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PART 3—
THE ROLE OF THE GROUP IN BIODIVERSITY



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Page 223 PART 3— THE ROLE OF THE GROUP IN BIODIVERSITY

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Page 225 The World Beneath Our Feet: Soil Biodiversity and Ecosystem Functioning Diana H. Wall Natural Resource Ecology Laboratory, College of Natural Resources, Colorado State University, Fort Collins, Colorado 80523 Ross A. Virginia Environmental Studies Program, Dartmouth College, Hanover, NH 03755 Ecological Services Provided by Soil Biodiversity The Importance of soil fertility as a national resource was aptly noted by Franklin D. Roosevelt: “The nation that destroys its soils destroys itself” (Roosevelt 1937). Since then, the importance of soils and the organisms within them for many vital ecosystem processes has been identified, for example, cleansing of water, detoxification of wastes, and decay of organic matter. Indeed, it is now recognized that the functioning of soils, the dark material beneath our feet, is critical for the survival of life on the planet in its present form. Almost every phylum known above the ground exists below the surface of the ground (Brussaard and others 1997). Soil biota include the microorganisms (bacteria, algae, and fungi), protozoa (single-celled animals), microscopic invertebrates that are less than 1 mm long (such as rotifers, copepods, tardigrades, nematodes, and mites), larger invertebrates up to several centimeters long such as those easily seen by the naked eye—ants, snails, earthworms, spiders, termites and so on, and vertebrates. One cubic meter of soil can harbor millions of species of microorganisms and microscopic invertebrates—organisms whose identities and contributions to sustaining our biosphere are largely undiscovered. Life in soil is recognized as an important part of Earth's overall biodiversity, yet few studies measure the taxonomic diversity of soil or the relationship of soil biodiversity to ecosystem structure and function (Pimentel and others 1997; Swift and Anderson 1994). Understanding of the relationship between biodiversity and ecosystem function in soils is critically needed if we are to manage and predict

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Page 226 the impacts of human activity on ecosystems effectively and ensure soil sustainability. Species in soils perform ecological services that directly control the sustainability of human life. Soil microorganisms and invertebrates (such as fungi, bacteria, nematodes, and earthworms) provide for the purification of air and water, for the decay and recycling of organic matter and hazardous wastes, and for soil fertility. Soil organisms mediate critical ecosystem processes, particularly those in biogeochemical cycling (Swift and Anderson 1994; Matson and others 1987). Soils store vast amounts of carbon, and it is the biota in soils that most influences local and global processes involving the cycling of carbon and nitrogen, including several greenhouse gases (Coleman and Crossley 1996; Huston 1993). The organisms in soil—through their direct, indirect, and modifying effects on these ecosystem processes (Lavelle and others 1995)—provide humans with numerous services (table 1). Pimentel and others (1997) valued the function of soil biodiversity at $25 billion per year on the basis of the contributions of soil biodiversity to topsoil formation in agricultural lands; this value would increase considerably if natural terrestrial systems were included. A single ecosystem service, such as the generation and renewal of soil and soil fertility (table 1), involves many ecosystem processes and countless organisms representing diverse phyla. These range from large vertebrates to invertebrates and smaller macrofauna such as earthworms and ants that channel through the soil, algae living on the soil surface, and microorganisms involved in the decay of organic matter (Pankhurst and Lynch 1994). The decay of a small animal (such as a piglet) in the soil requires many phyla and can involve 100–500 species of Arthropoda (Richards and Goff 1997). Knowledge of the succession of species participating in the decay of humans is used in forensic medicine to determine the time of death (Goff 1991). Information on the number and types of soil species and phyla required to decompose plant material or invertebrates might be avail- TABLE 1 Some Ecosystem Services Provided by Soil Biota Biota Ecosystem Services Regulation of major elemental cycles Retention and delivery of nutrients to plants Generation and renewal of soil, and soil fertility Detoxification and decomposition of wastes Modification of the hydrological cycle Mitigation of floods and droughts Translocation of nutrients, particles, and gases Regulation of atmospheric trace gases (production and consumption) Regulation of animal and plant populations Control of potential agricultural pests Foundation of life from which humanity has derived elements of its agricultural, medicinal, and industrial enterprises Source: Modified from Daily (1997).

