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.

Author’s personal copy

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Inorganic and Organic Constituents

and Contaminants of Biosolids:

Implications for Land Application

R. J. Haynes,* G. Murtaza,†,‡ and R. Naidu§

Contents

1. In

s in

065

ol oe foeerialiateCA

troduction

Agronomy, Volume 104 # 2009

-2113, DOI: 10.1016/S0065-2113(09)04004-8 All rig

f Land, Crop and Food Sciences/CRC CARE, The University of Queensland, St Lucr Environmental Risk Assessment and Remediation, Division of Information Technong and the Environment, University of South Australia, Mawson Lakes Campus, South

of Soil and Environmental Sciences, University of Agriculture, Faisalabad, PakistanRE, Salisbury, South Australia, Australia

Else

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ia,logAus

166

2. S
ewage Treatment Processes 168
3. C
omposition of Biosolids 169
3
.1. O rganic matter 169
3
.2. In organic components 174
4. N
utrient Content and Release 175
4
.1. N itrogen 175
4
.2. P hosphorus 179
4
.3. O ther nutrients 181
5. H
eavy Metal Contaminants 182
5
.1. T otal concentrations 183
5
.2. E xtractable fractions 185
5
.3. A pplication to the soil 187
5
.4. P lant response and metal uptake 202
5
.5. In gestion by animals 207
6. O
rganic Contaminants 208
6
.1. O rganic compounds present 211
6
.2. P otential transfer to groundwater, plants, and animals 227
7. S
ynthesis and Conclusions 234
Refe
rences 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

vier Inc.

reserved.

Australiay,tralia,

165

166 R. J. Haynes et al.


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 isprimarily derived from domestic sources being a combination of humanfeces, urine, and graywater (from washing, bathing, and meal preparation).Sewage also contains discharges from commercial and industrial enterprisesand often some stormwater. As the wastewater is treated, it goes through aseries of processes that reduce the concentrations of organic material thatwere originally present. Primary sludge (principally fecal material) resultsfrom settling of solids as they enter the treatment plant. Secondary sludgeoriginates from the conversion of suspended and soluble organic matter insewage into bacterial biomass. The biomass is removed and combined withthe primary sludge to produce material termed sewage sludge. This materialthen undergoes treatment (usually anaerobic but sometimes aerobic diges-tion) to reduce the volume and stabilize the solid organic matter componentas well as to reduce the presence of disease-causing organisms. The finalproduct 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 isgenerally seen as the most economical and beneficial way to deal with

Inorganic and Organic Constituents and Contaminants of Biosolids 167


biosolids (Shammas and Wang, 2007a). Indeed, about 60% of all biosolidsproduced in both United States and United Kingdom are land applied(Pepper et al., 2006). Biosolids contain organic matter and nutrients andwhen applied to farmland can improve productivity and reduce the need formanufactured fertilizer inputs (Singh and Agrawal, 2008). Biosolids havealso been used successfully as a topsoil substitute for landscaping (Wu, 1987)and to enhance revegetation process on disturbed sites (e.g., mined land andtailings dumps) (Sopper, 1992). The organic matter acts as a soil condi-tioner, improving soil physical conditions and stimulating soil microbialactivity while macro- and micronutrients present serve as a source of plantnutrients. However, there are potential hazards with land application since arange of contaminants can be present in biosolids including heavy metals,recalcitrant organic compounds, and pathogens (Hue, 1995; Jenson andJepsen, 2005; Mininni and Santori, 1987; Pepper et al., 2006; Singh andAgrawal, 2008). Their presence greatly influences public perceptionsregarding 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 andenvironmental aspects of land application is testament to the relevance andimportance of the topic. Several workers have reviewed agronomicand environmental aspects of land application of biosolids (During andGath, 2002; Epstein, 2003; Hue, 1995; Singh and Agrawal, 2008) and thepresence of pathogens in biosolids was recently discussed (Pepper et al.,2006). However, a detailed understanding of the nature and content ofbiosolids, and how this develops during sewage treatment, helps greatly inpredicting their effects on the soil and the wider environment. In thischapter we provide an overview of findings on the nature of inorganicand organic constituents and contaminants of biosolids in relation to theimpact that land application has on soil properties, crop growth, and thewider environment.

Biosolids are well characterized materials and the nature and content oforganic and inorganic constituents, their nutrient content, and nutrientrelease characteristics are well documented and are reviewed here. Simi-larly, voluminous literature exists on the fate of contaminant heavy metalsduring wastewater treatment and, more particularly, the fate of biosolids-borne heavy metals in soil following land application. Consequently, anoverview of this information is also presented here. By comparison, researchinto organic contaminants in biosolids is in its infancy and the majority ofstudies are surveys of the presence and concentrations of various compoundsfound in a range of biosolids samples. Current knowledge on the occur-rence of organic contaminants is therefore reviewed and using the scarcedata that exists, their fate during wastewater treatment and in the soil afterland application of biosolids is discussed.



2. Sewage Treatment Processes

Prior to treatment, the influent sewage water is screened to removelarge objects and then undergoes grit removal in which heavy inorganiccoarse, sand-like, material is removed by settling. The water is then pumpedto large sedimentation tanks where it undergoes primary treatment. Thisinvolves sedimentation in which most of the suspended solids are removedas sludge material which sinks to the floor of the tanks. The sludge is thenpumped as a slurry (primary sludge) to storage tanks. The liquid remainingenters secondary treatment which is designed to degrade the remainingdissolved and colloidal organic content in the sewage.

During the secondary stage, most of the organic matter remaining in thewaste water is consumed by microbes under aerobic conditions. This isaccomplished by bringing together wastewater, bacteria (and othermicrobes), and oxygen and can be achieved by either fixed film orsuspended growth systems. In fixed film methods (e.g., trickling filters androtating biological contactors) the microbial biomass grows on a mediumand the sewage passes over its surface. The microorganisms remove andoxidize the organic material. The most common suspended growth systemis the activated sludge process. Primary-treated sewage combined withmicroorganisms is aerated by bubbling O2 through a tank. A biologicalfloc (composed of saprophytic bacteria and associated protozoa and rotifers)develops which removes and oxidizes the organic material. The treatedsupernatant is runoff and a portion of the settled sludge is returned to thehead of the aeration system to reseed the new sewage entering the tank.Secondary treatment commonly removes about 60–90% of dissolved andsuspended organic matter. The waste sludge from this process (secondarysludge) consists predominantly of saprophytic bacterial biomass, some othermicroorganisms and adhering microbial by-products. It is removed andnormally mixed with the sludge from the primary treatment process.

The accumulated sludges are then treated before disposal. Treatmentsusually include thickening, stabilization, and then dewatering. Thickeningis used to increase the solids content and reduce the volume that needs to behandled. It increases the solids content of sludge from 1–2% to 4–5% andcan reduce volumes to as low as 20% of unthickened sludge. The mostcommon stabilization treatments are anaerobic and aerobic digestion. Thesludge is digested to reduce the amount of organic matter and the number ofdisease-causing microorganisms present in the solids. In anaerobic digestion,(Taricska et al., 2007) sludge is passed into a closed container held at eitherthe mesophilic (e.g., 36 �C) or thermophilic range (e.g., 55 �C). Bacteriadecompose 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 andbacterial activity breaks down organic matter to produce CO2 and it alsogenerates heat to kill pathogens (Shammas and Wang, 2007b).

Other lesser used stabilization methods include lime stabilization andthermal treatment. Lime stabilization involves mixing the sludge with limeto achieve a pH of 12 or more and maintaining it for 2 h or longer. Thealkaline conditions produced drastically reduces microbial activity andcauses death of many pathogens. Thermal treatment subjects the sludge tohigh temperatures (e.g., 150–180 �C) and pressures up to 3 mPa in a closedreaction vessel. This results in rupture of cell walls of microorganismspresent (including pathogens) and causes chemical oxidation of organicmatter.

Following digestion, the treated sludge is often dewatered to reduce thevolume and mass for transport. Belt filter presses, vacuum filtration, orcentrifugation are used to increase the solids content of sludge to 25–45%whereupon the material takes on the properties of a solid rather than aliquid. It can also be composted to further reduce volume, produce a morestabilized product, and reduce the incidence of pathogens (Parr et al., 1978).Composting usually involves blending dewatered biosolids with a bulkingagent (e.g., bark chips) and composting the product in windrows. Heat isgenerated during the intense microbial activity of composting and thermo-philic temperatures (�55 �C) can be reached which cause death of manypathogenic organisms.

3. Composition of Biosolids

3.1. Organic matter

3.1.1. Nature of organic matterBiosolids samples are typically made up of 40–70% organic matter (asmeasured by loss of mass on ignition). They typically have an organic Ccontent ranging from 20–50%, a total N content of 2–5%, and a C/N ratioof 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 ofdecomposition 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, humicmaterial, and fatty acids. Live and deadmicroorganisms constitute a substantialproportion of the organic material and provide a large surface area forsorption of lipophilic organic contaminants in the sludge. Because much ofthe insoluble inorganic matter settles out during primary sedimentation,



the organic matter content of primary sludge (47–70%) is normally less thanthat of secondary sludge (62–82%) (Alvarez et al., 2002; Solis et al., 2002).The organic matter content of mixed sludge typically declines duringdigestion 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 organicmatter content of 65% for mixed sludge but a content of only 56% afteranaerobic digestion. There is a further decline (as much as 30–60%) inorganic 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 thecase if a slowly decomposable organic bulking agent (e.g., shredded barkchips) is added prior to composting (Nomeda et al., 2008).

Humification is a natural process by which plant and animal residuesdecompose in the soil and a dark colored, more or less stable portion oforganic matter remains. The humic materials remaining are high molecularweight organic molecules made up of a core of phenolic polymers producedfrom the products of biological degradation of plant and animal residues andthe synthetic activity of microorganisms (Stevenson, 1994). They exist asheterogeneous, complex, three-dimensional amorphous structures. Thehumic fraction of biosolids differs from that of soils because the former hasundergone a relatively short period of decomposition/transformation by atechnological process rather than a long-term transformation under naturalsoil conditions.

Characterization of humic substances is complex and involves a widerange of techniques including elemental and functional group analyses, gelfiltration 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 soilhumic substances, humic substances from biosolids are characterized bylower molecular weights, higher contents of S- and N-containing groups,lower C/N ratios and contents of acidic groups, much lower metal bindingcapacities and stability constants, a prevalence of aliphaticity, extendedmolecular heterogeneity, and lower degrees of polycondensation and humi-fication (Amir et al., 2004; Ayuso et al., 1997; Boyd et al., 1980; Leinweberet 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 twoseparate sources (primary and secondary sludge). For example, Smernik et al.(2003b) showed that organic matter in biosolids consisted of two spatiallyand chemically distinct ‘‘domains’’ derived from partially degraded plantmaterial (i.e., human feces) and bacterial residues, respectively. Results ofa comparative study of the humic substances from anaerobically andaerobically digested biosolids (Hernandez et al., 1988) showed that the



type of digestion process has little effect on elemental composition orfunctional 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 volumeof 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 thebiosolids-derived humic substances more closely approach those of nativesoil humic substances (Amir et al., 2004, 2005a; Garcia et al., 1991a;Jouraiphy et al., 2005; Sanchez-Monedero et al., 2002; Zbytniewski andBuszewski, 2005). For example, Amir et al. (2004) demonstrated that duringcomposting there was a steady decrease in C content, a more substantialdecrease in N content, an increase in C/N ratio, and a decrease in aliphaticcompounds 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., aliphaticside chains of lipidic and N-containing peptide structures) occurs. This leadsto 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 Ccontent (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 substancescan sometimes increase during the initial stages of composting (Zbytniewskiand Buszewski, 2005) as more complex organic substances are broken downand, in addition, organic metabolites are excreted by the decomposermicrobial community. However, over the composting period (usually50–150 days), there is typically an overall decline in soluble C concentrations(both absolute concentrations and those as a percentage of total organic Ccontent) until they account for about 1–2%of organicC (Garcia et al., 1991b;Zbytniewski and Buszewski, 2005). Indeed, a decline in water-solubleorganic C is often used as an indicator of compost maturity since freshcompost consists of many easily degradable and water-soluble substances,whereas mature compost is rich in stable, decomposition-resistant, highmolecular weight, humic substances (Zmora-Nahum et al., 2005).

3.1.2. Application to the soilFollowing 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 someconcomitant decomposition of native soil organic matter (Terry et al., 1979).



Nevertheless, when biosolids are applied to soils at high rates and/orrepeatedly, there is typically a substantial increase in soil organic mattercontent (Gupta et al., 1977; Kladivko and Nelson, 1979; Moffet et al., 2005;Navas et al., 1998; Rostagno and Sosebee, 2001). The effect is particularlypronounced on degraded soils with a low initial organic matter content(Garcia-Orenes et al., 2005). Indeed, using 14C-labeled biosolids, Terryet al. (1979) showed that a major portion of biosolids-C was resistant todecomposition in the soil and had a turnover rate in the order of hundredsof years.

Not only is the soil organic matter content increased, but also the qualityof organic matter is changed. That is, as expected based on the abovediscussion, amending soils with biosolids generally causes an increase inaliphaticity and N, H, and S contents and a decrease of C/N ratios, O andacidic functional group contents and metal binding capacities of soil humicmaterials (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 amendedsoil humic substances return to those of the unamended soil since thebiosolids-derived humic materials undergo further humification andbecome incorporated within the soil humic fraction (Senesi et al., 2007).Amending soils with composted biosolids, however, has a much lesser effecton the characteristics of soil humic substances compared to uncompostedmaterial.

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 andRomanya, 2006). Han and Thompson (1999) also observed that the molec-ular weight distribution of soluble organic matter in soils shifted to lowerweights (e.g., <14,000 Da) following biosolids applications. The signifi-cance of dissolved organic matter to the mobility of biosolids-derived heavymetals is discussed in Section 5.3.5.

The cation exchange capacity (CEC) of the soil is often increasedfollowing 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 isattributable to the high CEC of biosolids organic matter conferred by themany negatively charged functional groups present on humic material. Theextent of the increase will depend on such factors as soil texture, initial soilorganic matter content and CEC, nature of biosolids, and period since lastapplication. Over time, there will be a subsequent decrease in CEC as theadded biosolids organic matter decomposes (Clapp et al., 1986).

The increase in organic matter content following biosolids applicationoften results in a concomitant improvement in soil physical properties



(Clapp et al., 1986; Khaleel et al., 1981). There is often an increase in waterstable aggregation (Epstein, 1975; Gupta et al., 1977; Kladivko and Nelson,1979; Pagliai et al., 1981) due to the binding properties of organic matterand the associated microflora. Because of increased aggregation, total porespace is typically increased resulting in measured deceases in bulk densityand 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 tooccur (Epstein, 1975; Gupta et al., 1977; Tsadilas et al., 2005) and as a resultthere can be decreased runoff and water erosion (Moffet et al., 2005;Rostagno and Sosebee, 2001). Water-holding capacity often increases atboth field capacity and wilting point (Kladivko and Nelson, 1979; Guptaet al., 1977; Table 1) but the amount of available water (held between fieldcapacity 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 increasein the size and activity of the soil microbial community as well as theactivities of soil enzymes. Such stimulation of microbial activity can occurfollowing biosolids applications and/or inhibitory effects can occur due tothe 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 applicationshave been recorded in both laboratory incubations and field studies. Forexample, in a two-month incubation experiment Dar (1996) showed thatbiosolids amendment at 0.75% increased soil microbial biomass by 8–28%,arginine ammonification rate by 8–12%, and dehydrogenase and alkalinephosphatase 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 organicmatter content and some soil physical properties

Biosolids

rate

(Mg ha� 1)

Organic

mattera

content (%)

Bulk

density

(g cm� 3)

Field

capacity

(g g� 1)

Wilting

point

(g g� 1)

Available

water

(g g� 1)

Final

infiltration

rate (cm h� 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 levelp< 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 incubationstudies (Gomah et al., 1990; Hattori, 1988; Kizilkaya and Hepsen, 2004;Topac et al., 2008).

In field experiments, increases in microbial biomass C and N, basalrespiration, metabolic quotient (qCO2), and FDA hydrolysis rate havebeen noted following biosolids applications (Fernandes et al., 2005;Garcia-Gil et al., 2004; Sanchez-Monedero et al., 2004) as have increasesin the activities of dehydrogenase, protease, urease, amylase, catalase,b-glucosidase and alkaline phosphatase (Fernandes et al., 2005; Furczakand 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 monthsfollowing biosolids applications (i.e., during the rapid phase of decomposi-tion (Garcia-Gil et al., 2004). Even where levels of heavy metals in biosolidsare high, there can be an initial increase in microbial activity during theinitial phase of decomposition which is then followed by a later phase wheremicrobial activity is inhibited (Kizilkaya and Bayrakli, 2005).

The stimulating effect on soil microbial activity of the application ofcomposted biosolids has been shown to be lower but more persistent thanthat of uncomposted biosolids ( Jimenez et al., 2007; Pascual et al., 2002;Sanchez-Monedero et al., 2004). Nevertheless, substantial increases inmicrobial biomass C and N, basal respiration rate, potentially mineralizableN, and the activities of some soil enzymes have been noted following fieldapplications 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 andRomanya, 2006). Han and Thompson (1999) also observed that the molec-ular weight distribution of soluble organic matter in soils shifted to lowerweights (e.g., <14,000 Da) following biosolids applications. The signifi-cance of dissolved organic matter to the mobility of biosolids-derived heavymetals is discussed in Section 5.3.5.

3.2. Inorganic components

The inorganic content of biosolids, as measured by ash content, commonlyranges 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 theeffective removal of many of the inorganic components from wastewaterduring primary and secondary treatment. The inorganic component ofbiosolids consists mainly silt- and clay-sized particles that arise from arange 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 originatingfrom plant material; 1–4%), cosmetics, and other products washed downresidential drains.

X-ray fluorescence analysis on dried sludge by Thawornchaisit andPakulanon (2007) indicated that oxides of Si, Al, and Fe (with a combinedtotal of 62%) were the three main inorganic constituents of biosolids. X-raydiffraction analysis of biosolids has been performed by a number of workers( Jaynes and Zartman, 2005; Mun et al., 2005; Sommers, 1977). Jaynes andZartman (2005) observed an inorganic matrix consisting mainly of Quartz(SiO2) and feldspars (crystalline minerals that consist of aluminum silicatescontaining K, Na, Ca, or Ba) and kaolinite, mica, and expandable clays werealso present. Sommers (1977) identified quartz, calcite, dolomite, feldspars,and layer silicates while Mun et al. (2005) found quartz was the dominantmineral but there were also significant amounts of feldspars, muscovite, andchlorite. In biosolids ash, Hartman et al. (2007) identified quartz andhaematite as the predominant minerals. Jaynes and Zartman (2005) alsofound significant amounts of poorly crystalline Al and Fe phosphates(thought to be formed during anaerobic digestion) and talc residuesoriginating from cosmetics.

