Inorganic and Organic Constituents and Contaminants of Biosolids: Implications for Land Application

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From: R. J. Haynes, G. Murtaza, and R. Naidu, Inorganic and Organic Constituents
and Contaminants of Biosolids: Implications for Land Application. In Donald L.
Sparks, editor: Advances in Agronomy, Vol. 104, Burlington: Academic Press, 2009,
pp. 165-267. ISBN: 978-0-12-374820-1
© Copyright 2009 Elsevier Inc.
Academic Press.

C H A P T E R F O U R
Inorganic and Organic Constituents
and Contaminants of Biosolids:
Implications for Land Application
R. J. Haynes,* G. Murtaza,†,‡
and R. Naidu§
Contents
1. Introduction 166
2. Sewage Treatment Processes 168
3. Composition of Biosolids 169
3.1. Organic matter 169
3.2. Inorganic components 174
4. Nutrient Content and Release 175
4.1. Nitrogen 175
4.2. Phosphorus 179
4.3. Other nutrients 181
5. Heavy Metal Contaminants 182
5.1. Total concentrations 183
5.2. Extractable fractions 185
5.3. Application to the soil 187
5.4. Plant response and metal uptake 202
5.5. Ingestion by animals 207
6. Organic Contaminants 208
6.1. Organic compounds present 211
6.2. Potential transfer to groundwater, plants, and animals 227
7. Synthesis and Conclusions 234
References 237
Abstract
Large amounts of biosolids are produced as a by-product of municipal waste-
water treatment. They are composed of about 50% organic and 50% inorganic
material. The organic component is partly decomposed and humified material
Advances in Agronomy, Volume 104 # 2009 Elsevier Inc.
ISSN 0065-2113, DOI: 10.1016/S0065-2113(09)04004-8 All rights reserved.
* School of Land, Crop and Food Sciences/CRC CARE, The University of Queensland, St Lucia, Australia
{
Centre for Environmental Risk Assessment and Remediation, Division of Information Technology,
Engineering and the Environment, University of South Australia, Mawson Lakes Campus, South Australia,
Australia
{
Institute of Soil and Environmental Sciences, University of Agriculture, Faisalabad, Pakistan
}
CRC CARE, Salisbury, South Australia, Australia
165
Author’s personal copy

derived from human feces and bacterial biomass while the inorganic component
is derived from materials such as soil, sediment, and inorganic residuals
(e.g., silica). The major contaminants in biosolids are heavy metals (e.g., Cu,
Zn, Cd, Pb, Ni, Cr, and As) plus a range of synthetic organic compounds.
Following land application, biosolids-borne metals are typically immobile in
soils. They can be toxic to soil microflora, small amounts may leach with soluble
organic matter, they can be accumulated in plants and sometimes transferred to
grazing animals (mainly by soil ingestion). Regulations and guidelines for
biosolids applications are still principally based on total metal loadings and in
the future the use of bioavailable metal concentrations in biosolids-treated soils
should be considered. The significance, effects, and fate of biosolids-borne
organic contaminants in soils are not well understood and require further
study. In the majority of cases, neither heavy metal nor organic contaminants
are considered a significant hazard to the soil–plant system. Indeed, land
applications of biosolids can be highly beneficial to crop production since
they supply substantial amounts of N, P, Ca, and Mg and added organic matter
can improve soil physical properties and stimulate soil microbial activity. To
avoid ground/surface water pollution, application rates should be based on the
N need of the crop and potential N mineralization rate of biosolids-N, and the
high P loadings need to be managed.
1. Introduction
Biosolids are derived from the treatment of wastewater (sewage) that is
primarily derived from domestic sources being a combination of human
feces, urine, and graywater (from washing, bathing, and meal preparation).
Sewage also contains discharges from commercial and industrial enterprises
and often some stormwater. As the wastewater is treated, it goes through a
series of processes that reduce the concentrations of organic material that
were originally present. Primary sludge (principally fecal material) results
from settling of solids as they enter the treatment plant. Secondary sludge
originates from the conversion of suspended and soluble organic matter in
sewage into bacterial biomass. The biomass is removed and combined with
the primary sludge to produce material termed sewage sludge. This material
then undergoes treatment (usually anaerobic but sometimes aerobic diges-
tion) to reduce the volume and stabilize the solid organic matter component
as well as to reduce the presence of disease-causing organisms. The final
product is termed biosolids.
The safe disposal of biosolids is a major environmental concern through-
out the world. Disposal alternatives include dumping at sea, incineration,
landfilling, and land application (Epstein, 2003). Land application is
generally seen as the most economical and beneficial way to deal with
166 R. J. Haynes et al.

biosolids (Shammas and Wang, 2007a). Indeed, about 60% of all biosolids
produced in both United States and United Kingdom are land applied
(Pepper et al., 2006). Biosolids contain organic matter and nutrients and
when applied to farmland can improve productivity and reduce the need for
manufactured fertilizer inputs (Singh and Agrawal, 2008). Biosolids have
also been used successfully as a topsoil substitute for landscaping (Wu, 1987)
and to enhance revegetation process on disturbed sites (e.g., mined land and
tailings dumps) (Sopper, 1992). The organic matter acts as a soil condi-
tioner, improving soil physical conditions and stimulating soil microbial
activity while macro- and micronutrients present serve as a source of plant
nutrients. However, there are potential hazards with land application since a
range of contaminants can be present in biosolids including heavy metals,
recalcitrant organic compounds, and pathogens (Hue, 1995; Jenson and
Jepsen, 2005; Mininni and Santori, 1987; Pepper et al., 2006; Singh and
Agrawal, 2008). Their presence greatly influences public perceptions
regarding the safety of land applications.
That an enormous volume of literature has been, and is continuing to be,
published on the nature and content of biosolids and the agronomic and
environmental aspects of land application is testament to the relevance and
importance of the topic. Several workers have reviewed agronomic
and environmental aspects of land application of biosolids (During and
Gath, 2002; Epstein, 2003; Hue, 1995; Singh and Agrawal, 2008) and the
presence of pathogens in biosolids was recently discussed (Pepper et al.,
2006). However, a detailed understanding of the nature and content of
biosolids, and how this develops during sewage treatment, helps greatly in
predicting their effects on the soil and the wider environment. In this
chapter we provide an overview of findings on the nature of inorganic
and organic constituents and contaminants of biosolids in relation to the
impact that land application has on soil properties, crop growth, and the
wider environment.
Biosolids are well characterized materials and the nature and content of
organic and inorganic constituents, their nutrient content, and nutrient
release characteristics are well documented and are reviewed here. Simi-
larly, voluminous literature exists on the fate of contaminant heavy metals
during wastewater treatment and, more particularly, the fate of biosolids-
borne heavy metals in soil following land application. Consequently, an
overview of this information is also presented here. By comparison, research
into organic contaminants in biosolids is in its infancy and the majority of
studies are surveys of the presence and concentrations of various compounds
found in a range of biosolids samples. Current knowledge on the occur-
rence of organic contaminants is therefore reviewed and using the scarce
data that exists, their fate during wastewater treatment and in the soil after
land application of biosolids is discussed.
Inorganic and Organic Constituents and Contaminants of Biosolids 167

2. Sewage Treatment Processes
Prior to treatment, the influent sewage water is screened to remove
large objects and then undergoes grit removal in which heavy inorganic
coarse, sand-like, material is removed by settling. The water is then pumped
to large sedimentation tanks where it undergoes primary treatment. This
involves sedimentation in which most of the suspended solids are removed
as sludge material which sinks to the floor of the tanks. The sludge is then
pumped as a slurry (primary sludge) to storage tanks. The liquid remaining
enters secondary treatment which is designed to degrade the remaining
dissolved and colloidal organic content in the sewage.
During the secondary stage, most of the organic matter remaining in the
waste water is consumed by microbes under aerobic conditions. This is
accomplished by bringing together wastewater, bacteria (and other
microbes), and oxygen and can be achieved by either fixed film or
suspended growth systems. In fixed film methods (e.g., trickling filters and
rotating biological contactors) the microbial biomass grows on a medium
and the sewage passes over its surface. The microorganisms remove and
oxidize the organic material. The most common suspended growth system
is the activated sludge process. Primary-treated sewage combined with
microorganisms is aerated by bubbling O2 through a tank. A biological
floc (composed of saprophytic bacteria and associated protozoa and rotifers)
develops which removes and oxidizes the organic material. The treated
supernatant is runoff and a portion of the settled sludge is returned to the
head of the aeration system to reseed the new sewage entering the tank.
Secondary treatment commonly removes about 60–90% of dissolved and
suspended organic matter. The waste sludge from this process (secondary
sludge) consists predominantly of saprophytic bacterial biomass, some other
microorganisms and adhering microbial by-products. It is removed and
normally mixed with the sludge from the primary treatment process.
The accumulated sludges are then treated before disposal. Treatments
usually include thickening, stabilization, and then dewatering. Thickening
is used to increase the solids content and reduce the volume that needs to be
handled. It increases the solids content of sludge from 1–2% to 4–5% and
can reduce volumes to as low as 20% of unthickened sludge. The most
common stabilization treatments are anaerobic and aerobic digestion. The
sludge is digested to reduce the amount of organic matter and the number of
disease-causing microorganisms present in the solids. In anaerobic digestion,
(Taricska et al., 2007) sludge is passed into a closed container held at either
the mesophilic (e.g., 36
C) or thermophilic range (e.g., 55
C). Bacteria
decompose organic matter in the absence of O2 to produce CO2
and methane (biogas), the latter gas is used as a fuel to heat the digester.