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Page 227 able, but the data from many isolated field studies and from taxonomic work have not been synthesized. Soil organisms contribute to the detoxification of pollutants on a global and a local scale—for instance, detoxifying the pollutants in our yards, farms, golf courses, and parks. These organisms, through their metabolism, are critical to detoxifying and purifying many pollutants before they are leached into groundwater and reach aquatic ecosystems (Abrams and Mitchell 1980; Sayler 1991). Finding environmentally sound ways to use organisms to renew polluted soils and decompose the garbage in our landfills is a growing industry (bioremediation) that depends on the ecosystem services of soil organisms (Sayler 1991). Ecosystem services such as the mitigation of floods and droughts through prevention of soil erosion, the buffering and modification of the hydrological cycle and the translocation of nutrients, particles, and gases are a reult of many species accomplishing different, but linked, tasks. For example, soils are a temporary habitat for predominantly aboveground organisms (such as vertebrates, lizards, rabbits, gophers, and birds) (Anderson 1987) and invertebrates (such as ants, spiders, and beetles) that move through the soil acting as cultivators or bioturbators, some species ingesting soil and others burrowing in and moving it. Those activities affect soil porosity, the retention of soil water and its movement vertically and horizontally, the transfer of materials throughout the soil profile, and the hydrological cycle. Soil bioturbators, while changing the physical and chemical environment of the soil, also transfer other, smaller organisms and soil particles within the soil, constantly creating new soil aggregates and new surfaces as habitats for microorganisms and facilitating topsoil formation. In this way, the soil biota “plows” the soil, mixing organic matter and nutrients essential for life throughout the soil profile. Soil organisms have long been recognized as essential for agricultural food production. Nitrogen-fixing bacteria, mycorrhizal fungi, and rhizobacteria that have beneficial relationships with plants, in consort with the decomposers, supply elements essential for plant growth. In addition, through predator and prey interactions and parasitism, soil organisms control vast numbers of agricultural pests (insects, microorganisms, and fungi) (Kerry 1987). For example, the Steinernematid and Heterorhabditid nematodes that parasitize insects above and in the ground are used as a biological control of armyworms, carpenter worms, flea beetles, crown borers, cutworms, cockroaches, leaf miners, mole crickets, root weevils, stem borers, and white grubs (Kaya 1993). Many invertebrate species yet to be discovered are expected to have enormous potential as biological control agents. Soil Biodiversity Assessment Despite the essential nature of services provided by the soil biota, the systematics of the majority of these organisms has not been determined. Information is lacking on how species' abundance, distribution, and interactions influence ecosystem functioning and whether there are key taxa essential for ecosystem processes. Our ecological knowledge is insufficient to make needed inferences about factors controlling the distribution and activity of the species of soil biota

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Page 228 over broad geographic ranges and whether removal or introduction of species alters ecosystem processes. The identification of individual soil organisms to the species level is severely hampered because • the sheer abundance of soil biota is overwhelming to describe—1 m2 of a pasture can contain 10 million nematodes, 45,000 oligochaetes, and 48,000 mites and collembola (Overgaard-Nielsen 1955); • few scientists have soil taxonomic or soil ecological expertise; estimates are that only 3% of the world's scientists study microscopic and invertebrate organisms in soils; • in situ identification of most soil organisms is difficult, so sampling and extraction techniques must be used to remove the organisms from soil, and these techniques should not affect the features used to identify and describe the individuals; • organisms range in size from microscopic to macroscopic; • organisms can have many different structures during their life cycle; • methods of sampling and identification must vary with the size of the taxonomic group, for example, earthworms and bacteria (Hall 1996; Oliver and Beattie 1996); and • promising molecular techniques for most soil organisms are still in their infancy (Blair and others 1996; Hall 1996). Together, those factors often make the identification and enumeration of soil biota seemingly an insurmountable obstacle for soil research. Perhaps the most important part of this problem is that the decline in human resources in taxonomy overall as a result of diminished institutional support for systematic research, particularly by agricultural and natural resource agencies, has been especially severe for soil taxa (Brussaard and others 1997; Freckman 1994). There is a poor understanding of the ecological roles played by soil species. Factors contributing to the dearth of knowledge are many and include the following: • The diversity of soil organisms spans many phyla (from microorganisms to arthropods to vertebrates), and this makes interactions and ecological roles difficult to assess. • The temporal (seasonal and annual population changes) and spatial scale of the soil habitat that is relevant for an organism (from soil aggregate to landscape) varies among groups. • Soil species can live at considerable depths (Freckman and Virginia 1989; Silva and others 1989), or can be restricted to microhabitats such as near the surface of roots (rhizosphere). • The specific taxa participating in soil food webs can change with the soil physiochemical environment, the quality of organic matter, plant species diversity, landscape characteristics and climate. All these make it difficult to compare the ecological roles of soil taxa in different ecosystems.