4. Nutrient Content and Release

4.1. Nitrogen

The N content of biosolids can vary greatly (Sommers, 1977) but is typicallyin the range of 2.8–3.8% (Epstein, 2003; Hue, 1995). Accumulation of totalN in the surface soil, 15 years after an application of 500 ton ha� 1 ofbiosolids to a forest soil is evident in Fig. 1. Because 50–90% (often quotedas 80%) of N in biosolids is in organic form (Sommers, 1977), informationon the N mineralization rate is necessary to predict N availability followingland application. Because nitrification (the microbial conversion of NH4

þto NO3

�) is predominantly an aerobic process, in anaerobically digestedbiosolids the content of mineral N consists of about 99% NH4

þ–N and 1%NO3

�–N (USEPA, 1995). However, in aerobically digested biosolids thebulk of the mineral N is present as NO3

�–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 reportedmineralization rates of 4–48% in 16 weeks (Ryan et al., 1973), 14–25% in13 weeks (Magdoff and Chromec, 1977), 40–42% in 15 weeks (Epsteinet al., 1978), 15% in 16 weeks (Parker and Sommers, 1983), and 24–68%in 32 weeks (Lindermann and Cardenas, 1984). The N mineralized tends tobe greater from aerobically than anaerobically digested biosolids (Hseu and

20100 201000

50

100

150

0

50

100

150

Total N (mg g−1) Total P (mg g−1)

Sludge-treatedControl Control

Sludge-treated

Soi

l dep

th (

cm)

Figure 1 Total N and P concentration with depth in a forest soil treated with500 Mg ha� 1 municipal biosolids 15 years previous to sampling and in a control(untreated) plot. From Harrison et al. (1994); copyright Elsevier.



Huang, 2005; Magdoff and Chromec, 1977) and composting greatlydecreases biosolids-N mineralization potential (Epstein et al., 1978; Parkerand Sommers, 1983).

In biosolids, N mineralization potential has been related to total organicN content and more particularly to various indices of protein content.A large proportion of biosolids organic N is thought to be proteinaceousin origin and this fraction represents a labile pool of organic N (Hattori andMukai, 1986; Lerch et al., 1992). Hattori and Mukai (1986) found acorrelation between mineralization of biosolids-N and crude protein con-tent while Hattori (1988) found a correlation with proteinase enzymeactivity in the biosolids-amended soil. Lerch et al. (1992) also found acorrelation between N mineralization and low molecular weight amines(assumed to be proteins) in biosolids while Rowell et al. (2001) found acorrelation with the alkyl index and the alkyl toO-alkyl ratio (as determinedby solid state13C NMR spectroscopy). This was explained as a reflection ofproteins 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 thematerial as a whole.

In soils, N mineralization is carried out by the heterotrophic microbialcommunity and is therefore highly dependent on environmental factorswhich affect microbial activity (e.g., soil type, temperature, water content,aeration). Thus, under field conditions, the proportion of the potentially



mineralizable pool of organic N that is actually released will be highlyvariable 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 that20%, 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 ofa number of factors including biosolids composition and rate of application,soil type, pH, soil temperature, soil water content, and aeration (Artiola andPepper, 1992; Barbarick et al., 1996; Sims and Boswell, 1980). Based onfield trials in Wisconsin, Keeney et al. (1975) suggested an organic N decayrate series of 15–20%, 6%, 4%, and 2% for the first, second, third, and fourthyears 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 trialsin 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 areconservative and that often more than 20% of biosolids organic N ismineralized in the first year (Barbarick and Ippolito, 2000; Barbaricket al., 1996; Cogger et al., 1998).

The agronomic response to applied biosolids-N will be greatly affectedby a range of environmental and soil conditions. Binder et al. (2002), forexample, showed in a series of field trials that irrigated maize yield responsewas relatively consistent between years with maximum yields being attainedat about 441 kg organic N ha� 1 (Fig. 2). However, dryland sorghum yieldswere less consistent. In 1996, there was no significant yield response becauseof high residual soil NO3

� and mineralizable N originating from a previoussoybean crop and a previous 3-year fallow (Fig. 2). Yields in 1997 and 1998were similar and considered representative of more common rotations andclimatic conditions in south east Nebraska. In 1999, cool weather restrictedN mineralization rate and sorghum responded to much higher rates ofbiosolids-N (Fig. 2).

For anaerobically digested biosolids, the NH4þ initially present and that

which is ammonified soon after application is at risk of volatilization loss ifbiosolids are surface applied. Ammonia volatilization is favored when highconcentrations 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 tendsto favor volatilization and losses ranging from 25–80% of the initial NH4

þcontent have been recorded (Adamsen and Sabey, 1987; Beauchamp et al.,1978; Robinson and Polglase, 2000; Robinson and Roper, 2003; Terryet al., 1981). Incorporation of biosolids into the soil will minimize suchlosses. Over a period of several weeks following biosolids application,nitrification will typically proceed induced by indigenous autotrophicnitrifier bacteria present in the soil.

Typical year 1997, 1998

After soybean 1996

Cool/dry year 1999

Sorghum

Maize

Rel

ativ

e yi

eld,

%

100

90

80

70

60

50

100

90

80

70

60

50

40

30100 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 tothe amount of organic N applied with biosolids in the year of application. From Binderet al. (2002); copyright American Society of Soil Science.



It is important that the rate of biosolids-N supply matches crop Nrequirements (i.e., that an ‘‘agronomic biosolids rate’’ is used; USEPA,1993) since excess N will accumulate in the soil profile as the mobileNO3

� anion. This can be lost from the soil as N2/N2O via denitrificationunder anaerobic soil conditions or can be leached down the profile intogroundwater. Indeed, a frequently quoted hazard of biosolids applications isexcessive movement of NO3

� to groundwater (Keeney, 1989). To estimate



an agronomic rate that supplies the amount of N required by the crop andminimizes the amount of residual NO3

� available for leaching, the poten-tially available N (PAN) concentration may be calculated:

PAN ¼ NNO3þ XNNH4

þ YNorg;

where X is the fraction of NH4 that does not volatilize and Y is the fractionof 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 forplant uptake and 100% of NH4 is also available (i.e., X¼ 1) unless biosolidsare 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 oftenestimated at 0.20 in the year of application. Pierzynski (1994) suggestedfigures of 0.25 for aerobically digested sludge, 0.15% for anaerobicallydigested sludge, and 0.05–0.10 for composted biosolids.

Several workers have developed models specifically to describe NO3�

leaching from biosolids-amended soils (Andrews et al., 1997; Joshua et al.,2001; Vogeler et al., 2006). However, in general, applications of biosolids atagronomic rates cause minimal NO3

� leaching (Correa et al., 2006; McLarenet al., 2003; Surampalli et al., 2008). The greater the proportion of biosolids-N initially present in NH4

þ form (which is rapidly nitrified following soilapplication) the greater the potential for NO3

� leaching since there is moreNO3

� in the soil profile (Shepherd, 1996; Smith et al., 1998). Deep injectionof biosolids exacerbates leaching losses because less drainage is required toleach N below the root zone (Shepherd, 1996). Timing of applications willbe an important consideration so that N supply from biosolids is in syn-chrony with crop uptake requirements. For example, applying biosolids inautumn prior to winter rains (during a period where crop growth and Nuptake is slow) is likely to favor leaching losses of NO3

� (Shepherd, 1996).Nitrogen mineralization will occur whenever conditions are favorablewhich on an annual basis is likely to be over a longer period than that forN uptake by the crop. As a result, mineral N will inevitably be producedduring periods when there is little chance of plant uptake. It will thereforebe advisable, where repeated biosolids applications are being made, tomeasure soil profile mineral N prior to biosolids applications and reducethe 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 theP (>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-tricalciumphosphate (Shober et al., 2006). Using combined sequential chemicalextraction, 31P NMR and XANES, Ajiboye et al. (2007) concluded thatreadily soluble P forms in biosolids mainly originated from easily soluble Caand 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 soilbefore it is plant available. In lime-stabilized biosolids, recalcitrant calciumphosphates (e.g., hydroxyapatite, tricalcium phosphate) become majorcomponents (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, agriculturalcrops sequester about four times as much N as P leading to an overallincrease in soil P in relation to N. Pierzynski (1994) calculated that if atypical biosolids sample (containing 13 g kg� 1 PAN and 10 g kg� 1 total P)were applied to supply 150 kg N ha� 1, it would also apply 115 kg P ha� 1

which is approximately three times more than would typically be recom-mended for maize. The imbalance between N and P in biosolids typicallyleads to a substantial increase in extractable soil P levels (Kelling et al.,1977b; Maguire et al., 2000; Peterson et al., 1994), often to levels muchgreater than those necessary for adequate P nutrition of crops. This can leadto 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 soilis clearly illustrated in Fig. 1.

Current recommendations in both United Kingdom and United Statesare that the relative effectiveness of biosolids-P, compared to solublefertilizer P, is 50% (MAFF, 1994; USEPA, 1995). O’Connor et al. (2004)assessed phytoavailability of 12 different biosolids samples in a greenhousestudy, relative to triple superphosphate (TSP), and confirmed that mostbiosolids produced by conventional methods had a relative phytoavailabil-ity in the range of 25–70% TSP. Biosolids produced in water treatmentplants where Fe, Al, or Ca is added during treatment to lower soluble P (tomeet effluent limitations) have a lower P availability (i.e., <25% TSP)(O’Connor et al., 2004). Indeed, in such biosolids, the solubility andavailability of P is characteristically low (Lee et al., 1981; Lu andO’Connor, 2001, Maguire et al., 2000; Soon and Bates, 1982) since thephosphate is strongly adsorbed to the surfaces of Fe and Al hydrous oxidesand 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 toproduce low P concentrations in the treated effluent wastewater. Itemploys aerobic and anaerobic zones to selectively enrich for bacteria



which take up large amounts of phosphate and store it intracellularly aspolyphosphate 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 surfaceapplications of biosolids have been practiced, runoff of particulate matterhigh in P is a potential danger since P inputs to aquatic freshwater systemscan increase the rate of eutrophication (Carpenter et al., 1998). The higherthe water-soluble P content of biosolids, the greater the risk of runoff lossesof P (Elliott et al. (2005).

Due to its strong adsorption onto soil colloids, it is usually consideredthat 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 sorptioncapacity) with a low pH (because of increased P solubility) and/or wheresoils have become P saturated, especially following heavy animal manureapplications (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 (Heckrathet al., 1995; Hesketh and Brookes, 2000; McDowell et al., 2001). Suchleaching is thought to occur principally by macropore flow (e.g., in cracks,earthworm burrows, and root channels) and much may be as particulateorganic matter and as phosphate sorbed to clay particles. Indeed, particle-facilitated transport of P has been found to play an important role inP 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 pointvalues, due to repeated biosolids applications, could therefore induceincreased 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 cmsoil 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 comparisonwith that for N, yet demand for it by crops is often comparable. For thatreason, biosolids is generally considered a poor source of K and supplemen-tary fertilizer K applications often need to be made. The reason for this isthat most K compounds are water soluble and remain in the sewage effluentor aqueous fraction during sludge dewatering. Nevertheless, the K inbiosolids 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 tothat of animal manures (Hue, 1995). Biosolids also supplies micronutrientssuch as B, Cu, Zn, Mn, Fe, Mo, and Ni (Epstein, 2003) and this may beimportant where micronutrient deficiencies occur in the soils where land



application is being practiced. Nevertheless, as discussed below, metals suchas Zn and Cu may sometimes be present in biosolids at levels that areconsidered unacceptable.

Addition of biosolids also results in an increase in electrical conductivity(EC) in soil solution (increased salinity) and alterations to soil pH (Clappet al., 1986). The EC of biosolids can be measured in a number of differentways including directly on the wet sludge, or after drying in either satura-tion paste extracts or 1:5 solid: water extracts. This contributes to variabilityin reported values which generally lie between 3 and 12 dS m� 1 (Garcia-Orenes et al., 2005; Moffet et al., 2005; Navas et al., 1998; Rostagno andSosebee, 2001). Such values are generally considerably greater than thoseencountered in nonsaline soils (i.e., 0–2 dS m� 1 in saturation paste extractsand 0–0.15 dS m� 1 in 1:5 soil: water extracts). The high EC in biosolids isattributable to the high concentrations of ions such as Mg2þ, Ca2þ, and Cl�that are present. During heavy rains/irrigation, soluble salts will leach downbelow the root zone and EC in the surface soil will return to that prior tobiosolids application.

Increases, decreases, and no effect of biosolids application on soil pHhave been noted (Clapp et al., 1986; Epstein, 2003; Singh and Agrawal,2008). Changes will be dependent on many soil and biosolids propertiesincluding the initial pH and buffering capacity of both materials. Thebuffering capacity of the biosolids will be largely controlled by factorscontributing to the CEC of the material and the content of Ca and Mgoxides. The initial pH of biosolids varies greatly but can often be in therange 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 whilethat of alkaline soils (e.g., >8) will tend to be decreased. However, in arange of soils a progressive decline in pH following biosolids application hasoften 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 ofnutrients 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 andsemimetals (often defined as having an atomic number greater than 20or 21) that have been associated with contamination and/or potentialtoxicity 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 canaccumulate 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 industrialproducts may, at the end of their life, be discharged as wastes. Key urbaninputs include drainage waters, business effluents (e.g., car washes, dentaluses, 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 (Bergbacket al., 2001; Comber and Gunn, 1996; Sorme and Lagerkvist, 2000).Household effluents can also be important. For example, at an Englishtreatment works, Comber and Gunn (1996) found domestic inputs ofCu 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 camefrom 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 insewage varies enormously between treatment works and depends greatlyon local factors such as type and number of industries in the region,regulations regarding the quality of industrial discharges allowed to sewersand public awareness of the environmental impacts of metal contaminateddischarges. Heavy metal content of sewage often fluctuates due to irregularinputs 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 onetreatment works can also vary greatly in heavy metal loadings with time.

Although waste water treatment plants are expected to control thedischarge of heavy metals to the environment, they are chiefly designedfor 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 sedimentand are then removed with the sediment. During secondary treatment twomain processes lead to removal of metals. These are (i) bioaccumulation inwhich metals are accumulated into the living bacterial cells and (ii) biosorp-tion in which heavy metals are sorbed onto negatively charged sites onbacterial cell walls and on extracellular polysaccharide gels (Brown andLester, 1979; Urrutia, 1997). The heavy metals are then removed in themicrobial sludge which is mixed with the primary sludge. The heavy metalconcentrations in primary and secondary sludges (on a dry weight basis) aretypically 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 treatmentcan vary greatly for different metals in the same treatment plant as well asbetween 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%, Cd78–91%, Zn 60–94%, Ni 11–70%, and Cr 67–79%. Many factors influenceremoval 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, andbiological factors (Brown and Lester, 1979; Chipasa, 2003; Stovelandet al., 1979). Thus, removal efficiency is not a predictable property.

A large number of studies in many parts of the world have surveyed theheavy 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 beensummarized previously (Epstein, 2003; Hue, 1995). Taking account of thegreat variability in heavy metal inputs which occurs between water treat-ment plants, some ‘‘typical’’ concentrations of metals encountered in bio-solids samples (in mg kg� 1 values) are shown in Table 2. It is evident thatZn is commonly present in highest concentrations and that substantialconcentrations of Pb, Cu, and Cr are also often present. In the UnitedStates and Canada, heavy metal concentrations in biosolids (particularlythose of Cd, Cr, Pb, and Ni) have been shown to be decreasing during

Table 2 Typical concentrations of heavy metals commonlyencountered in biosolids

Element

Concentration

(mg kg� 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 toenforcement by municipalities of regulations regarding the maximum metalloadings in effluents that can be discharged into the sewerage system. As aresult, industrial pretreatment of effluents has become common. However,for Zn and Cu, concentrations in biosolids have remained similar over thelast two decades (Epstein, 2003) because, as noted previously, they are oftennot principally of industrial origin. While heavy metal concentrations inbiosolids have generally been decreasing and in most situations they arebelow regulatory limits (see below), their addition to soils still causesdisquiet. This is because, unlike organic contaminants, most heavy metalsdo not undergo microbial or chemical degradation and therefore elevatedconcentrations persist in the soil for extremely long periods of time.

Concerns regarding the heavy metal load in biosolids have resulted inguidelines and regulations being developed in many parts of the world toregulate land applications. These are generally based on the maximumallowable metal concentration limits (mg kg� 1 dry weight) in biosolidsand/or the allowable loading limits (kg ha� 1 yr� 1) of metals added inbiosolids to soil (Epstein, 2003). The most quoted limits are those of theUSEPA (USEPA, 1993) and the European Union also has its own standards.In general, USEPA and UE limits for metal concentration limits in biosolidsare broadly similar but maximum loading limits are generally lower for theEU guidelines. Nevertheless, limits can vary quite widely with countriessuch as Sweden, Denmark, Germany, and the Netherlands generally havinglower limits than USEPA or EU guidelines (Smith, 2001). USEPA metalconcentration limits in biosolids are: Zn, 2800; Cu, 1500; Ni, 420; Pb, 300;Cd, 39; and As, 41 mg kg� 1 (USEPA, 1993). USEPA regulations are riskbased and therefore provide an opportunity to modify values as betterscientific data becomes available (Epstein, 2003).

5.2. Extractable fractions

Total concentrations of heavy metals indicate the extent of contaminationbut provide little insight into the potential mobility or bioavailability ofthese metals once the biosolids are soil applied. Depending on their nature,individual metals are associated in a variable manner with different phasesmaking up the biosolids. Sequential chemical fractionation procedures arewidely used to characterize the forms of metals present (chemical specia-tion). These methods involve chemical extractions using a sequence ofreagents of increasing strength. For each reagent, a particular chemicalform(s) is assigned to the metals extracted. Drawbacks of these methodsinclude (i) lack of specificity, selectivity, and validation; (ii) postextractionreadsorption; and (iii) sensitivity to procedural variables (e.g., sample size,pH, temperature, contact time, concentration of extractant, etc.) (Kot andNamiesnik, 2000). Despite such limitations, sequential extractions are



considered the best available method of gaining knowledge on the forms inwhich metals are present in biosolids.

A wide range of sequential fractionation schemes have been proposed fordetermination of heavy metal forms present in biosolids (Kot andNamiesnik, 2000; Marchioretto et al., 2002; Sims and Kline, 1991; Tessieret al., 1979). One of the simplest and most commonly used methods today isthat specified by the Community Bureau of Reference (CBR) (Ure et al.,1993) in which the sample is extracted with (i) acetic acid to release theeasily available ‘‘exchangeable’’ forms present in soluble and exchangeableforms and those associated with carbonate phases, (ii) hydroxylammoniumchloride to release the ‘‘reducible’’ fraction associated with Fe andMn oxidecements and nodules (forms that could become available under anoxicconditions), and (iii) hydrogen peroxide to extract the ‘‘oxidizable’’ fractionthat is strongly bound to organic matter constituents. Following the sequen-tial extraction, the amounts remaining in the ‘‘residual’’ fraction (iv) aremeasured after digestion with aqua regia and these are considered to behighly unavailable and associated with residual solids that occlude metalsin their crystalline structures. The amounts present in fractions (i) and (ii) areconsidered ‘‘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; Sprynskyyet al., 2007; Wang et al., 2005, 2006a,b). To generalize from the results ofthese studies, Cu is typically found to be concentrated (about 80% of totalCu content) in the oxidizable fraction bound to organic matter. This is inaccordance with the high stability constant of the Cu complexes withorganic matter (Ashworth and Alloway, 2004). By contrast, Zn isdistributed preferentially (usually 40–60%) in the available exchangeableplus oxidizable fractions. Greater than 50% of total Pb content is typicallyfound in the residual fraction with substantial amounts (15–30%) also beingpresent in the oxidizable fraction. Ni and Cd have a similar distribution with60–70% of total content being present in the unavailable oxidizable andresidual forms (usually more or less equally distributed between the twofractions). Co is similarly distributed between unavailable and availablefractions with significant amounts (30–50%) being present in the organicfraction. Cr is concentrated in the unavailable forms (usually more than 90%of total content) with over 50% in the residual fraction and a significantproportion also organically bound. For Fe, 80–90% of total content is inunavailable forms with greater than 60% in the residual form and 10–20% inthe organic fraction. However, for Mn, 70–80% of total content is inavailable forms with greater than 50% in the exchangeable form. In sum-mary, Zn and Mn are the metals preferentially found in the mobile fractionsof 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 tothe organic components of biosolids.

