Applied Soil Ecology 21 (2002) 149–158
Toxicity of cadmium to soil microbial activity: effect of sewage
sludge addition to soil on the ecological dose
José L. Moreno a,∗ , Teresa Hernández a , Aurelia Pérez b , Carlos Garcı́a a
a
Department of Soil and Water Conservation and Waste Management, Centro de Edafologı́a y Biologı́a Aplicada
del Segura (CEBAS-CSIC), P.O. Box 4195, 30100 Murcia, Spain
b Universidad Miguel Hernández, Carretera de Beniel km 3.2, 03312 Orihuela, Alicante, Spain
Received 21 January 2002; received in revised form 28 May 2002; accepted 29 May 2002
Abstract
Cadmium has a toxic effect on soil microbial activity which plays an important role in nutrient cycling and, therefore,
in maintaining soil fertility. In addition, the mobility of this heavy metal in soil is affected by the addition of urban wastes
such as sewage sludge. This study was conducted to determine the effect of sewage sludge amendment of a semiarid soil,
previously polluted with Cd, on the toxic effect of this heavy metal on soil microbial biomass and its activity. Dehydrogenase
activity, ATP content, microbial soil respiration and microbial biomass carbon were used as bioindicators of the toxic effect
of Cd. The inhibition of microbial activity and biomass by different Cd concentrations ranging from 0 to 8000 mg Cd kg−1
soil was described by three mathematical models in order to calculate three ecological doses of Cd: ED50 , ED10 , and ED5 . In
general, higher ED values were calculated for the sewage sludge amended soil than for unamended soil. Thus the Cd toxicity
to microbial activity of the sewage sludge amended soil can be considered lower than that of the unamended soil. Moreover,
increased ED values with time after soil Cd contamination were observed.
© 2002 Elsevier Science B.V. All rights reserved.
Keywords: Soil microbial activity; Cadmium; Sewage sludge; Ecological dose
1. Introduction
Cadmium is one of the most dangerous of soil
pollutants because this heavy metal may easily move
from soil to food plants through root absorption, and
fairly large amounts can accumulate in their tissues
without showing stress (Oliver, 1997). In this way,
Cd may enter the food chain and so affect human
health, as poisoning of hundreds of inhabitants of
Japan demonstrated (Adriano, 1986). The residents
close to a mining operation had been ingesting Cd
over a 30-year period in their rice, in which Cd had
∗ Corresponding author. Fax: +34-968-396213.
E-mail address: [email protected] (J.L. Moreno).
accumulated through the polluted river water used for
irrigation. Furthermore, Cd may have a toxic effect
on the soil microbial communities, which play an
important role in soil nutrient cycling and, therefore,
in soil fertility. There have been many studies on this
topic (Brookes, 1995; Nannipieri et al., 1997; Giller
et al., 1998). However, little is known about the effect
of Cd on the microbial activity of semiarid soils.
Urban wastes, such as sewage sludge, are increasingly used to amend soils, especially those with a
low organic matter content, to improve their fertility
(Garcı́a et al., 1994). However, addition of sewage
sludge to soil may either contribute to heavy metal
pollution of soils (Adriano, 1986; Alloway, 1995)
or affect heavy metal mobility in the soil (Saviozzi
0929-1393/02/$ – see front matter © 2002 Elsevier Science B.V. All rights reserved.
PII: S 0 9 2 9 - 1 3 9 3 ( 0 2 ) 0 0 0 6 4 - 1
150
J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
et al., 1983). On the other hand, organic amendment
increases activity and diversity of soil microbial populations, and over time microbial tolerance to heavy
metals can be observed (Barkay et al., 1985).