In aerobic digestion, air is continuously pumped into the digester and
bacterial activity breaks down organic matter to produce CO2 and it also
generates heat to kill pathogens (Shammas and Wang, 2007b).
Other lesser used stabilization methods include lime stabilization and
thermal treatment. Lime stabilization involves mixing the sludge with lime
to achieve a pH of 12 or more and maintaining it for 2 h or longer. The
alkaline conditions produced drastically reduces microbial activity and
causes death of many pathogens. Thermal treatment subjects the sludge to
high temperatures (e.g., 150–180
C) and pressures up to 3 mPa in a closed
reaction vessel. This results in rupture of cell walls of microorganisms
present (including pathogens) and causes chemical oxidation of organic
matter.
Following digestion, the treated sludge is often dewatered to reduce the
volume and mass for transport. Belt filter presses, vacuum filtration, or
centrifugation are used to increase the solids content of sludge to 25–45%
whereupon the material takes on the properties of a solid rather than a
liquid. It can also be composted to further reduce volume, produce a more
stabilized product, and reduce the incidence of pathogens (Parr et al., 1978).
Composting usually involves blending dewatered biosolids with a bulking
agent (e.g., bark chips) and composting the product in windrows. Heat is
generated during the intense microbial activity of composting and thermo-
philic temperatures (!55
C) can be reached which cause death of many
pathogenic organisms.
3. Composition of Biosolids
3.1. Organic matter
3.1.1. Nature of organic matter
Biosolids samples are typically made up of 40–70% organic matter (as
measured by loss of mass on ignition). They typically have an organic C
content ranging from 20–50%, a total N content of 2–5%, and a C/N ratio
of about 10–20 (Alonso et al., 2006, 2009; Alvarez et al., 2002; Cai et al.,
2007a; Solis et al., 2002; Sprynskyy et al., 2007; Wang et al., 2005, 2006a).
The organic matter originates principally from human feces (primary sludge)
and bacterial cells (secondary sludge) and has undergone some degree of
decomposition and humification during anaerobic or aerobic digestion.
The organic fraction of biosolids has been identified as a mixture of fats,
proteins, carbohydrates, lignin, amino acids, sugars, celluloses, humic
material, and fatty acids. Live and dead microorganisms constitute a substantial
proportion of the organic material and provide a large surface area for
sorption of lipophilic organic contaminants in the sludge. Because much of
the insoluble inorganic matter settles out during primary sedimentation,

the organic matter content of primary sludge (47–70%) is normally less than
that of secondary sludge (62–82%) (Alvarez et al., 2002; Solis et al., 2002).
The organic matter content of mixed sludge typically declines during
digestion as organic matter is decomposed and lost as CO2 (Alvarez et al.,
2002; Solis et al., 2002). Solis et al. (2002), for example, recorded an organic
matter content of 65% for mixed sludge but a content of only 56% after
anaerobic digestion. There is a further decline (as much as 30–60%) in
organic matter content if the biosolids are composted (Alvarez et al., 2002;
Liu et al., 2007a,b; Solis et al., 2002), although this will not necessarily be the
case if a slowly decomposable organic bulking agent (e.g., shredded bark
chips) is added prior to composting (Nomeda et al., 2008).
Humification is a natural process by which plant and animal residues
decompose in the soil and a dark colored, more or less stable portion of
organic matter remains. The humic materials remaining are high molecular
weight organic molecules made up of a core of phenolic polymers produced
from the products of biological degradation of plant and animal residues and
the synthetic activity of microorganisms (Stevenson, 1994). They exist as
heterogeneous, complex, three-dimensional amorphous structures. The
humic fraction of biosolids differs from that of soils because the former has
undergone a relatively short period of decomposition/transformation by a
technological process rather than a long-term transformation under natural
soil conditions.
Characterization of humic substances is complex and involves a wide
range of techniques including elemental and functional group analyses, gel
filtration chromatography, electrophoresis, pyrolysis, thermochemolysis,
and ultraviolet/visible, infrared, nuclear magnetic resonance (NMR), elec-
tron spin resonance (ESR), and fluorescence spectroscopies (Senesi et al.,
2007). These techniques have shown that in comparison with native soil
humic substances, humic substances from biosolids are characterized by
lower molecular weights, higher contents of S- and N-containing groups,
lower C/N ratios and contents of acidic groups, much lower metal binding
capacities and stability constants, a prevalence of aliphaticity, extended
molecular heterogeneity, and lower degrees of polycondensation and humi-
fication (Amir et al., 2004; Ayuso et al., 1997; Boyd et al., 1980; Leinweber
et al., 1996; Mao et al., 2003; Rowell et al., 2001; Senesi et al., 1991;
Smernik et al., 2003a, 2004; Soler Rovira et al., 2002). Part of the hetero-
geneity of the humic material probably arises because it is derived from two
separate sources (primary and secondary sludge). For example, Smernik et al.
(2003b) showed that organic matter in biosolids consisted of two spatially
and chemically distinct ‘‘domains’’ derived from partially degraded plant
material (i.e., human feces) and bacterial residues, respectively. Results of
a comparative study of the humic substances from anaerobically and
aerobically digested biosolids (Hernandez et al., 1988) showed that the

type of digestion process has little effect on elemental composition or
functional group content.
Composting organic wastes is an established method of obtaining chem-
ical stabilization, biological maturation, and sanitization and involves con-
trolled, aerobic, decomposition of organic waste to form a smaller volume
of relatively stable humus-like material (Senesi and Plaza, 2007). Thus,
composting of sewage sludge results in further decomposition and humifi-
cation and as a result the chemical and physicochemical properties of the
biosolids-derived humic substances more closely approach those of native
soil humic substances (Amir et al., 2004, 2005a; Garcia et al., 1991a;
Jouraiphy et al., 2005; Sanchez-Monedero et al., 2002; Zbytniewski and
Buszewski, 2005). For example, Amir et al. (2004) demonstrated that during
composting there was a steady decrease in C content, a more substantial
decrease in N content, an increase in C/N ratio, and a decrease in aliphatic
compounds which was accompanied by an increase in the relative abun-
dance of aromatic structures. These changes occur because during compost-
ing, oxidative degradation of readily accessible compounds (e.g., aliphatic
side chains of lipidic and N-containing peptide structures) occurs. This leads
to a more oxidized, polycondensed aromatic structure.
Digested biosolids contain a significant portion of water-soluble ‘‘labile’’
organic matter. This fraction often makes up 2–3% of total organic C
content (Garcia et al., 1991b; Zbytniewski and Buszewski, 2005) and con-
sists of sugars, aliphatic organic acids, amino acids, and soluble low molecu-
lar weight polyphenolic humic substances. The amounts of such substances
can sometimes increase during the initial stages of composting (Zbytniewski
and Buszewski, 2005) as more complex organic substances are broken down
and, in addition, organic metabolites are excreted by the decomposer
microbial community. However, over the composting period (usually
50–150 days), there is typically an overall decline in soluble C concentrations
(both absolute concentrations and those as a percentage of total organic C
content) until they account for about 1–2% of organic C (Garcia et al., 1991b;
Zbytniewski and Buszewski, 2005). Indeed, a decline in water-soluble
organic C is often used as an indicator of compost maturity since fresh
compost consists of many easily degradable and water-soluble substances,
whereas mature compost is rich in stable, decomposition-resistant, high
molecular weight, humic substances (Zmora-Nahum et al., 2005).
3.1.2. Application to the soil
Following application of biosolids to soils, there is a rapid phase of decom-
position as the easily decomposable fractions are degraded. This is accom-
panied by a period of intense microbial activity in the sludge-amended soil
(see below). This can lead to a ‘‘priming effect’’ and result in some
concomitant decomposition of native soil organic matter (Terry et al., 1979).

Nevertheless, when biosolids are applied to soils at high rates and/or
repeatedly, there is typically a substantial increase in soil organic matter
content (Gupta et al., 1977; Kladivko and Nelson, 1979; Moffet et al., 2005;
Navas et al., 1998; Rostagno and Sosebee, 2001). The effect is particularly
pronounced on degraded soils with a low initial organic matter content
(Garcia-Orenes et al., 2005). Indeed, using 14C-labeled biosolids, Terry
et al. (1979) showed that a major portion of biosolids-C was resistant to
decomposition in the soil and had a turnover rate in the order of hundreds
of years.
Not only is the soil organic matter content increased, but also the quality
of organic matter is changed. That is, as expected based on the above
discussion, amending soils with biosolids generally causes an increase in
aliphaticity and N, H, and S contents and a decrease of C/N ratios, O and
acidic functional group contents and metal binding capacities of soil humic
materials (Adani and Tambone, 2005; Boyd et al., 1980; Garcia-Gil et al.,
2004; Han and Thompson, 1999; Piccolo et al., 1992; Plaza et al., 2005,
2006). These effects are most evident at high rates of addition of biosolids.
With increasing time after application, the characteristics of the amended
soil humic substances return to those of the unamended soil since the
biosolids-derived humic materials undergo further humification and
become incorporated within the soil humic fraction (Senesi et al., 2007).
Amending soils with composted biosolids, however, has a much lesser effect
on the characteristics of soil humic substances compared to uncomposted
material.
Increases in concentrations of dissolved organic matter in soil solution,
and its downward movement in the soil profile, following biosolids applica-
tions have been noted by a number of workers (Ashworth and Alloway,
2004; Han and Thompson, 1999; Neal and Sposito, 1986; Toribio and
Romanya, 2006). Han and Thompson (1999) also observed that the molec-
ular weight distribution of soluble organic matter in soils shifted to lower
weights (e.g., 14,000 Da) following biosolids applications. The signifi-
cance of dissolved organic matter to the mobility of biosolids-derived heavy
metals is discussed in Section 5.3.5.
The cation exchange capacity (CEC) of the soil is often increased
following land application of biosolids (Clapp et al., 1986; Epstein et al.,
1976; Gaskin et al., 2003; Navas et al., 1998; Udom et al., 2004). This is
attributable to the high CEC of biosolids organic matter conferred by the
many negatively charged functional groups present on humic material. The
extent of the increase will depend on such factors as soil texture, initial soil
organic matter content and CEC, nature of biosolids, and period since last
application. Over time, there will be a subsequent decrease in CEC as the
added biosolids organic matter decomposes (Clapp et al., 1986).
The increase in organic matter content following biosolids application
often results in a concomitant improvement in soil physical properties

(Clapp et al., 1986; Khaleel et al., 1981). There is often an increase in water
stable aggregation (Epstein, 1975; Gupta et al., 1977; Kladivko and Nelson,
1979; Pagliai et al., 1981) due to the binding properties of organic matter
and the associated microflora. Because of increased aggregation, total pore
space is typically increased resulting in measured deceases in bulk density
and increases in total porosity (Garcia-Orenes et al., 2005; Navas et al., 1998;
Rostagno and Sosebee, 2001; Table 1). Because of the increased porosity,
increases in infiltration rate (Table 1) and hydraulic conductivity also tend to
occur (Epstein, 1975; Gupta et al., 1977; Tsadilas et al., 2005) and as a result
there can be decreased runoff and water erosion (Moffet et al., 2005;
Rostagno and Sosebee, 2001). Water-holding capacity often increases at
both field capacity and wilting point (Kladivko and Nelson, 1979; Gupta
et al., 1977; Table 1) but the amount of available water (held between field
capacity and wilting point) is often not greatly affected (Gupta et al., 1977;
Kladivko and Nelson, 1979; Tsadilas et al., 2005).
Addition of an organic substrate to a soil generally results in an increase
in the size and activity of the soil microbial community as well as the
activities of soil enzymes. Such stimulation of microbial activity can occur
following biosolids applications and/or inhibitory effects can occur due to
the presence of heavy metals and other pollutants (see, Section 5.3.6).
Where there is little or no inhibition of microbial activity from pollutants,
substantial increases in microbial activity induced by biosolids applications
have been recorded in both laboratory incubations and field studies. For
example, in a two-month incubation experiment Dar (1996) showed that
biosolids amendment at 0.75% increased soil microbial biomass by 8–28%,
arginine ammonification rate by 8–12%, and dehydrogenase and alkaline
phosphatase enzyme activities by 18–25% and 9–23%, respectively,
compared to unamended soils. Increases in the activities of other soil
Table 1 Effect of annual biosolids applications over a 3-year period on soil organic
matter content and some soil physical properties
Biosolids
rate
(Mg haÀ 1
)
Organic
mattera
content (%)
Bulk
density
(g cmÀ 3
)
Field
capacity
(g gÀ 1
)
Wilting
point
(g gÀ 1
)
Available
water
(g gÀ 1
)
Final
infiltration
rate (cm hÀ 1
)
0 2.57aa
1.41b 27.46a 14.23a 53.13a 1.95a
10 2.86b 1.32a 29.46b 16.01b 53.25a 1.95a
30 3.38c 1.3a 30b 16.51c 53.41a 3.6b
50 3.75d 1.27a 33.85c 18.39d 58.62b 4.05b
a
Numbers in the same column followed by different letters differ significantly at probability level
p 0.05 to the LSD test.
From Tsadilas et al. (2005); copyright Taylor Francis.

