the influence of wastewater irrigation on the transformation andâ„¢bioavailability of heavy...

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The Influence of Wastewater

Irrigation on the Transformation

and Bioavailability of Heavy Metal

(Loid)s in Soil

Anitha Kunhikrishnan,*,†,} Nanthi S. Bolan,*,† Karin Muller,‡

Seth Laurenson,§ Ravi Naidu,*,† and Won-Il Kim}

Contents

1. In

s in

065

re foaliaeraidems Mandican-s

troduction

Agronomy, Volume 115 # 2012

-2113, DOI: 10.1016/B978-0-12-394276-0.00005-6 All rig

r Environmental Risk Assessment and Remediation, University of South Australia, Ma

tive Research Centre for Contamination Assessment and Remediation of the E, Australiaodelling, The NZ Institute for Plant and Food Research Ltd., Hamilton, New ZealEnvironment, AgResearch Ltd, Invermay, New Zealandl Safety Division, Department of Agro-Food Safety, National Academy of Agricultui, Gyeonggi-do, Republic of Korea

Else

hts

wso

nvi

and

ral

216

2. S
ources of Wastewater and Heavy Metal(Loid)s in Soils 219
2
.1. W astewater production and quality 219
2
.2. H eavy metal(loid) sources 227
3. E
ffects of Wastewater Irrigation on Soil Properties Affecting Heavy
Metal(Loid) Interactions
231
3
.1. S oil chemistry 231
3
.2. S oil biology 242
3
.3. S oil physics 245
4. E
ffect of Wastewater Irrigation on Heavy Metal(Loid) Dynamics
in Soils
248
4
.1. A dsorption 248
4
.2. C omplexation 249
4
.3. R edox reactions 253
4
.4. M ethylation/demethylation 256
4
.5. L eaching and runoff 258
5. B
ioavailability of Wastewater-Borne Heavy Metal(Loid)s in Soils 261
5
.1. C hemical extraction 262
5
.2. B ioassay 265
6. C
onclusions and Research Needs 271
Refe
rences 273
vier Inc.

reserved.

n Lakes,

ronment,

Science,

215

216 Anitha Kunhikrishnan et al.

Abstract

With pressure increasing on potable water supplies worldwide, interest in using

alternative water supplies including recycled wastewater for irrigation purposes

is growing. Wastewater is derived from a number of sources including domestic

sewage effluent or municipal wastewater, agricultural (farm effluents) and

industrial effluents, and stormwater. Although wastewater irrigation has many

positive effects like reliable water supply to farmers, better crop yield, pollution

reduction of rivers, and other surface water resources, there are problems

associated with it such as health risks to irrigators, build-up of chemical

pollutants (e.g., heavy metal(loid)s and pesticides) in soils and contamination

of groundwater. Since the environment comprises soil, plants, and soil organ-

isms, wastewater use is directly associated with environmental quality due to

its immediate contact with the soil–plant system and consequently can impact

on it. For example, the presence of organic matter in wastewater-irrigated sites

significantly affects the mobility and bioavailability of heavy metal(loid)s in the

soil. Wastewater irrigation can also act as a source of heavy metal(loid) input to

soils. In this chapter, first, the various sources of wastewater irrigation and

heavy metal(loid) input to soil are identified; second, the effect of wastewater

irrigation on soil properties affecting heavy metal(loid) interactions is

described; and third and finally, the role of wastewater irrigation on heavy

metal(loid) dynamics including adsorption and complexation, redox reactions,

transport, and bioavailability is described in relation to strategies designed to

mitigate wastewater-induced environmental impacts.

1. Introduction

In many parts of the world, continued extraction of freshwater forvarious activities including irrigation have led to unsustainable rates of waterconsumption, which has not been assisted by declining rainfall and increasedrationing of water to the ecosystem (Brown, 2007; Seckler et al., 1998).Considerable pressure is now being placed on communities, particularlyprimary producers, to improve water-use efficiency and use alternativewater supplies including recycled wastewater sources for irrigation, in amuch better way. Although using wastewater for irrigation raises concernsabout public exposure to pathogens and contamination of soil, surface water,and groundwater, under controlled management these water sources can beemployed safely and profitably for irrigation (Plate 1) (Qadir et al., 2007).

Wastewaters originate from a number of sources including domesticsewage (municipal wastewater), agricultural, urban and industrial effluents,and stormwater. Wastewater irrigation has many beneficial effects, includinggroundwater recharging (Asano and Cotruvo, 2004) and nutrient supply toplants (Anderson, 2003). There are, however, some detrimental effects, such asbuild-up of salts, pesticides, and heavy metal(loid)s. At sites irrigated with

A

B

Plate 1 Recycled water irrigation in horticultural crops (A) Carrots, (B) Olives (Bolanet al., 2011a).

Wastewater and Bioavailability of Heavy Metal(Loid)s in Soil 217

wastewater, mobilization and transport of pesticides and heavy metal(loid)sinto groundwater have been noted, as well as their enhanced bioavailability tosoil biota and higher plants. For example, dissolved organic matter (DOM)present in wastewater and sewage sludge has been shown to facilitate thetransport of both pesticides and heavy metal(loid)s (Ashworth and Alloway,2004; Bolan et al., 2011a; Muller et al., 2007; Sedlak et al., 1997; Tam andWong, 1996; Thevenot et al., 2009). Wastewater irrigation and sludge appli-cation have also been shown to act as a source of heavy metal(loid) input tosoils (Barman et al., 2001; Eriksson and Donner, 2009; Murtaza et al., 2008).

The term “heavymetal(loid)” in general includes elements (bothmetals andmetalloids) with an atomic density greater than 6gcm�3 [with the exception


of arsenic (As), boron (B), and selenium (Se)]. This group includes bothbiologically essential [e.g., cobalt (Co), copper (Cu), chromium (Cr), manga-nese (Mn), and zinc (Zn)] and nonessential [e.g., cadmium (Cd), lead (Pb), andmercury (Hg)] elements (Sparks, 2003). Heavy metal(loid)s reach the soilenvironment through both pedogenic (or geogenic) and anthropogenic pro-cesses. Anthropogenic activities, primarily associated with the disposal ofindustrial and domestic waste materials including wastewaters and biosolids,are the major sources of metal(loid) enrichment in soils (Adriano, 2001).

Although the role of wastewater irrigation on the transport of pesticideshas been reviewed recently (Muller et al., 2007), no comprehensive reviewhas focused on its role in the mobilization, transport, and bioavailability ofheavy metal(loid)s in soil. This review aims to classify the different sourcesof wastewater irrigation and heavy metal(loid) input to soil. It describes theinfluence of wastewater irrigation on soil properties affecting heavy metal(loid) interactions and explains the role of wastewater irrigation on heavymetal(loid) dynamics including adsorption and complexation, redox reac-tions and bioavailability (Fig. 1). Whilst some literature reviews haveexamined metal(loid) input through inorganic fertilizers, sewage sludge,

Treated sewage Stormwater Farm dairy effluent Piggery effluent Winery effluent

Changes in soil

properties

pH, EC, CEC TOC, DOC Sodicity and salinity Aggregate stability BD and total porosity HC and infiltration

Wastewater

irrigation

Metal(loid) sink

Adsorption Complexation Precipitation Redox reaction

Metal(loid) source

Plant uptake Microorganisms Earthworms Volatilization/demethylation Leaching

M+M+

M+

M+

M+ M+

M+ M+ M+

Figure 1 Schematic representation of wastewater sources and their effect on metal(loid) transformation and fate in soils by acting as a source and sink for metal(loid)s andby altering soil properties.


and atmospheric deposition (Adriano, 2001; Smith, 2009), most reviews onwastewater irrigation have focused on environmental issues of nutrients andsalt accumulation (Bolan et al., 2009; Bond 1998; Carpenter, 1998). Therehas been no comprehensive review on the input of metal(loid)s via effluentand wastewater, and the subsequent transformation and bioavailability ofeffluent-borne metal(loid)s in soils. Unlike wastewater irrigation, a numberof studies have assessed the environmental implications of metal(loid)s derivedfrom sewage sludge and manure slurry application to soils (Cornu et al., 2001;McBride, 2002; McGrath et al., 1994). Since sewage sludge is derived duringwastewater treatment and there are major resemblances in the compositionand chemical properties between these two resources, some of the informa-tion on the distribution and bioavailability of metal(loid)s is inferred fromsewage sludge and manure research. An improved knowledge of wastewaterirrigation’s long-term effects on metal(loid) dynamics in soils can enhance thedevelopment of strategies to mitigate environmental impacts and maximizethe benefits of wastewater as a viable irrigation source.

2. Sources of Wastewater and Heavy

Metal(Loid)s in Soils

2.1. Wastewater production and quality

As indicated above wastewaters originate from a number of sources includingdomestic sewage, agricultural and industrial effluents, and stormwater.Recycled water is defined as wastewater that is treated and reused to supple-ment water supply (US EPA, 1992). The beneficial utilization of treatedwastewater for agriculture is the major water reuse application worldwide(US EPA, 2004). This water source can have the advantage of being aconstant, reliable water source and furthermore reduces the amount ofwater extracted from the environment (Toze, 2006). Approximately 70%of the world’s water resources including all the water from underground andredirected from rivers is used for agricultural irrigation, so reusing treatedwastewater for agricultural and landscape irrigation (Plate 1) reduces both theamount of water that has to be extracted from natural water sources and theuncontrolled discharge of wastewater to the environment (Pedrero et al.,2010). Thus, treated wastewater is a valuable water source for recycling andreuse, especially in the Mediterranean countries and other arid and semi-aridregions including Australia with increasing water shortages (Pedrero et al.,2010). Table 1 summarizes the amount of wastewater generated and reusedannually in selected countries. For example, 88% and 70% of the recycledwater in Spain and Israel, respectively, is used for agricultural purposes(Kanarek and Michail, 1996; Lallana et al., 2001).

