Impact of agricultural inputs on soil organisms--a review.
External inputs to agricultural production systems include mineral fertilisers such as urea, ammonium nitrate, sulfates, and phosphates; organic fertilisers 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 maximising productivity and economic returns.
Mineral fertilisers 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 fertiliser products with a value of approximately AU$2 billion. Common types of mineral fertilisers 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 summarise 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.
Most mineral fertiliser 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 fertilisers 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 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 fertilisation 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 fertiliser 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 fertiliser 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 fertilisation (Ryan et al. 2000).
Many field experiments have shown a lack of response of the microbial biomass and earthworms to mineral fertilisers (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 fertilisers (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 fertilisation (Table 4). For example, although the total number of nematodes was not affected by N fertilisation 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 fertilisation 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 fertiliser 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 fertilisation 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 fertilisers 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 fertilisation and showed that the microbial biomass was directly influenced by residue management and indirectly by NPK fertilisation 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 fertiliser inputs was investigated in a long-term fertilisation experiment without plants (Pernes-Debuyser and Tessier 2004). The comparison of various N, P, and K fertilisers, liming, and manure treatments revealed that ammonium fertilisers 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 fertiliser 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 fertilisation could be due to the toxicity of metal contaminants contained in mineral fertilisers. In general, N and K fertilisers contain very low levels of contaminants, whereas P fertilisers 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 fertilisers. 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 fertiliser 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 fertilisers that can be toxic to soil organisms.
Since most organic fertilisers 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 fertilisers, 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 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 fertiliser 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.
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.
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 fertiliser 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.
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.
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 (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 (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).
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.
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 fertilisers 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 fertilisation. 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 fertilisers 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.
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 immobilisation, 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.
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 fertiliser section, however, is that any practice that increases levels of soil organic matter will also increase soil biological activity.
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.
(D) Corresponding author. Email: firstname.lastname@example.org
Table 1. Some common inorganic fertilisers and their abbreviations as used in this review Name Abbreviation Ammonium nitrate AN Ammonium sulfate AS Calcium nitrate CaN Diammonium phosphate DAP Elemental sulfur [S.sup.0] Phosphate rock PR Sodium dihydrogen phosphate NaHP Superphosphate SP Triple superphosphate TSP Urea U Table 2. Common methods to assess the amount, activity, and diversity of soil organisms Name Method description Reference Enumeration/amount Microbial C C in microbial biomass Vance et al. by fumigation- (1987), Islam extraction or and Weil (1998) microwave methods Microbial N N in microbial biomass Vance et al. by fumigation- (1987), Amato extraction methods and Ladd (1988) Microbial P P in microbial biomass Brookes et al. by fumigation- (1982), Kouno extraction methods et al. (1995) Adenosine triphosphate Extractable ATP Contin et indicates size of al. (2001) microbial biomass Total bacterial DNA PicoGreen dsDNA Angersbach and Earp (2004) CFU Colony forming units; e.g. plate counting Sarathchandra techniques, e.g. et al. (1993) Gram--ve bacteria, actinomycetes, fungi AMF Arbuscular mycorrhizal e.g. Ryan et fungi; usually root al. (2000) colonisation observed Soil fauna Nematodes, collembola e.g. Martikainen (springtails), et al. (1998), enchytraeids, Van Zwieten et earthworms al. (2003) (sometimes in combination with avoidance tests) Activity Soil respiration C[O.sub.2]-release Alef (1995) from incubated soil Metabolic quotient Ratio of soil Anderson and respiration to Domsch (1990) microbial C; higher values can indicate physiological stress Soil enzyme activities Dehydrogenase, acid Tabatabai (1994) and alkaline phosphatase, amidase, urease, arylsulfatase, etc. FDA hydrolysis Fluorescein diacetate Adam and (FDA) hydrolysis as Duncan (2001) a measure of total microbial activity Acetylcholin-esterase Activity of the enzyme e.g. Panda and activity in earthworms that hydrolyses Sahu (2004) acetylcholine in the nervous system; reduced activity indicates toxicity Diversity PLFA and FAME Phospholipid fatty Drenovsky et acid and fatty acid al. (2004) methyl esters and analysis: indicate changes in microbial community composition DGGE Microbial diversity e.g. Marschner assessed by DNA et al. (2004) extraction, amplification with polymerase chain reaction (PCR) and differentiation by denaturing gradient