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Page 229 Other than for earthworms, termites, and other larger soil invertebrates, the use of species composition in ecosystem studies is not yet widespread, because the taxonomy of nearly all groups is incomplete, and for most species only the adult stage is described. As a result, the approach to studying the link between organisms and ecosystem processes has been to place soil organisms in functional groups at a gross level—for example, considering all oribatid mites and springtails that feed on fungi to be fungivores, all mesostigmatid mites to be predators of other micro-fauna, and so on. The taxonomic and ecological limitations of this approach have been emphasized (Moore and others 1996; Walter and others 1988). We lack knowledge of the feeding strategies of more than 90% of the soil biota. There is only baseline knowledge of the soil biodiversity in a few ecosystem types, mainly those with high economic value—agricultural, grazing, and forestry (Daily 1997; Pimentel and others 1997). The soil biodiversity estimates of those types generally exclude aboveground organisms that have only one phase of their life cycle in soil or that use the soil as a habitat. Groups of invertebrates, such as wasps and bees, or vertebrates (Anderson 1987; Ingham and Detling 1984; Naiman and Rogers 1997) are studied primarily by “aboveground scientists”, and the interchange of information about the functions of such organisms between these scientists and soil ecologists is rare. Some vertebrates, a group generally thought of as living predominantly aboveground, live entirely in the ground. For example, Caecilians (Wake 1983), one species of which was found living at a depth of 30 ft. We note that the recent summaries of taxonomic progress in the major soil biotic groups, which we have outlined below, do not include these predominantly aboveground organisms (Brussaard and others 1997; Groombridge 1992; Hawksworth and Ritchie 1993; O'Donnell and others 1994; Systematics Agenda 2000 1994). A brief assessment of the summaries and methods for studying soil organisms follows. We discuss the groups of soil organisms in order of increasing body size. Viruses, Bacteria, and Fungi There have been dramatic advances in the methods for assessing bacterial and fungal biodiversity, although no method can give the “best” quantitative estimate of diversity, because for these taxa and such invertebrates as the nematodes, the reproductive biology of the groups does not permit the application of a “species concept” (de Leij and others 2000; Zak and Visser 1996). For those groups, characteristics to define species are genetic, ecological, chemotaxonomic, and physiological (Snelgrove and others 1997). Molecular methods and chemosynthetic approaches are expanding our knowledge of bacterial and fungal diversity and of trophic relationships with soil invertebrates (Anderson 1975; Hawksworth 1991). Fungi and other microorganisms can be combined into functional groups on the basis of differences in the enzymes required to use particular carbon compounds (for example, cellulose, lignin, and sugar) (Zak and Visser 1996). The specificity of analysis has increased, allowing functional groups to be separated at finer levels of resolution, and enabling the types and numbers of microorganisms and their rates of use of primary and secondary compounds to be analyzed (Zak and Visser 1996). Another method, relying on biochemical markers of diversity termed

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Page 230 FAME (fatty acid, methyl esterase) profiles the relative abundance and diversity of broad groups of microorganisms (Stahl and Klug 1996). Viruses are rarely considered even though they can be potential biological control agents for soilborne pathogens of plants or plant pests. The importance of these microorganisms to ecosystem function is detailed in Lynch and others (this volume). Microfaunal and Mesofaunal Invertebrate Groups These groups are Protozoa, Rotifera (wheel animals), Tardigrada (water bears), Nematoda (roundworms), Acari (mites), and Collembola (springtails). No method extracts all taxa, and methods vary widely in their ability to extract these organisms from soil quantitatively and qualitatively. Many groups can be classified to the species level on the basis of morphological characteristics. More molecular methods are becoming available for classification and for assessing interspecific and intraspecific variation on a geographic basis (Avanzati and others 1994; Courtright and others in press; Oliver and Beattie 1996) Macrofaunal Invertebrate Groups These include Arachnida, Chilopoda, Diplopoda, Insecta, Annelida (segmented worms), and Mollusca (see figure 1). They can be more easily classified to the species level, and their ecological roles are known in general (Brown and Gauge 1990). In temperate regions, their ecological roles include direct processing of organic matter, predatory regulation of population size, modification of soil structure, and production and consumption of atmospheric gases, such as methane. Organisms that cannot be readily identified to the species level include enchytraeid worms, many of the larger mites, some spiders, larval beetles and larval flies. Knowledge of these soil taxa varies dramatically between different locales, and only a few locations have well-described invertebrate macrofauna. Stable-isotope techniques have been used successfully to study trophic relationships and interactions in freshwater habitats and have great promise in soils (Barios and Lavelle 1986; Boutton and others 1983), particularly if extended to the microfauna. Current Estimates of Soil Biodiversity. Almost all aboveground phyla have representatives in soils, but there are no global assessments of the biodiversity in soils and only a few global estimates of individual taxonomic groups (Brussaard and others 1997). In figure 1, we present our estimate of soil biodiversity described to date on the basis of the literature or “best guesses” by specialists working on particular groups. Caution should be exercised when considering these numbers, inasmuch as the size of the soil samples used varies greatly with the size of the organism. Earthworm diversity might have been assessed from 1-m2 samples to a depth of 40 cm, whereas the protozoan diversity might have been described from a 5-g sample scraped from the soil surface. There are estimates of the total number of species that exist in these groups. For some of the taxa, all or the vast majority of projected species are soil-dwellers; estimates of total species in these groups include

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Page 231 Figure 1 Size classification of organisms in the decomposer food web of soilsa, by body width (based on Swift, Heal and Anderson 1979), with number of species described to date for each group. (a) Species in litter and decaying logs are included in the estimate of soil-dwelling species. (b) Torsvik and others (1994) measured 4,000 independent bacterium-sized genomes in 30 g of forest soil, using DNA analysis. They calculated that this equated to 13,000 species. In a review of Torsvik's study, Dykhuizen (1998) suggested that there could be as many as 500,000 species in the soil sample. Those numbers of species based on DNA analysis are much higher than those described from traditional bacterial isolation and culturing techniques (for example, the 1,718 according to Akimov and Hattori 1996) because culturable bacteria might represent only 0.1–1% of the species in a population. (c) Maddison (1995) gives the number of total described Diplura species (800). Because most Diplura are found in soil, we assumed this number to equal the number of soil-dwelling species. (d) Hoffman (1982a) gives the number of total Chilopoda species described (2,500). Because Chilopods are found only in soil, leaf litter, rotting wood, and caves, we assumed this number to equal the number of soil-dwelling species. 1Torsvik and others 1994. 2Akimov and Hattori 1996. 3Brussaard and others 1997. 4Walters, personal communication. 5Ravlin 1996a. 6Maddison 1995. 7Scheller 1982. 8Bignell, personal communication. 9Bignell and Eggleton 1998. 10Brusca 1997. 11Hoffman 1982a. 12Hoffman 1990.