Table 1 Wastewater generated and reused annually in selected countries

Country Wastewater generated (GL) Wastewater reused (GL) % Reuse

Argentina 200.3 90.7 45.28*Australia 1634 262.9 16.09

Bahrain 45 8 17.77

Bolivia 135.8 – –

Chile

Greece

295.6

–

–

0.7

–

–

Egypt 10,012 200 1.998*India 13,870 1460 10.53

Jordan 82 64.9 79.15

Kuwait 119 52 43.69

Libya 546 40 7.332

Mexico

New Zealand

13,340

67

280

16

2.104

23.88

Oman 78 8.6 11.03

Peru 34.7 18.6 53.60

Saudi Arabia

Spain

730

24,094

122.6

1100

16.79

4.574

Syria 825 550 66.67

Tunisia 240 33.8 14.08

UAE

US

881

–

185.3

2271

21.03

–

Source: FAO AQUASTAT Database, *Mekala et al. (2008).


By 2020, it is expected that 65% of the irrigation water used in Israeliagriculture will be sewage effluents (Assouline et al., 2002). Other arid andsemi-arid countries, such as Jordan and Tunisia, reclaim the vast majority ofmunicipal wastewater for agricultural irrigation. Wastewater has beenrecycled in agriculture for centuries as a means of disposal in cities such asBerlin, London, Milan, and Paris (AATSE, 2004). In Pakistan, 26% ofnational vegetable production is irrigated with wastewater (Ensink et al.,2004). In Hanoi, 80% of vegetable production is derived from urban andperi-urban areas that receive a secured supply of recycled water (Lai, 2000).In Ghana, irrigation involving diluted wastewater from rivers and streamshas been reported (Keraita and Drechsel, 2004). In Mexico, about 260,000ha are irrigated with wastewater (Mexico CAN, 2004). Agriculture, beingthe largest user of recycled water in Australia, accounts for approximately66% (�280 GL) of all recycled water used (ABS, 2006).

In many countries, municipal wastewater is not collected and treated butdischarged directly into surface water bodies or used in agriculture withoutappropriate consent. In most developing countries, 90% of all wastewater is


discharged untreated into local waterways (Johnston, 2003). In the rest ofthe world, most of the wastewater is collected and treated to remove solids,pathogens, oils, and other contaminants. Two main sets of regulations existfor wastewater treatment and reuse: the California Health Laws (Title 22,State of California, 2001) and the World Health Organization Guidelines(WHO, 1989). The WHO guidelines are frequently used in developingcountries. Permissible water quality criteria stipulated are less restrictive thanthose described in the California Health Laws, which recommend wastestabilization pond systems as the preferred treatment method as opposed to aconventional energy intensive treatment system (Crook, 1991).

Wastewater treatment can be grouped into three main processes:(i) primary treatment, which includes physical processes such as grit removaland settling out of coarse material to the bottom of the tank as primarysludge. In some treatment plants, flocculants such as aluminum sulfate(i.e., alum) are added; (ii) secondary treatment, which aims to removesoluble and colloidal biodegradable organic matter (OM) and suspendedsolids. Secondary treatment generally consists of an aerobic biological pro-cess whereby microorganisms oxidize OM in the wastewater; (iii) tertiarytreatment or advanced treatment technologies, these referring to any physi-cal, chemical, or biological treatment process used to accomplish a degree oftreatment greater than that achieved by secondary treatment, such as ozoni-zation, rapid gravity filtration, and ultraviolet radiation. The conditions inwhich wastewater is stored following treatment may further influence itschemical composition (Droste, 1997; Yu et al., 1997).

Specific composition of a waste stream is dependent on its origin and thedegree of treatment it receives. Heterogeneity in influent waste streams mayinclude domestic, industrial sources (paper and printing manufacturing,timber processing plants, leather, and textile industries) and agriculturalsources (dairy, poultry, meat, and vegetable processing operations). Waste-water quality defines certain biological, chemical, and physical character-istics that influence its suitability for a specific use (Ayers andWestcot, 1985;WHO, 2006). Nutrient loading (N, P, K, and S), organic loading, dissolvedconstituents, such as dissolved salts and solids, types and concentrations ofmicroorganisms, and heavy metal(loid)s, trace organic compounds includ-ing pharmaceuticals, and pH are all quality criteria. Wastewater characteri-zation is further complicated by daily and seasonal variation. There is atwofold risk associated with applying wastewaters to agricultural cropswith respect to anthropogenic contaminants including metal(loid)s. First,wastewater-borne metal(loid)s can be assimilated by plants and subsequentlyenter the food chain. Second, application of wastewater can also impact onheavy metal(loid)s that have been applied to soil and crops prior to thewastewater irrigation. There is, however, limited information in the literatureon both issues. In Tables 2 and 3, important wastewater types for agriculturalirrigation and their main organic and inorganic components are summarized.

Table 2 Heavy metal(loid) concentrations in various wastewater and waste sludge sources

Metal

(loid)s

Treated

sewage

Storm

water

Dairy

effluent

Piggery

effluent

Pulp and

paper

secondary

sludge

Feedlot

manure

Dairy

cattle

slurry

Beef cattle

slurry

Poultry

litter

Broiler

litter

Swine

slurry

Deep-pit

poultry

litter

Threshold

values

LTVa

(mgL�1) (mgkg�1) (mgL�1)

Cr 0.035 �0.012

0.04 - – 20 – 5.64 4.69 – 9.9 2.82 6 0.1

Cd 0.002 �0.0003

0.04 – – 4.5 – 0.33 0.26 3 4.93 0.3 2 0.01

Pb 0.003 �0.002

0.073–

1.78

– – 42 – 5.87 7.07 11 - 2.48 13 2

Ni 0.011 �0.002

0.053 – – 35 – 5.4 6.4 15 2.46 10.4 14 0.2

Cu 0.002 �0.001

0.022–

7.033

0.5–10.5 0.26 206 16.5 62.3 33.2 748 6.1 351 19 0.2

Zn 0.059 �0.021

0.056–

0.929

– 0.58 513 6480 209 133 718 743 575 252 2

As – 0.058 – – 0.17 1.44 2.6 43 34.6 1.68 – 0.1

Hg – 3.22 – – 0.3 – – – – – – – 0.002

Reference Antanaitis and

Antanaitis

(2004)

Barrett

et al.

(1993)

Bolan et al.

(2003a)

Lowe

(1993)

Hart and

Speir

(1992),

Carnus

(1994)

Wallingford

et al.

(1975)

Nicholson

et al.

(1999)

Nicholson

et al.

(1999)

Moore

et al.

(1998)

Jackson

and

Miller

(2000)

Nicholson

et al.

(1999)

Bomke

and

Lowe

(1991)

Nicholson

et al.

(1999)

a LTV (long-term trigger values) in irrigation water (long-term use—up to 100years) (ANZECC and ARMCANZ, 2000).

Table 3 Composition of wastewaters or sludges from selected sources

Parameter

Milk

factory

wastewater

Meat

processing

secondary

effluent

Raw

meat

effluent

Tannery

secondary

effluent

Dairy

effluent

Piggery

effluent

Textile

effluent

Untreated

wastewater

Preliminary-

treated

wastewater

Primary-

treated

wastewater

Pulp and

paper

secondary

sludges Biosolids

TDS – – – – – – 1480 1152 844.8 780 – –

Suspended

solids

–– 20–100 1155 120 – – 471 132 121 105 – 65

BOD5 1700 20–100 646 30 – – 645 – – – – –

COD – 80–400 1544 410 – – 2430 – – – – –

Total N 70 40–200 – 130 190 1300 – 1415 1260 1124 32,000 8.8

Total P 35 5–30 – 1.6 30 600 – 6.4 5.10 5.14 8075 2.8

Fat 400 0–30 110 – – 8.3 – – – – – –

Na 560 50–250 – 2700 50 – – 205 193 154 4586 –

K 13 20–150 – – 220 500 – 60 52 42 2905 1.8

Ca 8 3–250 – 340 110 – 1.24 55 51 47 17,000 4.9

Mg 1 3–10 – 36 30 – 1.04 48 – – 2000 1.7

Free

Chlorine

– – – – – – 1.14 – – – – –

Nitrate – – – – – – 7.97 – – – – 420

Phosphate – – – – – – 2.63 – – – – –

Sulfide – – – – – – 0.58 – – – – –

Reference Hart and

Speir

(1992),

Carnus

(1994)

Hart and

Speir

(1992),

Carnus

(1994)

Hart and

Speir

(1992),

Carnus

(1994)

Hart and

Speir

(1992),

Carnus

(1994)

Hart and

Speir

(1992),

Carnus

(1994)

Hart and

Speir

(1992),

Carnus

(1994)

Yusuff and

Sonibare

(2004)

Yusuff and

Sonibare

(2004)

Yusuff and

Sonibare

(2004)

Yusuff and

Sonibare

(2004)

Hart and

Speir

(1992),

Carnus

(1994)

Nash et al.

(2011)

Units are mgL�1 except for pulp and paper sludges and biosolids (mgkg�1).


2.1.1. Municipal wastewaterIn both developed and developing countries, land application of municipalwastewater (both treated and untreated) is a common practice. Municipalwastewater is composed of domestic and industrial wastewater (Hussain et al.,2002; Pettygrove and Asano, 1984). Domestic wastewater consists of dis-charges from households, institutions, and commercial buildings. Wherecountry or state legislation permits, this wastewater is applied to land. How-ever, this depends on the crop it is applied to and the level of treatment.Secondary-treated wastewater typically contains low levels of contaminants asthese tend to settle under gravitation with solid fractions in the treatmentlagoons. Settling of suspended solids also lowers both the chemical andbiochemical oxygen demand. Municipal wastewater also contains high con-centrations of nutrients, especially nitrogen (N) and phosphorus (P), traceelements, such as iron (Fe) and Mn and dissolved salts, particularly sodium(Na), chloride (Cl), and in some cases bicarbonates. These parameters arecritical when wastewater is reused in agriculture.

