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This article was downloaded by: [University of Alberta]On: 26 November 2014, At: 18:40Publisher: Taylor & FrancisInforma Ltd Registered in England and Wales Registered Number: 1072954 Registered office:Mortimer House, 37-41 Mortimer Street, London W1T 3JH, UK
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Metal‐Associated Forms and Speciation inBiosolid‐Amended OxisolsM. L. Silveira a , A. C. Chang b , L. R. F. Alleoni c , G. A. O'Connor d & R.Berton ea Range Cattle Research and Education Center, University of Florida , Ona,Florida, USAb Environmental Sciences Department , University of California , Riverside,California, USAc Department of Soils and Plant Nutrition , University of Sao Paulo ,Piracicaba, Brazild Soil and Water Science Department , University of Florida , Gainesville,Florida, USAe Campinas Agronomic Institute , Campinas, BrazilPublished online: 25 Apr 2007.
To cite this article: M. L. Silveira , A. C. Chang , L. R. F. Alleoni , G. A. O'Connor & R. Berton (2007)Metal‐Associated Forms and Speciation in Biosolid‐Amended Oxisols, Communications in Soil Science and PlantAnalysis, 38:7-8, 851-869, DOI: 10.1080/00103620701263700
To link to this article: http://dx.doi.org/10.1080/00103620701263700
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Metal-Associated Forms and Speciationin Biosolid-Amended Oxisols
M. L. Silveira
Range Cattle Research and Education Center, University of Florida,
Ona, Florida, USA
A. C. Chang
Environmental Sciences Department, University of California,
Riverside, California, USA
L. R. F. Alleoni
Department of Soils and Plant Nutrition, University of Sao Paulo,
Piracicaba, Brazil
G. A. O’Connor
Soil and Water Science Department, University of Florida,
Gainesville, Florida, USA
R. Berton
Campinas Agronomic Institute, Campinas, Brazil
Abstract: The objective of this study was to determine the effects of pH and ionic
strength on the distribution and speciation of zinc (Zn), copper (Cu), and cadmium
(Cd) in surface soil samples from two Brazilian Oxisols amended with biosolids.
Soils and biosolids were equilibrated in an experimental dual-chamber diffusion
apparatus that permits the soils and biosolids to react through a solution phase via
diffusion across a membrane. After equilibrium was reached, soil and biosolids
samples were sequentially fractionated to identify various solid forms of Zn, Cu, and
Cd. Metal concentrations in the solution phase were determined and mass balance
Received 9 June 2005, Accepted 16 March 2006
Address correspondence to M. L. Silveira, UF/IFAS Range Cattle Research and
Education Center, 3401 Experiment Station, Ona, Florida 33865, USA. E-mail:
Communications in Soil Science and Plant Analysis, 38: 851–869, 2007
Copyright # Taylor & Francis Group, LLC
ISSN 0010-3624 print/1532-2416 online
DOI: 10.1080/00103620701263700
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calculated. Equilibrating pH had no major effect on Cu solubility from biosolids and, at
pH range from 4 to 7, most Cu remained in the biosolids. Soluble Zn and Cd concen-
tration increased with decreasing pH because of the increased solubility of the
biosolids. Copper and Zn were primarily associated with the residual fraction and Fe
oxides in one soil, but were primarily associated with chemically unstable fractions,
or adsorbed to the surface of oxides, in the other soil. In both soils, Cd was
primarily associated with readily bioavailable fractions. The effect of pH on the
metal distribution was more evident than the ionic strength effect. Free ions were the
predominant metal species in solution, especially at lower pH values.
Keywords: Biosolids, cadmium, chemical speciation, copper, heavy metals, oxisols,
sequential fractionation, zinc
INTRODUCTION
Biosolids are residues generated during primary, secondary, or advanced
treatment of domestic sanitary sewage through one or more controlled
processes that reduce pathogens and attractiveness to vectors. The term
biosolids is related to the definition of sewage sludge found in Part 31,
Water Resources Protection of the Natural Resources and Environmental Pro-
tection Act, 1994 PA 451, as amended; however, biosolids are only that
portion of sewage sludge that undergoes adequate treatment (pathogen
reduction) to permit application to land (USEPA 1994).
Land-applied biosolids can improve physical–chemical soil properties,
such as pH, cation-exchange capacity, and aggregate stability (Tsadilas
et al. 1995), and serve as organic sources of nitrogen (N), phosphorus (P),
sulfur (S), and micronutrients for plant nutrition. However, biosolids can
also contain high concentrations of heavy metals that can accumulate to
problematic concentrations in agricultural soils receiving biosolids for long
periods.
Metals present in the biosolids can be solubilized for reaction with soils,
and the kinetics of metal dissolution determine the rate at which equilibrium is
reached (Gerritse et al. 1983). The chemical reactivity of metals is determined
by the chemical equilibrium between metals in solution and in solid phases.
Although metal solubility can be initially reduced by soil sorption reactions,
long-term solubility is controlled by chemical forms in the solid phases
(Martinez and McBride 1998). Therefore, knowledge of metal-associated
forms in the soil and the biosolids and chemical speciation in solution
is essential for understanding soil-metal chemistry in the environment
(Mattigod and Page 1983).
