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ARTICLE IN PRESS Advances in Environmental Research xx (2002) xxx–xxx 1093-0191/03/$ - see front matter ᮊ 2002 Elsevier Science Ltd. All rights reserved. doi:10.1016/S1093-0191(02)00135-1 Bioavailability and toxicity of cadmium to microorganisms and their activities in soil: a review K. Vig, M. Megharaj* , N. Sethunathan, R. Naidu ,11 CSIRO Land and Water, PMB 2, Glen Osmond, Adelaide, SA 5064, Australia Received 28 January 2002; received in revised form 1 November 2002; accepted 13 November 2002 Abstract Significant quantities of cadmium (Cd) have been added to the soils globally due to various anthropogenic activities, raising concerns for environmental health. Microorganisms play a unique role in the soil ecosystem, because of their contributions to soil fertility. Contrasting trends, reported on the toxic effects of heavy metals including Cd on soil microorganisms and their activities, are attributable to short-term studies often limited to a single soil type and conducted under controlled laboratory conditions. There is a paucity of reliable field data on Cd alone, since most field studies on Cd-microorganism interactions in soils are based on sewage sludge containing multimetals and organic substances. No single parameter can be used to generalize Cd toxicity and different parameters can provide contrasting results. A battery of relevant tests, rather than just one single assay, involving important microbial activities should therefore be included in ecotoxicity studies. The bioavailability of Cd and associated toxicity to soil biota vary with time, soil type, speciation, ageing, Cd-source, organisms and the environmental factors. The available fraction or soil solution Cd, and not the total concentration of Cd, seems to correlate well with the toxicity parameters. ᮊ 2002 Elsevier Science Ltd. All rights reserved. Keywords: Cadmium; Soil; Microorganisms; Bioavailability 1. Introduction Cadmium (Cd), a potentially toxic heavy metal with no known biological function, occurs widely in nature in small amounts, with an average content of 0.2 mg kg in the geosphere (Lindsay, 1979) . The concen- y1 tration of Cd in rocks range from 0.1 to 11 mg kg . y1 Naturally occurring Cd ores usually occur in association with Zn ore sphalerite and are recovered as byproducts of Zn mining. Anthropogenic activities such as industrial waste disposals, fertiliser application and sewage sludge disposals on land have also led to accumulation of Cd *Corresponding author. Tel.: q61-8-8302-5044; fax: q61- 8-8302-3057. E-mail address: megharaj.mallavarapu@unisa.edu.au (M. Megharaj). Present address: Centre for Environmental Assessment and 1 Remediation, University of South Australia, Mawson Lakes, SA 5095, Australia. in soil and its leaching under certain soil and environ- mental conditions (Alloway, 1990; Naidu et al., 1997), with eventual increase in its concentration in food crops. The FAOyWHO recommended maximum tolerable intake of Cd is 70 mg day . With the estimated half- y1 life of Cd in soil varying between 15 and 1100 years (Kabata Pendias and Pendias, 1992), its accumulation in the environment and its entry into the food chain are of great concern. Soil solution Cd is considered as the bioavailable form from the standpoint of its ecotoxicity. The concen- tration of Cd in soil solution varies significantly with soil properties and nature of management practices imposed on the system by farmers and other land users. In soils with no anthropogenic inputs of Cd, soil solution concentrations could range from 0.3 to 22.5 mgl y1 (Helmke, 1999), depending on the geological source of the soil. In agricultural soils, Cd concentrations rarely exceed 10 mg l in the soil solutions (Kookana et al., y1 ARTICLE IN PRESS 2 K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx 1999). The concentration of total dissolved Cd increases with ionic strength and varies with the concentration of dissolved organic carbon (Fotovat and Naidu, 1998). More recent speciation studies by Krishnamurti and Naidu (2000) demonstrated that over 60% of soil solution Cd may be present in association with dissolved organic matter. Cadmium, in association with chlorides, hydroxyl, sulfhydryls and thiol groups, forms soluble complexes and these complexes largely govern the biological activ- ity of Cd. Given that soil microbiological activity has a great potential as an early and sensitive indicator of stress in soil (Brookes, 1995), our aim here is to (i) compile the global information on Cd toxicity to soil microbes and their activities; (ii) compare its toxicity between field and laboratory studies; (iii) determine the relationship between Cd bioavailability and toxicity; and (iv) deter- mine the factors that control Cd bioavailability and toxicity to soil microorganisms. 2. Soil microorganisms and their significance Soil serves as a habitat for diverse groups of micro- organisms, comprised of algae, bacteria, fungi and other organisms. The major groups and their predominant functions related to soil health and fertility are highlight- ed below. 2.1. Algae and cyanobacteria Algae are the primary producers and form the base of the food chain. Algae account for 4–27% of the total microbial biomass in the soil (McCann and Cullimore, 1979) and are involved in maintaining soil fertility and oxygen production (Bold and Wynne, 1978). Minerali- zation of algae releases nutrients which are utilized by other organisms in the soil for their growth and devel- opment. Cyanobacteria are more widespread than other free-living microorganisms (Burns and Hardy, 1975) and contribute to the nitrogen economy of agricultural soils, because of their ability to fix atmospheric nitrogen. Since microalgae resemble higher plants in terms of their intracellular organisation and cyanobacteria are prokaryotes, bioassays with these groups can serve as valuable indicators of toxic effects of pollutants to other fundamentally similar cells (Megharaj et al., 2000a).In spite of their importance, algae in the soil ecosystem are the neglected group among soil microbiota (Megh- araj et al., 2000a). However, the role of algae in remediation of industrial wastewater and other contam- inated resources through sorption and accumulation has been gaining importance in recent years. 