2.1.2. Farm wastewaterFarm effluents such as those emanating from dairy sheds and piggeries arebeing increasingly employed as sources of irrigation water and nutrients(Bolan et al., 2009; McDonald, 2007). For example, in New Zealand, dairyand piggery effluents generate annually about 9000Mg of N, 1250Mg of Pand 14,000Mg of K (Bolan et al., 2004a). Effluents from farms differ in theircomposition depending on the animal production system from which theyare derived (chicken, pigs, beef, dairy). Generally, farm wastewater is rich inorganic and inorganic components (Tables 2 and 3) (Wang et al., 2004).Copper and Zn are commonly used as feed additives, growth promoters, fordisease prevention or treatment, and their concentration in the final waste-water can be significant (Bolan et al., 2004b; Sims and Wolf, 1994).

In many countries including Australia and New Zealand, farm effluentshave traditionally been treated biologically using two-pond systems andthen discharged to land or stream. Bolan et al. (2009) have suggested that landapplication of farm effluent facilitates the recycling of valuable nutrients,carbon (C), and water, and if managed well, helps to mitigate surface waterpollution. In many instances, this may be the cheapest and most socially/culturally accepted form of final treatment. Application of farm effluents canincrease pasture yield due to the net loading of nutrients and water (Bolanet al., 2004c; Wang et al., 2004). This, however, is influenced by the rate,method and time (season) of application, soil fertility, and climatic conditions(Ball and Field, 1982). According to Bolan et al. (2009) returning dairy andpiggery effluents directly to land has become the most common method oftreatment in most parts of the world.

However, in many regions, the amount of farm effluents generated on aper farm basis exceeds the quantity that can be safely accommodated by the


available agricultural land and repeated annual applications of large amountsof effluent can cause soil nutritional side effects and environmental damage(Balota et al., 2010). For example, Giacomini et al. (2009) observed that halfof the piggery effluent N is lost from soil–plant systems through leachingand volatilization, resulting in environmental pollution. Balota et al. (2010)and McDonald (2006) indicated that safe farm effluent application in agri-culture is necessary to minimize the environmental damage.

2.1.3. Effluents from the agricultural industryRecycling of water from agricultural industry is another common sourceof wastewater. For example, in Australia, agricultural drainage effluent iscollected and reused as a source of irrigation water (Dillon, 2000). Similarly,wastewater from farm animal treatment plants (abattoirs) is increasingly usedas a source of irrigation water (Luo et al., 2004). Wastewater from intensiveagricultural industries (fish processing plants) and rural industry (abattoirs) ischaracterized by high chemical and biological oxygen demand and nutrientsrelative to many other wastewaters (McLaren and Smith, 1996) (Tables 2and 3). Mittal (2004) recently reviewed effluent wastewater from abattoirsand concluded that water quality depended on animal and processingtype and water usage, that is, dilution. One concern associated with landapplication of abattoir waste is the high level of pathogens that have thepotential to contaminate receiving water bodies either directly as pointdischarge or indirectly in runoff. These effluents also contain elevated levelsof grease, blood, and organic chemicals added during processing and clean-ing operations (Kretzschmar, 1990). In many countries including Australiaand New Zealand, abattoir effluent is usually disposed of to land due to highcosts associated with independent treatment systems and environmentalconcerns over surface water discharge (Quinn and Fabiansson, 2001).Afonso and Borquez (2003) reported that the wastewaters generated duringfish meal production contain a high organic load, but unlike other industrialeffluents they do not contain any known toxic or carcinogenic materials.

In some regions, winery wastewater is also a viable water source that isincreasingly being recycled by grape growers and pastoralists for irrigation(Stevens, 2009). Reuse is driven primarily through the obligations of thewinery to dispose of their wastewater, preferably in a sustainable and cost-effective manner. Generally, winery wastewater is treated on-site throughsystems of only small flow volume capacity and the chemical composition ofthis wastewater source varies considerably between wineries and mayrequire a management approach that is site specific when irrigated to land(Kumar and Christens, 2009). In general, winery wastewater contains highsalt concentrations thereby accounting for a considerably greater salt loadingrelative to irrigating with river, ground, or town supply water. Specific ions,in particular Na and potassium (K), originating from cleaning products,grape lees, and waste juice may also have confounding effect on soils beyond


that imposed by salinity alone. A high concentration of either Na or K inirrigated waters is undesirable and when continually applied to soils candisplace more desirable cations [i.e., calcium (Ca) and magnesium (Mg)]from the soil exchange complex (Pils et al., 2007). This in turn raises thepotential for adverse changes to soil structure (Jayawardane et al., 2011). Themanagement of salts is imperative so water conservation benefits are notcompromised by a decline in soil and plant health and off-site pollution(McCarthy, 1981; Neilsen et al., 1989).

2.1.4. Effluents from pulp and paper millsPulp and paper mill effluent either from thermomechanical pulp mill orchemi-thermomechanical pulp mill is often irrigated to land after primarytreatment (Smith et al., 2003; Wang et al., 1999). Pulp mill effluent has highchemical and biochemical oxygen demand and some wood derived organiccompounds, metal(loid)s, fatty and resin acids, and relatively high C:Nratios (Tables 2 and 3). Kookana and Rogers (1995) reviewed the effect ofpulp mill effluents on soil properties. An extensive investigation into thedifferent organic chemicals in pulp mill effluents and their behavior in soilcan be found in this excellent review paper. Effluent from pulp mills is a richsource of OM, N, P, Ca, Mg, and trace elements (Kannan and Oblisami,1990), and consequently the application of pulp mill effluents on land isbecoming a common method for recycling nutrients (Rubilar et al., 2008).The presence of chlorinated organic compounds, most notably chlorinesubstituted phenolic compounds, chlorinated lignins, dioxins, chloroben-zenes, and non-chlorinated organic compounds (Deriziotis, 2004; Gergovet al., 1988), has raised concerns about land application (Kannan andOblisami,1990; Lavric et al., 2004).

2.1.5. StormwaterUrban stormwater harvesting has emerged in recent years as a viable option toreduce pressures on existing water sources and to alleviate adverse environ-mental impacts associated with stormwater runoff (Roy et al., 2008a). This is arelatively abundant, local source of water, available throughout most urbanareas. In Australia, for instance, approximately 10,300 million liters of storm-water are generated annually (Laurenson et al., 2010). In many cases, urbanstormwater runoff contains a broad range of pollutants that are transported tonatural water systems (Aryal et al., 2010). Stormwater pollutants originatefrom many sources and activities and can occur as either particulate ordissolved forms. Many toxic chemicals, such as pesticides and herbicides arefound in stormwater, along with oil, grease, and heavy metal(loid)s such asCd, Cr, Cu, Ni, Pb, and Zn (Wong et al., 2000). Nutrients such as N and Pare also important pollutants in stormwater.

The harvesting of stormwater from industrial zones prior to its entry intonatural waterways is likely to reduce the subsequent impact of point source


discharge on surface waters by reducing pollutant loads (Davis et al., 2009;Henderson et al., 2007). If suitably designed, a stormwater harvesting systemwill also provide urban stream health benefits by mitigating frequent flows tostreams and serve as a public amenity (Bratieres et al., 2008; Hatt et al., 2009).Stormwater harvesting and storage can be achieved in a number of waysincluding biofiltration, porous pavement, rain garden, and groundwaterrecharge (Davis et al., 2009; Department of Planning and Local Government,2009; Henderson et al., 2007; Kim et al., 2003; Laurenson et al., 2011).

2.2. Heavy metal(loid) sources

Heavy metal(loid)s reach the soil environment through both pedogenic(geogenic) and anthropogenic processes. Most heavy metal(loid)s occur nat-urally in soil parent materials, chiefly in forms that are not readily available forplant uptake (Adriano, 2001; Alloway, 2004; Bolan et al., 2008). Due to theirlow solubility, the heavy metal(loid)s present in the parent materials are oftennot bioavailable and have a minimum impact on soil organisms. Apart fromSe (Dhillon and Dhillon, 1990; Doblin et al., 2006) and As (Chakrabortyand Saha, 1987; Mahimairaja et al., 2005; Mukherjee et al., 2008; Naidu andSkinner, 1999; Naidu et al., 2008), other heavy metal(loid)s (e.g., Cr, nickel(Ni), Pb) derived via geogenic processes have limited impact on the soilecosystem. Unlike pedogenic inputs, heavy metal(loid)s added throughanthropogenic activities typically have high bioavailability (Adriano et al.,2004; Lamb et al., 2009; Naidu et al., 1996). Anthropogenic activities primar-ily associated with industrial processes, manufacturing, and the disposal ofdomestic and industrial waste materials including wastewater are the majorsource of metal(loid) enrichment in soils (Adriano, 2001). Fertilizer, manure,effluents, and organic amendments addition to agricultural soils are consid-ered to be the major sources of most minor elements including heavy metal(loid)s that are essential for plant growth (Bolan et al., 2004b; Loganathanet al., 2008; Park et al., 2011).