Solis et al. (2002) showed that for all metals (on a mean basis) theavailable (exchangeable plus reducible) fractions were higher in secondarythan primary sludge. During anaerobic digestion of combined sludge therewas a general increase in the percentage of metals in the unavailableoxidizable and residual fractions and during composting of the biosolidsthere was a further increase in the percentage of metals present in theunavailable fractions. A number of other workers have followed heavymetal 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) observedsimilar results for Cr, Cu, Mn, Fe, Ni, and Pb. However, Nomeda et al.(2008) showed that available fractions of Pb, Zn, and Cd increased withtime but those of Cu decreased. Liu et al. (2007a,b) observed that duringcomposting, the available fractions of Pb and Zn increased while those ofCu, Ni, and Cr were little affected. Thus, although it is clear that heavymetal levels are concentrated during composting, the effects on distributionof metals among fractions are much less clear and may vary dependingon 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 becocomposted with an absorbent material such as zeolite (e.g., crushedclinoptilolite rock) added at 10–25%w/w. This results in substantialdecreases in the amounts of metals being present in the potentially availableexchangeable and reducible fractions (Sprynskyy et al., 2007; Zorpas et al.,2008) since the metals are adsorbed to the zeolite surfaces. Cocompostingwith a sodium sulfide/lime mixture (3%w/w) was also shown by Wanget al. (2008) to reduce the percentage of metals in the available fractions.A number of methods have also been developed to remove heavy metalsfrom contaminated biosolids prior to land application. These include chem-ical extraction, bioleaching, electroreclamation, and supercritical fluidextraction (Babel and del Mundo Dacera, 2006).

5.3. Application to the soil

5.3.1. Heavy metal extraction from soilsIt 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 ofbiosolids constituents that adsorb metals limits the usefulness of total metalcontent 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 asestimated by the TCLP leaching procedure. Indeed, biosolids applicationto the soil not only increases the concentrations of heavy metals present butalso 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 organiccomponents 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 fordifferent 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 selectivelyextract metals associated with particular soil phases (Ure et al., 1993).Despite the limitations of such fractionation schemes, their use gives someindication of the fate of biosolids-borne heavy metals once they enter thesoil 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 andinorganic soil constituents.

A wide range of soil test extractants have been employed to determineheavy metal availability (McLaughlin et al., 2000a; Ure, 1995). The mostcommonly 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 theUnited Kingdom. Correlations between DTPA- and EDTA-extractablemetals and metal uptake by crops are generally reasonable (Bidwell andDowdy, 1987; Brun et al., 1998; Hooda et al., 1997; Hseu, 2006; Sanderset al., 1986, 1987; Sukkariyah et al., 2005a). Dilute acids (e.g., 0.05–0.1 MCH3COOH, 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; Sukkariyahet al., 2005a). These latter salt solutions extract metals in soil solution plusthose in short-term equilibrium with that solution. Complexing reagentsand dilute acids extract larger amounts of metals which include a ‘‘poten-tially available’’ fraction. They, in affect, overestimate phytotoxicity andassess potential rather than immediate toxicity (McLaughlin et al., 2000b).

McLaughlin et al. (2000b) suggested that in the future regulations andguidelines should consider extractable fractions of heavy metals in soils.That is, it is the concentration of biologically active (extractable) heavymetals present in biosolids-treated soil that is toxic to plants and soil biota(Merrington et al., 2003), yet present regulations are based on total loadingsof metals (see, Section 5.1). McLaughlin et al. (2000b) considered thatmetals extracted with dilute salt solutions and those extracted with more



harsh reagents (complexing agents or dilute acids) could be used together toestimate immediately toxic and potentially toxic metals, respectively.Certainly, extractable metal concentrations are likely to give a betterindication of bioavailability than values based on total concentrations.Monitoring of extractable metal levels on long-term sites, where biosolidsapplications are continuing and/or have been terminated, will give valuabledata on the long-term trends in bioavailability of various total loadings. Suchdata could well be used in the future to develop guidelines and regulationsbased on extractable soil metal levels.

5.3.2. Effects of biosolids properties on availabilityFollowing land application, the properties of the biosolids effect metalavailability both directly (through heavy metal content and sorptive capacityof inorganic and organic components) and indirectly (through propertiessuch as pH, mineralizable N content, and EC) (Merrington et al., 2003). It isusually assumed that biosolids properties dominate metal bioavailability inthe short and medium term in the zone of incorporation but with time,biosolids properties have progressively less influence and soil propertiesultimately control availability (Smith, 1996). The effect of biosolids materi-als on heavy metal retention by amended soils is complex and this is at leastpartially because a suite of metals is added, and competition between themfor adsorption sites occurs. Bergkvist et al. (2005), for example, found Cdsorption was slightly smaller in biosolids-amended soils compared to controleven though organic C content was 70% higher and oxalate-extractable Fewas roughly doubled. They attributed this to competition for sorption sitesbetween Cd and biosolids-derived Fe and other metals such as Zn. McBrideet al. (2006) found that addition of high Fe, high Al, and biosolids to soilshad no long-term effect on their affinity for Cd. By contrast, Vaca-Paulinet al. (2006) observed that biosolids-amended soils showed increasedadsorption capacity for Cu and Cd and attributed this to the complexingability of the biosolids-derived organic matter.

Strong metal retention by the inorganic fraction is attributable to thehigh adsorption capacity of Fe, Al, and Mn hydrous oxides and silicates(Basta et al., 2005; Merrington et al., 2003). The inorganic solids present inbiosolids are initially present, at least partially, in noncrystalline form(Baldwin et al., 1983; Rogers and McLaughlin, 1999) and the higher surfacearea of noncrystalline Fe and Al oxides results in them having a higheradsorption capacity than their crystalline counterparts (Rogers andMcLaughlin, 1999). In general, the order of affinity of metals for adsorptionsurfaces on Al and Fe oxide surfaces follow the order Cu>Pb>Zn>Co >Ni>Cd although for Fe oxides Pb>Cu has been reported andsometimes also Ni>Co ( Jackson, 1998; Sparks, 2003). In addition, car-bonate, phosphate, and sulphite present in biosolids can form sparinglysoluble solid phases with many metals and thus account for a substantial



portion of some metals present biosolids (Karapanagiotis et al., 1991). Forexample, during anaerobic digestion, low solubility Cu and Zn sulfidescharacteristically form (Nagoshi et al., 2005).

The organic component also has the ability to bind to heavy metals. Theheterogeneous nature of humic substances and the large number of func-tional groups present means that binding of metals can be regarded asoccurring at a large number of reactive sites with binding affinities thatrange from weak forces of attraction (ionic) to stable coordinate linkages(McBride, 2000; Sparks, 2003). Indeed, mechanisms involved in metalbinding to organic matter are complex and probably involve simultaneouschelation, complex formation, adsorption, and coprecipitation (Stevensonand Vance, 1989). Because of the many variables involved, there are manyinconsistencies in reported selectivity orders of metals with organic matter.A generalized order is Cr3þ>Pb2þ¼Hg2þ>Cu2þ>Cd2þ>Zn2þ¼Co2þ>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 decompositionphase. It has been suggested that heavy metals bound to biosolids organicmatter could be released to soil solution during decomposition and as aresult 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 biosolidsadditions and this is followed by a reduction in availability (as estimated bymetal extractability and/or plant uptake) as well as a reduction in organicmatter content (Bidwell and Dowdy, 1987; Hseu, 2006; Logan et al., 1997;McBride et al., 1999; Walter et al., 2002). Nonetheless, the initial highavailability may well be partially due to the rapid decomposition of biosolidsorganic matter and the consequent release of metals. Evidently, the metalsreleased 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 metalcations become increasingly immobile at high pH. This is because boththeir adsorption onto reactive oxide surfaces and precipitation reactions arefavored at high pH (Sparks, 2003). As noted in Section 4.3, since the initialpH of biosolids is typically in the range of 6–8, their application will have aliming effect on acid soils thus raising their pH (Kidd et al., 2007) andtending to reduce metal availability.

The mineralizable N content of biosolids is, however, an importantproperty in relation to their effects on soil pH. During ammonification oforganic N to NH4

þ–N, one OH� ion is released per unit of N while duringnitrification of NH4

þ–N to NO3�–N, two Hþ ions are released. The overall

process of conversion of organic biosolids-N to NO3�–N is therefore

acidifying. Thus, Hooda and Alloway (1994) observed a progressive



decrease in soil pH following biosolids application to soil which wasaccompanied by an accumulation of soil NO3

�–N and an increase in uptakeof Cd, Ni, Pb, and Zn by ryegrass growing in the soil. Such an increase inmetal bioavailability accompanying acidification induced by nitrification ofbiosolids-derived N has also been observed by others (De Haan, 1975;Hooda and Alloway, 1993). It is therefore important to monitor pH andapply 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 anincrease in soluble salts in soil solution. High soluble salts will tend to reducesoil solution pH (by exchange between cations in soil solution and Hþ andAl3þ on soil cation exchange sites) thus increasing the solubility of heavymetal cations. In addition, high concentrations of solution Cl� can increasemobilization, availability, and plant uptake of Cd through the formation ofCd–chloro complexes (Weggler-Beaton et al., 2000).

5.3.3. Effects of soil properties on availabilitySoil properties such as pH, redox potential, EC, clay, hydrous oxide, andorganic matter content will also influence heavy metal availability. Themost widely recognized factor is soil pH. With the exception of As andSe, heavy metal retention by soils increases with increasing pH (McBride,1994). As noted above, with an increase in pH, the charge on the variablecharge adsorption surfaces (e.g., Fe, Al, and Mn hydrous oxides) becomesincreasingly negative thus favoring metal cation adsorption and the high pHalso favors surface precipitation of the metals onto the surfaces (Bradl, 2004;McBride, 2000). In general, the more mobile metals such as Ni, Cd, and Znare more sensitive to increasing pH than other metals such as Pb and Cu thatare more strongly complexed with soil organic colloids (Smith, 1996).

Manipulation of soil pH has been found to be the most effective way ofcontrolling heavy metal bioavailability in biosolids-treated soils (Allowayand Jackson, 1991). Indeed, a large number of workers have shown that thebioavailability of metals to plants in biosolids-amended soils decreases as pHis raised either by liming or applying lime-stabilized sludges (Basta andSloan, 1999; Milner and Barker, 1989; Oliver et al., 1998). Liming arange of biosolids-treated soils to pH 7 was shown by Jackson andAlloway (1991) to reduce Cd content of lettuce by an average of 41% andcabbage by 43%.

Redox potential is often considered an important factor although bothincreases and decreases in heavy metal solubility have been recordedfollowing 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 anumber of different processes occur following the onset of anaerobiosisand these often interact to affect metal solubility. In freely-drained soils,



Fe andMn occur in their high oxidation states as oxides and hydrous oxides.However, as soils become anaerobic, due to waterlogging, the redoxpotential decreases and oxide minerals begin to dissolve as soluble Mn2þand Fe2þ forms (Stum, 1992; Stum and Sulzberger, 1992). This can not onlyresult 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 becomeanaerobic the pH tends to converge to neutrality irrespective of initial pH,whether acidic or alkaline (McBride, 1994). For acidic soils this increase inpH can result in release of organic matter into soil solution and metalsbound to the organic molecules are also thought to be released (Gryboset al., 2007). This also tends to increase metal solubility. Nonetheless, theincrease in pH up to about 7, favors adsorption/surface precipitation ofmetal cations thus favoring removal of metals from solution (Kashem andSingh, 2001a). In addition, at low redox potential sulfate ions are reduced tothe sulfide form which may form complexes with metals such as Cd, Zn,and Ni (Hesterberg, 1998; Van Den Berg et al., 1998). Most metal sulfidesare insoluble even under acidic conditions and so this process also tends toreduce soluble metal concentrations.

Oxidation state of the contaminant itself also affects solubility.For example, selenite [Se(IV)] is much more strongly adsorbed to soilcolloid surfaces than selenate [Se(VI)] and the presence of selenite is favoredunder 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 andmetal 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 conditionsAs is present mainly as As(III) (O’Neill, 1995). Compared to other Asspecies, As(III) exhibits the greatest mobility and plant availability becauseof its presence as the neutral species H3AsO3 (Ascar et al., 2008; Marin et al.,1993). Nonetheless, strongly reducing conditions in biosolids-amended soilscan lead to precipitation of As as As2S3 (Carbonell-Barrachina et al., 1999).

The ability of soils to adsorb and sequester metals is also an importantfactor. This is dependent on their content of inorganic (clay and Fe, Mn andAl hydrous oxide content) and organic (soil humic material) binding agents.For example, sandy soils with low oxide content and low organic matterhave low sorption capacities and will have greater metal availabilities thanloamy 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 biosolidsto three different soils, a limed volcanic ash-derived Andept, an alkalineVertisol, and a limed manganiferous Oxisol. DTPA-extractable soil metallevels, lettuce growth, and tissue metal concentrations were measured.



The Andept had the highest metal adsorption capacity and the Oxisol thelowest. As a result, lettuce Cd, Mn, Ni, and Zn concentrations were highestin the Oxisol and Mn levels reached phytotoxic levels. Hue et al. (1988)concluded that the Andept could tolerate the highest biosolids loading rateand the Oxisol the lowest. The calcite (CaCO3) content of soils can also beimportant. In calcareous soils, calcite represents an effective sorbent formetal ions. The initial reaction is thought to be chemisorption but metalswith an ionic radius similar to that of Ca (Cd 2þ, Mn 2þ, Fe 2þ) can alsoreadily enter the calcite structure and form coprecipitates (Gomez de Rioet al., 2004; McBride, 2000).

5.3.4. Metal availability over timeThe long-term (>10 years) bioavailability of heavy metals in biosolids-amended soils is of great importance in relation to environmental effectsof land application of biosolids. As noted previously (Section 5.3.2), follow-ing a one-off application of biosolids the extractability of metals generallydeclines over time (Hseu, 2006; Sukkariyah et al., 2005a; Walter et al.,2002). Sukkariyah et al. (2005a), for example, showed DTPA-extractableCu and Zn levels progressively decreased following one-time applicationsof biosolids at rates ranging from 42 to 210 Mg ha� 1 (Table 3). Seventeenyears after application, extractable concentrations of Cu and Zn haddecreased by 58% and 42%, respectively. The decrease is attributable tometals reverting to more recalcitrant forms in the soil such as occlusion in Feoxides 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 kg� 1

DTPA-extractable

Zn mg kg� 1

Biosolids rates

Mg ha� 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 probabilitylevel.

From Sukkariyah et al. (2005a); copyright American Society of Agronomy.



those of control for many decades after applications have ceased (Allowayand Jackson, 1991; Basta et al., 2005; McBride, 1995; McGrath, 1987).Results from a long-term market garden experiment at Woburn (UK) serveto illustrate this point. Sludge was applied in the 1940s until the 1960s andCaCl2-extractable Cd changed little from 1950 until the early 1980sremaining significantly higher than the control soils over the entire intervalmonitored (McGrath and Cegarra, 1992). Similarly, EDTA-extractable Cu,Pb, Zn, Ni, and Cr changed little following termination of biosolidsapplication and treated soils maintained a much greater proportion ofmetal in EDTA-extractable form than the control. Such results occurreddespite there being a significant loss of biosolids organic matter over theperiod indicating that heavy metals released from the decomposing organicmatter were rapidly adsorbed by inorganic components of biosolids/soiland/or native soil organic matter. Certainly, biosolids-derived heavy metalsare strongly sorbed to soil components making them characteristicallyimmobile in soils. Indeed, the vast bulk of the added metals remain in thetopsoil in the layer of incorporation and there is a marked reduction inconcentration 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 leachingThe results of Sukkariyah et al. (2005b) serve to illustrate the immobility ofbiosolids-borne heavy metals in soil. They found that more than 85% oftotal applied Cu and Zn was still in the layer of incorporation (0–15 cm)17 years after a one-time biosolids application. Results for MehlichI-extractable Cu and Zn at that site are shown in Fig. 3. It is evident thatextractable Cu and Zn are concentrated in the 0–15 cm layer but there issome indication of a small amount of movement down into the 15–20 cmlayer. Mass balances calculated for several long-term experiments do suggestsome losses of heavy metals from the topsoil (McBride, 1995). Lateralmovement in the soil due to tillage (McGrath and Lane, 1989) or physicalmixing with the lower soil layer by plowing (Sloan et al., 1998) can beresponsible for a significant part of the losses from the original amended soillayer. Nevertheless, mass balances calculated for sites where little or notillage has been performed have shown less than 100% recovery (McBrideet 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 beennoted in field experiments (Barbarick et al., 1998; Baveye et al., 1999; Bellet al., 1991; Keller et al., 2002; Schaecke et al., 2002). Leachate samplingbelow field plots and/or undisturbed monolith lysimeters receiving biosolidshas also revealed elevated metal concentrations (Keller et al., 2002; Lamyet al., 1993; McBride et al., 1997, 1999; Richards et al., 1998; Sidle andKardos, 1977). In addition, column leaching studies have shown that heavymetals 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)

Dep

th, c

m

210 Mg ha−1

126 Mg ha−1

Control

210 Mg ha−1

126 Mg ha−1

Control

Zn

Cu

/ / / /

/ / / /

Figure 3 Distribution of Mehlich-I extractable Cu and Zn with soil depth 17 yearsafter biosolids application. From Sukkariyah et al. (2005a,b); copyright American Societyof 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-mixinglayer represents a small fraction (i.e., <1–2%) of the total amount of metaladded (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 ina substantial redistribution into the subsoil layers and/or groundwater. Inaddition, in some studies, water quality standards have been exceeded in soilsolution at depths below the zone of incorporation (McBride et al., 1999;Richards et al., 1998). Dilution by other unpolluted water will normallyprevent water quality standards being exceeded in receiving groundwater.The most danger will occur where large areas of land above small, shallowwater bodies are treated with biosolids.

A major contributor to heavy metal mobility in soils is thought to be theformation of complexes with dissolved organic matter released from thebiosolids (Brown et al., 1997; Christensen, 1985; Gerritse et al., 1982; Lamyet al., 1993; McBride et al., 1997). The amount of dissolved organic matterin soil solution and leaching through the profile characteristically increasesfollowing biosolids application and it acts as a ‘‘carrier’’ for heavy metals.Elevated concentrations of both heavy metals and dissolved organic matterare 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 amendedwith biosolids at 0, 20, and 100 Mg ha� 1, there was a substantial increase indissolved organic C at about day 23 which was attributed to a flush ofmicrobial activity. This was accompanied by a similar increase in solubleZn and an increase in calculated activity of Zn-organic matter species (Fig. 4).

The formation of strong soluble organic matter–heavy metal complexesin soil solution has been found to reduce heavy metal adsorption to solid soilphases. Neal and Sposito (1986), for example, found that sewage sludge canprovide sufficient dissolved organic matter to reduce adsorption of Cd ontosoil surfaces. Wong et al. (2007) showed dissolved organic matter had astronger 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 Pbby soils. Thus, both heavy metal solubility and mobility is increased.Dissolved organic matter originating from the biosolids may well have asecond effect in increasing metal mobility. That is, dissolved organic mattermolecules 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 partiallyblock potential sorption sites for metals thus tending to increase theirsolubility and availability.

In drainage waters from biosolids-amended soils, the bulk of heavymetals have been found to be associated with soluble organic matter.