2.2. Bacteria and other microbial populations in soil Bacteria and fungi are probably the most widely studied groups of soil microorganisms. In soil, fungi, although numerically much less abundant than bacteria, can account for twice the weight of bacteria and actinomycetes combined (Jenkinson and Ladd, 1981). Soil fungi can occur free-living or in association with plant roots. The best-known function of fungi is decom- position of complex compounds of plant and animal origins, such as cellulose, lignin, and chitin. 3. Effect of Cd on microorganisms and their activities Although toxic effects of heavy metals on soil micro- biota are well-recognized (Baath, 1989; Giller et al., 1998), contrasting trends have been reported in the literature on the effect of Cd on soil microbiota. These are largely because of the differences in soil types and source of metal contamination (e.g. sewage sludge, soluble metal salts) which would have profound effect on Cd chemistry in soil and its impact on soil microbiota. Table 1 summarizes some information on the effects of Cd on soil microorganisms. In a recent study lasting over 180 days in our laboratory, we found that the concentration of soil solution Cd in freshly contaminated soil ( 3 mg Cd kg soil) decreased exponentially with y1 time to -0.0006 mg l within 50 days of ageing. If y1 the data, presented in Fig. 1, reflect changes in Cd bioavailability under field conditions, then most of the short-term laboratory based studies have presumably overestimated the effects of Cd on soil microbiota when extrapolating their findings to metal contaminated field soils. Usually, any pollutant that permitted a full recov- ery of a microbial parameter within 30 days of exposure is not considered a risk while any negative impact beyond 60 days is considered a significant risk (Domsch et al., 1983). 3.1. Microbial population Estimates of microbial population in soil by conven- tional plate count method have shown that bacteria are more sensitive to Cd than fungi. Based on laboratory incubation experiments, the effect of Cd on the bacterial population depended on its concentration used and the soil type. Even at environmentally unrealistic concentra- tions Cd was not toxic at all (1000 mg kg , Fritze et y1 al., 2000) or toxic only for 4 weeks (5000 mg kg , y1 Kozdroj, 1995) to the population of heterotrophic bac- teria in soils. But in several other studies Cd was inhibitory to the bacterial population when spiked at much lower levels (Zibilske and Wagner, 1982; Dar and Mishra, 1994; Dar, 1996). Fungal numbers increased with increasing Cd concentration in both sewage sludge- ARTICLE IN PRESS 3K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx Table 1 Effect of Cd on microbial populations and their activities (selected reports) Soil typey treatment Cd (mg kg soil) y1 % inhibition(y)y References stimulation(q) Field studies Sandy loam soil (received sewage sludge 20 years ago), pH 6.8, 1.9% C, 9% clay Cd 8.6, Ni 27, Cu 102, Zn 289 y, N-fix 25% Brookes et al. (1986a) Sandy loam soil (received sewage sludge 20 years ago), pH 6.8, 1.9% C, 9% clay Cd 4.7, Ni 6.7, Cu 56.5, Zn 139.2 y, biomass 42–60% Brookes et al. (1986b) Clay-loam, pH 6.5, 26.3% clay Cd 60, Pb 1088, Zn 3049 y, cellulose decomp 7% Khan and Frankland (1984) Oak forest near abandoned zinc smelter, pH 5.0–6.2, 0.5–0.7% C Cd 26, Cu 15.0, Pb 21.6 and Zn 478 y, bacteria 86%, fungi 60%, actinomy- cetes 86%, nitrosomonas 94%, nitrobacter 40%, DHA 93%, urease 88% Pancholy et al. (1975) Lab amendments pH 5.1–6.1, 1.5–2.9% OC, 10–21% clay CdCl 562 2 y, arylamidase 55–82% Acosta-Martinez and Tabatabai (2001) pH 6.2–7.6, 2.7–5.3% OC, 26–34% clay 2810 y, arylsulfatase 23–55% AL-Khafaji and Tabatabai (1979) pH 7.6, 3.2% OC, 30% clay 281 y, arylsulfatase 7% pH 4.8, 5.8% OM 1000 NE, ammon Bewley and Stotzky (1983) 500 y, nitr Silt loam, pH 6.75, 1.8% C, 28% clay Cd (NO ) 10 32 y, denitr Bollag and Barabasz (1979) Silt loam, (1.31% OC)q1% dry sludgeq 1% alfalfa CdCl 8.7 2 y, CO 10% 2 Chang and Broadbent (1981) Sandy loam, pH 7.9, 0.47% OC; Loam, pH 8.1, 1.6% OC; Clay-loam, pH 7.7, 0.7% OC CdCl 50 2 y, CO 21–30%, biomass 17–25%, N- 2 min 42–53% Dar and Mishra (1994) Sandy loam, pH 7.9, 0.47% OC; Loam, pH 8.1, 1.61% OC; Clay-loam, pH 7.7, 0.72% OC CdCl 50 2 y, bacteria, biomass, DHA, alkaline phos- phatase, arg ammon Dar (1996) Sandy, pH 7.0, 1.6% OM, 2% clay CdCl 150 2 y,CO 9% 2 Doelman and Haanstra (1984) Sandy peat, pH 4.4, 12.8% OM 400 y, CO 10% 2 Sandy, pH 7.0, 1.6% OM CdCl 150 2 y, urease 10% 6 weeks Doelman and Haanstra (1986) Sandy peat, pH 4.4, 12.8% OM 1980 y, urease 10% 6 weeks 40 y, urease 10% 1.5 years Calcareous soil pH 7.4, 5.4% OC, 4.3% Ca equiv, 23.5% clay; CdCl 10 2 q, nitrate production 15–16% in both the soils Dusek (1995) Non-calcareous soil pH 7.6, 2.6% OC, 1.2% Ca equiv, 18.6% clay 100 and 500 y, nitrate production 13–37% Silty, pH 5.6, 2.6% OC, 28% clay 562 y, amidase 6% Frankenberger and Tabatabai, (1981) Loamy, pH 7.0, 3.2% OC, 30% clay Podzolized sandy soil with 3 cm thick humus, pH 3.8, 50.3% C CdCl 200–4000 2 y, CO 15–58% 2 Fritze et al. (1995) Forest humus, pH 3.95, 52% C CdO or CdCl 2 y, CO 16–24% 2 Fritze et al. (2000) 1000qPumice NE, biomass Sandy loam soil, pH 5.4, 1.65% OC, 16% clay; 10 y, C-min Gupta et al. (1984) Sandy loam soil, pH 5.4, 