Metals present in soils or biosolids can be associated with different com-
ponents. Sequential extraction techniques have been widely used to assess
heavy-metal distributions in the solid phase (Shuman 1985, 1991). This
procedure is especially useful to determine risks of soil contamination, to
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predict the potential soil’s ability to release metals into solution (Krishnamurti
and Naidu 2002), and to determine the long-term behavior of metals intro-
duced into the soil and their redistribution with time (Candelaria and Chang
1997). The soluble and exchangeable fractions have received special
attention because they are considered readily available to plants.
Heavy-metal mobility and retention are correlated with the chemical and
mineralogical soil characteristics (Matos et al. 2001; Appel and Ma 2002;
Covelo, Couce, and Vega 2004). Soils contain a variety of surface functional
groups, such as hydrous oxide minerals, organic matter, and alumosilicates,
responsible for metal sorption reactions. Crystalline and poorly crystallized
iron oxides are by far the most active components, from a chemical standpoint,
in the geochemical cycling of trace elements in soils (Martinez and McBride
1998). Thus, the impact of contaminants in tropical soils with high Fe
hydroxide concentrations, such as Oxisols, can be particularly distinctive,
necessitating a better understanding of the processes controlling heavy-metal
interactions in these soils. Although many studies have evaluated heavy-
metal behavior in temperate soils (Shuman 1985, 1986; Candelaria and
Chang 1997), relatively few experiments have been conducted on tropical
soils (Naidu 1997; Matos et al. 2001; Silveira, Alleoni, and Chang 2006).
In natural aqueous systems, many different organic and inorganic ligands
are present in solution (Sposito 1983). The soil solution is influenced by
reactions that take place at the solid–solution interface, and the concentrations
of metals in the aqueous phase can be used as potential indicators of metal
uptake by plants. Metals can exist as free ions (hydrated) or interact with
other ions or molecules forming outer-sphere complexes (ion pairs) or
inner-sphere complexes. Free ions are considered the most readily available
chemical specie of metal in the soil solution. On the other hand, the soil
solution can contain 100–200 different soluble complexes (Sposito 1994),
and the presence of ligands can improve the complexation and solubility of
metals. The stability of complexes is markedly affected by environmental con-
ditions, and pH has been identified as the most important factor regulating
metal chemical speciation in biosolid-amended soils (Obrador et al. 1997).
The main objectives of this study were 1) to determine the solid-phase
fractionation and the chemical speciation in solution of (Zn), (Cu), and (Cd)
in biosolid-amended Oxisols and 2) to evaluate the effects of pH and ionic
strength on metal-associated forms in the soil and the biosolids. An exper-
imental chamber system allowed the solutions and separate solid phases
(soils and biosolids) to be characterized independently after the system
reached the equilibrium.
MATERIALS AND METHODS
Surface samples (0–0.2 m) of a Typic Eutrorthox (TE) and a Typic
Haplorthox (TH) soil were collected from intensively cultivated sugarcane
Forms and Speciation in Biosolid-Amended Oxisols 853
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fields in the state of Sao Paulo, Brazil. Selected chemical and physical
properties of the soils are given in Table 1.
Biosolids, anaerobically digested and air dried, were collected in 1974
from Metropolitan Sanitary District of Greater Chicago, IL (MSDGC).
After, collection, dried biosolids were stored in a cold room at 48C until
used. This biosolid, sample was chosen because it was contaminated by
heavy metals, particularly Cd, due to industrial-waste discharges (Table 2).
Although Zn and Cu concentrations in this biosolid do not exceed the ceiling
limits (7500 mg kg21 and 4300 mg kg21 for Zn and Cu, respectively)
(USEPA 1994), Cd concentration was nearly twice (152 mg kg21) the
ceiling concentration for this metal (85 mg kg21). Accordingly, this material
could not be land applied. This biosolid sample represents a distinctive
material and does not reflect the characteristics of the biosolids currently
produced in the United States. Limiting the discharge of industrial wastes
has significantly decreased heavy-metal concentrations in biosolids produced
by the MSDGC. For instance, biosolid samples collected in 2001 from
Table 1. Selected chemical and physical soil attributes
Soil TE TH
pH, H2O 5.0 4.0
C (g kg21) 20 8.3
Ca (mmolc dm23) 40 0.4
Mg (mmolc dm23) 17 0.1
K (mmolc dm23) 5.8 0.02
Na (mmolc dm23) 0.1 0
Al (mmolc dm23) 0 1.43
CEC (mmolc dm23) 62.9 2.0
Sand (g kg21) 170 830
Silt (g kg21) 230 50
Clay (g kg21) 600 120
Fet (g kg21) 220 9
Note: TE, Typic Eutrorthox; TH, Typic Haplorthox;
Fet, total Fe concentration determined after sulfuric
acid digestion.
Table 2. Soil and biosolids total Zn, Cu, and Cd concentrations
Soil/biosolids Zn (mg kg21) Cu (mg kg21) Cd (mg kg21)
TE soil 184 257 0.22
TH soil 13 13 0.15
Biosolids 4319 1263 152
Note: TE, Typic Eutrorthox; TH, Typic Haplorthox.
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the MSDGC exhibited 74, 56, and 96% less Zn, Cu, and Cd, respectively, than
in the samples collected in 1974. Despite its uniqueness, the heavily contami-
nated biosolids was used herein to simplify estimates of the rates at which Zn,
Cu, and Cd are released from the biosolids and reacted with the soil and
solution phases.
