## Richard Bardgett

Print publication date: 2005

Print ISBN-13: 9780198525035

Published to Oxford Scholarship Online: April 2010

DOI: 10.1093/acprof:oso/9780198525035.001.0001

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# Organism interactions and soil processes

Chapter:
(p.57) 3 Organism interactions and soil processes
Source:
The Biology of Soil
Publisher:
Oxford University Press
DOI:10.1093/acprof:oso/9780198525035.003.0003

# Abstract and Keywords

This chapter illustrates how the activities of soil biota, especially their trophic interactions, influence the processes of decomposition and nutrient cycling, and examines the significance of this for material flow and plant production in terrestrial ecosystems. The focus is on the availability of nitrogen and phosphorus since they are the two nutrients that most limit primary productivity in natural and managed terrestrial ecosystems. First, the issue of how soil microbes regulate the internal cycling of nutrients in terrestrial ecosystems is discussed. This is followed by a discussion of how soil animals influence nutrient cycling and plant growth through their feeding activities on microbes and other fauna.

# 3.1 Introduction

In the previous chapter, the main players in the soil food web and the factors that regulate patterns of soil biodiversity, were introduced. The next issue to be considered is the functional significance of these organisms and their interactions with one another for ecosystem processes, especially nutrient cycling. Decomposer organisms are essential for the functioning of terrestrial ecosystems largely because they decompose dead organic material in soil, converting this into carbon dioxide and other soluble nutrient forms that provide resources for other biota and primary production. In natural ecosystems, most N and a portion of the P required for plant growth are supplied through decomposition of organic matter, relying therefore on the activities of soil biota. Even in agricultural systems that typically receive large inputs of fertilizers, the supply of plant nutrients via the decomposition of organic matter is central to maintaining sustainable levels of crop production.

The aim of this chapter is to illustrate how the activities of soil biota, especially their trophic interactions, influence these processes of decomposition and nutrient cycling, and examine the significance of this for material flow and plant production in terrestrial ecosystems. The focus will be on the availability of N and P since they are the two nutrients that most limit primary productivity in natural and managed terrestrial ecosystems (Chapin et al. 2002). First, the issue of how soil microbes regulate the internal cycling of nutrients in terrestrial ecosystems will be discussed. This will be followed by a discussion of how soil animals influence nutrient cycling and plant growth through their feeding activities on microbes and other fauna. (The influence of plant-parasitic organisms on nutrient cycling will not be considered here, since this issue is discussed in more detail in Chapter 4). (p.58) The issue of how soil organisms and their interactions affect ecosystem properties by modifying the physical environment in which they live will then be discussed. Finally, a related and very topical issue that is high on the political and scientific agenda will be considered, that is, the question of whether variations in the diversity and architecture of soil food webs have important consequences for ecosystem functioning. In other words, is there a relationship between soil biological diversity and ecosystem functioning?

# 3.2 Microbial control of soil nutrient availability

In total, 80–90% of primary production enters the soil system as dead plant litter and roots, and the primary decomposers of this material are the bacteria and fungi. Initial fragmentation and ingestion of fresh organic material by detritivores increases its surface area for microbial colonization, but the microbes, via the production of enzymes, do most of the chemical alteration of this material. As noted in Chapter 2, microbes have the capacity to produce a vast array of enzymes that can degrade almost all plant-derived compounds.

In many ecosystems, fungi are the most abundant primary decomposers. They are better equipped than bacteria for carrying out the decaying of insoluble plant material since their hyphal networks can penetrate new substrates and proliferate both within and between dead plant cells. They can also produce a range of enzymes that are capable of degrading complex cell wall components, such as lignin, enabling them to gain access to more labile compounds that occur within the cell. Another advantage of fungi over bacteria is their ability to transport nutrients through their hyphal network to zones of exploitation. This means they have the capacity to exploit new substrates, even when nutrients such as N and P are limiting (Boddy 1999). In contrast, bacteria are either immobile or move passively through soil. Therefore, they tend to exhaust substrates that become available in their immediate microenvironment, and then become inactive until resources once more become available, for example, due to the exudation of soluble C compounds from newly growing roots.

## 3.2.1 Nitrogen mineralization

The mineralization of nutrients—the process by which soil microbes break down soluble and insoluble organic matter and convert it into inorganic forms—is of critical importance for ecosystem function since, in many ecosystems, it directly determines the availability of nutrients to plants. For example, in fertile ecosystems, such as deciduous forests, plant N supply is strongly correlated with rates of N mineralization in soil (Nadelhoffer et al. 1985) (Fig. 3.1). A summary of the soil N cycle is given in Fig. 3.2.

(p.59)

Fig. 3.1 Above-ground plant N uptake in relation to net N mineralization in the surface 0–20 cm soil. Symbols designate dominant genera on sites: O = oak, B = birch, M = maple, P = pine, S = spruce. (Data from Nadelhoffer et al. 1985.)

Fig. 3.2 Schematic diagram of the terrestrial N cycle.

(p.60) Most N in the soil (some 96–98%) is contained in dead organic matter as complex insoluble polymers such as proteins, nucleic acids, and chitin. These polymers are too large to pass through microbial membranes, so microbes produce extracellular enzymes (e.g. proteinases, ribonucleases, and chitinases) that break them down into smaller, water soluble subunits that can be absorbed by microbial cells (e.g. amino acids). This material is called dissolved organic N (DON) and it can constitute a large proportion of the total soluble N pool, especially in infertile ecosystems with mor soils, such as Boreal forest and arctic tundra (Northup et al. 1995; Jones and Kielland 2002). Even in agricultural soils that receive regular dressings of inorganic N fertilizer, concentrations of DON in soil can be equal to, or even higher than, those of inorganic N (Bardgett et al. 2003).

The majority of DON in soil solution is absorbed by free-living soil microbes, which use the C and N contained within it for their growth. How microbes use DON, however, depends mainly on whether they are C- or N-limited. Under conditions when microbial growth is C-limited, microbes use the C from DON to support their energy needs for growth and they excrete plant available ammonium as a waste product into the soil; that is, N is mineralized by the microbial biomass. When DON is not sufficient to meet microbial N demand, however, microbes absorb additional inorganic N from soil solution; that is, N is immobilized by the microbial biomass, thereby reducing the availability of inorganic N for plant uptake. Besides plant uptake, ammonium produced by microbial mineralization has several potential fates: it can be absorbed onto negatively charged surfaces of clay minerals and organic matter and, when oxygen is plentiful, it can be oxidized by nitrifying bacteria to produce nitrate (Box 3.1).