In response to the apparent need to easily quantify the influence of pollutants on microbe-mediated
ecological processes in various ecosystems, the concept of an ecological dose (ED50 ) was developed,
which is the toxicant concentration that inhibits a
microbe-mediated ecological process by 50% (Babich
et al., 1983). However, a 50% reduction in a basic
ecological process may be too extreme for the continued functioning of a soil and so, lower percentage
values of inhibition (5, 10 or 25%) equivalent to ED5 ,
ED10 or ED25 need to be established. These values
may be more suitable criteria for protecting soil quality and for assessing sensitivity of a soil ecosystem
to a stress (Doelman and Haanstra, 1989; Kostov and
Van Cleemput, 2001). Different researchers have proposed several mathematical models to calculate the
ecological dose value. Haanstra et al. (1985) used a
sigmoidal curve to describe the inhibition of urease
and phosphatase activity as a function of the natural
logarithm of the concentration of heavy metals, while
Speir et al. (1995) used two models based on enzyme
inhibition kinetics.
In this study, we measured different microbial indicators of untreated soils and of soils treated with
sewage sludge containing a wide range of Cd concentrations to evaluate the effect of this organic amendment on the sensitivity of soil microbial activity to Cd
toxicity. The microbial biomass and activity indicators used were: dehydrogenase activity, ATP content,
microbial respiration, and biomass carbon. Using a
statistical test of non-linear correlation we evaluated
which of the previously mentioned models best fit the
experimental data. We calculated the ED values for
Cd using all three models and compared changes in
these values in both unamended and sewage sludge
amended soils over time.
2. Materials and methods
2.1. Soil and sewage sludge
The experiment was carried out with a soil from
Beniel, Murcia (SE Spain). The soil samples were
Table 1
Some chemical and physical characteristics of unamended and
amended soil
Parameters
Soil
Sewage sludge
amended soil
Sand (%)
Silt (%)
Clay (%)
WHC (33.8 kPa), %
pH (1:2.5 soil:water)
EC (1:5 soil:water) (␮S cm−1 )
TOC (%)
Total N (g kg−1 )
11.3
35.1
53.6
33.8
8.6
250
0.9
2.0
–
–
–
–
8.3
320
1.8
3.2
Total content of heavy metals (mg kg−1 soil)
Cd
0.37
Ni
5.0
Cu
8.0
Pb
6.0
Zn
52.0
Cr
26.0
Hg
0.40
0.38
6.0
10.0
13.0
75.0
28.0
0.45
collected from the plough layer (0–30 cm) of an agricultural plot without any crop. Soil samples were
mixed and dried at room temperature to facilitate the
sieving. Dehydrated aerobic sewage sludge from a
waste-water treatment plant in Murcia, which treats
4000 m3 of waste-water from the municipality per
day, was used as organic soil amendment. The characteristics of the unamended and amended soil are
shown in Table 1.
2.2. Soil amendment and incubation
Two hundred grams of sieved soil (<2 mm) and
soil amended with sewage sludge (to a final value
of 5%, w/w) were put into a semi-closed microcosm
(Naseby and Lynch, 1998). Microcosms with added
sewage sludge and without added sewage sludge
were treated with Cd. Two control microcosms (with
and without sewage sludge) lacking Cd treatment
were also prepared. The Cd treatments consisted of
CdSO4 solution additions to achieve several final Cd
concentrations in soil: 25, 50, 100, 250, 500, 1000,
2000, 4000, 6000, and 8000 mg Cd kg−1 soil. All
Cd treatments were incubated in the dark at 25 ◦ C
and 70–80% humidity over three time periods (3 h,
12 days, and 40 days). Soil moisture content was
adjusted to 50–60% of its water-holding capacity
(WHC) during the incubation time as needed. After
J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
the above time periods concluded, the content of each
microcosm was put in plastic bags and stored at 4 ◦ C
before use for microbial parameter assay
2.3. Soil microbial parameters
All of the following bioindicators were measured
in triplicate and, means of the triplicate measurements
were used to calculate ED values.
Microbial biomass C was determined by the
fumigation–extraction method (Vance et al., 1987).