gel electrophoresis (DGGE) Biolog Soil microbial Konopka et substrate al. (1998) utilisation potential Table 3. Potential effects of inorganic and organic agricultural inputs on soil organisms Time frame (A) Direct effects * Increased amount and/or activity after Short- to long-term removal of nutrient limitations * Decreased activity due to high nutrient availability * Decreased amount and/or activity due to toxicity Indirect effects * Change in pH Long-term * Change in soil physical properties (aggregation, porosity) * Change in productivity, residue inputs, and soil organic matter levels (A) Short-term, 1 season; long-term, more than 1 season. Table 4. Effects of inorganic fertilisers on soil organisms as observed in pot experiments and field studies Reference and Soil type and Vegetation/ location characteristics (A) test plant Ryan and Ash Red-brown earth, Pot exp with white (1999), Australia pH 6, OC 26 g/kg clover and ryegrass Rubio et al. (2003), Volcanic soil, pH 5.5, Pot exp with wheat Chile 18% SOM Sarathchandra et al. Not given Pasture (1993), NZ Lovell and Hatch 36% clay Pasture (1997), UK Lupwayi et al. Gray Luvisols and Wheat-canola (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, Pasture NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric Pasture (2001), NZ Dystrochrept, sandy loam, pH 5.7, OC 63 g/kg Typic Hapludand, silt > 10 years loam, pH 4.9, OC 77 g/kg Gupta et al. (1988), Grey Luvisols, Canola-fallow Canada pH 5.4 and 5.7, rotation OC 11 and 34 g/kg Gupta and Germida Grey Luvisol, Pasture (1988); Gupta pH 5.7, OC 28 et al. (1988), g/kg Canada Ladd et al. (1994), Red-brown earth, Wheat rotations Australia 14% clay, pH 6.8, OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, Pasture NZ pH 5.6-6.0, OC 32-11 t/ha in 0-7.5 cm Ryan et al. (2000), Mostly red-brown Pasture Australia earths, mean pH 6.2, mean OC 2.6 g/kg Moore et al. (2000), Hapludoll, 22% clay Maize-soybean-oat- USA and 31% sand, meadow mean pH 6.9, rotations mean OC 21 g/kg Haplaquoll, 33% clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, Sugarcane (2002); Graham 58% clay, pH 5.8, and Haynes 2005, OC 42 g/kg S. Africa Reference and Soil type and Time location characteristics (A) frame Ryan and Ash Red-brown earth, 5 weeks (1999), Australia pH 6, OC 26 g/kg Rubio et al. (2003), Volcanic soil, pH 5.5, 6 months Chile 18% SOM Sarathchandra et al. Not given 2 weeks (1993), NZ 3 years Lovell and Hatch 36% clay 10 weeks (1997), UK Lupwayi et al. Gray Luvisols and 1-2 seasons (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, 2 years NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric 4 years (2001), NZ Dystrochrept, sandy loam, pH 5.7, OC 63 g/kg Typic Hapludand, silt 100 P (SP) loam, pH 4.9, OC 77 g/kg Gupta et al. (1988), Grey Luvisols, 2 years Canada pH 5.4 and 5.7, OC 11 and 34 g/kg Gupta and Germida Grey Luvisol, 5 years (1988); Gupta pH 5.7, OC 28 et al. (1988), g/kg Canada Ladd et al. (1994), Red-brown earth, 9-13 years Australia 14% clay, pH 6.8, OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, 7-23 years NZ pH 5.6-6.0, OC 32-11 t/ha in 0-7.5 cm Ryan et al. (2000), Mostly red-brown 15-17 years Australia earths, mean pH 6.2, mean OC 2.6 g/kg Moore et al. (2000), Hapludoll, 22% clay 19 years USA and 31% sand, mean pH 6.9, mean OC 21 g/kg Haplaquoll, 33% 44 years clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, 59-60 years (2002); Graham 58% clay, pH 5.8, and Haynes 2005, OC 42 g/kg S. Africa Reference and Soil type and Fertiliser location characteristics (A) (kg/ha) Ryan and Ash Red-brown earth, 180 N (B) (AN) (1999), Australia pH 6, OC 26 g/kg 200 P (B) (NaHP) Rubio et al. (2003), Volcanic soil, pH 5.5, 17-86 P (TSP, PR) Chile 18% SOM Sarathchandra et al. Not given 0-120 P (SP, PR) (1993), NZ Lovell and Hatch 36% clay 40 N (AN) (1997), UK Lupwayi et al. Gray Luvisols and 20 S ([S.sup.0] or AS) (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, 61 P (SP) NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric 400 N (U) (2001), NZ Dystrochrept, sandy loam, pH 5.7, OC 63 g/kg Typic Hapludand, silt loam, pH 4.9, OC 77 g/kg Gupta et al. (1988), Grey Luvisols, 50-100 S ([S.sup.0]) Canada pH 5.4 and 5.7, OC 11 and 34 g/kg Gupta and Germida Grey Luvisol, 44S ([S.sup.0]) (1988); Gupta pH 5.7, OC 28 et al. (1988), g/kg Canada Ladd et al. (1994), Red-brown earth, 0-80 N (AN) Australia 14% clay, pH 6.8, OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, 12-37 P, 12-26 S NZ pH 5.6-6.0, OC 32-11 t/ha in 0-7.5 cm Ryan et al. (2000), Mostly red-brown 27 P (SP, DAP), 17 N Australia earths, mean (U) pH 6.2, mean OC 2.6 g/kg Moore et al. (2000), Hapludoll, 22% clay 180 N (U) USA and 31% sand, mean pH 6.9, mean OC 21 g/kg Haplaquoll, 33% 180 N (U) clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, 140 N, 20 P, 140 K (2002); Graham 58% clay, pH 5.8, and Haynes 2005, OC 42 g/kg S. Africa Effect on soil organisms (amount and activity; % of control) Reference and Soil type and Negative location characteristics (A) Ryan and Ash Red-brown earth, (1999), Australia pH 6, OC 26 g/kg % Clover root length colonised by AMF (20-30%) Rubio et al. (2003), Volcanic soil, pH 5.5, % Root colonisation Chile 18% SOM (60-90%) Sarathchandra et al. Not given (1993), NZ Lovell and Hatch 36% clay (1997), UK Lupwayi et al. Gray Luvisols and Diversity (Biolog) (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric Microbial C (80%), (2001), NZ Dystrochrept, diversity (Biolog), sandy loam, Meloidogyne pH 5.7, OC 63 (10%). g/kg plant-associated (68%) and fungal-feeding (82%) nematodes Typic Hapludand, silt Microbial C, loam, pH 4.9, OC microbial 77 g/kg diversity (Biolog), total nematodes Gupta et al. (1988), Grey Luvisols, Fungal CFUs (38%), Canada pH 5.4 and 5.7, protozoa OC 11 and 34 (4-25%), g/kg microbial C (49-98%), respiration (67-86%) Gupta and Germida Grey Luvisol, Microbial C (60%), (1988); Gupta pH 5.7, OC 28 respiration (54%), et al. (1988), g/kg hyphal length Canada (24%), fungal CFUs (23%), protozoa (29-71%) Ladd et al. (1994), Red-brown earth, Microbial C Australia 14% clay, pH 6.8, (68-86%) OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, NZ pH 5.6-6.0, OC 32-11 t/ha in 0-7.5 cm Ryan et al. (2000), Mostly red-brown % Clover and grass Australia earths, mean root length pH 6.2, mean OC colonised by 2.6 g/kg AMF (67-79%) Moore et al. (2000), Hapludoll, 22% clay USA and 31% sand, mean pH 6.9, mean OC 21 g/kg Haplaquoll, 33% clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, Dehydrogenase, (2002); Graham 58% clay, pH 5.8, arylsulfatase, and Haynes 2005, OC 42 g/kg alkaline S. Africa phosphatase Effect on soil organisms (amount and activity; % of control) Reference and Soil type and No change location characteristics (A) Ryan and Ash Red-brown earth, % Clover and rye (1999), Australia pH 6, OC 26 g/kg grass root length colonized by AMF % Ryegrass root length colonised by AMF Rubio et al. (2003), Volcanic soil, pH 5.5, Chile 18% SOM Sarathchandra et al. Not given Microbial P, (1993), NZ earthworms Microbial P, earthworms Lovell and Hatch 36% clay Microbial C and N (1997), UK Lupwayi et al. Gray Luvisols and Microbial C (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, Microbial P and S NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric Total nematodes (2001), NZ Dystrochrept, sandy loam, pH 5.7, OC 63 g/kg Typic Hapludand, silt loam, pH 4.9, OC 77 g/kg Gupta et al. (1988), Grey Luvisols, Canada pH 5.4 and 5.7, OC 11 and 34 g/kg Gupta and Germida Grey Luvisol, Bacterial and (1988); Gupta pH 5.7, OC 28 actinomycetes et al. (1988), g/kg CFUs Canada Ladd et al. (1994), Red-brown earth, C and N Australia 14% clay, pH 6.8, mineralisation OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, Earthworms NZ pH 5.6-6.0, OC 32-11 t/ha in 0-7.5 cm Ryan et al. (2000), Mostly red-brown Australia earths, mean pH 6.2, mean OC 2.6 g/kg Moore et al. (2000), Hapludoll, 22% clay Microbial C USA and 31% sand, mean pH 6.9, mean OC 21 g/kg Haplaquoll, 33% Microbial C clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, Protease, respiration (2002); Graham 58% clay, pH 5.8, and Haynes 2005, OC 42 g/kg S. Africa Reference and Soil type and Positive location characteristics (A) Ryan and Ash Red-brown earth, (1999), Australia pH 6, OC 26 g/kg Rubio et al. (2003), Volcanic soil, pH 5.5, Chile 18% SOM Sarathchandra et al. Not given Fungi (480%), Gram (1993), NZ -ve bacteria -140% Lovell and Hatch 36% clay Nitrification, (1997), UK ammonification Lupwayi et al. Gray Luvisols and (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric Paratvlenchus (2001), NZ Dystrochrept, (1677%) sandy loam, pH 5.7, OC 63 g/kg Typic Hapludand, silt loam, pH 4.9, OC 77 g/kg Gupta et al. (1988), Grey Luvisols, Microbial S Canada pH 5.4 and 5.7, (136-168%), acid OC 11 and 34 phosphatase g/kg (106-130%) Gupta and Germida Grey Luvisol, Microbial S (178%), (1988); Gupta pH 5.7, OC 28 acid phosphatase et al. (1988), g/kg (141%) Canada Ladd et al. (1994), Red-brown earth, Australia 14% clay, pH 6.8, OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, Microbial P NZ pH 5.6-6.0, OC (168-300%), 32-11 t/ha in microbial N 0-7.5 cm (106-163%), total nematodes (123-223%) Ryan et al. (2000), Mostly red-brown Australia earths, mean pH 6.2, mean OC 2.6 g/kg Moore et al. (2000), Hapludoll, 22% clay USA and 31% sand, mean pH 6.9, mean OC 21 g/kg Haplaquoll, 33% clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, microbial C (2002); Graham 58% clay, pH 5.8, (119-136%), and Haynes 2005, OC 42 g/kg FDA hydrolysis S. Africa rate, acid phosphatase Reference and Soil type and Other changes location characteristics (A) Ryan and Ash Red-brown earth, (1999), Australia pH 6, OC 26 g/kg Rubio et al. (2003), Volcanic soil, pH 5.5, Chile 18% SOM Sarathchandra et al. Not given (1993), NZ Pasture production [up arrow] Lovell and Hatch 36% clay (1997), UK Lupwayi et al. Gray Luvisols and (2001), Canada Black Chernozems, pH 5.7 and 7.3, OC 28 and 47 g/kg Perrott et al. (1992), Yellow-brown earth, Herbage [up arrow] NZ silt loam, pH 5.8, OC 76 g/kg Sarathchandra et al. Umbric OC [down arrow], pH (2001), NZ Dystrochrept, [down arrow] by 0.4 sandy loam, units pH 5.7, OC 63 g/kg Typic Hapludand, silt loam, pH 4.9, OC 77 g/kg Gupta et al. (1988), Grey Luvisols, pH [down arrow] by Canada pH 5.4 and 5.7, 0.15 units OC 11 and 34 g/kg Gupta and Germida Grey Luvisol, pH [down arrow] by 1.0 (1988); Gupta pH 5.7, OC 28 units, OC [down et al. (1988), g/kg arrow] Canada Ladd et al. (1994), Red-brown earth, pH [down arrow] by Australia 14% clay, pH 6.8, 0.4-1.0 units OC 10 g/kg Parfitt et al. (2005), Typic Dystrudepts, NZ pH 5.6-6.0, OC 32-11 t/ha in 0-7.5 cm Ryan et al. (2000), Mostly red-brown Australia earths, mean pH 6.2, mean OC 2.6 g/kg Moore et al. (2000), Hapludoll, 22% clay pH [down arrow] by USA and 31% sand, 0.1-0.9 units; mean pH 6.9, microbial C mean OC 21 g/kg related to OC Haplaquoll, 33% clay and 24% sand, mean pH 6.2, mean OC 34 g/kg Graham et al. Chromic Vertisol, OC [up arrow], pH [down (2002); Graham 58% clay, pH 5.8, arrow] by 0.7 units and Haynes 2005, OC 42 g/kg S. Africa (A) Soil classification, texture, pH, and OC content as far as given. (B) mg/kg soil. Table 5. Comparative effects of inorganic and organic fertilisers on soil organisms as concluded from field experiments FYM, farmyard manure Reference Soil type and Vegetation Time frame and location characteristics (A) (years) Peacock et al. Typic Fragiudalf, silt Maize-pasture 5 (2001), USA loam, pH 6.0, OC rotation 15 g/kg Leita et al. Calcic Cambisol, Cover crop 12 (1999), 29% clay, 47% Italy sand, pH 7.8, OC 7,9 g/kg Witter et al. 35% clay, 21% sand, Arable crops 34-36 (1993), pH 6.2, OC Sweden 10 g/kg Dick et al. Haploxeroll, Wheat fallow 55 (1988), USA coarse-silty, rotation pH 6.5 Parham et al. Paleustoll, 23% clay, Wheat 69 (2002, 38% sand, pH 5, 2003), USA OC 6.7 g/kg 100 Colvan et al. Clay loam, pH 5.2, Pasture 100 (2001); OC 34 g/kg O'Donnell et al. (2001), UK Houot and Agrudalf, 22-30% Wheat-sugar 112 Chaussod clay, 40-44% beet (1995), sand, pH 8.0-8.3 France Reference Fertiliser Negative and location (kg/ha.year) Peacock et al. 18 N (AN) Gram -ve bacteria (2001), USA (85%) FYM (252 N) Leita et al. 100 N (AN), 75 P (1999), (SP), 150 K (PS) Italy Compost (500-1500 N) FYM (500 N) Witter et al. 80 N (CaN) (1993), 80N(AS) Microbial C(13%) Sweden FYM (4 t/ha.year) Sewage sludge Microbial C (69%) 4 t/ha.year) Dick et al. 90N Amidase, urease Straw + manure Parham et al. 67 N, 15 P, 28 K Fast-growing (2002, bacteria 2003), USA FYM (269 N/ha.4 years Colvan et al. 35 N (AS) Microbial P, C (2001); 60 P O'Donnell et 67 K al. (2001), 35 N, 60 P, 67 K UK FYM (20 t/ha.year) Houot and 87 N, 40 P, 75 K Chaussod FYM (1 t/ha.year) (1995), France Effect on soil organisms (amount and activity; % of control) Reference No change Positive and location