For total elemental analyses, subsamples of both the biosolids and soil
samples were oven dried, grounded in an agate mortar, and passed through
a 100-mesh sieve. Approximately 0.250 g of soil and 0.100 g of biosolids
then were digested in triplicate according to USEPA 3052 protocol (USEPA
1996). For the biosolids samples, 1 mL of H2O2 30% was added to increase
the solubilization of the organic fraction. Certified soil (NIST 2709, San
Joaquin soil) and biosolid (NIST 2781, domestic sludge) samples were
digested using the same protocol for quality assurance of the analyses. Recov-
eries ranged from 90 to 110% (data not shown). Stock solutions of all reagents
(analytical grade) were prepared using deionized, distilled water. The
glassware was soaked overnight in 2 M nitric acid (HNO3) and rinsed with
deionized, distilled water prior to use.
Experiment Design
Soil samples were equilibrated with biosolids in a dual-chamber diffusion
apparatus (DCDA), modified from De Pinto (1982). The DCDA was
initially used to characterize P release from municipal wastewater particulates
(DePinto et al. 1981, Young et al. 1982) and from Great Lakes tributary
suspended sediments (DePinto, Young, and Martin 1981) and to investigate
the solution-phase speciation and solid-phase distribution of heavy metals in
biosolid-amended soils (Candelaria and Chang 1997; Berton, Chang, and
Page 1999). The apparatus consists of two polycarbonate chambers of
300 cm3 of volume, separated by a 0.45-mm membrane (Figure 1). The
design of the DCDA allows the two solid phases (soil and biosolids) to
react with each other only through the solution phase via diffusion across
Figure 1. Schematic representation of the dual-chamber diffusion apparatus
(DCDA).
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the membrane. In this way, the solid phases remain separated in each side of
the chamber, and after the equilibrium is reached, the chemical changes in one
solid phase can be easily characterized without disturbance or contamination
by the other solid phase (DePinto 1982). To maximize the diffusion rate across
the membrane, the magnets originally used by DePinto (1982) were replaced
by a stirring apparatus, which allowed more uniform mixing of the solutions.
The kinetic processes that take place in the soil and biosolid phases when
incubated in the DCDA can be functionally divided in three phases (Figure 2):
rate 1) the rate of soluble metals released from the biosolids, rate 2) the rate of
diffusion across the membrane, and rate 3) the rate of adsorption of metals by
the soil. According to Candelaria and Chang (1997), the heavy metal sorption
step (rate 3) is relatively fast, so the reaction rate in the DCDA should be
limited by either the diffusion of metals across the membrane (rate 2) or the
desorption of metals from the solid phase (rate 1). Available metal released
by the biosolids in the DCDA is expected to diffuse across the membrane
and become rapidly immobilized by the soil. By performing a mass balance
on the system, heavy metal distribution and dynamics can be examined.
Two grams of soil or biosolids (dry wt. eq.) were placed in each side of the
chamber. The large biosolids–soil ratio (1:1) was chosen primarily for
simplicity to characterize overall changes in heavy-metal distribution and
was not intended to mimic recommended biosolids rates for field application.
The DCDA was filled with 500 mL of a background solution: calcium nitrate
[Ca(NO3)2; 0.005 M]þ calcium sulfate (CaSO4; 0.003 M), which approxi-
mated the ionic strength and composition of saturation water extracts of
both Oxisols (data not shown). The ionic strength was calculated based on
the electrical conductivity of the solution according to the Debey–Huckel
equation (Lindsay 1979). Calcium, nitrate, and sulfate were chosen because
these ions dominated the soil solutions of the control soils and because of
the lower tendency of nitrate and sulfate to form inner-sphere complexes
with colloid surfaces (specific adsorption) (Sposito 1989).
The pH of the background solution was adjusted to range from 4 to 7
using either 0.15 M HNO3 or calcium carbonate (CaCO3). Additional sets
Figure 2. Kinetic rates that determine the changes in metal distribution in the solid/solution phases.
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of samples were equilibrated in DCDA and contained a background solution
with an ionic strength ten times greater than the control soil solution. One
DCDA served as a control and contained only the background solution and
2 g of soil. One drop of toluene was added to each chamber to control
microbial growth. The suspension was mixed constantly, and pH and electrical
conductivity (EC) were measured daily until equilibrium was established (�5
to 7 d). Equilibrium was judged to have been obtained when solutions in both
sides of the chambers exhibited the same pH and EC. Additionally, after 5 to 7
days of reaction, a subsample of the solution from each side of the chambers
(soil and biosolids) was analyzed for Zn, Cu, and Cd. At the end of the
experiment, both sides of the chamber exhibited the same heavy-metal
concentrations.
Following equilibration, the suspensions were centrifuged, and the super-
natant was filtered through a 0.45-mmmembrane filter. The solid phases (soils
and biosolids) were recovered and freeze dried before analysis. An aliquot of
the filtered solution phase was acidified with concentrated HNO3 and stored in
a cold room at 48C until analysis. The concentrations of iron, manganese, zinc,
calcium, Cu2þ, potassium, magnesium, sodium, nickel, cadmium, lead, and
aluminum in solutions were quantified by atomic absorption spectropho-
tometer or by coupled plasma atomic emission spectrometry (ICP/AES).The concentration of anions [sulfate (SO4
22), chloride (Cl2), nitrate (NO32),
and phosphate (PO432)] were measured in non-acidified solutions by a
Dionex AS-11 ion chromatograph (Sunnyvale, CA). Dissolved organic
carbon was determined using a Shimadzu TOC analyzer (TOC 5050,
Kyoto, Japan). Certified water standards (NIST 1640, trace elements in
natural water) were used to ensure the quality assurance of the analyses.