Three grams of sample from each microcosm were
fumigated with chloroform and other 3 g were not
fumigated. C was extracted with K2 SO4 (0.5 M) solution from fumigated and non-fumigated samples, and
the C content was measured in the centrifuged samples using a soluble-organic C analyzer (Shimadzu
TOC-5050A). Microbial biomass C (MBC) was calculated by the expression: MBC = C extracted × 2.66
(Vance et al., 1987), where C extracted is the difference between C extracted from fumigated samples
and C extracted from non-fumigated samples.
Soil respiration was analyzed by placing 50 g of
soil with moisture content corrected to 50–60% of its
water-holding capacity, in hermetically sealed flasks
and incubating for 31 days at 28 ◦ C. The CO2 produced was periodically measured (after 1, 2, 3, 4
days and then weekly) using an infrared gas analyzer
(TORAY PG-100, Toray Engineering Co. Ltd., Japan).
The data were summed to give a cumulative amount of
CO2 evolved after 31 days of incubation, and soil respiration was expressed as mg CO2 C kg−1 soil per day.
ATP was extracted from soil using the Webster et al.
(1984) procedure and measured as recommended in
Ciardi and Nannipieri (1990). Twenty millilitres of a
phosphoric acid extractant was added to 1 g of soil,
and the closed flasks were shaken in a cool bath.
Then the mixture was filtered through Whatman paper and an aliquot was used to measure the ATP content by the luciferin–luciferase assay in a luminometer
(Optocomp 1, MGM Instruments, Inc.).
Soil dehydrogenase activity (DH activity) was
determined using 1 g of soil, and the reduction of
p-iodonitrotetrazolium chloride (INT) to p-iodonitrotetrazolim formazan (INTF) was measured as a
modification of the method reported by Von Mersi
and Schinner (1991). Soil DH activity was expressed
as gINTF g−1 soil.
151
2.4. Mathematical models
The two kinetic models proposed by Speir et al.
(1995) and the sigmoidal dose-response model proposed by Haanstra et al. (1985) were used to calculate the ED values and evaluate the suitability of these
models for describing Cd inhibition of soil biological
and biochemical properties. The algebraic expressions
of kinetic models were:
c
(Model 1)
v=
(1 + bi)
v=
c(1 + ai)
(1 + bi)
(Model 2)
The constants a, b and c were always positive, with
b > a. The constant c represents the uninhibited value
of the tested parameter, and the constant a and b depend on the curve slope. Model 1 describes the full
inhibition of v (tested parameter) by i, the concentration of inhibitor (Cd concentration) and Model 2
describes the partial inhibition. For data fitting both
Model 1 and Model 2, it was possible to calculate the
ecological dose values from the relationships:
ED50 =
1
b
ED10 =
1
9b
ED5 =
5
95b
Model 1 and Model 2 describe a concave rectangular
hyperbolic relationships between v and i, but the latter has an asymptote ca/b, parallel to, but above the
x-axis. The mathematical equation for the sigmoidal
dose-response model was:
a
(Model 3)
y=
{1 + exp[b(x − c)]}
where y is the tested parameter, x the natural logarithm
of Cd concentration, a the uninhibited value of y, b
a slope parameter indicating the inhibition rate, and c
the natural logarithm of ED50 . The ED values were
calculated using the following expressions:
ED50 = exp c
2.2
ED10 = exp c −
b
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J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
ED5 = exp c −
2.9
b
Model 3 describes a logistic curve which is the relationship between the measured activity and the
natural logarithm of the inhibitor concentration. Diagrammatic representations of these three models are
presented in Speir et al. (1995) and Haanstra et al.
(1985). The values of the constant a, b and c of the
models were estimated using the Marquadt’s iterative
search algorithm of the statistical program STATGRAPHICS Plus version 2.1 (Statistical Graphics
Corp.). The value of the coefficient of determination
(r2 ) of the non-linear regression was only determined
at P < 0.05.