1.4% OC, 16% clay 30.7 NE, N-min, catatlase Agricultural sandy loam, pH 7.8, 4% OM CdSO 1800 4 y, biomass 78%, DNA 36% Griffiths et al. (1997) ARTICLE IN PRESS 4 K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx Table 1 (Continued) Soil typey treatment Cd (mg kg soil) y1 % inhibition(y)y References stimulation(q) Sandy, pH 7.0, 1.6% OM, 2% clay; CdCl 55 2 q, glutamic acid decomp Haanstra and Doelman (1984) Silty loam, pH 7.7, 2.4% OM, 19% clay; 55 q, glutamic acid decomp Clay, pH 7.5, 3.2% OM, 60% clay; 55 NE, glutamic acid decomp Sandy peat, pH 4.4, 12.8% OM, 5% clay 1000 NE, glutamic acid decomp Clayey loam, pH 8.52, 0.87% C CdSO 2248 4 q, N-min 28% Hassen et al. (1998) 5 different soils: River sand, Gley, Gray lowland, Andosol and Humic Andosol (pH 5.7–8.5) CdCl 3372 2 y, CO 37–65% 2 Hattori (1989) 5 different soils—River sand, Gley, Gray lowland, Andosol and Humic Andosol (pH 5.7–8.5)q2% sludge CdCl 1124–3372 2 y, CO , N-min, bacteria, actinomycetes, 2 proteinase Hattori (1989) q, fungal, b-glucosidase Gley soil, pH 5.8, 96.4% sand, 1.3% silt, 2.2% clayq1% glucoseycellulose CdCl 11.24–1124 2 y,CO 2 Hattori (1991) q, fungi, ATP NE, bacteria Gley soil, pH 5.8, 96.4% sand, 1.3% silt, 2.2% clayq2% sewage sludge; CdCl 1124 2 y, bacteria, Gley Hattori (1992) y, CO both the soils 2 q, bacteria, Andosol Andosol, pH 6.4, 36.3% sand, 36.1% silt, 27.6% clayq2% sewage sludge q, fungus both the soils Brown earth-loamy sand, pH 4.6, 12.3% clay; CdCl 10–100 2 y, cellulose decomp 13.5–35% Khan and Frankland (1984) Brown earth-loamy sand, pH 5.4, 10.3% clay y, cellulose decomp, root biomass Sandy loam from long-term liming experi- ment, pH 4.5–7.0 CdSO 3.1–4.3 4 NE, biomass Knight et al. (1997) Red soil, pH 4.51, 1.74% OM, 1.08% TOC CdCl 5–100 2 y, biomass 14–91% Khan et al. (1997) Cd(CH COO) 5–100 32 y, biomass 6–14% Sandy loam, pH 6.5, 12.6% OC, 8% clay, 12% silt, 80% sand CdCl 5000 2 y, heterotrophic bacteria initially fol- lowed by similar levels after 4 weeks Kozdroj (1995) q, arg ammon Forest soil, pH 4.8, 2.3% OC, 87% sand, 8% silt, 5% clay CdSO 500 4 y, CO , DHA, ATP, acid phosphatase 2 Landi et al. (2000) NE, biomass 50 y, acid phosphatase Forest litter from forest floor, pH 3.4–4.3 CdCl 250 2 y, CO 14% 2 Laskowski et al. (1994) CdyCa or CdyMg or CdyK 50y500 mgg y1 y, CO 8–17% 2 Soil litter CdSO 562 4 y, N-fix 36% Lighthart (1980) Cdqcitrate 562 y, N-fix 19% Agricultural, pH 5.8–7.8, 2.6–5.5% C, 23–34% clay CdSO 562 4 y, nitr 67–94% Liang and Tabatabai (1978) Clay; Cd(NO ) 61.7 32 y,NO toNO 2y McKenney and Vriesacker (1985) ARTICLE IN PRESS 5K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx Table 1 (Continued) Soil typey treatment Cd (mg kg soil) y1 % inhibition(y)y References stimulation(q) Sandy 54.1 Montepaldi soil, pH 8.1, 66% sand, 21% silt, 13% clay, 1.7% TOC; CdSO 3–4000 4 y, ATP, DHA, urease Moreno et al. (2001) Castelporziano soil, pH 4.8, 87.5% sand, 8% silt, 4.6% clay, 2.3% TOC Montepaldi )Castelporziano Forest litter MOR—brown podzolized, pH 4.5–5.2, 58% OM; CdCl 400 2 y,CO 2 Niklinska et al. (1998) Forest litter MULL—rendzina pH 3.8–4.5, 70% OM 25 y,CO 2 Phaeosem, pH 6.9, 2.2% OM; Cd (CH COO) 1.7–228.8 32 y,CO 2 Reber (1989) Sandy hortisol, pH 7.0, 2.6% OM; Acidic cambisol, pH 5.6, 1.6% OM Agricultural soil, 1.3% OC Cd(NO ) 150 32 y, DHA 48% Rogers and Li (1985) Fir needle litter, 78% OM CdCl 1000 2 y, CO 24%, cellulase 29%, amylase 2 34% Spalding (1979) NE, invertase, xylanase, b-glucosidase, polyphenoloxidase pH (CaCl ) 4.4–6.6, 0.9–2.8% C, 1–17% 2 clay; CdCl 200 2 y, potential nit rate 50–80% Smolders et al. (2001) pH (CaCl ) 6.6, 1.1% C, 10% clay 2 2 y, potential nit rate 14% pH 4.6–7.0, 1.99–5.32% C, 24–36% clay CdSO 2810 4 y, pyrophosphatase 19–50% Stott et al. (1985) Surface soils, pH 5.1–7.8, 2.6–5.5% OC, 17–42% clay CdSO 562 4 y, urease 49–67% Tabatabai (1977) Agricultural, pH 5.1 29 y, CO 36% 2 Walter and Stadelman (1979) 58 q, ammon Sandy luvisol, pH (CaCl ) 6.0, 0.9% OC, 2 7% clay CdCl (single or successive dose) 100 2 y, CO , DHA 2 Wilke (1991) Silt loam (Alfisol), pH 6.0, 2.1% OM CdCl (spiked in sewage sludge) 0.1 2 y, ATP 50% Zibilske and Wagner (1982) 1.0 altered fungal community 111 y, bacteria 80% NE, no effect; OC, organic carbon; OM, organic matter; N-fix, nitrogen fixation; arg ammon, arginine ammonification; nit, nitrification; decomp, decomposition; DHA, dehydrogenase. ARTICLE IN PRESS 6 K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx Fig. 1. Effect of ageing on soil solution Cd in a Xeralf from South Australia (Naidu, unpublished). amended (Hattori, 1989) and unamended (Hattori, 1991) soils, spiked with Cd (0–1124 mg kg ).Itwas y1 the amount of water-soluble Cd, and not the total Cd applied that determined its impact on the microorgan- isms (Hattori, 1992). Most of the published work on metal toxicity to algae is based on studies conducted with aquatic algae, while information on terrestrial algae is limited. Cyanobacter- ial colonization and autotrophic nitrogen fixation have been shown to decrease in a metal contaminated soil (EDTA extractable soil metal concentrations of Cd, 3; Ni, 2.5; Cu, 20; Zn, 50 mg kg ) even 20 years after y1 the last metal inputs through sewage sludge were made (Brookes et al., 1986a). But, the effect of Cd, applied alone, on soil algal populations is little understood. 