2.2.1. Fertilizer productsOf the heavy metal(loid)s present in fertilizers, the presence of elevatedconcentrations of Cd is of greatest concern as it is highly toxic to humansand can accumulate in soils, plants and animals (Alloway 1990; USPHS 2000).Phosphate fertilizers are considered to be the major source of heavy metal(loid) input, especially Cd, in agricultural and pasture soils in Australia andNew Zealand due to the extensive use of high Cd-containing phosphatefertilizers (Loganathan et al., 2003; McLaughlin et al., 1996; Naidu et al.,1997). Increased concentrations of Cd in fertilizers that are applied to landhave been reported to result in increased Cd concentrations in grain crops(Bolan et al., 2011b; Grant et al., 2010; He and Singh 1994; McLaughlin et al.,1996). Results of studies in several countries have shown that some heavy


metal(loid)s in P fertilizers may be available to plants (Huang et al., 2003,2005; Mortvedt, 1996).

There have been increasing efforts to reduce the accumulation of Cd insoils by using low Cd-containing P fertilizers. This is achieved by eitherselective use of phosphate rocks (PRs) with low Cd or treating the PRsduring processing to remove Cd. Superphosphate fertilizer manufacturers inmany countries are introducing voluntary controls on the Cd content of Pfertilizers. A number of PRs with low Cd contents are available which canbe used for the manufacture of P fertilizers, but sources with higher Cdcontents continue to be used in many countries for practical and economicreasons (Bolan and Duraisamy, 2003; Loganathan et al., 1995, 2003).

2.2.2. BiosolidsOrganic amendments such as biosolids (e.g., Cd) and poultry manure (e.g.,As) have been regarded as a major source of heavy metal(loid) accumulationin soils, and a large volume of work has been carried out to examine themobilization and bioavailability of heavy metal(loid)s derived from thesesources (Bolan et al., 2004b; Haynes et al., 2009; McBride, 1995). The mostcommonly detectable heavy metal(loid)s in biosolids, Pb, Ni, Cd, Cr, Cu,and Zn originate primarily from the contamination of these wastes withindustrial wastewater (Haynes et al., 2009). Heavy metal(loid) concentrationsare governed by the nature and the intensity of the industrial activity, as wellas the type of process employed during the biosolid treatment (Mattigod andPage, 1983; Oviasogie and Ndiokwere, 2008; Wang et al., 2003a).

Gove et al. (2001) reported that biosolid application (250kgN ha�1

yr�1) to a sandy or sandy loam soil resulted in loadings of approximately6, 2, 5, and 0.2mgkg�1 Zn, Cu, Pb, and Ni, respectively. Illera et al. (2000)demonstrated that biosolid application to soil had little effect on the totalconcentration of Ni and Cr but resulted in Cd, Cu, Pb, and Zn increasingconsiderably as a consequence of the high content of these metal(loid)s inbiosolids. It is known that these heavy metal(loid)s are typically immobilizedin soils, but they can be toxic to soil micro flora and can accumulate in plantsand grazing animals (Haynes et al., 2009). Kao et al. (2006) reported thatthe addition of biosolid accumulated Cu, Pb, and Zn but reduced microbialbiomass, indicating that microbial activities were disrupted by the heavymetal(loid)s.

2.2.3. ManureManures from intensive animal industries are a major source of organicamendments for agricultural land. In Australia, beef and dairy cattle aloneproduce approximately 4 million Mg of manure every year (Bunemann et al.,2006). Similarly, in USA, of about 0.9 billion Mg organic and inorganicagricultural recyclable by-products generated, approximately 45.4 millionMgare dairy and beef cattle manure, and 27 million Mg are poultry and swine


manure (Camberato et al., 1997; Walker et al., 1997). As heavy metal(loid)sare increasingly employed as a feed additive in intensive animal productionsystems,manure application is likely to be an important source of certainmetal(loid) input to soils (Bolan et al., 2004b; Moscuzza and Fernandez-Cirelli,2009). Heavy metal(loid)s in manure by-products are also derived fromingestion of contaminated soil by the animal, and during manure collectionand handling. Heavy metal(loid)s vary considerably between manure types,animal categories, and farms (Menzi et al., 1993).

Heavy metal(loid) content in manures derived from intensive animalproduction systems is related to feed mineral content and the animal conver-sion efficiency (Nicholson et al., 1999). Increases in metal(loid) concentrationin animal feed have often resulted in corresponding increases in their con-centration in manure by-products (Mohanna et al., 1999; Moscuzza andFernandez-Cirelli, 2009; Nahm, 2002). A number of heavy metal(loid)s areadded to livestock and poultry feedstuff not only as essential nutrients but alsoas supplement to improve health and feed efficiency. Diets of poultry andlivestock include heavy metal(loid)s (As, Co, Cu, Fe, Mn, Se, Zn) to preventdiseases, improve weight gains, and increase egg production (Mondal et al.,2007; Tufft and Nockels, 1991). Not all of the heavy metal(loid)s consumedby animals are absorbed by their digestive tracts; consequently, the manure isoften metal(loid) enriched (Sistani andNovak, 2006). For example, swine canexcrete approximately 80–95% of the total daily Cu and Zn intake (Brumm,1998; Moral et al., 2008). Adding As to feed as an additive to control coccidi-osis in poultry has resulted in an increase in As level in poultry litter (Churchet al., 2010; Garbarino et al., 2003; Morrison, 1969; Sims and Wolf, 1994).

Regular application of manures and slurries has often been shown toresult in the accumulation of heavy metal(loid)s. Brink et al. (2003) reportedthat application of swine effluent (annual mean of 10hacm) to Bermudagrass pasture added annual averages of 0.6, 2.2, 0.3, and 0.86kgha�1 Cu, Fe,Mn, and Zn, respectively. Evers (2002) also reported application of 9Mgha�1 broiler litter added averages of 5.85, 5.0, 9.4, and 6.55kgha�1 Zn, Feand Cu, respectively. In another study, Jinadasa et al. (1997) reported thathigh Cd levels in soils and vegetables throughout Sydney, Australia, weredue to repeated applications of poultry manures. In New Zealand, landapplication of dairy pond effluent, based on a N loading of 150kgN ha�1, islikely to add a maximum of 31.5 and 73.7kg Cu ha�1 through effluent andmanure sludge application, respectively (Bolan et al., 2003a).

2.2.4. WastewaterWastewaters act both as a source and sink for heavy metal(loid)s in soils.Depending on the source and level of treatment, wastewater may contain arange of heavy metal(loid)s (Table 2) and continuous application is likely toresult in these heavy metal(loid)s accumulating in soils. Heavy metal(loid)sare usually removed during common treatment processes and most of them


end up in the biosolid fraction of the treatment process with very low metal(loid) concentrations present in the treated effluents (Kulbat et al., 2003;Sheikh et al., 1987; Ziolko et al., 2011). Generally therefore in highlytreated wastewater, the concentration of heavy metal(loid)s is low andconsidered safe when used for irrigation and other recreational purposes.Consequently, heavy metal(loid)s are of little concern for irrigation of cropswhen using treated effluents as a source of wastewater. Irrigation withuntreated or partially treated wastewaters has the potential to cause heavymetal(loid)s to accumulate in the soils and become bioavailable for crops(Cui et al., 2004; Qadir et al., 2000). If the wastewater is derived from anindustrial source or is less treated, then the effect of heavy metal(loid)swould need to be inspected (Toze, 2006).

Wastewater also contains a range of components including dissolvedand particulate OM, soluble organic and inorganic anions which can inter-act with heavy metal(loid)s, thereby altering their mobility and subsequentbioavailability. Wastewater often contains high levels of nutrients, whichcan be beneficial to crop production (Liu et al., 2005) (Table 3). Bolan et al.(2004c) have indicated that the application of dairy farm effluent irrigationmay provide an attractive means of increasing pasture growth throughincreased nutrient loading and a cost-effective amelioration technique,provided that the associated environmental risks of contamination of soilsand groundwater by heavy metal(loid)s are minimized.

Direct irrigation of untreated sewage effluents is a common practiceespecially in Asian countries and many studies have reported the risksassociated with this system. Heavy metal(loid)s are easily accumulatedin the edible parts of leafy vegetables, as compared to grain or fruit crops(Mapanda et al., 2005). Arora et al. (2008) assessed the levels of differentheavy metal(loid)s like Fe, Mn, Cu, and Zn, in vegetables irrigated withwater from different sources. The results indicated a substantial build-up ofheavy metal(loid)s in vegetables and the range was 116–378, 12–69, 5.2–16.8, and 22–46mgkg�1 for Fe, Mn, Cu, and Zn, respectively. Theyreported that the highest mean levels of Fe and Mn were detected in mintand spinach, whereas the levels of Cu and Zn were highest in carrot.In another study, Latif et al. (2008) examined the heavy metal(loid) con-tamination of different water sources, soils, and vegetables and reported thatthe concentrations of heavy metal(loid)s in sewage and industrial effluents-irrigated vegetables were above critical levels. Similarly, Singh et al. (2010a)quantified the concentrations of heavy metal(loid)s (Cd, Cr, Cu, Ni, Pb,and Zn) in soil, vegetables, and the wastewater used for irrigation. Theirstudy demonstrated that the wastewater used for irrigation had high con-centrations of Zn followed by Pb, Cr, Ni, Cu, and Cd and its continuousapplication for more than 20years has led to accumulation of heavy metal(loid)s in the soil. They observed that concentrations of Cd, Pb, and Nihave crossed the safe limits for human consumption in all the vegetables.


They also noticed that the percentage contribution of fruit and vegetablesto daily human intake for Cu, Ni, Pb, and Cr was higher than that of leafyvegetables, while the reverse was true for Cd and Zn.

3. Effects of Wastewater Irrigation on Soil

Properties Affecting Heavy Metal(Loid)

Interactions

Wastewater irrigation affects metal(loid) dynamics by directly influen-cing their reactions in soils and indirectly by altering soil propertiescontrolling their fate. Research shows that wastewater irrigation can resultin significant changes to soil physical, chemical, and biological properties(Muller et al., 2007). In the following section, an overview on the impacts ofwastewater irrigation on soil properties relevant to the fate of metal(loid)s inthe soil environment is provided. The effect of wastewater irrigation onmetal(loid) reactions in soils will be discussed in Section 4.

