Impact of Agricultural Inputs on Soil Organisms

 
11/1/2006

Australian Journal of Soil Research

By L. Van Zwieten

External agricultural inputs such as mineral fertilizers, organic amendments, microbial inoculants, and pesticides are applied with the ultimate goal of maximizing productivity and economic returns, while side effects on soil organisms are often neglected. We have summarized the current understanding of how agricultural inputs affect the amounts, activity, and diversity of soil organisms. Mineral fertilizers have limited direct effects, but their application can enhance soil biological activity via increases in system productivity, crop residue return, and soil organic matter. Another important indirect effect especially of N fertilization is soil acidification, with considerable negative effects on soil organisms. Organic amendments such as manure, compost, biosolids, and humic substances provide a direct source of C for soil organisms as well as an indirect C source via increased plant growth and plant residue returns. Non-target effects of microbial inoculants appear to be small and transient. Among the pesticides, few significant effects of herbicides on soil organisms have been documented, whereas negative effects of insecticides and fungicides are more common. Copper fungicides are among the most toxic and most persistent fungicides, and their application warrants strict regulation. Quality control of organic waste products such as municipal composts and biosolids is likewise mandatory to avoid accumulation of elements that are toxic to soil organisms.

Additional keywords: fertilizer, compost, manure, biosolids, pesticide, soil biology.

Introduction

Agricultural inputs

External inputs to agricultural production systems include mineral fertilizers such as urea, ammonium nitrate, sulfates, and phosphates; organic fertilizers such as animal manures, composts, and biosolids; various other organic products such as humic acids and microbial inoculants, and pesticides including herbicides, insecticides, nematicides, fungicides, veterinary health products, and soil fumigants. All these products are applied with the ultimate goal of maximizing productivity and economic returns.

Mineral fertilizers are a major physical input into Australian agricultural production and account for over 12% of the value of material and services inputs used (Fertilizer Industry Federation of Australia Inc., www.fifa.asn.au). In 1999, Australian farmers used around 5.25 million t of fertilizer products with a value of approximately AU$2 billion. Common types of mineral fertilizers and their abbreviations as used in this review are shown in Table 1. Manures from intensive animal industries are a major source of organic amendments for agricultural land. In Australia, beef and dairy cattle alone produce approximately 4 million t of manure every year. Human waste is another important source. Sydney, Australia's largest urban area, produces 185 000 t of biosolids each year (Sydney Water Annual Report 2004). Nearly all of this is used for land amendment, either as dewatered solids, lime-stabilised solids, or in composts with green wastes. Pesticides are a diverse group of inorganic and organic chemicals. More than 380 active constituent pesticides are currently registered in Australia (Record of Approved Active Constituents at: www.apvma.gov.au). Pesticide inputs constitute a major cost for Australian agriculture. For herbicide inputs alone it was estimated at AU$571 million to annual winter crops in the 1998-99 growing season (Jones et al. 2005).

Soil organisms: groups, activities, methods

Soil organisms consist of the microflora (bacteria and fungi) and the soil fauna (protozoa and invertebrate groups such as nematodes, mites, and earthworms). They influence the availability of nutrients for crop production via a range of activities such as the decomposition of crop residues, immobilisation of nutrients, mineralisation, biological nitrogen fixation, and bioturbation. The soil fauna is crucial for the initial comminution and mixing of residues into the soil, whilst the microflora has a greater suite of enzymes for chemical breakdown of organic material (Paul and Clark 1996). Bacteria and fungi are often considered as a labile pool of nutrients (C, N, P, S) called the soil microbial biomass that has a pivotal role in nutrient immobilisation and mineralisation. The release of nutrients from the microbial biomass is partly regulated through grazing by the soil fauna.

The effect of agricultural inputs on soil organisms can be measured either as changes in the amount of single organisms, organism groups or methodologically defined pools such as the microbial biomass, or as changes in biological activity, e.g. soil respiration and enzyme activities. The most commonly used methods are listed and explained in Table 2. Variable effects of a given amendment on different organisms may change the composition of the microbial (or faunal) community without changing total amounts or activities. However, most studies have focussed on the soil microbial biomass as the central pool in nutrient cycling.

Concept of this review

In this paper we summerize the current understanding of the effects of inorganic and organic agricultural inputs on soil organisms. The underlying concept is that these inputs can affect soil organisms through direct or indirect effects (Table 3). Direct effects via changes in nutrient availability or toxicity will become already apparent in the first season after the application or in the longer term if repeated additions are required to reach a threshold above which effects are seen. Indirect effects will usually take more than one season to establish, especially when changes in soil organic matter levels are involved. In the case of long-term data, it can be difficult to separate direct and indirect effects.

Existing data are presented for the different amendments separately but discussed together. The evidence from Australia is rather limited, and therefore the review includes literature from overseas, in an attempt to establish the main principles and to draw some conclusions applicable to agro-ecosystems in Australia.

Mineral fertilizers

Most mineral fertilizer in Australia and elsewhere is applied to systems with regular and significant nutrient exports in harvested products, i.e. to grasslands and land under arable cropping. Experimental approaches to assess the effect of mineral fertilizers range from laboratory incubations, pot experiments, and 1-season studies in the field to long-term field experiments and sampling of paired sites under different management, thus covering time frames from days to more than 1 O0 years. In an attempt to separate direct from indirect effects, in the following sections we have compiled studies according to their experimental approach and time frame.