3.2. Microbial biomass Microbial biomass serves as a pool of nutrients and is a sensitive indicator of microbial changes in soil. Generally, microbial biomass in soils, measured by the fumigation extraction method, is not adversely affected by Cd even at abnormally high concentrations of 500– 1000 mg Cd kg (Fritze et al., 2000; Landi et al., y1 2000). But, Cd, applied at 500 mg kg , albeit innocu- y1 ous to microbial biomass as estimated by the fumigation extraction method, effected a decrease in ATP content (Landi et al., 2000). Evidently, microbial biomass may not always be a reliable indicator of metal stress (Megharaj et al., 2000b). A distinct decrease in the ATP content was caused by Cd (Moreno et al., 2001) whereas Hattori (1991) reported an increase in ATP content following increased Cd addition to soil. The reason for such contrasting effects is not clear. The metabolic quotient (qCO ), the ratio between 2 respiration and the microbial biomass, has also been used as an ecophysiological indicator of heavy metal stress in soil (Anderson and Domsch, 1990; Brookes, 1995). qCO decreased in the heavily Cd contaminated 2 soil over that in control during the first 2-d incubation (Landi et al., 2000). Reports on qCO values in metal 2 contaminated soils have been contradictory, with higher values in contaminated soils than that in uncontaminated soils in some reports (Chander and Brookes, 1991; Moreno et al., 1999) and vice versa in other reports (Baath et al., 1991; Insam et al., 1996; Landi et al., 2000). It is appropriate to use qCO as an indicator for 2 assessing metal effects in not very dissimilar soils (Insam et al., 1996). In a recent study (Renella et al., 2002), Cd, applied singly and in combination, at 12 mg kg (4 times the y1 EU limit), was the least toxic among the three metals Cu, Cd and Zn. But, the concentrations of metals used in this study were different (Cd, 12 mg kg soil; Cu, y1 140 mg kg ; Zn, 300 mg kg soil), making the y1 y1 comparison of the acute effects of these metals in this short-incubation study difficult. 3.3. Microbial community structure Gross measurements of microbial diversity have been used to assess environmental stress (Atlas, 1984), but such studies are hampered by problems of sampling, extraction and culturing leading to bias towards certain groups within microbial communities. Pollution may lead to a decrease in microbial diversity due to the extinction of species which lack sufficient tolerance to the stress imposed, and enhanced population of other species which thrive under stress (Atlas, 1984). The modification of the metabolic behaviour of the whole microbial community should reflect a shift in its quantitative and qualitative composition. Biolog soil community fingerprints, based on inoculation of Biolog plates with soil suspensions, have been used to deter- mine the changes in soil biochemical properties insti- gated by heavy metals (Knight et al., 1997; Fritze et al., 2000). However, the reproducibility of the results obtained is governed by the inoculum density and microscale heterogeneity, which may make Biolog assay unsuitable for comparing environmental samples from different origins. Another approach to detect the changes in the microbial community is to determine the phos- pholipid fatty acid (PLFA) composition in the soil. This method is based on the extraction and identification of signature lipid biomarkers from the cell membranes of microorganisms. Its quantitative analysis can provide taxonomic information at the species and subspecies level and serves as a measure of the viable biomass. Evidence from field studies suggests that under long- term metal stress there is a change in the genetic structure of the soil microbial community, without nec- essarily any increase in the metal tolerance (Giller et al., 1998). Fritze et al. (2000) reported an increase in ARTICLE IN PRESS 7K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx the actinomycete PLFA signatures and a decrease in the fungal PLFA signatures in laboratory-incubated forest soils amended with 400 and 1000 mg Cd kg (as y1 CdCl or CdO). Griffiths et al. (1997) also reported a 2 shift in microbial communities in an agricultural sandy loam soil, spiked with CdSO at 1800 mg kg soil, by y1 4 %GqC profiles, DNA hybridization and PLFA patterns. Generally, actinomycetes are more resistant to Cd than Gram negative and Gram positive bacteria in that order (Doelman, 1986). There is a need for more research in species diversity in Cd polluted soils. 3.4. Microbial activities Microbial activities may reflect the functions of total (respiration) or specific ( nitrification) groups of micro- organisms in the soils (Domsch et al., 1983). The most commonly studied microbial reactions impacted by inor- ganic or organic pollutants include: C-mineralization, N-mineralization, CO production and enzyme activities. 2 3.4.1. Mineralization of carbon Mineralization of organic carbon to CO commonly 2 known as ‘soil respiration’ is a good index of total activity of microflora involved in organic matter decom- position (Anderson, 1982). Therefore, soil respiration has been the most studied parameter on the effects of metals on microbial activities in soil (Baath, 1989). Total soil respiration reflects the metabolic behaviour of a range of microorganisms, which are not equally sensitive to pollutants. Tests based on this parameter are therefore not always valid (Vallaeys et al., 1997). Usually, there is no clear trend between metal contam- ination and respiration in agricultural soils (Giller et al., 1998), whereas negative effects have been reported in forest soils (Baath, 1989). Heavy metals may reduce the substrate availability for soil respiration by forming complexes with the substrates or by killing the micro- organisms (Landi et al., 2000). Even lower levels (10 mg Cd kg soil) of spiked Cd can decrease CO y1 2 evolution from an acid sandy loam soil (Gupta et al., 1984). There are also reports (Niklinska et al., 1998; Landi et al., 2000) of inhibition of carbon mineralization by Cd but only at unrealistically high levels (400–500 mg Cd kg soil). The impact of pollutants on soil y1 respiration differs with soil type. The level of Cd required to produce a 9% decrease in CO evolution 2 was 150 mg kg in a sandy soil (pH 7.0, organic y1 matter 