Both immobilization and mineralization of N occur simultaneously in soil, but net N mineralization occurs when microbes are predominately C-limited, whereas net immobilization occurs when their growth is limited by N. The balance between mineralization and immobilization is determined by a range of factors such as the availability of DON and predation by soil animals, but one factor that is of high importance is the relative demand by microbes for N and C, which is determined by the C : N ratio of the organic substrate they are utilizing. Heterotrophic bacteria and fungi have C : N ratios ranging from 4 : 1 to 12 : 1, but as they break down organic matter they respire some 50% of the C contained within it and use the remaining 50% for biomass production (Kaye and Hart 1997). Therefore, there is a widely cited critical substrate C : N ratio of around 30 : 1 which is needed to meet microbial needs for N (Kaye and Hart 1997). If substrate C : N ratios are higher, which is often the case for plant litters, microbes become N-limited and hence use (or immobilize) exogenous sources of inorganic N while decomposing the substrate (Kaye and Hart 1997). In such situations, plants and microbes should theoretically compete for N in soil solution. Several factors influence substrate quality in ecosystems, such as inherent soil fertility and plant community composition, and factors such as grazing that modify the quality of plant litter inputs to soil within ecosystems; the importance of these factors for soil nutrient cycling will be discussed in Chapters 4 and 5.

(p.61)

(p.62) Not all DON in soil solution is used by microbes; significant quantities may be leached from soil in drainage waters (Perakis and Hedin 2002) and there is growing evidence that plants can uptake DON directly from soil in the form of amino acids, thereby bypassing microbial mineralization (reviewed by Lipson and Näsholm 2001). This has been shown to be the case in many ecosystem types, but is thought to be especially prevalent in less fertile ecosystems, such as the arctic tundra (Chapin et al. 1993; Kielland 1994; Schimel and Chapin 1996; Lipson and Monson 1998; Raab et al. 1999; Henry and Jefferies 2003) and Boreal forest (Näsholm et al. 1998; Nordin et al. 2001). In contrast, in agricultural situations, while plants clearly have the capacity to uptake amino acids directly (e.g. Näsholm et al. 2000, 2001; Streeter et al. 2000) this is thought to be of limited importance for plant nutrition due to very rapid microbial mineralization of DON (Hodge et al. 1998, 1999; Owen and Jones 2001; Bardgett et al. 2003). An exception to this, however, could be where amino acid availability is especially high, for example, in resource hot spots or patches in soil. Many plants have the capacity to exploit such resource patches by root proliferation, and hence, in these situations, the acquisition of organic N could be a significant pathway of plant nutrition (Hodge et al. 1998).

## 3.2.2 Nitrogen fixation

Another route through which plants gain N is via the process of N fixation. This is especially the case in natural, unpolluted terrestrial ecosystems where N fixation is often the primary route by which N enters the system (Cleveland et al. 1999). Microbes that have the capacity to reduce molecular N2 to NH3, and incorporate it into amino acids for protein synthesis, drive this process of N fixation. These microbes are either free-living or (p.63) they live in symbiotic association with plants, forming nodules in the root where they receive carbohydrate from the plant to meet their energy needs and in turn they supply the plant with amino acids formed from reduced N.

Table 3.1 Rhizobia and their host plants. (Adapted from Paul and Clark 1996)

Genus

Species

Host plant

Rhizobium

Meliloti

Alfalfa (Medicago, Melilotus)

leguminosarum

Peas (Pisum)

Vetches (Vicia)

Clovers (Trifolium)

Beans (Phaseolus)

loti

Trefoil (Lotus)

fredii

Soyabean (Glycine)

japoniucum

Soyabean (Glycine)

Tropical legumes (Arachis, Leucaena)

Azorhizobium

caulinodans spp

Stem nodules (Sesbania)

Non-legumes (Parasponia)

The most widely known N fixers are probably the legumes that are associated with bacterium of the genus Rhizobium (Table 3.1) and non-legumes such as alder that have the actinomycete Frankia as their endophyte (Table 3.2). These symbiotic associations can supply large amounts of N into soil of both natural and agricultural ecosystems, and their benefits for soil fertility have been recognized since pre-Roman times. In agricultural grassland, for example, legumes can contribute up to 150 kg N ha−1 yr−1 (Newbould 1982), and in natural situations, especially in early primary succession, they are often the major contributors to N accumulation (Walker 1993). The primary route by which this N enters the soil from legumes is through the breakdown of litter inputs that are enriched with N. Another point of entry is via the exudation or leakage of N-rich exudates from roots. (p.64) Details on the establishment of symbiosis, the biochemistry of N fixation, and the factors that influence rates of N fixation, are given in Paul and Clark (1996).

Table 3.2 Distribution of actinorhizal plants. (Adapted from Paul and Clark 1996)

Continent

Native genera

North America

Alnus, Ceanothus, Cerocarpus, Chamaebatia, Comptonia, Coriaria, Cowania, Datisca, Dryas, Elaeagnus, Myrica, Purshia, Shepherdia

South America

Alnus, Colletia, Coriaria, Discaria, Kentrothamnus, Myrica, Retanilla, Talguenea, Trevoa

Africa

Myrica

Eurasia

Alnus, Coriaria, Datisca, Dryas, Elaeagnus, Hippophae, Myrica

Oceania, including Australia

Allocasuarina, Casuarina, Ceuthostoma, Coriaria, Discaria, Gymnostoma, Myrica

Another potentially important pathway of N fixation is through free-living bacteria that can fix N in the course of decomposing litter and soil organic matter (Vitousek and Hobbie 2000). These non-symbiotic N-fixers are ubiquitous in terrestrial ecosystems (Paul and Clark 1996), and fix relatively small, but significant amounts of N (<3 kg N ha−1 yr−1) (Cleveland et al. 1999). In extreme environments with sparse cover of vascular plants, cyanobacteria are often the main source of N. These N-fixing organisms are especially abundant in dry and/or cold regions, where the growth of vascular plants is limited. For example, in arid ecosystems they occur in soil crusts (Box 3.2) composed of cyanobacteria, lichens, and mosses (Belnap 2003). Similarly, cyanobacteria are abundant as crusts or felts on the soil in the Antarctic (Wynn-Williams 1993) and in recently deglaciated terrain, such as that found at Glacier Bay (Fig. 3.3) (Chapin et al. 1994). In all these situations, they are vital for creating and maintaining fertility, often fixing appreciable amounts of N (1–10 kg N ha−1 yr−1) (Vitousek et al. 2002); they also have an important role in stabilizing surface soil particles in extreme environments (Wynn-Williams 1993; Belnap 2003).