Peacock et al. Total PLFAs Gram +ve bacteria (2001), USA (120%) Gram +ve Total PLFAs bacteria (170%), Gram -ve bacteria (115%) Leita et al. Microbial C (1999), Microbial C Italy (200-350%) Microbial C (243%) Witter et al. Microbial C (142%) (1993), Sweden Microbial C (209%) Dick et al. Acid and alk. (1988), USA phosphatase, arylsulfatase, glucosidase Acid and alk. phosphatase, arylsulfatase, glucosidase, amidase, urease Parham et al. Bacterial and fungal Microbial C, acid (2002, CFU, alk. phosphatase, 2003), USA phosphatase, slow-growing phosphodiesterase, bacteria pyrophosphatase Fungal CFU, acid Microbial C, alk. phospbatase, phosphatase, fast-growing phosphodiesterase, bacteria pyrophosphatase, bacterial CFU, slow-growing bacteria Colvan et al. Phosphatase (2001); Phosphatase Microbial P O'Donnell et Phosphatase al. (2001), Microbial P, UK phosphatase Microbial P, phosphatase Houot and Microbial C Chaussod Microbial C (1995), France Other Reference changes and location Peacock et al. pH [down arrow] by (2001), USA 0.6 units OC [up arrow] Leita et al. Microbial C (1999), correlated to OC Witter et al. pH [down arrow] with (1993), AS and sewage sludge; Sweden microbial C and respiration related to OC Dick et al. pH [down arrow] by (1988), USA 0.6 units pH [up arrow] by 0.6 units, OC [down arrow] Parham et al. pH [down arrow] by (2002, ~0.5 units 2003), USA pH [up arrow] by ~0.6 units Colvan et al. pH [down arrow] by (2001); 2 units O'Donnell et al. (2001), UK pH [up arrow] by Houot and 0.6 units Chaussod Microbial C related (1995), to OC France (A) Soil classification, texture, pH, and OC content as far as given. Table 6. Intended benefits of organic amendments Reason for organic amendment Examples (a) Supply bulk nutrients for Animal manures, sewage sludge, and plant production other composted organics supply N, P, K for plant uptake (b) Increase availability of Bacteria solubilise P and S from soil existing soil nutrients minerals (Grayston and Germida 1991; Gyaneshwar et al. 2002). Mycorrhizae extend root exploration and uptake of immobile nutrients (Dodd and Thomson 1994) (c) Increase the availability Humic acid products may increase of applied fertilisers fertiliser P availability (Delgado et al. 2002) (d) Fix N from air Symbiotic and free-living [N.sub.2]- fixing bacteria (Brockwell 2004; Kennedy et al. 2004) (e) Improve soil chemical Manure, sewage sludge, and compost can fertility increase soil organic matter and cation exchange capacity. Humic substances can enhance micronutrient availability (Chen et al. 2004) (f) Improve soil physical Mulches prevent erosion and improve condition water infiltration and water storage (Buerkert et al. 2000). Manures and mycorrhizae enhance aggregate stability and pore structure (Tisdall and Oades 1982) (g) Improve soil biology Manures and composts can add significant quantities of readily decomposable C substrate for microbes, and add microbes as well (Semple et al. 2001). 'Helper' bacteria can stimulate mycorrhizal and rhizobial symbioses (Garbaye 1994) (h) Plant growth promoters Rhizobacteria and possibly humic substances can supply plant growth-promoting hormones (Bowen and Rovira 1999) (i) Direct suppression of Composted manure and brewed compost plant disease leachates may suppress plant diseases (Scheuerell and Mahaffee 2002). Mycorrhizal fungi can control nematodes and root diseases (Siddiqui and Mahmood 1995; Whipps 2004) (j) Indirect suppression of Rhizobacteria can be added to seed or plant disease soil to enhance plant resistance to disease. Organic substrates may stimulate plant-beneficial microbial populations (k) Decontaminate polluted Microbially catalysed reactions in soils soil can breakdown organic pollutants or precipitate metals making them unavailable for plant uptake or water transport (Romantschuk et al. 2000) (l) Degrade crop residues and Microbial inoculants may enhance other compostable breakdown of crop residues and waxes materials that cause water repellency (Damodaran et al. 2004; Roper 2004) Table 7. Effects of animal manures, biosolids, and composts on soil organisms Reference and location Soil type and Compared treatments characteristics Trochoulias et al. Red basaltic soil Poultry manure, (1986), Australia gypsum + dolomite, others Poll et al. (2003), Luvic Phaeozem (FAO), Long-term annual Germany 8% clay, 72% sand, application of pH 5.6, OC 10 g/kg farmyard manure, control Dinesh et al. (2000), 5 soils; pH 5.7-6.4, 3 years poultry India OC 6-9 g/kg manure, FYM, sesbania and gliricidia residues, control Wu et al. (2004), 3 soils: Calcaric Manure, mineral China Cambisols, Haplic fertiliser,combined Greyxems, and manure and Calcic Kastanozems fertiliser (FAO); 25-30% clay, pH 8.3-8.4, OC 7-17 g/kg Min et al. (2003), Mesiq Achic 5 years of dairy USA Hapludult, fine manure slurry, loamy, pH 6.6, OC mineral fertiliser, 15.8 g/kg control Thomsen et al. (2003), 3 soils: 11-34% clay, Lab. incubation of Denmark 11% sand, pH soil amended with 6.4-7.4, OC sheep manure at 13.7-15.4 g/kg various soil matric potentials and clay contents Yang et al. (2003), Thermic Typic Surface mulches of USA Xerothent; coarse grass clippings, loam; pH 8.1, OC lucerne stems, 2 g/kg composted manure, eucalyptus, oleander or pine chip waste, chipped construction waste Villar et al. (2004), 3 Typic Haplumbrepts; Single application of Spain sandy loam to poultry manure, NPK sandy clay loam; fertiliser to soil pH 4.6-6.3; OC after wildfire 7.1-19.9 g/kg Baker et al. (2002), 3 soils: Aeric Biosolids (30120 t/ha) Australia Kandiaqualf, Typic Natraqualf, Oxic Ustropept Munn et al. (2001), 6 soils: 6-54% clay, Single application of Australia pH 4.2-6.4, OC biosolids from 5 7-32 g/kg treatment plants applied to one soil. Biosolids from 1 plant applied to 6 soils Abaye et al. (2005), Typic Udipsamment; Long-term FYM, England sandy loam; 8% metal-contaminated clay; pH 6.5-7.1; sewage sludge, NPK OC 5.2-14.8 g/kg mineral fertiliser Chaudhuri et al. Acid lateritic soil; Several combinations (2003), India pH 5.2; OC 5.4 g/kg of sludge and coal ash, control, NPK fertiliser Usman et al. (2004), Calcareous soil; Short-term incubation, Germany 3.5% clay, 87% sewage sludge, sand; pH 8.16; OC composted turf and 3.1 g/kg plant residues Alvarez et al. (1999), Thermic Vertic 2 years of sewage Argentina Argiudol; clay sludge applied to 42%, sand 14%; pH de-surfaced soils 5.8; OC 17 g/kg Barbarick et al. 2 soils: Aridic