Biodiversity and Conservation 5, 135-168 (1996)

The diversity of soil communities, the 'poor man's tropical rainforest' P A U L S. G I L L E R Department of Zoology, University College, Lee Maltings, Prospect Row, Cork, Ireland

Received 13 January 1995; revised and accepted 29 May 1995

This paper reviews the various factors that facilitate the high biodiversity of soil communities, concentrating on soil animals. It considers the problems facing soil ecologists in the study of soil communities and identifies the important role such communities play in terrestrial ecosystems. The review also considers diversity and abundance patterns. A range of factors are identified that may contribute to the biodiversity of soil and their role is reviewed. These include diversity of food resources and trophic specialization, habitat favourableness, habitat heterogeneity in space and time, scale and spatial extent of the habitat, niche dynamics and resource partitioning, productivity, disturbance and aggregation.

Biodiversity of soil organisms appears high, largely attributable to the nested set of ecological worlds in the soil - the relationship between the range of size groupings of soil organisms relative to the spatial heterogeneity perceived by these various groups - that provide a large 'area for life' for the micro- and mesofauna. The role of aggregation and how it relates to the spatial scale under consideration and to species interactions amongst soil animals is largely unknown at present. The role of disturbance is equivocal and man's activities more often than not seem to lead to a reduced biodiversity of soil communities. This paper also identifies areas where further work is desirable to improve our understanding of the structure and functioning of soil communities.

Keywords: biodiversity; soil communities; resource partitioning; disturbance; heterogeneity: community ecology

Introduction

Fundamental to the understanding and management of natural and disturbed or threatened ecosystems is a description of the biodiversity and an understanding of the processes that promote diversity patterns and control the functioning of communities. However , rather little is known about these features for the soil. Research at the community level has concentrated on component communities of individual taxonomic groups, the 'peripheral ' soil communities associated with dung and carrion and higher latitude rather than low latitude soil communities, especially the very simple soil animal communities of the polar regions. As such, soil community ecology has fallen somewhat behind advances in understanding of other types of communities. In addition, virtually no attention seems to have been paid to conservation activities and the formation and management of nature reserves of below-ground communities and soil habitats in general (Usher, 1988). From both theoretical and applied perspectives, this state of affairs is surprising f rom three points of view.

Firstly, in terms of the importance of soils to global biodiversity. The estimates for total global biodiversity lie somewhere between 5-80 million species (Wilson, 1992).

0960-3115 © 1996 Chapman & Hall

136 Giller

Most of these are invertebrates, most invertebrates are arthropods and most arthropods are terrestrial insects. The majority of terrestrial insects are soil dwellers for at least some stage in their life cycle (Ghilarov, 1977; Behan-Pelletier, 1993). It is not surprising therefore, that most authors who have written on the subject describe soil communities as being amongst the most species rich components of terrestrial ecosystems (Anderson. 1975, 1978a; Ghilarov, 1977: Stanton, 1979). The volume of soil under 1 m 2 of woodland could contain over 200 species of arthropod (Usher and Parr, 1977) and up to 1000 species of soil animals in toto (Anderson, 1975). Mature forest soils appear to have a phylogenetic diversity greater than any habitat other than perhaps coral reefs (Behan-Pelletier and Bisset, 1992).

Secondly, in terms of the ecosystem processes that occur in the soil, 60 to 90% of terrestrial primary production is decomposed in the soil, which thus performs an importan! 'ecological service'. Detritivorous soil animals play a key role in decomposition through fragmentation of litter, cropping of microbial populations thus stimulating their metabolism, inoculation with microbial spores via faeces or carried on the cuticle and other functions (Behan-Pelletier and Hill, 1983; Thomas and MacLean, 1988; Dangerfield, 1990). Macrofauna can process up to 30% of the annual dead organic matter input to most soils (Dangerfield, 1990) and termites alone have been recorded to consume about 20% of litterfall in Ghanaian forests (Usher, 1975). Soil fauna appear to be the major regulatory agents of soil processes affecting the physical and chemical fertility of soils (Behan- Pelletier and Hill, 1983; Lavelle and Pashanasi, 1989; Behan-Pelletier, 1993) forming resistant macro-aggregate structures by ingesting and egesting soil, mixing it with organic residues, digging burrows and transporting soil to the surface as well as vertical transportation of organic matter to deeper soil and synthesis of humus. In fact, there is a strong inverse relationship between mass of accumulated organic material on the soil surface and total soil faunal biomass (Schaeffer and Schauerman, 1990).

Thirdly, soil fauna can offer a suite of bioindicators for classification of soils and detection of disturbances and pollution (e.g. mites, Lebrun, 1979: nematodes, Bongers, 1990: Freckman and Ettema~ 1993).

Soil communities have been described as 'the poor man's tropical rainforest" (Usher et al., 1979). I would agree on three counts - they appear to sustain a relatively high biodiversity, only a proportion of all the species have been described and very little is known about their community structure and dynamics. Why do we know so little about soil communities and how can they sustain such diversity? These are the main topics of this review. I will concentrate on soil animals, particularly the larger-sized animal groups, which reflects both the true bias in the ecological literature and my own limited area of expertise!

The study of soil communities

How can we explain the relative lack of attention given to the study and conservation of soil communities? Usher (1985) points to the 'opaqueness of the system' and to the lack of 'attractive (colourful, cuddly and cute) species'. More crucially, there are severe taxonomic and scale problems which affect our ability to work with soil communities.

Diversity of soil communities 137

Taxonomy

According to Whitford (1992), there are no examples where the soil biota of a specific area of land has been completely described at the species level. As an example, only 53 % of the expected 48500 species of North American soil arthropods have been described (Behan-Pelletier and Bisset, 1992). Some groups are better known than others; beetle, spider and ant species are relatively well known (78% identified) but acarine species are poorly known (< 20% identified). However, the immatures of most arthropods are very poorly known; e.g. only 10% of the North American Coleoptera larvae have been identified (Arnett, 1990). These stages are often the most active in the soil, so this paucity of taxonomic knowledge at the species level poses a real difficulty. Most studies have therefore been based on higher taxonomic levels (family, order or class) or on 'morphospecies' described on the basis of external morphological characteristics (Stanton, 1979; Thomas and MacLean, 1988). Resolution of problems associated with loss of diversity is best served by detailed knowledge of systematics and ecology at the species level. Likewise, little advance in our understanding of soil communities, especially the complex tropical ones, can be made whilst research is based at the higher taxonomic levels (Usher, 1988). As the prospect of significant improvement in our taxonomic knowledge for many parts of the world is not great for the foreseeable future, a possible alternative approach may lay in the recognition and utility of functional groups among soil biota (see later, Table 3). At least this approach relates to ecological rather than simply morphological similarity amongst taxa.

Scale and abundance

Soil fauna can be divided into microfauna (including protozoa, some nematodes and other minor phyla); mesofauna (including mites, Collembola, nematodes, primitive insects, Enchytraeidae); and macrofauna (including other insects, Myriapoda, Lumbricidae, Crustacea and gastropods) on the basis of size (Luxton and Peterson, 1982; Whitford, 1992). The small size of the majority of soil organisms (most components of the mesofauna for example are 160-1000ktm in length (Behan-Pelletier, 1993)) obviously contributes to their difficulty of study and our lack of knowledge. There is also a clear negative correlation between body size and population density (Table 1).

