OSTRZEŻENIA!

At the landscape level, patterning of soil biota is mostly related to factors such as soil pattern and topography (Chapter 1), and to factors such as dis-turbances and vegetation patchiness. At the broadest level, it has been argued by Ponge (2003) that the characterization of soil into mull and mor humus types provides a framework for understanding the distribution of soil biodiversity. Fertile mull soils associated with productive ecosystems, such as deciduous forest and temperate grassland, tend to support a high level of plant, animal, and microbial diversity owing to ample provision of resources and a high level of heterogeneity, caused largely by the activities of the organisms themselves. In contrast, the harsh climatic conditions and abundance of recalcitrant organic matter typical of unproductive, mor type soils mean that fewer species are present (Ponge 2003). This kind of patterning is also evident on a more local scale, such as in forests, where differences in the quality of litter beneath coexisting tree species produce

100 90 80 70 60 50 40 30 20 10 0

Rank order decreasing

Local

Frequency of occurrence (%)

World

Fig. 2.11 Frequency of detection of 95 ciliate species in 150 soil samples from 1 ha grassland in Scotland, and of the same species in 606 samples taken worldwide, showing a significant agreement between local and global diversity, that is, that those species that were locally rare were similarly rare on a global scale. (Redrawn with permission from Finaly 2002, Science 296, 1061–1063. Copyright 2002, AAAS)

zones of influence that explain the patchy distribution of soil organisms and process rates (Saetre and Bääth 2000). Similarly, in semiarid ecosystems it has been shown that patterns of soil biodiversity are strongly related to the patchy spatial pattern of vegetation. At these sites, biodiversity in soil below plants is greater than in adjacent exposed soil, suggesting that plants serve as ‘resource islands’ providing resources for soil biota (Herman et al.

1995). In contrast, in ecosystems where vegetation is sparse, abiotic factors determine spatial patterns of soil biodiversity. For example, in the McMurdo Dry Valleys site in Antarctica, patterns in the diversity of soil nematode communities across the landscape depend on abiotic factors such as organic C, salts, and soil moisture (Courtright et al. 2001) (Fig. 2.12).

Another important factor affecting the distribution of soil biodiversity at the landscape level is physical disturbance, which generally leads to dramatic reductions in soil biodiversity. There are numerous such cases. For example, disturbance of soil through tillage has been shown to reduce the abundance and diversity of earthworms (Springett 1992), although the

Fig. 2.12 Dry Valleys in Antarctica. Here, patterns in the diversity of soil nematode communities across the landscape depend on abiotic factors such as organic C, salts, and soil moisture.

(Image by David Hopkins.)

scale of these effects depends on soil type, climate, and tillage operation (Chan 2001). Other studies have examined gradients of disturbance resulting from the conversion of natural vegetation to agriculture, showing that the diversity of certain soil biota is reduced as a consequence. For example, the conversion of primary tropical forest to agriculture has been shown to dramatically reduce microbial biomass (Dlamini and Haynes 2004), and the diversity of macrofauna (Lavelle and Pashanasi 1989), termites (Eggleton et al. 2002), microarthropods (Tain et al. 1992), and nematodes (Bloemers et al. 1997) (Fig. 2.13). Further, studies done in many parts of the world report that earthworm communities of agro-ecosystems have lower species richness and a lower number of native species than do undisturbed ecosystems (Fragoso et al. 1997; Dlamini and Haynes 2004).

The diversity of Collembola is also commonly found to be greater in native prairie than in prairie that has been influenced by agriculture (Brand and Dunn 1998), and the diversity of soil nematodes has been shown to decrease with agricultural improvement of native pastures in New England Tablelands (Yeates and King 1997). Measurement of changes in the relative abundance, or evenness, of phospholipid fatty acids (PLFA) suggests that the broad scale phenotypic diversity of microbial communities in soil is also reduced as a consequence of disturbances caused by intensive agricultural management of temperate grasslands (Bardgett et al. 2001b) (Fig. 2.13). Studies of mycorrhizal fungi also reveal that disturbances lead to dramatic declines in species richness. For example, Daniell et al. (2001) used molecular tools to show that AM fungal diversity is much lower in arable systems than in adjacent, undisturbed woodlands. Similarly, species diversity of EM fungi is reduced by clear-cutting of forests (Hagerman et al.