Single application of (2004), USA Argiboroll, Aridic biosolids Argiustoll; pH 7.3, 5.9; OC 1.5 g/kg Garcia Gil et al. Calcareous sandy Single application of (2004), Spain loam; sand 41%, sewage sludge clay 29%; pH 8.1; OC 10.1 g/kg Speir et al. (2003), Typic Udipsamment; Compost of biosolids, New Zealand coarse sand; pH wood waste 6.1; OC 39 g/kg and green waste Canali et al. (2004), Sandy loam; pH 7.8; Composts of distillery Italy OC 17.3 g/kg waste and livestock manure, poultry manure, mineral fertiliser control Wells et al. (2000), Luvic Ferrasol Composts of woody Australia (yellow earth); 77% material with either sand, 15% clay; pH manure (poultry and 5.4; OC 11.9 g/kg horse) or sewage sludge, several mineral fertiliser treatments Franco et al. (2004), 10 soils: 4 Glucose, maize stalks, Italy Inceptisols, 3 or maize stalk Mollisols, 3 compost added to Entisols; 10-60% soils contaminated clay; pH 5.2-8.3; with crude oil OC 13.6-58.5 g/kg Lalande et al. (2003), Orthic Humo-Ferric Single application of Canada Podzol; loamy co-composted sand; pH 5.4; OC papermill sludge 26-35 g/kg and hog manure applied alone or in combination with mineral fertilisers Zaller and Kopke Fluvisol; pH 5.35 9-year study of (2004), Germany traditionally composted FYM, 2 types of biodynamically composted manure Miyittah and Typic Hapludand; pH Composts of soymilk Inubushi (2003), 4.87; OC 64.9 g/kg residues, cow Japan manure, poultry manure, and sewage sludge Tiquia et al. (2002), Silt loam; 29% sand, Soil mulched with USA 29% clay; pH 5.5; composted yard OC 29 g/kg waste, ground wood pallets, bare soil control, with or without chemical fertiliser Reference and location Effects Trochoulias et al. Manured treatment had highest (1986), Australia microbial C Poll et al. (2003), Manure addition enhanced Germany microbial biomass and xylanase and invertase activity Dinesh et al. (2000), Organic manures increased India microbial biomass, activity, diversity, and C turnover Wu et al. (2004), Manure [+ or -] N and P fertiliser China treatments restored OC and microbial C to the level of the native sod Min et al. (2003), Dairy manure slurries increased OC USA and microbial biomass and decreased metabolic quotient compared with mineral fertiliser treatments Thomsen et al. (2003), Manure increased soil respiration in Denmark all combinations of soils and matric potentials. Microbial biomass increased most with the addition of manure to the sandiest soil Yang et al. (2003), Only grass clippings stimulated USA dehydrogenase activity in the soil measured after 1 year. Eucalyptus yardwaste and grass clippings caused shifts in bacterial populations and increased bacterial diversity but only at the soil surface Villar et al. (2004), Poultry manure application Spain increased microbial biomass C, particularly at high dose. Little or no changes as a consequence of inorganic fertilisation Baker et al. (2002), Increase in earthworm abundance Australia Munn et al. (2001), Symbiotic effectiveness of Australia rhizobium dependent on soil type and level and source of biosolids, not on basis of heavy metal concentrations Abaye et al. (2005), Microbial biomass-C and total England bacterial numbers greater in the FYM-treated soil than in NPK and sludge-amended soils. Relatively small heavy-metal concentrations decreased microbial C and bacterial numbers, increased metabolic quotient, and changed microbial community 40 years after metal inputs ceased Chaudhuri et al. Microbial C and soil enzyme (2003), India activities increased with all amendments; highest at equal proportions of coal ash and sludge. Mobile fractions of Cd and Ni correlated with microbial C Usman et al. (2004), Compared with compost, sewage Germany sludge caused greater increases in soil respiration, microbial C, and metabolic quotient, especially with increasing application rate Alvarez et al. (1999), Microbial biomass not affected by Argentina sludge, but metabolic activity and organic matter mineralisation enhanced. Increased soil respiration from sludge-amended soil represented 21% of C applied that year and 15% of C applied the year before Barbarick et al. 6 years after application, amended (2004), USA plots had increased microbial respiration, nitrogen mineralisation, root colonisation by AM, microbial biomass. No change in metabolic quotient Garcia Gil et al. Microbial biomass, basal (2004), Spain respiration, metabolic quotient, and enzymatic activities increased in soil 9 months after sludge application, but increases had disappeared after 36 months, presumably due to the loss of energy sources Speir et al. (2003), Soil basal respiration, microbial C, New Zealand and anaerobically mineralisable N were significantly increased in the amended plots. No effects on rhizobial numbers or microbial biosensors (Rhizotox C and lux-marked Escherichia coli) Canali et al. (2004), Parameters related to potentially Italy mineralisable C showed significant differences among the treatments. No differences were observed in biodiversity indexes Wells et al. (2000), Both composted treatments higher Australia in microbial C than mineral fertiliser treatments, but trial was of systems so there were also other differing factors Franco et al. (2004), The addition of organic substrates Italy (glucose, maize stalks, and maize stalk compost) to contaminated soils had no synergistic effect on the decomposition of crude oil but produced a marked increase in microbial biomass, although the increase was smaller than in uncontaminated soils. Compost decreased the stress conditions caused by oil contamination as measured by a reduction in metabolic quotient Lalande et al. (2003), Activities of [beta]-glucosidase, Canada [beta]-galactosidase, acid phosphatase, urease, and fluorescein diacetate hydrolysis, microbial C and soil respiration all increased compared with the control. Addition of fertiliser to compost resulted in a greater increase in enzyme activities than compost alone but had little effect on microbial biomass. Enzyme activities and microbial biomass decreased in the second season Zaller and Kopke FYM increased microbial biomass, (2004), Germany dehydrogenase activity, decomposition (cotton strips), but not saccharase activity, microbial basal respiration, or metabolic quotient. Biodynamic manure preparation decreased soil microbial basal respiration and metabolic quotient compared to non-biodynamic manure. After 100 days, decomposition