3.1. Soil chemistry

Soil chemistry plays an important role in the successful utilization of wastewa-ter as a source of irrigation. On the one hand, the fertilization effect ofwastewater and the resultant independence from costly fertilizers has enhancedthe development of wastewater irrigation systems in many countries. On theother hand, induced salinity and sodicity often limit its viability. Conse-quently, the impact of wastewater irrigation on soil chemical properties iscomparatively well investigated for a variety of soil types, climatic conditions,and crops (Tables 4 and 5). A selection of soil chemical properties affectedby wastewater irrigation and relevant for the transformation, transport, andcomplexation of metal(loid)s, is described, viz. the soil pH, soil organic matter(SOM), cation exchange capacity (CEC), salinity, and sodicity.

3.1.1. Soil pHIn the majority of studies, the soil pH significantly increased after long-termirrigation with wastewater from different sources (Friedel et al., 2000; Hassanliet al., 2008; Magesan et al., 1999; Qishlaqi et al., 2008; Roy et al., 2008b;Schipper et al., 1996; Walker and Lin, 2008). In some studies, however, soilpH was unaffected by long-term wastewater irrigation (Friedel et al., 2000;Gwenzi and Munondo, 2008; Magesan et al., 1999), while others reporteddecreased soil pH (Rattan et al., 2005; Rosabal et al., 2007; Xu et al., 2010).For example, vinasse irrigation (pH 5.02) for 40years significantly decreasedthe soil pH from 7.1 to 6.7 and from 6.2 to 5.9 in sampling depths of 0.1 and1m, respectively (Rosabal et al., 2007). Angin et al. (2005) explained the

Table 4 Selected references on the effect of wastewater on soil properties

Site Wastewater (WW) Experiment Soil properties

Suggested

mechanism References

Location Soil type Land-use Type (years)

(mm d�1)

TSS

(gm�3)

SAR pH Depth

(m)

Approach Bulk density

(gcm�3)

Total

porosity

(%)

Ksat

(mm h�1)

Taupo, New

Zealand

Silt loam Pasture WW

12

4.72

28 n.d. 7.2 0–0.05 Disc

infiltrometer

0.75* (0.84) 71 (68) 16 (8) Different land

management of

control and

treated area

Vogeler

(2009)

Levin, New

Zealand

Sand Pasture WW

22

9

185 n.d. 6.2 0–0.05 Disc

infiltrometer

1.1* (1.2) 58 (54) 34* (13)* Vogeler

(2009)

Levin, New

Zealand

Sand Pinus radiata Secondary-

treated

WW

7

8

185 n.d. 6.2 0–0.1 Repacked

column

1.1 (1.2) n.d. 35 (39) n.a. Magesan et al.

(1999)

Rotorua, New

Zealand

Sand Pinus radiata Tertiary-

treated

WW

5

8

6 n.d. 7.2 0–0.1 Repacked

column

0.6 (0.6) n.d. 21 (23) n.a. Magesan et al.

(1999)

Amman, Jordan Clay Barley WW

0yr

2yr

5yr

15yr

n.s.

111 5.4 7.6 0.1–0.2 Repacked

column

1.3

1.3

1.3

1.3

n.d. 8*

7*

6*

3*

Retention of

DOC, clay

dispersion, and

change in pore

size distribution

Gharaibeh

et al.

(2007)

Mizra, Israel Clay Orchard WW (drip

irrigation)

23

15

170 5.1 7.6 0–0.2 Intact core 1.1

(0.9)

n.d. 0.02

(0.06)

Physical blocking of

pores, clay

swelling

Bhardwaj et al.

(2007)

Rotorua, New

Zealand

Sandy

loam

Pinus radiata WW

2.8

n.s.

25 1.5 7.5 2.3 Monolith n.d. (0.74–1) n.d. 6* (30)a Biological clogging

of soil pores

Cook et al.

(1994)

Levin, New

Zealand

Sand Pinus radiata Primary-

treated

WW

4

8

185 n.d. 6.2 0–0.1 Intact cores 0.8* (1.2) 70* (54) 185 (159) Increased

macroporosity

due to

stimulation of

microbial

communities

Magesan

(2001)

Rotorua, New

Zealand

Sandy

loam

Pinus radiata Tertiary-

treated

WW

4

8

6 n.d. 7.2 0–0.1 Intact cores 0.7 (0.6) 72 (72) 114 (39) Increased

macroporosity

due to

stimulation of

microbial

communities

Magesan

(2001)

Madurai

Corporation,

India

Sandy

loam

Farm WW

10

n.s.

712 2.0 7.3 0–0.1b Intact cores 1.2 (1.3) 38 (35) 135 (91) Input of OM

improved

structure

Mathan

(1994)

Madurai

Corporation,

India

Sandy

loam

Farm WW

15

n.s.

712 2.0 7.3 0–0.1b Intact cores 1.1 (1.3) 55 (35) 163 (91) Input of OM

improved

structure

Mathan

(1994)

Pennsylvania

State

University,

USA

Diverse:

silt

loam–

silty

clay

loam

Farm WW

44

7

5 n.d. 7 0–0.2b Landscape

approach

1.5* n.d. 38* (116) Excessive water,

impact of water

drops and

machinery, soil

transport across

landscape

Walker and

Lin (2008)

Isfahan, Iran Aridisol,

silty

clay

Sugar beet,

corn,

sunflower

Secondary-

treated

WW

n.d. n.d. 7.8 0–0.15

0.15–

0.3

Double ring

infiltrometer

1.3* (1.2)

1.3* (1.2)

49* (54)

49* (53)

35 (38)

35 (37)

Biological and

physical

clogging

Abedi-Koupai

et al.

(2006)

a Ponded infiltration rate.b Measured in increments for more than 1m depth in original paper. Values within brackets are from the control site (before irrigation of wastewater).

Table 5 Selected references on the effect of wastewater on soil properties (pH, SOC, and C-input)

Location Soil type

Wastewater Type

(years) Land-use Depth (m) pH SOC (%)

C-input

(tha�1yr�1) References

Shiraz, Iran – Untreated domestic

WW, (20yr

estimated)

Wheat 0–0.2 8.4 (7.3) 15.9 (0.4) n.d. Qishlaqi et al.

(2008)

Marvdasht,

Iran

Silty clay

Silty loam

Silty loam

Treated municipal

effluent, 25months

(39 ML ha�1yr�1)

Trees 0–0.3

0.3–0.6

0.6–0.9

8.8

(8.0)

8.8

(8.2)

8.8

(8.1)

0.0003

(<dL)

0.0006

(<dL)

0.0006

(<dL)

n.d. Hassanli et al.

(2008)

Mezquital

Valley,

Mexico

Silt loam,

Mollic

Leptosol

WW (11 ML ha�1yr�1)

0

25

65

80

Maize 0–0.15 7.5

7.9

7.6

7.7.

1.8

2.2

1.9

2.3

1.5 Friedel et al. (2000)

Mezquital

Valley,

Mexico

Clay loam,

Eutric

Vertisol

WW (11 ML ha�1yr�1)

0

25

65

80

Maize 0–0.15 7.4

7.9

8.1

7.6.

1.1

1.7

2.2

2.7

1.5 Friedel et al. (2000)

Amman,

Jordan

Clay silt,

Vertisol

Municipal WW

0yr

2yr

5yr

15yr

Barley 0–0.2

0.2–0.4

0–0.2

0.2–0.4

0–0.2

0.2–0.4

0–0.2

0.2–0.4

8

8.1

7.7

7.9

8.1

8.2

7.9

8.2

SOM

0.7

0.5

1.1

0.7

1.0

0.7

1.3

0.7

Gharaibeh et al.

(2007)

Rotorua,

New

Zealand

Sand; Typic

Udivitrand

WW; 5yr

(29 ML ha�1 yr�1)

Forest

plantation

0–0.1

0.1–0.2

6.9 (5.9)

6.7

(6.0)

7.4 (7.1)

4.5 (4.5)

2.2 Magesan et al.

(1999)

Levin, New

Zealand

Sand,

Psamment

WW, 7yr

(31 ML ha�1 yr�1)

Forest

plantation

0–0.1

0.1–0.2

n.d.

n.d.

2.7 (2.4)

0.9

(0.4)

2.3 Magesan et al.

(1999)

LaHabana,

Cuba

Ultisol Vinasse, 40yr; n.d. Sugarcane 0–0.1

0.1–0.2

0.2–0.3

0.3–0.4

0.4–0.5

0.5–0.6

6.7 (7.1)

7.0

(7.5)

6.9

(7.4)

6.8

(7.4)

6.9

(7.5)

5.9

(6.2)

2.9

(2.0)

2.7

(1.6)

2.3

(0.9)

1.2

(0.5)

0.4

(0.2)

0.2

(0.3)

n.d. Rosabal et al. (2007)

Erzurum,

turkey

n.s. Raw WW; long-term Cabbage,

potato

0–0.3

0.3–0.6

0.6–0.9

Decrease þ164%

þ109%

þ118%

Angin et al. (2005)

Berlin,

Germany

Haplic Luvisol Primary treated WW;

100yr

(20 ML ha�1 yr�1)

Pasture Topsoil 5.8

(5.6)

2.2

(0.6)

2.3 Filip et al. (2000)

Delhi, India Loamy sand;

sandy loam

Sewage effluents

5, 10, 20yr

Rice, grain,

vegetables

0–0.15 7.5 (7.9) 0.65 (0.39) n.d. Rattan et al. (2005)

Ramat-

Hacovech,

Israel

Sand Effluents, 10yr Citrus 0–0.25 7.5 (7.2) 1.3 (0.6)

OM

n.d. Lado et al. (2005)

Mizra, Israel Clay Effluents, 12yr Citrus 0–0.25 7.8 (7.5) 3.9 (3.4)

OM

n.d. Lado et al. (2005)

Harare,

Zimbabwe

Loamy sand

Gleyic Lixisol

Pasture n.d.