Recent findings also point to cyanobacteria as being the primary N source in Boreal forest, where N fixation has traditionally been thought to be extremely limited. DeLuca and colleagues (2002) showed that a N-fixing symbiosis between a cyanobacterium (Nostoc sp.) and the ubiquitous feather moss Pleurozium schreberi fixes significant quantities of N (1.5–2.0 kg N ha−1 yr−1), and acts as a major contributor to N accumulation and cycling in Boreal forests (Fig. 3.4). This finding is especially significant since P. schreberi is the most common moss on Earth, found throughout North and South America, Greenland, Asia, Europe, and Africa, and in Boreal forests, it often forms a continuous carpet, accounting for up to 80% of the ground cover (DeLuca et al. 2002).

## 3.2.3 Microbial phosphorus mineralization

In contrast to N, most P in soil is in a range of insoluble inorganic forms that are unavailable for plant uptake, for example, as occluded P (Fig. 3.5). As a result, many of the reactions that govern plant availability of P are geochemical, rather than biological (Box 3.3). Soil P availability is also strongly affected by past management. For example, P fertilization has a substantial and incredibly long-term effect on soil P availability, in that it greatly elevates plant available phosphate $( PO 4 3 − )$ concentrations in soil, especially in surface horizons. Despite this, soil microbes are closely involved in the cycling of P, in that they participate in the solubilization of inorganic P and (p.65) in the mineralization of organic P, the latter being governed by the production of plant and microbial phosphatases in situations of low P availability. These enzymes act by cleaving ester bonds in organic matter to liberate phosphate $( PO 4 3 − )$ that can be taken up by the plant or microbial biomass.

(p.66)

Fig. 3.3 Black algal crusts (left) adjacent to the pioneer N-fixer Dryas drummondii (right) that colonize recently deglaciated terrain in Glacier Bay, Alaska. (Image by Richard Bardgett.)

Fig. 3.4 Carpets of feather moss Pleurozium schreberi develop on the floor of Boreal forests, such as those in Alaska, where in association with cyanobacteria it fixes significant quantities of N, and acts as a major contributor to N accumulation and cycling. (Image by Richard Bardgett.)

(p.67)

Fig. 3.5 Schematic diagram of the terrestrial P cycle, showing main transformations of P that occur in soil.

As with N, the mineralization of organic P is partly regulated by the C : P ratio of substrates. In general, when the C : P ratio rises above ∼100, P is immobilized by microbes which have a relatively high P requirement (1.5–2.5% P by dry weight compared to 0.05–0.5% for plants). As a result, microbes compete aggressively with plants for available P in soil. The importance of microbial immobilization of P is illustrated by the fact that microbes often contain as much as 20–30% of the total soil organic P pool (Jonasson et al. 1999a), which is much larger than the proportion of C (∼1–2%) or N (∼2–10%) held in microbes. Also the capacity of the soil microbial biomass to act as a sink for P is strongly affected by environmental stresses, such as wetting and drying cycles, that are known to result in the death of a portion of the soil microbial biomass and a resultant flush of available P, and presumably other nutrients, into soil solution (Turner and Haygarth 2001). Microbial P is, therefore, an important source of potentially available P in soil. Mycorrhizal fungi also play an important role in the transfer of P in soil systems, as will be discussed in the next section.

(p.68)

## (p.69) 3.2.4 The role of mycorrhizal fungi in plant nutrient supply

Mycorrhizae are distributed across a wide range of ecosystems, but there is a distinct pattern in the distribution of different mycorrhizal types according to biome, soil type, and limiting resources (Fig. 3.6) (Read 1983). In general, ericoid mycorrhizal fungi dominate in high latitude and altitude ecosystems, which are dominated by dwarf-shrub heath vegetation on peaty N-limited soils; Ectomycorrizal fungi dominate Boreal and broad-leaved forest ecosystems, that typically occur on brown earth and podzolic soils; whereas vegetation of tropical forest, grassland, and desert predominately have AM fungal associations. This pattern provides a simple framework to characterize the functional role of mycorrhizae in relation to plant community types.

The primary benefit to plants of having mycorrhizal associations is enhanced acquisition of nutrients from soil, especially of N and P. Indeed, mycorrhizal fungi often constitutes one of the largest C sinks for primary productivity (Harley 1971), and allocation to these fungi is especially high when nutrients are limiting (Johnson et al. 2003c). It has long been known that mycorrhizal fungi enhance ammonium uptake from soil, and more recently it has been shown that both ericoid and ectomycorrhizal fungi have the capacity to uptake DON directly from soil, in the form of amino acids, pure protein, and even as highly recalcitrant proteins that are co-precipitated with tannins (Read 1994). They can use these more complex proteins (p.70) because they produce carboxyl-proteinase enzymes that have the ability to cleave proteins into constituent amino acids. Interestingly, the production and activity of these enzymes is optimal at soil pHs of 2–4, a range that coincides with that of the mor soils where these fungi typically occur (Leake and Read 1990). There is also emerging evidence that AM fungi can both enhance decomposition of, and increase N capture from, complex organic material in soil (Hodge et al. 2001); the significance of this for the growth and nutrition of AM plants is yet to be established.

Fig. 3.6 Relationships between habitat types, limiting resources, and mycorrhizal associations.(Adapted from Read 1983.)

Another route by which mycorrhizal fungi can influence soil N availability is by increasing the fixation of N by plant-microbial N-fixing associations. AM fungal infection, for example, is known to increase rates of nodulation and N fixation in legumes (Haystead et al. 1988) and in plants with actinorhizal associations (Gardner et al. 1984), thereby potentially increasing the supply of N to soil. AM fungi can also transport fixed N from legumes, through soil, to neighbouring plants via hyphal networks. For example, N transfer between agricultural legumes and grasses has been shown to occur via AM hyphal networks (e.g. Haystead et al. 1988; Zhu et al. 2000) and it has been proposed—but not tested—that this route is of greatest significance in N-limited, early successional plant communities where legumes are often strongly mycorrhizal (Haystead et al. 1988). Similar transfers of N are also thought to occur via ectomycorrhizal fungi between actinorhizal alder (Alnus sinuata) and its later successional neighbours (Allen 1991).