enzymes such as urease, amidase, proteinase, b-glucosidase, and arylsulpha-
tase in response to biosolids addition have also been noted in incubation
studies (Gomah et al., 1990; Hattori, 1988; Kizilkaya and Hepsen, 2004;
Topac et al., 2008).
In field experiments, increases in microbial biomass C and N, basal
respiration, metabolic quotient (qCO2), and FDA hydrolysis rate have
been noted following biosolids applications (Fernandes et al., 2005;
Garcia-Gil et al., 2004; Sanchez-Monedero et al., 2004) as have increases
in the activities of dehydrogenase, protease, urease, amylase, catalase,
b-glucosidase and alkaline phosphatase (Fernandes et al., 2005; Furczak
and Joniec, 2007; Garcia-Gil et al., 2004; Sastre et al., 1996). The stimula-
tory effect on microbial activity is most intense during the first few months
following biosolids applications (i.e., during the rapid phase of decomposi-
tion (Garcia-Gil et al., 2004). Even where levels of heavy metals in biosolids
are high, there can be an initial increase in microbial activity during the
initial phase of decomposition which is then followed by a later phase where
microbial activity is inhibited (Kizilkaya and Bayrakli, 2005).
The stimulating effect on soil microbial activity of the application of
composted biosolids has been shown to be lower but more persistent than
that of uncomposted biosolids ( Jimenez et al., 2007; Pascual et al., 2002;
Sanchez-Monedero et al., 2004). Nevertheless, substantial increases in
microbial biomass C and N, basal respiration rate, potentially mineralizable
N, and the activities of some soil enzymes have been noted following field
applications of composted biosolids ( Jimenez et al., 2007; Speir et al., 2004;
Zaman et al., 2004).
Increases in concentrations of dissolved organic matter in soil solution,
and its downward movement in the soil profile, following biosolids applica-
tions have been noted by a number of workers (Ashworth and Alloway,
2004; Han and Thompson, 1999; Neal and Sposito, 1986; Toribio and
Romanya, 2006). Han and Thompson (1999) also observed that the molec-
ular weight distribution of soluble organic matter in soils shifted to lower
weights (e.g., 14,000 Da) following biosolids applications. The signifi-
cance of dissolved organic matter to the mobility of biosolids-derived heavy
metals is discussed in Section 5.3.5.
3.2. Inorganic components
The inorganic content of biosolids, as measured by ash content, commonly
ranges from 30–60% ( Jaynes and Zartman, 2005; Sommers et al., 1976;
Terry et al., 1979). This high ash content (i.e., about 50%) results from the
effective removal of many of the inorganic components from wastewater
during primary and secondary treatment. The inorganic component of
biosolids consists mainly silt- and clay-sized particles that arise from a
range of sources including local soil and sediment materials, broken glass

washed into stormwater drains, inorganic residues in human feces
(e.g., relatively high concentrations of SiO2 are found in foods originating
from plant material; 1–4%), cosmetics, and other products washed down
residential drains.
X-ray fluorescence analysis on dried sludge by Thawornchaisit and
Pakulanon (2007) indicated that oxides of Si, Al, and Fe (with a combined
total of 62%) were the three main inorganic constituents of biosolids. X-ray
diffraction analysis of biosolids has been performed by a number of workers
( Jaynes and Zartman, 2005; Mun et al., 2005; Sommers, 1977). Jaynes and
Zartman (2005) observed an inorganic matrix consisting mainly of Quartz
(SiO2) and feldspars (crystalline minerals that consist of aluminum silicates
containing K, Na, Ca, or Ba) and kaolinite, mica, and expandable clays were
also present. Sommers (1977) identified quartz, calcite, dolomite, feldspars,
and layer silicates while Mun et al. (2005) found quartz was the dominant
mineral but there were also significant amounts of feldspars, muscovite, and
chlorite. In biosolids ash, Hartman et al. (2007) identified quartz and
haematite as the predominant minerals. Jaynes and Zartman (2005) also
found significant amounts of poorly crystalline Al and Fe phosphates
(thought to be formed during anaerobic digestion) and talc residues
originating from cosmetics.
4. Nutrient Content and Release
4.1. Nitrogen
The N content of biosolids can vary greatly (Sommers, 1977) but is typically
in the range of 2.8–3.8% (Epstein, 2003; Hue, 1995). Accumulation of total
N in the surface soil, 15 years after an application of 500 ton haÀ 1 of
biosolids to a forest soil is evident in Fig. 1. Because 50–90% (often quoted
as 80%) of N in biosolids is in organic form (Sommers, 1977), information
on the N mineralization rate is necessary to predict N availability following
land application. Because nitrification (the microbial conversion of NH4
þ
to NO3
À) is predominantly an aerobic process, in anaerobically digested
biosolids the content of mineral N consists of about 99% NH4
þ–N and 1%
NO3
À–N (USEPA, 1995). However, in aerobically digested biosolids the
bulk of the mineral N is present as NO3
À–N (Sommers, 1977). Mineraliza-
tion of biosolids-N in soils has been widely studied in laboratory incuba-
tions. Such studies with anaerobically digested sludge have reported
mineralization rates of 4–48% in 16 weeks (Ryan et al., 1973), 14–25% in
13 weeks (Magdoff and Chromec, 1977), 40–42% in 15 weeks (Epstein
et al., 1978), 15% in 16 weeks (Parker and Sommers, 1983), and 24–68%
in 32 weeks (Lindermann and Cardenas, 1984). The N mineralized tends to
be greater from aerobically than anaerobically digested biosolids (Hseu and

Huang, 2005; Magdoff and Chromec, 1977) and composting greatly
decreases biosolids-N mineralization potential (Epstein et al., 1978; Parker
and Sommers, 1983).
In biosolids, N mineralization potential has been related to total organic
N content and more particularly to various indices of protein content.
A large proportion of biosolids organic N is thought to be proteinaceous
in origin and this fraction represents a labile pool of organic N (Hattori and
Mukai, 1986; Lerch et al., 1992). Hattori and Mukai (1986) found a
correlation between mineralization of biosolids-N and crude protein con-
tent while Hattori (1988) found a correlation with proteinase enzyme
activity in the biosolids-amended soil. Lerch et al. (1992) also found a
correlation between N mineralization and low molecular weight amines
(assumed to be proteins) in biosolids while Rowell et al. (2001) found a
correlation with the alkyl index and the alkyl to O-alkyl ratio (as determined
by solid state13C NMR spectroscopy). This was explained as a reflection of
proteins in the alkyl region of the CPMAS NMR spectra and Rowell et al.
(2001) suggested that N mineralization from biosolids is mainly a conse-
quence of catabolism of the protein pool rather than decomposition of the
material as a whole.
In soils, N mineralization is carried out by the heterotrophic microbial
community and is therefore highly dependent on environmental factors
which affect microbial activity (e.g., soil type, temperature, water content,
aeration). Thus, under field conditions, the proportion of the potentially
20100 20100
0
50
100
150
0
50
100
150
Total N (mg g−1) Total P (mg g−1)
Sludge-treated
Control Control
Sludge-treated
Soildepth(cm)
Figure 1 Total N and P concentration with depth in a forest soil treated with
500 Mg haÀ 1
municipal biosolids 15 years previous to sampling and in a control
(untreated) plot. From Harrison et al. (1994); copyright Elsevier.

mineralizable pool of organic N that is actually released will be highly
variable depending on soil and seasonal conditions. Furthermore, minerali-
zation will proceed over a period of several years.
For agronomic and environmental purposes, it is often assumed that
20%, 10%, and 5% of biosolids organic-N is mineralized in the first, second,
and third year, respectively, after application (USEPA, 1995). As expected,
actual field mineralization rates are variable and depend on the interaction of
a number of factors including biosolids composition and rate of application,
soil type, pH, soil temperature, soil water content, and aeration (Artiola and
Pepper, 1992; Barbarick et al., 1996; Sims and Boswell, 1980). Based on
field trials in Wisconsin, Keeney et al. (1975) suggested an organic N decay
rate series of 15–20%, 6%, 4%, and 2% for the first, second, third, and fourth
years after application but Kelling et al. (1977a) found a decay rate of 45,
25–30, and 10–15% over a 3-year period. In California, Pratt et al. (1973)
found a decay rate of 35, 10, 6, and 5% over a 4-year period. From field trials
in Nebraska, Binder et al. (2002) found a decay series of 40, 20, 10, and 5%
over a 4-year period. Most data suggests that the USEPA guidelines are
conservative and that often more than 20% of biosolids organic N is
mineralized in the first year (Barbarick and Ippolito, 2000; Barbarick
et al., 1996; Cogger et al., 1998).
The agronomic response to applied biosolids-N will be greatly affected
by a range of environmental and soil conditions. Binder et al. (2002), for
example, showed in a series of field trials that irrigated maize yield response
was relatively consistent between years with maximum yields being attained
at about 441 kg organic N haÀ 1 (Fig. 2). However, dryland sorghum yields
were less consistent. In 1996, there was no significant yield response because
of high residual soil NO3
À and mineralizable N originating from a previous
soybean crop and a previous 3-year fallow (Fig. 2). Yields in 1997 and 1998
were similar and considered representative of more common rotations and
climatic conditions in south east Nebraska. In 1999, cool weather restricted
N mineralization rate and sorghum responded to much higher rates of
biosolids-N (Fig. 2).
For anaerobically digested biosolids, the NH4
þ initially present and that
which is ammonified soon after application is at risk of volatilization loss if
biosolids are surface applied. Ammonia volatilization is favored when high
concentrations of NH4
þ are present in an environment with a pH above 7.
The typically high pH of 6–8 in biosolids (see, Section 4.3) therefore tends
to favor volatilization and losses ranging from 25–80% of the initial NH4
þ
content have been recorded (Adamsen and Sabey, 1987; Beauchamp et al.,
1978; Robinson and Polglase, 2000; Robinson and Roper, 2003; Terry
et al., 1981). Incorporation of biosolids into the soil will minimize such
losses. Over a period of several weeks following biosolids application,
nitrification will typically proceed induced by indigenous autotrophic
nitrifier bacteria present in the soil.