Definition and delimitation of soil communities

A community can be defined as a group of organisms, generally of wide taxonomic affinities, occurring together, many of which will interact within a framework of horizontal and vertical linkages (Giller and Gee, 1987). Recognition of a community requires some notion of a unique collection of species and some concept of boundaries delimiting one community from another. This is more readily possible in dung and carrion communities, but has proved difficult in soil communities proper. Multivariate techniques, especially ordination, can separate soil community types, but mostly at a large spatial scale. Clear differences in mesofauna are apparent between distinct, strongly contrasting, soil types such as mull and mor forest soils (from deciduous and coniferous forests respectively) or mor and mor/moder forest soils (Curry, 1978; Teuben and Smidt, 1992). There are also clear differences in soil macrofauna between forested and grassland systems and other types of land use (Fig. 1; Lavelle and Pashanasi, 1989). Similarly, nematode component communities (sensu Giller and Gee, 1987) separate through canonical discriminant

138

Table 1. Density range per m ~ of selected soil faunal components across the range of size classes. (Data from Anderson, 1978a; Peterson, 1982: Lynch et al., 1988: Thomas and MacLean, 1988)

Fauna Density range m :

Microfauna Protoza t x 107-9 × 10 ~ Nematodes 8.1 × 11)C3 × l&

Mesofauna Mites 5 x 103-1.7 × 10" Collembola l x 102-6.7 x 11) ~

Macrofauna Large oligochaetes 10-500 Diptopoda 10-100 Aranae 400-800 Formicidae 5 × 10C7 x 1() ~ Other insects 5 x 10 ~

Gil ler

analysis into clusters of similar agricultural management (Freckman and Ettema, 1993). Studies of simple Antarctic communities illustrate the difficulty in the identification of representative communities. PCA analysis has shown the importance of vegetation (mosses and lichens) for microarthropods in these harsh conditions, but there are no significant changes in species composition between clusters, only in the relative abundance of species (Usher et al., 1982). More detailed analysis of Prostigmata mites at 159 locations in maritime Antarctica failed to yield discrete PCA clusters that could be interpreted as representative communities (Usher, 1988). Separation of multivariate species clusters (involving 17 taxa) was evident on a larger geographical scale in this study, which mirrored the results of a world-wide analysis of 1479 species of Collembola (Blackith and Blackith, 1975).

The influence of vegetation in the separation of soil communities is ambiguous. On grassland soils, there appears to be no relationship between vegetation type and microarthropod community structure (Curry, 1978). In Norwegian forests, the relationship between Coltembola and vegetation becomes stronger as environmental conditions become extreme (Hagvar, 1982) and the best examples of plant-soil microarthropod associations seem to come from impoverished soils such as blanket bog and steppe (Curry, 1978).

We are thus left with a complex picture, where represenlative soil communities can be identified between strongly contrasting soils or in extreme and/or impoverished soils, and clear biogeographical differences occur, but on a more local scale, there are difficulties in delimiting soil communities. Nevertheless, analysis of patterns and processes that relate to diversity are, of necessity, based on the community concept and level of organization.

Diversity of soil communities 139

) [ ] Pontoscolex coreth 3. 1 [ ] other earthworms

• c ~ f (20) 0,,3~" @ ~ I []~] termiteSants

TS 5.5 [ ] Coleoptera 7 .6 (23) ~ ~ [ ] Myriapoda (18) h~S~~'J~"C"'~""""':-':'Z'~;:~x I~ Arachnida

/ i f; P K c x & (3z) (22)

~ . . ' : ~ ~!:J ~. /

~ ~PF .........~.. j 159.2 s 3.9 (27)

t : ° ~ s r s (41) ~,s.~ ~3,~ ~ s 9

Figure 1. The distribution of biomass among the main soil macrofaunal taxa in different types of land use in the humid tropics of Peru. The relative position of the land use types relates to their location in relation to the first two PCA axes. Size of circles is proportional to the biomass. Figures represent mean biomass (g m 2) and taxa richness (in brackets). Land use type codes are as follows; PF - primary forest; SF - secondary forest; LI - low input cultivation (rice); TS - traditional system (cassava); HI - high input (maize); K - Kudzu fallow; PK - peach-palm + Kudzu; C - Centrosema pubescens pasture; BD - improved pasture. (After Lavelle and Pashanasi, 1989.)

Diversity and community ecology

It can be argued that one of the most important developments in community ecology has been the emergence of a structural model of communities based on the Hutchinsonian concept of the niche (Giller and Gee, 1987; Primark, 1992). This model provides the basis for a large body of theory which regards the community as a notional space filled to saturation with niches, each overlapping its neighbour to a limited extent (Giller, 1984). The saturated community represents a state of equilibrium, governed largely by competit ive interactions between species, and the structure of such communities should be predictable. However, some of the testable predictions found little support in certain communities and in others the support was dependent on the temporal and spatial scale of analysis. Non-equilibrium communities have also been described, where species abundances change continually in relation to changing environmental conditions or lottery dynamics (Hanski, 1991). A variety of views have identified different processes as influential in shaping the structure of one community or another (Cornell and

140 Giller

Lawton, 1992). This diversity of views leads to the worrying prospect that each community might be unique in structure, moulded individually by one or more of a wide range of factors. The structure of such communities might be entirely unpredictable. The hope for a predictive science of community ecology thus hinges on the possibility that communities might be rationally arranged on a restricted number of axes that are related to some measurable characteristics of the organisms or their environment (Giller and Gee, 1987) Thus communities should be thought of as lying somewhere along a gradient of equilibrium to non-equilibrium states which vary in the relative importance of various characteristics (Fig. 2).

Where do soil communities sit on this gradient? In brief, it is not at all clear. Much work has been conducted on simple high latitude soil communities, various components of more complex ones or on species pairs. We do not know to what extent these results can be generalized to an understanding of more complex (and whole) communities (Usher. 1985 ). Peripheral soil communities in dung and carrion have received considerable attention at the community level, based as they are on easily identifiable resources that are delimited in space and time. Some evidence would tend to put such communities towards the non-equilibrium end of the gradient (Doube, 1987; Hanski, 1991) although considerable levels of species interaction occur, especially in low latitude systems (see later). The problems of the nature of resources in soils, taxonomic constraints and scale of approach have made it difficult to integrate the study of soil communities into the more general ecological theories at the community level.

------• St ructural stabi l i ty

Ratio of disturbance per iodic i ty to communi ty

response t ime ~ Ac.cnmmc

Opportunist ic - - exploi tat ion [

Predictabil i ty of species composi t ion

- - 3

Openness of communi ty I v=

- - ~ Periods of resource . _ , J l imitat ion

~ lmportance of species interactions

Importance of ~ stochastic events

Scale of analysis

Non-equi l ibr ium ~ ~ Equi l ibr ium

Figure 2. The relative importance (thickness of bars) of various characteristics of communities along a continuum from a non-equilibrium to an equilibrium state. Shaded bars indicate the main characteristics discussed in this review. (After Giller and Gee, 1987. I

Diversity o f soil communi t ies 141

Diversity and abundance patterns in soil animal communities

Latitudinal trends of species richness rising towards the tropics are characteristic of nearly all kinds of organisms (Brown, 1988; Wilson, 1992). However, data on latitudinal diversity gradients amongst soil organisms are sparse and often incomplete - examples include mites (Stanton, 1979); Collembola (Peterson, 1982a); ants (Lynch et al., 1988; Wilson, 1992); Scarabaeid dung beetles (Hanski and Cambefort, 1991) and termites (Wood, 1976). Not all groups support the general trend; soil nematodes appear to reach maximum diversity in temperate zones (Bernard, 1992; Huston, 1994). An interesting study using standardized litter samples found o~ diversity (species richness per sample) to be comparable between temperate and tropical samples (Stanton, 1979) but the slightly steeper slope of the species area relationships in the tropical area, especially for deciduous and pine forest habitats, indicated more species would occur in a fixed area than in temperate systems, thus tropical soil systems showed greater [3 diversity (Fig. 3).

Other diversity gradients have been identified with depth in soil (Anderson, 1978a;

50

ffl w

U W

ffJ U. O 10 n-

iJJ w

z

A

f

..s : " 7 " ,, .,.. , , . ,4..,"..

' . . . . . . l b ' ' ' s ' 0 . . . . .

AREA 1 = 6 0 0 c m 2

Figure 3. Species-area relationships for litter mites in three habitat types in tropical (Costa Rica) and temperate (Wyoming) study sites (field/grassland, D - Costa Rica; F- Wyoming; deciduous forest, B - Costa Rica; C - Wyoming; and pine forest, A - Costa Rica; E - Wyoming). (Adapted from Stanton, 1979.)

142 Giller

Brown, 1988), with distance from disturbances and perturbation (Usher and Parr, 1977: Nestel et al., 1993) and along habitat gradients (Peterson, 1982a) such as vegetation type (e.g. nematodes from reed beds to ungrazed pasture), soil types (e.g. for enchytraeids, nematodes and Collembola from mull to mor soils) and soil moisture (for Collembola from moist to dry).