Laboratory incubations

Laboratory incubations allow the study of short-term effects under controlled conditions, i.e. in the absence of plants, climatic variation, and external inputs or losses. We found, however, very limited and often contradictory results from laboratory studies. For example, the addition of 200 mg N/kg soil as ammonium sulfate to 2 pasture soils of varying P status from New Zealand resulted in a decrease in microbial P, no change in the turnover of added C, and an increase in N mineralisation during 168 days of incubation (Saggar et al. 2000). An earlier study from New Zealand had, however, found an increase in soil respiration and microbial P, but no effect on microbial N and a decrease in various enzyme activities upon addition of 500 mg P/kg soil as calcium diphosphate (Haynes and Swift 1988). The addition of N, P, K, and S at 100, 20, 100, and 20mg/kg soil, respectively, to a range of soils from southern Australia, followed by incubation for 20 days, resulted in minor changes (increase or decrease) of soil respiration and microbial C, N, and P that remained within 20% difference from the non-amended controls (Bunemann, unpublished). Remarkably, changes in microbial C, N, and P were not interrelated.

Contradicting evidence such as an increase in microbial P while microbial C and N are unaffected might be interpreted as shifts in the composition of the microbial community. This possibility has been investigated in recent studies using biochemical markers and molecular techniques. The addition of N did not change the community composition as indicated by the phospholipid fatty acid (PLFA) profile in a study where total soil respiration was unaffected, but peroxidase activity and the preferential use of older, more stable soil organic matter increased after N addition (Waldrop and Firestone 2004). In 2 studies from Germany, ammonium addition did not change the composition of the microbial community during 28 days of incubation (Avrahami et al. 2003a), but led to community shifts after 16 weeks of incubation (Avrahami et al. 2003b). Using molecular techniques in a range of pot experiments, Marschner et al. (2004) showed that soil pH and N and P fertilization can affect the microbial community composition, but that substrate availability, e.g. in the form of root exudates in the rhizosphere, appears to be the main factor determining the community composition in the rhizosphere. It is thus important to consider the potential feedback from improved plant nutrition when examining fertilizer effects on soil organisms.

Pot experiments and field studies

Pot experiments (Table 4) have mainly been used to investigate the effect of mineral P and N fertilizer on root colonisation by arbuscular mycorrhizal fungi (AMF). Whereas the addition of mineral N did not affect AMF, increasing additions of inorganic P decreased the rate of root length colonisation in 2 cases (Ryan and Ash 1999; Rubio et al. 2003). A decrease in AMF root colonisation was also observed in pastures after 15-17 years of mineral P and N fertilization (Ryan et al. 2000).

Many field experiments have shown a lack of response of the microbial biomass and earthworms to mineral fertilizers (Table 4), even in cases where pasture production increased (e.g. Perrott et al. 1992; Sarathchandra et al. 1993). Where a decrease in microbial C was observed, it was usually accompanied by a decrease in soil pH after application of N or S fertilizers (e.g. Gupta et al. 1988; Ladd et al. 1994; Sarathchandra et al. 2001). Other methods such as microbial enumeration by plate counts (Sarathchandra et al. 1993), enzyme activities (Graham and Haynes 2005), and nematode counts (Parfitt et al. 2005), which are possibly more sensitive than measurements of microbial biomass, show variable changes due to mineral fertilization (Table 4). For example, although the total number of nematodes was not affected by N fertilization and a concomitant decrease in pH, some nematode species increased, whereas others were decreased (Sarathchandra et al. 2001).

The absence of changes in microbial C in response to N fertilization and a related decrease in pH in the 2 long-term field experiments studied by Moore et al. (2000) are interesting, because in the same study microbial C was found to be correlated to levels of organic C (OC) as induced by different crop rotations. Several long-term field experiments in which mineral and organic fertilizer inputs have been compared (Table 5) have likewise shown good correlations between the microbial biomass and soil organic C (Witter et al. 1993; Houot and Chaussod 1995; Leita et al. 1999). Although soil organic C levels are often increased by mineral fertilization compared with the non-fertilised control, even greater increases in soil organic C are usually achieved in treatments receiving organic amendments. This is also reflected in the fact that whereas mineral fertilizers show variable effects on soil organisms, organic amendments have only been reported to have insignificant or positive long-term effects (Table 5). The only exception was a decrease in microbial C after sewage sludge application, which also decreased soil pH (Witter et al. 1993). These observations point towards the role of C inputs, either with the organic amendment, or indirectly via increased plant growth and resulting plant residue input.

Graham et al. (2002) investigated the amounts of microbial C and N under sugarcane after 59 years of differential crop residue management and NPK fertilization and showed that the microbial biomass was directly influenced by residue management and indirectly by NPK fertilization through increased residue inputs. A follow-up study in the same trial revealed the interaction of soil acidification with negative effects and organic matter accumulation with positive effects on soil organisms and enzyme activities (Graham and Haynes 2005). The long-term field experiment studied by Houot and Chaussod (1995) exemplifies that agro-ecosystems can be relatively slow to respond to changes in management and thus illustrates the value of long-term field experiments. The excellent correlation between microbial C and soil organic C found after > 100 years of constant management practices remained disturbed 2 years after a change in crop rotation and crop residue management. The time required to reach a new equilibrium is a factor that may confound the results from many short-term studies.