It is important that the rate of biosolids-N supply matches crop N
requirements (i.e., that an ‘‘agronomic biosolids rate’’ is used; USEPA,
1993) since excess N will accumulate in the soil profile as the mobile
NO3
À anion. This can be lost from the soil as N2/N2O via denitrification
under anaerobic soil conditions or can be leached down the profile into
groundwater. Indeed, a frequently quoted hazard of biosolids applications is
excessive movement of NO3
À to groundwater (Keeney, 1989). To estimate
Typical year
1997, 1998
After soybean
1996
Cool/dry year
1999
Sorghum
Maize
Relativeyield,%
100
90
80
70
60
50
100
90
80
70
60
50
40
30
100 200 300 400 500 600 700 800
Organic N in applied biosolids, kg ha−1
0
Year applied
1996
1997
1998
1999
Figure 2 Relative yield response of irrigated maize and rainfed sorghum in relation to
the amount of organic N applied with biosolids in the year of application. From Binder
et al. (2002); copyright American Society of Soil Science.

an agronomic rate that supplies the amount of N required by the crop and
minimizes the amount of residual NO3
À available for leaching, the poten-
tially available N (PAN) concentration may be calculated:
PAN ¼ NNO3
þ XNNH4
þ YNorg;
where X is the fraction of NH4 that does not volatilize and Y is the fraction
of organic N (Norg) that is expected to be mineralized during the first season.
It is generally assumed that 100% of biosolids NO3 (NNO3
) is available for
plant uptake and 100% of NH4 is also available (i.e., X ¼ 1) unless biosolids
are surface applied in which case an estimate of the proportion of NH4
volatilized is made. As noted above, Y is difficult to estimate but is often
estimated at 0.20 in the year of application. Pierzynski (1994) suggested
figures of 0.25 for aerobically digested sludge, 0.15% for anaerobically
digested sludge, and 0.05–0.10 for composted biosolids.
Several workers have developed models specifically to describe NO3
À
leaching from biosolids-amended soils (Andrews et al., 1997; Joshua et al.,
2001; Vogeler et al., 2006). However, in general, applications of biosolids at
agronomic rates cause minimal NO3
À leaching (Correa et al., 2006; McLaren
et al., 2003; Surampalli et al., 2008). The greater the proportion of biosolids-
N initially present in NH4
þ form (which is rapidly nitrified following soil
application) the greater the potential for NO3
À leaching since there is more
NO3
À in the soil profile (Shepherd, 1996; Smith et al., 1998). Deep injection
of biosolids exacerbates leaching losses because less drainage is required to
leach N below the root zone (Shepherd, 1996). Timing of applications will
be an important consideration so that N supply from biosolids is in syn-
chrony with crop uptake requirements. For example, applying biosolids in
autumn prior to winter rains (during a period where crop growth and N
uptake is slow) is likely to favor leaching losses of NO3
À (Shepherd, 1996).
Nitrogen mineralization will occur whenever conditions are favorable
which on an annual basis is likely to be over a longer period than that for
N uptake by the crop. As a result, mineral N will inevitably be produced
during periods when there is little chance of plant uptake. It will therefore
be advisable, where repeated biosolids applications are being made, to
measure soil profile mineral N prior to biosolids applications and reduce
the biosolids application rate accordingly (Pierzynski, 1994).
4.2. Phosphorus
The P content of biosolids is often in the range of 1.2–3.0% (Sommers,
1977, Sommers et al., 1976). In anaerobically digested sludges, almost all the
P (80%) is present in inorganic form (Ajiboye et al., 2007; Hinedi et al.,
1989a,b; Shober et al., 2006; Smith et al., 2006) mainly as phosphate

adsorbed to ferrihydrite and Al hydroxides, hydroxyapatite and b-tricalcium
phosphate (Shober et al., 2006). Using combined sequential chemical
extraction, 31P NMR and XANES, Ajiboye et al. (2007) concluded that
readily soluble P forms in biosolids mainly originated from easily soluble Ca
and Al phosphates while recalcitrant forms were associated with Fe and Al.
In aerobically digested sludge, the organic P content is greater (e.g., 50%)
and this is present predominantly as phosphate monoesters and diesters
(Hinedi et al., 1989a). Organic P must undergo mineralization in the soil
before it is plant available. In lime-stabilized biosolids, recalcitrant calcium
phosphates (e.g., hydroxyapatite, tricalcium phosphate) become major
components (Shober et al., 2006).
A typical biosolids sample may contain 3.2% N and 1.4% P (Hue, 1995)
and although the biosolids provides about twice as much N as P, agricultural
crops sequester about four times as much N as P leading to an overall
increase in soil P in relation to N. Pierzynski (1994) calculated that if a
typical biosolids sample (containing 13 g kgÀ 1 PAN and 10 g kgÀ 1 total P)
were applied to supply 150 kg N haÀ 1, it would also apply 115 kg P haÀ 1
which is approximately three times more than would typically be recom-
mended for maize. The imbalance between N and P in biosolids typically
leads to a substantial increase in extractable soil P levels (Kelling et al.,
1977b; Maguire et al., 2000; Peterson et al., 1994), often to levels much
greater than those necessary for adequate P nutrition of crops. This can lead
to an increased potential for off-site movement of P via runoff and leaching.
The accumulation of total P in the surface layers of a biosolids-amended soil
is clearly illustrated in Fig. 1.
Current recommendations in both United Kingdom and United States
are that the relative effectiveness of biosolids-P, compared to soluble
fertilizer P, is 50% (MAFF, 1994; USEPA, 1995). O’Connor et al. (2004)
assessed phytoavailability of 12 different biosolids samples in a greenhouse
study, relative to triple superphosphate (TSP), and confirmed that most
biosolids produced by conventional methods had a relative phytoavailabil-
ity in the range of 25–70% TSP. Biosolids produced in water treatment
plants where Fe, Al, or Ca is added during treatment to lower soluble P (to
meet effluent limitations) have a lower P availability (i.e., 25% TSP)
(O’Connor et al., 2004). Indeed, in such biosolids, the solubility and
availability of P is characteristically low (Lee et al., 1981; Lu and
O’Connor, 2001, Maguire et al., 2000; Soon and Bates, 1982) since the
phosphate is strongly adsorbed to the surfaces of Fe and Al hydrous oxides
and calcium carbonate. Heat-dried biosolids also have low P availability
(Chinault and O’Connor, 2008). By contrast, biological P removal bioso-
lids have a high P phytoavailability (75% TSP) (O’Connor et al., 2004).
These biosolids are produced by a modified activated sludge process used to
produce low P concentrations in the treated effluent wastewater. It
employs aerobic and anaerobic zones to selectively enrich for bacteria

which take up large amounts of phosphate and store it intracellularly as
polyphosphate under cyclic anaerobic and aerobic conditions.
Surface runoff is the major pathway for P loss from soils to surface waters
(Daniel et al., 1998; Sharpley et al., 1994). Particularly where surface
applications of biosolids have been practiced, runoff of particulate matter
high in P is a potential danger since P inputs to aquatic freshwater systems
can increase the rate of eutrophication (Carpenter et al., 1998). The higher
the water-soluble P content of biosolids, the greater the risk of runoff losses
of P (Elliott et al. (2005).
Due to its strong adsorption onto soil colloids, it is usually considered
that there is a low risk of P leaching down the soil profile. However,
leaching can be a concern particularly in sandy soils (with low P sorption
capacity) with a low pH (because of increased P solubility) and/or where
soils have become P saturated, especially following heavy animal manure
applications (van Riemsdijk et al., 1987). Some studies have, however,
shown that if soil test P values exceed a certain critical ‘‘change point’’
value, soluble P increases and significant leaching losses can occur (Heckrath
et al., 1995; Hesketh and Brookes, 2000; McDowell et al., 2001). Such
leaching is thought to occur principally by macropore flow (e.g., in cracks,
earthworm burrows, and root channels) and much may be as particulate
organic matter and as phosphate sorbed to clay particles. Indeed, particle-
facilitated transport of P has been found to play an important role in
P leaching (de Jonge et al., 2004; Djodjic et al., 2000; Laubel et al., 1999;
Siemens et al., 2004). The elevation of soil test P values above change point
values, due to repeated biosolids applications, could therefore induce
increased P leaching particularly for biosolids low in reactive Fe and Al
(Elliott et al., 2002). Certainly, Sui et al. (1999) detected significant down-
ward movement of surface-applied biosolids-P into the 0–5 and 5–25 cm
soil layers after 6 years of annual applications.
4.3. Other nutrients
The K content of biosolids is very low (e.g., 0.15–0.40%), in comparison
with that for N, yet demand for it by crops is often comparable. For that
reason, biosolids is generally considered a poor source of K and supplemen-
tary fertilizer K applications often need to be made. The reason for this is
that most K compounds are water soluble and remain in the sewage effluent
or aqueous fraction during sludge dewatering. Nevertheless, the K in
biosolids is normally assumed to be 100% available for plant uptake
(Pierzynski, 1994).
The Ca (2.1–3.9%) and Mg (0.3–0.6%) content of biosolids is similar to
that of animal manures (Hue, 1995). Biosolids also supplies micronutrients
such as B, Cu, Zn, Mn, Fe, Mo, and Ni (Epstein, 2003) and this may be
important where micronutrient deficiencies occur in the soils where land

application is being practiced. Nevertheless, as discussed below, metals such
as Zn and Cu may sometimes be present in biosolids at levels that are
considered unacceptable.
Addition of biosolids also results in an increase in electrical conductivity
(EC) in soil solution (increased salinity) and alterations to soil pH (Clapp
et al., 1986). The EC of biosolids can be measured in a number of different
ways including directly on the wet sludge, or after drying in either satura-
tion paste extracts or 1:5 solid: water extracts. This contributes to variability
in reported values which generally lie between 3 and 12 dS mÀ 1 (Garcia-
Orenes et al., 2005; Moffet et al., 2005; Navas et al., 1998; Rostagno and
Sosebee, 2001). Such values are generally considerably greater than those
encountered in nonsaline soils (i.e., 0–2 dS mÀ 1 in saturation paste extracts
and 0–0.15 dS mÀ 1 in 1:5 soil: water extracts). The high EC in biosolids is
attributable to the high concentrations of ions such as Mg2þ, Ca2þ, and ClÀ
that are present. During heavy rains/irrigation, soluble salts will leach down
below the root zone and EC in the surface soil will return to that prior to
biosolids application.
Increases, decreases, and no effect of biosolids application on soil pH
have been noted (Clapp et al., 1986; Epstein, 2003; Singh and Agrawal,
2008). Changes will be dependent on many soil and biosolids properties
including the initial pH and buffering capacity of both materials. The
buffering capacity of the biosolids will be largely controlled by factors
contributing to the CEC of the material and the content of Ca and Mg
oxides. The initial pH of biosolids varies greatly but can often be in the
range of 6–8 (Epstein, 2003; Merrington et al., 2003; Navas et al., 1998).
Thus, in general, pH of acidic soils (e.g., 6) will tend to be increased while
that of alkaline soils (e.g., 8) will tend to be decreased. However, in a
range of soils a progressive decline in pH following biosolids application has
often been observed and this is attributable to nitrification of biosolids NH4
þ
(Clapp et al., 1986; Harrison et al., 1994; Navas et al., 1998; see, Sec-
tion 5.3.2). Changes in pH will have indirect effects on the availability of
nutrients as well as heavy metals (see, Section 5.3.3).
5. Heavy Metal Contaminants
Heavy metal is a term commonly used as a group name for metals and
semimetals (often defined as having an atomic number greater than 20
or 21) that have been associated with contamination and/or potential
toxicity to animals or plants. Common elements considered include Cu,
Zn, Co, Ni, Pb, Hg, Cd, Cr, Se, and As.