The abundance patterns of soil fauna are clearer; abundance (per unit volume soil) of a whole range of invertebrate groups (e.g. protozoa, nematodes, Collembola, mites and earthworms) (Peterson, 1982a) appears to peak in temperate areas and is far lower in the tropics with a further decrease in tundra and arctic communities. Biomass of tropical soil invertebrate populations are lower and more similar to boreal taiga forests than to temperate deciduous forests and grasslands (Peterson, 1982b). Tropical soil communities thus appear to have more rare species (e.g. in Stanton's (1979) study, 60-75% of taxa were rare in tropical communities sampled, but only 38-66% were rare in comparable temperate ones).

The third pattern is the relatively high degree of dominance (by abundance or biomass) of certain groups within the entire soil faunal community (Table 2). The abundance distribution of nematode species in soil samples is often strongly skewed, with a fe~' dominants and a majority of species at relatively low numbers (de Goede and Bongers, 1994). In several temperate ecosystems, earthworms make up a large proportion of the entire animal biomass e.g. in oak deciduous forests in both Belgium and New Zealand (Brockie and Moeed, 1986). This is common in mull-type soils and mild climatic conditions (Peterson. 1982b: Schaeffer and Schauerman, 1990). More exposed habitats of the coldest climatic zones and moder or mor soils have small biomasses and are dominated by mesofauna (Enchytraeidae, Collembola, Acari, Diptera (Peterson, 1982b; Schaeffer and Schauerman, 1990)). Mediterranean type systems are also dominated by mites and Collembola (Askidis and Stamou. 1991). Sub-tropical arid regions are dominated by protostigmatid mites and termites (Peters(m, 1982b: Silva et at., 1989). Dominance diversity curves within specific taxonomic groups appear to follow a lognormal or logseries distribution in more species rich temperate/subtropical assemblages (e.g. Collembola and dung beetles (Fig. 4)) but show steeper geometric series distributions in simple communities. Lognormal distributions are indicative of a few abundant and relatively fe~' very rare species and a majority of intermediates, whilst geometric series distributions suggest a very strong degree of dominance (Gilter, 1984).

Explanations of the high biodiversity of soil communities

Explanations for high diversity have traditionally been based on the equilibrium concept of community structure and a high degree of resource/niche partitioning (Giller, 1984) with habitat, food and time amongst the most important niche dimensions (Schoener, 1974). The size of soil organisms in relation to spatial and temporal heterogeneity and living area is important in this context. Such resource partitioning and coexistence of species is enhanced through the favourabteness, stability, productivity, heterogeneity and the area of the habitat. If such explanations are valid we might expect specialization, reduced niche overlap and evidence of competition or niche shift when resource levels are limiting or decline. Other explanations of high diversity lend themselves to the non-equilibrium concept of the community and involve disturbance (temporal

Tab

le 2

. Bio

mas

s es

timat

es (

mg

d w

t m

-2)

and

% c

ompo

sitio

n fo

r m

ajor

soi

l ani

mal

gro

ups

in a

ran

ge o

f bi

omes

. Fi

gure

s fo

llow

ed b

y (?

) ar

e te

ntat

ive.

(A

dapt

ed f

rom

Pet

erso

n, 1

982b

)

Tem

pera

te

Tem

pera

te

Tem

pera

te

Tro

pica

l co

nife

rous

de

cidu

ous

Tro

pica

l T

undr

a gr

assl

and

gras

slan

d fo

rest

fo

rest

(m

ull)

fo

rest

Mes

ofau

na

Nem

atod

e 16

0 (4

.8%

) 44

0 (7

.8%

) 50

(?)

(2.

6%)

120

(5%

) 33

0 (4

.1%

) 50

(?)

(2.

8%)

Enc

hytr

aied

ae

1800

(54.

5%)

330

(5.4

%)

20 (

?) (

1%)

480

(20%

) 43

0 (5

.4%

) 20

(?)

(1.

1%)

Col

lem

bola

15

0 (4

.5%

) 90

(1.

6%)

10 (?

) (0

.5%

) 80

(3.

3%)

110

(1.4

%)

20 (

?) (

1.1%

) C

rypt

ostig

mat

a 60

(1.6

%)

110

(2%

) 20

(?)

(1%

) 45

0 (1

8.8%

) 18

0 (2

.2%

) M

esos

tigm

ata

20 (0

.6%

) 60

(?)

(1%

) 10

(?)

(0.

5%)

80 (

?) (

3.3%

) 40

(0.5

%)

Pros

tigm

ata

10 (0

.3%

) 40

(?)

(0.

7%)

50 (

?) (

2.6%

) 30

(?)

(1.

2%)

10 (?

) (0

.1%

) T

otal

aca

ri

100

(?)

(5.5

%)

Mac

rofa

una

Lar

ge

Olig

ocha

eta

330

(10%

) 31

00 (5

5.3%

) 17

0 (8

.9%

) 45

0 (1

8.8%

) 53

00 (6

6.2%

) 34

0 (1

8.9%

) D

iplo

poda

0

(?)

1000

(17.

8%)

10 (

?) (

0.5%

) 50

(2.

1%)

420

(5.2

%)

20 (1

.1%

) D

ipte

ra l

arva

e 47

0 (1

4.2%

) 60

(1%

) 10

(?)

(0.5

%)

260

(10.

8%)

330

(4.1

%)

0 (?

) Is

opte

ra

0 0

(?)

1000

(?)

(52.

6%)

0 0

1000

(?)

(55.

5%)

Chi

lopo

da

20 (

?) (

0.6%

) 14

0 (2

.5%

) 5

(?)

(0.2

5%)

70 (2

.9%

) 13

0 (1

.6%

) 5

(0.3

%)

Car

abid

ae a

nd

Stap

hylin

idae

50

(?)

(1.

5%)

80 (

?) (

1.4%

) 10

(?)

(0.5

%)

120

(5%

) 90

(1.1

%)

10 (?

) (0

.6%

) A

rane

ae

10 (0

.3%

) 30

(?)

(0.

5%)

30 (

?) (

1.6%

) 50

(2.

1%)

40 (0

.5%

) 20

(1.

1%)

Gas

trop

oda

0 (?

) 10

0 (?

) (1

.8%

) 10

(?)

(0.5

%)

20 (

?) (

0.8%

) 27

0 (3

.4%

) 10

(?)

(0.6

%)

Form

icoi

dea

0 (?

) 10

0 (1

.8%

o)

300

(?)

(15.

8%)

10 (?

) (0

.4%

) 10

(?)

(0.

1%o)

30

(?)

(1.

7%)

Tot

al

3300

56

00

1900

24

00

8000

18

00

144 Giller

heterogeneity), predation and aggregation patterns. Here, niche partitioning is not a necessary precondition for coexistence, and niche overlap may appear quite large between coexisting species. In the following sections, I review the role various factors appear to play in explaining the biodiversity of soil communities.

Diversity of food resources and trophic speciafization

The enigma of soil animal diversity, as described by Anderson (1975, 1978a), is that despite the high diversity of soil communities, it is widely reported that most soil species are trophically non-specialists and there appears to be relatively little partitioning of food

!0000

I00

10000

100

a

5 I O f 5 2 0 2 5

b

-,,.,

m w

. . , J i . . . . . - - _

0 30 60 90 120

RANK Figure 4. Rank-abundance curves for Collembola and dung beetles. (a) Five collembolan communities (left to right at the 100 level); simple Signy Island moss-turf community; Lynch island grass sward: temperate Wharram Quarry nature reserve; Scots pine forest; mixed coniferous forest. (b) Data from three successional communities on a heathland (left to right at 100 level) in Eriophorum angustifolium, Eriophorum vagination and Caluna vulgaris. (After Usher, 1985.) (c) Dung beetle assemblage from Mkuzi game reserve in South Africa. (After Doube, 1991: Gaston, 1994.)