1.6%) and 400 mg kg in a sandy peat soil y1 (pH 4.4, organic matter 12.8%)(Doelman and Haanstra, 1984). The inhibitory effect of Cd on microbial respi- ration in forest litter soil was noticed only when amend- ed with nutrients such as Ca, Mg or K (Laskowski et al., 1994), probably due to its increased solubility (Schierl et al., 1986). In another study on the effects of different forms of Cd on soil respiration, Fritze et al. (2000) observed a similar inhibitory effect on respira- tion in forest humus soil spiked with 1000 mg Cd kg as CdO or CdCl . Evidently, in a sandy loam y1 2 soil (planted to wheat) spiked with 100 mg Cd kg , y1 Cd, added in insoluble forms as carbonate and oxide, was less toxic to cellulose decomposition than Cd added as sulfate or chloride (Khan and Frankland, 1984). Hattori (1989) reported a negative correlation between the amount of C mineralized and the concentration of water-soluble Cd in Cd- and sludge-amended soils. Thus, the toxicity of Cd varies with the soil type and the nature of Cd salt. 3.4.2. Nitrogen mineralization The impact of metals on nitrogen mineralization rates has been studied extensively. Nitrification in the field appears to be a sensitive indicator of metal pollution (Baath, 1989). In laboratory incubation studies, Cd was generally inhibitory to nitrification in soils (Necker and Kunze, 1986; Hattori, 1989). However, the inhibitory effect varied with the soil and test conditions. Liang and Tabatabai (1978) reported an inhibition of nitrifi- cation at 562 mg Cd kg in soils. Likewise, nitrifica- y1 tion, but not ammonification, was inhibited by 500 mg Cd kg and above (Bewley and Stotzky, 1983). Cad- y1 mium, applied at 50 mg kg , adversely affected nitro- y1 gen mineralization to a greater extent in a sandy loam soil than that in a clay-loam soil (Dar and Mishra, 1994). Interestingly, there was a significant decrease (14%) in nitrification in a soil (pH 6.6) spiked with 2 mg Cd kg (Smolders et al., 2001). Also, there was y1 evidence that Cd promoted nitrification at 10 mg kg y1 and inhibited it at 100 mg kg in both calcareous and y1 non-calcareous soils (Dusek, 1995). Nitrogen minerali- zation was stimulated under field conditions by Cd as opposed to inhibition under the laboratory incubation (Necker and Kunze, 1986) , however, the concentration used in field was less than that in the laboratory. The stimulation under long-term impact may be due to the slow development of an adapted microflora tolerant to Cd. Cadmium has also been shown to be inhibitory to denitrification in a silt loam soil at 10 mg kg (Bollag y1 and Barabasz, 1979), and in clay and sandy loam soils, but only at )50 mg kg soil (McKenney and Vrie- y1 sacker, 1985), probably due to its decreased bioavailability. 3.4.3. Enzymes Soil enzyme activity is used as a sensitive indicator of the effect of pollutants, including metals in soils (Giller et al., 1998). In some of the soil enzyme assays, abiotic activity, however, originates from dead cells, which cannot readily be distinguished from the enzy- matic activity of the living cells (Burns, 1982). ARTICLE IN PRESS 8 K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx Heavy metals can reduce enzyme activity by inter- acting with the enzyme–substrate complex, denaturing the enzyme protein, interacting with its active sites (Nannipieri, 1994) or by affecting the synthesis of the enzymes within the microbial cells. Metal-induced changes in the community structure can also modify the enzyme activity (Nannipieri, 1994). When Cd binds to active sites, enzymes such as alkaline phosphatase are inactivated and metabolism is disrupted (McGrath, 1999). Cadmium, besides being an enzyme inhibitor, can have deleterious effects on membrane structure and function by binding to the ligands such as phosphate and the cysteinyl and histidyl groups of proteins (Collins and Stotzky, 1989). The effect of Cd on the soil enzymes varies with the enzyme studied and the soil type as shown in Table 1. Dar (1996) reported a decrease in dehydrogenase activity (DHA) and alkaline phosphatase activity at 50 mg Cd kg in a laboratory study with different soil y1 types. Landi et al. (2000) found a negative effect of Cd on DHA, but at a much higher concentration of 500 mg Cd kg whereas acid phosphatase activity decreased at y1 50 mg Cd kg . The concentration of Cd (mg kg ) y1 y1 required to cause a 10% decrease in urease activity after 6 week incubation amounted to 150 in sandy loam soil, 360 in silt loam, 950 in clay soil and 1980 in sandy peat soils (Doelman and Haanstra, 1986). The Cd concentration required to cause a similar inhibitory effect on phosphatase activity, was 10 times greater in clay soil than in sandy soil (Doelman and Haanstra, 1989). Rogers and Li (1985) found that Cd was 4 times more toxic to DHA in unamended soil than in soil amended with 1% alfalfa. Alfalfa amendment apparently increased the microbial population resistant to Cd or decreased its availability. The toxicity of Cd to urease and dehydrogenase activities decreased in soil with low pH (4.8) and high organic carbon (2.3%)(Moreno et al., 2001). 3.4.4. Biodegradation of organic contaminants Bioaugmentation with focus on using metal-resistant microorganisms for promoting the degradation of organ- ic contaminants by microorganisms is an emerging area of research in the remediation of cocontaminated sites. Toxicity of metals such as Cd to soil microorganisms adversely affected the biodegradation of organic con- taminants in cocontaminated soil (Said and Lewis, 1991). But, in a dual bioaugmentation strategy by coinoculation, four Cd-resistant bacterial isolates were used to reduce the soluble concentrations of Cd and then allow the degradation of 2,4-dichlorophenoxyacetic acid (2,4-D) by a Cd-sensitive 2,4-D-degrading bacte- rium, Ralstonia eutropha JMP134 in a cocontaminated soil (Roane et al., 2001). None of the four metal- resistant bacteria could degrade 2,4-D, but each of these isolates supported the degradation of 500-mg 2,4-D ml by R. eutropha JMP134. Cadmium detoxification y1 by the metal-resistant bacteria involved a plasmid- independent intracellular mechanism. 