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Page 232 Collembolla, 20,000–24,000 (Ravlin 1996b); Diplura, 1,647 (Ravlin 1996a); Enchytraeidae, 1,200 (Behan-Pelletier, personal communication); and Isoptera, 3,000 (Bignell and Eggleton 1998). In other groups, some of the species live outside the soil; estimates of total species in these groups include bacteria, 1,000,000; fungi 1,500,000; algae, 400,000; Nematoda, 1,000,000; protozoa, 200,000 (Hammond and others 1995); acari, 348,500–900,000 (Walter, personal communication; Walter and others 1998); and Diplopoda, 50,000–60,000 (Hoffman 1982b, 1990). These estimates of total existing species could be low; the soil component of biodiversity has traditionally been underappreciated and poorly described relative to aboveground species because of soil organisms' abundance and microscopic size and the dearth of soil taxonomists. The gulf between the numbers of species already described in soils and the projections of total numbers of species confirms what has already been heralded by others (Andre and others 1994; Wilson 2000; May 2000): that microscopic groups—such as bacteria, fungi, nematodes (Baldwin and others 2000; Bernard 1992), acari (Behan-Pelletier and Bisset 1993), Symphyla, and Enchytraeids (Healy 1980) are desperately in need of taxonomic and ecological attention. The broad groups listed in figure 1 probably all have a global distribution. Information has not been synthesized to make general statements about diversity and geographic distribution (Brussaard and others 1997; Dighton and Jones 1994). For example, Lavelle and others (1995) analyzed earthworm community assemblages containing 8–11 species across 53 global climatic locations and found that neither species richness nor species diversity increased with decreasing distance from the tropics. However, they noted changes in the proportion of species feeding on soil and litter with decreasing latitude. That might not be unusual; Hammond (1995) noted that, for seasonally dry ecosystems, the increase in noristic diversity is not necessarily paralleled by an increase in the terrestrial diversity of invertebrates, fungi, or microorganisms. There are more biogeographic assessments for individual groups listed in figure 1 on the political and regional scale than on a global scale (Brussaard and others 1997; Folgarait 1996; Pearce and Waite 1994). Species distribution patterns are influenced by chemical and physical factors—such as soil texture, organic matter, and soil moisture—as well as by climate and vegetation (Anderson 1975; Swift and others 1979; Wright and Coleman 1993). For example, some species of earthworms and Enchytraeids in the Oligochaetes rarely occur in deserts (Dash 1990; James 1995); and in the United States, where there are six indigenous genera of earthworms, the species distribution has been limited geographically (to such areas as those not affected by the Wisconsin glaciation, forests, mud flats, and riparian areas) (James 1995; Reynolds 1995). On the basis of political boundaries, Coomans (1989) listed 228 terrestrial nematode species in Belgium, and Behan-Pelletier and Bissett (1993) estimated North American soil arthropods as follows: isopods, 92 described, 62 undescribed; chilopods, 850 described, 400 undescribed; diplopods, 850 described, 400 undescribed; pauropods, 70 described, 47 undescribed; symphyla, 33 described, 22 undescribed; spiders, 1,700 described, 250 undescribed; Opilionids, 235 described, 250 undescribed; acari, 2,500 described, 14,500 undescribed; and

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Page 233 Protura, 26 described, 104 undescribed. Lindquist and Behan-Pelletier (personal communication) estimate that the totality of described and undescribed species represents 10% of the world's total soil arthropods. Lawton and others (1996) in a Cameroon tropical-forest study, identified 115 species of termites and 432 nematode morphospecies in 185 genera. More than 90% of the nematodes remain unidentified because costs for further descriptions by taxonomists are expensive and time-consuming. There is no compilation of endemic species in soils, but there are summaries of the endemic species in some groups listed in figure 1, such as earthworm species in the United States Qames 1995). It is likely that in ecosystems that have endemic species and low species diversity, such as in soils of the Antarctic dry valleys, the species are more vulnerable to loss by disturbance, but whether their loss would affect ecosystem function must be confirmed through experimentation (Freckman and Virginia 1998). Effect of Disturbance on Soil Biodiversity Whether the presence or absence of a single soil species can affect ecosystem structure or function (for example, such biogeochemical processes as rates of decomposition and plant production) is largely unresolved (Beare and others 1995) except for agroecosystems, in which a single species becomes noted as a crop pest, affecting transfer of nutrients and plant production (Brussaard and others 1997; Swift and Anderson 1994). The effects of introductions of soil species on ecosystem functioning constitutes a new field of study by aboveground ecologists, but it is a well-established field among agriculturalists. An example of the introduction of an alien species from Asia into a natural ecosystem in the United States is the earthworm species Amynthas hawayanus, which reduced New York forest-floor organic matter and increased water runoff and soil erosion (Burtelow and others 1998). The effects of introductions of alien species into fields have been studied by agriculturalists worldwide because of the potential devastating economic loss to crops (Gotten and Riel 1993; Swift 1997). Of growing importance is research on the introductions of soil organisms that might enhance plant productivity while protecting plants from pests (Cook 1993). Better known, however, are the quarantines to restrict movement of soil biota and plant pests and thereby prevent the spread of exotic or established soil-pest species. The European Community lists nine plant-parasitic nematode species that are targeted in the hope of prohibiting introduction into Europe (EEC Council Directive 77/93/EEC, see Gotten and Riel 1993); California has 25 nematode species listed in its quarantine regulations. Despite such efforts, even the tightest regulations have been only partially successful. The US Department of Agriculture, with millions of dollars allocated to prevent the spread of Globodera rostochiensis (the potato cyst nematode) from Long Island, NY, where it became established in 1941, was unsuccessful in preventing its spread to other North American areas (figure 2). A second species of the potato cyst nematode, Globodera pallida, has not established itself in the United States but since 1972