Another potential indirect effect of fertilizer inputs was investigated in a long-term fertilization experiment without plants (Pernes-Debuyser and Tessier 2004). The comparison of various N, P, and K fertilizers, liming, and manure treatments revealed that ammonium fertilizers decreased pH and CEC, causing a degradation of hydraulic properties, whereas basic amendments increased pH and CEC. Aggregate stability was lowest in acid plots, intermediate in basic plots, and highest in plots treated with manure. A short-term study suggested that ammonium nitrate enhanced soil porosity by 18%, compared with 46% increase in a manure treatment. Since soil respiration almost doubled in the mineral fertilizer treatment compared with the unfertilised control, the authors discussed a potential priming effect of N addition on the decomposition of soil organic matter. Although such a priming effect is often observed (Kuzyakov et al. 2000), it seems to be rather short-lived, which might explain why we did not find much evidence for it (Table 4).

A decreased amount or activity of soil organisms after mineral fertilization could be due to the toxicity of metal contaminants contained in mineral fertilizers. In general, N and K fertilizers contain very low levels of contaminants, whereas P fertilizers often contain significant amounts of cadmium, mercury, and lead (McLaughlin et al. 2000). Metal contaminants are, however, most prevalent in waste products from urban and industrial areas and will be dealt with more in-depth in the section on organic fertilizers. Long-term chronic toxicity due to gradually accumulating metals appears to be far more common than immediate, acute toxicity (Giller et al. 1998). Quality control of fertilizer products is therefore required. This applies in particular to any new products. For example, the application of rare earth elements such as lanthanum, which is increasing in China, was shown to decrease soil respiration and dehydrogenase activity at high application rates (Chu et al. 2003). Such observations warrant more detailed investigation into processes of accumulation, bioavailability, and threshold levels of elements contained in fertilizers that can be toxic to soil organisms.

Organic fertilizers

Since most organic fertilizers are waste products, their application rate is often determined by availability rather than demand. Most amendments are applied primarily to benefit plant growth. In contrast to mineral fertilizers, however, effects on the soil's physical, chemical, and biological properties are sometimes intended as well (Table 6). In the following sections, we try to establish some links between the properties of various organic inputs and their effects on soil organisms.

Compostable organics

Compostable and composted materials vary widely in characteristics such as dry matter content, pH, salinity, carbon content, plant nutrient concentrations, non-nutrient elements, and microbial types, numbers, and activity. Although studies of amendments vary widely in nature of materials, application rates, and experimental conditions (Albiach et al. 2000), amendment with raw and composted organics generally results in increased microbial proliferation in the soil (Table 7). The duration of observed increases in soil organisms depends on the amount and proportions of readily decomposable carbon substrates added and the availability of nutrients, particularly nitrogen (Hartz et al. 2000; Adediran et al. 2003). However, microbial characteristics of amended soils often return to their baseline within a few years (Speir et al. 2003; Garcia Gil et al. 2004). Sustained changes in microbial biomass, diversity, and function are more likely where organic amendments are ongoing, as is the case in organic and biodynamic farms (Mader et al. 2002; Zaller and Kopke 2004). Ryan (1999) argues, however, that an increase in microbial populations may not be seen when system productivity is limited by nutrient input or water supply.

Manures and sewage sludge generally have higher salinity than municipal garden wastes, and salts can build up in soil with repeated heavy applications (Hao and Chang 2003; Usman et al. 2004). Sewage sludges (biosolids) often contain heavy metals such as copper, zinc, or cadmium, especially where industries contribute to the waste stream. Heavy metals can affect microbial processes more than they affect soil animals or plants growing on the same soils. For example, nitrogen-fixing rhizobia were far more sensitive to metal toxicity than their host plant clover. This resulted in N deficiency of clover due to ineffective rhizobia in sludge-amended soils (Giller et al. 1998). Sewage sludge and livestock manure may also contain active residues of therapeutic agents used to treat or cure diseases in humans and animals (Jjemba 2002). Green wastes from farms and gardens are typically lower in nutrient concentrations than manures or sewage sludges, but may contain residues of synthetic compounds such as herbicides, insecticides, fungicides, and plant growth regulators. Composting degrades some but not all such compounds, depending on the nature of the pesticide and the specific composting conditions (Buyuksonmez et al. 2000). Negative effects of heavy metals (Giller et al. 1998) can persist for many years following cessation of application (Abaye et al. 2005), since metals persist in soil practically indefinitely (McLaughlin et al. 2000). Such observations warrant strict regulations of organic fertilizer quality and applied quantity, especially of waste products such as sewage sludge and biosolids, in order to minimise contamination of agricultural land with toxic metals.

Humic substances

Humus in soil has traditionally been separated into humin, humic acid, and fulvic acid based on extraction with an alkaline solution and subsequent precipitation after addition of an acid (Swift 1996). The fractions typically rank in their resistance to microbial decomposition in the order humic acid > fulvic acid > humin (Qualls 2004). Concentrated sources of organic material such as peat, composts, and brown coal (oxidised coal, lignite, leonardite) also contain humic substances and are often marketed on the basis of their humic and fulvic acid contents as determined by similar procedures. Contents of humic acids vary, however, widely (Riffaldi et al. 1983). Some of the chemically extracted humic and fulvic acid separates are themselves sold as soil amendments. In discussion of organic amendments, a clear distinction must be made between products containing humic substances and those products that are humic (or fulvic) acids extracted from the primary sources listed above.