(Continued)

Table 5 (Continued)

Location Soil type

Wastewater Type

(years) Land-use Depth (m) pH SOC (%)

C-input

(tha�1yr�1) References

Treated WW

26yr

(15–37 ML ha�1 yr�1)

0–0.3

0.3–0.6

0.6–0.9

5.0 (3.9)

5.0 (3.9)

5.4 (3.8)

17.2 (7.1)

4.3 (3.8)

3.3 (2.9)

Gwenzi and

Munondo

(2008)

Isfahan, Iran Silty clay Municipal WW

(150, 300, 600t ha�1

yr�1)

0–0.3

150t ha�1

300t ha�1

600t ha�1

0.3–0.6

150t ha�1

300t ha�1

600t ha�1

(7.6)

7.4

7.3

7.2

(6.9)

6.6

6.5

6.3

(0.6)

0.9

0.9

1.1

(0.6)

0.7

0.8

0.8

58

116

173

Khoshgoftarmanesh

and Kalbasi

(2002)

Ebro Valley,

Spain

Xeric

Petrocalcid

Typic

Xerofluvent

Vegetable canning WW;

4yr

Annual

rotation

0–0.3 n.d.

n.d.

1.2

(1.2)

0.8

(1.0)

0.06

0.26

Virto et al. (2006)

Values within brackets are from the control site (before irrigation of wastewater).


lower soil pH of a long-term wastewater-irrigated soil by the increasedmineralization of OM, while Xu et al. (2010) ascribed the effect to the appliedwastewater’s acidity. In general, the effect of wastewater irrigation on soilpH depends on the pH of the wastewater source and the pH bufferingcapacity of soil.

pH is an important factor that controls the accumulation, mobility, andbioavailability of heavy metal(loid)s in wastewater-irrigated soils. The pH isoften reported to show good correlation with soil adsorption of heavy metal(loid)s (Naidu et al., 1997; Tyler and McBride, 1982). Qishlaqi et al. (2008)examined the negative impacts of untreated wastewater irrigation on soilsand crops and reported that pH increased by 2–3 units and heavy metal(loid)s (notably Pb and Ni) accumulated in topsoil above maximum per-missible limits. Although they found a positive relationship (P<0.01)between pH and total contents of Pb and Ni in soils, they reported only1.3–7.7% of Ni and 0.07–1.69% of Pb was phytoavailable. Roy et al.(2008b) reported that even though a higher pH in soils irrigated withpaper mill wastewater was observed, the data did not show a significantpositive correlation with the metal(loid) ions.

3.1.2. Soil organic matterDepending on the level of treatment, wastewater comprises about 0.1%suspended and dissolved organic and inorganic compounds (Feigin et al.,1991; Lado et al., 2005). Generally, therefore wastewater irrigation addsmore OM to a soil than freshwater irrigation or rain. Hence, various studiesconducted in long-term farm dairy/sewage effluent irrigated areas reportedsignificantly increased SOM contents in the topsoil (Barkle et al., 2000;Bhandral et al., 2007; Filip et al., 2000; Friedel et al., 2000; Gwenzi andMunondo, 2008; Marecos do Monte, 1998; Qishlaqi et al., 2008; Rattanet al., 2005; Siebe and Fischer, 1996; Walker and Lin, 2008; Xu et al., 2010)and in the sub-soil (Walker and Lin, 2008). Such SOM-increases have alsobeen attributed to indirect positive effect on biomass production throughthe nutritional benefit of wastewater irrigation leading to more residues inthe soil (Ramırez-Fuentes et al., 2002). However, our attempts to correlatethe C inputs associated with wastewater irrigation with the observed netC-increases failed because of: (i) the frequent lack of information on theC-input through wastewater application; and (ii) the difference in soil types,irrigation periods, effluent characteristics, and climate of the studies consid-ered (Tables 4 and 5).

In a short-term laboratory experiment of 40days, Travis et al. (2010)observed no change on SOM contents in three soil types irrigated withgraywater. Falkiner and Smith (1997) even observed that total C in the top0.1m of a sandy loam under forest plantation after 4years of irrigation withsecondary-treated municipal wastewater was significantly reduced. ThisSOM loss was explained by accelerated decomposition rates caused by the


frequent wetting and drying cycles at the site. In three Luvisols and threeVertisols, long-term municipal wastewater irrigation led to SOM losses in1m depth ranging between 0.6 and 4.9Mgha�1 compared with freshwater-irrigated soils, while the impacts in the topsoils were inconsistent (Jueschkeet al., 2008). Jueschke et al. (2008) postulated that the observed lossesresulted from increased microbial activity stimulated through the applica-tion of easily degradable and complex organic substances with wastewater.In another study, the SOM content in 0–0.75m depth of a dairy factoryeffluent-irrigated allophanic soil was the same as in the non-irrigated controlsite after 22years, but a redistribution of SOM from the top 0.1m down to0.50m depth was observed. This was partly attributed to leaching and themodified earthworm fauna, dominated by the earthworm Aporrectodea longa,a species that forms permanent burrows to lower depths (Degens et al.,2000). In a more detailed study, Herre et al. (2004) evaluated the impact oflong-term wastewater irrigation (90years) on the quality of SOM in two soiltypes, Leptosols and Vertisols in the Mezquital valley in Mexico. Theyfound that the quality of SOM (i.e., differences in carbon mineralization)had changed and carbon mineralization in the irrigated soils significantlyincreased. Consequently, the DOM concentrations in the irrigated soils alsoincreased. This effect was more pronounced in the Leptosols than theVertisols, suggesting the importance of clay (Leptosols: 26–35%, Vertisols:39–56%) in stabilizing SOM (Friedel et al., 2000; Herre et al., 2004).

Increased DOM concentrations in the soil solution of wastewater-irrigatedsites have often been observed and explained by the direct input of wastewater-borne DOM and the indirect solubilization of SOM resulting from increasedpH (Amiel et al., 1990; Fine et al., 2002; Menneer et al., 2001). Bhandral et al.(2007) noticed an increase in DOM concentration soon after effluent irrigationto a pasture soil, which varied with the type of effluent (Fig. 2). In anotherstudy, the DOM concentrations in wastewater-irrigated soils not onlyincreased significantly but the aromaticity of the DOM in soil solutionsdecreased at the same time (Jueschke et al., 2008). Increased DOM con-centrations in soil solutions may affect soil physical properties, such as soilaggregate stability and water binding potential (Frenkel et al., 1992). It alsoprovides organic substrate for soil microorganisms and mobile sorbents tothe system.

OM content is also one of the most important factors that control theaccumulation, mobility, and bioavailability of heavy metal(loid)s in waste-water-irrigated soils. Increase in SOM content can lead to increased soiladsorption capacity by which accumulation of heavy metal(loid)s will beenhanced. Qishlaqi and Moore (2007) carried out statistical analysis of thesources and accumulation of heavy metal(loid)s in agricultural soils andnoticed that SOM was the most important factor controlling the distribu-tion of heavy metal(loid)s. It was revealed that soil samples with high SOMcontent accumulated significantly higher concentrations of heavy metal

50

75

100

125

150

175

0 10 20 30 40 50 60

Days after treatment application

Con

cent

ratio

n (m

g kg

–1 s

oil)

TFDE UFDE

TPFE TME

Water Control

Figure 2 Soil DOM concentration at 10cm depth following the winter application ofwater and a range of effluents types to sheep grazed pasture (TFDE, treated farm dairyeffluent; UFDE, untreated farm dairy effluent; TPFE, treated piggery dairy effluent;TME, treated meat effluent; Bhandral et al., 2007).


(loid)s compared with other samples. Similarly, increase in DOM in soils asa result of wastewater irrigation controls the mobility and bioavailability ofmetal(loid)s (Bolan et al., 2011b; Jackson et al., 2006; Khan et al., 2006).

3.1.3. Cation exchange capacityA long-term increase in SOM content resulting from wastewater irrigation,which is sometimes accompanied by an increased soil pH, can result inan increase of the CEC (Angin et al., 2005; Falkiner and Smith, 1997). Thishas been observed, for example, in a 5-year study conducted in Portugalcomparing the impact of potable water, primary effluent, and secondaryeffluent on various chemical parameters including CEC (Marecos doMonte, 1998). Qishlaqi et al. (2008) reported that the CEC of a sandytopsoil that has been irrigated with raw wastewater for about 20yearsincreased by about 880%. Others, however, did not observe a significantincrease in CEC in spite of significantly increased SOM contents throughwastewater irrigation (Gharaibeh et al., 2007). Madyiwa et al. (2002) studiedthe effects of combined sewage sludge and effluent application on soilproperties of a sandy soil under pasture. The relatively high metal(loid)(Cu, Ni, Pb, and Zn) concentrations within the top 10cm compared to thelower horizons in the irrigated area confirmed the immobility of most heavymetal(loid)s. They argued that considering the lower clay content in top20cm, the high CEC resulting from high OM content of these layersattributed to metal(loid) immobilization. They confirmed that the four


metal(loid)s in their study were strongly correlated to CEC (R2¼0.94–0.99)and OM (R2¼0.88–0.99) in the sewage effluent irrigated soil.