4. Field studies As mentioned earlier, data related to field studies with Cd alone are negligible. Field studies with Cd come mainly from industrially contaminated sites or from agricultural fields amended with sludge containing multimetals or exposed to extensive fertilizer applica- tions. Baath (1989) and Giller et al. (1998) have extensively reviewed these aspects of metal toxicity in forest and agricultural soils. As in laboratory amend- ments, toxicity observed in the field also varies with the soil type as shown in Table 1. 5. Possible reasons for variability in Cd toxicity data Reports on the ecotoxicity of Cd, like other metals, on biodiversity and microbial activities have seldom been consistent for many reasons. Some of the discrep- ancies in the literature on ecotoxicity of Cd in soils may be related to bioavailability, sensitivity of micro- organisms, and the methodologies used. 5.1. Bioavailability of Cd A major factor governing the toxicity of a metal in soil is its bioavailability. Bioavailability is considered as the fraction of the total contaminant in the interstitial water and soil particles that is available to the receptor organism. This suggests that there is a continuum between 0 bioavailability and 100% bioavailability and within this spectrum the pool of contaminant available to receptor organisms may vary depending on the nature of organisms and the perturbations imposed by the environment (Naidu et al., 2000). There has been limited work on bioavailability of Cd and its relationship to ecotoxicity in soils spiked with Cd. Bioavailability of Cd is largely governed by (i) soil type, (ii) Cd speciation, (iii) ageing, (iv) nature of Cd applied and (v) nature of microorganisms. 5.1.1. Soil type Physico-chemical and biological properties of the soil play major roles in the availability of the metal to the organisms. Change in pH affects the sorption of Cd by soils and thereby its concentration in soil solution (Naidu et al., 1997). Increases in soil pH decrease Cd in solution or make it less available. Organic matter and clay content of soil can also significantly influence the concentration of soil solution Cd. The effect of soil contamination with Cd on microbial processes (phos- phatase activity (Doelman and Haanstra, 1989)), urease ARTICLE IN PRESS 9K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx activity (Doelman and Haanstra, 1986), as described earlier under Section 3.4.3) is generally not of great magnitude especially in clay soil, probably due to its low bioavailabilty (Khan et al., 1997). Application of Cd (55–400 mg kg ) delayed the decomposition of y1 glutamic acid in both sandy and silty loam soils, but not in clay and sandy peat soils (Haanstra and Doelman, 1984). The mineralization of carbon and nitrogen in different soils were related to the amount of water- soluble Cd in soil (Hattori, 1989, 1992). Evidently, it is not the amount of metal spiked in soil that causes toxicity, but the amount ‘available’. Complexation of Cd with organics andyor the formation of insoluble chelates was a possible reason for its low bioavailability in clay-loam and loam soils (Lighthart et al., 1983; Dar, 1996). For the microorganisms, the free divalent ion, considered as the toxic species of Cd, in soil solution is the readily bioavailable fraction. Clay is the dominant abiotic component in soils that decreases the toxicity of Cd, followed by organic matter content (Kuo and Baker, 1980; Doelman and Haanstra, 1984). Babich and Stotz- ky (1977) also reported decreased toxicity of Cd to several fungi in soil with an increasing content of montmorillonite. Presence of different mineral nutrients, cations or anions in the soil can also alter the toxicity of Cd by complexation, sorption or desorption processes (Naidu et al., 1994, 1997; Bolan et al., 1999). Because soil is such a complex system, it is difficult to make broad generalizations on the effect of ligands in solution on the sorption of Cd (Harter and Naidu, 1995) and hence on the availability of Cd. However, various studies suggest that bioavailable fraction of Cd in soil decreases with time and with increase in pH, clay and organic matter contents in the soil. 5.1.2. Speciation The total metal content of a soil is distributed among all the possible chemical forms (speciation) in the solid, liquid or the biotic phases of the soil (Krishnamurti, 2000). Speciation of the metal ion in the soil solution may play a significant role in its bioavailability. Only few studies have tried to identify the particular species of the Cd, which may contribute to its bioavail- ability. The availability of particulate-bound Cd in a soil–plant system decreased in the order: exchangeable -carbonates -metal–organic complexes -organics - Fe and Mn oxides -mineral lattices (Krishnamurti et al., 1995). Comparison of the toxicity of pore water in two soils spiked with Cd(NO ) to a soil alga, Chloro- 32 coccum sp. clearly showed that organically complexed Cd is bioavailable and contributed to the toxicity to the alga (Krishnamurti, unpublished data). This contradicts the long-held notion that Cd–organic complexes are not bioavailable to soil biota. There is a need for more research on speciation in relation to toxicity to micro- organisms and their activities. 