3. Results
The ED50 , ED10 and ED5 values calculated with
all three models for soil dehydrogenase activity,
Table 2
Values of r2 (P < 0.05) obtained from the regression analysis
of Model 1 (M1), Model 2 (M2) and Model 3 (M3) which best
describe the inhibition of soil dehydrogenase activity by Cd, and
ED values (mg Cd kg−1 soil) predicted from these models
Treatment
Model
r2
ED50
ED10
soil ATP content, soil respiration and soil microbial
biomass C are shown in Tables 2–5. These ED values
were calculated for both the unamended and amended
soil after the different incubation periods mentioned
(3 h, 12 days and 40 days). Plots showing the curve
of the models, which best fitted the measured soil
bioindicators, are shown in Figs. 1–4.
All three mathematical models described well the
inhibition of soil dehydrogenase activity (Table 2).
Thus, the values of the coefficient of determination (r2 )
were high (0.87–0.98), except for the sewage sludge
amended soil 12 days after the addition of Cd (0.7 for
Model 3). Fig. 1(E) shows both a very low Cd inhibition of DH activity of the sewage sludge soil and a high
dispersion of data. The ED50 values calculated with the
three models were similar after 3 h, but ED10 and ED5
were higher using Model 3 than Model 1 or Model 2.
After 12 and 40 days of Cd exposure, the lowest values of ED10 and ED5 calculated with Model 3 were
observed. ED50 values for the sewage sludge amended
Table 3
Values of r2 (P < 0.05) obtained from the regression analysis
of Model 1 (M1), Model 2 (M2) and Model 3 (M3) which best
describe the inhibition of soil ATP by Cd, and ED values (mg
Cd kg−1 soil) predicted from these models
ED5
Treatment
3 h of incubation
Soil
M1
M2
M3
Soil + SSa
M1
M2
M3
12 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
40 days of incubation
Soil
M1
M2
M3
Soil + SS
a
M1
M2
M3
0.8687
0.8898
0.8910
2500.0
2209.9
2340.9
277.8
245.5
530.5
131.6
116.3
322.0
0.9725
0.9800
0.9734
1666.7
1628.7
1629.1
185.2
181.0
224.3
87.7
85.7
115.1
0.8122
0.8080
0.8348
2500.0
2312.7
981.6
277.8
257.0
13.5
131.6
121.7
3.2
0.6184
0.6267
0.7000
7692.3
1397.0
5171.4
854.7
155.2
4.5
404.9
73.5
0.4
0.9211
0.9247
0.9580
1428.6
833.3
669.7
158.7
92.6
12.2
75.2
43.9
3.2
0.8915
0.9375
0.9560
2500
526.3
1494.2
277.8
58.5
26.1
131.6
27.7
6.7
Soil + SS: sewage sludge amended soil.
Model
3 h of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
12 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
40 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
r2
ED50
ED10
ED5
0.9264
0.9264
0.9584
1592.4
1592.4
1917.2
176.9
176.9
514.0
83.8
83.8
330.1
0.8164
0.8126
0.8121
1306.3
1097.7
1121.4
145.1
122.0
86.5
68.8
57.8
36.6
0.9681
0.9681
0.9711
1193.3
1193.3
1226.2
132.6
132.6
177.3
62.8
62.8
92.5
0.9084
0.9084
0.9154
1930.1
1930.5
1975.2
214.5
214.5
305.1
101.6
101.6
162.8
0.9655
0.9655
0.9691
1414.4
1414.4
1451.8
157.2
157.2
217.1
74.4
74.4
114.6
0.9328
0.9328
0.9434
2551.0
2551.0
2714.3
283.4
283.4
518.5
134.3
134.3
297.1
J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
Table 4
Values of r2 (P < 0.05) obtained from the regression analysis
of Model 1 (M1), Model 2 (M2) and Model 3 (M3) which best
describe the inhibition of soil microbial respiration by Cd concentration, and ED values (mg Cd kg−1 soil) predicted from these
models
Treatment
Model
3 h of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
12 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
40 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
r2
ED50
ED10
ED5
0.9725
0.9795
0.9764
373.8
270.6
317.9
41.5
30.1
22.2
19.7
14.2
9.1
0.9347
0.9351
0.9387
4514.7
3351.2
4119.3
501.6
372.4
290.4
237.6
176.4
119.0
0.8971
0.8965
0.8938
453.9
323.7
426.6
50.4
36.0
40.7
23.9
17.0
18.5
∗ n.f.