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Page 242 Natural Investment in Diversity: The Role of Biological Communities in Soil Frans A.A.M. De Leij David B. Hay James M. Lynch School of Biological Sciences, University of Surrey, Guildford, Surrey GU2 5XH, UK Introduction There is a persistent view that the commercial application of “environmental biotechnology” relies largely on the exploitation of single species, well-defined bio-chemical pathways, or the expression of novel gene sequences by genetically modified organisms. Consequently, the role of species assemblages and in particular the importance of interactions between the members of natural communities have been largely neglected in a commercial or industrial context. This oversight is important because most ecosystem processes appear to be governed by the activity of species at the guild and community level and not by species that function in isolation. Indeed, most well-studied bioremediation experiments and many examples of effective biological control show that natural, or “intrinsic,” processes are at least as important in achieving economic ends as specific and targeted bio-technological interventions. This paper suggests that exploitation of microbial communities is a potentially rewarding alternative to the “classical” or “single-species” biotechnological approach. We emphasize the need for ecosystem study in the context of biotechnology development. The development of sampling and monitoring techniques has high priority for research. Monitoring is an important technology for tracking “intrinsic” beneficial processes, and sampling will provide data to improve our fundamental understanding of ecological processes in an applied context. In this paper, the commercial uses of microorganisms and microbial communities are examined in the context of bioremediation and biological control, but they could

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Page 243 equally well be scrutinized in relation to such processes as soil formation and biodegradation. Biological Control Development of Single Isolates as Biological Control Agents The commercial development of Bacillus thuringiensis is one of the most outstanding examples of successful technology based on a specific biological resource. B. thuringiensis has taken the dominant share of the biopesticide market: a little over 1% of total pesticide sales. That figure could rise to 10% during the next decade (Hokkanen and Lynch 1995). Although B. thuringiensis is often referred to as a biological control agent, its application and action have much more in common with chemical pesticides. The toxin crystal produced by the bacterium is the active ingredient of commercial products, and the ecology of the organism itself is largely irrelevant in pest control—in stark contrast with familiar examples of biological control in which the dynamics of host-parasite or host-predator interactions determine product efficacy. Where biological solutions have been sought to combat pests in well-defined (and controlled) conditions, commercial success has sometimes been based on the exploitation of single species or isolates. For example, Peniophora gigantea is used to control Heterobasidion annosum in pine forests, and Agrobacterium radiobacter is commercially valuable for the control of A. tumifaciens (crown gall) in tree nurseries (Deacon 1991). Nevertheless, it is the specific match between the environmental requirements of the biocontrol agent and the conditions in which pest populations thrive that results in high product efficacy. Such matches appear to be the exception rather than the rule. Involvement of Many Organisms in “Natural” Biological Control Soil communities comprise a large variety of microbial species. It is estimated that a gram of natural soil might contain as many as 4,000 or 5,000 “species” with DNA-sequence similarities of less than 70% (Tørsvik 1990). This level of diversity is comparable in a variety of soil habitats, but microorganisms from different localities generally show markedly different patterns of species composition. The role of these microbial “species” and of species diversity in ecosystem function is largely unknown, but it is well established that several components of the normal soil flora serve to regulate the activities of pathogens. Plant-parasitic nematode populations in soil are regulated by a large variety of egg parasites, female parasites, nematode-trapping fungi, bacteria, and possibly viruses (Jatala 1986). Specificity is common, and in suppressive soils it is the combined activities of a consortia of antagonists that achieve control. Nematode-trapping fungi—which produce adhesive knobs, adhesive rings, or adhesive hyphal networks—are adapted to “catch” the free-living nematodes. Likewise, the nonmotile spores of Pasteuria penetrans (a bacterial parasite of root-knot and cyst nematodes) attach to juvenile nematodes. Fusarium oxysporum, Catenaria auxiliaris, and Nematophtora gynophila parasitize young females before egg-laying commences, and Peacilomyces