Mycorrhizal fungi also have an important role to play in the uptake of soil P by host plants. These fungi grow deep into the soil matrix, accessing soil P that is beyond the roots. They do this through two important mechanisms. First, they are known to produce phosphatase enzymes that cleave ester bonds that bind P to C in organic matter, thereby releasing phosphate $( PO 4 3 − )$ that can be taken up by the fungi and passed on to the plant. Second, they produce low molecular weight organic acids, such as oxalates, which enhance the availability of soil P by increasing weathering rates of P contained in clay minerals, and by complexing cations (e.g. Ca, Fe, and Al) that would otherwise bind to phosphates, thereby taking them out of soil solution.

# 3.3 Influence of animal–microbial interactions on nutrient availability

Mineralization of nutrients is governed directly by the activities of bacteria and fungi. The ability of microbes to do this, however, is affected strongly by soil animals that live alongside them, and also by food web interactions that determine the transfer of nutrients through the plant-soil system. There are many ways that soil animals can affect microbial nutrient mineralization, (p.71) but for convenience they are typically divided into three broad pathways. First, soil animals affect nutrient cycling through their selective feeding on microbes, which alters microbial activity, abundance, and community structure. Second, soil animals affect nutrient mineralization by altering the form of the organic matter in soil, in that they fragment and mix organic matter inputs to soil, increasing its susceptibility to microbial attack. Third, it has been suggested that soil animals, notably protozoa, can have non-nutritional effects on plant growth, in the form of hormonal effects on root morphology. Combined, these interactions between microbes and animals drive processes of energy flow and nutrient cycling, and, therefore contribute to plant nutrient acquisition and plant growth. In the following section, the effects of these biotic interactions on the structure and activity of microbial communities, and the consequences of this with regard to nutrient availability to plants will be discussed.

## 3.3.1 Selective feeding on microbes by soil animals

While soil animals are increasingly considered to be both generalist and opportunist in their feeding behaviour (Scheu and Falca 2000; Ponsard and Arditi 2000; Maraun et al. 2003), there is also evidence that many of them are very selective in this regard. For example, it is well known that certain nematodes have mouth parts that are adapted to feeding on either bacteria (bacterivores) or fungi (fungivores) (Yeates et al. 1993), and there is evidence that certain protozoa use chemical cues to discriminate between bacterial and algal prey (Verity 1991). Other faunal groups, notably the collembolans, have been shown to be even more selective in their feeding behaviour, actually choosing to feed on particular fungal species over others (e.g. Bardgett et al. 1993b). One of the best examples of this is that of Newell (1984a,b) who showed that selective grazing by the collembolan Onychiurus latus on the fungus Marasmius androsaceus in coniferous leaf litter resulted in a reduction in the activity of this palatable fungus, and an increase in the abundance of an unpalatable fungus Mycena galopus present in the litter. This change in fungal community structure then had a knock-on effect at the ecosystem scale, in that it reduced decomposition rates of coniferous litter, since M. galopus decomposes litter at a slower rate than M. androsaceus. This kind of selection of prey can be based on numerous factors, but it is mostly related to the palatability of the fungus, which varies with species, age, and physiological status. For example, the collembolan Folsomia candida has been shown to feed on metabolically active hyphae in preference to dead or inactive hyphae (Moore et al. 1985) and also selects regions of the fungal thallus with high N content (Leonard 1984). Preferential grazing is also thought to arise as a result of the avoidance of toxins that are produced by some fungal colonies (Parkinson et al. 1979), and some species of Collembola have been shown to locate and select their fungal food source by volatile compounds released from fungal mycelium (Bengtsson et al. 1988). (p.72) Preferential grazing by collembolans on fungus has dramatic effects on the activity and abundance of the prey, and, as shown above, can also alter fungal community structure. It is well known, for example, that an intermediate level of grazing by collembolans actually enhances the activity and growth of selected fungi (Hanlon and Anderson 1979; Bengtsson and Rundgren 1983; Bardgett et al. 1993c). This stimulation of fungal growth, often referred to as compensatory growth, is due to new fungal growth after senescent hyphae are grazed, and regrowth after periodic grazing of actively growing mycelia. For example, a study by Hedlund et al. (1991) showed that grazing of the fungus Mortierella isabellina by the collembolan O. armatus induced switching from a ‘normal’ hyphal mode, with appressed growth and sporulating hyphae, to fan-shaped sectors of fast growing and non-sporulating mycelium with extensive areas of aerial hyphae. These authors also showed that the activities of specific amylase enzymes were several times greater in grazed cultures than in those cultures that were ungrazed. In contrast, heavy grazing of fungal communities can reduce their activity, as shown by Warnock et al. (1982) who found that heavy grazing of mycorrhizal fungi by collembolans counteracted their mutualistic relationship with the host plant.

The ingestion and exposure of microbes to intestinal fluids can greatly influence microbial activity. For example, microbial communities have been shown to be more abundant and active following passage through the gut of earthworms (Brown 1995), and dormant bacteria can be activated during this intestinal journey due to the removal of endospore cell coats and subsequent germination of bacterial spores (Fischer et al. 1997). However, the earthworm gut can also be a hostile environment for microbes. As shown by Moody et al. (1996), the germination of fungal spores can be reduced following ingestion by earthworms, due to exposure to intestinal fluids.

## 3.3.2 Effects of microbial-feeding fauna on nutrient cycling and plant growth

The effects of microbial-feeding microfauna on microbial activity, nutrient mineralization, and primary production are generally positive (Mikola et al. 2002). Enhanced C mineralization results from increased turnover rate, activity, and respiration of grazed microbial populations (Anderson et al. 1981; Kuikman et al. 1990; Bardgett et al. 1993c; Mikola and Setälä 1998a; Cole et al. 2000), whereas enhanced N mineralization is mainly due to direct animal excretion of excess N (Woods et al. 1982). In general, grazers have lower assimilation efficiencies than the microbes upon which they graze, and therefore they excrete into soil nutrients that are not required for production, in forms that are biologically available (e.g. protozoa preying on bacterial populations are assumed to release about one-third of the N consumed). This release of nutrients into the soil system is effectively a remobilization (p.73) of the nutrients that were bound up in the microbial biomass, and has been termed the ‘microbial loop’ (Clarholm 1985).