Humic substances can stimulate microbial activity directly through provision of carbon substrate, supplementation of nutrients, and enhanced nutrient uptake across cell walls (Valdrighi et al. 1996). Several studies showed that increasing amounts of compost or brown coal-derived humic acid stimulated aerobic bacterial growth, but had only slight effects on actinomycetes and no effect on filamentous fungi (Vallini et al. 1993; Valdrighi et al. 1995, 1996). Differences in microbial response were related to the molecular weight of the humic acids, with the lower weight fractions, typical of composts, causing greater microbial stimulation than the higher molecular weight fractions extracted from brown coal (Garcia et al. 1991; Valdrighi et al. 1995). Application of humic substances may induce changes in metabolism, allowing organisms to proliferate on substrates which they could not previously use (Visser 1985). Both heterotrophic and autotrophic bacteria can be stimulated by humic acid addition, mostly through the enhanced surfactant-like absorption of mineral nutrients, although heterotrophs also benefit from the direct uptake of organic compounds (Valdrighi et al. 1996). Vallini et al. (1997) showed that nitrifiers (chemotrophs) cannot use humic acids as an alternative carbon and energy source. Microbial activity may even be inhibited if humic acid is the sole carbon source (Filip and Tesarova 2004).

The principal indirect effects of humic substances on soil organisms are through increased plant productivity by mechanisms as listed in Table 6, but excessive applications can negatively affect plant growth (Fagbenro and Agboola 1993; Vallini et al. 1993; Valdrighi et al. 1995; Atiyeh et al. 2002), possibly through reduced availability of chelated nutrients (Chen et al. 2004). Field studies vary widely in the applied amounts of humic substances and in outcomes. Kim et al. (1997a) found no effect of commercial humate applied at 8.2t/ha on microbial activity or microbial functional groups (total fungi, actinomycetes, total Gram-negative bacteria, fluorescent pseudomonads, and P. cupsici) in a sandy soil used to grow bell peppers. Similarly, after 5 years of annual applications of 100 L/ha liquid humic acid to a horticultural soil, Albiach et al. (2000) found no effect on microbial biomass or enzyme activity. They ascribed the lack of effect to the low rates recommended by the manufacturer because of high product costs. Municipal solid waste compost and sewage sludge were more affordable and led to significant increases in microbial biomass in the same study. Only fungi were stimulated by humate added to soil being restored post-mining (Gosz et al. 1978), whereas Whiteley and Pettit (1994) found that lignite-derived humic acid inhibited decomposition of wheat straw. Chen et al. (2004) calculated from laboratory studies that 67.5 kg/ha of humic substances were needed for effective application to a sandy soil, but thought beneficial effects to plants may only occur in semi-arid or arid areas when applied in combination with irrigation and mineral nutrients.

Microbial inoculants

Inoculation with natural or genetically engineered microbial formulations can be broadly categorised according to whether they are intended to (a) exist on their own in the bulk soil, (b) populate the rhizosphere, (c) form symbiotic associations with plants, or (d) promote microbial activity on leaf or straw surfaces. To achieve the desired effect in the field, the inoculant organism must not only survive but establish itself and dominate in the soil or rhizosphere. Survival depends firstly on the quality of the inoculant itself, i.e. purity, strain trueness, viable numbers, the degree of infectivity, and level of contaminants (Abbott and Robson 1982; Kennedy et al. 2004). Secondly, the establishment and proliferation of inoculant in the soil environment are determined by many edaphic and climatic factors, the presence of host organisms (for symbionts and endophytes) and, most importantly, by competitive interactions with other microorganisms and soil fauna (Stotzky 1997; Slattery et al. 2001; McInnes and Haq 2003). Effects of inoculation on indigenous soil organisms can therefore either result from direct addition effects and interactions with indigenous soil organisms, or from indirect effects via increases in plant growth by one or several of the mechanisms listed in Table 6.

Positive effects of inoculants on the soil microbial biomass may be short-lived (Kim et al. 1997b), and increases in biomass or activity can even be due to the indigenous population feeding on the newly added microorganism (Bashan 1999). The most successful and widely studied inoculants are the diazotroph bacteria (Rhizobium, Bradyrhizobium, Sinorhizobium, Frankia) used for symbiotic fixation of [N.sub.2] from air. Provided soil conditions are favourable for rhizobia survival (Slattery et al. 2001), inoculation can increase microbial C and N in the rhizosphere compared with uninoculated soils (Beigh et al. 1998; Moharram et al. 1999). Population changes can be limited to the season of inoculation if the newly added organism is not as well adapted to the soil conditions as the indigenous population (McInnes and Haq 2003).