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Page 244 lilacinus, Cylindrocarpon destructans, and Verticillium chlamydosporium are classified as specialized egg parasites. Many other organisms inhibit soil nematode populations in a nonspecific way through toxins, competition, and predation (Jatala 1986). The combined action of the whole community of specialized and nonspecialized organisms is responsible for keeping nematode populations below the economic threshold in nematode-suppressive soils. There is also considerable species diversity within “functional groupings.” Kerry (1988) reported that as many as 150 species of fungi were isolated from eight cyst-nematode species, parasitizing 97% of adult female nematodes in suppressive soils; and Jatala (1986) reported that there are at least 100 species of nematode-trapping fungi. At any time, consortia of antagonists provide nematode suppression. The exact composition or activity of these consortia, however, is determined by spatial and temporal characteristics of the environment (Crump 1987). Different parts of the diverse community of organisms involved in nematode suppression are necessary to provide effective nematode control at any given time during the crop cycle. Despite many studies that document the combined activities of several antagonists in the control of plant-parasitic nematodes, the commercial drive for biopesticide products has been in the development of solutions based on single microbial species or even single isolates. That approach is based on little understanding of host-parasite population dynamics, and the selection of agents for development is complicated by issues of ease of culture and product shelf-life. Biological control of cyst nematodes provides a good example; Nematophtora gynophila and Pasteuria penetrans are thought to be important agents of nematode control in natural soil (Davies and others 1992; Kerry and others 1982), but Verticillium chlamydosporium is the only organism chosen for commercial development, mainly because it is easy to mass-produce it in vitro. Beneficial Use of Simple Consortia of Species if Vitro Culture is Possible. Experimental tests have shown that the fungus Verticillium chlamydosporium can reduce plant-parasitic nematode populations in soil by as much as 90% (De Leij and others 1992b) but that this can be achieved only under specific environmental conditions: temperatures must be close to 20°C, appropriate host plants must be available for fungal colonization, and nematode population densities must be low (De Leij and others 1992a,b,c). Thus, V. chlamydosporium has a specific “window of opportunity”, and its utility as a commercial biopesticide in the field is not large. Furthermore, the high multiplication rate of cyst and root-knot nematodes means that parasitism rates as high as 90% (common in laboratory tests) are insufficient to prevent nematode population increases and economic damage to crops. Higher levels of nematode control can be achieved with combined application of V. chlamydosporium and Pasteuria penetrans (an obligate bacterial parasite of nematodes). This approach provides control that is comparable with the use of nematicides and is much more efficacious than the use of either organism in isolation (De Leij and others 1992a). V. chlamydosporium is unable to penetrate the

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Page 245 plant root and parasitizes only egg masses on the root surface (De Leij and others 1992a); P. penetrans also parasitizes nematodes that develop deep inside the root. Even though this simple consortium approach shows promise, it is commercially unattractive because of production constraints. Furthermore, the combination of nematode host specificity of P. penetrans and the relative small environmental window of opportunity of V. chlamydosporium is commercially unattractive. Management of the System as a Means of Inducing Biological Control Agricultural management provides an alternative to biological intervention. Simple practices, such as crop rotation, prevent the buildup of pests and diseases to economically damaging levels. In continuous cropping, initial accumulation of pests and diseases is tolerated to allow populations of natural antagonists to reach levels that provide long-term control. Augmentation of soil with organic manure is widely used to increase nonspecific biological activities that suppress pest and disease populations. Hams and Wilkin (1961), for example, reported that augmentation of soils with farmyard manure or green manure reduced plant damage attributed to plant-parasitic nematodes; the introduction of organic substrates probably promotes general microbial activity that is antagonistic to nematode populations. Similarly, amending soil with organic residues that have relatively high carbon-to-nitrogen ratios can control fusarium root rot; the free nitrogen required by the microbial biomass to degrade these organic amendments leads to insufficient nitrogen availability for pathogen growth (Snyder and others 1959). Others (for example, Park and others 1988) have suggested that induction of fusarium-suppressive soils is a more specific process whereby nonpathogenic Fusarium oxysporum isolates interact with siderophore-producing fluorescent pseudomonads to provide conditions that are nonconducive for the pathogen. Lemanceau and co-workers (1992, 1993) showed that pathogenic F. oxysporum isolates were more sensitive than nonpathogenic isolates to the iron deficiency induced by Pseudomonas putida, and this difference resulted in effective biological control. In general, disease suppression in soils is attributed to biological processes. Experiments have shown that suppression can be transferred to nonsuppressive soils by adding small quantities of suppressive soil to soils that are conducive to disease (for example, Stirling and Kerry 1983). However, attempts to attribute disease suppression to specific components of the natural microbial community have largely met with failure. Processes and species interactions at the community level, rather than the specific ecosystem services or functions of individual species, are likely to be responsible for disease control and pest control in suppressive soils. Research on and economic exploitation of processes at the community level are therefore potentially rewarding. It is also an environmentally sound, sustainable, and in many situations realistic approach. Suppressive soils need not be only “hunting grounds” for potential biological control products; an understanding of suppressive-soil community ecology is likely to lead to augmentative and manipulative management practices that are of considerable economic benefit.