was faster in plots which received biodynamic FYM than in plots which received no or non-biodynamic FYM Miyittah and Soil respiration increased rapidly Inubushi (2003), initially, but patterns differed Japan among the composts. Composted soymilk treatment gave higher C[O.sub.2]-evolution and lower metabolic quotient than the other composts Tiquia et al. (2002), Microbial respiration rate was USA highest in soils mulched with composted yard wastes. Mulching with compost strongly influenced the structure of the microbial rhizosphere community Table 8. Impact of herbicides on non-target soil organisms Reference and Soil type and Active chemical location characteristics Sannino and Gianfreda 22 soils from Campania Atrazine (2001), Italy region, sandy to clayey soils, OC 1.2-34.2 g/kg Seghers et al. (2003), Soils from a field site in Atrazine, Belgium Melle, Belgium. No metolachlor further descriptions Panda and Sabo (2004), Soil from upland non- Butachlor India irrigated paddy field. Sandy-loam with pH 6.8, OM 2.7% Araujo et al. (2003), Two Brazilian soils: sandy Glyphosate Brazil clay, pH 5.9, OM 2.3%, and clay soil, pH 5.2, OM 2% Busse et al. (2001), Three soils: 18-34% clay, Glyphosate USA (forestry) pH 5.4-5.9, OM 2.0-6.7% Sannino and Gianfreda Described above Glyphosate, (2001), Italy paraquat Dalby et al. (1995), Yellow, duplex loam Glyphosate and Australia (Palexeralf) 2,4-DB Reid et al. (2005), UK Three soils: Icknield Isoproturon (silty clay loam, OM 4.8%), Shellingford (sandy loam, OM 4.3%) and Evesham (clay, OM 6.0%) Mosleh et al. (2003), 43.9% clay, 28.7% sand, pH Isoproturon France 8.16, OC 7.7% Das et al. (2003), Clayey Typic Fluvaquent, Oxyfluorfen, India pH 7.1, OC 5.8% oxadiazon Strandberg and Review paper covering Pendimethalin Scott-Fordsmand several soil types (2004), Denmark Amorim et al. (2005), OECD standard soil: pH 6, Phenmedipham OECD standard soil OM 8%. 17 other test and several European soils: pH 3.2-6.9, OM test soils 1.7-15.9% Heupel (2002), Laboratory standard soil Phenmedipham Germany: standard described as a loamy sand soil Kinney et al. (2005), Remmit fine sandy loam Prosulfuron USA (Ustollic camborthids) Reference and Effects location Sannino and Gianfreda Significant activation of soil urease (2001), Italy activity (up to 100-fold increase), and suppression of invertase enzyme Seghers et al. (2003), Altered community structure of Belgium several groups of bacteria and actinomycetes Panda and Sabo (2004), Very toxic, suppressing growth, India sexual maturation and cocoon production of the earthworm Drawida willsi following single dose at recommended rate Araujo et al. (2003), Bacteria reduced. Fungi and Brazil actinomycetes increased. Microbial activity increase by 9-19%. Increased glyphosate degradation with repeated application Busse et al. (2001), Short term changes to community USA (forestry) structure. Increased microbial activity and no long-term changes to community structure Sannino and Gianfreda Activation of urease and invertase (2001), Italy soil enzymes, but glyphosate suppressed phosphatase activity (up to 98%) Dalby et al. (1995), No effect of single dose to soil on Australia growth or survival of the earthworms Aporrectodea trapezoides, A. caliginosa, A. longa or A. rosea Reid et al. (2005), UK Catabolic activity induced in soils not previously treated with this herbicide Mosleh et al. (2003), Affected earthworms at very high France soil concentrations (not agricultural rates) with LC50 for Eisenia fetida > 1000 mg/kg Das et al. (2003), Both herbicides stimulated India microbial populations, and increased availability of phosphorus in rhizosphere soil of rice Strandberg and Soil nematodes and other Scott-Fordsmand invertebrates reduced, (2004), Denmark plant-rhizobium symbiosis reduced at herbicide rates as low as 0.5-1.0 kg/ha Amorim et al. (2005), Enchytraeid worms avoided these OECD standard soil chemicals in standard avoidance and several European test procedures. In some soil test soils types, avoidance behaviour exhibited at low concentrations (1 mg/kg) Heupel (2002), Dose-dependent avoidance of the Germany: standard collembolan Isotoma anglicana, soil Heteromurus nitidus, Lepidocyrtus violaceus, Folsomia candida, and Onychiurus armatus Kinney et al. (2005), Significant reduction in production USA of [N.sub.2]O and NO following N-based fertiliser application: significant reduction in nitrification Table 9. Impact of insecticides and nematicides on non-target soil organisms Reference and Soil type and Active chemical location characteristics Hart and Brookes Silty clay loam, Aldicarb, chlorfenvinphos (1996), UK pH 6.4, OC 13.6 g/kg Van Zwieten et al. Sand, sandy clay Arsenic (2003), Australia loam and clay (contaminated site) loam soils. No further details Ghosh et al. (2004), Range of clay loam Arsenic India to clay soils, pH 6.9-7.5, OC 8.7-10.7% Amorim et al. (2005), Described Benomyl OECD standard soil previously and several European test soils Ribera et al. (2001), OECD artificial Carbaryl France: OECD soil was standard soil prepared Sannino and Gianfreda Described Carbaryl (2001), Italy previously Kanungo et al. (1998) Typic Haplaquept Carbofuran (deltaic alluvium), pH 6.7, OM 17% Panda and Sahu (2004), Described Carbofuran, malathion India previously Pandey and Singh Sandy loam, Chlorpyrifos, quinalphos (2004), India pH 6.75, OC 0.49% Menon et al. (2005), Loamy sand, pH Chlorpyrifos, quinalphos India 8.2, and sandy loam, pH 7.7, both soils from semi-arid regions Endlweber et al. No data on soil Chlorpyrifos, dimethoate (2005), Germany types provided Edvantoro et al. 11 soils, sandy to DDT, arsenic (2003), Australia clayey, pH contamination (contaminated site) 4.9-6.0, OC 0.14-5.2% Megharaj et al. (2000), Sandy soil, pH DDT Australia 7.1, OC (contaminated site) 1.77-3.6% Singh and Singh Silty sand, pH Diazinon (2005), India 6.98-7.22, OM 0.63-0.93% Martikainen et al. Soil from a Dimethoate (1998), Finland pesticide free grain field in central Finland, no further description Dalby et al. (1995), Described Dimethoate Australia previously Singh and Singh Described Imidacloprid (2005), India previously Capowiez and Berard Artificial soil Imidacloprid (2006), France with pH 8.3 Singh and Singh Described Lindane (2005), India previously Loureiro et al. Silty sand, pH Lindane, dimethoate (2005), Portugal 5.03, OM 1.28%. Also soils from metal contaminated mine site, pH 4.14-4.47, OM 