0

1

2

3

4

5

6

Day 0Day 23

0

0.2

0.4

0.6

0.8

1

0Days of incubation

Sol

uble

Zn

(mg

kg−1

)

100 Mg ha−120 Mg ha−1Control

Zn

(µm

ol L

−1)

Control 20 Mg ha−1 100 Mg ha−1

50 100 150 200 250 300 350

Figure 4 Water-soluble Zn dynamics during incubation of amended and biosolids-a ended soils and calculated activities of Zn-dissolvedorganic matter species (mmol L�1) at days 0 and 23. From Antoniadis et al. (2007); co yright American Society of Agronomy.


mp



Using gel filtration chromatography, Dudley et al. (1987) found that in soilextracts from 80–100% of water-soluble Cu, 48–100% of Zn, and 39–100%of Ni was in organically complexed form. Using differential pulse anodicstripping voltametry, McBride et al. (1999) determined that only 30% ofwater-soluble Zn, 18% of Cd, and 10% of Cu was present as ionic orinorganic complexes and the remainder was presumed to be complexedwith dissolved organic matter. Using the same method, Al-Wabel et al.(2002) concluded that >99% of soluble Cu and Zn in leachates was presentin organically complexed form. Heavy metals have, however, also beenshown to be present in drainage water associated with suspended clay-sizedparticles (Keller et al., 2002). The metals become adsorbed to the surfaces ofFe oxide and layer silicate clays present in this leached particulate matter.Keller et al. (2002) calculated that movement of particulate matteraccounted for about 20% of Cu, Zn, and Cd leaching from a biosolids-amended soil.

An important factor thought to contribute to leaching of metals ispreferential flow of water and dissolved metals down the soil profile indownward oriented macropores (e.g., cracks, earthworm channels, rootchannels) (Camobreco et al., 1996; Keller et al., 2002; Lamy et al., 1993).This water bypasses the soil matrix thus minimizing the chances that thedissolved metals will be adsorbed to soil surfaces. Preferential flow is proba-bly the main pathway of movement of suspended particulate matter andassociated metals (Keller et al., 2002).

The period of greatest risk of metal leaching is soon after biosolidsapplication. This is when soluble organic matter is present in high concen-trations and when preferential flow down surface-connected macropores ismost likely. Indeed, leaching losses of metals are normally greatest duringthis initial period (Antoniadis et al., 2007; Camobreco et al., 1996; Kelleret 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 drainagefrom soils immediately following land application of biosolids.

5.3.6. Soil microbial/biochemical effectsElevated concentrations of heavy metals in soils are known to affect soilmicrobial populations and associated activities (Baath, 1989; Brookes, 1995;McGrath, 1994). Baath (1989) concluded that the following order oftoxicity to soil microbes is most commonly found (in mg kg� 1 values):Cd>Cu>Zn>Pb. However, he showed an enormous disparity betweenindividual studies as to the exact concentrations at which metals becometoxic. Giller et al. (1998) suggested that much of the variability in derivingtoxic concentrations of heavy metals occurs through comparison of resultsfrom short-term laboratory incubation studies with data from long-termexposures 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 oflong-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 structuraland functional diversity since toxicities can suppress and/or kill sensitiveparts of the community. Nevertheless, rediversification can occur in thesurviving tolerant communities (Barkay et al., 1985). The other is anincrease in respiration per unit of microbial biomass (metabolic quotient;qCO2) which is thought to occur because stressed microorganisms direct arelatively larger amount of available energy into maintenance of variousbiochemical functions (Giller et al., 1998). Thus, in general heavy metalcontamination of soils has been shown to result in a decline in microbialbiomass 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 negativeeffects on soil enzyme activity (Belyaeva et al., 2005; Kizilkaya and Bayrakli,2005). Enzyme reactions can be inhibited by heavy metals through anumber of mechanisms including by (i) complexing with the substrate, (ii)combining with the protein-active groups of the enzymes, or (iii) reactingwith the enzyme–substrate complex (Dick, 1997).

In the case of biosolids application to soils, the addition of organicmaterial increases organic matter content and consequently the sizeand activity of the microbial community also tend to be stimulated(Section 3.1.2). However, if biosolids contain a high heavy metal loadthen metal toxicities may have an inhibitory effect on soil microbial activity.Indeed, many workers have observed an inhibitory effect in soils wherebiosolids high in heavy metals have been applied and these negative effectscan remain for decades after application (Giller et al., 1998; Stoven et al.,2005).

Numerous short- and long-term studies have been carried out wherebiosolids contaminated with one or more heavy metals (or biosolidsenriched with one or more heavy metals) have been applied to soils andthe size and activity of the microbial community measured. Short-termincubation experiments have generally shown a reduction in microbialbiomass C and N, usually an increase in metabolic quotient and a variableeffect on enzyme activity (Bhattacharyya et al., 2008; Kao et al., 2006; Rostet al., 2001). Long-term (>8 years) field trials have shown similar resultswith a depression in microbial biomass C and microbial biomass Cexpressed 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) sampledsoils in fields that had been irrigated with heavy metal contaminatedwastewater (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 quotientgenerally increased along the gradient while microbial biomass C declined(Table 4). Observed effects on soil enzyme activities have been variable withBhattacharyya 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 dehydrogenaseand phosphatase activities were not consistently affected by a combinationof 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 highcombined concentrations of Cr, Cd, Cu, Hg, Ni, Pb, and Zn.

Not only is the size and activity of the soil microbial community affectedby 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 thepotential toxicity of water-soluble contaminants in soils and this techniquewas employed by Horswell et al. (2006) to determine the effects of Cu-,Ni-, and Zn-spiked biosolids on the microbial community in the litter layerof a forest soil. They found that increased Cu caused a decline in biolumi-nescence response of the fungal biosensor, increased Zn caused decline inresponse of the bacterial biosensor while increased Ni had little effect oneither. 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, anddecreased diversity, were induced in both bacterial (Sandaa et al., 1999a,2001) and archaeal (Sandaa et al., 1999b) community structures. Usingmolecular techniques Macdonald et al. (2007) showed that in an 8-yearstudy 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 resultsshowed that fungi, and to a lesser extent archaea, were more negativelyaffected by Zn addition than was the bacterial community. Results fromseveral long-term experiments have shown that Rhizobium leguminosarum, aN2-fixing symbiotic bacteria of white clover, is considerably more sensitiveto the toxic effects of heavy metals than the host plants and that the hostplant 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 freeliving rhizobium particularly in response to high Zn (Chaudri et al., 2008).

Thus, the general effect of heavy metal contamination of soils inducedby 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 Stransformations. Such decreased enzyme activity will tend to reduce theturnover of C, N, P, and S in the soil. The potential effect of a change in

Table 4 DTPA-extractable Cd, Cu, and Zn, microbial Biomass C, metabolic quotient nd the activities of dehydrogenase and cellulose insoils on a gradient of increasing Cd loading from 30 years of irrigation with heavy m tal-contaminated wastewater

Site

DTPA-extractable

(mg kg� 1)

Microbial

biomass C

(mg kg� 1)

Metabolic quotient

(�102mg CO2–C

h� 1mg� 1)

D hydrogenase activity

( g product kg� 1 h� 1)

Cellulase activity

(mg product kg� 1 h� 1)

Cd Cu Zn

1 0.48 5.18 3.42 207 5.47 1 5 14.8

2 0.48 5.00 4.62 199 5.17 4 1 10.3

3 0.52 5.28 2.44 177 4.91 3 9 16.8

4 1.05 13.0 9.32 142 7.51 2 2 14.3

From Zhang et al. (2008); copyright Elsevier.


, ae

e

m

.5

.2

.8

.0



species composition and loss of some microbial species is controversial. Thisis because there is a high degree of functional redundancy among soilmicroflora and functionally similar organisms can have different environ-mental tolerances (Nannipieri et al., 2003). Thus, as a stress is imposed onthe soil, there will be a progressive extinction of sensitive species among afunctional group; below a certain threshold it is speculated that there will beinsufficient individuals to sustain a particular function (Giller et al., 1997).Such a situation does not appear to have occurred in biosolids-treated soils.With increasingly stringent regulations being enforced regarding the con-centrations of heavy metals that can be released into sewerage systems, thenegative effects of biosolids applications on soil microbial activity, inducedby their heavy metal loads, is likely to decline over the ensuing years.

5.4. Plant response and metal uptake

At agronomically realistic rates of biosolids application (e.g., 2–8 Mg ha� 1),heavy metals do not normally represent a serious limitation to crop growtheven though relatively large amounts of metals can be added to the soil;indeed as noted previously, the majority of experiments demonstrate apositive effect of biosolids application on crop growth ( Juste and Mench,1992; Singh and Agrawal, 2008). For example, Juste and Mench (1992)reviewed a large number of field trials in the United States and Europe andconcluded that phytotoxicity due to biosolids-borne metals was only rarelyobserved in grain crops. At one site, Cd toxicity, Ni toxicity, or thecombined effects of Cd and Ni caused yield reductions in maize. At veryhigh rates of biosolids application, metal toxicities are likely to occur. Bertiand Jacobs (1996), for instance, applied large amounts of biosolids to crop-lands over a 9-year period (cumulative applications of 240–690 Mg ha� 1)and recorded yield reductions in maize, sorghum, and soybean due to thecombined phytotoxic effects of Zn and Ni. Leguminous crops are usuallymore sensitive to metal loadings in biosolids than nonlegumes (Giordanoet al., 1975; McGrath et al., 1988). Such effects are explicable in terms of thedetrimental effects of heavy metals on Rhizobium nodule function.

There are a number of reasons for the low phytotoxic effect of biosolids-borne metals. These can include metal sorption by metal oxides and organicmatter (present in both the biosolids and the soil), increased pH, formationof insoluble salts (e.g., with silicate, sulfate, and phosphate), and antagonisticeffects of between biosolids metals (Bell et al., 1991; Emmerich et al., 1982;Juste and Mench, 1992).

Despite this, heavy metal loadings are often considered the major poten-tially detrimental effect of land-applied biosolids on the environment. Notonly do they accumulate in the soil but also phytotoxic effects on plants canoccur. More concerning is the accumulation of metals in crops in edibleplant parts. These aspects are considered below.



5.4.1. Metal toxicity and toleranceElevated concentrations of both essential (e.g., Cu, Zn) and nonessential(e.g., Cd, Pb, Hg) heavy metals in soils can inhibit plant growth. Toxicity iscaused by a range of interactions at both the cellular and molecular levels(Clemens, 2006; Clemens et al., 2002; Hall, 2002; Shaw et al., 2006). It canresult from binding of metals to sulfhydryl groups in proteins leading toan inhibition of activity or disruption of structure, from displacement of anessential element resulting in deficiency effects or from stimulation of theformation of free radicals and reactive oxygen species resulting in oxidativestress (Dietz et al., 1999; Hall, 2002; Van Assche and Clijsters, 1990).

Metal tolerance is ubiquitous and the main components of metalhomeostasis in plants involve transport, chelation, and sequestration pro-cesses (Clemens, 2001, 2006; Clemens et al., 2002; Hall, 2002). Theregulated activities of these processes ensure proper delivery and distributionof metals at both the organizational and cellular levels (Clemens, 2001). Lossof one of these critical processes (e.g., synthesis of chelating agents for Cdwithin cells) leads to genotypes that are more sensitive (hypersensitive) thanwild-type plants (Howden et al., 1995). By contrast, some plant species andgenotypes can grow naturally in soils containing concentrations of metalsthat would be toxic to the majority of plants. These hypertolerant speciespossess naturally selected higher levels of tolerance. Potential mechanisms ofdetoxification are primarily involved in avoiding the buildup of toxicconcentrations at sensitive sites within the cell (Hall, 2002) thus preventingthe damaging effects outlined above. Some plants not only tolerate highconcentrations of metals but also hyperaccumulate them. Hyperaccumula-tor is a term used to describe plants capable of accumulating more than1000 mg g� 1 Ni, Co, Cu, Pb, for Cd 100 and Zn 10,000 mg g� 1 (Love andBabu, 2006). Metal hyperaccumulation is a rare phenomenon limited toabout 400 different species belonging to a wide range of taxa (Baker andBrooks, 1989). Hyperaccumulation is mainly observed with Ni, Zn, Co,and Se but there are also reports of it for Pb, Cd, and As (Baker and Brooks,1989). The existence of hyperaccumulator plants is the basis for the phytor-emediation technique in which plants are used to absorb metals and henceremove them from contaminated soils (Chaney et al., 1997).

Possible strategies of heavy metal tolerance in plants are diverse and canbe extracellular, at the plasma membrane, or intracellular. Extracellularmechanisms include the effects of ectomycorrhizae ( Jentsche andGodbold, 2000) and arbuscular mycorrhizae (Kaldorf et al., 1999; Leyvalet al., 1997) in restricting metal movement into the plant host roots. This hasbeen attributed to absorption of metals by the hyphal sheath, chelation byfungal exudates, adsorption onto external mycelium, and reduced access tothe plant root cytoplasm ( Jentsche and Godbold, 2000; Leyval et al., 1997).Root exudates (e.g., organic acids such as citrate, malate, and oxalate) have



also been suggested as agents for detoxification through chelation of metalsin the rhizosphere (Dong et al., 2007). In solution culture experiments,addition of a range of organic compounds, such as organic acids, aminoacids, and peptides, to the growing medium can alleviate heavy metaltoxicities (Hall, 2002) and the Ni chelating ability of citrate and histidinein root exudates has been associated with its nonaccumulation in species ofThlaspi (Salt et al., 2000). Another possible mechanism is the accumulationof heavy metals in root cell walls. Indeed, in many plants that accumulateheavy metals, 60–80% of metals (e.g., Cd, Cu, Zn, Pb, Co, Ni) accumulatein roots and root cell walls are the main region of accumulation (Liu et al.,2007a,b; Nishizono et al., 1987; Sousa et al., 2008; Zornoza et al., 2002).In some cases, this accumulation in the cell wall may reflect active efflux ofmetals from root cells through the plasma membrane (see below). Further-more, for some plants (e.g., lettuce) that accumulate metals (e.g., Cd) inleaves, the Cd accumulates principally in the cell walls (Ramos et al., 2002)suggesting it has been deposited there by active efflux across leaf cell plasmamembranes.

There are three suggested mechanisms at the plasma membrane. Firstly,plasma membrane function may be rapidly affected by high concentrationsof metals with increased leakage of solutes (e.g., K) occurring (Hall, 2002).Tolerance could involve protection of plasma membranes against metaldamage or improved repair mechanisms (Hall, 2002). The cell membranecould also play an important role in reducing entrance of metals into cells(Clemens, 2006). However, this strategy is thought to be of minor signifi-cance since many essential heavy metals must be taken up from the soil forvarious metabolic functions and the chemical properties of nonessentialmetals are similar to those of essential ones and their uptake is thought tooccur through the same processes/mechanisms (Shaw et al., 2006). Analternative strategy is active efflux of toxic metal ions. In bacteria, mostresistance systems are, in fact, based on energy-dependent efflux of toxicions either by ATPases or chemostatic cation/proton antiporters (Silver,1996). It seems likely that such a mechanism also occurs in higher plants(Hall, 2002).

It is generally accepted that the principal mechanism of detoxificationand tolerance is achieved by sequestration of metals inside cells (Shaw et al.,2006). Chelation of metals in the cytosol by organic ligands is an importantintracellular mechanism of detoxification and tolerance. Chelates bind withmetals and buffer cytosolic ionic metal concentrations. Carboxylic acids(e.g., citrate, malate, oxalate) and amino acids such as histidine are potentialligands (Clemens, 2001) and metallothioneins (low molecular weight, cys-teine-rich proteins) may also play a role (Cobbett and Goldsbrough, 2002).Nonetheless, the most studied ligands are phytochelatins which are metal-binding peptides synthesized in plants in response to exposure to metal ions(Cobbett and Goldsbrough, 2002; Rauser, 1999). Hg, Cd, As, and Cu show



the greatest tendency to induce phytochelatins and Zn, Pb, Ni, and Se havea lesser effect (Shaw et al., 2006). For Cd, the phytochelation pathwayconsists of two parts. Firstly, metal-activated synthesis occurs and metals inthe cytoplasm are complexed. The second stage is active efflux of the metal–phytochelatin complex across the tonoplast and sequestration in the vacu-ole. The vacuole is known to be the site of accumulation of metals such asZn and Cd (Ernst et al., 1992). As already noted, active efflux of metalsacross the plasma membrane with their accumulation in cell walls may beanother important mechanism.

It is evident that the general strategy of plants is to avoid toxicity byminimizing the buildup of excess metals in the cytosol. This is achievedthrough the use of a variety of mechanisms. It is likely that a specificmechanism, or combination of mechanisms, is employed for a specificmetal in particular plant species. In situations where soils have becomehighly contaminated with heavy metals, the use of tolerant species/cultivarsthat employ these mechanisms can be an important consideration. How-ever, since heavy metal toxicities in plants are rarely encountered in bioso-lids-amended soils, this aspect is of minor importance to land application ofbiosolids. Accumulation of heavy metals in plants and plant parts can,nevertheless, be of concern.

5.4.2. Metal accumulation in plantsOne of the major concerns regarding land application of biosolids is thatheavy metals may accumulate in plants and subsequently enter the foodchain and/or have toxic effects on humans or grazing animals ingestingthem. Indeed, the main route of entry of metals into human food chain istheir accumulation in edible portions of crop plants and this can pose apotential threat to human health (McLaughlin et al., 1999). Chaney (1980)classified metals into four groups when considering their potential healthrisks. Group 1 (Ag, Cr, Sn, Ti, Y, and Zr) were considered to pose littlerisk because their low solubility in soils results in them not being taken upby plants to any great extent. Group 2 (As, Hg, and Pb) are also stronglyadsorbed to soil surfaces and while they may be taken up by plants they arenot readily translocated to edible portions. They therefore pose little risk tohuman health. Group 3 (Cu, Mn, Ni and Zn) are accumulated in plants butare phytotoxic at concentrations that pose little risk to human health.Members of Group 4 (Cd, Co, and Se) can pose human health risks atplant tissue concentrations that are not phytotoxic. In general, metals thathave most commonly given rise to human health concerns in relation tofood safety are Cd, Hg, Pb, As, and Se (Reilly, 1991). McLaughlin et al.(1999) considered that Cd and Se are of greatest concern in relation toterrestrial food chain contamination. In particular, excessive human intakeof Cd is of concern since it accumulates in the body and impairment ofkidney function is the main adverse effect.



Concentrations of extractable heavy metals in soils are likely to give thebest estimate of metal phytoavailability. Even so, at a broad level, metalconcentrations in plants grown on biosolids-treated soils are a function ofthe annual biosolids loading rate ( Juste and Mench, 1992). Cumulativemetal input to the soil is also a major factor determining plant tissue metalconcentrations (Chang et al., 1997; Soon et al., 1980). Nonetheless,although plant metal concentrations generally increase with increasingbiosolids rates, concentrations in plant tissues often exhibit a plateauresponse at high loadings (Basta et al., 2005). Indeed, a number of studieshave showed that metal uptake reaches a maximum with increasing bioso-lids application for wheat, maize, and a range of vegetables (Barbarick et al.,1995; Brown et al., 1998; Logan et al., 1997, Sukkariyah et al., 2005a).Although a number of explanations have been forwarded for this (Bastaet al., 2005, McBride, 1995) the most likely explanation is that metalavailability in biosolids-treated soils shows a plateau at high loadingscorresponding to metal availability in the biosolids (Basta et al., 2005;Corey et al., 1987). Plant tissue metal concentrations will, however, notonly be related to biosolids application rate but will also be both metal andplant specific. That is, the mobility of individual metals in plants differs andthe ability of individual plant species and cultivars to absorb and translocatemetals also differs.