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Page 246 Bioremediation Intrinsic and Augmentative Approaches The degradative enzymatic capabilities of microorganisms and microbial species assemblages play an important role in the remediation of polluted environments (see Crawford and Crawford 1996 for a review of principles and applications of microorganism exploitations and Lynch and Wiseman 1998 for details of microbial and microbial-product use in ecotoxicant monitoring). Natural communities show considerable potential to recover from small- and even medium-scale pollution effects, and with time biotic and abiotic factors interact to reduce contaminants to nondetectable levels. This is “intrinsic bioremediation”, remediation that relies solely on natural processes with little or no intervention (see Ellis and Gorder 1997 for review). As a commercial technology, the “do nothing” (but monitor) approach does not require investment in physical removal or discharge technology and is easily integrated with other pollution-control and remediation technologies. In many situations, this is a realistic and economical approach to pollution abatement. As a consequence, biodegradation by naturally occurring populations of microorganisms is a major mechanism, for example, in the removal of petroleum from coastal waters (well documented in the Prince William Sound, Alaska, after the Exxon Valdez oil spill). Rapid acclimation of the resident microbial population is common after hydrocarbon contamination (for example, Braddock and others 1995) and is evidence of the capacity of the community to respond to pollution. Detailed research has also shown that a 10-fold increase in population size of hydrocarbon degraders can follow substantial petroleum contamination of coastal waters (Atlas 1995a). Furthermore, petroleum degradation rates can be increased (by a factor of 3–5) by enrichment with inorganic fertilizers (Atlas, 1995b: Coffin and others 1997; Pritchard and others 1995). Similarly, in terrestrial environments, there is considerable evidence that natural microbial species assemblages respond to pollution in ways that ameliorate or remove contaminants and that this activity can be enhanced by manipulation of the physiochemical conditions to augment remediation (Liu and Suflita 1993). Intrinsic cleanup does not require extensive knowledge of the abiotic and biotic processes and interactions by which remediation is achieved. Nevertheless, understanding of “the system” is beneficial where it leads to the ability to enhance decontamination rates by manipulating and controlling environmental conditions or augmentation of biological processes for human benefit. Complex Interactions, Biological Diversity, and the Exploitation of Intrinsic Bioremediation Processes Molecular studies have shown that diverse microbial species assemblages (and genes) are involved in the complete catabolism of complex substrates (for example, Vallaeys and others 1995). Metabolic capabilities are often widely dispersed among distinct taxonomic groups and environments (Mueller and others 1994), and the metabolic capabilities of microbial communities as a whole are

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Page 247 characterized by considerable functional overlap (Pritchard and others 1995). This could be indicative of “functional redundancy,” but it is more likely that patterns of susceptibility and resistance to pollutants interact with metabolic and cometabolic activities to create a mosaic of functions underpinning microbial community integrity in the context of environmental heterogeneity. Furthermore, studies of mixed cultures (enrichments of xenobiotic-degrading microorganisms in liquid culture, for example) have demonstrated the importance of secondary use of substrates in microbial populations. Complex interactions—including the provision of specific cofactors, removal of toxic products, modification of growth rates, cometabolism, and gene transfer—have now been implicated in microbial communities and are important in degradation (Weightman and Slater 1979). Much can be learned from degradation studies of pesticides: the herbicide dalapon, for example, is not readily degraded by any organism, but enzyme products in small and well-structured microbial communities can bring about complete substrate metabolism (Senior and others 1976). Diversity of metabolic function is undoubtedly important in the ability of microbial communities to achieve bioremediation, but it is also important because of indirect and “cascade” interactions that enable complete degradation. Genetic diversity also underpins community-level responses to environmental change, species compensation, and complementarity (see Frost and others 1995); and genetic diversity is likely to be particularly important in the degradation of substrates in changeable, fluctuating, and perturbed environmental conditions (which are often characteristic of polluted environments). Difficulty of High-Technology Bioremediation Solutions in Natural Environments Recombinant-gene technology appears to offer appealing prospects for the design of microorganisms for use in bioremediation. Genetic manipulation has been used to expand the array of substrates that can be used by wild-type microorganisms and to restructure existing metabolic pathways (thereby avoiding the production of deleterious metabolites; see Lui and Suflita 1993). Nevertheless, the “inundative approach”in which single species, strains, or isolates of bacteria (recombinant or wild-type) are cultured in vitro and released into the environment—has proved difficult for achieving viable populations of pollution degraders in situ. Experimental tests often show that recombinants are likely to be outcompeted by wild-type parental strains (Fleming 1994; Recorbet and others 1992; Vahjen 1997), and even indigenous species often fail to establish in field trials, because of abiotic or biotic factors (de Leij and others 1992c; Kerry and others 1993). Thus, the commercial use of recombinants and wild-type “superstrains” is not likely to be great in the context of bioremediation. As Hamer (1993) has stated, the utility of genetically engineered microorganisms in bio-remediation processes is likely to be restricted to specific in situ and ex situ applications because recombinants fail to match the degradative abilities of natural microbial species assemblages (despite the addition of metabolic capabilities) and fastidiousness is likely to preclude their use in all but the most highly protected and controlled environments.