Recoveries ranged from 90 to 110% (data not shown).
Chemical speciation of Zn, Cu, and Cd in solution was calculated using
the PC-GEOCHEM model (Parker, Norvell, and Chaney 1995). The contri-
bution of organic ligands was estimated based on the DOC concentration
using the mixture model (Mattigod and Sposito 1979; Sposito et al. 1982).
In this model, the organic carbon concentration is used to estimate the
approximate quantitative distribution of various organic acids. One
advantage of the mixture model is that organic acids that complex metals
can be characterized over a wide range of pH and ionic strength (Mattigod
and Sposito 1979).
Soils and biosolids were analyzed for total Zn, Cu, and Cd (EPA 3052)
(USEPA 1996). Mass balance of metals in solid and solution phases was cal-
culated, and the recoveries ranged from 95 to 109% (data not shown). Heavy-
metal sequential fractionation was assessed using the method proposed by
Silveira et al., (2006), adapted for soils with high Fe oxide concentrations
(Table 3). Metals extracted by this procedure were operationally defined as
exchangeable (Ex), carbonate and surface adsorbed (Surf. Ox.), organic
bound (OM), Mn oxide associated (Mn Ox.), amorphous (Fe Ox.) and crystal-
line Fe (Fe Cryst.) oxides sorbed, and residual fraction (Res.).
Forms and Speciation in Biosolid-Amended Oxisols 857
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RESULTS AND DISCUSSION
Zn, Cu, and Cu Distributions between the Solid and the Solution
Phases
There was good agreement between the total Zn, Cu, and Cd concentrations of
the soils and biosolids and the sum of fractions assessed by the sequential frac-
tionation protocol (mass balance�90–110%, data not shown). The agreement
suggests that the scheme efficiently removed the metals associated with the
operationally defined fractions.
The majority of Zn, Cu, and Cd remained in the biosolids after equili-
brium with both soils at pH � 5 (Figure 3). The solution concentrations of
the metals increased with decreasing pH, attributed to greater metal solid-
phase dissolution rates of the biosolids at lower pH values.
There was no variation in the biosolids metal solubility when pH was
reduced from 7 to 6; however, the amount of Cu, Zn, and Cd released from
the biosolids increased when pH was decreased to 5 and 4. The percentage
of metals bound to the biosolids decreased as pH was reduced from 5 to 4:
from 87% to 72% for Cu, from 57 to 22% for Zn, and from 65 to 30% for
Cd (Figure 3). However, the amount of metal adsorbed by the soil did not
Table 3. Heavy-metal sequential extraction (1 g of soil or biosolids) used
Fraction Abbreviation Reagent
Extraction time/temperature
Exchangeable Ex 15 mL 0.1 M Sr(NO3)2 2 h, 258CSurface oxides/carbonate
Surf Ox 15 mL 1 M NaOAc (pH 5) 5 h, 258C
Organic matter OM 5 mL 0.71 M (pH 8.5) 30 min, 908CMn oxides Mn Ox 50 mL NH2OH . HCl
0.05 M (pH 2)
30 min, 258C
Poor-crystalline
Fe oxides
Fe Ox 40 mL 0.2 M ammonium
oxalateþ 0.2 M oxalic
acid (pH 3)
2 h, Extraction in
the dark
Crystalline Fe
oxides
Fe Cryst 250 mL 6 M HCl 24 h, 258C
Residual Res HNO3þHFþH2O Microwave
digestion (EPA
3052)
Notes: Samples were centrifuged at 1225 G for 10 min, and the suspension was
filtered through Whatmann No. 42. Between each step, samples were washed with
5 mL of 0.1 M NaCl, centrifuged, and filtered to avoid readsorption of metals by the
solid phase. All extracts were acidified, using concentrated HNO3, except the Fe Ox
fraction, in which 1 drop of toluene was added. Blanks were used to assess possible
contamination.
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consistently increase as metal concentrations increased in the soil solution.
Overall, the metal sorption of the TE soil was greater than the TH soil,
possibly due to the greater clay and Fe- and Mn-oxide contents of the TE
soil (Table 1).
The pH effect on metal solubility varied with the studied metal species.
Biosolid-Cu extractability was less affected by pH than biosolids-Zn or -Cd
extractabilities. At pH 4, the quantity of metal bound to the biosolids at
Figure 3. Percentage of total Zn, Cu, and Cd in the biosolids after the equilibrium
with soils in the DCDA. Bars indicate one standard error (n ¼ 4) (IS ¼ pH 6, but 10
times ionic strength).
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equilibrium varied from 63 to 81% for Cu to only 22% and 30% for Zn and Cd,
respectively. Clearly, biosolids exhibited greater binding affinity for Cu than
for Zn and Cd.
Increasing the ionic strength of the background solution ten-fold (IS
treatment) had no effect on metal distributions between the solid and liquid
phases. Thus, increasing the ionic strength (at pH 6) did not change the metal
release from the biosolids or the metal concentrations in the soil solution.