The transfer coefficient (TC) (the concentration of metal in the plant tothat in the soil) gives an indication of its mobility. From a number of studies,Antoniadis et al. (2006) calculated TC values of Cr 0.0005, Pb 0.02, Ni0.06, Cu 0.21, Cd 0.94, and Zn 1.05. Similarly, Alloway (1995) also notedthat TC values were highest for Cd and Zn (1–10) and least for Cr and Pb(0.01–0.10). In a review of a number of long-term (>10 years) biosolidstrials, Juste and Mench (1992) concluded that Cd, Ni, and Pb were themost likely to accumulate in plants while for Cr and Pb, plant uptake wasinsignificant. Generally, metals are present in highest concentrations in rootsand lowest concentrations in seeds (Love and Babu, 2006). However, Cdcan accumulate in leaves so on Cd-contaminated soils, leafy vegetables canbe as much, or more, of a risk than seed or root crops (Alloway and Jackson,1991).

The propensity for plants to accumulate and translocate metals to edibleand harvestable parts depends greatly on plant species as well as agronomic,climatic, and soil factors. Different plant species and cultivars are known toaccumulate different quantities of metals when grown in the same con-taminated soil. Antoniadis et al. (2006) broadly classified crop species inrelation to metal accumulation. Crops with a high plant uptake includedlettuce, spinach, celery, kale, ryegrass, sugarbeet, and turnip while the lowuptake group included potato, maize, peas, leek, onion, tomato, and berryfruit. Nonetheless, plant species differ in their capacity to accumulateindividual heavy metals. For example, on biosolids-amended soils Davis



and Carlton-Smith (1980) found the trend for metal accumulation wastobacco> lettuce> spinach> celery> cabbage for Cd; kale> ryegrass>celery for Pb; sugarbeet> some varieties of barley for Cu and sugarbeet>mangold> turnip for Zn. For Cd, Alloway et al. (1990) showed Cd accu-mulation followed the order: lettuce> cabbage> radish> carrots. Maizecultivars have been shown to differ considerably in their uptake of bothCd and Zn (Hinesly et al., 1982; Logan and Miller, 1985). For root cropsgrown in biosolids-amended soils, it may be important to wash of alladhering soil (or peel them) to avoid soil contamination with heavy metals.Indeed, where crops are being grown on biosolids-amended soils, routinemonitoring of heavy metal concentrations in edible plant parts is an impor-tant consideration. Many countries have set limits for metal concentrationsin foods and it is important that these are not exceeded.

5.5. Ingestion by animals

Pasture is an attractive option for land application of biosolids since there isgreater accessibility for a greater proportion of the year compared to arableland. Nevertheless, a major concern is transfer to, and accumulation of,heavy metals in the grazing animal and the human food chain (Hill, 2005;Hillman et al., 2003). The three main pathways by which this may occur are(i) transport of metals from soil to plants and then ingestion by animals(biosolids–soil–plant–animal), (ii) direct contamination of plants subse-quently fed to animals (biosolidsplant–animal), and (iii) ingestion of con-taminated soil by grazing animals (biosolids–soil–animal) (Fries, 1996).

Accumulation of metals into foliage of grasses is generally low sincemetals usually accumulate in the roots (Love and Babu, 2006). Thus,pathway (i) is likely to be of minor importance. However, where surfaceapplication of biosolids is practiced, adhesion of biosolids-derived metals topasture herbage occurs (Aitken, 1997; Klessa and Desira-Buttegieg, 1992).The greatest intake will occur when biosolids are applied directly to estab-lished pasture and animals have immediate access. Intake will be reduced ifaccess to pasture is delayed so there is time for biosolids to be washed offleaves, foliar concentrations are reduced by dilution through plant growthand the biosolids can move onto/into the soil surface (Aitken, 1997;Hillman et al., 2004). Herbage may also become contaminated throughrain splash or from deposition of dust (Aitken, 1997). There are waitingperiods imposed in most countries/states/regions following surface applica-tion of biosolids to pastures before land can be grazed again. These aregenerally formulated to minimize the risk of exposure of grazing animals topathogens rather than to chemical contaminants and can range from as shortas 3 weeks in the United Kingdom, Spain, and the Netherlands to 1 year inDenmark (Hill, 2005).



When pasture is the main diet of grazing animals, the chance for animalingestion of biosolids-borne metals is greatest since soil ingestion becomesan additional important pathway. Research has shown that when pasture isthe sole animal feed source, soil ingestion is generally inversely related to theavailability of forage (Beyer and Fries, 2003). Lowest soil ingestion rates(1–2% of dry matter intake) occur in spring when grass growth is greatestwhile when forage is sparse (autumn and winter) soil intake can be as great as18%. For sheep that graze close to the ground soil ingestion rates of as highas 30% have been reported on occasions (Abrahams and Steigmajer, 2003).In addition, in winter, poaching, or pugging of pasture can also occur(Drewry et al., 2008). This is when the hooves of grazing sheep and cattletrample the foliage into muddy soil thus causing soil contamination offorage. For grazing sheep and cows (fed no additional forage) average annualsoil ingestion commonly amounts to 4.5–5.1% of dry matter intake (Beyerand Fries, 2003). In general, soil/biosolids ingestion is the main pathway fortransfer of metals from pasture to the grazing animal (Beresford andHoward, 1991; Hillman et al., 2003; O’Riordan et al., 1994; Raffertyet al., 1994).

There is differential accumulation of individual metals in tissues ofgrazing ruminants with Cd and Pb accumulating in the liver and kidneyto a greater extent than muscle and fat tissues (Hillman et al., 2003).Cu tends to accumulate in the liver while Zn and Fe have no clear patternsof accumulation. The percentage of metals retained is generally low (Baxteret al., 1982; Johnson et al., 1981). For example, Johnson et al. (1981) fedsteers a diet containing 11.5% biosolids for 106 days. Retention of ingestedmetals averaged 0.09%, 0.06%, and 0.30% for Cd, Hg, and Pb, respectively,and no retention of Cu and Zn was detected. Nonetheless, this lowpercentage retention increased tissue Cd, Hg, and Pb concentrations inliver and kidney by 5–20-fold. Overall, the risk to the human food chainfrom biosolids-derived heavy metals via grazing animals is considered low(Hillman et al., 2003).

6. Organic Contaminants

In the last 50-years production of synthetic organic chemicals forindustrial and domestic uses has increased enormously. For example,between 1950 and 1970, production increased from 7 million tons to 63million tons (Rogers, 1996). As a consequence, the occurrence and con-centration of organic contaminants in effluents, sewage, and biosolids hasalso increased. The presence and level of organic contaminants in biosolidsdepends greatly on the quality of the wastewater, the different local pointsources, the physicochemical properties of particular organic compounds,



and operational parameters of the wastewater treatment plant. Concentra-tions of organic pollutants are generally greater in industrial sewage than indomestic effluent (Bodzek and Janoszka, 1999; Smith, 2000). Over 300organic chemicals from a diverse range of classes of compounds have beenidentified in biosolids and their concentrations vary from the pg kg� 1 to theg kg� 1 level ( Jacobs et al., 1987; Smith, 2000).

Although water treatment plants were designed primarily to removeorganic matter, and the mechanisms of degradation of bulk organic compo-nents are well studied and understood, the processes by which syntheticorganics are degraded have received relatively little study. An organic con-taminant could undergo a number of processes including (i) sorption to solidsurfaces, (ii) volatilization, (iii) chemical degradation, and (iv) biodegradation.Generally, the more hydrophobic a compound is, the more susceptible toaccumulation onto sewage sludge particles it will be. During primary sedi-mentation, hydrophobic contaminants may partition onto settled primarysludge solids. The tendency to accumulate in sludge solids can be assessedusing the octonol–water partition coefficient (Kow) (Byrns, 2001). Contami-nants with log Kow values less than 2.5 have low sorption potential and thosewith values greater than 4 have high sorption potential. Similarly, the ten-dency for volatilization can be gauged using Kow and Henry’s Law constant(Rogers, 1996).

Although there is a scarcity of data on the behavior of organic contami-nants during the water treatment and sludge digestion processes, somegeneralizations can be made (Rogers, 1996; Scow, 1982). For example,molecules with highly branched hydrocarbon chains are generally lesssusceptible to biodegradation than unbranched compounds and short chainsare not as quickly degraded as long chains. In addition, unsaturated aliphaticcompounds are generally more susceptible to degradation than saturatedanalogs. In general, due to their characteristically low water solubility andhigh lipophilicity, organic contaminants partition into sludges during sedi-mentation resulting in their accumulation in biosolids in concentrationsseveral orders of magnitude greater than influent wastewater concentrations(Bhandari and Xia, 2005).

Some organic contaminants are known, or are suspected to be endocrinedisrupting chemicals (EDCs) and this has magnified interest in their pres-ence in sewage, their fate during waste water treatment, and their possiblepresence in biosolids. The endocrine system is found in nearly all animalsand is a complex network of glands that discharge hormones that regulatethe body’s functions including growth, development, and maturation aswell as the way various organs operate. EDCs possess the ability to alter ordisrupt endocrine system function mimicking, antagonizing, or interferingwith biosynthesis or biodegradation of endogenous hormones(Sonnenschein and Sato, 1998). Indeed, there is growing concern aboutthe apparently increasing incidence of reproductive disorders and abnormal



development in wildlife and reduced fertility in human males, problems thatmay be caused by EDCs that have been released anthropogenically into theenvironment (Ashby et al., 1997; Sonnenschein and Soto, 1998).

A wide range of chemicals have been found, or are suspected to becapable of disrupting the endocrine system (Birkett, 2003a). These include:

(i) persistent organochlorines and organohalogens (e.g., PCBs, dioxins,furans, brominated fire retardants),

(ii) pesticides (e.g., DDT, atrazine, vinclozolin, TBT),(iii) alkyl phenols (e.g., nonylphenol, octylphenol),(iv) phytoestrogens (e.g., isoflavoids, lignins, B-sitosterol), and(v) natural and synthetic hormones (e.g., b-estradiole, ethynylestradiol).

The last group of chemicals (above) is known as estrogenic EDCs andhas received particular attention in recent times. Estrogens are a group offemale sex hormones involved in the estrous cycle. They are excreted inurine and eliminated in feces and both naturally occurring estrogens andxenoestrogens (man-made analogs) have been identified in sewage, bioso-lids, and waste water treatment plant effluents (Birkett, 2003b).

Analysis of organic contaminants in biosolids presents a number ofchallenges. The organic molecules are physically and/or chemically boundto the biosolids solid phase matrix and must be extracted with relativelyharsh reagents/methods prior to analysis. Because extraction yields a com-plex mixture of organic compounds, the various organic fractions need tobe segregated (cleaned) prior to analysis. Furthermore, concentrations ofcontaminants are often low, and sometimes below the limits of currentanalytical techniques. As a result, preconcentration is often necessary priorto analysis. Traditional methods entail soxhlet extraction, concentrationusing rotary evaporation and cleanup by column chromatography(Rogers, 1996). Contemporary methods, often preferred today, involvepressurized liquid extraction (PLE) followed by combined cleanup andconcentration using solid phase extraction (SPE) (Cirelli et al., 2008;Jones-Lepp and Stevens, 2007). Quantification of organic compoundspresent is performed by one of a number of powerful chromatographicmethods such as high performance liquid chromatography (HPLC), gaschromatography (GC), liquid chromatography (LC), and superficial fluidchromatography (SFC) and these are often coupled with molecule analysisby mass spectrometry (MS).

In recent times, as methods of extraction, cleanup and analysis havebecome more rapid, an increasing number of surveys have reported theconcentrations of various organic contaminants in wastewater and biosolidssamples. However, as yet, there are only a few studies where the fate of suchcontaminants during wastewater treatment processes has been reported.Similarly, very little information is available on the behavior and degradationof biosolids-borne organic contaminants in soils following land application.



Organic contaminants certainly have the potential to adversely impacton the soil/crop/animal system receiving land application of biosolids(Bhandari and Xia, 2005). Nevertheless, the relative risk of organic chemi-cals in biosolids is generally considered to be minimal due to the relativelylow concentrations present and the many transformations (especially bio-degradation) that can occur in soils (Epstein, 2003).The European Unionhas proposed limit values for several organic contaminants (or groups ofcontaminants) in biosolids and several European countries (e.g., Sweden,Germany, Denmark) have enforced a number of them. The EuropeanUnion limits are for sum of halogenated organic compounds, linear alkyl-benzene sulfonates, di-(ethylhexyl) phthalate, nonylphenol and nonylphe-nol ethoxylates, polynuclear aromatic hydrocarbons, polychlorinatedbiphenyls, and polychlorinated dibenzo-p-dioxins and -furans. Below, themost frequently detected organic contaminants in biosolids are considered.Their origin in sewage, potential toxicity, fate during wastewater treatment,and persistence or otherwise in the soil following land application areconsidered.

6.1. Organic compounds present

6.1.1. Phthalic acid esters (PAEs)These materials are manufactured in large quantities and are used predom-inantly as plasticizers to make plastics more flexible (Bhandari and Xia,2005). The large majority of phthalates are used to plasticize polyvinylchloride (PVC) to produce products ranging from kitchen and bathroomflooring to medical tubing, toys, footwear, electrical cables, packaging,and roofing. Di-(2-ethylhexyl) phthalate (DEHP) is the most widely usedphthalate. Other common PAEs include Di-n-octyl phthalate (DnOP),butylbenzyl phthalate (BBP), Di-n-butyl phthalate (DnBP), diethylphthalate (DEP), and dimethyl phthalate (DMP). Phthalates are used innon-PVC applications such as paints, rubber products, adhesives, cos-metics and toiletries (e.g., nail polish, perfumes), epoxy resins, adhesives,and printing inks. The widespread industrial and domestic use of productscontaining phthalates results in large amounts being washed down drainsand into sewage systems (Marttinen et al., 2003; Palmquist and Hanaeus,2005). Some PAEs are considered to be EDCs and/or carcinogens(Birkett, 2003a). Indeed, DEHP is considered to be an EDC (Hoyer,2001) and has been shown to cause malformations of the reproductivesystem in male rats (Gray et al., 1999). BBP and DBP have the strongestestrogenic potencies (Harris et al., 1997).

A wide range of bacteria can degrade PAEs under both aerobic andanaerobic conditions (Staples et al., 1997). The extent of biodegradationduring anaerobic digestion is apparently related to the size of the alkyl sidechain and compounds with the larger C-8 side chain are much more



resistant to microbial attack ( Jianlong et al., 2000; Staples et al., 1997). As aresult, the higher molecular mass PAEs such as DEHP and DnOP areconsiderably more persistent to anaerobic microbial degradation thanlower molecular mass compounds (e.g., DPP and DEP). DEHP is generallyconsidered persistent during sewage treatment (especially with anaerobicsludge digestion) (Bhandari and Xia, 2005; Scrimshaw and Lester, 2003).

Total concentrations of PAEs in biosolids typically range from 12 to200 mg kg� 1 (Amir et al., 2005c; Cai et al., 2007b; Gibson et al., 2005;Harrison et al., 2006; Marttinen et al., 2003; Oliver et al., 2005). DEHP isconsistently the most abundant in biosolids usually accounting for 25–95%of total PAEs present (Amir et al., 2005c; Cai et al., 2007b; Gibson et al.,2005; Oliver et al., 2005). Other PAEs such as DnOP, BBP, DnBP, DEP,and DMP are commonly present in biosolids in low concentrations(<10 mg kg� 1) (Cai et al., 2007b; Zheng and Zhou, 2006). Thus, landapplication of biosolids generally introduces PAEs to the soil environment.Under aerobic soil conditions PAEs are, however, readily microbiallydegraded ( Jianlong et al., 1997, 2004; Shanker et al., 1985; Wang et al.,1997) with the less degradable high molecular mass compounds havinglonger half-lives. Results of Jianlong et al. (2004) are shown in Fig. 5.Degradation of PAEs in soil (added at 100 mg of phthalate g� 1) decreasedwith increasing alcohol chain length with DMP being degraded completelywithin 15 days while less than 50% of DOP was degraded after 30 days.

0

20

40

60

80

100

0

Time (d)

PA

Es

conc

entr

atio

n (m

g/L)

5 10 15 20 25 30

Figure 5 Degradation of four phthalate esters in a soil after addition at 100 mg g� 1:(▪) DOP, (□) DBP, (△) DEP, (▲) DMP. From Jianlong et al. (2004); copyrightAmerican Chemical Society.



6.1.2. Polycyclic aromatic hydrocarbons (PAHs)PAHs are a complex group of organic compounds containing two or morefused aromatic rings in linear, angular, and cluster arrangements that containonly C and H atoms. PAHs occur in oil, coal, and tar deposits. They are alsoproduced as by-products of incomplete combustion of C-containing fuelssuch as wood, coal, diesel oil, fat, and tobacco (Wild and Jones, 1995).Important combustion processes include food preparation, internal com-bustion engines in motor vehicles, house fires, coal-fired power stations,cement works, iron and Al smelters as well as volcanoes, and forest fires.Thus, PAHs enter the environment mostly from releases to the air and thendirect aerial fallout. They enter the waste water system from a multitude ofsources including runoff from land, from roads containing car exhaustparticles, and direct discharge of petroleum products from garages. PAHsare hydrophobic compounds and their persistence within ecosystems ischiefly due to their low water solubility and their tendency to becomeadsorbed to solid particles (Cerniglia, 1992). In addition, the fused aromaticrings possess dense clouds ofP electrons on both sides of the ring structuresmaking them resistant to nucleophilic attack ( Johnsen et al., 2005). Due tothis effect, the resonance energy of PAH compounds (a measure of the extrastability of a conjugated system compared to the corresponding number ofdouble bonds) increases with increasing number of aromatic rings present.In general both water solubility and bioavailability of PAHs decrease andlipophilicity increases almost logarithmically with increasing molecular mass( Johnsen et al., 2005; Zhang et al., 2006). Due to their toxic, mutagenic,estrogenic, and carcinogenic properties, 16 PAH compounds have beenidentified as priority pollutants by the USEPA and seven of them areconsidered carcinogenic (IARC, 1983).

Due to their lipophilicity, PAHs rapidly become associated with solidsludge particles during waste water treatment and, as a result, significantquantities are typically present in biosolids. Indeed, degradation of PAHsunder anaerobic conditions is generally slow (Zhang et al., 2006) so thatanaerobic sludge digestion is not very effective at removing them. Never-theless, some degradation of PAHs (particularly 2- and 3-ring compounds)can occur anaerobically under sulfate-reducing and denitrifying conditionsusing NO3

� and SO42�, respectively, as electron acceptors (Christensen

et al., 2004; Johnsen et al., 2005; Zhang et al., 2006). The mean concentra-tion of total PAHs in biosolids from industrialized countries is typically inthe range of 1–100 mg kg� 1 (Blanchard et al., 2004; Dai et al., 2007;Harrison et al., 2006; Stevens et al., 2003, Villar et al., 2006). Concentrationsof individual PAHs vary markedly between wastewater treatment plants andregions although the dominant compounds are usually those with 3, 4 or 5rings (Bodzek and Janoszka, 1999; Cai et al., 2007b; Oleszczuk, 2007). TheEuropean Union proposed that for land application, the sum content of



11 PAHs in biosolids should not exceed 6 mg kg� 1 (CEC, 2000). In biosolidsfrom industrial countries this sum can often exceed the limit (Blanchard et al.,2004; Cai et al., 2007b; Harrison et al., 2006; Oleszczuk, 2007).

Land application of biosolids inevitably elevates concentrations of soilPAHs and their slow degradation follows (Beck et al., 1995; Oleszczuk,2006). It is thought that the strong binding of PAHs to biosolids organicmatter initially limits their decomposition. However, as the sludge organicmatter begins to breakdown, sorptive processes are weakened and PAHdegradation proceeds (Baran and Oleszczuk, 2003; Oleszczuk, 2006).During decomposition, there is a rapid disappearance of low molecularmass PAHs and a slower degradation of higher molecular mass compounds(Beck et al., 1995; Oleszczuk, 2006). After cessation of biosolids applicationsat two long-term monitoring sites, Beck et al. (1995) observed 90% of totalPAHs had been lost from soil at one site and about 65% at the other. Half-lives ranged from 2 years for naphthalene to over 7 years for fluorantheneand over 9 years for benzol[ghi]perylene and coronene. PAH degradation isgreater under arable cropping, with tillage (which stimulates organic matterdecomposition) than under undisturbed soils (pasture, trees) (Oleszczuk,2006; Saison et al., 2004).