5.1. Total concentrations
A significant proportion of the anthropogenic emissions of heavy metals can
accumulate in sewage. Industrial wastewater is often the major source.
Wastewater from surface treatment processes (e.g., electroplating, galvaniz-
ing) can be a source of metals such as Cu, Zn, Ni, and Cr while industrial
products may, at the end of their life, be discharged as wastes. Key urban
inputs include drainage waters, business effluents (e.g., car washes, dental
uses, other enterprises), atmospheric deposition, and traffic related emissions
(vehicle exhausts, brake linings, tires, asphalt wear, petrol/oil leakage, etc.)
which are transported with stormwater into the sewage system (Bergback
et al., 2001; Comber and Gunn, 1996; Sorme and Lagerkvist, 2000).
Household effluents can also be important. For example, at an English
treatment works, Comber and Gunn (1996) found domestic inputs of
Cu and Zn were large representing 64 and 46%, respectively, of total inputs.
The bulk of the Cu originated from Cu piping while most of the Zn came
from household activities (since it is a component of skin creams, ointments,
makeup, deodorant, talcum powder, shampoo, and aftershave).
The presence or absence of elevated heavy metal concentrations in
sewage varies enormously between treatment works and depends greatly
on local factors such as type and number of industries in the region,
regulations regarding the quality of industrial discharges allowed to sewers
and public awareness of the environmental impacts of metal contaminated
discharges. Heavy metal content of sewage often fluctuates due to irregular
inputs from industrial and urban sources and as a result influent concentra-
tions can vary greatly on an hourly, daily, or monthly basis (Brown et al.,
1973; Oliver and Cosgrove, 1974). As a result the biosolids produced at one
treatment works can also vary greatly in heavy metal loadings with time.
Although waste water treatment plants are expected to control the
discharge of heavy metals to the environment, they are chiefly designed
for removal of organic matter. Heavy metal removals are a side benefit.
Metal removal occurs both during primary and secondary treatment. Dur-
ing primary treatment, as suspended solids slowly settle out, metals asso-
ciated with/adsorbed to the solid particles are concentrated in the sediment
and are then removed with the sediment. During secondary treatment two
main processes lead to removal of metals. These are (i) bioaccumulation in
which metals are accumulated into the living bacterial cells and (ii) biosorp-
tion in which heavy metals are sorbed onto negatively charged sites on
bacterial cell walls and on extracellular polysaccharide gels (Brown and
Lester, 1979; Urrutia, 1997). The heavy metals are then removed in the
microbial sludge which is mixed with the primary sludge. The heavy metal
concentrations in primary and secondary sludges (on a dry weight basis) are
typically similar in order of magnitude but concentrations are typically

30–70% greater in primary sludges (Alonso et al., 2009; Alvarez et al., 2002;
Solis et al., 2002).
The extent of removal of metals during primary and secondary treatment
can vary greatly for different metals in the same treatment plant as well as
between plants. For example, in a treatment plant in Poland, Chipasa (2003)
recorded removal efficiencies of Zn 84%, Cu 51%, Pb 33%, and Cd 15%
and noted that these were directly proportional to metal influent concen-
trations. From a variety of sources, Lester et al. (1979) and Stoveland et al.
(1979) reported removal efficiencies of Cu 71–96%, Pb 91–95%, Cd
78–91%, Zn 60–94%, Ni 11–70%, and Cr 67–79%. Many factors influence
removal efficiency including initial concentrations of metals in influents,
characteristics of individual metals (e.g., pH/solubility relationships),
operating parameters of the plant and other physical, chemical, and
biological factors (Brown and Lester, 1979; Chipasa, 2003; Stoveland
et al., 1979). Thus, removal efficiency is not a predictable property.
A large number of studies in many parts of the world have surveyed the
heavy metal content of biosolids samples (e.g., Kuchenrither and McMillan,
1990; Ozaki et al., 2006; Sajjad et al., 2005) and much of this data has been
summarized previously (Epstein, 2003; Hue, 1995). Taking account of the
great variability in heavy metal inputs which occurs between water treat-
ment plants, some ‘‘typical’’ concentrations of metals encountered in bio-
solids samples (in mg kgÀ 1 values) are shown in Table 2. It is evident that
Zn is commonly present in highest concentrations and that substantial
concentrations of Pb, Cu, and Cr are also often present. In the United
States and Canada, heavy metal concentrations in biosolids (particularly
those of Cd, Cr, Pb, and Ni) have been shown to be decreasing during
Table 2 Typical concentrations of heavy metals commonly
encountered in biosolids
Element
Concentration
(mg kgÀ 1
dry weight)
Arsenic 1–20
Cadmium 1–70
Chromium 50–500
Cobalt 5–20
Copper 100–800
Lead 100–600
Mercury 1–10
Nickel 10–200
Selenium 5–10
Zinc 1000–3000
Calculated from Hue (1995), Mininni and Santori (1987), and Epstein
(2003).

the 1980s and 1990s (Epstein, 2003; Hue, 1995). This is attributable to
enforcement by municipalities of regulations regarding the maximum metal
loadings in effluents that can be discharged into the sewerage system. As a
result, industrial pretreatment of effluents has become common. However,
for Zn and Cu, concentrations in biosolids have remained similar over the
last two decades (Epstein, 2003) because, as noted previously, they are often
not principally of industrial origin. While heavy metal concentrations in
biosolids have generally been decreasing and in most situations they are
below regulatory limits (see below), their addition to soils still causes
disquiet. This is because, unlike organic contaminants, most heavy metals
do not undergo microbial or chemical degradation and therefore elevated
concentrations persist in the soil for extremely long periods of time.
Concerns regarding the heavy metal load in biosolids have resulted in
guidelines and regulations being developed in many parts of the world to
regulate land applications. These are generally based on the maximum
allowable metal concentration limits (mg kgÀ 1 dry weight) in biosolids
and/or the allowable loading limits (kg haÀ 1 yrÀ 1) of metals added in
biosolids to soil (Epstein, 2003). The most quoted limits are those of the
USEPA (USEPA, 1993) and the European Union also has its own standards.
In general, USEPA and UE limits for metal concentration limits in biosolids
are broadly similar but maximum loading limits are generally lower for the
EU guidelines. Nevertheless, limits can vary quite widely with countries
such as Sweden, Denmark, Germany, and the Netherlands generally having
lower limits than USEPA or EU guidelines (Smith, 2001). USEPA metal
concentration limits in biosolids are: Zn, 2800; Cu, 1500; Ni, 420; Pb, 300;
Cd, 39; and As, 41 mg kgÀ 1 (USEPA, 1993). USEPA regulations are risk
based and therefore provide an opportunity to modify values as better
scientific data becomes available (Epstein, 2003).
5.2. Extractable fractions
Total concentrations of heavy metals indicate the extent of contamination
but provide little insight into the potential mobility or bioavailability of
these metals once the biosolids are soil applied. Depending on their nature,
individual metals are associated in a variable manner with different phases
making up the biosolids. Sequential chemical fractionation procedures are
widely used to characterize the forms of metals present (chemical specia-
tion). These methods involve chemical extractions using a sequence of
reagents of increasing strength. For each reagent, a particular chemical
form(s) is assigned to the metals extracted. Drawbacks of these methods
include (i) lack of specificity, selectivity, and validation; (ii) postextraction
readsorption; and (iii) sensitivity to procedural variables (e.g., sample size,
pH, temperature, contact time, concentration of extractant, etc.) (Kot and
Namiesnik, 2000). Despite such limitations, sequential extractions are

considered the best available method of gaining knowledge on the forms in
which metals are present in biosolids.
A wide range of sequential fractionation schemes have been proposed for
determination of heavy metal forms present in biosolids (Kot and
Namiesnik, 2000; Marchioretto et al., 2002; Sims and Kline, 1991; Tessier
et al., 1979). One of the simplest and most commonly used methods today is
that specified by the Community Bureau of Reference (CBR) (Ure et al.,
1993) in which the sample is extracted with (i) acetic acid to release the
easily available ‘‘exchangeable’’ forms present in soluble and exchangeable
forms and those associated with carbonate phases, (ii) hydroxylammonium
chloride to release the ‘‘reducible’’ fraction associated with Fe and Mn oxide
cements and nodules (forms that could become available under anoxic
conditions), and (iii) hydrogen peroxide to extract the ‘‘oxidizable’’ fraction
that is strongly bound to organic matter constituents. Following the sequen-
tial extraction, the amounts remaining in the ‘‘residual’’ fraction (iv) are
measured after digestion with aqua regia and these are considered to be
highly unavailable and associated with residual solids that occlude metals
in their crystalline structures. The amounts present in fractions (i) and (ii) are
considered ‘‘available’’ and those in (iii) and (iv) ‘‘unavailable.’’
This method has been extensively used for characterization of biosolids
(Alonso et al., 2006, 2009; Alvarez et al., 2002; Fuentes et al., 2004, 2008;
Perez-Cid et al., 1999; Scancar et al., 2000; Solis et al., 2002; Sprynskyy
et al., 2007; Wang et al., 2005, 2006a,b). To generalize from the results of
these studies, Cu is typically found to be concentrated (about 80% of total
Cu content) in the oxidizable fraction bound to organic matter. This is in
accordance with the high stability constant of the Cu complexes with
organic matter (Ashworth and Alloway, 2004). By contrast, Zn is
distributed preferentially (usually 40–60%) in the available exchangeable
plus oxidizable fractions. Greater than 50% of total Pb content is typically
found in the residual fraction with substantial amounts (15–30%) also being
present in the oxidizable fraction. Ni and Cd have a similar distribution with
60–70% of total content being present in the unavailable oxidizable and
residual forms (usually more or less equally distributed between the two
fractions). Co is similarly distributed between unavailable and available
fractions with significant amounts (30–50%) being present in the organic
fraction. Cr is concentrated in the unavailable forms (usually more than 90%
of total content) with over 50% in the residual fraction and a significant
proportion also organically bound. For Fe, 80–90% of total content is in
unavailable forms with greater than 60% in the residual form and 10–20% in
the organic fraction. However, for Mn, 70–80% of total content is in
available forms with greater than 50% in the exchangeable form. In sum-
mary, Zn and Mn are the metals preferentially found in the mobile fractions
of biosolids while the others are mainly concentrated in immobile forms.

