003), termites (Gillison et al., 2003) or macrofauna as a whole (Lavelle & Pashanasi, 1989; Decaëns et al., 1994; Mathieu et al., 2005) (Fig. 10b). A shared limitation of these studies is that they mainly compare systems that strongly differ not only in disturbance levels but also in a number of other environmental factors. Cultivation is for instance known to severely reduce food supply (and thus productivity) for soil food webs through intense mineralization of soil organic matter and reduction of plant detritus inputs. A decrease in species richness with agricultural intensification may thus not be considered as a sufficient feature to support a negative monotonic response of soil biodiversity to perturbation gradients (Bardgett et al., 2005b). We still need to explore this question through experimental approaches if we wish todisentangle the relative impact of stress and disturbance in structuring soil communities.
Figure 10. Effects of land use on (a) the functional diversity of soil microflora in New Zealand (after Degens et al., 2000; NV, native vegetation; PF, pine forest; in the box-plot representation horizontal line = median, limits of the grey rectangle = confidence interval, ‘error bars’= interquartile range; isolated points = outliers) and (b) species richness of earthworms in north-western France (after Decaëns et al., 2003) (* indicates that an ANOVA showed significant land-use effects at P < 0.05).
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Responses of soil biodiversity to successional gradients
The process of vegetation succession is classically described as a succession of three distinct phases: a build-up phase during which plant biomass aggrades, a maximal biomass phase and a decline or retrogression phase (Wardle, 2002). While the build-up phase is usually followed by a marked accumulation of soil organic matter and an increase in nitrogen availability, the retrogression phase corresponds to a drop in primary productivity and in organic matter and phosphorus availability for soil food webs.
During the aggradation phase, the biomass and diversity of soil microorganisms increase and fungi tend to dominate communities (Ohtonen et al., 1999; Schipper et al., 2001; Bardgett et al., 2005a). On the other hand, soil animal diversity may present different non-synchronous peaks, making any generalisation difficult. This feature has been demonstrated by Doblas-Miranda et al. (2008) in a post-glacial primary succession in New Zealand (Fig. 11a) and by Scheu & Schulz (1996) and Decaëns et al. (1997, 1998) in a post-pastoral secondary succession in Germany and France, respectively (Fig. 11b). Bihn et al. (2008) found ant taxonomic richness (number of genera) to increase during secondary forest succession in the Atlantic Forest of southern Brazil, although this pattern was more consistent for epigeic than for hypogeic taxa. Conversely, Scheu (1992) and Decaëns et al. (1997) observed little change in earthworm species richness along secondary successional gradients under temperate climatic conditions (Fig. 11b), with, however, a slight peak in pre-forested stages and an increasing proportion of epigeics (i.e. surface-dwelling species) relative to anecics (i.e. deep-burrowing species). Ponge & Delhaye (1995) also reported a stable richness of earthworms among different stages in the natural cycle of a virgin beech forest in north-western France. This great variability of responses among soil biotic groups underlines different responses to top-down and bottom-up controls and to abiotic factors during the chronosequence (Doblas-Miranda et al., 2008).
Figure 11. Variation in the diversity of different groups of soil organisms along successional chronosequences: (a) nematodes and macroinvertebrates in a post-glacial primary succession in New Zealand (after Doblas-Miranda et al., 2008); successional stages range from 60 to 120,000 years, the black arrow indicates the onset of the introgression phase); (b) earthworms and macrofauna in post-agriculture secondary successions in France (Fr; after Decaëns et al., 1997, 1998) and in Germany (Ger; after Scheu, 1992; AG, grassland; S1–S3, afforestation stages; FO, forest); (c) microbial communities (AWCD, Average Well Color Development), springtails and macrodetritivorous invertebrates in cycles of forest management in France and Germany (after Chauvat et al., 2003; Hedde et al., 2007; stand age in years); and (d) earthworms and springtails in ageing gradients of temperate grasslands in France and Germany (after Hedde, 2006; Chauvat et al., 2007; CA, annual crop; P, pasture; plotage in years).
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Changes in soil biodiversity during the retrogression phase have seldom been addressed in scientific studies. Doblas-Miranda et al. (2008) found that nematode and macroinvertebrate species richness only weakly decreased, except for epigeic macroinvertebrates, during forest senescence (Fig. 11a). Bernier & Ponge (1994) also showed that vegetation senescence in old temperate forests leads to an increased litter palatability for soil fauna, which favours earthworms to the detriment of epigeic arthropods.
A number of agricultural or sylvicultural systems of production use management cycles that drive the system toward succession-like dynamics. In these anthroposystems, soil communities may present similar temporal patterns to those described for true successions. In a forest cycle in Germany, Chauvat et al. (2003) observed a steady drop in microbial diversity after tree harvesting, followed by a monotonic increase during the growing phase (Fig. 11c). In a similar system in France, Hedde et al. (2007) reported a relative stability of the macrodetritivore species richness in the first phases of the cycle, and then a decrease in the older stands (Fig. 11c). In grazed systems, the diversity of soil fauna usually increases during the first years after vegetation sowing and then stabilizes (Hedde, 2006; Chauvat et al., 2007) (Fig. 11d), which reflects a positive response to an increase in resource availability followed by a state of niche saturation in older plots.