2001) and late successional forests have more diverse mycorrhizal commu-nities than do earlier successional forests that have been subjected to natural or anthropogenic disturbance (Horton and Bruns 2001). The depletion of species richness resulting from conversion of land to farming is also reinforced by comparison of the number of AM species typically found in arable lands (6 species; Helgason et al. 1998) with the number found in semi-natural grasslands (24; Vanderkoornhuyse et al. 2002) and tropical forest (30; Husband et al. 2002).

It is important to note that there are likely to be numerous, interacting rea-sons for declines in diversity of soil biota resulting from the conversion of natural ecosystems for agriculture. Physical disturbance resulting from site preparation (e.g. deforestation, burning, and cultivation) and declines in the amount and complexity of organic residues returned to the soil are likely to be the most important factors (Beare et al. 1997b). However, con-version of native vegetation to agricultural land also typically results in a substantial loss of soil organic matter owing to the small amounts of organic matter returned to soil, and regular tillage favouring enhanced organic matter mineralization. This is especially the case in farming systems

where crop residues are burned rather than left as mulch (Dominy et al.

2002; Graham et al. 2002; Dlamini and Haynes 2004). The organic matter returned to soil in agricultural systems is also of much lower complexity than that in native ecosystems, often being from a single crop; as will be discussed in Chapter 4, the diversity of decomposers is typically greater in more complex litter mixtures than in simple soil substrates (e.g. Hansen

(a) (b)

(c) (d)

Near primary Old secondary Slash and burn Mechanical

Number

AgroforestryShort fallowsMixed crop

Mechanical clearance

Fig. 2.13 Effects of anthropogenic disturbances on the diversity of soil biota: (a) effects of forest disturbances on the number of species and genera of nematodes in soil of tropical forests, Cameroon (data from Bloemers et al. 1997.); (b) effects of land management on the number of taxonomic units of macrofauna in Peruvian Amazonia (data from Lavelle and Pashanasi 1989); (c) termite diversity across a gradient of anthropogenic disturbance in the humid forest zone of West Africa (data from Eggleton et al. 2002);

and (d) changes in PLFA evenness index of soil microbial communities along a grazing-related disturbance gradient on grasslands in the Snowdonia National Park, UK. Site 1 ungrazed control, 2 long-term ungrazed, 3  short-term ungrazed, 4  lightly grazed, 5 moderately grazed, and 6  heavily grazed (data from Bardgett et al. 2001b).

and Coleman 1998; Hansen 2000). This is also shown by the observation that declines in termite diversity with increasing intensification from primary forest to farmland are strongly linked to diminished plant com-munity complexity (Gillison et al. 2003; Jones et al. 2003) (Fig. 2.14), which is a strong determinant of the spatial and temporal heterogeneity of plant detritus (Beare et al. 1995). Declines in decomposer diversity resulting from the conversion of natural ecosystems to farmland may also be linked to changes in soil microclimate. For example, Critchley et al. (1979) studied relationships between soil microclimate and soil invertebrate populations in cultivated and native bush plots in the humid tropics, and found that diurnal temperatures in cultivated soils ranged from 26 to 32C, whereas in the bush soils the temperature was almost constant at 25C. Bush soils also had a higher moisture content than cultivated ones; consequently, the activity and abundance of most surface-dwelling fauna was greater in native bush than in adjacent cultivated areas (Table 2.3).

Disturbance does not always have negative effects on soil biota. Indeed, some disturbed soils can have highly diverse soil communities, especially when the disturbances enhance spatial heterogeneity, thereby providing opportunities for more species to coexist. A good example of this is soils of the riparian zone (the zone between terrestrial and aquatic ecosystems) that possess an unusually high diversity of fauna and microflora (Ettema et al.