Solid-Phase Fractionation of Zn
In the control TE soil samples, approximately 93% of the total Zn was associ-
ated with the Fe-cryst and Res fractions (Figure 4). In the TH soil, Zn was not
detected in these fractions; rather, the Fe-ox and Mn-ox fractions were the most
important for Zn retention (Figure 4). This differential pattern of Zn distribution
can be attributed to differences in the soils mineralogy and clay contents
Figure 4. Distribution of Zn in soil and biosolids samples (TE ¼ Typic Eutrorthox,
TH ¼ Typic Haplorthox, IS ¼ ionic strength).
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(Kalbasi and Racz 1978). Similar results were reported by Abd-Elfattah and
Wada (1981) and Shuman (1979), who found Zn to be dominantly associated
with clay minerals and Fe and Al hydroxides in temperate soils.
Although the majority of Zn was found in the Fe-cryst and Res fractions,
the Surf. Ox. and Mn-Ox fractions contained more Zn after the soils were
equilibrated with biosolids than in the control soil samples. The increase in
the Zn-Surf. Ox. in both soils was likely due to metal associations with the
carbonate fraction, which became more important at higher pH values.
Decreases in pH increased Zn association with the exchangeable fraction.
The percent of exchangeable Zn found in the TE soil samples agree with
the results obtained by Sims (1986), who found 1 to 53% of the total Zn in
the exchangeable form, in the pH range from 4.1 to 7.5. Exchangeable Zn
was greater in the TH soil samples, ranging from 3 to 82% of the total Zn.
The effect of pH reduction on the increase of exchangeable forms was more
evident in the TH soil than in the TE soil samples. Possibly, this reflects the
major contribution of Fe oxides and presence of kaolinite in the TE soil and
the greater affinity of Zn for less labile chemical forms.
Overall, Zn distribution among the various fractions for IS treatment was
similar to that found in the soil samples incubated at pH 6.0; however, the
absolute amount of Zn adsorbed by both soils was reduced at greater ionic
strength. The increase in Ca concentration, due to the increase in the
solution ionic strength, most likely resulted in competition with Zn for adsorp-
tion sites in the soil. Shuman (1986) found that increasing the ionic strength of
the solution decreased Zn adsorption. Adsorption isotherms had similar shapes
at ionic strengths ranging from 0.005 to 0.1 mol L21, but Zn adsorption
increased approximately tenfold at corresponding lower ionic strength
treatment isotherm (0.005). In the IS treatment, Zn content associated with
the TH Surf. Ox. and Mn-ox fractions slightly increased. Exchangeable Zn
decreased in both soils in the IS treatment, compared to the pH 6 treatment.
In the TH soil, the percent exchangeable Zn decreased from 26% in the pH
6 treatment to 3% in the IS treatment.
For the biosolids, reducing pH decreased the amount of Zn associated
with the Surf. Ox. fraction. This was likely due to the dissolution of carbonates
as pH decreased and subsequent sorption in the exchangeable-Zn fraction. No
exchangeable-Zn was observed at pH 7, whereas at pH 4, the values varied
between 18 and 24% of total Zn.
Solid-Phase Fractionation of Cu
Copper in the control soils was dominantly present in nonsoluble forms
(Figure 5). Although soil Cu concentrations increased after the biosolids
incubation, the metal distributions among the fractions were similar to
those in the control soil samples. The pH effect on the amount of Cu in
the soil solution and/or on soil adsorption and distribution was less
marked than observed for Zn. Jeffery and Uren (1983) found no marked
Forms and Speciation in Biosolid-Amended Oxisols 861
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dependence of soil-available Cu on soil pH, except for Cu-contaminated
areas. Similar results were reported by Sims (1986), who found no signifi-
cant change in soil Cu distribution in the pH range from 4.1 to 7.5. This
could be explained by the fact that Cu is differentially adsorbed by the
soil in relation to Zn and Cd. McBride and Blasiak (1979) suggested that
soil Fe and Al hydroxides first adsorbed Cu and second Zn because of the
difference in the pH50 (the pH at which 50% of maximum metal adsorption
occurs). Thus, Cu is typically more strongly associated with Fe and Al
hydroxides than Zn and Cd.
On average, 79% of the total Cu was associated with the Res fraction in
the TE soil samples, whereas most Cu was associated with the amorphous Fe
and Mn oxides and OM fractions in the TH soil (Figure 5). McLaren and
Figure 5. Distribution of Cu in soil and biosolids samples (TE ¼ Typic Eutrorthox,
TH ¼ Typic Haplorthox, IS ¼ ionic strength).
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Crawford (1973), studying surface samples of 24 soils with contrasting attri-
butes, found most (�77%) of the total Cu bound to clay minerals and 35 %
associated with the Fe-ox fraction. Abd-Elfattah and Wada (1981) also
reported that amorphous Fe oxides were the most selective components in
the Cu adsorption.
In both biosolid-amended and control samples, about 1 to 3% of the total
Cu in the TE soil and 12 to 20% in the TH soil was associated with the OM
fraction. These percentages are less than those reported by Sims (1986),
who found 6 to 73% of the total Cu in the OM fraction. However, soils
studied by Sims (1986) had greater OM concentrations than the Oxisols
studied here. Further, the sequential extraction method used by Sims (1986)
did not account for the surface oxide fraction, which could have led to over-
estimation of the quantity of Cu associated with the OM fraction. Contrary to
the results obtained by Sposito, Lund, and Chang (1982), no increase in the
OM fraction Cu concentration was observed in the biosolid-treated soil
samples compared to the control soil samples. However, differences in the
sequential fractionation procedures used make comparison of the results
difficult. Sposito, Lund, and Chang (1982) extracted the OM fraction immedi-
ately after the exchangeable/sorbed fraction and subsequently the carbonate
fraction. In this study, metals associated with the carbonate fraction were
assessed before the organic fraction. The IS treatment did not affect Cu distri-
bution in soils and biosolids.