The significance of the ‘microbial loop’ is that the nutrients released from microbial biomass by grazers increase the availability and uptake of nutrients by plants (Clarholm 1985), and, in some cases, enhance plant growth. For example, a classic study by Ingham et al. (1985) showed that the addition of microbial-feeding nematodes to soil in pot experiments enhanced N uptake by, and growth of, the grass Bouteloua gracilis (Fig. 3.7(a)). Similarly, Bardgett and Chan (1999) showed that, when acting together, Collembola and nematodes increased concentrations of ammonium in the soil solution of an upland organic soil, leading to increased plant nutrient uptake by, but not growth of, the grass Nardus stricta (Fig. 3.8). Studies of trees also show similar responses to the addition of animals. For example, Setälä and Huhta (1991) showed that leaf, stem, and shoot biomass of birch seedlings (Betula pendula) were increased when the plants were grown in the presence of a diverse soil fauna (Fig. 3.7(b)), and Setälä (1995) showed that grazing of ectomycorrhizal fungi associated with B. pendula by soil fauna resulted in enhanced growth of, and nutrient uptake by, B. pendula, despite reductions in the biomass of ectomycorrhiza. There is therefore ample evidence to suggest that faunal grazing on microbes can enhance both microbial activity and the availability of nutrients for plants, thereby influencing net primary productivity (NPP).

Evidence is also emerging that interactions between decomposer organisms can influence the performance of higher trophic groups that live above-ground, especially herbivores that benefit from enhanced plant nutrition. For example, Scheu et al. (1999) showed that plant-sucking aphids (p.74) performed better where the host plant was grown in the presence of microbial-feeding Collembola or earthworms, as compared to systems without these organisms. Similarly, bacterial-feeding microfauna were shown to indirectly increase the numbers and biomass of aphids on barley shoots through their positive effects on soil N turnover and the nutritional status of the plant (Bonkowski et al. 2001a), and the presence of earthworms has been shown to increase foliage consumption by a leaf chewer (Mamestra brassicae), due to a positive effect of these soil animals on soil N availability and foliar N content (Newington et al. 2004). It is therefore apparent that important indirect linkages and feedbacks operate between the consumer organisms of the above-ground and below-ground food webs.

Fig. 3.7 Influence of soil trophic interactions on plant growth. (a) Adding bacterial-feeding nematodes to soil increased shoot and root growth of the perennial grass Bouteloua gracilis (data from Ingham et al. 1985). (b) Addition of diverse fauna to soil increased root and shoot biomass of birch seedlings (Betula pendula). Filled bars = shoot biomass; open bars = root biomass. (Data from Setälä and Huhta 1991.)

Fig. 3.8 Effects of animal treatments (N = nematodes, C = Collembola, and N + C = nematodes and Collembola in combination) on the amount of $(NO 3 − )$ (μg g−1 soil) and $( PO 4 3 − )$ −P (μg g−1 soil) leached from microcosms, and shoot N and P content (mg g−1). Values are means ±SE. For each measure, values with the same letter are not significantly different. (Data from Bardgett and Chan 1999.)

It is important to recognize that effects of soil animals on nutrient mineralization and plant nutrient uptake are case specific. For example, soil fauna have been shown to both reduce or enhance microbial activity and (p.75) mineralization depending on the season (Teuben 1991) and the abundance of grazers (Hanlon and Anderson 1979; Hanlon 1981). The effects of interactions between fauna and microbes on nutrient release are also highly dependant on resource quality. In N-limited soils, for example, nutrients released by fauna grazing on fungi are often rapidly re-utilized by microbes, and hence do not become available for plant uptake (Visser et al. 1981; Bardgett et al. 1993c; Cole et al. 2002b). This suggests that in nutrient limited conditions, fungal-feeding animals alone are not able to limit the ability of microbes, and especially fungi, to sequester nutrients and hence compete with plants. Effects of animal–microbial interactions on nutrient cycling are also likely to be spatially constrained, occurring largely in hot spots of activity around resource patches (Bonkowski et al. 2000).

## 3.3.3 Non-nutritional effects of microbial grazers on plant growth

Beneficial effects of protozoa on plant growth are well documented (Clarholm 1985; Kuikman et al. 1990; Jentschke et al. 1995; Alphei et al. 1996; Bonkowski et al. 2000), and are usually assigned to nutritional effects via the ‘microbial loop’. In recent years, however, there is ample evidence to suggest that protozoan grazing could also have non-nutritional effects on plant growth. The idea here is that protozoa indirectly influence root architecture by influencing bacterial production of plant growth-promoting hormones. A number of studies show that plants grown in the presence of protozoa develop an extensive and highly branched root system, due to strong branching of lateral roots, and that these effects resemble those caused by plant growth-promoting hormones, such as auxins (Jentschke et al. 1995; Bonkowski et al. 2000, 2001b; Bonkowski and Brandt 2002). While it has long been known that protozoa can themselves release auxins (Nikolyuk and Tapilskaja 1969), recent studies reveal that selective grazing by protozoa in the rhizosphere significantly stimulates the growth of auxin-producing bacteria; this has been proposed as the most likely mechanism of protozoan effect on root growth (Bonkowski and Brandt 2002; Bonkowski 2004) (Fig. 3.9). Changes in root architecture may also be due to other effects of protozoa grazing on the bacterial community. In particular, stimulation of nitrifying bacteria may lead to hot spots of nitrate concentration; as well as being a source of N for plant growth, nitrate acts as a signal for lateral root elongation and may act to direct root growth towards nutrient patches (Zhang and Forde 1998). Overall, these studies reveal that protozoan effects on plant growth are more complex than previously assumed, in that a combination of nutritional and non-nutritional responses to microbial grazing are at play (Bonkowski 2004).

## 3.3.4 Multitrophic controls on soil processes

Effects of biotic interactions in soil on decomposition and nutrient release are complex and involve a diversity of species from more than one trophic (p.76) group. For example, the effects of feeding of fungal-feeding Collembola on nutrient cycling have been shown to become apparent only when they are interacting with another trophic group of soil fauna, namely microbial-feeding nematodes (Bardgett and Chan 1999). Likewise, it has been demonstrated that combinations of soil animals, as opposed to a single group of soil fauna, had a synergistic effect on the microbial community in microcosms of coniferous forest humus, resulting in enhanced leaching of mineral nutrients (Setälä et al. 1991). Animals also prey on each other which complicates the interpretation of how soil fauna affects ecosystem processes. Not much is known about the importance of predation in soil food webs, but the studies to date reveal that manipulation of predatory fauna can have dramatic effects on processes of decomposition and nutrient mineralization and that these effects can be both positive and negative, depending on the particular circumstances. This occurs because predatory fauna can induce trophic cascades that lead to positive, neutral, and negative effects on the activity or biomass of microbes. For example, Santos et al. (1981) found that reductions in predatory mites in soil food webs of desert soils led to increased abundance of their prey, bacterial-feeding nematodes, which in turn led to reduced microbial growth and decomposition of plant litter. In other studies, however, reductions in abundance of microbial-feeding nematodes by predation have led to neutral (Laakso and Setälä 1999a) or negative (Bouwman et al. 1994; Mikola and Setälä 1998a,b; Setälä et al. 1999) (p.77) effects on nutrient mineralization, presumably due to differential response of microbes to changes in their predation. In sum, it is clear that predatory fauna have the potential to induce cascading effects on soil food webs that ultimately influence rates of nutrient mineralization and hence plant nutrient availability.