5.1.3. Ageing The bioavailability of a heavy metal decreases with the duration of its contact with soil (Naidu et al., 2003) due to its decreased desorbability over time, as illustrat- ed with soil solution Cd in Fig. 1. The concentration of soil solution Cd in a freshly contaminated soil decreased exponentially from 3 mg kg soil at the start to y1 negligible levels within 50 days of ageing (Naidu, unpublished data). A decrease in solution Cd can lead to its decreased toxicity to microorganisms (Hattori, 1989, 1992; Dar and Mishra, 1994; Dar, 1996). In a recent laboratory incubation study (Vig, unpublished data), toxicity of Cd to nitrate reductase decreased with time in a sandy loam soil at 100 mg Cd kg . In contrast, Doelman and y1 Haanstra (1986) reported an increase in toxicity of Cd to urease activity with time (1.5 years) in a laboratory study with 5 soils spiked with 55–8000 mg Cd kg y1 soil, probably due to increased mobility of Cd by chelation (Doelman and Haanstra, 1984). 5.1.4. Nature of Cd applied Usually, studies on Cd toxicity have been conducted using Cd salts such as CdCl , CdSO and 24 Cd(CH COO) . The nature of the anions can modify 32 the toxicity of Cd, as, for instance, the nitrate seems to counteract the negative effect of Cd on soil respiration (Saviozzi et al., 1997). Soil respiration was inhibited by CdCl and CdSO , and not by Cd(NO )(Rother et 24 32 al., 1982). Acetates of Cd, Pb and Zn were less toxic than the respective sulfates to ammonification and nitri- fication. Cadmium acetate (14–91%) effected a more pronounced decrease in microbial biomass carbon than CdCl (6–14%) when spiked at 0–100 mg Cd kg y1 2 (Khan et al., 1997). The low toxicity of CdCl was 2 probably due to the fixation of the major part of the Cd in the soil. But, acetate in Cd(CH COO) acted as a 32 metal-complexing ligand, thereby reducing the adsorp- tionyfixation of the applied Cd on the soil constituents (Elliott and Denneny, 1982) and increasing its bioavail- ability and toxicity to biomass. 5.1.5. Nature of microorganisms The bioavailability of Cd, like that of other metals, in soils is governed by the nature of microorganisms and microbially mediated processes. Sorption of Cd to a Gram positive soil bacterium, Rhodococcus erythro- polis A177 was caused mostly by its binding to the cell wall, with a minor part being taken up inside the cells (Plette et al., 1999). Binding of Cd to the cell walls depended on its solution concentration, pH, calcium concentration and ionic strength. Bacterial sorption of Cd was more pronounced in sandy soil than in clay soil under similar conditions (total Cd, pH and calcium concentration). ARTICLE IN PRESS 10 K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx Fig. 2. A generalized diagram showing heavy metal sensitivity to different components of soil biota. Soil microorganisms can alter the metal bioavailabil- ity by releasing specific compounds that form complex with Cd andyor by changing the soil pH. Rayner and Sadler (1989) demonstrated complexation of Cd to polyphosphate granules in the cell membrane and intra- cellular Cd binding proteins in a tolerant strain of Pseudomonas putida, grown in a defined medium con- taining 3 mM Cd . Mechanisms of complexing Cd 2q have evolved over time to allow the cell to survive high intracellular Cd concentrations. Extracellular accumula- tion of metals on microbial cells generally involves either adsorption of metals to the polysaccharide coat- ing, or adsorption to binding sites such as carboxyl, phosphate, sulfahydryl or hydroxyl groups on the cell surface. There is also considerable evidence of microbial methylation of Cd, As, Se, Hg and Pb by bacteria of common genera, such as Bacillus and Pseudomonas (Hughes and Poole, 1989). However, as the studies involving microbial transformations have been per- formed mainly in vitro with pure cultures, the effect of these processes on metal bioavailability and dynamics in soils is not clear. Physiology of the microbes includ- ing the transport of metals across the membranes are often not considered during the assessment of contami- nant bioavailability. The parameters that control the membrane transport of contaminants are species-specific and therefore, the same bioavailable pool of contami- nants in soils may not necessarily result in the same rates of impact on different microorganisms. Competitive interactions between living cells and ligands and the physiological status of the organism should also be considered in the relationship between toxicity and free ion metal concentration (Campbell et al., 2000). Toxic responses of a genetically modified, luminescent bacterium, Escherichia coli revealed a pos- itive relationship between free metal ion activity and toxicity for Cu, but not for Zn and Cd. The stable chloride complexes appeared to contribute to the toxicity of Cd under the test conditions. Nederlof and Van Riemsdijk (1995) attributed the sorption of metal ions by soil organisms to the competition for binding of that metal ion by all reactive soil components (including the organisms). In addition, the total amount of bioavailable metal in soil is influenced by complexation with dis- solved organic and inorganic ligands. The toxicity of Cd to the mineralization of C–ace- 14 tate to C–CO by Pseudomonas putida MT 2 increased 14 2 with increasing pH of soil pore water in model system (Vanbeelen and Fleurenkemila, 1997). Plette et al. (1996) reported a similar increase in the metal toxicity with increasing pH to soil bacteria in a nutrient medium. A close relationship existed between the amount of metals that can be bound by the organism and the amount of metal that can potentially cause an effect. Metal toxicity was closely related to the binding of metal ions to membrane proteins, prior to membrane transport. Evidence suggested that speciation and bioa- vailability of Cd govern the effects of this metal on the biota in the system. Microbial activity in the rhizosphere of plants is several orders of magnitude greater than that in the bulk soil. Root exudates can lower the rhizosphere soil pH generally by one or two units over that in bulk soil. Such changes in pH along with exudation of organic ligands from plant roots may impact metal bioavailabil- ity (Naidu et al., 2003). Since microorganisms are known to mobilize Cd in soils (Chanmugathas and Bollag, 1987) intense microbial interactions in the rhizosphere may lead to increased availability of metals. 