Aerobic composting of biosolids prior to land application can substan-tially lower PAH concentrations (Amir et al., 2005d; Cai et al., 2007b;Moeller and Reeh, 2003; Oleszczuk, 2007). Cai et al. (2007c), for example,showed that 56 days composting resulted in removal rates of 64–94%.Removal rates are in the same order for 2, 3, 4, 5, and 6 ring PAHs(Cai et al., 2007c; Oleszczuk, 2007).

6.1.3. Chlorobenzenes (CBs)Chlorobenzenes are a group of cyclic aromatic compounds in which one ormore hydrogen atoms have been replaced by a chlorine atom. There are 12different CBs: monochlorobenzene (MCB), dichlorobenzene (DCB) (threeisomers), trichlorobenzene (TCB) (three isomers), tetrachlorobenzene(TCB) (three isomers), pentachlorobenzene (PeCB), and hexachloroben-zene (HCB). CBs (particularly 1,3-DCB, TCBs, and HCB) are used asintermediates in the synthesis of pesticides and other chemicals. 1,2-DCBis used in paintstrippers, engine cleaners, and other solvents, TCBs andPeCBs are used as components of dielectric fluids, TCBs are used as solventsand degreasers and 1,4-DCB is widely used as a toilet deodorant. DCBs,1,2,4-TCB and HCB have been classified as priority pollutants by bothUSEPA and the European Union and some CBs (e.g., HCB) are consideredas carcinogens ( Jones and Wild, 1991; Wang et al., 1995).

CBs are a major group of organics found in biosolids (Rogers, 1996;Wang and Jones, 1994a). Concentrations vary significantly between waste-water sources, sludge type, and treatment technique but are generally higherin biosolids derived from industrial than domestic sewage sources (Rogers



et al., 1989). Typically total CBs in biosolids range in concentrationbetween 0.1 and 40 mg kg� 1 although concentrations as high as2000 mg kg� 1 have been recorded (Beck et al., 1995; Cai et al., 2007b;Harrison et al., 2006; Rogers et al., 1989; Smith, 2000; Wang et al., 1995).In general, DCBs are the most commonly detected and are found in highestconcentrations (Harrison et al., 2006; Rogers et al., 1989; Wang et al., 1995)although in a range of Chinese biosolids samples, Cai et al. (2007b) detectedTCBs and HCB in higher concentrations than DCBs.

Very little is known about the fate of CBs during wastewater treatment.Katsoyiannis and Samara (2004), however, recorded 91% removal of HCBfrom the aqueous phase in a sewage treatment plant in Greece. Henry’s lawconstants for CBs suggest that they are likely to be volatilized from aquaticsystems. The solubility of CBs in water is, however, low and decreases withincreasing chlorination and the octanol to water coefficient increases withincreasing chlorination. Thus, CBs are strongly sorbed to organic matterand become associated with sludge during the sedimentation phases ofwastewater treatment (Beck et al., 1995). They are, therefore, present inbiosolids. CBs can be degraded aerobically by a consortium of bacteria viaoxidative dechlorination, followed by ring fission and mineralization (Wangand Jones, 1994a). Less chlorinated CBs are more readily biodegraded.Biodegradation has also been reported under anaerobic conditions (Yuanet al., 1999) but at a much slower rate. Thus, during anaerobic sludgedigestion, CB degradation is probably slow (Rogers, 1996).

Once introduced to the soil by land application of biosolids, volatiliza-tion is thought to be the major loss mechanism for CBs (Wang and Jones,1994a,b; Wang et al., 1995). Wang and Jones (1994b) showed the half-lifefor CBs added in biosolids to soil ranged from 13 to 209 days and theseincreased with increasing chlorination of individual CBs. Wang et al. (1995)showed that only about 10% of added CBs became recalcitrant in soils (i.e.,became strongly adsorbed to soil constituents). HCB was found to be themost persistent CB in soils (Beck et al., 1995; Wang et al., 1995).

6.1.4. Polychlorinated biphenyls (PCBs)PCBs consist of a biphenyl ring (two benzene rings) with 10 positionswhere chlorine substitution may occur. There are 209 possible PCB con-geners and most PCB mixtures consist of about 130 of these. Until the1960s PCBs found wide use as coolants and insulating fluids for transformersand capacitors, stabilizing additives in flexible PVC coatings of electricalwiring and electronic components, pesticide extenders, adhesives, woodfloor finishes, paints, and printing inks. However, because PCBs have beenfound to be persistent organic pollutants that bioaccumulate in animals, andthey are EDCs (Birkett, 2003a), industrial usage of PCBs has largely beencurtailed since the 1970s. Nonetheless, they still remain a major class of EUpriority pollutants (Rogers, 1996).



PCBs enter sewage effluent streams mainly via atmospheric depositionand catchment runoff. Total PCB concentrations in biosolids commonlyrange from 20 to 2000 mg kg� 1 (Alcock and Jones, 1993; Blanchard et al.,2004; Eljarrat et al., 2003; Frost et al., 1993; Katsoyiannis and Samara, 2004;McGrath et al., 2000, Stevens et al., 2003). It has been proposed by theEuropean Union that the sum of seven congeners (28, 52, 101, 118, 138,153, and 180) should not exceed 800 mg kg� 1 (CEC, 2000) and often EUPCB levels in biosolids are below this level (Alcock and Jones, 1996, Bersetand Holzer, 1999; Katsoyiannis and Samara, 2004; Stevens et al., 2003).Indeed, the concentrations of PCBs in biosolids (Blanchard et al., 2004),and in soils (Alcock et al., 1993), have generally been decreasing over thelast 20 years as PCB use has declined. PCBs exhibit a wide range of toxiceffects which may vary depending on the specific PCB being considered(Kannan et al., 2000; Masuda, 2005). Many are considered to be EDCs(Raychoudhury et al., 2000) and they are also carcinogens (Knerr andSchrenk, 2006; Ludewig et al., 2008).

Because of their lipophilic and hydrophobic properties, PCBs whichenter wastewater treatment plants tend to be adsorbed to particulate organicmaterial and are removed during sedimentation. Removal efficiencies typi-cally range from 40% to 80% (Blanchard et al., 2001, 2004) with about 50%of that removed during primary sedimentation and the rest by secondarysedimentation. Removal efficiency generally increases with increasing chlo-rination level since the highly substituted PCBs are more lipophilic (Becket al., 1996). As a result, it is the higher chlorinated congeners that are foundin the highest concentrations in biosolids. Both Alcock and Jones (1993)and Stevens et al. (2003) found that congeners 28, 52, 101, 138, 153, and180 were most prevalent in UK biosolids. Although anaerobic cultures havebeen shown to have the capacity for reductive dehalogenation of PCBs(Mohn and Tiedje, 1992), very little degradation is observed duringanaerobic sludge digestion (Buisson et al., 1990).

Following biosolids application to soils, there is a slow loss of PCBs. In along-term study, Alcock et al. (1995) found that total PCB concentrationsdeclined exponentially over a 31-year period (since the last biosolids appli-cation) but sludge-amended plots still contained five times more PCBs thanthe control. They also showed that the 3/4-Cl congeners were lost morerapidly than the 5-Cl homologues and above. Alcock et al. (1993) suggestedthat volatilization is the major loss mechanism of PCBs from soils sinceaerobic biodegradation is characteristically slow.

6.1.5. Organochlorine pesticides (OCPs)The term organochlorine refers to a wide range of organic compoundswhich contain chlorine. The organochlorine pesticides are a diverse groupof synthetic chemicals that have mainly been used against insect pests ofagricultural crops and diseases of humans and domestic animals that are



carried by insects (Lal and Saxena, 1982). The most well known OCPsinclude, DDT, DDD, kelthane, chlorobenzilate, chloropropylate,methoxychlor, aldrin, dieldrin, heptachlor, lindane, endosulfan, isodrin,isobenzan, endrin, chlordane, toxaphene, mirex, and kepone. Their usehas been minimized or terminated in most technologically advancedcountries because of their persistent nature, susceptibility to biomagnifica-tion and toxicity to higher animals (Durdane, 2006; Jorgenson, 2001;Rosario et al., 2007). Some OCPs such as DDT, toxaphane, and methoxy-chlor are known to be strong EDCs (Birkett, 2003a). Some are alsocarcinogenic and/or have immunosuppressant properties (Belpommeet al., 2007).

OCPs are hydrophobic and associate strongly with the solid phase of rawsewage. They are therefore predominantly removed during sedimentationwith a removal efficiency of between 30% and 95% (Garcia Gutierrez et al.,1982, 1984; McIntyre et al., 1981). OCP concentrations have been declin-ing in the environment in recent years because of restrictions regarding theiruse and as a consequence individual compounds are commonly present inbiosolids in only very low concentrations and often some are not detectable(Berset and Holzer, 1999; Falandysz and Strandberg, 2004; Katsoyiannis andSamara, 2004; McIntyre and Lester, 1984; Stevens et al., 2003; Wang et al.,2007; Webber et al., 1996). For example, in a study of Biosolids fromSwitzerland, Berset and Holzer (1999) observed the most frequent com-pounds found were DDT and its reaction products (DDD and DDE),lindane, aldrin, heptachlor epoxide, and endosulfan. When present, con-centrations of DDT, DDD, DDE, and aldrin ranged from 7.5 to 15.5, 15.4to 47.9, 36.6 to 97.2, and 4.4 to 28.9 mg kg� 1, respectively. In Polishbiosolids, Falandysz and Strandberg (2004) also found DDT and its meta-bolites to be present in highest concentrations (330–490 mg kg� 1) withsignificant amounts of dieldrin also being present (8.6–9.9 mg kg� 1). InGreece, Katsoyiannis and Samara (2004) measured heptachlor epoxide,DDD, DDE, dieldrin, and endrin in biosolids in mg kg� 1 amounts butdid not detect DDT, aldrin, or isodrin.

Microbial degradation of OCPs is characteristically slow but promotedunder anaerobic conditions (Hill and McCartney, 1967; Neilson, 1996;Olaniran et al., 2001). Thus, some degradation may occur during anaerobicsludge digestion. Under aerobic soil conditions their degradation is veryslow (Aislabie and Lloyd-Jones, 1995; Aislaibie et al., 1997; Lal and Saxena,1982) and therefore, if they are present in biosolids, they will remain forlong periods following land application.

6.1.6. ChlorophenolsChlorophenols are organochlorines of phenol that contain one or morecovalently bonded chloride atoms. They consist of 19 different compoundsand pentachlorophenol (PCP) is the most widely used in the group. Because



of their broad spectrum biocidal properties, chlorophenols are used asdisinfectants, mothproofing agents, miticides, termiticides, herbicides, andfungicides. They are used as preservation agents for wood, paints, andleather and some are important intermediates in the production of pharma-ceuticals, dyes, and herbicides. PCP is known to be neurotoxic and immu-nosuppressant and is a suspected carcinogen and EDC (USEPA, 2008).In addition, dioxins (see below) are common by-products/contaminantsin PCP formulations and these are known to be extremely persistent andtoxic compounds. Because of concerns about the toxicity of PCP, andassociated dioxins, its use has been severely limited (e.g., banned fromresidential indoor use) in most technologically advanced countries sincethe 1980s. As a result, concentrations in freshwater and marine environ-ments are generally falling (Muir and Eduljee, 1999). Chlorophenols areconsidered priority pollutants by both the European Union and the USEPA(Ruzgas et al., 1995).

Chlorophenols may enter wastewater through runoff from residentialand agricultural land following atmospheric deposition and/or a variety ofuses (e.g., herbicides, fungicides, termiticideswood preservatives), throughdomestic sewers and wastewater from manufacturing processes. Duringwastewater treatment, chlorophenols can be removed by sedimentation,volatilization, and biodegradation. Ettala et al. (1992) showed that biodeg-radation during activated sludge treatment accounted for a large proportionof the removal. Further degradation will occur during sludge digestionparticularly if it occurs under anaerobic conditions. Indeed, biodegradationis known to be favored under anaerobic conditions with the higher chlori-nated phenols (e.g., pentachlorophenol) being sequentially dechlorinated totetra- tri-, di- and monochlorophenol (Chang et al., 1995; Mikesell andBoyd, 1988; Togna et al., 1995). Phenols substituted in the 2,4 and 6positions are generally more readily degraded than 3 and 5 chloro com-pounds. Nevertheless, significant amounts of di-, tri-, tetra- and pentachlor-ophenols have all been measured in biosolids samples (Bright and Healey,2003; DeWalle et al., 1982; Wild et al., 1993). Concentrations of individualcompounds are often in the range of 0.02–10 mg kg� 1 but total concentra-tions in the range of 10–60 mg kg� 1 have been measured by some(Wild et al., 1993). Following land application of biosolids, chlorophenolswould be expected to degrade relatively quickly since microbial degradationunder aerobic soil conditions is well documented (Cho et al., 2000; Okekeet al., 1996).

6.1.7. Polychlorinated dibenzodioxins (PCDDs)and polychlorinated dibenzofurans (PCDFs)

Dioxin is a generic term for a mixture of 219 different almost planar tricyclicaromatic compounds belonging to the PCDD and PCDF groups. The mostextensively studied is the PCCD 2,3,7,8-tetrachlorodibenzo-p-dioxin



(TCDD) which was a contaminant in Agent Orange. Dioxin concentra-tions are usually expressed as International Toxicity Equivalency Factors(I-TEQs) relative to the most toxic dioxin (TCDD). PCCDs and PCDFsare not produced commercially but are formed as unwanted by-productsfrom the production or use of many organochlorine compounds such aschlorophenols and their derivatives, chlorinated diphenyl ethers, poly-chlorinated biphenyls, and chlorophenoxy herbicides. Earlier and/or ongo-ing use of pentachlorophenol (principally as a wood preservative; see above)is often considered as the major source of dioxins in industrialized countries.Chlorine bleaching of pulp and paper can be an important source and theyare also produced in graphite electrode sludge from chlorine-alkali plants(Rappe, 1994). In addition, dioxins are produced during combustion pro-cesses including chlorinated waste incineration, iron and steel production,coal-fired power generation, and in motor vehicle emissions. Dioxinsbioaccumulate in food chains and are classified as persistent organic pollu-tants and under the Stockholm Convention and signatories are obliged toeliminate or minimize all sources. In mammals, they are known to bemutagenic, carcinogenic, immunotoxic, teratogenic, hepatotoxic, andEDCs (Birkett, 2003a; Birnbaum, 1994; Boening, 1998; Peterson et al.,1993).

There is often little difference in PCDD/F concentrations in wastewaterfrom industrial and domestic areas (Rappe et al., 1998; Stevens et al., 2001).It is thought that a major source of higher chlorinated dioxins is that they arewashed out of textiles and fabrics in washing machines and enter householdwastewater (Horstmann and McLaughlan, 1995; McLachlan et al., 1996).By contrast, the lower chlorinated dioxins enter wastewater treatmentplants principally by transport of atmospheric deposition in surface runoff(McLachlan et al., 1996; Oleszek-Kudlak et al., 2005). PCDD/Fs generallyhave low volatility, high octanol water coefficients, and low aqueoussolubilities and consequently they sorb strongly onto organic solids andare mostly removed during sedimentation. Although PCDD/Fs can beslowly degraded by reductive dechlorination under anaerobic conditions(Field and Sierra-Alvarez, 2008; Wittich, 1998), total concentrations havebeen shown to increase during anaerobic sludge digestion due to loss ofsludge biomass (Disse et al., 1995; Oleszek-Kudlak et al., 2005). Concentra-tions of dioxins in biosolids range from 2 to 1270 ng I-TEQ kg� 1 with20–200 ng I-TEQ kg� 1 being common (Dudzinska and Czerwinski, 2002;Eljarrat et al., 1999; Hagenmaier et al., 1992; Oleszek-Kudlak et al., 2005;Rappe et al., 1998; Sewart et al., 1995; Stevens et al., 2001). The predomi-nant isomers present are the higher chlorinated PCDD/Fs with octachlor-odibenzo-p-dioxin (octCDD) commonly being present in highestconcentrations (e.g., 300–30,000 ng kg� 1) (Oleszek-Kudlak et al., 2005).Other dioxins commonly present in biosolids include octaCDF, hep-taCDD, and heptaCDF (Eljarrat et al., 2003; Martinez et al., 2007;



Oleszek-Kudlak et al., 2005). In general, PCDD/F levels in biosolids inEurope have decreased since the 1980s (Eljarrat et al., 2003; McLachlanet al., 1996) and this is presumably a consequence of a ban on the use ofpentachlorophenol.

Once introduced to the soil via biosolids application, PCDD/Fs are verypersistent, with half-lives in excess of 10 years (McLachlan et al., 1996;Wilson et al., 1997). It is likely that some slow biodegradation occurs sincelower chlorinated dioxins can be degraded by aerobic bacteria and can alsobe attacked cometabolically by white rot fungi (Field and Sierra-Alvarez,2008). However, photodegradation and volatilization have a negligibleinfluence on the fate of PCDD/F in soils (McLachlan et al., 1996) and asa result they are exceptionately persistent.

6.1.8. Organotin compoundsOrganotin compounds are organometallic compounds based on tin withhydrocarbon substituents. A large number of compounds exist whichbelong to four classes; tetraorganotins (R4Sn), triorganotins (R3SnX), dior-ganotins (R2SnX2), and monoorganotins (RSnX3). R is usually a butyl,octyl, or phenyl group and X a chloride, fluoride, oxide, hydroxide,carboxylate, or thiloate group. Monoorganotins have limited use mainlyfor thermal and UV stabilization in PVC products. Diorganotins are alsoused as stabilizers for PVC and as well as catalysts in production of polyure-thane foams and in silicone vulcanization. Triorganotins have biocidalproperties and are used in timber protection, protection of textiles, leatherand other materials, and for crop and animal protection in agriculture. Somewere also used as antifouling agents in marine paints but this use is nowprohibited in most countries. Tetraorganotins are mainly used as intermedi-ates in preparation of other organotins. Because of their high toxicity(particularly triorganotins) toward aquatic animals, at concentrations ofonly a few ng L� 1, there is concern about their adverse effects on freshwaterand marine environments (Chiavarini et al., 2003; Diez et al., 2005; Rudel,2003). In mammals, organotins are known to be neurotoxic, carcinogenic,immunotoxic, and also affect reproducibility (Appel, 2004; Dopp et al.,2007).