Fig. 3.9 Schematic diagram of grazer-induced hormonal effects on root growth. (1) plant roots release exudates thereby: (2) stimulating growth of bacteria; and (3) bacterial predators; selective grazing by protozoa favours auxin-producing and nitrifying bacteria; inducing lateral root growth; (6) leading to enhanced exudation; and (7) bacterial growth. (Redrawn with permission from the New Phytologist Trust; Bonkowski 2004)

# 3.4 Effects of animals on biophysical properties of soil

Soil fauna also affect ecosystem processes of nutrient cycling via physical alteration of decomposing material and the soil environment. Two groups of organisms are responsible for this: (1) the litter transformers that consume plant detritus and egest this material into soil as fecal pellets, thereby affecting rates of decomposition and nutrient release; and (2) the ecosystem engineers that build physical structures in soil that provide habitats for microbes and other organisms, and also alter the movement of materials through soils and across ecosystems (Lavelle et al. 1995).

## 3.4.1 Consumption of litter and the production of fecal pellets

Litter consumers are animals such as microarthropods and some macro-fauna that consume plant detritus and egest this material into soil as fecal pellets. Such fecal pellets have a higher surface-to-volume ratio compared to the original leaf litter, which enhances its rate of decomposition (Webb 1977). Fecal pellets also provide a highly favourable environment for microbial growth, especially of bacteria, again leading to increased rates of decomposition and nutrient release (Hassall et al. 1987; Zimmer and Topp 2002). Decomposition is also enhanced through the action of endosymbiotic microbes that reside in the guts of many soil animals, such as earthworms, isopods, and termites. These endosymbiotic microbes produce extracellular enzymes that degrade cellulose and phenolic compounds thereby further enhancing the degradation of ingested material (Zimmer and Topp 1998).

Earthworms consume vast amounts of litter fall in many ecosystems, but especially those in mull type soils (Table 3.3). Indeed, in mixed forest ecosystems, earthworms can consume the entire annual litter fall of the forest (Nielson and Hole 1964; Satchell 1967). This organic matter is consumed along with soil mineral particles, and these two fractions are mixed together in the earthworm gut and then egested as surface or subsurface casts. Estimates from various habitats around the globe show that earthworms produce between 2 and 250 ton of cast ha−1, whereas typical values for many temperate ecosystems probably range from 20 to 40 ton per hectare (Edwards and Bohlen 1996). It is widely documented that these casts contain significantly greater numbers of microbes and have higher (p.78) enzyme activities than the surrounding soil, either because of enrichment of organic matter with microbes in the worm’s intestine or because the cast provides a better substrate for microbial growth, being rich in organic matter and available nutrients. As a result, rates of decomposition and C mineralization in soil are significantly enhanced by the presence of earthworms (Cortez et al. 1989). Similarly, rates of N mineralization have been found to be greater in earthworm casts relative to surrounding soil, largely due to them being enriched with inorganic N and N-rich excretory products and mucus from the earthworm (Lavelle and Martin 1992). P availability is also reported to be much greater in casts than surrounding soil, largely due to the stimulation of phosphatase activity (Sharpley and Syers 1976). While these positive effects of casting on microbial activity are often transitory (Lavelle and Martin 1992), their net effect, along with other earthworm activities, is the stimulation of total soil nutrient availability (Anderson et al. 1983; Scheu 1987) and enhanced plant nutrient uptake (Edwards and Bohlen 1996).

Table 3.3 Amount of organic matter ingested or incorporated into soil by earthworm populations of different habitats (Adapted from Bohlen 2002.)

Ecosystem

Type of organic matter

Amount consumed or incorporated (kg ha−1 yr−1)

Maize field

Maize residues

840

Orchard

Apples leaves

2000

Mixed forest

Canopy tree leaves

3000

Oak forest

Oak leaves

1071

Alfalfa field

Alfalfa residues

1220

Tallgrass prairie

Total organic matter

740–8980

Savanna

Total organic matter

1300

## 3.4.2 Physical engineering of the soil structure

Soil fauna also affect ecosystem properties by acting as ecosystem engineers (Lavelle et al. 1997), substantially modifying the physical structure of the soil profile, and hence the habitat and activities of other organisms and the passage of materials through soils. Engineers tend to be macrofauna, which exploit plant litter at the soil surface, such as earthworms and termites, and create macropores and channels as a consequence of their feeding and burrowing activities. Macrofauna also improve soil porosity and drainage through their burrowing activities. For example, one of the main effects of earthworms on soil porosity is to increase the proportion of macropores and channels in soil (Knight et al. 1992; Lavelle et al. 1997). These structures collectively act to enhance the movement of both water and soluble nutrients through soil and to interconnected waterways. Indeed, studies that have eliminated earthworms from pasture, through the use of pesticides, (p.79) have observed dramatic decreases in water infiltration of up to 93% (Sharpley et al. 1979). Such effects of earthworms on infiltration, especially by deep-burrowing species, are known to increase leaching of nutrients from soil to groundwaters. Studies of grain-crop agro-ecosystems, for example, showed that the inoculation of soils with earthworms produced a 4- to 12-fold increase in leachate volumes and a 10-fold increase in N contained within them (Subler et al. 1997). Similarly, the inoculation of earthworms (Aporrectodea caliginosa) into limed coniferous forest soils produced a 50-fold increase in the concentration of nitrate and cations in soil solution (Robinson et al. 1992, 1996).

Nests of soil-nesting ants provide extensive macroporosity to the soil that affects infiltration rates of water and soluble nutrients. Bulk flow along nest galleries also provides an important route for recharge of deep soil moisture in arid and semi-arid environments. For example, in semi-arid Western Australia, ant biopores were found to transmit water down the soil profile when the soil was saturated and water was ponding on the surface (Lobry de Bruyn and Conacher 1994). Similarly, water infiltration into soils with nest entrances of funnel ants (Aphaenogaster barbigula), which can reach densities of up to 88,000 ha–1, averaged 23.3 mm min–1 but was only 5.9mmmin–1 in nest entrance free soil (Eldridge 1993). Although not quantified, the above effects of ants on water infiltration are likely to have strong influences on transfers of water and nutrients to groundwaters, and also on the movement of water and nutrients to adjacent ecosystems (Bardgett et al. 2001a).