Diversity of soil communities 145

type. Such views, however, tend to be based on only a few comparative studies of the whole soil biota (David et aL, 1993). Designation of species to various trophic groups is frequently based on the morphology of mouthparts or the alimentary canal rather than on known feeding habits (e.g. for nematodes, Freckman and Ettema, 1993) and as such possibly provides incorrect classification. Autecological studies on actual feeding behaviour have demonstrated considerable problems inherent in generalizations (e.g. mites, Behan- Pelletier and Bisset, 1992; nematodes, Yeates, 1987). Whilst it is probably true that most species appear to be panphytophagous (consuming a wide variety of dead, higher plant material and microbial flora) (Luxton, 1982; Usher et al., 1982), there are a variety of functional groups that have been identified in various soil communities and component communities (Table 3).

Little opportunity for trophic specialization is likely where non-discrete litter is the food resource. However, preferences and feeding habits may change dependent on the state of decomposition of the medium (e.g. collembolan and mite species (Teuben and Smidt, 1992)). Anderson (1978a) concluded panphytophagous mites become more generalist feeders as availability of fungal hyphae decreases in relation to higher plant material as the litter decomposes. Selectivity of food material may vary seasonally (Luxton, 1982), with habitat (Collembola, Teuben and Smidt, 1992), with microhabitat (CoUembola, Peterson, 1971; Teuben and Smidt, 1992; cryptostigmatid mites, Anderson, 1978a) and with ontogenetic shifts as animals age (mites, Behan-Pelletier and Bisset, 1992). It appears that for many taxa, rigid classification of trophic status and feeding niche is difficult.

For species using discrete food items, such as fungi or seeds, some degree of specialization seems to be common (Fig. 5). Microphytophagous Collembola frequently show clear preferences from amongst fungal species (Mills and Sinha, 1971; Vegter, 1983) and species body size has been shown to correlate with seed size utilized amongst ants (Mehlhop and Scott, 1983; Davidson, 1985). Trophic specialization on the basis of body size has also been recorded for cryptostigmatid mites, where smaller species are able to

o~oo /X Orchesello cincto

• Tomocerus minor / \ / \

P O L L E N /x. X ,~, ~'--" ~ ~ ) \ P L E U R O C O C C U S

I 0 0 ' ~ ~ Y v v v v v v ~ 0 0 20 40 80 80 100

FUNGUS

Figure 5. Gut contents of two species of Entomobryidae Collembola showing relative preferences from amongst three food types. (After Vegter, 1983.)

-Ix

Tab

le 3

. Cla

ssif

icat

ion

sche

mes

for

tro

phic

fun

ctio

nal

grou

ps i

n so

il co

mm

unit

ies

and

com

pone

nt c

omm

unit

ies

Fre

ckm

an a

nd

Scha

effe

r and

L

eadl

ey-B

row

n.

Lux

ton.

198

2 E

ttem

a, 1

99

3

Scha

uerm

an,

Food

typ

e 19

78 (

gene

ral)

(g

ener

al)

(nem

atod

es)

1990

(mit

es)

Beh

an-P

elle

tier

, 19

93 (

gene

ral)

Beh

an-P

elle

tier

an

d H

ill,

1983

(o

riba

tida

e)

Fung

al s

pore

s an

d ba

cter

ia

Gen

eral

hig

her

plan

t de

trit

us

Faec

al f

eede

rs

Car

rion

fee

ders

L

ivin

g pl

ant

tissu

es

Liv

ing

anim

al t

issu

es

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Diversity o f soil communities 147

discriminate among fungal hyphae and fungal species but large mite species cannot (Anderson, 1978b).

Where food resources are in discrete patches (e.g. dung and carrion) selectivity is also apparent. Dung beetles show some degree of preference for dung types (Doube, 1987; Gittings, 1994, see also Fig. 8). Similar preferences are shown by carrion feeders (e.g. Kneidal, 1984). Variation in the way the resource is used can also enhance coexistence on such discrete resources. Seven functional groups have been identified from amongst dung beetles for example (Doube, 1990).

Overlap in food type might be expected to lead to interspecific competition if food was a limiting factor. Schoener (1983) concluded that competition is found very often when it is looked for experimentally. Competition has frequently been inferred in soil communities and in soil arthropod assemblages in particular, but has been demonstrated less frequently (Usher, 1985). Examples include species pairs of Collembola (Kaczmarek, 1975; Longstaff, 1976; Greenslade and Greenslade, 1980; Sparkes, 1982 and Hagvar, 1990), cryptostigmatid mites (Anderson, 1978b) and ants (Culver, 1974; Davidson, 1985). Strong interspecific competition for ephemeral resources of dung in sub-tropical habitats (Anderson and Coe, 1974; Ridsdill-Smith et al., 1982; Giller and Doube, 1989) and carrion in tropical and temperate habitats (Wilson et al., 1984; Hanski, 1987 and Trumbo, 1990) has been readily identified. Most of these studies have indicated strong asymmetric competition (e.g. Fig. 6).

The conclusion to be drawn from the body of evidence is that, whilst some clear specialization is evident on trophic resources, this is probably insufficient to explain fully the biodiversity of soil communities.

Habitat favourableness

On a global scale, environments supporting diverse species appear to have mild, warm conditions with little seasonal variation, that can be tolerated by many species. Depauperate habitats often have harsh and/or unpredictable abiotic conditions that require special adaptations and would be stressful to many species that occur in other habitats (Brown, 1988). In stable, deterministic environments, close species packing and rich speciation is possible (May, 1974; Giller, 1984). Thus some more 'favourable' environments maintain higher biodiversity than others. This concept of habitat favourableness can be readily applied to soil communities. For example, soil is thermally buffered from atmospheric changes in temperature (Whitford, 1992) and temperature variation decreases with depth such that below 20cm little diurnal variation occurs (Jackson and Raw, 1966). Also, at a certain depth, the soil atmosphere is saturated with water, even in deserts, reducing the problem of desiccation (Ghilarov, 1977).

Not all soils offer favourable conditions though. Surficial soils, exposed to direct insolation do exhibit daily fluctuations in both temperature and moisture, but soils with vegetative covering are largely buffered against this. This may explain the especially deep distribution of soil animals in dry natural grasslands of temperate and tropical regions (Peterson, 1982c). Tundra habitats on the other hand, have a strong concentration of organisms near the surface (Peterson, 1982c) possibly due to the occurrence of permafrost at depth. Harsh climates, such as those in the Antarctic, do result in very species poor communities (Usher et al., 1982; Usher, 1985). Regularly flooded forest soils, like arctic ones, offer predictable but harsh conditions, with resultant low diversity (Irmler, 1979).

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Low pH conditions also lead to lower diversity and abundance of soil fauna. Porous, sandy soils with mor-type (coniferous) humus often have high soil acidity. These usually have poor Lumbricid and other saprophagous macrofauna and are dominated by mesofauna. Podzols and mull litter types have high abundance and diversity of earthworms and macrofauna and low mesofauna (Teuben and Smidt, 1992; David et al., 1993).

Habitat heterogeneity in space and time

Not only do soils offer a generally favourable environment, but they also offer an extremely heterogeneous one both spatially and temporally, which can potentially enhance niche partitioning and hence greater species packing.

Diversity of soil communities 149

Spatial heterogeneity. Horizontal heterogeneity is related to gradients in soil type, soil water content and relative humidity, nutrient concentration, texture, pore space, pH, soil atmosphere, depth of litter layers, overlying vegetation, temperature variation and degree of bioturbation by macrofauna (Jackson and Raw, 1966; Usher, 1976; Ghilarov, 1977; Leadley-Brown, 1978; Schaefer and Schauerman, 1990; Teuben and Smidt, 1992; Huston, 1994). Soil gradients occur on macro- and mesoscales from continents, mountain slopes and hillsides to around individual plants and animal droppings (Huston, 1994). On a microscale, size of pore space, soil texture in terms of size and shape of soil particles and aggregates of particles (crumbs) and ratio of air to water in pore space vary within and between soil types (Jackson and Raw, 1966; Leadley-Brown, 1978). Vertical heterogeneity is also related to soil texture, structural diversity of litter layers, pore size and atmosphere (Jackson and Raw, 1966; Anderson, 1975; Ghilarov, 1977).