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Page 248 Wilson and Lindow (1993) refer to some 27 experimental releases of recombinant microorganisms in field trials. Much of the work has been done to assess the potential risks posed by releasing genetically modified organisms into the environment, but it is now clear that it remains a major technological challenge to achieve the establishment and persistence of recombinants in natural environments. The genetic and physiological “programming” of microorganisms to achieve viable and controlled phenotypic expression under variable (and largely unpredictable) physiochemical and biotic conditions is no small task (Delorenzo 1994). High Priority of Ecological Research for Commercial Exploitation of Microorganisms in Bioremediation Exploiting intrinsic bioremediative processes does not require a full understanding of the biochemical, physiological, and ecological interactions by which pollutant removal (or transformation) is achieved. But it is clear that the better these processes are understood, the easier it will be to improve intrinsic remediation efficiencies. A considerable body of evidence suggests that bioremediation rates in situ are determined by soil, sediment, and substrate chemistry. However, recent studies of pollutant mineralization (of hexadecane, phenanthrene, and naphthalene, for example) show that environmental factors (especially temperature, disturbance, and mixing) are at least as important as purely chemical interactions in determining rates of biodegradation by microorganisms (Sugai and others 1997), emphasizing the need for ecosystem-level study of system interactions in bioremediation processes. Similarly, the study of interactions between microorganisms and other fauna has high priority. In terrestrial habitats, competition and grazing by microbial predators are thought to be important determinants of soil biodegradation rates (Travis and Rosenberg 1997), but models of microorganism-substrate interactions have yet to include a robust analysis of distribution and dispersal of such interactions in field situations (Dighton and others 1997). As a whole, there is a need for fundamental research to improve understanding of the complex interactions that determine the removal of pollutants by natural communities. Such research might lack the glamour of manipulative genetic technologies (to produce recombinants with “designer” functions), and the approach does not have the same value as “intellectual property,” but it is likely to be much more profitable. As Price (1997) concludes in a recent review of bioremediation of marine oil spills, “understanding fundamental microbial ecology is the priority for commercial clean-up technology.” Discussion Diverse communities are likely to comprise commercially valuable individual species and strains. This is well documented and often cited to support “species conservation” (Wilson 1992). However, diverse microbial species assemblages can act in concert (via complex ecosystem-level interactions) to achieve ecologically and economically valuable processes. Furthermore, it is the innate capacity of

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Page 249 microbial assemblages to respond to change that makes the community as a whole valuable as a resource in combating pollutants, plant pests, and diseases. The gene frequency of degradative biochemical pathways appears to be characterized by high levels of “functional overlap”. Similarly, shared preferences for specific hosts are well documented among potential biological control agents. This could suggest “redundancy” in the normal community state but have considerable economic value in “variable” natural systems. Natural microbial communities comprise large numbers of unidentified or unculturable genotypes (Ward and others 1990) that can belie ecosystem integrity and contribute “function” after community disturbance. In fact, the nature and scale of microbial ecosystem processes have led some authors to conclude that such concepts as redundancy and species value have little meaning in the context of microbial community ecology (for example, Finlay and others 1997). We conclude that intrinsic processes and ecosystem interactions can have important value for human society. Often, it is the community function as a whole (not a particular or “valuable” species) that is important. The maintenance of biodiversity itself is important for benefits derived from microbial communities. Microbial diversity needs to be conserved not only for the benefit of individual species and genotypes that function within the community as a whole, but also because microbial “ecosystem services” are often carried out at the community level. Besides development of appropriate environmental-management strategies aimed at preserving and stimulating the activity of naturally occurring communities, monitoring techniques have high priority for research and technology development. Intrinsic cleanup and biocontrol processes must be tracked so that further steps can be taken when natural abatement fails, and sampling data are also likely to improve our fundamental understanding of ecosystem processes. References Atlas RM. 1995a. Bioremediation of petroleum pollutants. Intl Biodet Bioremed 35:317–27. Atlas RM. 1995b. Petroleum biodegredation and oil-spill bioremediation. Mar Poll Bull 31:178–82. Braddock JF, Lindstrom JE, Brown EJ. 1995. Distribution of hydrocarbon-degrading microorganisms in sediments from Prince William Sound, Alaska, following the Exxon Valdez oil spill. Mar Poll Bull 30:125–32. Chakrabarty AM, Friello DA, Bopp LH. 1978. Transposition of plasmid DNA segments specifying hydrocarbon degredation and their expression in various microorganisms. Proc Nat Acad Sci USA 75:3109–12. Crawford RL, Crawford DL (eds). 1996. Bioremediation: principles and applications. Cambridge UK: Cambridge Univ Pr. 400p. Coffin RB, Cifuentes LA, Pritchard RB. 1997. Assimilation of oil-derived carbon and remedial nitrogen applications by intertidal food chains on a contaminated beach in the Prince William Sound. Mar Environ Res 44:27–39. Crump DH. 1987. Effect of time sampling, method of isolation and age of nematode on the species of fungi isolated from females of Heterodera schachtii and H. avenae. Rev Nématol 10:369–73. Davies KG, Flynn CA, Laird V, Kerry BR. 1990. The life-cycle, population dynamics, and host specificity of a parasite of Heterodera avenae, similar to Pasteuria penetran. Rev Nématol 13:303–9. Deacon JW. 1991. Significance of ecology in the development of biological control agents against soil-borne plant pathogens. Biocont Sci Tech 1:5–20. de Leij FAAM, Davies KG, Kerry BR. 1992a. The use of Verticillium chlamydosporium Goddard

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