Soil fauna can also indirectly affect the hydrology of ecosystems by altering rates of organic matter accumulation. A fine example of this comes from subantarctic Marion Island, where the soil-borne larvae of a flightless moth, Pringleohaga marioni, process some 1.5 kg plant litter or peat m–2 yr–1, thereby stimulating N and P mineralization 10-fold and 3-fold, respectively (Smith and Steenkamp 1990). Recently, the house mouse, Mus musculus, was introduced onto the island, which feeds on the moth larvae, thereby reducing annual litter turnover by some 60% of that originally processed by the moth (Smith and Steenkamp 1990). According to Smith and Steenkamp (1990), if mouse numbers continue to increase, rates of peat accumulation will increase, which in turn will dramatically alter the hydrological regime and vegetation of the island.

# 3.5 Functional consequences of biological diversity in soil

In recent years, there has been an explosion of studies that examine the ecosystem consequences of declines or changes in biological diversity, and these studies of diversity–function relationships have spurred considerable debate: one group of workers report that ecosystem productivity is positively (p.80) related to species diversity, whereas the other group argues that there is no relationship between biodiversity and productivity, and that ecosystem function is explained by species identity rather than diversity per se (i.e. the absolute number of species with the community) (see Loreau et al. 2002). Most studies on diversity–function relationships, however, have been done on plant and aquatic communities, with the issue of how soil biological diversity affects ecosystem properties receiving only minor attention. This section considers the potential importance of soil biological diversity, and asks whether changes in the diversity of soil animal communities could significantly affect ecosystem properties.

A very common view amongst soil ecologists is that since species richness in soil is so high (Chapter 2) and there are a large number of trophically equivalent organisms, most species must be functionally redundant, that is, they are replaceable with other species without influencing general soil functions, such as nutrient and C mineralization (Andrén et al. 1995; Lawton et al. 1996; Groffman and Bohlen 1999; Setälä et al. 2005). In other words, the loss of individual species from the soil community will not result in a change in ecosystem properties, since another species will replace its role. An opposing view, however, is that this redundancy in soil is largely assumed without evidence (Behan-Pelletier and Newton 1999) and that within trophic groups differences among species are sufficient to produce species-specific impacts on ecosystem functioning (Wall and Virginia 1999).

Support for either of these ideas regarding the importance of diversity in soil is scarce. However, one study of special note is that of Laakso and Setälä (1999b). These authors used model ecosystems to manipulate trophic group structure (comparing treatments composed of fungal-feeding and microbi-detritivorous animals, and their combination), and species richness (one versus five species) within trophic groups, and examined effects on nutrient mineralization and plant growth. It was found that trophic group identity, and richness, and species richness within trophic groups, had no detectable effect on either nutrient mineralization or plant growth. The study also showed that compositional effects of soil faunal species within trophic groups were far less important than across trophic group effects, and that certain species, in particular the enchytraeid worm Cognettia sphagnetorum, had a functionally irreplaceable role in the decomposer system. Similar conclusions also arise from a modelling study by Hunt and Wall (2002). They used a simulation model based on short-grass prairie, that included 15 functional groups of microbes and soil fauna, and they found that while the ‘whole community’ was important for maintaining plant productivity in these models, the deletion of no single faunal group affected this measure significantly. Furthermore, deletion of only three functional groups affected N mineralization by up to 10%. This result was taken to suggest that ecosystems could sustain the loss of some functional groups with little decline in ecosystem function, due largely to compensatory (p.81) responses in the abundance of surviving groups. Together, therefore, these studies support the notion that there is a high degree of functional redundancy in soil food webs at the species level—supporting the redundant species hypothesis—and also that some soil biological processes are driven by particular animal species that are functionally irreplaceable. Few studies have examined the role of microbial diversity as a driver of soil processes, although a study by Setälä and McLean (2004) tackled this issue by examining how the number of fungal taxa (using a gradient from 1 to 43 taxa) influenced decomposition of raw forest humus in microcosms. They found that decomposition, measured as CO2 production, increased with increasing fungal diversity, but only in the species-poor end of the gradient, again suggesting a considerable level of functional redundancy among decomposer fungi.

Another microcosm study by Cragg and Bardgett (2001) came to similar conclusions, that is, differences in composition of the food web are more important than changes in diversity per se. These authors manipulated species richness and composition within a model soil community of up to three species of fungal-feeding Collembola. They showed that, when in monoculture, individual species differed markedly in their effect on plant litter decomposition and nutrient release, but the effects of two and three species combinations on these processes were due to differences in the composition of the collembolan community, and in particular the presence of one, fast-growing collembolan, rather than the number of species present (Cragg and Bardgett 2001). In another microcosm experiment, Cole et al. (2004) examined the effects of individual species of microarthropods, and variations in microarthropod diversity of up to eight species, on soil microbial properties and plant uptake of an added organic 15N source (glycine). They detected effects of increasing species richness of microarthropods on microbial biomass and mycorrhizal colonization of plants, but no effects on plant 15N uptake. However, these effects on microbial properties of soil could be interpreted in relation to the presence of individual species, rather than diversity per se. Together, the findings of these studies suggest that soil processes are mainly driven by the physiological attributes of the dominant animal species present. This supports the notion that the effects of declining species diversity within a trophic group on soil processes are likely to be idiosyncratic, depending on which species are removed from the community. Such findings concur with studies of above-ground communities which point to the role of vegetation composition and dominant plant species as a major driving force of ecosystem function at local scales (Grime 1997; Hooper and Vitousek 1997; Wardle et al. 1997a).