Cu and, to a lesser extent Pb and Co, have a particular affinity for binding to
the organic components of biosolids.
Solis et al. (2002) showed that for all metals (on a mean basis) the
available (exchangeable plus reducible) fractions were higher in secondary
than primary sludge. During anaerobic digestion of combined sludge there
was a general increase in the percentage of metals in the unavailable
oxidizable and residual fractions and during composting of the biosolids
there was a further increase in the percentage of metals present in the
unavailable fractions. A number of other workers have followed heavy
metal fractions during the composting of biosolids with variable results.
Amir et al. (2005b) found that potentially available fractions of Cu, Zn, Pb,
and Ni tended to decrease over time while Zorpas et al. (2008) observed
similar results for Cr, Cu, Mn, Fe, Ni, and Pb. However, Nomeda et al.
(2008) showed that available fractions of Pb, Zn, and Cd increased with
time but those of Cu decreased. Liu et al. (2007a,b) observed that during
composting, the available fractions of Pb and Zn increased while those of
Cu, Ni, and Cr were little affected. Thus, although it is clear that heavy
metal levels are concentrated during composting, the effects on distribution
of metals among fractions are much less clear and may vary depending
on conditions of composting, presence or absence of a bulking agent
(e.g., sawdust, bark), and other factors such as changes in pH.
Where biosolids have a high loading of heavy metals, the material can be
cocomposted with an absorbent material such as zeolite (e.g., crushed
clinoptilolite rock) added at 10–25% w/w. This results in substantial
decreases in the amounts of metals being present in the potentially available
exchangeable and reducible fractions (Sprynskyy et al., 2007; Zorpas et al.,
2008) since the metals are adsorbed to the zeolite surfaces. Cocomposting
with a sodium sulfide/lime mixture (3% w/w) was also shown by Wang
et al. (2008) to reduce the percentage of metals in the available fractions.
A number of methods have also been developed to remove heavy metals
from contaminated biosolids prior to land application. These include chem-
ical extraction, bioleaching, electroreclamation, and supercritical fluid
extraction (Babel and del Mundo Dacera, 2006).
5.3. Application to the soil
5.3.1. Heavy metal extraction from soils
It has often been observed that heavy metal availability in biosolids-
amended soils is closely related to total metal content of the added biosolids
( Jamili et al., 2007; Jing and Logan, 1992). Nonetheless, the presence of
biosolids constituents that adsorb metals limits the usefulness of total metal
content as an indicator of potential metal availability (Merrington et al.,
2003). For example, Richards et al. (1997) found total metal contents of a

range of biosolids samples was not closely related to metal mobility as
estimated by the TCLP leaching procedure. Indeed, biosolids application
to the soil not only increases the concentrations of heavy metals present but
also alters the adsorption capacity of the soil (Alloway and Jackson, 1991).
As already noted, biosolids are composed of about 50% inorganic and 50%
organic material. The relative importance of the inorganic and organic
components in retention of heavy metals by biosolids is a matter of contro-
versy (Basta et al., 2005; Merrington et al., 2003) but is likely to differ for
different biosolids samples as well as for different metals.
Total loadings of heavy metals in biosolids-amended soils are not neces-
sarily a good indicator of potential metal availability. Sequential fraction-
ation schemes, as discussed in Section 5.2, are often employed to selectively
extract metals associated with particular soil phases (Ure et al., 1993).
Despite the limitations of such fractionation schemes, their use gives some
indication of the fate of biosolids-borne heavy metals once they enter the
soil system. In particular, fractionations are useful in studying the partition-
ing of metals between potentially available (toxic) and residual, occluded
(nontoxic) fractions and the association of metals between organic and
inorganic soil constituents.
A wide range of soil test extractants have been employed to determine
heavy metal availability (McLaughlin et al., 2000a; Ure, 1995). The most
commonly used extractants are the organic metal complexing agents diethy-
lenetriaminepentaacetic acid (DTPA) and ethylenediaminetetraacetic acid
(EDTA). The DTPA test is favored in the United States and EDTA in the
United Kingdom. Correlations between DTPA- and EDTA-extractable
metals and metal uptake by crops are generally reasonable (Bidwell and
Dowdy, 1987; Brun et al., 1998; Hooda et al., 1997; Hseu, 2006; Sanders
et al., 1986, 1987; Sukkariyah et al., 2005a). Dilute acids (e.g., 0.05–0.1 M
CH3COOH, HCl, and HNO3) are also used as heavy metal extractants
(McLaughlin et al., 2000a). Dilute salt solutions (e.g., 0.1 M CaCl2, Ca
(NO3)2, NH4NO3) are also effective extractants for predicting metal avail-
ability (Alloway and Jackson, 1991; Juste and Mench, 1992; Sukkariyah
et al., 2005a). These latter salt solutions extract metals in soil solution plus
those in short-term equilibrium with that solution. Complexing reagents
and dilute acids extract larger amounts of metals which include a ‘‘poten-
tially available’’ fraction. They, in affect, overestimate phytotoxicity and
assess potential rather than immediate toxicity (McLaughlin et al., 2000b).
McLaughlin et al. (2000b) suggested that in the future regulations and
guidelines should consider extractable fractions of heavy metals in soils.
That is, it is the concentration of biologically active (extractable) heavy
metals present in biosolids-treated soil that is toxic to plants and soil biota
(Merrington et al., 2003), yet present regulations are based on total loadings
of metals (see, Section 5.1). McLaughlin et al. (2000b) considered that
metals extracted with dilute salt solutions and those extracted with more

harsh reagents (complexing agents or dilute acids) could be used together to
estimate immediately toxic and potentially toxic metals, respectively.
Certainly, extractable metal concentrations are likely to give a better
indication of bioavailability than values based on total concentrations.
Monitoring of extractable metal levels on long-term sites, where biosolids
applications are continuing and/or have been terminated, will give valuable
data on the long-term trends in bioavailability of various total loadings. Such
data could well be used in the future to develop guidelines and regulations
based on extractable soil metal levels.
5.3.2. Effects of biosolids properties on availability
Following land application, the properties of the biosolids effect metal
availability both directly (through heavy metal content and sorptive capacity
of inorganic and organic components) and indirectly (through properties
such as pH, mineralizable N content, and EC) (Merrington et al., 2003). It is
usually assumed that biosolids properties dominate metal bioavailability in
the short and medium term in the zone of incorporation but with time,
biosolids properties have progressively less influence and soil properties
ultimately control availability (Smith, 1996). The effect of biosolids materi-
als on heavy metal retention by amended soils is complex and this is at least
partially because a suite of metals is added, and competition between them
for adsorption sites occurs. Bergkvist et al. (2005), for example, found Cd
sorption was slightly smaller in biosolids-amended soils compared to control
even though organic C content was 70% higher and oxalate-extractable Fe
was roughly doubled. They attributed this to competition for sorption sites
between Cd and biosolids-derived Fe and other metals such as Zn. McBride
et al. (2006) found that addition of high Fe, high Al, and biosolids to soils
had no long-term effect on their affinity for Cd. By contrast, Vaca-Paulin
et al. (2006) observed that biosolids-amended soils showed increased
adsorption capacity for Cu and Cd and attributed this to the complexing
ability of the biosolids-derived organic matter.
Strong metal retention by the inorganic fraction is attributable to the
high adsorption capacity of Fe, Al, and Mn hydrous oxides and silicates
(Basta et al., 2005; Merrington et al., 2003). The inorganic solids present in
biosolids are initially present, at least partially, in noncrystalline form
(Baldwin et al., 1983; Rogers and McLaughlin, 1999) and the higher surface
area of noncrystalline Fe and Al oxides results in them having a higher
adsorption capacity than their crystalline counterparts (Rogers and
McLaughlin, 1999). In general, the order of affinity of metals for adsorption
surfaces on Al and Fe oxide surfaces follow the order Cu Pb Zn
Co Ni Cd although for Fe oxides Pb Cu has been reported and
sometimes also Ni Co ( Jackson, 1998; Sparks, 2003). In addition, car-
bonate, phosphate, and sulphite present in biosolids can form sparingly
soluble solid phases with many metals and thus account for a substantial

portion of some metals present biosolids (Karapanagiotis et al., 1991). For
example, during anaerobic digestion, low solubility Cu and Zn sulfides
characteristically form (Nagoshi et al., 2005).
The organic component also has the ability to bind to heavy metals. The
heterogeneous nature of humic substances and the large number of func-
tional groups present means that binding of metals can be regarded as
occurring at a large number of reactive sites with binding affinities that
range from weak forces of attraction (ionic) to stable coordinate linkages
(McBride, 2000; Sparks, 2003). Indeed, mechanisms involved in metal
binding to organic matter are complex and probably involve simultaneous
chelation, complex formation, adsorption, and coprecipitation (Stevenson
and Vance, 1989). Because of the many variables involved, there are many
inconsistencies in reported selectivity orders of metals with organic matter.
A generalized order is Cr3þ Pb2þ ¼ Hg2þ Cu2þ Cd2þ Zn2þ ¼
Co2þ Ni2þ ( Jackson, 1998; Jin et al., 1996; Stevenson, 1994).
As noted previously, there is often a flush of organic matter decomposi-
tion following application, and this is followed by a slow decomposition
phase. It has been suggested that heavy metals bound to biosolids organic
matter could be released to soil solution during decomposition and as a
result metal bioavailability would increase over time (Hooda and Alloway,
1994; McBride, 1995). In fact, it is often observed that heavy metal avail-
ability is greatest immediately (the first few months) following biosolids
additions and this is followed by a reduction in availability (as estimated by
metal extractability and/or plant uptake) as well as a reduction in organic
matter content (Bidwell and Dowdy, 1987; Hseu, 2006; Logan et al., 1997;
McBride et al., 1999; Walter et al., 2002). Nonetheless, the initial high
availability may well be partially due to the rapid decomposition of biosolids
organic matter and the consequent release of metals. Evidently, the metals
released from decomposing organic matter are rapidly readsorbed by inor-
ganic and/or organic components in the soil/biosolids.
Biosolids pH will have a substantial controlling influence on the avail-
ability of metals following land application. In general, most heavy metal
cations become increasingly immobile at high pH. This is because both
their adsorption onto reactive oxide surfaces and precipitation reactions are
favored at high pH (Sparks, 2003). As noted in Section 4.3, since the initial
pH of biosolids is typically in the range of 6–8, their application will have a
liming effect on acid soils thus raising their pH (Kidd et al., 2007) and
tending to reduce metal availability.
The mineralizable N content of biosolids is, however, an important
property in relation to their effects on soil pH. During ammonification of
organic N to NH4
þ–N, one OHÀ ion is released per unit of N while during
nitrification of NH4
þ–N to NO3
À–N, two Hþ ions are released. The overall
process of conversion of organic biosolids-N to NO3
À–N is therefore
acidifying. Thus, Hooda and Alloway (1994) observed a progressive

decrease in soil pH following biosolids application to soil which was
accompanied by an accumulation of soil NO3
À–N and an increase in uptake
of Cd, Ni, Pb, and Zn by ryegrass growing in the soil. Such an increase in
metal bioavailability accompanying acidification induced by nitrification of
biosolids-derived N has also been observed by others (De Haan, 1975;
Hooda and Alloway, 1993). It is therefore important to monitor pH and
apply lime, if necessary, to maintain a relatively high pH (e.g., 6.5) follow-
ing biosolids application.
As noted in Section 4.3, the high EC of biosolids may result in an
increase in soluble salts in soil solution. High soluble salts will tend to reduce
soil solution pH (by exchange between cations in soil solution and Hþ and
Al3þ on soil cation exchange sites) thus increasing the solubility of heavy
metal cations. In addition, high concentrations of solution ClÀ can increase
mobilization, availability, and plant uptake of Cd through the formation of
Cd–chloro complexes (Weggler-Beaton et al., 2000).
5.3.3. Effects of soil properties on availability
Soil properties such as pH, redox potential, EC, clay, hydrous oxide, and
organic matter content will also influence heavy metal availability. The
most widely recognized factor is soil pH. With the exception of As and
Se, heavy metal retention by soils increases with increasing pH (McBride,
1994). As noted above, with an increase in pH, the charge on the variable
charge adsorption surfaces (e.g., Fe, Al, and Mn hydrous oxides) becomes
increasingly negative thus favoring metal cation adsorption and the high pH
also favors surface precipitation of the metals onto the surfaces (Bradl, 2004;
McBride, 2000). In general, the more mobile metals such as Ni, Cd, and Zn
are more sensitive to increasing pH than other metals such as Pb and Cu that
are more strongly complexed with soil organic colloids (Smith, 1996).
Manipulation of soil pH has been found to be the most effective way of
controlling heavy metal bioavailability in biosolids-treated soils (Alloway
and Jackson, 1991). Indeed, a large number of workers have shown that the
bioavailability of metals to plants in biosolids-amended soils decreases as pH
is raised either by liming or applying lime-stabilized sludges (Basta and
Sloan, 1999; Milner and Barker, 1989; Oliver et al., 1998). Liming a
range of biosolids-treated soils to pH 7 was shown by Jackson and
Alloway (1991) to reduce Cd content of lettuce by an average of 41% and
cabbage by 43%.
Redox potential is often considered an important factor although both
increases and decreases in heavy metal solubility have been recorded
following waterlogging and the onset of anaerobic soil conditions
(Charlatchka and Cambier, 2000; Chuan et al., 1996; Grybos et al., 2007;
Kashem and Singh, 2001a,b; Xiong and Lu, 1992). This is because a
number of different processes occur following the onset of anaerobiosis
and these often interact to affect metal solubility. In freely-drained soils,