Dung and carrion also offer considerable structural heterogeneity, dependent largely on size. Corpses of large animals contain a variety of potential food resources including skin, sinew, horn, bone and various body tissues, providing a wider range of niches for specialists than smaller corpses (Doube, 1987). Dung beetle species can also differ in location within a large dung pat (Holter, 1982; Gittings, 1994). On a macroscale, dung beetle species show preferences for habitat, soil type and vegetation cover (Doube, 1987; Hanski and Cambefort, 1991; Giller and Doube, 1994) as well as different types and ages of dung (Doube et al., 1988; Gittings, 1994). A greater diversity of animals depositing dung or providing carrion will naturally lead to greater diversity of coprophagous and necrophagous species in the habitat.

Spatial heterogeneity clearly provides a major explanatory factor for high biodiversity in soil communities. The ecological literature is full of examples showing increased diversity in more heterogeneous environments (Brown, 1988) although studies on soil communities are rare enough. Figure 7 provides one such example where diversity of soil litter mites is positively correlated with habitat structure measured on the basis of 24 microhabitat categories including size of leaf and wood fragments, type of roots and fungi, animal faeces and remains and size of soil pore cavities (Anderson, 1978a).

Temporal heterogeneity. Clear seasonal differences in activity have been described for a wide range of soil animals including nematodes and Enchytraeidae (Peterson, 1982c), Collembola (Vegter, 1983), neo-tropical grassland Coleoptera (Villalobos and Lavelle, 1990), Scarabaeid dung beetles (Holter, 1982; Gittings, 1994), temperate woodland Carabid and Staphylinid beetles (Dennison and Hodkinson, 1984) and temperate and tropical ants (Lynch et al., 1980; Levings, 1983). These seasonal patterns may be associated directly with regular changes in weather or other environmental conditions (Irmler, 1979; Levings, 1983; Schaefer and Sehauerman, 1990; Villalobos and Lavelle, 1990) or may represent 'active temporal partitioning' to avoid competitors. There is little experimental evidence for this latter notion, but further studies on seasonal activity patterns along diversity gradients, notably in areas where widespread species are co-occurring with fewer contemporary species than in other areas, would be useful to test for shifts in activity patterns in the absence of potential competitors.

Considerable information is available on seasonal variation in dung communities. Aphodius dung beetle species from temperate grasslands of NW Europe are fairly well spaced out along the seasonal axis (Holter, 1982) suggesting overdispersion of niches (as a result of species interactions), but seasonal distribution of metabolic activity is more

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uneven than seasonal species frequencies in this case. Increasing evidence suggests that seasonal and annual changes in populations of tropical animals, including arthropods, are as large as those in temperate ones (Levings, 1983: Wolda, 1987). However, in more sub-tropical pastures, annual rainfall patterns confine most dung beetle activity to certain temporal windows, thus preventing significant temporal partitioning on a seasonal axis (Giller and Doube, 1994). The role of seasonal partitioning is thus equivocal in terms of the community ecology of dung beetles at least.

Differential activity patterns in relation to short term changes in conditions is evident in some soil animal groups, but most information again comes from soil surface and dung fauna. Dung beetle activity, for example, varies in relation to local weather conditions in both temperate (Gittings, 1994) and sub-tropical (Giller and Doube, 1994) grasslands. Diurnal partitioning is also seen amongst dung beetles (Doube, 1987; Hanski and Camberfort, 1991 ) and seed harvesting ants (Levings, 1983, Mehlhop and Scott, 1983) with clear nocturnal, diurnal and crepuscular forms. This is considered to contribute to partitioning of limited resources, but can do so only if resource use during one time period does not influence resource availability during another.

Successional patterns are well established for soil communities (Usher and Parr, 1977)

Diversity o f soil communities 151

and contribute to temporal niche partitioning. Successional changes in nematode communities from colonizer taxa to more specialized late successional species have been documented in the longer term (several weeks to months) following disturbances from a variety of anthropogenic factors (Bongers, 1990; Ettema and Bongers, 1993). More natural succession in nematode communities involved an increase in abundance and diversity following succession from bare sand to forest soils (De Goede et al., 1993). Similar successional changes in relation to overall changes in habitat have been shown for termites following clearcutting of forest (Usher et al., 1982; Usher, 1988) and mites and Collembola on natural grasslands (Parr, 1980) or artificial soils (Huhta et al., 1979; Hermosilla, 1982).

Successional changes can also occur in the shorter term in relation to changes in the nature of resources, as litter and ephemeral resources decompose. For example, as boreal litter decomposes, a gradual replacement of bacteria and yeast by filamentous fungi occurs, which is mirrored by a sequential pattern of microarthropod colonization (Silva et al., 1989). On a smaller scale, plithiricorid mites living within conifer needles show successional change (Hayes, 1966). Dung beetles also show some degree of successional change as the dung pad ages, related to changes in the physico-chemical characteristics (in terms of moisture content, fibre and organic matter, as well as the size of the pad) which leads to some turnover of species that show clear preferences for different aged dung (Gittings, 1994). The difference in species composition over time has been clearly demonstrated in field successional experiments on different dung types using Canonical Corresponding Analysis (CCA) analysis. Time is a major environmental parameter explaining differences in the distribution of dung beetle species amongst dung sampled over time (Fig. 8a). Clear early, mid and late successional species of Aphodius beetles can be identified (Fig. 8b).

Whether all these successional patterns are related to the amelioration of species interactions over evolutionary or ecological time or are simply the result of physiological differences is not known at present.

Finally, there will also be stochastic temporal variation in resource levels due to year to year differences in standing crops of litter and variation in the timing of leaf fall etc. Spatial variation in soil organic matter or litter accumulation (dependent on wind conditions, plant distribution) and variation in ephemeral resources like dung and carrion in space and time will add to the overall spatial and temporal heterogeneity of soil habitats. With a mosaic of patches of detrital resources and, at a larger scale, of vegetation and soil patches at different stages of development, there is enhanced heterogeneity in resources and conditions associated with the soil environment. This should, in turn, lead to greater biodiversity within the area as a whole. In this way, the soil mirrors the patchy nature related to high biodiversity of tropical rainforests (Connell, 1978) and relatively species rich deep sea habitats (Grassle and Grassle, 1994).

Scale, spatial extent and environmental heterogeneity. Scale is critical to understanding heterogeneity (Huston, 1994). Levins (1979) suggested that generation time (and hence size) is per sea niche component in a variable environment, thus the scale of spatial and temporal heterogeneity may separate not only coexisting species of different size, but also separate different communities within a certain area (Fenchel, 1987). The physical space of soils can potentially harbour a hierarchy of communities overlapping to some degree (perhaps along the lines of the micro- meso- macrofaunal categories spanning several orders of magnitude of size), each representing a few trophic levels and characteristics of

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scales of time and space. This size-based partitioning is likely to be related to pore size, water films surrounding particles and food sources in soils (Yeates, 1987). In fact, spatial and temporal heterogeneity of the soil will increase in relation to decreasing size of soil organisms.

Natural objects have a graininess or nested irregularity to them, which allows them to be considered in terms of fractal geometry (Sugihara and May, 1990). Based on theoretical considerations viewing nature as fractal and on the concept of the species area relationship, the number of species is predicted to increase in an area as heterogeneity increases (Williamson, 1988). The species area relationship is described by Log S -- zLogA. Area is in fact related to the square of the length; Log A -- 2LogL. L, the length, varies in relation to the so called fractal dimension of the object in question. This has been illustrated by considering the length of some coastline on a map. The length is given by the number of times (N) a 'ruler' of length l can be stepped around the coastline, i.e. N × 1. As I is reduced, the steps reflect more of the detailed structure of the coastline, so the total length L increases (Gaston, 1994). The relationship between N, L and a constant E is given by E = Nl D (Mandlebrot, 1983), where E is the fractal extent and D the fractal dimension of the coastline. For a smooth curve, D = 1 and D increases with increasingly complex boundaries towards a maximum of 2. Where D = 1.5, a ten-fold decrease in measurement scale will increase the apparent length of the object by a factor of 10 °5, i.e., 3. In the context of the complex boundary of a leaf for example, an order of magnitude decrease in measurement scale will increase the perceived surface area by 3.16-10 times (Sugihara and May, 1990). Thus for organisms, a decrease in body length by one order of magnitude will result in the equivalent of 3.16-10 times more available living space.