Manganese and amorphous Fe oxides, and OM, were the main fractions
responsible for the Cu retention in biosolids. As pH decreased, the percentages
of metal associated to the exchangeable fraction increased slightly (from 0 to
3%), and the percentage in the Surf. Ox. fraction decreased (from 12 to 6%).
Petruzzelli et al. (1994) reported similar percentages of exchangeable Cu.
According to these authors, although exchangeable fraction may represent a
lower percentage of total Cu than the other fractions, small changes in pH
can affect Cu bioavailability.
Solid-Phase Fractionation of Cd
Cadmium distribution in soils was greatly affected by the biosolids incubation.
Compared to control samples, Cd concentration was 52-fold greater in the TE
soil and 27-fold greater in the TH soil after the biosolids treatment (Figure 6).
Compared to Cu and Zn, Cd in the biosolids was more readily available and
easily dissolved in the solution. Possibly, the mechanisms of Cd retention in the
biosolid matrix were more susceptible to the effect of pH changes (Figure 6).
The majority of Cd (�76% of total Cd) was associated with the Surf. Ox.
and OM fractions. Because of the biosolids high pH, the Surf. Ox. fraction
may contain carbonates, which contribute to Cd retention in the solid phase
(Sposito 1983). As pH decreases, the carbonates are solubilized, and metals
associated with this fraction are either redistributed into other solid phases or
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solubilized to the solution. At pH 4, 19 to 26% of total Cd was associated with
the exchangeable fraction compared to 2.6 to 3.4% at pH 6.
The increased Cd dissolution from biosolids was not followed by a pro-
portional increase in Cd soil adsorption. In the TH soil, for all pH values,
about 2% of the total Cd was adsorbed, independent of the metal concentration
in solution. In the TE soil, this value ranged from 5 to 8% of total Cd. Possibly
the ability of soils to retain Cd was limited, because this metal has less
tendency to be adsorbed compared to Zn and Cu (Alloway 1990). Increasing
solution ionic strength did not affect Cd dissolution from biosolids; however,
the distribution of this metal among the solid phases shifted. Cadmium associ-
ated with OM was increased, followed by a decrease in Ex-Cd for both soils.
In both control soils, Cd was found mainly in the Surf. Ox. and OM
fractions. When pH was decreased, Cd was found primarily associated with
exchangeable fraction, which supports the results obtained by Kuo,
Heilman, and Baker (1983).
Figure 6. Distribution of Cd in soil and biosolids samples (TE ¼ Typic Eutrorthox,
TH ¼ Typic Haplorthox, IS ¼ ionic strength).
M. L. Silveira et al.864
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Table 4. Total concentration and chemical speciation of Zn, Cu, and Cd in solution
Treatment
Total Zn
(mg L21)
Free Zn
(% of total)
Total Cu
(mg L21)
Free Cu
(% of total)
Total Cd
(mg L21)
Free Cd
(% of total)
TEþ Biosolids
pH 7 0.06 80.5 nd 0 0.01 88.0
pH 6 0.5 92.2 0.01 82.7 0.03 90.2
IS 1.0 87.9 0.01 80.8 0.07 86.3
pH 5 3.7 96.7 0.1 95.2 0.3 95.5
pH 4 13.4 98.3 1.1 97.8 0.5 97.8
THþ Biosolids
pH 7 0.1 81.5 nd 2.3 0.02 88.1
pH 6 0.8 92.6 0.05 26.6 0.06 90.8
IS 1.3 87.4 0.02 81.0 0.08 85.4
pH 5 8.0 97.6 0.2 96.5 0.3 96.9
pH 4 12.7 98 0.9 97.4 0.5 97.4
Note: TE, Typic Eutrorthox; TH, Typic Haplorthox; nd, not detected (below the detection limit).
Form
sandSpecia
tionin
Biosolid
-Amended
Oxiso
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2014
The Oxisols in the present study exhibited contrasting attributes, but Cd
distribution was similar in both soils. In control samples of TH and TE, Cd
was associated with the Surf. Ox. and OM fractions (82% of total Cd)
(Figure 6), both of which are potentially bioavailable fractions. The Surf. Ox
and OM fractions, and consequently the soil Cd distribution, can be modified
with soil acidification and/or OM mineralization. Thus, the mechanisms of
Cd retention in these soils are chemically unstable, and Cd can be eventually
released to the solution (Kuo, Heilman, and Baker 1983). Biosolid addition to
the soil increased Ex-Cd. The exchangeable fraction was 4% of total Cd in
the control soil samples but increased to 11% and 80% of total Cd, at pH 7
and pH 4, respectively, after incubation with biosolids.