Inoculant application research is increasingly focussing on co-inoculation with several strains or mixed cultures enabling combined niche exploitation, cross-feeding, complementary effects, and enhancement of one organism's colonisation ability when co-inoculated with a rhizosphere-competent strain (Goddard et al. 2001). An example is the use of phosphorus-solubilising bacteria to increase available phosphorus along with mycorrhizae that enhance phosphorus uptake into the plant (Kim et al. 1997b). Saini et al. (2004) achieved maximum yields of sorghum and chickpea at half the recommended rates of inorganic fertilizer when a combination of mycorrhizae, [N.sub.2]-fixing bacteria, and phosphorus-solubilising bacteria was added. Increases in microbial biomass C, N, and P in soils of inoculated treatments were strongly correlated with N and P uptake of the plants. Garbaye (1994) suggested that specific 'helper' bacteria may improve the receptivity of the root to the fungus to enhance mycorrhizal colonisation and symbiotic development with plant roots (e.g. Founoune et al. 2002). Similarly, legume root nodulation can be enhanced by co-inoculation with Azospirillum, which increases root production and susceptibility for rhizobium infection and may also increase secretion of flavonoids from roots that activate nodulation genes in Rhizobium (Burdman et al. 1996). Conn and Franco (2004) found a significant reduction in indigenous actinobacterial endophytes upon inoculation of soil with a commercial multi-organism product, compared with no change in diversity after inoculation with a single species. Trial with 'effective microorganisms' (EM), a proprietary combination of photosynthetic bacteria, lactic acid bacteria, and yeasts used as a soil and compost inoculant, showed enhanced soil microbial biomass, plant growth, and produce quality (Daly and Stewart 1999; Cao et al. 2000). The interactions of microbial inoculants with indigenous soil organisms are likely to be complex, and a better mechanistic understanding is necessary to predict short- and long-term effects.

Pesticides

The results from our literature survey on the effects of selected pesticides on soil organisms are shown in Table 8 (herbicides), Table 9 (insecticides and nematicides), Table 10 (fungicides), and Table 11 (veterinary health products, fumigants, and biological/non-chemical products). Although more than 380 active constituent pesticides are currently registered in Australia, this current review has found data on the effects of only 55 of these on soil organisms. There is clearly a paucity of data in both the Australian and international literature on the effects of a large number of pesticides on soil organisms. Additional data may be available in the chemical reviews of the Australian Pesticide and Veterinary Medicines Authority (www.apvma.gov.au), but much of the information is contained within confidential company reports. Some of the chemicals such as DDT and chloropicrin are no longer registered for use in Australia; however, data have been included in this review as their use continues in many countries.

Herbicides

The herbicides (Table 8) generally had no major effects on soil organisms, with the exception of butachlor, which was shown to be very toxic to earthworms at agricultural rates (Panda and Sahu 2004). The authors showed, however, that butachlor had little effect on acetylcholinesterase activity. Butachlor is not registered for use in Australia. Phendimedipham induced avoidance behaviour in earthworms (Amorim et al. 2005) and collembola (Heupel 2002). These effects are expected to be relatively short lived, as phendimedipham is broken down moderately rapidly (25-day half-life) in soil (Tomlin 1997). Other effects of herbicides on soil organisms were mainly isolated changes in enzyme activities. Glyphosate, for example, was shown to suppress the phosphatase activity by up to 98% (Sannino and Gianfreda 2001) in a laboratory study; however, urease activity was stimulated by glyphosate as well as atrazine.

Insecticides

Insecticides (Table 9) were generally shown to have a greater direct effect on soil organisms than herbicides. Organophosphate insecticides (chlorpyrifos, quinalphos, dimethoate, diazinon, and malathion) had a range of effects including changes in bacterial and fungal numbers in soil (Pandey and Singh 2004), varied effects on soil enzymes (Menon et al. 2005; Singh and Singh 2005), as well as reductions in collembolan density (Endlweber et al. 2005) and earthworm reproduction (Panda and Sahu 1999). Carbamate insecticides (carbaryl, carbofuran, and methiocarb) had a range of effects on soil organsism, including a significant reduction of acetylcholinesterase activity in earthworms (Ribera et al. 2001; Pandey and Singh 2004), mixed effects on soil enzymes (Sannino and Gianfreda 2001), and inhibition of nitrogenase in Azospirillum species (Kanungo et al. 1998). Persistent compounds including arsenic, DDT, and lindane caused long-term effects, including reduced microbial activity (Van Zwieten et al. 2003), reduced microbial biomass, and significant decreases in soil enzyme activities (Ghosh et al. 2004; Singh and Singh 2005).

Fungicides

Fungicides (Table 10) generally had even greater effects on soil organisms than herbicides or insecticides. As these chemicals are applied to control fungal diseases, they will also affect beneficial soil fungi and other soil organisms. Very significant negative effects were found for copper-based fungicides, which caused long-term reductions of earthworm populations in soil (Van Zwieten et al. 2004; Eijsackers et al. 2005; Loureiro et al. 2005). Merrington et al. (2002) further demonstrated significant reductions in microbial biomass, while respiration rates were increased, and showed conclusively that copper residues resulted in stressed microbes. Other observed effects included the reduced degradation of the insecticide DDT (Gaw et al. 2003). These negative effects are likely to persist for many years, as copper accumulates in surface soils and is not prone to dissipative mechanisms such as biodegradation. Negative effects were also found for benomyl, which caused long-term reductions in mycorrhizal associations (Smith et al. 2000). Two fungicides, ehlorothalonil and azoxystrobin, have recently been shown to affect on a biocontrol agent used for the control of Fusarium wilt (Fravel et al. 2005), illustrating potential incompatibilities of chemical and biological pesticides.