This theoretical foray is instructive, as the same concept can equally well apply to soil organisms. Here, the available living space lies on the surfaces of soil particles, in pore spaces and along the length of passageways/tunnels between solid particles (Ghilarov, 1977). As mentioned above, diversity increases with area; this species-area effect is largely due to the underlying heterogeneity of the physical environment (Brown, 1988; Williamson, 1988). As the heterogeneity will increase with a decrease in the scale of resolution of measurement, so diversity is predicted to increase as size of organisms decreases. In the complex fractal world of the soil environment, there is effectively greater niche space for the smaller organisms that characterize soil habitats and hence a greater opportunity for specialization, species packing and increased diversity. At the smallest scale of resolution for example, 4000-5000 bacterial species can be found in a single gram of beech forest soil (Goksyr and Torsvik, 1990, in Wilson, 1992). Scale and spatial extent

Figure 8. Canonical correspondence analysis of representative experiments on the distribution and succession of temperate pasture dung beetles amongst dung pats of different types. (a) and (b) show CCA biplots of species arrangements in relation to successional occurrence and dung type (from a wildlife park; gi - giraffe; zeb - zebra; cow) and dung age (Day). The successional gradient is represented by the vector 'Day' and the centroid for each dung type is shown. (c) and (d) show relative successional occurrences of various species along the 'Day' vector in the CCA biplots. (After Gittings, 1994.) Species codes as follows; sphac - Aphodius sphacelatus; prod - A. prodromus; scara - Sphaeridium scarabaeoides; dep- A. depressus; errat- A. erraticus; lun - S. lunatum; ater- A. ater; loss- A. fossor; tim - A. fimetarius; spin - Geotrupes spiniger. Rare species are indicated by numbers (a) - 1, S. bipustulatum; (b) - 1, A. ater; 2 - S. bipustulatum.

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Diversity o f soil communities 155

can thus explain a considerable amount of the diversity that characterizes soil communities.

Niche theory in soil communities

Traditional niche theory would predict that greater biodiversity should result in more specialized niches (smaller niche width) and a greater degree of resource partitioning (reduced niche overlap) (Giller, 1984). This was discussed for food resources in an earlier section. In addition, in structurally diverse habitats, coexisting species often do appear to specialize and avoid competitive exclusion by differential use of the physical structure (Brown, 1988). For soil communities, niche width appears to be variable amongst soil arthropods, with both specialists and generalists often within the same taxonomic group coexisting within an area (Kaczmarek, 1975; Usher, 1976; Curry, 1978). In addition, whilst most information to date indicates some degree of niche overlap (e.g. Usher, 1976; Krishnamorthy, 1985) there are many clear examples of niche partitioning amongst soil organisms, often on two niche axes (i.e. differential niche overlap).

Two convincing examples of differential niche overlap are shown in Fig. 9. Collembolan species (Fig. 9a) appear to separate in terms of horizontal moisture gradient and vertical distribution. Other studies of collembolans along moisture gradients indicate an even spacing of cuticular resistance to transpiration amongst species (Vegter, 1983) suggestive of overdispersion of niches in niche space. In the oribatid mite example (Fig. 9b), the two species Northrus silvestris and Platynothrus peltifer maintain high population densities at the same sites in the same sample station at the same time. However different life stages of the species show differences in relation to pH and organic matter contents. There are also some food preferences; N. silvestris for fungi and decaying oak leaves, P. peltifer for remains of green plants. These differences are associated with differences in carbohydrase activity (Siepel, 1990). Amongst macrofauna, granivorous ants are often found to partition seed resources on the basis of size and species of seed (Mehlhop and Scott, 1983; Davidson, 1985) as well as seasonal and vertical partitioning (Lynch et al., 1980). Niche differentiation has also been recorded in earthworms in respect of vertical zonation in relation to organic matter and soil moisture (e.g. Krishnamorthy, 1985) Aphodius dung beetles partition dung along two dimensions of time (season and age of dung pats) and two dimensions of space (amongst pats and within pats) (Holter, 1982). To this one should add the preferences for soil type, vegetation and dung type as discussed earlier.

Body size can also represent an important niche dimension. For coexisting granivorous ants, niche overlap was consistently lowest between pairs of species with the greatest difference in body size (Mehlhop and Scott, 1983). The ratio between successive sizes of species within a nematode assemblage ranged between 1.11-1.4 with a mean of 1.285

Figure 9. Examples of niche partitioning amongst soil animals. (a) Relative distribution of Collembola species in relation to depth and soil moisture. Shaded circles refer to habitat generalists and the open circles to habitat specialists. Circle area is related to density of the species. (After Kaczmarek, 1975). (b) Canonical correspondence analysis ordination for adult, nymphal and larval stages of two species of soil mite Northrus silvestris and Platynothrus peltifer in relation to six physico/chemical parameters (organic matter content, pH~:cL, phosphate (P) content, potassium (K) content and C:N ratio of organic matter. (After Siepel, 1990.)

156 Giller

(Yeates, 1987) extremely close to Hutchinson's rule for limiting similarity between coexisting species (Giller, 1984). There is however, relatively little other work of this kind amongst the soil fauna.

Indirect evidence of the role of species interactions and resource levels in niche dynamics has been found amongst soil arthropods. Clear niche shift was seen in distribution patterns and diets amongst a pair of mite species which appeared to decrease overlap in the vertical dimension in mixed compared with single species cultures (Anderson, 1978b). Niche differences also increased over short periods if resources became limiting. Similar changes have been documented for ants. When food levels (in the form of small seeds) decreased, medium sized ant species showed an increase in niche width but a decrease in niche overlap with smaller ant species (Mehlhop and Scott, 1983). Amongst earthworms, high spatiotemporal overlap has been recorded amongst species when worms are in a quiescent non-feeding state, but such overlap was only found amongst one species pair during periods when the worms were active (Krishnamoorthy, 1985). On a larger scale, the increase in dominance of nematodes with increased environmental harshness as one moves further from the equator has been attributed to a form of ecological release due to the decrease in coexisting arthropods and other fauna (Proctor. 1990).

Pianka's (1978) niche overlap hypothesis suggests that as species richness and hence potential diffuse competition increase within a community, so pairwise niche overlap should decrease. There is some evidence that the degree of niche overlap varies inversely with number of coexisting species of earthworms (Krishnamoorthy, 1985) and specialization amongst soil organisms (in terms of food resources) appears to be higher in soils with greatest species diversity (Teuben and Smidt, 1992) and in tropical compared with temperate communities (in terms of microhabitat) (Stanton, 1979).

Two interesting ideas have been floated to explain the patterns of increased specialization in relation to biodiversity. The first is Rappoport's rule (Stevens, 1989: Wilson, 1992) which attempts to explain the increased specialization and richness in tropical systems. The general trend is for the geographical and altitudinal ranges of individual species to decrease towards the equator (Wilson, 1992). The more stable climates, with more muted seasonality, of the tropics are predicted to allow more kinds of species to specialize on narrower components of the environment, outcompete generalists and persist for longer than in temperate zones, thus allowing for greater species packing. The generally favourable soil environment, in addition to the small size and low vagility of the majority of soil organisms in relation to the high heterogeneity, could promote diversity in a similar way in soil communities. Secondly, Brown (1988) predicts thal increased diversity leads to reduced niche width/greater specialization and reduced abundance of species (as in tropical soils) based on the ~cost of commonness'. This concept relates to two factors. A trade-off may exist between less efficient use of a wide range ot resources and more efficient exploitation of a narrow range. Few abundant generalisl species tend to be replaced by many less common specialists as productivity increases. Also, if common species are more prone to natural enemies (predators, parasites and pathogens), then as productivity increases, enemies will tend to facilitate new species. To date, no one has really considered the validity of these ideas for soil ecosystems.