It is thought organotins entering wastewater streams originate fromindustrial manufacture of PVC, runoff from biocidal applications forwood preservation and agriculture, and even normal leaching andweathering of PVC pipes (Fent, 1996). Significant quantities of organotinsare present in wastewater and concentrations usually range from a few ng toa few (Horstmann and McLaughlan, 1995; McLachlan et al., 1996) mg L� 1

(Chau et al., 1992; Fent and Muller, 1991). The compounds are primarilyassociated with the suspended solids in untreated wastewater (Fent, 1996)and they are mainly removed during primary and secondary sedimentation.They become enriched in biosolids since they are not readily degraded



during sludge digestion (Fent, 1996; Voulvoulis and Lester, 2006). Suchenrichment is clearly evident in Fig. 6 where concentrations of organotinsin digested sludge at four different sampling times were about twice those inundigested sludge. The most commonly encountered organotins in bioso-lids are monobutyltin (0.016–43.56 mg kg� 1), dibutyltin (0.41–8.56 mgkg� 1), tributyltin (0.005–237 mg kg� 1), monophenyltin (0.1 mg kg� 1),

0

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2

22

February 1988 January 1989

February 1988 January 1989

23 28 11 17

22 23 11 17

Con

cent

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)

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Figure 6 Concentrations of monobutyltin (MBT), dibutyltin (DBT), tributyltin(TBT), and total organotins (Total) in raw sewage (top) and digested sewage sludge(below). From Fent and Muller (1991); copyright American Chemical Society.



diphenyltin (0.1–0.4 mg kg� 1), and triphenyltin (0.3–3.4 mg kg� 1)(Arnold et al., 1998; Chau et al., 1992; Fent, 1996; Fent and Muller,1991; Voulvoulis et al., 2004). The fate of organotins in soils followingbiosolids applications and their effects on the soil system is not well known(Fent, 1996). However, it is known that organotins can be slowly biode-graded by bacteria and fungi and that this can involve sequential removal oforganic moieties to yield less toxic derivatives (Gadd, 2000).

6.1.9. Brominated flame retardantsFlame retardants incorporated into potentially flammable materials, such asplastics, rubbers, and textiles, to slow down and/or inhibit the initial phaseof a developing fire. Brominated fire retardants are an extremely diversegroup of compounds including aromatics, cyclic aliphatics, phenolic deri-vatives, aliphatics, and phthalic anhydride derivatives (Hyotylainen andHartonen, 2002). Their flame retardancy mechanism is that with the appli-cation of heat, they decompose before the matrix of flammable polymer,thus preventing the formation of flammable gases (Rahman et al., 2001).The brominated fire retardants most commonly used are polybrominateddiphenyl ethers (PBDEs), polybrominated biphenyls (PBBs), hexabromo-cyclododecane (HBCD), and tetrabromobisphenol (TBBPA) (de Witt,2002). PBDE congeners are named by the number and position of brominesanalogous to PCBs (e.g., BDE-47, BDE-99, etc.). Retardants such asPBDEs, PBBs, and HBCD are additives mixed into polymers and are notchemically bound to the plastic or textile so they may separate or leach fromthe product. Others such as TBBPA are reactive and are chemically bondedto the material and are less likely to be released to the environment. PBDEsand PBBs are lipophilic and resistant to degradative processes and thereforebioaccumulate in wildlife and humans (Hakk and Letcher, 2003; Law et al.,2003, 2006). Mechanisms of toxicity are similar to those of PCBs(Pijnenburg et al., 1995). They are known to be neurotoxins, carcinogen,EDCs and have estrogenic activity (Martin et al., 2004; Rahman et al., 2001;Siddiqi et al., 2003). In 2003, the European Union placed a ban on use ofpentaBDE and in Europe production has shifted toward HBCD andTBBPA.

Sewage from industrial and domestic areas generally show brominatedfire retardant concentrations of a similar order indicating that the majorsource is diffuse leaching from treated products into wastewater streamsfrom both households and industries (Law et al., 2006; Wang et al., 2007).The water solubilities and vapor pressures of PBBs, TBBPA, HBCD, andPBDEs are very low (Rahman et al., 2001) and they strongly bind to solidparticles particularly organic matter (Litz, 2002). As a result, they areremoved during sedimentation and North (2004) estimated 96% removalefficiency in a wastewater treatment plant. Concentrations of PBDEsin biosolids have been reported to range from 0.4 to 2600 mg kg� 1, with



100–1000 mg kg� 1 being common; congeners 47, 99, 100 and 209 are mostprevalent (Christensen et al., 2003; de Witt, 2002; Fabrellas et al., 2004;Hagenmaier et al., 1992; Hale et al., 2003; Law et al., 2006; Moche andThanner, 2004; Oberg et al., 2002). Concentrations of HBCD of 0.4–650,TBBPA of 0.4–300, and PBB of 0.4–10 mg kg� 1 have also been reported(Law et al., 2006; Oberg et al., 2002; Sellstrom et al., 2005).

Following land application of biosolids, PBDEs have been shown to bepersistent in soils (Eljarrat et al., 2008; Sellstrom et al., 2005). Nevertheless,losses of added PBDEs from soils have been measured (Litz, 2002) andthese were greater under anaerobic conditions. Biodegradation is gener-ally not thought to be an important pathway for the PBDEs but photolysismay play an important role (Fang et al., 2008; Rahman et al., 2001).Nevertheless, microbial degradation of PBDEs can occur under aerobicand particularly anaerobic conditions (He et al., 2006; Rayne et al., 2003;Vonderheide et al., 2006).

6.1.10. Surfactants and related residuesSurfactants are one of the most ubiquitous groups of anthropogenic organiccompounds. The global market is more than 18 million tons per year andabout 40% of this is as soaps and household detergents (Cirelli et al., 2008).They are also used by a range of industries in the production of cosmetic,personal care, household, painting, coating, textile, dyes, polymer, food,agrochemical, and oil products. A fundamental property of surfactants istheir ability to form micelles (colloidal-sized clusters) in solution. Theproperty is due to the presence of both hydrophobic and hydrophilic groupson each molecule. This gives surfactants their detergency and solubilizationproperties. Common formulations include anionic, cationic, amphoteric,and nonionic surfactants. The most common surfactant used is the anioniclinear alkylbenzene sulfonate (LAS).

After use, large quantities of soaps and detergents are released intosewage. During passage through the sewerage system, chemical andbiological reactions can result in considerable reductions in concentrationsof surfactants present (Matthijs et al., 1995; Moreno et al., 1990). Manystudies at waste water treatment plants have shown that LASs are efficientlyremoved by physical, chemical, and biological processes and concentrationsin effluent water are characteristically low. Indeed, due to the amphiphilicnature of surfactants, they are readily adsorbed to suspended solids insewage. As a result, a significant proportion of LAS in sewage is adsorbedto particulate matter and is removed during sedimentation. Thus, sedi-ment in primary settling tanks is generally rich in LAS with concentrationsranging from 5 to 15 g L� 1 (Brunner et al., 1988; De Henau et al., 1982).Most of the remaining LAS in the solution phase are removed by micro-bial degradation resulting in a reduction of 95–99% of LAS load in the



liquid phase of most water treatment plants (Cirelli et al., 2008; Painter andZabel, 1989).

In general, LAS, cationic and alkylphenol ethoxylate surfactants are allrelatively resistant to degradation in anaerobic environments (Cirelli et al.,2008). By contrast, they are generally rapidly degraded under aerobicconditions. Biodegradation of LAS involves degradation of the linear alkylchain, the sulfonate group and finally the benzene ring and is carried out bya consortium of bacteria (Cirelli et al., 2008; Perales et al., 2003). Thus,biosolids that have been aerobically digested generally have a low LASconcentration (e.g., 100–500 mg kg� 1) compared to anaerobically treatedsludge (e.g., 5000–15,000 mg kg� 1) (Cirelli et al., 2008; Jenson, 1999).Since anaerobic digestion is the predominant treatment of sludge, biosolidsoften contain a substantial surfactant load. Once reintroduced into anaerobic environment, such as soil, surfactants such as LAS are rapidlydegraded (Cirelli et al., 2008; De Wolf and Feijtel, 1998).

Under both aerobic and anaerobic conditions the alkylphenol ethoxy-lates (nonionic surfactants, e.g, nonylphenol exothylate, nonylphenoldiethoxylate) undergo almost complete primary degradation but underanaerobic conditions degradation by-products tend to persist (Ejlertssonet al., 1999). The nonylphenol group is most resistant (Thiele et al., 1997)and has caused concern since it has been shown to be present in anaerobi-cally digested biosolids in significant quantities (i.e., up to several thousandmg kg� 1) (Maguire, 1999; Xia and Jeong, 2004) and it is known to be to anEDC (Thiele et al., 1997; Ying et al., 2002). Nevertheless, nonylphenol israpidly biodegraded under aerobic conditions in soil ( Jenson, 1999) anddegradation is stimulated by the addition of organic substrates such asbiosolids (Roberts et al., 2006).

Surfactants are used both in flushing/washing of heavy metal-contaminated soils and during bioremediation of soils contaminated withorganics to promote desorption of contaminants bound to the solid soilphase (Haigh, 1996; Mulligan et al., 1999, 2001). However, such interac-tions are unlikely to occur to any significant extent in biosolids-amendedsoils. This is because the concentrations of surfactant necessary to achievemicellization in soil pore water are much greater than those typical inbiosolids-treated soils (Cirelli et al., 2008; Haigh, 1996).

Other detergent-derived residues which have also been measured inbiosolids include quaternary ammonium-based surfactants which are widelyused as fabric softeners (Rogers, 1996). These materials are, however, biode-gradable under aerobic conditions (Sullivan, 1983). Fluorescent whiteningagents are also a component of most modern laundry detergents and thesehave been measured in biosolids at concentrations of 5–100 mg kg� 1 (Poigeret al., 1993, 1998). Such materials are thought to be relatively resistant tobiodegradation in the soil (Devane et al., 2006).



6.1.11. Pharmaceuticals and personal care productsPharmaceuticals and personal care products (e.g., musks) in sewage ariseprimarily from human excreta and urine which contain residues or meta-bolites, wash water containing topically applied chemicals, and from anydeliberate disposal of unwanted and/or expired prescription medicines( Jones-Lepp and Stevens, 2007; Rogers, 1996). Materials used in veterinaryapplications may also be important. Recent advances in chemical analyticalmethodology have revealed that a wide variety of these chemicals can bepresent in waste water and biosolids (Bright and Healey, 2003; Kinney et al.,2006; Koplin et al., 2002). For example, chemicals belonging to groups suchas synthetic musks (2000–500,000 mg kg� 1) (used in perfumes and scents),fluoroquinolones (1000–2500 mg kg� 1) (broad-spectrum antibiotics),macrolides (1–150 mg kg� 1) (a group of antibiotics), and sulfonamides(1–200 mg kg� 1) (antibacterial drugs) have all been identified in biosolids( Jones-Lepp and Stevens, 2007). Specific chemicals such as acetaminophen(e.g., tylenol, panadol), acetylsalicylic acid (asprin), albuterol (a bronchodi-lator used in asthma treatment), and gemfibrozil (lowers blood fats andcholesterol) have also been detected (Bhandari and Xia, 2005; Jones-Leppand Stevens, 2007).

There is little information available on the environmental behavior,biological activity, or ecotoxicology of these materials. It is believed that alarge proportion of pharmaceutical chemicals will undergo microbial trans-formations during sewage treatment processes (Kinney et al., 2006;Richardson and Bowron, 1985). The extent of such transformations andthe biological activity of any metabolites produced are, however, unknown.At present, there is no evidence that residues of pharmaceutical chemicalspresent in biosolids are likely to be harmful to either human health or to theenvironment. Nonetheless, in the future much more accurate data on theirconcentrations, fate in wastewater treatment facilities, environmental fateand transport, and potential effects on humans and the environment will berequired.

6.1.12. Natural and synthetic hormonesThe major hormones associated with the estrous cycle in womenare 17b-estradiol, estradiole, and estrone and all enter sewage in measure-able amounts. The xenoestrogen, 17a-ethinylestadiol, is an analog for17b-estradiol and is used as a key component of oral contraceptives. It isalso present in municipal wastewater. Concentrations of estrogenic sub-stances in wastewater consistently fall in the 40–100 ng L� 1 range (Laytonet al., 2000). The efficiency of biodegradation of both natural and syntheticestrogenic organics and the consequent concentrations in wastewater treat-ment plant effluents and in biosolids varies according to sewage treatmenttechniques with aeration notably increasing efficiency (Esperanza et al., 2004;



Joss et al., 2004; Lorenzen et al., 2004; Servos et al., 2005). The incompletedecomposition of these compounds is of concern since very low concentra-tions in water can contribute to feminization of male fish (Purdom et al.,1994; Routledge et al., 1998). Purdom et al. (1994), for example, foundconcentrations of estradiol as low as 1 ng L� 1 induced vitellogen (an eggyoke precursor protein) production in male fish. The fate of estrogeniccompounds (17a-ethinylestadiol, estradiol, and estrone) in soils treated withbiosolids was followed by Lorenzen et al. (2006). They found that underaerated soil conditions at soil temperatures typical of the Canadian growingseason, these compounds were rapidly degraded (Fig. 7) in three different soilsof different texture and therefore the risk of leaching to groundwaters wasminimal.

6.1.13. Other organic compoundsA wide range of other organic compounds have been measured, in quan-tities in the mg kg� 1 order, in biosolids samples around the world. Theseinclude (a) short-chained and chlorinated aliphatics (e.g., chlorobutane,chloroethane, chloromethane, chloropropane, pentanone, hexanone,butanol, acrylonitrile, propenol), (b) monocyclic and heterocyclic hydro-carbons (e.g., benzene, benzoic acid, analene, styrene, toluene, xylene),(c) nitrosamines and nitroaromatics (e.g., N-nitrosodiphenylamine,

0

20

40

60

80

100

0 1 2 3 4 5 6 7Days

% E

xtra

ctab

le r

adio

activ

ity

LoamSandy loamSilt loam

Figure 7 Relative quantity of extracted radioactivity in three agricultural soils varyingin texture following applications of C14-labeled ethynylestradiol at 120 ng kg� 1. FromLorenzen et al. (2006); copyright NRC Canada.



N-nitrosodiethanolamine), (d) organophosphates (e.g., diazinon, mala-thion, flame retardants), and (e) phenoxy herbicides (e.g., 2,4-D, 2,4,5-T,MCPP, MCPA) (Bright and Healey, 2003; Harrison et al., 2006; Rogers,1996). In general, the fate of such compounds during water treatmentprocesses and sludge digestion and their subsequent breakdown in soilsfollowing land application of biosolids are not well known.

6.2. Potential transfer to groundwater, plants, and animals

With the transfer of organic contaminants to soils when biosolids are landapplied there is the chance of their transfer to the wider environment. Thelikelihood of this will be highly dependent on the amounts and types oforganic compounds introduced to soils and their rates of loss via microbialdecomposition/volatilization/photolysis. As is evidenced by the above dis-cussion, at present, the rates of loss of these compounds from soils are notwell characterized. Below, the general principals of transfer of soil-borneorganic contaminants to groundwater, plants, and animals are discussed.

6.2.1. GroundwaterWilson et al. (1996) used a range of models to assess the potential of organiccontaminants in biosolids to leach and cause groundwater contamination.In general, highly mobile chemicals were considered those with low Kow

and low organic C/water partition coefficients (Koc). A provisional list ofleachable compounds included chloroanalines, nitrobenzene, nitrochloro-phenol, dinitrochlorophos, tetrachloroethane, chloroethyl ether, trichloro-phon, linuron, atrazine, and simazine. Hydrophobic organic compoundssuch as PCBs, PAHs, dioxins, and phthalates are not generally mobile insoils since they strongly sorbed to solid soil and biosolids particles (particu-larly organic surfaces). Thus, organic contaminants in sewage that readilypartition into the solid phase and accumulate in biosolids are generally likelyto be characteristically immobile in soils. The effect of increasing soilorganic matter content is typically to decrease the mobility of these hydro-phobic compounds (Petruzzelli et al., 2002). Even so, the presence ofdissolved organic matter can result in mobilization of a very small percent-age (e.g., <5%) of the organic contaminant (Kretzschmar et al., 1999).

As discussed previously, dissolved organic matter released from thebiosolids can leach down the soil profile. Other work has shown thatdissolved organic matter is involved in transport of organic contaminantsin soils (Kretzschmar et al., 1999). In column and batch studies, complexa-tion or association of strongly hydrophobic substances (e.g., polychlorinatedbiphenyls, polyaromatic hydrocarbons, and organochlorine pesticides) withdissolved organic matter has been shown to result in enhanced aqueoussolubility, decreased sorption, and enhanced transport (Chin et al., 1991;Chiou et al., 1986; de Jonge et al., 2002; Dunnivant et al., 1992; Enfield



et al., 1989; Kan and Tomson, 1990; McCarthy and Jimenez, 1985). Indeed,dissolved organic matter, derived from biosolids, has been shown toenhance transport of organic contaminants through soil (Hassett andAnderson, 1982; Muszkat et al., 1993; Raber and Kogel-Knabner, 1997;Vinten et al., 1983). Thus, biosolids addition to soils can potentially pro-mote leaching of organic contaminants present in the biosolids down thesoil profile. Nonetheless, soluble organic matter can also sorb to soil surfacesand this would enhance sorption and retard transport of the associatedorganic contaminants (Totsche et al., 1997).

The potential for losses of organic contaminants via surface runoff andleaching certainly exists. Preferential flow of biosolids colloids down soilmacropores could contribute to leaching of organic contaminants. Move-ment of organic contaminants sorbed to particulate matter is favored wheremacropore flow occurs (de Jonge et al., 2002). As with heavy metals, thegreatest risk of leaching of organic contaminants is likely to be immediatelyafter land application of biosolids when soluble organic matter concentra-tions are elevated and when preferential macropore flow of biosolids parti-cles is most probable. Minimizing water inputs and drainage during thisperiod is therefore an important consideration. Nonetheless, even whereexceptional circumstances promote runoff and/or macropore flow(e.g., storm events), these are likely to be isolated incidents where dilutionof ground or surface water from unpolluted water will be large.

6.2.2. PlantsThere are four main pathways by which organic chemicals in soil can enterplants (Ryan et al., 1988; Topp et al., 1986). These are: (a) root uptake fromsoil solution and subsequent translocation to shoots (i.e., liquid phasetransfer), (b) absorption by shoots (or roots) of volatilized organics fromsurrounding air (i.e., vapor phase transfer), (c) uptake by external contami-nation of shoots by soil or dust and subsequent retention in the cuticle orpenetration through it, and (d) uptake and transport in oil channels whichare found in some oil-containing plants (e.g., carrots). The first three routesare shown in Fig. 8 which demonstrates the importance of the abovegroundportion of the plant to uptake of organics. Only a few studies have beenpublished detailing the behavior of organic pollutants in crops after applica-tion of contaminated sewage sludge to land but it is generally consideredthat a combination of (a) and (b) accounts for most of the uptake (Beck et al.,1996; Duarte-Davidson and Jones, 1996).

6.2.2.1. Root uptake and translocation Except for a few hormone-likechemicals such as phenoxy acid herbicides, uptake of organic compounds byplant roots is generally considered as a passive process (Collins et al., 2006).PAHs, chlorobenzenes, PCBs, and PCDD/Fs have all been found to accu-mulate significantly in plant roots (Duarte-Davidson and Jones, 1996).

Transportation in thetranspiration streamwithin the xylem

Dry and wet deposition ofparticles followed bydesorption into leaf

Suspension of soilparticles by windand rain

Desorption from soilfollowed by root uptakefrom soil solution

Evaporation andvolatilization from leaf

Volatilization fromsoil

Gaseous deposition to leafvia cuticle and stomata

Figure 8 Principal uptake pathways for the uptake of organic chemicals by plants.From Collins et al. (2006); copyright American Chemical Society.



Lipophilic organic compounds possess a greater tendency to partition intoplant root lipids (including lipids in cell walls and cell membranes) thanhydrophilic ones (Collins et al., 2006). Indeed, it is generally consideredthat there is a linear relationship between Kow and retention of organiccompounds in roots ( Duarte-Davidson and Jones, 1996); compounds withKow> 4 have a high potential for retention on plant roots. Such compoundswill therefore have a high potential for contamination of the surface of rootcrops although fortunately the outer layers are usually removed prior tohuman consumption. Plant species with a higher lipid content may tend tohave a greater root accumulation of organic compounds (Gao and Zhu, 2004;Schwab et al., 1998). Gao and Zhu (2004), for example, found the ability of12 plant species to accumulate phenanthrene and pyrene in roots was posi-tively correlated with root lipid content. The higher retention of organiccompounds in the peels of carrots and potato compared to their innerportions has been attributed to the higher lipid content of the peel (Fismeset al., 2002, Wild and Jones, 1992).