Veterinary health products, soil fumigants, and non-chemical products

Veterinary health products (Table 11) include a range of nematicides, hormones, and antimicrobials. Data on the potential effect of these compounds on soil organisms are quite limited. The antimicrobials tylosin, oxytetracycline, and sulfachloropyridazine reduced Gram-positive bacterial populations and inhibited microbial respiration (Vaclavik et al. 2004), which is in accordance with changes in the microbial community structure after tylosin addition (Westergaard et al. 2001). The broad-spectrum anti-parasite Ivermectin was shown to be toxic to collembola at concentrations as low as 0.26 mg/kg soil; however, it was far less toxic to enchytraeid worms (Jensen et al. 2003) and earthworms (Svendsen et al. 2005).

Soil fumigants are designed to eliminate harmful soil organisms and any competition for soil resources between soil organisms and the crop. In spite of this, soil fumigants have not always been found to have significant effects on soil organisms (Table 11). Confirmed long-term effects on various soil functions (Karpouzas et al. 2005) are, however, a serious concern. The long-term effects of fumigants were shown to be reduced by the addition of composted steer manure, with normal biological activity being observed 8-12 weeks following high application rates of the fumigant (Dungan et al. 2003). In the absence of the organic amendment, little recuperation (resilience) of soil function was detected even after 12 weeks.

Microorganisms have been used to control plant diseases for over 100 years (Winding et al. 2004). However, risks of biological control agents are often forgotten. Although the selected microbes may occur naturally in the environment, there are concerns that altering the proportion of soil microbes will affect non-target species including mycorrhizal and saprophytic fungi, soil bacteria, plants, insects, aquatic and terrestrial animals, and humans (Brimner and Boland 2003). In a recent review of non-target effects of bacterial control agents suppressing root pathogenic fungi, Winding et al. (2004) concluded that significant non-target effects occurred that were, however, generally short lived. Residues from genetically modified maize expressing a protein from Bacillus thuringiensis (Bt) that is toxic to corn borers were found to decompose similarly to residues from conventional maize (Cortet et al. 2006), although the Bt toxin did inhibit some decomposition processes under laboratory conditions (Accinelli et al. 2004). Other methods for pest control include technologies such as solarisation (Table 11). This method uses plastic sheeting to heat-sterilise the surface soil. Several authors found reductions in microbial biomass and bacterial diversity (Gelsomino and Cacco 2006; Patricio et al. 2006).

Pesticide formulation

In addition to the active ingredient, the formulation of a pesticide may also influence soil organisms. This is, however, an aspect that is rarely investigated. Little is known about the environmental fate of adjuvants after application on agricultural land. Adjuvants constitute a broad range of substances, of which solvents and surfactants are the major types. Non-ionic surfactants such as alcohol ethoxylates (AEOs) and alkylamine ethoxylates (ANEOs) are typical examples of pesticide adjuvants (Krogh et al. 2003). Tsui and Chu (2003) demonstrated that the surfactant in the Roundup formulation polyoxyethylene amine (POEA) was significantly more toxic to Microtox bacterium than glyphosate acid or the IPA salt of glyphosate. Even Roundup was found to be less toxic. The toxicity of glyphosate acid was concluded to be a result of its inherent acidity. In another study, dos Santos et al. (2005) demonstrated that the presence of ethylamine in a glyphosate formulation had major effects on Bradyrhizobium, whereas the active ingredient (glyphosate) had little if any effect. In formulation, effects included reduced nodulation in a soybean crop.

General discussion

Main findings and knowledge gaps

In agreement with the main focus of this journal, we attempted to base our review primarily on results from Australia and New Zealand. However, we found that the existing database on the effect of agricultural inputs on soil organisms in this region was far too limited to draw sound conclusions. Even when considering the global literature, we identified several knowledge gaps.

There was little evidence for significant direct effects of mineral fertilizers on soil organisms, whereas the main indirect effects were shown to be an increase in biological activity with increasing plant productivity, crop residue inputs, and soil organic matter levels, and a depression with decreasing soil pH as a result primarily of N fertilization. This is in accordance with a review by Wardle (1992) who suggested that soil organic matter is the main factor governing levels of microbial biomass in soil, followed by soil pH. Long-term field experiments comparing mineral and organic fertilizers illustrated the role of indirect and direct carbon inputs into the soil in supporting biological activity. There is, however, a lack of such experiments in Australia and New Zealand.

Although direct C addition with the various organic amendments plays a major role in stimulating soil organisms, the role of C quality is not yet well understood. Compostable organics are an extremely diverse commodity with many potential benefits to soil organisms but also potential harmful effects, particularly with long-term application. Proper composting negates many potential harmful effects but not all. The toxic components that are not degraded or deactivated need to be identified and their specific effects better quantified. Australian standard AS 4454-2003 (Composts, soil conditioners and mulches) specifies threshold limits of heavy metals, pathogens, and organic compound contaminants based on demonstrated effects on plants and animals, not microorganisms, which may have a much lower threshold (Giller et al. 1998). As more and more of this material is used as a soil amendment rather than landfill, more research must be done on the long-term effects of the various contaminants on microorganisms.