Diversity of soil communities 157

Productivity

Many authors have predicted that an increase in ecosystem productivity and energy should result in proliferation of species (Brown, 1988). This has been incorporated into the so-called ESA theory - energy, stability and area theory of biodiversity (Currie, 1991; Wilson, 1992). More solar radiation, greater climatic stability (over a range of time scales) and larger area should lead to greater biodiversity. More diverse communities have greater numbers of specialists at lower levels of abundance. Such species can only survive if they maintain a sufficient population size to avoid extinction- the total population size and area of extent of productivity are important. There is clearly low diversity in areas of low productivity like tundra and desert ecosystems, even though the area is often extensive and the systems are temporally relatively stable, simply due to the harsh conditions (Brown, 1988). How such concepts relate to soil communities is not clear at present. One might expect that ecosystems with high and varied production of litter etc. would encourage high diversity amongst soil communities, but hard data are difficult to find. It is more the variety of resources than the level of productivity that seems important. Indeed, as in other communities, there are instances where, beyond some point, decreased diversity follows increased productivity of soils, e.g. manuring and high fertilization or through intensive agriculture (see below).

Disturbance

The previous few sections are related more to the equilibrium concepts of community ecology and explanations of diversity patterns. Non-equilibrium concepts such as the 'Intermediate Disturbance Hypothesis' of Connell (1978) can also be considered for soil communities. Under intermediate levels of disturbance (in terms of frequency, magnitude, etc.) the environment is not too harsh to prevent colonization by many species but competitive exclusion of inferior competitors is prevented by reduction in the dominant species during the disturbances. Thus a non-equilibrium community can be established, with a higher diversity than would arise if the community was allowed to reach a stable equilibrium state (Giller, 1984; Giller and Gee, 1987). The 'Gradual Change Hypothesis', originally used to explain the 'Paradox of the Plankton' (Hutchinson, 1966) also predicts a high diversity as conditions change at a rate fast enough to prevent any one or a few species dominating the community before other species are favoured under the new conditions. The 'Predation Hypothesis' has also been proposed to explain the maintenance of non-equilibrium communities of high species richness, where key-stone predators maintain the dominant competitor at low density, allowing inferior competitors to coexist (Paine, 1966; Giller, 1984).

The Gradual Change Hypothesis is unlikely to operate in the rather benign soil environment, where as discussed above, environmental fluctuations tend to be buffered. Is there, however, any evidence of disturbances or predation maintaining high biodiversity in soil communities? On the contrary, most of the evidence to date points to disturbances of soil communities leading to reduced diversity. Land management disturbances in particular lead to dramatic changes. Quarrying (Usher and Parr, 1977), tillage (Abbott and Parker, 1980; House and Alzugaray, 1989), increased stocking rates of domestic animals (King et al., 1976), manure application (Weil and Kroontje, 1979) and burning of tussock heath vegetation (Thomas and MacLean, 1988) all lead to changes in community composition and reduced diversity and biomass. The reduced diversity relates not only to

158 Gil ler

reduced species richness, but also to increased dominance (e.g. Well and Kroontje, 1978: Freckman and Ettema, 1993). Clearing of forest followed by intensive cultivation in humid tropics led to a decrease in biomass to 6%~ of diversity to 17% and taxon richness to 50% of that of primary forest communities (Lavelle and Pashanasi, 1989). The effect of forest removal and cultivation on macrocoleoptera leads to a clear gradient of diversity from undisturbed to most disturbed areas (Table 4). Lebrun (1979) has produced a useful review of the ecological effects of agricultural practices on soil fauna. Natural disturbances, like the annual inundation of tropical forest soils by river flooding also creates low diversity communities (Irmler, 1979). Disturbances due to application of pesticides or through pollution also have negative effects on soil communities. Experimental use of carbofuran on Canadian Red Maple forest floor reduced macrofaunal abundance and diversity, which in turn led to a reduction in the rate of decomposition by over 40% (Weary and Merrian. 1978, in Lebrun, 1979). Copper pollution has also been shown to reduce nematode and microarthropod abundance by between 47-92% (Parmalee et al.~ 1993).

The reason for such reductions in reported diversity may be related to the fact that these disturbances are continuous, directed or long lasting rather than the punctuated events required under the intermediate disturbance hypothesis. The effects on soil arthropods in particular are due to exposure to unfavourable environmental conditions, such as desiccation, mechanical destruction, soil compaction and reduced pore volume and disruption of access to food resources (Abbott and Parker, 1980: House and Alzugaray. 1989). There is however, some indication that intermediate levels of disturbance may have some positive effects on components of the soil community. Whilst omnivore-predatory nematodes may decline under copper pollution by over 90%, other nematode trophic groups increase in abundance at intermediate levels of copper concentration, suggested to be due to the negative effect on their predators (Parmalee et al., 1993). Similarly. insecticides have negative effects on mesostigmatid mites but Collembola increase, again perhaps due to reduced predation (Usher, 1985). Acidification of the soil results in similar effects again (Baath et al., 1980). Finally, nematode trophic diversity appears to be greater in perennial cropping systems that had a longer period since the last cultivation than other types of system (Neher and Campbell, 1994). The potential therefore exists for positive effects from certain types and degrees of disturbance on soil communities but there is a lack of information and experimental work.

A similar conclusion must be drawn concerning the effect of predation on biodiversity o! soil fauna. There are large numbers of predatory taxa within the detritus food chains and also large predators in grazing food chains that also feed on soil animals. The impact ol predators on other trophic levels may be considerable (Anderson, 1975) as indicated in the examples above, where predators were removed by some kind of disturbance. Experimental removal of spiders from woodland soils led to significant increases in Collembola and centipedes, the preferred prey (Clarke and Grant, 1968). Usher (19851 reviewed some of the information available at that time on the effects of predators. There is a lot of descriptive work identifying the major soil invertebrate predators, and more limited information on predation rates and functional responses, but Usher concluded thai there was still little known about the effects of predators on the density of populations of soil organisms in the field. More detailed information is available from applied studies. such as the biological control of dung breeding flies using predatory beetles and mites (Doube et al., 1986: Doube. 1987). Predators can also cause local extinction of components of the soil fauna, as found in the recent invasion of the New Zealand flatworm

Diversity o f soil communities 159

Table 4. Variation in species richness (S) and dominance (% of total sample) of scarabaeid beetles along a disturbance gradient from virgin tropical rainforest to coffee plantations (adapted from Nestel et al., 1993)

Polyspecific shade Monospecific shade Unshaded Tropical rain coffee plantation coffee plantation coffee forest (> 10 species tree) (< 3 species tree) plantation

S 19 5 6 3 Dominance (n = 837) (n = 1279) (n = 142) (n = 244) 1st ranked 39% 79% 81% 99% 2nd ranked 16% 13% 15% 0.5% 3rd ranked 16% 7% 1% 0.5% Remaining species 29% 1% 3 % -

(Artioposthia triangulata) which destroys earthworm populations to Northern Ireland, Scotland and the Faroes (P. Enckell, personal communication). Given the key role earthworms play in modifying the structure of soils and the soil atmosphere, such impacts must have substantial negative effects on the rest of the soil community. Larger predators can also influence soil invertebrates. For example, house mice on a sub-antarctic island mainly feed on soil insects. It is predicted that the continued increase in the mouse population will, through increased predation pressure on the soil fauna, lead to reduced rates in nutrient cycling (Smith and Steenkamp, 1990). Crows can severely disrupt dung pats and thus influence local communities of dung inhabiting fauna in their search for large beetle and earthworm prey (personal observation). Apart from such anecdotal pieces of evidence, there appears to be virtually no information on the effects of predators on soil diversity per se; thus another avenue of promising research is wide open.

Aggregation

Recent theoretical models, based on ephemeral and patchily distributed resources, have predicted that coexistence is possible amongst a number of species even if many of the species are inferior competitors and resource limitation is evident (Hanski, 1987; Shorrocks and Rosewall, 1987; Ives, 1988, 1992). The essence of the models lies in the independent aggregation of species. The theoretical data, supported by less empirical data, suggest that when intraspecific aggregation in patches is greater than interspecific aggregation, such that different species tend to use different resource patches, other patches of resource are then available with no or few superior competitors, allowing inferior species to coexist in so called 'probability refuges' (Shorrocks and Rosewall, 1987; Ives, 1992). As such, species can effectively coexist without traditional resource partitioning. The high degree of spatial and temporal heterogeneity in soil communities outlined earlier would clearly lend itself to such processes.