Fe and Mn occur in their high oxidation states as oxides and hydrous oxides.
However, as soils become anaerobic, due to waterlogging, the redox
potential decreases and oxide minerals begin to dissolve as soluble Mn2þ
and Fe2þ forms (Stum, 1992; Stum and Sulzberger, 1992). This can not only
result in an increase in the solubility of Mn and Fe but also of other metals
(e.g., Zn, Cu, Co) which were previously adsorbed to, or occluded by,
these oxides (Chuan et al., 1996; Grybos et al., 2007). When soils become
anaerobic the pH tends to converge to neutrality irrespective of initial pH,
whether acidic or alkaline (McBride, 1994). For acidic soils this increase in
pH can result in release of organic matter into soil solution and metals
bound to the organic molecules are also thought to be released (Grybos
et al., 2007). This also tends to increase metal solubility. Nonetheless, the
increase in pH up to about 7, favors adsorption/surface precipitation of
metal cations thus favoring removal of metals from solution (Kashem and
Singh, 2001a). In addition, at low redox potential sulfate ions are reduced to
the sulfide form which may form complexes with metals such as Cd, Zn,
and Ni (Hesterberg, 1998; Van Den Berg et al., 1998). Most metal sulfides
are insoluble even under acidic conditions and so this process also tends to
reduce soluble metal concentrations.
Oxidation state of the contaminant itself also affects solubility.
For example, selenite [Se(IV)] is much more strongly adsorbed to soil
colloid surfaces than selenate [Se(VI)] and the presence of selenite is favored
under reducing conditions (Martinez et al., 2006; Neal and Sposito, 1989).
Se will therefore be less plant available under reducing conditions. Further-
more, under strongly reducing conditions Se may form elemental Se and
metal selenides (e.g., FeSe) both of which are insoluble (Elrashidi et al.,
1987; Masschelyen et al., 1991). Under oxidizing conditions both arsenate
[As(V)] and arsenite [As(III)] are present while under reducing conditions
As is present mainly as As(III) (O’Neill, 1995). Compared to other As
species, As(III) exhibits the greatest mobility and plant availability because
of its presence as the neutral species H3AsO3 (Ascar et al., 2008; Marin et al.,
1993). Nonetheless, strongly reducing conditions in biosolids-amended soils
can lead to precipitation of As as As2S3 (Carbonell-Barrachina et al., 1999).
The ability of soils to adsorb and sequester metals is also an important
factor. This is dependent on their content of inorganic (clay and Fe, Mn and
Al hydrous oxide content) and organic (soil humic material) binding agents.
For example, sandy soils with low oxide content and low organic matter
have low sorption capacities and will have greater metal availabilities than
loamy or clayey soils containing greater amounts of sorbents (e.g., clays,
oxides, and organic matter) provided the soils have similar pH values
(Basta et al., 2005). Hue et al. (1988) applied increasing rates of biosolids
to three different soils, a limed volcanic ash-derived Andept, an alkaline
Vertisol, and a limed manganiferous Oxisol. DTPA-extractable soil metal
levels, lettuce growth, and tissue metal concentrations were measured.

The Andept had the highest metal adsorption capacity and the Oxisol the
lowest. As a result, lettuce Cd, Mn, Ni, and Zn concentrations were highest
in the Oxisol and Mn levels reached phytotoxic levels. Hue et al. (1988)
concluded that the Andept could tolerate the highest biosolids loading rate
and the Oxisol the lowest. The calcite (CaCO3) content of soils can also be
important. In calcareous soils, calcite represents an effective sorbent for
metal ions. The initial reaction is thought to be chemisorption but metals
with an ionic radius similar to that of Ca (Cd 2þ, Mn 2þ, Fe 2þ) can also
readily enter the calcite structure and form coprecipitates (Gomez de Rio
et al., 2004; McBride, 2000).
5.3.4. Metal availability over time
The long-term (10 years) bioavailability of heavy metals in biosolids-
amended soils is of great importance in relation to environmental effects
of land application of biosolids. As noted previously (Section 5.3.2), follow-
ing a one-off application of biosolids the extractability of metals generally
declines over time (Hseu, 2006; Sukkariyah et al., 2005a; Walter et al.,
2002). Sukkariyah et al. (2005a), for example, showed DTPA-extractable
Cu and Zn levels progressively decreased following one-time applications
of biosolids at rates ranging from 42 to 210 Mg haÀ 1 (Table 3). Seventeen
years after application, extractable concentrations of Cu and Zn had
decreased by 58% and 42%, respectively. The decrease is attributable to
metals reverting to more recalcitrant forms in the soil such as occlusion in Fe
oxides or chemisorption to surfaces.
Despite the initial decrease in extractability, concentrations of extract-
able heavy metals in biosolids-amended soils can remain elevated above
Table 3 Long-term effect of biosolids application on DTPA-extractable Cu and Zn
DTPA-extractable
Cu mg kgÀ 1
DTPA-extractable
Zn mg kgÀ 1
Biosolids rates
Mg haÀ 1
1984 1995 2001 1984 1995 2001
0 1.4f a
3.7f 3.2f 1.6f 2.8f 2.7f
42 24.9e 23.1e 12.6e 19.2e 17.2e 9.1e
84 53.0d 44.3d 25.4d 38.9d 33.3d 19.8d
126 73.4c 64.8c 33.7c 52.4c 49.6c 27.9c
168 119.9b 78.7b 43.3b 73.2b 59.5b 35.5b
210 129.4a 92.8a 53.6a 78.2a 69.9a 49.7a
a
Values within columns followed by different letters are significantly different at the 0.05 probability
level.
From Sukkariyah et al. (2005a); copyright American Society of Agronomy.

those of control for many decades after applications have ceased (Alloway
and Jackson, 1991; Basta et al., 2005; McBride, 1995; McGrath, 1987).
Results from a long-term market garden experiment at Woburn (UK) serve
to illustrate this point. Sludge was applied in the 1940s until the 1960s and
CaCl2-extractable Cd changed little from 1950 until the early 1980s
remaining significantly higher than the control soils over the entire interval
monitored (McGrath and Cegarra, 1992). Similarly, EDTA-extractable Cu,
Pb, Zn, Ni, and Cr changed little following termination of biosolids
application and treated soils maintained a much greater proportion of
metal in EDTA-extractable form than the control. Such results occurred
despite there being a significant loss of biosolids organic matter over the
period indicating that heavy metals released from the decomposing organic
matter were rapidly adsorbed by inorganic components of biosolids/soil
and/or native soil organic matter. Certainly, biosolids-derived heavy metals
are strongly sorbed to soil components making them characteristically
immobile in soils. Indeed, the vast bulk of the added metals remain in the
topsoil in the layer of incorporation and there is a marked reduction in
concentration with depth (Alloway and Jackson, 1991; Brown et al., 1997;
Chang et al., 1983; Sloan et al., 1997; Sukkariyah et al., 2005b).
5.3.5. Heavy metal mobility and leaching
The results of Sukkariyah et al. (2005b) serve to illustrate the immobility of
biosolids-borne heavy metals in soil. They found that more than 85% of
total applied Cu and Zn was still in the layer of incorporation (0–15 cm)
17 years after a one-time biosolids application. Results for Mehlich
I-extractable Cu and Zn at that site are shown in Fig. 3. It is evident that
extractable Cu and Zn are concentrated in the 0–15 cm layer but there is
some indication of a small amount of movement down into the 15–20 cm
layer. Mass balances calculated for several long-term experiments do suggest
some losses of heavy metals from the topsoil (McBride, 1995). Lateral
movement in the soil due to tillage (McGrath and Lane, 1989) or physical
mixing with the lower soil layer by plowing (Sloan et al., 1998) can be
responsible for a significant part of the losses from the original amended soil
layer. Nevertheless, mass balances calculated for sites where little or no
tillage has been performed have shown less than 100% recovery (McBride
et al., 1999). Increased extractable heavy metal levels (e.g., for Cu, Zn, Ni,
Pb) at depths of 20–150 cm below the level of incorporation have been
noted in field experiments (Barbarick et al., 1998; Baveye et al., 1999; Bell
et al., 1991; Keller et al., 2002; Schaecke et al., 2002). Leachate sampling
below field plots and/or undisturbed monolith lysimeters receiving biosolids
has also revealed elevated metal concentrations (Keller et al., 2002; Lamy
et al., 1993; McBride et al., 1997, 1999; Richards et al., 1998; Sidle and
Kardos, 1977). In addition, column leaching studies have shown that heavy
metals can leach through many tens of cm of soil (Al-Wabel et al., 2002;

0 10 20 30 40 50 60 70
0–15
15–20
20–25
25–30
30–35
80–85
85–90
0 20 40 60 80
0–15
15–20
20–25
25–30
30–35
80–85
85–90
Concentration (mg kg−1)
Depth,cm
210 Mg ha−1
126 Mg ha−1
Control
210Mg ha−1
126Mg ha−1
Control
Zn
Cu
/ / / /
/ / / /
Figure 3 Distribution of Mehlich-I extractable Cu and Zn with soil depth 17 years
after biosolids application. From Sukkariyah et al. (2005a,b); copyright American Society
of Agronomy.