The main problem with evaluating effects of specific products such as humic substances lies in the variety of materials of various origins, and in the fact that the properties are often defined by extraction methods that vary among laboratories and product manufacturers. Very few studies have investigated how humic substances affect soil organisms, and a closer examination of the effects of humic substances in laboratory cultures and soil cultures is required for an improved process understanding.

Microbial inoculants have mainly been studied under the aspect of inoculant survival and efficiency rather than with respect to effects on indigenous soil organisms. Apart from rhizobial and some mycorrhizal inoculants, much of the potential for microbial inoculants is yet to be realised. Possibly, the conventional scientific approach has been too reductionist, producing single strain organisms that often cannot compete in complex field situations (Marx et al. 2002). Since there is evidence that multi-organism products may be in a better position to compete with indigenous microorganisms, it is necessary to investigate the mechanisms in order to derive a causal understanding. Non-target effects of inoculants appear to be small and transient. However, Winding et al. (2004) point out that not enough is known about some marketed products aimed at disease control whose antimicrobial effects may extend beyond the growth season.

Among the pesticides, herbicides appeared to have the least significant effects on soil organisms, whereas some insecticides and especially some fungicides proved to be quite toxic. Few studies have investigated long-term effects of pesticide application, and even less discuss measured or observed changes to soil processes. One example is the lack of bioturbation noted recently in a copper-contaminated orchard (Van Zwieten et al. 2004). Copper has been shown to reduce the burrowing activity of earthworms, which in turn led to increased soil bulk density in a vineyard (Eijsackers et al. 2005). Likewise, Gaw et al. (2003) described the lack of pesticide breakdown in soils where copper was a co-contaminant. There is clear evidence that soil organisms and thus soil functions can be affected by pesticides, but comprehensive data showing which of these changes are long-term and reduce soil health are lacking.

Methodological issues

A broad range of tests has been used to evaluate effects of agricultural inputs on soil organisms, measuring the amount, activity, and diversity of soil organisms (Table 2). The lack of standardised methods often precludes a direct comparison between the various studies. Even if a similar method is used, slight variations in environmental conditions during the assay may change the outcome considerably, resulting, for example, in threshold levels of metal toxicity that can vary among studies by several orders of magnitude (Giller et al. 1999). Microbial endpoints have therefore sometimes been deemed to have limited use in risk assessment (Kapustka 1999). Ideally, endpoints should be highly sensitive to the respective contaminant while at the same time being robust, i.e. showing little variation among soils in the absence of the contaminant. However, when testing 8 ecotoxicological endpoints on 2 sets of soils, one metal-contaminated and one non-contaminated, Broos et al. (2005) observed a negative relationship between sensitivity and robustness of an endpoint. Therefore, a reasonable compromise might be to use endpoints of average sensitivity and good robustness. In their study, the lag-times of substrate-induced respiration, clover yield, and N fixation in clover were the most suitable endpoints for metal toxicity.

The most commonly measured variable, the microbial biomass, generally appears to be less sensitive to the various agricultural inputs than microbial activities such as soil respiration and enzyme activities. In the context of using microbial parameters to monitor soil pollution by heavy metals, Brookes (1995) suggested that the ratio of microbial activity and biomass, i.e. the metabolic quotient (Table 2), is more sensitive as an indicator of stress than either of the measurements alone.

Interpretation of enzyme activities in soil is complicated by the fact that enzymes may remain active when stabilised on organic matter or mineral surfaces. In addition, enzyme assays are usually based on the hydrolysis of artificial substrates such as p-nitrophenyl phosphate, but enzyme activity against natural substrates and under soil rather than assay conditions may be different. Enzyme activity against an artificial substrate must therefore be viewed as a potential activity and cannot be translated into actual reaction rates, and soil respiration may be a more direct measurement of microbial activity.

Methods to determine the microbial diversity have greatly advanced in recent years with the development of DNA-based techniques. However, even these methods still suffer from shortcomings such as the dependence of results on the extraction protocol (Martin-Laurent et al. 2001). Inoculation research has benefited from recent methodological advances, especially the development of molecular methods that allow following specific microorganism after addition into the soil--plant system (Marx et al. 2002; Conn and Franco 2004). Another technique is to genetically 'tag' newly released organisms to monitor the effects of introducing genetically modified organisms into the rhizosphere (Hirsch 2005). At the cellular level, direct staining techniques and advanced microscopy can provide high-resolution data on the metabolic activity and growth of inoculants (Schwieger et al. 1997).

Although laboratory studies are important to investigate basic processes, only field studies can fully elucidate the complex interactions of plants, soil, and climatic variation. Extrapolation from short-term tests is often not possible, especially when the mechanisms behind observed changes are not fully understood. This is especially true when long-term chronic toxicity poses a different stress on soil organisms than the immediate shock effect in laboratory tests (Giller et al. 1999). Only long-term monitoring in the field can provide the information required to establish regulatory guidelines, and an improved understanding of the system is mandatory for a sound risk assessment.

Interpreting changes in measured variables: where is the limit?

Our review has shown that most agricultural management strategies and external inputs can cause changes in the measured variables, whether they represent the amount, activity, or diversity of soil organisms. The challenge lies in interpreting the findings: we need to establish the limits for changes that are acceptable in view of that fact that agricultural inputs are a necessity, and those that are unacceptable, e.g. because they decrease biodiversity, impede soil functions, and diminish system productivity. Ultimately, the question is: what do we want to protect?