0.6484
0.5383
–
140.9
11758.4
–
15.7
12.6
–
7.4
1.3
0.8980
0.9135
0.9093
1568.1
704.2
1194.9
174.2
78.2
53.9
82.5
37.1
19.0
n.f.
0.5088
n.f.
–
46.3
–
5.1
–
–
2.4
–
153
Table 5
Values of r2 (P < 0.05) obtained from the regression analysis
of Model 1 (M1), Model 2 (M2) and Model 3 (M3) which best
describe the inhibition of microbial soil biomass C by Cd, and
ED values (mg Cd kg−1 soil) predicted from these models
Treatment
Model
3 h of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
12 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
40 days of incubation
Soil
M1
M2
M3
Soil + SS
M1
M2
M3
r2
ED50
ED10
ED5
0.757
0.8415
0.9282
588.2
82.6
320.8
65.4
9.2
2.3
31.0
4.3
0.4
0.4657
0.5333
n.f.
8196.7
653.6
–
910.7
72.6
–
431.4
34.4
–
0.8112
0.8112
0.8318
2487.6
2487.6
3310.8
276.4
276.4
238.0
130.9
130.9
98.2
0.6749
0.7427
0.7450
2439.0
375.9
2796.4
271.0
41.8
34.1
128.4
19.8
7.8
0.3772
0.5151
0.7502
5649.7
63.3
5280.4
627.7
7.0
1799.6
297.4
3.3
1253.0
0.3791
0.3778
0.6497
7518.8
2770.1
6836.3
835.4
307.8
1419.4
395.7
145.8
836.5
∗
n.f. the model did not fit the experimental data; r2 values are
calculable but the level of significance for the regression was not
P < 0.05.
soil were higher than those for the unamended soil after 12 and 40 days of incubation, but were lower after 3 h of incubation. Only at 40 days were the ED10
and ED5 values estimated with Model 1 and Model 3
for the amended soil were higher than for unamended
soil. The ED50 value for unamended soil calculated
with the Model 3 decreased with the incubation time.
For soil ATP content, the coefficients of determination for the three models were high in all cases
(Table 3). In general, the ED50 values predicted from
the three models were similar but there was greater
discrepancy in the predicted ED10 and ED5 values.
The ED values for the amended soil were higher than
unamended soil except after 3 h of incubation. There
was an increase in ED values with incubation time in
the amended soil, but not in the unamended soil.
For soil microbial respiration, the r2 values for the
three models were high except for the sewage sludge
amended soil at 12 and 40 days (Table 4). In these
two cases the low Cd inhibition of soil microbial
respiration and dispersion of data produced a poor
fit the mathematical models used (Fig. 3). In those
cases where a great discrepancy was observed between the ED values predicted by the three models,
Model 2 gave the best fit to the measured data. With
this model, the ED values for the amended soil were
lower than those for the unamended soil at 12 days
and 40 days. However, this finding may not reflect
reality because the r2 values for this model are low in
some cases. The lowest ED10 and ED5 values were
calculated from Model 3. Increased ED values with
incubation time were observed in the unamended soil.
None of the models provided a good fit for soil
microbial biomass C measurements at the different
Cd concentrations, although in most cases Model 3
was the best for predicting the ED values (Table 5).
ED10 and ED5 values calculated from this model were
lowest after 3 h and 12 days of Cd exposure. A large
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J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
Fig. 1. Dehydrogenase activity both in unamended soil after 3 h, 12 days, and 40 days of incubation (A, B, and C, respectively) and in sewage
sludge amended soil after 3 h, 12 days, and 40 days of incubation (D, E, and F, respectively) with different levels of Cd contamination.