Other studies, however, do point to important effects of faunal diversity on soil processes. For example, Liiri et al. (2001) found that total N uptake and growth of birch trees increased asymptotically with increasing richness of microarthropod species, an effect that was presumably due to increased soil (p.82) N availability with increasing species richness. In the same experiment, however, increasing species richness did not modify the stability of the ecosystem, measured as the effect of drought on birch growth. Griffiths et al. (2000) also found some evidence to suggest that alteration of soil biological diversity can affect soil processes. These authors reduced total soil biodiversity using varying time periods of chloroform-fumigation, and found that certain measures, such as nitrification and methane oxidation, were reduced as a consequence. In contrast, other measures of biological function in soil increased with fumigation time (e.g. N mineralization), or were not affected (e.g. decomposition), suggesting that effects of fumigation and changes in soil biological diversity, are process specific. It was also shown in this study that the biological community of the fumigated soil was less resistant to a second stress (i.e. copper addition). In a related experiment, Griffiths et al. (2004) manipulated soil microbial community structure by re-inoculating sterile soils with different levels of dilution of a non-sterile soil suspension, resulting in a gradient of biodiversity. They then subjected these soils to heat and copper stress, and found that less biologically diverse soils (i.e. those receiving the most diluted inoculum) were the least resistant to these stresses. While these studies provide evidence for diversity effects, it has been argued that the methods of manipulation used make the data susceptible to covarying diversity factors that confound the interpretation of results (Griffiths et al. 2000; Mikola et al. 2002). In other words, it is difficult, if not impossible, to disentangle the effects of diversity from covarying factors on ecosystem processes; for example, chloroform-fumigation and dilution procedures selected for certain species and thus cannot separate species composition effects from species richness effects (Griffiths et al. 2000).

Heemsbergen et al. (2004) also detected soil biodiversity effects on decomposition processes in laboratory studies, but found that they were explained by the functional dissimilarity of component species, rather than by species number. In microcosms, these authors found no conclusive effects of varying species number of macro-detritivores on a range of decomposition processes, such as mass loss, respiration, leaf litter fragmentation, and nitrification. However, they detected a significant positive relationship between soil respiration and litter mass loss, against the functional dissimilarity of the animal community, defined as the degree to which component species are functionally different in those processes, determined from their performance in monocultures. These effects were attributed to facilitative interactions between species, which were greater when component species differed in their functional role. For example, facilitation was detected in communities that contained the earthworm L. rubellus, which is especially effective in transporting litter, which has been fragmented by isopods and millipedes, deep into soil where it is subject to microbial attack. In contrast, inhibition of decomposition occurred between the functionally similar Oniscus asellus (Isopoda) and Polydensmus deticulatus (Diplopoda), suggesting (p.83) competition for leaf litter between these species. Overall, these findings suggest that it is the degree of functional differences between species that is a driver of ecosystem processes, rather than species number. Therefore, to predict the consequences of species loss in soil requires an understanding of how individual species contribute to multiple species interactions in the community (Heemsbergen et al. 2004).

Apart from the above study, there appears to be little support in the literature for a predictable relationship between soil biological diversity and process rates that are crucial to ecosystem function. Rather, most evidence suggests that changes in the abundance of particular species and alteration in the nature of multiple species interactions that occur in soil are the main biotic control of ecosystem function. There is some evidence for a high degree of redundancy within soil food webs—an argument that is bolstered by the prevalence of omnivory in soil food webs (Scheu and Falca 2000; Ponsard and Arditi 2000; Maraun et al. 2003)—but it is also clear that some species are more redundent than others in diverse soil communities.

The absence of consistent diversity effects in experiments is likely to be due to a number of reasons related to the way that they are performed. Most notably, experiments are typically done under very artificial and structurally simple conditions and use a limited range of organisms that vary greatly in their life history and their performance in microcosms. As already discussed, soil food webs in nature are highly complex and involve a multitude of interactions that cannot be observed under simple laboratory conditions. Also, in such a heterogeneous environment as the soil, effects of biodiversity on soil processes will vary greatly in space and time, as will the relative importance of biodiversity to abiotic factors in soil. A study by Liiri et al. (2002) stresses this point; in a three-year field lysimeter study located in a Boreal forest, abiotic factors, such as moisture and temperature, were considerably more important as determinants of soil processes than were variations in the structure and biomass of the faunal community in soil. Similarly, Gonzalez and Seastedt (2001), in a comparative study of controls on decomposition in different ecosystems, found that the role of soil fauna varied between ecosystems, being disproportionately greater in tropical wet forests than in tropical dry or subalpine forests. These findings indicate, therefore, that while diversity and community composition of soil organisms might be important regulators of ecosystem processes at the local scale, abiotic factors such as climate are likely to be stronger regulators at larger regional scales (Mikola et al. 2002). This is analogous to the situation for plant productivity, which is driven mainly by abiotic factors at a larger scale, but is also affected by species composition and species richness at a smaller scale (Huston 1994).

In making the above conclusions it is important to note that the studies that have examined relationships between soil biodiversity and ecosystem (p.84) function have largely considered soil animals, notably detritivores. As a consequence, very little is known about the functional role of changes in microbial community composition and diversity. The general view is that redundancy within microbial communities varies for different functional groups of microbes; influences of microbial community composition and diversity are most likely to be observed for processes that are physiologically or phylogentically ‘narrow’, such as N-fixation or nitrification, whereas ‘broad’ processes such as N mineralization should be insensitive to microbial community composition (Schimel 1995). This view, however, has been challenged by Schimel et al. (2005). These authors argue that ‘broad’ processes of mineralization and immobilization can be ‘aggregated’ into individual components—based on specific enzyme activities and spatial distribution of microbes in microsites—that are sensitive to microbial community composition. In other words, different classes of enzymes involved (p.85) in these processes are produced by different groups of micro-organisms, and different micro-organisms may live and function in different types of microsites. These authors argue that as long as we view N mineralization as a simple ‘aggregate’ process, we will be blind to the specific roles of microbial community composition in regulating N cycling (Schimel et al. 2005). This emphasizes the importance of the need to disentangle the specific functions of microbes involved in important soil functions (Box 3.4).

# 3.6 Conclusions

Microbes have many important roles to play in the ecosystem processes of decomposition and nutrient mineralization, and these activities can determine the availability of nutrients to plants. However, a full understanding of the role of soil biota in driving these processes requires consideration of the interactions that occur between microbes and other organisms within the soil food web. There is ample evidence to prove that the feeding activities of microbial-feeding fauna can greatly influence the abundance and activities of microbes, ultimately affecting nutrient flux. Also, higher-level consumers can have cascading effects on soil food webs that ultimately affect microbes and the processes that they drive. Soil animals also modify greatly the physical structure of soil, thereby affecting the habitat and activities of other fauna and microbes, and also influencing the physical movement of water and solutes through soil. While variations in soil food web diversity have the potential to influence ecosystem properties, experimental evidence indicates that relationships are highly idiosyncratic, and that while there may be a high degree of redundancy within soil food webs, some species are much more important than others. There is emerging evidence, however, that biodiversity effects on soil processes can be predicted by the degree of functional differences among species. Much is still to be learnt about the functional role of soil food web interactions, especially of microbes, and also the relative importance of these biotic forces to abiotic ones as drivers of ecosystem function.