Antoniadis and Alloway, 2002; Ashworth and Alloway, 2004; Parakash et al.,
1997; Toribio and Romanya, 2006).
In most studies, the annual export of metals from the surface-mixing
layer represents a small fraction (i.e., 1–2%) of the total amount of metal
added (Holm et al., 1998; Keller et al., 2002; Lamy et al., 1993). Nonethe-
less, cumulative transport of metals over a long period of time could result in
a substantial redistribution into the subsoil layers and/or groundwater. In
addition, in some studies, water quality standards have been exceeded in soil
solution at depths below the zone of incorporation (McBride et al., 1999;
Richards et al., 1998). Dilution by other unpolluted water will normally
prevent water quality standards being exceeded in receiving groundwater.
The most danger will occur where large areas of land above small, shallow
water bodies are treated with biosolids.
A major contributor to heavy metal mobility in soils is thought to be the
formation of complexes with dissolved organic matter released from the
biosolids (Brown et al., 1997; Christensen, 1985; Gerritse et al., 1982; Lamy
et al., 1993; McBride et al., 1997). The amount of dissolved organic matter
in soil solution and leaching through the profile characteristically increases
following biosolids application and it acts as a ‘‘carrier’’ for heavy metals.
Elevated concentrations of both heavy metals and dissolved organic matter
are frequently found together in leachates below biosolids-treated soils
(Al-Wabel et al., 2002; Antoniadis et al., 2007; Ashworth and Alloway,
2004; Keller et al., 2002; Toribio and Romanya, 2006). Antoniadis et al.
(2007), for example, found that during a 310-day incubation of soils amended
with biosolids at 0, 20, and 100 Mg haÀ 1, there was a substantial increase in
dissolved organic C at about day 23 which was attributed to a flush of
microbial activity. This was accompanied by a similar increase in soluble
Zn and an increase in calculated activity of Zn-organic matter species (Fig. 4).
The formation of strong soluble organic matter–heavy metal complexes
in soil solution has been found to reduce heavy metal adsorption to solid soil
phases. Neal and Sposito (1986), for example, found that sewage sludge can
provide sufficient dissolved organic matter to reduce adsorption of Cd onto
soil surfaces. Wong et al. (2007) showed dissolved organic matter had a
stronger inhibitory effect on Zn sorption than that of Cd. Liu et al. (2007a,b)
also showed dissolved organic matter depressed sorption of Ni, Cu, and Pb
by soils. Thus, both heavy metal solubility and mobility is increased.
Dissolved organic matter originating from the biosolids may well have a
second effect in increasing metal mobility. That is, dissolved organic matter
molecules can also be sorbed to the inorganic component of soils (e.g.,
Al and Fe oxides) (Kalbitz et al., 2005; Shen, 1999) and this could partially
block potential sorption sites for metals thus tending to increase their
solubility and availability.
In drainage waters from biosolids-amended soils, the bulk of heavy
metals have been found to be associated with soluble organic matter.

0
1
2
3
4
5
6
Day 0
Day 23
0
0.2
0.4
0.6
0.8
1
0
Days of incubation
SolubleZn(mgkg−1
)
100Mg ha−1
20Mg ha−1
Control
Zn(µmolL−1
)
Control 20Mg ha−1
100Mg ha−1
50 100 150 200 250 300 350
Figure 4 Water-soluble Zn dynamics during incubation of amended and biosolids-amended soils and calculated activities of Zn-dissolved
organic matter species (mmol LÀ1
) at days 0 and 23. From Antoniadis et al. (2007); copyright American Society of Agronomy.

Using gel filtration chromatography, Dudley et al. (1987) found that in soil
extracts from 80–100% of water-soluble Cu, 48–100% of Zn, and 39–100%
of Ni was in organically complexed form. Using differential pulse anodic
stripping voltametry, McBride et al. (1999) determined that only 30% of
water-soluble Zn, 18% of Cd, and 10% of Cu was present as ionic or
inorganic complexes and the remainder was presumed to be complexed
with dissolved organic matter. Using the same method, Al-Wabel et al.
(2002) concluded that 99% of soluble Cu and Zn in leachates was present
in organically complexed form. Heavy metals have, however, also been
shown to be present in drainage water associated with suspended clay-sized
particles (Keller et al., 2002). The metals become adsorbed to the surfaces of
Fe oxide and layer silicate clays present in this leached particulate matter.
Keller et al. (2002) calculated that movement of particulate matter
accounted for about 20% of Cu, Zn, and Cd leaching from a biosolids-
amended soil.
An important factor thought to contribute to leaching of metals is
preferential flow of water and dissolved metals down the soil profile in
downward oriented macropores (e.g., cracks, earthworm channels, root
channels) (Camobreco et al., 1996; Keller et al., 2002; Lamy et al., 1993).
This water bypasses the soil matrix thus minimizing the chances that the
dissolved metals will be adsorbed to soil surfaces. Preferential flow is proba-
bly the main pathway of movement of suspended particulate matter and
associated metals (Keller et al., 2002).
The period of greatest risk of metal leaching is soon after biosolids
application. This is when soluble organic matter is present in high concen-
trations and when preferential flow down surface-connected macropores is
most likely. Indeed, leaching losses of metals are normally greatest during
this initial period (Antoniadis et al., 2007; Camobreco et al., 1996; Keller
et al., 2002; Lamy et al., 1993; Maeda and Bergstrom, 2000). For this reason,
it will be important to minimize water inputs (e.g., irrigation) and drainage
from soils immediately following land application of biosolids.
5.3.6. Soil microbial/biochemical effects
Elevated concentrations of heavy metals in soils are known to affect soil
microbial populations and associated activities (Baath, 1989; Brookes, 1995;
McGrath, 1994). Baath (1989) concluded that the following order of
toxicity to soil microbes is most commonly found (in mg kgÀ 1 values):
Cd Cu Zn Pb. However, he showed an enormous disparity between
individual studies as to the exact concentrations at which metals become
toxic. Giller et al. (1998) suggested that much of the variability in deriving
toxic concentrations of heavy metals occurs through comparison of results
from short-term laboratory incubation studies with data from long-term
exposures of microbial populations to heavy metals in field experiments.
This is because laboratory studies measure response to immediate acute

toxicity (usually from one large addition of metals) whereas monitoring of
long-term field experiments measures responses to long-term chronic tox-
icity which accumulates gradually.
Stress caused by heavy metal contamination typically has two interre-
lated effects on soil microbial communities. The first is a loss of structural
and functional diversity since toxicities can suppress and/or kill sensitive
parts of the community. Nevertheless, rediversification can occur in the
surviving tolerant communities (Barkay et al., 1985). The other is an
increase in respiration per unit of microbial biomass (metabolic quotient;
qCO2) which is thought to occur because stressed microorganisms direct a
relatively larger amount of available energy into maintenance of various
biochemical functions (Giller et al., 1998). Thus, in general heavy metal
contamination of soils has been shown to result in a decline in microbial
biomass C, an increase in metabolic quotient (Brookes, 1995; Giller et al.,
1998), and shifts in bacterial community structure (Frostegard et al., 1996;
Giller et al., 1998; Tom-Petersen et al., 2003). There are also often negative
effects on soil enzyme activity (Belyaeva et al., 2005; Kizilkaya and Bayrakli,
2005). Enzyme reactions can be inhibited by heavy metals through a
number of mechanisms including by (i) complexing with the substrate, (ii)
combining with the protein-active groups of the enzymes, or (iii) reacting
with the enzyme–substrate complex (Dick, 1997).
In the case of biosolids application to soils, the addition of organic
material increases organic matter content and consequently the size
and activity of the microbial community also tend to be stimulated
(Section 3.1.2). However, if biosolids contain a high heavy metal load
then metal toxicities may have an inhibitory effect on soil microbial activity.
Indeed, many workers have observed an inhibitory effect in soils where
biosolids high in heavy metals have been applied and these negative effects
can remain for decades after application (Giller et al., 1998; Stoven et al.,
2005).
Numerous short- and long-term studies have been carried out where
biosolids contaminated with one or more heavy metals (or biosolids
enriched with one or more heavy metals) have been applied to soils and
the size and activity of the microbial community measured. Short-term
incubation experiments have generally shown a reduction in microbial
biomass C and N, usually an increase in metabolic quotient and a variable
effect on enzyme activity (Bhattacharyya et al., 2008; Kao et al., 2006; Rost
et al., 2001). Long-term (8 years) field trials have shown similar results
with a depression in microbial biomass C and microbial biomass C
expressed as a percentage of organic C and an increase in metabolic quotient
(Bhattacharyya et al., 2008; Chander and Brookes, 1991; Fliebßach et al.,
1994; Stoven et al., 2005; Zhang et al., 2008). Zhang et al. (2008) sampled
soils in fields that had been irrigated with heavy metal contaminated
wastewater (polluted with Cd and to a lesser extent Zn and Cu) for

30 years along a gradient of increasing total soil Cd content (1–4) (Table 4).
Concentrations of extractable Cd, Cu, and Zn and metabolic quotient
generally increased along the gradient while microbial biomass C declined
(Table 4). Observed effects on soil enzyme activities have been variable with
Bhattacharyya et al. (2008) observing reductions in glucosidase, urease,
phosphatase, and sulphatase activities induced by high combined concen-
trations of Cd, Cr, Cu, and Pb, Zhang et al. (2008) finding dehydrogenase
and phosphatase activities were not consistently affected by a combination
of high Cd, Cu, and Zn (Table 4) and Stoven et al. (2005) finding dehydro-
genase activity was decreased but that of phosphatase was unaffected by high
combined concentrations of Cr, Cd, Cu, Hg, Ni, Pb, and Zn.
Not only is the size and activity of the soil microbial community affected
by heavy metal contamination originating from biosolids but also its com-
position is altered (Macdonald et al., 2007; Sandaa et al., 1999a,b). Biolumi-
nescence-based bacterial and fungal biosensors can be used to assay the
potential toxicity of water-soluble contaminants in soils and this technique
was employed by Horswell et al. (2006) to determine the effects of Cu-,
Ni-, and Zn-spiked biosolids on the microbial community in the litter layer
of a forest soil. They found that increased Cu caused a decline in biolumi-
nescence response of the fungal biosensor, increased Zn caused decline in
response of the bacterial biosensor while increased Ni had little effect on
either. In a 10-year field experiment where plots received different con-
centrations of biosolids spiked with a combination of Cd, Cu, Ni, and Zn,
molecular techniques were used to show that significant differences, and
decreased diversity, were induced in both bacterial (Sandaa et al., 1999a,
2001) and archaeal (Sandaa et al., 1999b) community structures. Using
molecular techniques Macdonald et al. (2007) showed that in an 8-year
study using Zn-spiked biosolids there were significant differences in micro-
bial community structure for all groups investigated (bacteria, fungi,
archaea, actinobacteria, and rhizobium/agrobacterium). Their results
showed that fungi, and to a lesser extent archaea, were more negatively
affected by Zn addition than was the bacterial community. Results from
several long-term experiments have shown that Rhizobium leguminosarum, a
N2-fixing symbiotic bacteria of white clover, is considerably more sensitive
to the toxic effects of heavy metals than the host plants and that the host
plant confers protection from metal stress to the rhizobium (Chaudri et al.,
1993; McGrath et al., 1995). The toxic effect is due to toxicity to the free
living rhizobium particularly in response to high Zn (Chaudri et al., 2008).
Thus, the general effect of heavy metal contamination of soils induced
by biosolids applications is a decrease in the size of the microbial commu-
nity, an increase in metabolic quotient, a change in species composition,
and often a decrease in activity of key enzymes involved in C, N, P, and S
transformations. Such decreased enzyme activity will tend to reduce the
turnover of C, N, P, and S in the soil. The potential effect of a change in

Inorganic and Organic Constituents and Contaminants of Biosolids: Implications for Land Application

Inorganic and Organic Constituents and Contaminants of Biosolids: Implications for Land Application

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