2.88-5.07% Panda and Sahu Sandy loam, pH Malathion (1999), India 6.8, OM 2.7% Reference and Effects location Hart and Brookes Aldicarb caused a long-term (1996), UK increase of 7-16% in microbial C. No other effects found on respiration or N mineralisation Van Zwieten et al. Arsenic co-contamination was (2003), Australia shown to inhibit the (contaminated site) breakdown of DDT, and a concomitant reduction in microbial activity was found Ghosh et al. (2004), Arsenic between 11-36 mg/kg India in soil reduced microbial biomass, respiration, fluorescein diacetate hydrolysis and dehydrogenase activity, and induced microbial stress measured by increased metabolic quotient Amorim et al. (2005), Enchytraeid worms avoids OECD standard soil benomyl in standard and several European avoidance test procedures test soils Ribera et al. (2001), Significant reductions in France: OECD acetylcholinesterase and standard soil other biotransformation enzymes in earthworms Sannino and Gianfreda Activation of urease and (2001), Italy invertase soil enzymes, but suppression of phosphatase enzyme Kanungo et al. (1998) Appears to have an inhibitory effect on nitrogenase activity in Azospirillum sp. at higher application rates Panda and Sahu (2004), Significant reduction in India acetylcholinesterase activity in earthworms (D. willsi) for up to 45 days (carbofuran) and 75 days (malathion) Pandey and Singh Reduced bacterial numbers, but (2004), India significantly increased fungal numbers with chlorpyrifos and slightly reduced fungal numbers (short-term) with quinalphos Menon et al. (2005), Reduced oxidative capability of India the soils as measured by reduced dehydrogenase activity and inhibited iron reduction Endlweber et al. Chlorpyrifos reduced (2005), Germany collembolan density to a greater extent than dimethoate. Both changed the dominance structure of the collembolan community, but had no effects on species composition Edvantoro et al. Bacterial and fungal numbers, (2003), Australia and biomass carbon were (contaminated site) reduced in contaminated soils compared to controls Megharaj et al. (2000), Reduced bacterial and soil algal Australia populations, but may have (contaminated site) increased fungal counts Singh and Singh Significant increase in (2005), India dehydrogenase activity and decrease in alkaline phosphomonoesterase for up to 30 days following treatment Martikainen et al. Short term reduction in (1998), Finland microarthropod numbers (Collembola), but recovery in numbers after time. Community structure remained differentiated. Slight reduction in soil microbial biomass (measured by ATP) Dalby et al. (1995), Single dose on soil had no Australia measured effect on the growth or survival of the earthworms Aporrectodea trapezoides, A. caliginosa, A. longa, or A. rosea Singh and Singh Significant increases in (2005), India dehydrogenase and phosphomonoesterase activities when used as seed dressing, effect lasting up to 60 days Capowiez and Berard Significantly reduced (2006), France burrowing activity in two earthworm species at sub-lethal concentrations (0.5-1 mg/kg), but no avoidance of the insecticide Singh and Singh Significant decreases in both (2005), India dehydrogenase and phosphomonoesterase enzyme activities for up to 90 days Loureiro et al. Collembola avoid soil with (2005), Portugal lindane (10-20 mg/kg) and dimethoate (5-20 mg/kg), while earthworms avoided dimethoate at 2.5 mg/kg Panda and Sahu Short-term impacts of standard (1999), India application rates of malathion on earthworm reproduction lasting for 105 days Table 10. Impact of fungicides on non-target soil organisms Reference and location Soil type and Active characteristics chemical Chen et al. (2001), Silt-loam Luvisol, pH 6.3, Benomyl USA OM 4% Smith et al. (2000), USA Smectitic silt loam soil Benomyl Loureiro et al. (2005), Portugal Martikainen et al. Previously described Benomyl (1998), Finland Previously described Benomyl Hart and Brookes Silty clay loam, pH 6.9, Benomyl, (1996), UK OC 1.36% triadimefon Chen et al. (2001), Previously described Captan USA Hu et al. (1995), USA A well-drained sandy clay Captan loam, OC 0.8% Van Zwieten et al. Ferrosols with kaolinitic Copper (2004), Australia clay minerals. pH (field and OECD 6.3-6.8, OC 4.9-7.1% soil) Merrington et al. Ferrosols with kaolinitic Copper (2002), Australia clay minerals. pH 4.5-6.0, OC 4.88-8.82% Loureiro et al. Previously described Copper, (2005), Portugal carbendazim Gaw et al. (2003, Mainly silty soils, pH Copper 2006), New Zealand 5.3-6.3, OC 1.5-9.4% Belotti (1998), Loamy to clayey soils, Copper Germany pH 5.5-7.2, OM 3.1-7.4% Eijsackers et al. Clayey soils, pH 5.7-6.8, Copper (2005), South Africa OM 0.6-1.3% Chen et al. (2001), Previously described Chlorothalonil USA Kinney et al. (2005), Previously described Mancozeb, USA chlorothalonil Fravel et al. (2005), In vitro studies not using Chlorothalonil, USA soil azoxystrobin Monkiedje et al. Silty clay soil, pH 7.2, OC Metalaxyl, (2002), Germany 1.69% mefenoxam Reference and location Effects Chen et al. (2001), Suppression of respiration, stimulation of USA dehydrogenase, effects were less noticeable with organic matter addition Smith et al. (2000), Significant long-term effects on USA mycorrhizal colonization (80% reduction), reduction in fungal to bacterial ratios and nematode numbers Loureiro et al. Earthworms avoid benomyl at 1 mg/kg soil (2005), Portugal Martikainen et al. Total numbers of enchytraeids and (1998), Finland nematodes, soil respiration and mineral N were not affected, but the collembolan community structure was affected Hart and Brookes Microbial biomass, SIR and mineralization (1996), UK of soil organic N to ammonium and then nitrate mostly unaffected by the pesticide treatments Chen et al. (2001), Suppression of respiration and USA dehydrogenase; increases in ammonium N Hu et al. (1995), USA Fungal length and density, and microbial C and N significantly reduced Van Zwieten et al. Earthworm populations avoid soils with (2004), Australia concentrations as low as 34 mg/kg. Lack (field and OECD of breakdown of organic carbon suggest soil) potential long-term implications Merrington et al. Increased metabolic quotient indicating (2002), Australia microbial stress at 280-340 mg Cu/kg. Significantly reduced microbial biomass and ratio of microbial biomass to OC Loureiro et al. Earthworm avoidance at concentrations (2005), Portugal > 100 mg/kg. Test