Terrestrial endpoints are often based on sensitive, threatened, and endangered species, such as the charismatic megafauna (Kapustka 1999). Measurements on soil organisms are, however, complicated by great spatial and temporal variation as well as complexity, since I g of soil can host more than 10 000 species of bacteria and an unknown diversity of fungi. In aquatic toxicology, an underlying assumption has sometimes been that if thresholds for toxic substances are based on the most sensitive species, then all species will be protected. However, the relative sensitivity of 2 species to chemical A may differ from that to chemical B. This concept is additionally complicated by the fact that an identified most-sensitive species may not be present in another ecosystem, making the application in regulatory terms questionable (Cairns 1986).

Protection of soil organisms based on their roles in nutrient cycling may be more practical and relevant for agroecosystems, even though it carries the risk that functional redundancy may mask changes in a population. Loss of specific functions that can only be carried out by very few species such as the loss of symbiotic nitrogen fixation due to application of metal-contaminated sewage sludge (Giller et al. 1998) or decreased decomposition due to detrimental effects of copper on earthworms (Van Zwieten et al. 2004) is obviously the biggest concern. Complete loss of function is, however, an exception rather than the rule.

When judging whether a change in a measured variable is of concern or not, the concept of Domsch et al. (1983) provides a good framework: a decrease in biological activity by up to 30% is deemed negligible, whereas a decrease by up to 90% could still be considered acceptable if it is followed by recovery within 30-60 days. This concept acknowledges the natural variation in many of the biological variables measured. It also places more emphasis on resilience than on resistance, where resistance is defined as the ability of the soil to withstand the immediate effects of perturbation, and resilience as the ability of the soil to recover from perturbation (Griffiths et al. 2001). Therefore, even laboratory tests should be run for a minimum of 30 days (Somerville et al. 1987). However, Giller et al. (1998) stress that a fundamental difference remains between acute toxicity (disturbance) and long-term chronic toxicity (stress), i.e. studying an adapting v. an adapted community. Thus, only long-term monitoring and field experiments can provide the information required to develop a sound risk assessment.

An increase in the amount, activity, or diversity of soil organisms is generally viewed as positive. However, an increase in the microbial biomass often goes along with increased nutrient immobilization, at least temporarily, and an increase in soil organic matter can increase populations of detrimental organisms such as parasitic nematodes and root diseases. As stated above, it is the resilience of the system that matters. In terms of biodiversity, a mild stress can actually increase species diversity by reducing competition effects, before diversity decreases at higher stress levels (Giller et al. 1998). This exemplifies the difficulties in interpreting changes, especially those in biodiversity.

Dahlin et al. (1997) observed that detrimental effects of metal contamination at one site were seen at metal concentrations below the background concentrations at the other site and asked in exasperation: 'Where is the limit?' One answer may be that there is no distinct threshold for metal toxicity, or for detrimental effects of other inputs, partly because the effects depend on site-specific characteristics such as climate and soil type. In testing procedures for the side effects of pesticides on soil microorganisms it has long been recognised that effects are more likely to be seen on light-textured soils that are low in organic matter than on heavier soils, and it is therefore recommended to use at least 2 contrasting soil types (Somerville et al. 1987). Likewise, changes in soil pH are more likely to have detrimental effects on soil organisms closer to the extreme points of the scale. For these reasons, it is mandatory to always choose a valid control, i.e. to allow for site-specific differences in the baseline, and to interpret changes in the context of the given site-specific characteristics.

An approach to assess the relative risk of pesticides to an agroecosystem (EcoRR) has been developed in Australia (Sanchez-Bayo et al. 2002). The methodology uses site-specific data and accounts for chemical dose, partitioning (air, soil, vegetation, surface and ground water), degradation, bioconcentration, and toxicity. Another model (PIRI) has been developed in Australia to assess the risk of pesticides entering groundwater (Kookana et al. 1998) and thus affecting the environment and human health. Neither of these models assesses, however, the risk of pesticides to soil organisms or even more broadly, soil quality.

Concluding remarks

The underlying principle for the protection of soil organisms should be to limit or prevent exposure of organisms to unacceptable hazards (McLaughlin et al. 2000). Our review has shown that some drastic negative effects such as those of copper fungicides and, to a lesser degree, soil acidification on soil organisms, have to be considered urgently if soil health is to be maintained. For some classes of inputs such as humic acids and various pesticides, the existing database is simply too small to draw sound conclusions. The main lesson learnt from the fertilizer section, however, is that any practice that increases levels of soil organic matter will also increase soil biological activity.

Acknowledgments

The senior author thanks the Grains Research and Development Corporation for support while this review was compiled. Questions and comments by 2 anonymous reviewers helped to improve the manuscript. We are also grateful to Kris Broos for providing us with relevant ecotoxicological references.

Manuscript received 30 August 2005, accepted 11 April 2006

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E. K. Bunemann (A,D), G. D. Schwenke (B), and L. Van Zwieten (C)

(A) School of Earth and Environmental Sciences, University of Adelaide, Adelaide, SA 5000, Australia.

(B) Tamworth Agricultural Institute, NSW Department of Primary Industries, Calala, NSW 2340, Australia.

(C) Wollongbar Agricultural Institute, NSW Department of Primary Industries, Wollongbar, NSW 2477, Australia.

Tables omitted.