Uptake of organic compounds into the root system, and subsequenttranslocation to shoots through the xylem, has been found to be favored forcompounds of intermediate solubility (Collins et al., 2006). Log Kow valuesfor maximum translocation have variously been reported as being in therange of 1.8–3.0 (Fismes et al., 2002; Smith and Jones, 2000; Trappand Pussemeir, 1991). For example, Collins et al. (2006) noted thatmono- and dichlorophenols, which have a range of log Kow values between2.15 and 3.65, have been shown to possess a high potential for root uptake



and translocation. The reason for the Kow optima is not completely under-stood but it appears that highly lipophilic compounds with high Kow values(e.g., log Kow> 4) are retarded at the root surface due to sorption whilepolar chemicals with low values (e.g., Kow< 0.5) are less able to crosshydrophobic lipid membranes (Collins et al., 2006; Duarte-Davidson andJones, 1996).

Water and solutes transported in the xylem may be sorbed by stemcomponents and/or accumulate in shoots and leaves. The organic chemicalsaccumulate in shoots as a result of equilibration of the aqueous phase in plantshoots with xylem constituents and sorption of the compound onto lipo-philic shoot solids (McFarlane, 1994). Plant transpiration streamflow rateand lipophilic solids content of the plant will greatly influence accumulationof compounds in shoots (Collins et al., 2006). The potential for substantialsorption of the compounds to stem components during xylem transportincreases with increasing lipophilicity of the chemical (Briggs et al., 1983;McGrady et al., 1987).

Plant species can vary greatly in their ability to absorb and translocateorganic compounds (Collins et al., 2006). Mattina et al. (2000, 2003, 2004)have shown that among a wide range of crops, plants within the Cucurbi-taceae family (e.g., Cucurbita pepo L, zucchini) are especially adept at uptakeof soil-bound DDE and chlordane. Similarly, Hulster et al. (1994) showedthat courgette and pumpkin absorb and translocate relatively more PCCD/Fs than most other crops. It has been suggested that these plants release rootexudates which mobilize organic compounds from the soil thus increasingtheir availability (Hulster et al., 1994; Mattina et al., 2000).

It is important to note here that soil characteristics are also importantfactors and that plant uptake of organic compounds is usually inverselyrelated to soil organic matter content (Beck et al., 1996; Ryan et al.,1988). This is because hydrophobic organic compounds with a high Kow

are strongly adsorbed to organic matter particles. As a result, plant uptake oforganic compounds (e.g., PAHs) from solution culture is characteristicallymuch higher than from soils at the same concentration of compound in soilsolution (Gao and Ling, 2006). Biosolids particles are about 50% organicmatter, so addition of biosolids to soils will tend to increase sorption (andcompounds will already be sorbed to biosolids particles). The desorptionprocess of these compounds from solid phase into soil solution will restricttheir uptake into plant roots. The availability of organic compounds in soilsolution is, therefore, usually the primary restriction to plant uptake ofcompounds including chlorobenzenes, organochlorines, and PAHs (Gaoand Ling, 2006; Mattina et al., 2003; Wang and Jones, 1994a). Indeed,organic compounds are usually present in biosolids in low concentrationsand they are so strongly sorbed by the soil/biosolids matrix that they exhibitvery low bioavailability and accumulate in the edible portion of food cropsat extremely low concentrations (O’Connor, 1996). Even where root



accumulation occurs, translocation to shoots is usually negligible due totheir low water solubility and high Kow values. Indeed, shoot concentrationfactors (ratio of concentration of organic contaminant in shoots to that insoil solution) are often an order of magnitude less than the equivalent valuesfor root concentration (Gao and Ling, 2006).

6.2.2.2. Uptake from leaves In general, accumulation of organic con-taminants into aboveground vegetation from biosolids-amended soils isthought to be dominated by vegetative uptake of contaminated vaporfrom the surrounding air (Beck et al., 1996; O’Connor, 1996). This hasbeen demonstrated to be the major uptake pathway into plant foliage for arange of organic compounds including PAHs, PCBs, and PCDD/Fs(Bohme et al., 1999; Simonich and Hites, 1994; Welsch-Pausch et al.,1995). Compounds with high volatility and lipophilicity will have thehighest bioconcentration into foliage. Compounds with a Henry’s lawconstant above 1� 10� 4 are generally considered susceptible to volatiliza-tion (Duarte-Davidson and Jones, 1996). Once volatilized from the soil intothe atmosphere, chemicals may subsequently diffuse into plant leaves via thecuticle or stomata. Chemicals entering the leaves will diffuse into intercel-lular air spaces and partition to aqueous and lipophilic phases of adjacentplant tissues (Collins et al., 2006). Lipophilic leaf tissues include the waxycuticle as well as membrane lipids, storage lipids, resins, and essential oils.Lipophilicity of the organic compound is therefore important and com-pounds with a log Kow of greater than 4 and high volatility have the greatestpotential for foliar uptake (Duarte-Davidson and Jones, 1996). Partition ofvolatilized chemicals between plant foliage and air has also been related tothe octanol/air partition coefficient (KOA) and a linear relationship betweenlog KOA and shoot concentrations of PAHs and chlorobenzenes has beennoted (Kipopoulou et al., 1999; Wang and Jones, 1994a). Differences incuticular permeabilities and leaf/air bioconcentration for volatile organiccompounds between plant species has been related principally to leaf lipidcontent (Collins et al., 2006).

Organic chemicals may also come in contact with foliage followingdeposition in association with dust, aerosols, or atmospheric particulatematter. Some of this may accrue from resuspension of soil/biosolids parti-cles. More particularly, chemicals may come in contact with foliage follow-ing direct application (e.g., surface application of biosolids to pastures).Once in contact with the leaves, particle-bound organic chemicals maydiffuse through the cuticle and become sorbed to lipophilic material orpermeate into the leaf interior (Collins et al., 2006). The permeability of thecuticle to organic chemicals in solution has been observed to be linearlyrelated to Kow and inversely related to its molecular size (Riederer et al.,2002). Uptake of chemicals from particulate deposits will, however, be



more complex since it will also depend on the capability of the compoundto desorb from particle surfaces.

6.2.2.3. Other roles of plants The presence of plants can be an importantfactor in increasing the rate of degradation of organic contaminants in soils(Alkorta and Garbisu, 2001; Pilon-Smits, 2005; Salt et al., 1998) includingthose amended with biosolids (Laturnus et al., 2007). As a result, phytor-emediation, which uses plants and associated rhizosphere microorganisms toremove or transform organic contaminants in soils, is an emerging technol-ogy (Alkorta and Garbisu, 2001; Susarla et al., 2002). Phytoremediation hasbeen successfully employed to treat soils contaminated with a range oforganics including chlorinated solvents, aromatic compounds, surfactants,and explosives (e.g., trinitrotoluene, TNT) (Susarla et al., 2002).

Plants release a range of carbonaceous compounds into the rhizosphereas root exudates. As a result, microbial densities are 1–4 orders of magnitudehigher than the surrounding bulk soil and the community also has a greaterrange of metabolic capabilities (Salt et al., 1998). Particular plant speciescould promote degradation of specific organic compounds. For example, ithas been suggested that some species preferentially release phenols capableof supporting PCB-degrading bacteria (Fletcher and Hegde, 1995). Rhizo-sphere microorganisms may also accelerate degradation by volatilizingorganics such as PAHs (Alkorta and Garbisu, 2001). Some bacteria canrelease biosurfactants that may make hydrophilic pollutants more watersoluble (Volkering et al., 1998) and plant exudates can contain lipophiliccompounds that increase pollutant water solubility and/or promote biosur-factant-producing microbial populations (Siciliano and Germida, 1998). Inaddition, both plant roots and microorganisms can release enzymes into soilswhich are involved in degradation of organic pollutants (e.g., laccases,dehalogenases, nitroreductases, nitrilases, and peroxidases) (Schnoor et al.,1995; Wolfe and Hoehamer, 2003). Enhanced biodegradation of organicsinduced by rhizosphere microflora has been reported for a number ofcompounds including pentachlorophenol (Ferro et al., 1994), surfactants(Knabel and Vestal, 1992), TCE (Shim et al., 2000; Walton and Anderson,1990), PAHs ( Joner et al., 2002), and phenanthrene (Corgie et al., 2004;Fang et al., 2001). It has been suggested that in biosolids-amended soils theabove mechanisms are of minor importance because biosolids already con-tain high concentrations of labile organic compounds and a large, activemicrobial community (Laturnus et al., 2007). These workers suggested thatimproved porosity and aeration induced by the presence of growing plantsfavors microbial activity and their degradative activities in biosolids-amended soils.

As already discussed, uptake of organic contaminants by plants is usuallylow, so phytoextraction as a mechanism of remediation of organic con-taminants is not usually a viable option. However, uptake can contribute to



phytoremediation. Following uptake, organic compounds have multiplefates. They may be transformed to less toxic compounds and bound inplant tissues in nonavailable forms, they may undergo partial or completedegradation (phytodegradation) or they may be volatilized (phytovolatiliza-tion) (Alkorta and Garbisu, 2001; Susarla et al., 2002). Phytodegradationoccurs when plant enzymes act on compounds and catalyze them partiallyor fully, thus reducing their concentrations in plant tissue. This can occur inboth the root and shoot tissue. Those chemicals where evidence of signifi-cant degradation has occurred include trichloroethylene, benzene, pyrene,and TNT (Hannink et al., 2002; Susarla et al., 2002). Phytovolatilizationoccurs when volatile organic pollutants with a high water solubility andvapor pressure are absorbed, translocated to the leaves and lost to theatmosphere as a gas. Examples of organics that can be volatilized from plantsinclude the solvents benzene and TCE and the fuel additive MTBE (Collinset al., 2006; Ma and Burken, 2003; Pilon-Smits, 2005).

In general, concentrations of organic pollutants in biosolids-amendedsoils will be considerably lower than those present in situations wherephytoremediation might be considered a strategy to clean up a contami-nated soil. Nonetheless, since biosolids are generally land applied to agricul-tural land, the presence of growing plants does seem likely to enhancethe degradation of many biosolids-borne organic contaminants that areintroduced to soils.

6.2.3. AnimalsWhen biosolids is applied to agricultural land the potential transport of toxicorganic chemicals to human food products is of concern. The pathwayscausing most apprehension are in animal production systems becausepersistent, lipophilic compounds bioconcentrate in body fat (i.e., meatproducts) and fat-containing products (e.g., milk). The pathways throughwhich biosolids-derived compounds enter the grazing animal, and factorsaffecting this entry, were discussed previously (see, Section 5.5). Since mostorganic compounds are not particularly mobile and do not accumulate inlarge amounts in plant tops, the biosolids–soil–plant–animal pathway isusually of minor importance. The greatest intake from herbage will occurwhen biosolids are applied directly to established pasture, biosolids adhere tothe herbage and animals have immediate access. Nevertheless, as with heavymetals, soil/biosolids ingestion during periods when forage is sparse is likelyto be the major pathway of entry of biosolids-derived organic compoundsinto the grazing animals (Fries, 1996). As noted previously, this typicallyoccurs most during winter when forage availability is least.

Of the many organic contaminants present in biosolids, lipophilic,halogenated hydrocarbons (e.g., halogenated biphenyls, chlorinated pesti-cides and hydrocarbons, and PCDD/Fs) are of primary concern since theyare resistant to degradation and bioconcentrate in fat of animals and animal



products (Fries, 1996). Among these, compounds with low degrees ofhalogenation are metabolized and do not accumulate but higher degreesof halogenation block metabolism and bioconcentration occurs(McLachlan, 1993; Shiu and Mackay, 1986). Other compounds commonlyfound in biosolids such as phthalates, PAHs, acid phenols, nitrosamines,volatile aromatics, and aromatic surfactants are metabolized and do notgenerally accumulate in animal tissues (Fries, 1996). Bioconcentration fac-tors (the ratio of concentration of compound in a tissue or product to theconcentration in the diet) for halogenated organics (e.g., DDE, dieldrin,hexachlorobenzene, heptachlor, PCBs, chlorodibenzo-p-dioxin, and chlor-odibenzofuran) are usually in the range of 5–6 in milk fat of cows and thebody fat of sheep and cattle (Fries and Marrow, 1975; Fries et al., 1969;Harrison et al., 1970; Parker et al., 1980; van den Hoek et al., 1975; Willettet al., 1990). To minimize bioaccumulation of halogenated hydrocarbons ingrazing animals it is important to allow time for biosolids to be washed offthe surface of pasture leaves prior to grazing and minimize the chances ofsoil ingestion during the winter months.

7. Synthesis and Conclusions

Wastewater derived from domestic sources (human feces, urine, andgraywater) plus that from commercial enterprises and industry, and thatfrom runoff into stormwater is treated in municipal wastewater treatmentplants in a series of processes primarily aimed at removing the dissolved andsuspended organic material. Biosolids are by-products of wastewater treat-ment and consist of approximately 50% inorganic and 50% organic material.The organic components originate principally from two sources: (i) humanfeces settled out during primary treatment (sedimentation) and (ii) bacterialcells settled out during secondary treatment in which bacterial activity isused to remove the suspended/dissolved organic load from the primarytreatment effluent. The organic components undergo a degree of degrada-tion and humification particularly during sludge stabilization (often anaero-bic digestion). The inorganic component mainly settles out duringsedimentation and originates from sources such as local soil and sediments,broken glass, and inorganic residuals of food/feces (e.g., silica).

Biosolids contain both inorganic and organic contaminants which origi-nate from the influent sewage wastewater. The major inorganic contami-nants are heavy metals (e.g., Cu, Zn, Cd, Pb, Ni, Cr, As) which mainlyoriginate from discharges from industry and from domestic graywaterincluding leaching from Cu and Pb pipes and Zn from domestic products(skin creams, deodorants, etc.). Heavy metals enter the sludge duringprimary treatment through their association/adsorption to sedimenting



particles and during secondary treatment through adsorption to bacterial cellwalls and/or accumulation into bacterial cells. Concerns regarding theheavy metal loads in biosolids have resulted in guidelines and regulationsbeing developed in many parts of the world which are usually based onmaximum allowable metal concentration limits in biosolids and/or theallowable loading limits of metals added in biosolids to the soil. A limitationof these approaches is that they consider total rather than biologically-active(extractable) concentrations of heavy metals in soils. Where agronomic ratesare used (2–8 Mg ha� 1), heavy metal toxicities limiting crop growth inbiosolids-amended soils are very rare. Nevertheless, a major concern is thatheavy metals may accumulate in edible portions of plants and subsequentlyenter the food chain and have toxic effects on grazing animals and/orhumans ingesting them. For the most part, heavy metals are not readilytranslocated to the aboveground edible portions of crops so toxicities fromingestion of food crops are not likely under current regulations. The mainpotential pathway for accumulation of heavy metals into the meat of grazinganimals is via direct soil/biosolids ingestion and substantial accumulation isunlikely under adequate grazing management. There is a growing body ofevidence that biosolids-induced heavy metal accumulation in soils can havenegative effects on soil microbial/biochemical activity but the significance/importance of this has yet to be fully understood.

With recent improvements in extraction techniques and analytical chro-matographic methods, there is an increasing body of research involvingsurveys of organic contaminants in biosolids. These include PAEs, PAHs,PCBs, chlorobenzenes, chlorophenols, dioxins, organotins, pesticides, bro-minated flame retardants, surfactants, pharmaceuticals, and natural andsynthetic hormones. Due to their low water solubility and high lipophilicitythey are generally believed to partition into sludges during sedimentation.Nonetheless, the fate of these chemicals during wastewater treatment pro-cesses is not well characterized. Many are not readily degradable underanaerobic conditions and therefore persist in biosolids following anaerobicdigestion. There is very limited information in the literature on the behav-ior and fate of biosolids-borne organic contaminants in agricultural soils.However, the literature available suggests some are rapidly lost followingland application since they are readily degraded under aerobic soil condi-tions, lost by volatilization or by photolysis. Some are, however, known orthought to be only slowly degraded (e.g., organotins, brominated flameretardants, and high molecular mass PAHs) and some are highly persistent(e.g., dioxins and organochlorines). Uptake of organic contaminants fromsoil into plants is characteristically low and the main pathway is via volatili-zation from the soil surface and then uptake of contaminated air by above-ground vegetation. Accumulation of organics into grazing animals couldoccur through ingestion of biosolids adhering to vegetation directly after isapplication and by direct soil/biosolids ingestion. Many organics are



metabolized in animals but halogenated biphenyls, chlorinated pesticidesand hydrocarbons, and PCDD/Fs are of primary concern since they areresistant to degradation and bioconcentrate in animal fat and animal pro-ducts (e.g., milk). The European Union has proposed limit values for severalorganic contaminants/groups of contaminants in biosolids although therelevance of these is not yet fully understood. A better understanding ofthe reactions and fate of organic contaminants in soils is required beforemore rigorous regulations can be drafted.

If sewage entering a wastewater treatment plant contains a substantialloading of heavy metals or organic contaminants, it is very likely thatbiosolids produced at the plant will also have a substantial contaminantload. Thus, to lower contaminant levels in biosolids, reductions in therelease of contaminants into the sewer system are required. The mostpracticable way of doing this is at point source outlets to the sewer. Indeed,in recent years, enforcement of tighter regulations has greatly reducedindustrial inputs of heavy metals to wastewater streams since industry isexpected to pretreat their wastewater (where appropriate) prior to itsrelease. In addition, the use of extremely persistent organics such as dioxinsand organochlorines has been effectively terminated making contaminationof wastewater with such compounds less pronounced. As a result, contami-nants in biosolids are becoming less of an issue and the positive effects ofrecycling the biosolids via land application onto agricultural or forestry landsare gaining momentum.

Indeed, land application is generally seen to be the most economical andbeneficial way to deal with biosolids. They contain high concentrations ofN, P, Ca, and Mg but K is usually low and needs to be supplemented forcrop production. Biosolids applications need to be based on the N needs ofcrops to be grown so that excessive applications can be avoided thusminimizing leaching and gaseous losses of N to the surrounding environ-ment. Because the bulk of N in biosolids is in organic form, an estimate of Nmineralization potential is required to calculate effective N rates. The Pcontent of biosolids is high relative to the plant requirement (based onoptimum N supply) and care needs to be taken to avoid runoff losses of Pand/or P leaching if available P becomes very high. Where available soilmicronutrient levels are low, supply of micronutrients such as Cu, Zn, Fe,and Mn in biosolids may be important. Biosolids applications have otherpositive effects. Addition of partly humified organic matter can increase soilorganic matter content, thereby improving soil physical conditions (e.g.,increased porosity and aggregation) and increasing the size and activity ofthe soil microbial biomass and soil enzyme activity.

Areas where a better understanding is required and where research needsto be concentrated include:



(1) Development of appropriate measures of the biologically active pool ofheavy metals in biosolids-amended soils and development of guidelinesusing such values. Continued monitoring of long-term field sites usingsuch measures is needed to understand long-term effects.

(2) A better understanding of the negative effects of sludge-borne heavymetals on soil microbial and biochemical activity and of the significanceof this to the soil system.

(3) A better understanding of the fate of organic contaminants in sewageduring wastewater treatment processes.

(4) An in-depth understanding of the fate of organic contaminants in soils(and the mechanisms of their degradation) following land application ofcontaminated biosolids. An understanding of the effects of accumulatedorganic contaminants on soil processes is also needed.

(5) Development of scientifically based critical concentrations/loadings fororganic contaminants.

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