Only the curves which best fitted the experimental data have been plotted for each case (Model 2 for D and Model 3 for the rest).
decrease in microbial biomass carbon with increasing
Cd concentration was observed in unamended soil
after 3 h of incubation (Fig. 4). The ED50 values for
the sewage sludge amended soil after 3 h and 40 days
were higher than those for the unamended soil. However, only after 3 h were the ED10 and ED5 values for
the amended soil higher than for the unamended soil.
In general, ED values increased with incubation time
both for unamended and amended soil.
4. Discussion
In most cases, the soil microbiological indicators
measured decreased with increasing Cd concentrations, which indicates that Cd has a toxic effect on
both the microbial metabolic processes and microbial
growth (Figs. 1–4). The microbial indicators also
decreased with the incubation time, presumably due
to the diminution with time of the substrates easily
available to microorganisms (Moreno et al., 1999).
The ED values obtained for Cd were greater than the
maximum concentration of Cd permitted in soils by
EU legislation (3 mg kg−1 soil). However, for safety,
a value 10 or 1000 times lower than ED50 (depending on the statistical reliability of this value) has been
proposed as a maximum limit for heavy metal in soils
(Doelman and Haanstra, 1989). Other authors have
used the ED10 and ED5 values to propose regulatory
criteria for heavy metals in soils (Scott-Fordsmand and
Pedersen, 1995). These ED10 and ED5 values may be
more suitable indicators of the sensitivity of an ecosystem to a given stress, because a 50% reduction in a
basic ecological process may be too extreme for its
J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
155
Fig. 2. ATP content in both unamended soil after 3 h, 12 days, and 40 days of incubation (A, B, and C, respectively) and in sewage sludge
amended soil after 3 h, 12 days, and 40 days of incubation (D, E, and F, respectively) with different levels of Cd contamination. Only the
curves which best fitted the experimental data have been plotted for each case (Model 1 for D and Model 3 for the rest).
continued functioning (Babich et al., 1983). Discrepancy observed in both ED10 and ED5 values calculated from the three models can be due to the great
elongation of initial part of the curve of Model 3 and
thus small variation in the rate of decrease of the studied bioindicator can produce a large variation in ED10
and ED5 values calculated from this model. Model 2
is a partial inhibition model and the minimum possible value of the studied bioindicator is always higher
than zero. When the asymptote value of this model
is high, the ED50 value calculated is lower than that
from the other two models.
In general for this experiment, the addition of
sewage sludge to the soil diminished the inhibitory
effects of Cd on biological parameters. The addition
of this kind of organic material can help in the remediation of Cd polluted soils provided the heavy metal
content of the organic material is very low (Alloway,
1995). The sewage sludge applied to soil may increase
the biodiversity of soil microorganisms and their
metabolic activity (Seaker and Sopper, 1988), favoring a selection of heavy metal resistant microorganisms with time (Kandeler et al., 2000). Furthermore,
the humic substances of sewage sludge contribute to
complexation and adsorption of Cd and thus restrict
the movement of this heavy metal trough the soil
profile and into the water table (Saviozzi et al., 1983).
In general, the ED values increased noticeably over
time since soil Cd treatment. Doelman and Haanstra
(1984) and Speir et al. (1995) also reported an increase
in ED50 over time for different heavy metals, when the
inhibition of soil microbial parameters by soil heavy
metal pollution was studied. It is possible that this is
due to an increase in Cd resistance in microorganisms
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J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
Fig. 3. Microbial respiration both in unamended soil after 3 h, 12 days, and 40 days of incubation (A, B, and C, respectively), and in sewage
sludge amended soil after 3 h, 12 days, and 40 days of incubation (D, E, and F, respectively) with different levels of Cd contamination.