1999, 2000). This diversity is maintained by a variety of disturbances (e.g.

periodic flooding, drought, freezing, abrasion, erosion, and occasionally toxic concentrations of nutrients) that create a spatial and temporal mosaic with few parallels in other systems (Ettema et al. 2000).

Ratio of plant species richness to plant functional type richness

1 2 3

Fig. 2.14 Relationships between termite species richness across a forest disturbance gradient in lowland Sumatra and the ratio of plant species richness to plant functional types. The num-bers next to data points indicate the land use systems:1. Primary rain forest; 2. Logged over rain forest; 3. Industrial softwood plantation; 4. Eight-year-old rubber plantation; 5. Old growth jungle rubber tree mosaic; 6. Imperata grassland; 7. 10-year-old cassava plantation (data from Gillison et al. 2003 and Jones et al. 2003, derived from Giller et al. 2005).

A note on above-ground theory

One intriguing question concerning landscape patterns of diversity is whether the same determinants of above-ground diversity operate below-ground. There are two factors that have traditionally been emphasized as key determinants of above-ground diversity: productivity or resource supply and consumption or physical disturbance (Fig. 2.15). Species richness is often found to be unimodally related to productivity, such that peak diversity is at intermediate productivity: the humpbacked model (e.g. Grime 1973; Al Mufti et al. 1977; Grace 1999), with declining diversity at higher levels of productivity being due to competitive exclusion.

Competitive exclusion can, however, be prevented by periodic mortality events that are caused by consumption or physical disturbance (Connell 1978; Huston 1979); these factors, likewise, display unimodal relationships with diversity, as hypothesized by the ‘intermediate disturbance hypothesis’

of Connell (1978). There are insufficient data available to test either of these ideas with any rigour in soil, and those available do not appear to lend much support for either the productivity–diversity or the disturbance–

diversity relationship operating in soil. It is known that productivity is an important determinant of soil food-web structure at the extremely unpro-ductive end of productivity gradients, for example, in caves, where detrital food chains become longer and more diverse in productive situations, with more energy being diverted to higher trophic levels (Moore and de Ruiter 2000). However, in a literature synthesis Wardle (2002) found no data to support the idea that diversity declines along the most favourable portion of the productivity gradient, that is, that soil organism diversity peaks at intermediate levels of productivity, thereby failing to conform to the humpbacked model. If the humpbacked model does not apply for soil

Table 2.3 Effects of cultivation of native bush plots on populations of microarthropods in a tropical Alfisol (0–10 cm). (Adapted from Critchley et al.

1979)

Land use Native bush plots Cultivated plots

Acari

Cryptostigmata 25,098 4221

Prostigmata 14,830 17,701

Mesostigmata 4570 1887

Astigmata 1609 725

Collembola

Isotomidae 10,607 2158

Entomobryidae 2826 1179

Onychiuridae 1982 348

Sminthuridae 1408 942

Poduridae 865 63

biota it suggests that below-ground communities could differ from their above-ground counterparts, in that soil biodiversity is not so strongly regulated by competition, and competitive exclusion does not occur when resource availability in soil is increased (Wardle 2002). Similarly, there is little support in the literature for optimization of soil animal diversity at intermediate levels of disturbance (see Bardgett et al. 2005); on the contrary, as already discussed, disturbances generally lead to dramatic reductions in soil biodiversity.

Another important determinant of above-ground diversity that has long been recognized is island size, which is part of the island biogeography the-ory of McArthur and Wilson (1967). Island size determines, among others things, the amount of resources available for organisms, and in general, larger islands have a greater variety of resources and hence more species—

the so-called species–area relationship. Wardle et al. (2003a) tested these ideas for decomposer communities by examining the diversity of groups of biota in the organic matter of ‘suspended soils’ produced by spatially sepa-rated epiphytes or treetop ‘islands’ in the crowns of canopy tree species in old-growth, warm, temperate forest in Northland, New Zealand. They found that larger ‘islands’ supported a greater diversity of macrofauna and microarthropods, which is consistent with the island size theory. It was also found that physical proximity of islands to sources of colonizing organisms did not influence the diversity of these faunal groups, which contrasts with the predictions of the island biogeography theory (Wardle et al. 2003a).