5.2. Tolerance and adaptation of microorganisms The effects of chemicals on microorganisms are difficult to evaluate due to the differential sensitivity of microorganisms (Fig. 2), complexity of the population dynamics, multitude of species, and conditions used for toxicity assays. Microorganisms within species of the same genus or within strains of the same species can differ in their sensitivity to metals. Giller et al. (1993) demonstrated that Rhizobium meliloti was less sensitive, in terms of growth, to Cd than R. leguminosarum and R. loti. Tolerance mechanisms in microorganisms toward spe- cific metals often include the binding of metals by cell wall or by proteins and extracellular polymers, forma- tion of insoluble metal sulfides, volatilization and enhanced export from cell (Hughes and Poole, 1989). The microbes also influence the metal availability by changing the pH, metal valence, chelation and other mechanisms (Francis, 1990). Microbes can sorb metals [...]... speciation, ageing, nature of applied metal salt, organisms and other environmental variables In the absence of a uniform methodology, inter-laboratory comparison of the ecotoxicity of Cd and other heavy metals has been very difficult In general, Cd shows more toxicity in sandy soils than in clay soils Soil solution Cd, and not the total concentration of Cd, seems to correlate well with ecotoxicity parameters... fixed Cd in soils (Chanmugathas and Bollag, 1987) Microbial mobilization of Cd, both aerobically and anaerobically, was faster in sandy loam and loam soils than in a silt loam soil The amount of Cd released was more pronounced under anaerobic conditions except in sandy loam soil Air-dried field soils exhibited a significantly greater rate and extent of microbial mobilization of Cd than did fresh soils... ecotoxicity parameters No single parameter can be adequate to generalize Cd toxicity, necessitating the need for a battery of assays with sensitive, reliable and ecologically relevant biological tools Parameters involving species abundance and diversity along with functional parameters give a better understanding of toxicity Despite their ubiquitous occurrence, toxicity of Cd to terrestrial algae is virtually... 1998 Mineralization of nitrogen in a clayey loamy soil amended with organic wastes enriched with Zn, Cu and Cd Bioresour Technol 64, 39–45 Hattori, H., 1989 Influence of cadmium on decomposition of sewage-sludge and microbial activities in soils Soil Sci Plant Nutr 35, 289–299 Hattori, H., 1991 Influence of cadmium on decomposition of glucose and cellulose in soil Soil Sci Plant Nutr 37, 39–45 Hattori,... environmental toxicity of individual metals such as Cd Most studies on metal toxicity to microorganisms are based on the effects of sewage sludge, which are known to contain a wide range of metals and complex substances Also, the results of such studies with sewage sludge cannot be extrapolated to agricultural field soils fertilized with phosphatic fertilizers rich in Cd or in laboratory-incubated soils spiked... with Cd alone Besides, several factors act on the physico-chemical properties of the field-contaminated soil and the field toxicity of Cd need not necessarily be of the same magnitude as in laboratory studies Also, it is difficult to compare the effect of the heterogeneously distributed pollutant in the field to that of a homogeneously distributed contaminant under laboratory conditions The soil storage... to treatment with Cd, had a marked effect on its toxicity to microbial respiration in forest litter samples, the effect varying with the soil type (Niklinska et al., 1998) However, most toxicological studies have ignored the effects of soil storage on the parameter studied 11 Similarly, subjecting soil to wet–dry cycles may give a different toxicity response to metal compared to freshly incubated soil. .. Determination of the effects of Cd on soil biological activity often yields conflicting results Such inconsistencies on its ecotoxicity can be attributed to overgeneralization of the outcomes from short-term laboratory studies that focus on a single soil type under controlled conditions Field data on the effects of individually applied Cd are limited Most of the field information on the effects of Cd. .. and rewetting as an indicator of metal toxicity in soils Soil Biol Biochem 33, 235–240 Mong, H.B., Brown, L.R., Kim, J.K., 1995 Effect of zinc and calcium on the intracellularly uptake of cadmium and growth of Escherichia coli J Microbiol 33, 302–306 Moreno, J.L., Garcia, C., Landi, L., Falchini, L., Pietramellara, G., Nannipieri, P., 2001 The ecological dose value (ED50) for assessing Cd toxicity on... for more research on changes in species diversity and microbial adaptations in relation to Cd pollution 5.3 Methodologies used Inter-laboratory comparison of metal toxicity studies has been difficult because of the different methods used for measuring bioavailable Cd In fact, there is no single standard technique available for assessing the ecotoxicological impacts of Cd on soil biota Limited studies . early and sensitive indicator of stress in soil (Brookes, 1995), our aim here is to (i) compile the global information on Cd toxicity to soil microbes and their. (iv) deter- mine the factors that control Cd bioavailability and toxicity to soil microorganisms. 2. Soil microorganisms and their significance Soil serves