Most field studies of soil populations have shown significant aggregation including Collembola and Acari (Usher, 1976), macrofauna like Isopoda, Diplopoda, termites, spiders, soil Coleoptera and ants (Abbott and Parker, 1980; Levings, 1983; Dangerfield, 1990; Villalobos and Lavelle, 1990), carrion feeders (Hanski, 1987) and dung beetles (Holter, 1982; Giller and Doube, 1994). On a large scale, this is closely related to habitat or soil type; on a smaller scale to physical/chemical factors like pore size, local temperature,

160 Giller

pH, food sources or microtopography etc. However, even in apparently homogeneous environments, significant aggregation occurs. This is shown for example in temperate and subtropical dung beetles distributed across identical resource patches in terms of size, age and composition (Fig. 10). The degree of aggregation seems to vary amongst various groups, e.g. Collembola and cryptostigmatid mites mostly show strong aggregation but mesostigmatid mites are more randomly distributed. Predatory species are surprisingly less aggregated than humus detritivores and fungus feeders (Usher, 1976),Litter layer and soil/litter interface species also tend to be more strongly aggregated than true soil-dwelling species (Dangerfield, 1990). The degree of aggregation within individual taxonomic groups also varies seasonally (Villalobos and Lavelle, 1990) or on a shorter time scale (Giller and Doube, 1994) in the Coleoptera at least.

Are there obvious reasons for such aggregations? Usher (1976) pointed to the five categories of aggregative patterns described by Hutchinson (1953):

(i) vectorial - aggregation in relation to the heterogeneity in physical and chemical factors in the soil at a range of scales as mentioned above.

(it) reproductive - related to oviposition patterns or the tendency for offspring to remain near parents. Lack of vagility amongst soil organisms, particularly larval forms, will enhance such patterns.

(iii) social - involving communication leading to clumping and avoidance of or protection from predators. There is no empirical evidence for this scenario amongst soil organisms.

(iv) stochastic - determined by random factors. This may play a role in terms of colonization patterns amongst ephemeral and patchy resources (Doube, 1987) and has led to the development of so-called lottery models to explain high diversity amongst competing coexisting species (Hanski, 1991). Non-territorial ground foraging ants tend to be patchily distributed although many species overlap in foraging ranges and nests are often in close proximity (Levings, 1983). The identity of the neighbouring species though is highly unpredictable, depending on the history of immigration and extinction from the local area. Presumably individual species will win some contests, lose others, and the possibility of facilitation (Davidson, 1985) will also encourage coexistence.

(v) coactive - determined by interactions of species in competition, leading to niche shifts etc. as illustrated earlier.

There does also seem to be evidence, however, for some degree of intrinsic disposition for aggregation over and above these organism/environment relations. Holter (1982) for example, has suggested that aggregation amongst Aphodius dung beetles may be related to reproductive behaviour, ensuring location of mates.

It is not clear exactly how the recent independent aggregation models fit into the above scheme - what is clear however, is that for such a process to function to enhance species diversity, positive species associations amongst resource patches should be low; i.e. different species should be aggregating in different resource patches. Presumably the aggregation could be accomplished through any of the above patterns. In caged experimental systems, artificial patchiness in the resource led to the inferior species maintaining a population and hence coexisting with the dominant competitor (Hanski, 1987). In a recent study of two functional groups of sub-tropical African paracoprid dung beetles (that bury dung beneath the pad), levels of aggregation were scale dependent, with

Diversity of soil communities 161

%

0 /

0 0

0 ~0~ 0 • °o /o o ~°~

i

0

a

/ logs 2 = - 0 2 3 + 1 . 7 4 1 o g X / (r =0.93,P.~O.Oy

:/;"

~°

A o •

1 2 3 log10

b

• log s z =--0-71 + 2-831og x (r= 0-98, P<O-O01) n log s z =--0.38 + 2-281og x (r:0-97, P<O.OI) • log s 2 =--0'23 + l'95tog x(r=O'84, P<O'02)

I-0

A/~,11 FG]~Z bee t / , ~ --I -0.

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162 Giller

aggregation increasing at higher spatial scales of analysis (Giller and Doube, 1994). There was strong intraspecific spatial aggregation for all species on spatial scales of 400 ha but not on finer scales of 100 m 2, where most species were randomly distributed. Similarly, some independent interspecific spatial aggregation occurred at the larger scale but not on the finer scale. Negative pairwise species associations were rare. Holter (1982) similarly found a lack of negative associations and high co-occurrence amongst temperate dung beetle assemblages. In the African example, intra-specific/functional group aggregation tended to be greater than inter-specific/functional group aggregation at all scales of analysis. Yel despite this, beetle densities in some 30% of experimental dung pads in the field would indicate that significant levels of intra- and inter-specific/functional group competition for dung is likely to occur, based on previous experimental evidence of competition between species within the two functional groups (Doube et al., 1988: Giller and Doube, 1989). Thus, aggregative behaviour of these dung beetle species is not predicted to diminish significantly the level of interspecific competition within and between the functional groups examined. When one adds the additional diffuse competition burden of endocoprid (laying eggs in dung pads) and diurnal telecoprid (ball-rolling) species, the importance of the level of aggregation shown in enhancing coexistence, and hence maintaining high biodiversity, is further lessened.

Even if soil communities do not sit comfortably within the independent aggregation models (and we do not have sufficient empirical data to decide one way or the other), the strong degree of aggregation amongst species within soil communities will increase the degree of environmental spatial heterogeneity and decomposition processes and hence, in this way alone, will contribute to enhanced species diversity in soil communities.

Endnote

There does appear to be some validity in considering soil communities as the "poor man's tropical rainforest' from a number of perspectives. Firstly, biodiversity of soil communities appears to be high in general compared with other communities although clearly more taxonomic work is imperative. This is largely attributable to the nested set of ecological worlds in the soil; the relationship between the range of size groupings of soil organisms relative to the spatial heterogeneity perceived by these various groups. The large "area for life" available to micro- and mesofauna also plays a part. The role of aggregation, how it relates to spatial scale under consideration and to species interactions in soil communities is largely unknown at present. As Usher (1985) questioned in a previous review, to what extent are the general theories of ecology applicable to the soil ecosystem and what might a study of soil communities contribute more widely to ecology'? Ten years later, we are still unable to come to grips with either question. What is becoming increasingly clear however,

Figure 10. Plots of log~ sample variance against logt0 mean number of dung beetle per pat. showing strong aggregation amongst identical resource patches. (a) Temperate Aphodius species: 6-12 pats per sample. Points represent A. fimetarius (x), A. rufus ( Q ), A. rufipes ( A ), A. contaminatus ([~) and the remaining 8 species (©) (from Holter, 1982). (b) South African crepuscular paracoprid beetles from two functional groups (FG; Doube, 1990); 81 pads per sample and 6 sampling occasions. Points represent large FG Ill (--), all FG III (---) and all FGV IV (--) dung beetles (from Giller and Doube, 1994).

Diversity o f soil communities 163

is that biodiversity of soil communities plays an important role in ecosystem processes such as decomposition, nutrient cycling and maintenance of soil fertility (see Lawton et al., this volume). The impact of human activities on soil ecosystems either directly or indirectly through changes in land use or vegetation, more often than not seems to lead to reduced biodiversity of soil communities - the longer term effects on the important 'ecological services' offered by soil communities remain to be seen.

Soil has been described as our most precious non-renewable resource (Marshall et al., 1982) but despite the perceived importance of the soil environment to the functioning of the biosphere, we know relatively little of the soil community structure and dynamics and the key controlling factors. Greater effort, especially in the fields of taxonomy, aspects of diversity and distribution, controlled field-based experiments, longer term studies on populations and community stability and application of ecological theory are needed to provide us with the necessary tools to predict the impact of man's activities on soil systems and processes. From this basis, we may be able to manage soil ecosystems more successfully in the future.

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