collembolan species far less sensitive to copper. Carbendazim avoidance by earthworms at 10 mg/kg Gaw et al. (2003, Reduced performance of soil functions 2006), New Zealand resulting in reduction of DDT degradation Belotti (1998), Bioavailable copper concentration of Germany 0.677 mg Cu/kg soil established as the critical concentration for soil impairment (irrespective of OC content) Eijsackers et al. Increasing copper resulted in reduced (2005), South Africa burrowing of earthworms and decreased growth of earthworms, resulting in increased soil bulk densities in a vineyard Chen et al. (2001), Suppression of respiration, stimulation of USA dehydrogenase Kinney et al. (2005), Significant reduction in production of [N. USA sub.2]O and NO following N-based fertiliser application: Significant reduction in nitrification Fravel et al. (2005), Both fungicides toxic to the biocontrol USA agent Fusarium oxysporum strain CS-20 which has been used to reduce incidence of Fusarium wilt Monkiedje et al. Reduced enzyme activity, in particular (2002), Germany dehydrogenase, for up to 90 days. Increased bacterial numbers with increasing doses, but toxic to N fixers at 1 mg/kg (mefenoxam) and 2 mg/kg (metalaxyl) Table 11. Impact of veterinary health products, fumigants, and other biological and non-chemical plant protection measures on non-target soil organisms Reference and Soil type and Active chemical location characteristics Veterinary health products Svendsen et al. OECD standard soil Ivermectin, fenbendazole (2005), Denmark substrate Jensen et al. (2003), Sandy-loamy soil, Tiamulin, olanquindox, Denmark pH 6.2, OC 1.6% metronidazole, ivermectin Radl et al. (2005), Marine sediments, Trenbolone (TBOH) from Germany 78% <0.01 mm, cattle production pH 7.5, redox 21 mV Vaclavik et al. Sandy loam, pH Tylosin, oxytetracycline, (2004), Denmark 6.1, OC 1.6% sulfachloropyridazine Westergaard et al. Sandy soil, pH Tylosin (2001), Denmark 6.8, OC 1.2% Fumigants Massicotte et al. Gravelly sandy Chloropicrin (1998), USA loam, no further descriptions Ingham and Thies Gravelly sandy Chloropicrin (1996), USA loam, no further descriptions Spokas et al. (2005), Sandy soil, pH Methyl isothiocyanate, USA 5.0-6.0, OC chloropicrin 1.11-1.39% Karpouzas et al. Silty sand, pH Metham sodium, methyl (2005), Greece 7.2-7.5, OM bromide 3.74-6.6% Dungan et al. (2003), Sandy loam, pH Propargyl bromide, 1,3 USA 7.2, OM 0.92% dichloropropene Klose and Ajwa Two sandy loam Propargyl bromide, (2004), USA soils, pH InLine, Midas, 7.75-7.82, Chloropicrin OC 0.6-0.7% Biological and non-chemical products Cohen et al. (2005), Clay loam soil, Brassica napus seed meal USA pH 7.4 Cortet et al. (2006), Three soils: silty Bt toxin Denmark and France sand, pH 6.2, OM 6.4%; silty clay, pH 8.1, OM 4.8%; clayey silt, pH 8.2, OM 1.5% Accinelli et al. Two soils: a loam, Bt toxin (2004), Italy pH 7.9, OC 0.92%; sandy loam, pH 8.1, OC 0.7% Gelsomino and Cacco A clay loam soil, Solarisation with (2006), Italy pH 7.2, OC 1.49% transparent polyethylene film Patricio et al. A fertile peat, Solarisation (2006), Brazil pH 5.9, OM around 20% Reference and Effects location Veterinary health products Svendsen et al. It was concluded that earthworm (2005), Denmark populations will not be affected in the field following normal use of these products Jensen et al. (2003), Threshold values for toxicity (10% Denmark reduced reproduction or EC 10 values) of antibacterials tiamulin, olanquindox and metronidazole were in the range of 61-111 mg/kg soil for springtails and 83-722 mg/kg soil for enchytraeids. Ivermectin more toxic with EC 10 values of 0.26 mg/kg soil for springtails and 14 mg/kg soil for enchytraeids Radl et al. (2005), N-acetyl-glucosaminidase activity was Germany almost 50% lower in sediments receiving trace quantities of TBOH. Biolog substrate utilisation was reduced, and this appeared to be permanent Vaclavik et al. Tylosin and sulfachloropyridazine (2004), Denmark significantly impact on gram positive bacteria, while oxytetracycline inhibits general microbial respiration at levels as low as lmg/kg soil Westergaard et al. Long term changes to microbial (2001), Denmark community structure, and short term reduction in total microbial numbers Fumigants Massicotte et al. Chloropicrin did not adversely affect the (1998), USA formation of ectomycorrhizae on young Douglas fir seedlings by naturally occurring fungi for up to 5 years following treatment Ingham and Thies Limited effects on fungal biomass and (1996), USA amoebae were found, but no major impact on the soil food web Spokas et al. (2005), Altered soil biology leading to USA 10-100-fold increases in N20 production lasting >48 days. No effects on methane production. Methyl isothiocyanate suppressed soil respiration in laboratory trials Karpouzas et al. Inhibited degradation of an (2005), Greece organophosphate nematicide applied to soil 9 months following fumigation, suggesting long term impacts Dungan et al. (2003), Significant decline in dehydrogenase USA activity; however, this recovered after 8 weeks in manure amended soil. Bacterial community diversity decreased with increasing fumigant concentration Klose and Ajwa Organic matter turnover and nutrient (2004), USA cycling, and thus, the long-term productivity largely unaffected in soils repeatedly fumigated with these products, except for a reduction in several key enzymes activities Biological and non-chemical products Cohen et al. (2005), Reduced Rhizoctonia root infection and USA Pratylenchus spp nematodes. Suggested mechanism was altered bacterial community supporting nitrous oxide production and induction of plant systemic resistance Cortet et al. (2006), No effect on litter bag decomposition or N Denmark and France mineralisation Accinelli et al. In laboratory studies, the presence of the (2004), Italy Bt toxin inhibited the breakdown of glufosinate-ammonium and glyphosate Gelsomino and Cacco Decreased bacterial diversity as measured (2006), Italy by DGGE fingerprinting over time, although short term increases in diversity were noted Patricio et al. Reductions in microbial C, and fungal and (2006), Brazil bacterial numbers, but no effect on fluorescent pseudomonas
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|Author:||Bunemann, E.K.; Schwenke, G.D.; Van Zwieten, L.|
|Publication:||Australian Journal of Soil Research|
|Date:||Nov 1, 2006|
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