Bioavailability and toxicity of cd to microorganisms and their activities in soil

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:
(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
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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-
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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)
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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)
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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.
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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
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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).
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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
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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
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11K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx
which may lead to increased proportions of metal in the
solid phase (Brown et al., 1994). Mong et al. (1995)
found that living cells of Escherichia coli took up more
Cd than dead cells, especially in the log phase. Probably
some metabolic reactions or their products promoted the
uptake of Cd.
Microorganisms can mobilize strongly bound or 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 pro-
nounced 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 that had been kept moist at 5 8C
under aerobic conditions.
Evidence suggests a change in the genetic structure
of the soil microbial community under long-term metal
stress (Giller et al., 1998). Such adverse effects impact-
ed by Cd can lead to a reduction in biodiversity and
resultant functions in the soil. A gradual change in the
microbial community structure was noticed in labora-

tory-incubated soils amended with Cd (Frostegard et
al., 1993; Griffiths et al., 1997). There is a need 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 ecotoxi-
cological impacts of Cd on soil biota. Limited studies
have considered the 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-contaminat-
ed 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 heterogene-
ously distributed pollutant in the field to that of a
homogeneously distributed contaminant under labora-
tory conditions.
The soil storage time, prior 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 toxicolog-

ical studies have ignored the effects of soil storage on
the parameter studied.
Similarly, subjecting soil to wet–dry cycles may give
a different toxicity response to metal compared to
freshly incubated soil as wet–dry cycle changes the soil
structure, pH (Paul et al., 1999) and chemical compo-
sition of the soil solution (Naidu and Haynes, 1999).
Merckx et al. (2001) observed no change in the DOC
concentration in soil solutions during 23 days of moist
incubation following spiking with Zn. But, after the two
air-drying and rewetting cycles, the DOC concentrations
significantly increased. Drying and wetting cycles can
also change the substrate availability to the microorgan-
isms. Decrease in biomass carbon, ATP content, enzyme
activities and nitrification in air-dried soils compared to
field moist soils has been reported (Sparling et al.,
1986; Bauhus and Khanna, 1994). A significantly great-
er rate and extent of microbial Cd mobilization were
noticed in air-dried field soils than in fresh soils that
had been kept moist at 5 8C (Chanmugathas and Bollag,
1987).
6. Conclusions
Determination of the effects of Cd on soil biological
activity often yields conflicting results. Such inconsis-
tencies on its ecotoxicity can be attributed to over-
generalization of the outcomes from short-term
laboratory studies that focus on a single soil type under
controlled conditions. Field data on the effects of indi-
vidually applied Cd are limited. Most of the field
information on the effects of Cd has been generated

from experiments using sludge containing multimetals,
and soil solution characteristics of such sludge-amended
soils are different from those in soils freshly spiked
with Cd alone. Only few studies have investigated the
relationship between soil properties, Cd loading and the
effect on Cd bioavailability. The bioavailable fraction,
the key factor in risk assessment analysis, varies with
time, soil type, speciation, ageing, nature of applied
metal salt, organisms and other environmental variables.
In the absence of a uniform methodology, inter-labora-
tory 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. No
single parameter can be adequate to generalize Cd
toxicity, necessitating the need for a battery of assays
with sensitive, reliable and ecologically relevant biolog-
ical 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
not known. Moreover, soil algae can probably be used
as an effective tool in ecotoxicity assays since these
organisms are known to be sensitive to a wide variety
of pollutants.
ARTICLE IN PRESS
12 K. Vig et al. / Advances in Environmental Research xx (2002) xxx–xxx
Acknowledgments
This work was supported by The Remediation of

Contaminated Environments Program, CSIRO Land and
Water, Adelaide, Australia. We thank Dr Peter Franz-
mann for his critical comments on the manuscript.
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