Only the curves which best fitted the experimental data have been plotted for each case (Model 1 for B; Model 2 for A, C, E and F and
Model 3 for D).
as observed by other authors (Bewley and Stotzky,
1983).
Dehydrogenase activity (DH activity) is related to
a group of intracellular enzymes present in active soil
microorganisms. This enzymatic activity has been
used as an index of overall microbiological activity in
toxicity assays (Nannipieri et al., 1997). DH activity
was more affected by increasing Cd concentration in
the amended soil after 3 h than in the unamended soil,
perhaps because under conditions of C starvation or
low content of energy sources for microorganisms,
metals taken up by an energy-driven transport system
cannot be accumulated in toxic concentrations inside
microorganisms (Giller et al., 1998). Differences in
C availability in the media have been shown to have
a profound effect on the toxicity of Cd, Cu and Zn
to Klebsiella sp. (Brynhildsen et al., 1988). Over
time, however, the amended soil had higher ED50
than unamended soil. Perhaps this is due to either a
complexation of Cd with the humic substances from
sewage sludge or a different evolution of the soil
microbial community.
Soil ATP is an independent measurement of microbial biomass when soil samples are pre-incubated under controlled conditions before analyses (Nannipieri
et al., 1990). However, soil ATP and microbial biomass
C exhibited different response patterns to Cd treatment. The correlation coefficients for soil ATP for
the three models were higher than those for microbial
biomass (Tables 3 and 5), and thus there was less discrepancy between ED values calculated for ATP from
the three models than for microbial biomass C.
In general, the soil microbial respiration measurements fit Model 2 best (Table 4). This model assumes
J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
157
Fig. 4. Microbial biomass C both in unamended soil after 3 h, 12 days, and 40 days of incubation (A, B, and C, respectively), and in sewage
sludge amended soil after 3 h, 12 days, and 40 days of incubation (D, E, and F, respectively) with different levels of Cd concentration. The
curve which best fitted the experimental data has been plotted for each case (Model 1 for B; Model 2 for C, D, E, F and Model 3 for A).
that a portion of the measured activity is not affected
by high soil Cd concentrations. This means that the
full inhibition of the measured activity is never 100%
of the control value (Dixon and Webb, 1979). Basically, Model 1 and Model 3 are full inhibition models,
although the latter model describes the inhibition of a
measured parameter with the natural logarithm of the
Cd concentration. In these cases the inhibition of measured parameters can be 100% of the control value.
The ED values predicted by Model 2 were lower than
those predicted by Models 1 and 3 probably due to
high asymptotic values (minimum value of measured
parameter). Thus, for Model 2, EDx represent the inhibitor concentration which reduces activity from the
maximum value of control (uninhibited) to x% of the
difference between this maximum value and the minimum possible activity.
In conclusion this study suggests that the addition
of sewage sludge to the soil diminished the toxic effect of Cd on microbial function, a finding which may
be of great use for the remediation of soils with severe Cd contamination. Examples of such soils would
be non-agricultural soils near smelters or mines. However, the addition of sewage sludge to the soil cannot
be considered a definitive solution for Cd soil contamination because this heavy metal still remain in soil and
its bioavailability can change with time. The values
of the ecological dose which produce a specific level
of inhibition of soil microbial parameters, are suitable
to set a maximum limit of Cd concentration in order
to avoid irreversible effects on soil functionality. ATP
content is the parameter of those studied, for which the
mathematical models used were the most suitable to
predict the ED values of Cd. However, for other soils
158
J.L. Moreno et al. / Applied Soil Ecology 21 (2002) 149–158
a different indicators may produce a better fit to these
models. So to safeguard the “health” of the soil, it is
necessary to evaluate multiple indicators when using
EDx values as a criterion to set a maximum permissible concentration of a contaminant in soil. Since ED
values changed with duration of incubation, it seems
that a long-term incubation is better to estimate ED
values for a soil, in order that the microbial community has time enough to adapt to the stress conditions.
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