3.1.4. SalinitySalinity is the most restricting factor for using wastewater as an irrigationsource, especially in Australia’s arid climate conditions. It refers to the totalconcentration of all salts in the irrigation water or soil solution and isdetermined by measuring the electrical conductivity (EC) and/or the totaldissolved solid (TDS) content in the water. Long-term wastewater irriga-tion adds large amounts of salts to a soil system (e.g., Bond, 1998; Falkinerand Smith, 1997; Gharaibeh et al., 2007; Gwenzi and Munondo, 2008;Menneer et al., 2001; Xu et al., 2010) as typical TDS concentrations in rawmunicipal sewage and tertiary-treated wastewater range from 200 to 3000mgL�1 (Feigin et al., 1991). Rana et al. (2010) indicated that long-termaddition of sewage water to agricultural lands enhanced EC values from0.99 dS m�1 for well irrigated to 1.65 dS m�1 for sewage irrigated soil.Salinity is likely to be at a minimum immediately after an irrigation eventwhen the soil water content is maximal. Water removal through evapo-transpiration can lead to salt accumulation in the topsoil (increased salinity),which may harm the crop depending on its salt tolerance (Mass andHoffman, 1977). Salts in wastewater can also reduce water availability tothe crop by changing the osmotic potential between plant and soil to theextent that the plant’s water, nutrient, and metal(loid) uptake and yield areaffected. The annual variation of the water balance has to be taken intoaccount when designing wastewater irrigation systems, and sufficient leach-ing for removal of excessive salts from the root zone has to be warranted(Smith et al., 1996). In addition to decreasing plant available water, salinitycan also impact on soil structure through flocculation/deflocculation pro-cesses (Shainberg and Letey, 1984).

McLaughlin et al. (1994) studied the causes of elevated Cd concentra-tions in field-grown potato tubers. They noticed that tuber Cdconcentrations were positively related to soil EC and extractable Cl (R2¼0.62, P<0.001) in the topsoil, with extractable Cl accounting for morevariation than EC. They observed that the tuber Cd was unrelated to tuberconcentrations of P or sulfur (S) but was positively related to concentrationsof Na. They concluded that the cause of elevated Cd concentrations intubers was due to the effect of Cl mobilizing Cd within the soil andincreasing the availability to plants irrigated with saline waters.

3.1.5. SodicityMunicipal wastewater, farm effluents, and effluents from agricultural indus-tries usually have high Na concentrations (e.g., secondary municipal waste-water, 50–250mgL�1; Feigin et al., 1991). Sodium has the opposite effect onsoils to that of elevated salt concentrations (Sumner, 1993). While high levels


of salinity promote flocculation, elevated Na levels enhance clay swelling,clay dispersion, and aggregate slaking. Clay dispersion can lead to structuralbreakdown of a soil and can have adverse effects on soil physical properties,such as soil porosity and permeability (Bond, 1998). The sodicity of irrigationwater can be quantified with the sodium adsorption ratio (SAR), which is thelevel of Na relative to other cations in the irrigation water:

SAR ¼ Naþð ÞffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiCa2þð Þþ Mg2þð Þ

2

q ð1Þ

with concentrations of Na, Ca, and Mg in meqL�1. The SAR of wastewa-ter can vary considerably, often within the range of 4.5–7.9 for secondarymunicipal wastewater (Arienzo et al., 2009; Feigin et al., 1991; Lado andBen-Hur, 2009). The threshold level of SAR in relation to dispersion variesbetween soil types (Sumner, 1995). A measure of a soil’s sodicity is theexchangeable sodium percentage (ESP), the prevalence of exchangeable Nacompared to other exchangeable cations, mainly Ca, Mg, K, hydrogen (H),and Aluminum (Al):

ESP % ¼ 100 exchangeable Nað ÞCEC

ð2Þ

where the Na concentration and CEC are in cmolc kg�1. Many studiesdemonstrated a positive correlation between a soil’s ESP and the SAR of theirrigation water (Harron et al., 1983; Jayawardane et al., 2011; Rengasamy andMarchuk, 2011). In Australia, soils that have more than 6% ESP are consideredto have structural stability problems (Sumner, 1995). This threshold is 15%ESPunder American conditions due principally to the differences in clay mineral-ogy (Halliwell et al., 2001), indicating that the thresholds are not absolutefigures. The impact of increased ESP on soil physical properties is very complexand dependent onmany other factors, such as clay content andmineralogy, ECof the soil solution, SOM, and DOM content, pH, and thus, cannot be readilypredicted (Sumner, 1993). The opposing effects of salinity and sodicity ofirrigation water on soil dispersion mean that while the likelihood of claydispersion increases with high SAR-values, this may be mitigated by theincreased flocculation due to high salt concentrations, an increased EC.

Increases in EC and SAR in soil solutions have been observed withdifferent types of effluents from municipal wastewater to pulp mill effluents(Patterson et al., 2008; Qian and Mecham, 2005; Seikh et al., 1998).In contrast, Hassanli et al. (2008) reported from a 25-month irrigation studyin Iran that the soil SAR decreased significantly under effluent irrigationcompared with borehole water irrigation. Curiously, the quality of theborehole water was often found to be inferior to the effluent quality (SAR,


15 and 8, respectively). Stewart et al. (1990) reported an increase in ESP from3.2% to 9.8% in 0.15–0.35m depth after effluents with an SAR of 5.4 havebeen irrigated to a plantation for 4years. Amongst several treatments, Falkinerand Smith (1997) observed a maximum increase in ESP from<2% to 25% in0.3–0.4m depth under a plantation after 4years of weekly secondary-treatedeffluent irrigation (SAR of effluent, 4.8). Menneer et al. (2001) irrigated twodifferent soil types with sodium-rich dairy factory wastewater over a periodof 5 years and reported an increased ESP of 31% compared to 0.4% at thesoil surface of unirrigated soils. A significant increase was also measured in anIranian trial after 1-year irrigation of municipal waste leachate(Khoshgoftarmanesh and Kalbasi, 2002).

The interrelationship between salinity and sodicity affects soil structureand thus, transport of heavy metal(loid)s. For example, Usman et al. (2005)investigated the effect of immobilizing substances (three clay minerals, ironoxides, and phosphate fertilizers) and NaCl salinity on the availability ofheavy metal(loid)s Zn, Cd, Cu, Ni, and Pb to wheat grown in sewagesludge-amended soil. The plants were irrigated either with deionized orsaline water containing 1600mgL�1 NaCl. They reported that the additionof metal(loid) immobilizing substances—specifically bentonite clay—significantly decreased metal(loid) availability to wheat. They noticed thatirrigation with saline water resulted in a significant increase in metal(loid)chloride species (MClþ and MCl2

0) with the highest concentration observedfor Cd, which was about 53% of its total soil solution concentration. Theyconcluded that saline water increased the availability of Cd and Pb to wheatand decreased the efficiency of bentonite to immobilize soluble Cd.

3.2. Soil biology

Soil biological properties as affected by wastewater application have beeninvestigated with variable results, depending on the experimental designand measurements monitored. For example, traditionally microbiologicalcounts have been reported, whereas in more recent studies molecularbiological approaches concentrating on gene expression and enzymaticactivities, are employed. Controversial results can also be explained by thedifferent nature of wastewater from different sources and the length ofwastewater irrigation. For example, while wastewater irrigation is generallyconsidered as a stimulant of microbial activities, long-term wastewaterirrigation can lead to the accumulation of metal(loid)s, salts, and organiccompounds such as pesticides in soils which might be toxic to soil fauna andflora (Muller et al., 2007). Antibiotics are bioactive compounds and canreach soils through wastewater irrigation, thereby affecting soil biologicalactivity (Kinney et al., 2006). Moreover, wastewater-borne microorganismsmight compete with indigenous microbial communities (Sidhu et al., 2001),thereby affecting the biotransformation of metal(loid)s.


3.2.1. Microbial communitiesA field trial using tertiary-treated domestic wastewater in a Pinus radiataforest on allophanic soils showed no significant differences in microbialbiomass, basal respiration, and sulfatase activity relative to mains waterirrigation possibly due to the low nutrient and carbon contents of thewastewater (Schipper et al., 1996). However, many other studies reporteda positive impact of long-term wastewater irrigation on total microbialbiomass and/or soil enzyme activities in different soils (Barkle et al., 2000;Brzezinska et al., 2006; Degens et al., 2000; Filip et al., 1999, 2000; Friedelet al., 2000; Goyal et al., 1995; Monnett et al., 1995; Ramırez-Fuentes et al.,2002). This phenomenon was ascribed to: first, the enrichment of the soilswith microbial available carbon and nutrient sources stimulating the soilmicrobial populations; and second, to favorable pH and moisture conditions(Filip et al., 2000). Shapir et al. (2000) reported that wastewater irrigationmainly affected the soil microbial communities of the topsoil layers of asandy soil. They emphasized that the observed increase in bacterial countsdid not always correlate with similar changes in bacterial activity.

Blume and Horn (1982) reported a shift in the microbiological popula-tion from aerobic to anaerobic microorganisms due to short-term oxygendepletion of the topsoil resulting from wastewater irrigation, as seen by adecrease in oxygen diffusion rate (Fig. 3; Bhandral et al., 2007). They alsonoticed a higher proportion of nitrifying and ammonifying microorganismsthan in the control. The stimulation of copiotrophic bacteria was observedin the same long-term wastewater irrigation area (Filip et al., 1999). Simi-larly, increased denitrification rates under wastewater irrigation were

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

TFDE UFDE TPFE TME Water Control

Con

cent

ratio

n (m

g cm

–2 m

in–1

)

Figure 3 Oxygen diffusion rate (ODR) values (mgcm�2min�1) from the 10cm soildepth following the winter application of water and a range of effluent types to sheepgrazed pasture (Bhandral et al., 2007).


reported and explained by higher substrate availability (Schipper et al., 1996)or denitrification enhancing surfactants (Friedel et al., 2000). Adenylateenergy charge ratios were reduced and attributed to the addition of Na andsalts with the wastewater irrigation (Friedel et al., 2000). Ramırez-Fuenteset al. (2002) assumed that the observed inhibition of N2O-oxidation in along-term wastewater-irrigated soil was due to the accumulation of salts,heavy metal(loid)s, and toxic organic compounds. Others also observedchanges in the structure and function of microbial communities due towastewater irrigation (Faryal et al., 2007; Oved et al., 2001). These examplesshow that changed environmental conditions in the topsoil of long-termwastewater-irrigated sites can have a selective impact on the compositionof microbial populations and soil functional diversity. It is noteworthy thatthe increase in microbial biomass in a long-term wastewater irrigation areain Berlin, Germany, remained detectable 20years after wastewater irrigationceased (Filip et al., 1999).