Metal Chemical Speciation in the Solution Phase
The distribution of soluble species, according to GEOCHEM-PC, showed that
between 81 to 98% of Zn, Cu, and Cd in solution were presented as free ions
(Table 4). There was a strong correlation between pH and the amount of ionic
species for both soils. Free-ion species percentages increased with increasing
acidity. Under high pH conditions, soil metal adsorption ability and the avail-
ability of complexing agents are increased (Salam and Helmke 1998),
favoring the complexes’ formation. There was a higher percentage of free-
Cd and -Zn ions in solution than Cu ions at the same pH values. The lower
free-Cu concentrations in solution compared to the other metals was due to
both the strong adsorption of Cu by the soil colloid surfaces (Cavallaro and
McBride 1980) as well as its higher affinity with complexing agents present
in the solution. Generally, increasing solution ionic strength favored the
formation of SO422 complexes with Zn, Cu, and Cd and decreasing concen-
trations of the free ionic species.
CONCLUSIONS
Chemical forms and total Zn, Cu, and Cd concentrations in the biosolids had a
marked effect on subsequent metal distribution in the soils. The pH markedly
affected Zn, Cd and, to a lesser extent, Cu distribution in soils and biosolids.
Increasing the solution ionic strength had minimal effects on the metal distri-
bution in soils or biosolids. Soluble-metal concentrations increased with
decreasing soil pH, due to the higher solubilization of the various metal
solid phases in the biosolids. At low pH values, the majority of the metals
in the solid phase were associated with exchangeable fractions, and free
ions were the predominant chemical specie in solution. Although Oxisols
exhibit distinctive chemical characteristics, Zn, Cu, and Cd partitioning was
similar to that in temperate soils amended with biosolids. The solubilities/exchangeabilities of the heavy metals studied were primarily dictated by the
characteristics of the biosolids and the pH of the soil solution.
M. L. Silveira et al.866
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ACKNOWLEDGMENT
Maria L. Silveira thanks the Brazilian Federal Agency CAPES for granting a
scholarship and supporting this study. The authors thank the staff of the
Environmental Sciences Department, University of California, Riverside,
for assistance with chemical analyses.
REFERENCES
Abd-Elfattah, A. and Wada, K. (1981) Adsorption of lead, copper, zinc, cobalt, andcadmium by soils that differ in cation exchange materials. Journal of Soil Science,32: 271–284.
Alloway, B.J. (1990) Heavy Metals in Soils; Wiley: New York.Appel, C. and Ma, L. (2002) Concentration, pH, and surface charge effects on cadmiumand lead sorption in three tropical soils. Journal of Environmental Quality, 31:581–589.
Berton, R.S., Chang, A.C., and Page, A.L. (1999) Metal ion activities in long-termbiosolids amended soils. Proceeding of International Conference on theBiogeochemistry of Trace Elements. International Society of Trace Element Biogeo-chemistry. Vienna, Austria, 292–293.
Candelaria, L.M. and Chang, A.C. (1997) Cadmium activities, solution speciation, andsolid phase distribution of Cd in cadmium nitrate and sewage sludge-treated soilsystems. Soil Science, 162: 722–732.
Cavallaro, N. and McBride, M.B. (1980) Activities of Cu and Cd in soil solutions asaffected by pH. Soil Science Society of American Journal, 44: 729–732.
Covelo, E.F., Couce, M.L.A., and Vega, F.A. (2004) Competitive adsorption ofcadmium, chromium, copper, nickel, lead, and zinc by humic umbrisols. Communi-cations in Soil Science and Plant Analysis, 35: 2709–2729.
DePinto, J.V. (1980) Phosphorus Removal in Lower Great Lakes Municipal TreatmentPlants; EPA-600/2-80-117, MERL: Cincinnati, OH.
DePinto, J.V., Young, T.C., and Martin, S.C. (1981) Alga-available phosphorus insuspended sediments from lower Great Lakes tributaries. Journal of Great LakesResearch, 7: 311–325.
DePinto, J.V. (1982) An experimental apparatus for evaluating kinetics of availablephosphorus release from aquatic particulates. Water Research, Elsevier Periodical16: 1065–1070.
Gerritse, G.E., van Driel, W., Smilde, K.W., and van Luit, B. (1983) Uptake of heavymetals by crops in relation to their concentration in soil solution. Plant and Soil, 75:393–405.
Jeffery, J.J. and Uren, N.C. (1983) Copper and zinc species in the solution and theeffects of soil pH. Australian Journal of Soil Research, 21: 479–488.
Kalbasi, M. and Racz, G.J. (1978) Association of zinc with oxides of iron andaluminum in some Manitoba soils. Canadian Journal of Soil Science, 58: 61–68.
Krishnamurti, G.S.R. and Naidu, R. (2002) Solid-solution speciation and phytoavail-ability of copper and zinc in soils. Environmental Science and Technology, 36:2645–2651.
Kuo, S., Heilman, P.E., and Baker, A.S. (1983) Distribution and forms of copper, zinc,cadmium, iron, and manganese in soils near a copper smelter. Soil Science, 135:101–109.
Forms and Speciation in Biosolid-Amended Oxisols 867
Dow
nloa
ded
by [
Uni
vers
ity o
f A
lber
ta]
at 1
8:40
26
Nov
embe
r 20
14
Lindsay, W.L. (1979) Chemical Equilibria in Soils; Wiley: New York.
Martinez, C.E. and McBride, M.B. (1998) Solubility of Cd, Cu, Pb and Zn in aged
coprecipitates with amorphous iron hydroxides. Environmental Science and Tech-
nology, 32: 743–748.
Matos, A.T., Fontes, M.P.F., Costa, L.M., and Martinez, M.A. (2001) Mobility of
heavy metals as related to soil chemical and mineralogical characteristics of
Brazilian soils. Environmental Pollution, 111: 429–435.
Mattigod, S.V. and Page, A.L. (1983) Assessment of metal pollution in soil. In Applied
Environmental Geochemistry; Thornton, I. (ed.); Academic Press: New York,
355–394.
Mattigod, S.V. and Sposito, G. (1979) Chemical modeling of trace metal equilibria in
contaminated soil solutions using the computer program GEOCHEM. In Chemical
Modeling in Aqueous Systems; Jenne, E.A. (ed.); American Chemical Society,
Symposium Ser. 9. ACS: Washington DC, 837–856.
McBride, M.B. and Blasiak, J.J. (1979) Zinc and copper solubility as a function of pH
in an acid soil. Soil Science Society of America Journal, 43: 866–870.
McLaren, R.G. and Crawford, D.V. (1973) Studies on soil copper, 1: The fractionation
of copper in soils. Journal of Soil Science, 24: 172–181.
Naidu, R., Kookana, R.S., Summer, M.E., Harter, R.D., and Tiller, K.G. (1997)
Cadmium sorption and transport in variable charge soils: A review. Journal of
Environmental Quality, 26: 602–617.
Obrador, A., Rico, M.I., Mingot, J.I., and Alvarez, J.M. (1997) Metal mobility and
potential bioavailability in organic matter–rich soil-sludge mixtures: Effect of soil
type and contact time. Science of the Total Environment, 206: 117–126.
Parker, D.R., Norvell, W.A., and Chaney, R.L. (1995) GEOCHEM-PC: A chemical
speciation program for IBM and compatible personal computers. In Soil Chemical
Equilibria and Reaction Models; Lippert, R.J. (ed.); Soil Science Society
America, Special Publication: Madison, Wisc.; Vol. 42, 253–269.
Petruzzelli, G., Ottaviani, M., Lubrano, L., and Veschetti, E. (1994) Characterization of
heavy metal mobile species in sewage sludge for agricultural utilization. Agrochi-
mica, 38: 277–284.
Salam, A.K. and Helmke, P.A. (1998) The pH dependence of free ionic activities and
total dissolved concentrations of copper and cadmium in soil solution. Geoderma,
83: 281–291.
Shuman, L.M. (1979) Zinc, manganese and copper in soil fractions. Soil Science, 127:
10–17.
Shuman, L.M. (1985) Fractionation method for soil microelements. Soil Science, 140:
11–22.
Shuman, L.M. (1986) Effect of liming on the distribution of manganese copper, iron,
and zinc among soil fractions. Soil Science Society of America Journal, 50:
1236–1240.
Shuman, L.M. (1991) Chemical forms of micronutrients in soils. In Micronutrients in
Agriculture, 2nd edn.; Mickelson, S.H. (ed.); Soil Science Society of America:
Madison, Wisc., 113–183.
Sims, J.T. (1986) Soil pH effects on the distribution and plant availability of
manganese, copper and zinc. Soil Science Society of America Journal, 50: 367–373.
Silveira, M.L., Alleoni, L.R.F., O’Connor, G.A., and Chang, A.C. (2006) Heavy metal
sequential extraction methods—A modification for tropical soils. Chemosphere, 64:
1929–1938.
M. L. Silveira et al.868
Dow
nloa
ded
by [
Uni
vers
ity o
f A
lber
ta]
at 1
8:40
26
Nov
embe
r 20
14
Sposito, G. (1983) The chemical forms of trace metals in soils. In Applied Environ-mental Geochemistry; Thornton, I. (ed.); Geology series, Academic Press:New York, 123–170.
Sposito, G. (1989) The Chemistry of Soils; Oxford University Press: New York.Sposito, G. (1994) Chemical Equilibria and Kinetics in Soils; Oxford University Press:New York.
Sposito, G., Bingham, F.T., Yadav, S.S., and Unouye, C.A. (1982) Trace metal com-plexation by fulvic acid extracted from sewage sludge, II: Development ofchemical models. Soil Science Society of America Journal, 46: 51–56.
Sposito, G., Lund, L.J., and Chang, A.C. (1982) Trace metal chemistry in arid-zonefield soils amended with sewage sludge: I. Fractionation of Ni, Cu, Zn, Cd, andPb in solid phases. Soil Science Society of America Journal, 46: 260–264.
Tsadilas, C.D., Matsi, T., Barbayiannis, N., and Dimoyiannis, D. (1995) Influence ofsewage sludge application on soil properties and on the distribution and availabilityof heavy metal fractions. Communications in Soil Science and Plant Analysis, 26:2603–2619.
U.S. Environmental Protection Agency. (1994) A Plain English Guide to the EPA Part503 Biosolids Rule; Environmental Protection Agency Office of WastewaterManagement: Washington, DC.
U.S. Environmental Protection Agency. (1996)Method 3052: Microwave Assisted AcidDigestion of Siliceous and Organically Based Matrices (compact disc); Environ-mental Protection Agency Office of Wastewater Management: Washington, DC.
Young, T.C., DePinto, J.V., Flint, S.E., Switzenbaum, M.S., and Edzwald, J.K. (1982)Algal availability of phosphorus in municipal wastewaters. Journal of WaterPollution Control Federation, 54: 1505–1516.
Forms and Speciation in Biosolid-Amended Oxisols 869
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