Additions of heavy metal(loid) salts to soils usually cause an immediatedecrease in respiration rates, but long-term responses are determined by theproperties of both the metal(loid) and the soil (Nwuche and Ugoji, 2008).According to Brookes (1995), high levels of Pb may have no effect on soilrespiration rates in clay soils but may decrease respiration rates in sandysoils that may be attributed to the difference in bioavailability of Pb betweensoil types. It has been reported that a neutral soil may contain high levels ofMn, Al, or Pb without any sign of toxicity to microorganisms whereastoxicity may develop with certain organisms at much lower metal(loid)concentrations in acid soils (Marschner and Kalbitz, 2003; Utgikar et al.,2003). Some heavy metal(loid)s contained in the wastewater, for exampleCu, Ni, and Zn, are essential trace elements for plants and microorganisms(Alloway, 1995). Even these trace elements, however, may become toxic athigher concentrations (Kosolapov et al., 2004). Copper at high concentra-tion has a detrimental effect on soil microorganisms and modification to thepopulation structure of microbial communities has been reported (Ranjardet al., 2006; Tom-Petersen et al., 2003).

DOM is considered the most dynamic C fraction in soils and it repre-sents a major source of energy and cellular C for the soil microbial commu-nity. Therefore, a close relationship exists between DOM and soil microbialactivity and this C fraction contributes substantially to the total CO2 fluxfrom soils (van Hees et al., 2005). Liu and Haynes (2010) investigated themicrobial activity of soils that had received dairy factory wastewater irriga-tion for greater than 60years. Soil organic C content was unaffected byirrigation but the size (microbial biomass C and N) and activity (basalrespiration) of the soil microbial community were increased. They con-cluded that these increases were attributed to regular inputs of soluble C(e.g., lactose) present as milk residues in the wastewater. Meli et al. (2002)investigated the dynamics of microbial biomass in the soil of a citrus orchard


which has been irrigated for 15years with lagooned municipal wastewater.They noticed that MBC, soluble C, cumulative respiration, and enzymaticactivity were significantly higher in the soils irrigated with wastewater thansoils irrigated with “clear” water; they also found that the metabolic quo-tient (qCO2) was significantly lower in wastewater-irrigated soil, indicatingthat the microbial biomass used the energy sources more efficiently.

3.2.2. EarthwormsIn a field trial, about 500mm dairy shed effluent applied during 270days to asilt loam soil under pasture had a positive effect on earthworm population(numbers, wet weight) compared with a control pasture (Yeates, 1976).This was explained by the increased moisture status of the irrigated soilleading to higher dry matter production and lower earthworm mortalityduring summer as a result of desiccation. The irrigation with dairy factoryeffluents for 22years to pasture on an allophanic soil led to lower earthwormnumbers but a higher biomass of earthworms than in the control, which wasaccompanied by a modified abundance of the five species present (Degenset al., 2000). Increased earthworm numbers were also recorded by Yeates(1995) in a 7-year experiment of sewage application to a 17-year-oldP. radiata plantation on a sandy soil. However, Blume and Horn (1982)warned that high wastewater irrigation rates have a detrimental impact onearthworms due to anaerobic soil conditions as observed under long-termwastewater irrigation “Rieselfelder” around Berlin.

3.3. Soil physics

Long-term wastewater irrigation can affect soil physical and hydraulicproperties (Daniel and Bouma, 1974; Jnad et al., 2000; Lado and Ben-Hur, 2010; Mathan, 1994; Vinten et al., 1983a,b; Vogeler, 2009). Changesin soil physical properties are primarily caused by the impact of wastewaterirrigation on soil chemical properties including soil pH, SOM content andquality, salinity, and sodicity. In this section, we considered potential risksof wastewater irrigation on soil structure, including changes in aggregatestability, bulk density, and hydraulic properties that in turn influence theretention and transport of heavy metal(loid)s.

3.3.1. Aggregate stabilityAggregate stability is an important soil property because it affects waterinfiltration and flow through soils. Wastewater irrigation impacts on soilaggregate stability through the continuous addition of DOM and salts to thesystem (Assouline et al., 2002; Gharaibeh et al., 2007; Menneer et al., 2001;Vogeler, 2009). DOM stabilizes aggregates through its binding actionand increases in microbiological activity (Vogeler, 2009), while Na accu-mulation through wastewater irrigation can lead to aggregate dispersion


(Misra and Sivongxay, 2009). Flocculation of fine soil particles under salineconditions has also been observed, highlighting the complex interactionsbetween sodicity and salinity (Ghadiri et al., 2007). Others found nosignificant effect of wastewater irrigation on aggregates stability (Bhardwajet al., 2007; Levy et al., 2003).

Sodium in wastewater below critical coagulation concentration cancause a reduction in aggregate stability, decrease in infiltration rate, andan increase in the risk of runoff. Alvarez-Bernal et al. (2006) studied theeffect of tannery wastewater on chemical and biological soil properties andobserved that aggregate stability and infiltration properties were adverselyaffected by increased Na content in the effluent.

3.3.2. Bulk density and total porosityChanges in soil bulk density and porosity induced by wastewater irrigationare dependent on the wastewater quality, in particular, the concentrationof dissolved and particulate constituents of the irrigation water. High con-centrations of total suspended solids (TSS) tend to increase soil bulk density,while wastewater with lower TSS contents has no significant impact onbulk density (Magesan, 2001; Sopper and Richenderfer, 1979; Vogeler,2009). Improvements in bulk density and soil porosity have been ascribed tothe addition of DOM (Vogeler, 2009).

Mathan (1994) reported significantly lower bulk density up to a depth of0.6m after 15years of wastewater irrigation to a sandy loam. Total porosity wasincreased by 67% in the topsoil of the wastewater-irrigated soil. An improve-ment in total porosity was measurable up to a depth of 1.2m. Similarly, bulkdensity was significantly decreased and total porosity increased after irrigationof primary-treated wastewater to a sandy soil for 7 years at a rate of 55mm perweek (Magesan, 2001). In contrast, the study by Jnad et al. (2000) demonstrateda decrease in the volume of large soil pores under subsurface drip irrigationwitheffluents. They attributed this shift in pore size distribution to an accumulationof suspended solids and an increased salt concentration leading to clay particledispersion. Amongst 11 soil physical properties including bulk density, fieldwater holding capacity, total porosity, clay content, and saturated hydraulicconductivity measured in the topsoil, Wang et al. (2003b) found that only totalporosity was affected by long-term wastewater irrigation. Shahalam et al.(1998), however, demonstrated that the impact of short-term wastewaterirrigation on porosity of a silt loam soil was not significant. Wastewater irriga-tion led to compaction and reduced total porosity, suggesting that total porositymight be a better soil quality indicator than bulk density (Wang et al., 2003b).

3.3.3. Soil hydraulic conductivity and infiltration rateThe impact of wastewater irrigation on the hydraulic conductivity andinfiltration rate is variable and depends on the soil type, clay content, presenceof CaCO3, antecedent moisture content, the quality of the wastewater and


the irrigation technique (Lado and Ben-Hur, 2009; Lado et al., 2005).A compactive force, for instance the kinetic energy of water drops hittingthe soil, may cause sealing of surface pores due to physical disintegration ofsoil aggregates (Mamedov et al., 2000), aggregate slaking controlled by thewetting rate of the soil (Mandal et al., 2008), TSS of the wastewater(de Vries, 1972) and/or by clay dispersion and the subsequent blocking ofsoil pores by clay particles (Agassi et al., 1981; Lado et al., 2005). For example,in wastewater-irrigated sites with a high ESP-value, the EC of the soilsolution is reduced during rain events which can lead to clay dispersion atthe soil surface, seal formation and subsequently to decreased infiltration rates(Lado et al., 2005; Mandal et al., 2008). In addition, DOM of wastewater canenhance soil water repellency (Assouline et al., 2002; Tarchitzky et al., 2007;Travis et al., 2010; Vogeler, 2009; Wallach et al., 2005). Soil water repellencyinhibits water infiltration (Muller et al., 2010). Reduced infiltration rates havealso been attributed to the collapse of soil structure caused by the dissolutionof SOM (Jnad et al., 2000; Lieffering andMcLay, 1996; Menneer et al., 2001),which can be initiated by alkaline wastewater (pH 11.5–13.5). In contrast,Magesan et al. (1996) showed increased infiltration rates due to increasedmacroporosity, which was explained by the increased biological activityfollowing the wastewater applications. Gharaibeh et al. (2007) observed thatthe length of wastewater irrigation might play a role as well. In their study, upto 5years of wastewater irrigation significantly decreased the infiltration rateof Vertisol, but 15years of wastewater irrigation increased the infiltration ratedue to the formation of large cracks.

Similarly, it has been suggested that various mechanisms affect hydraulicconductivity as a result ofwastewater irrigation (Table 4).Most studies reportedreduced soil hydraulic conductivity for wastewater-irrigated soils (Cook et al.,1994; Gharaibeh et al., 2007; Sopper and Richenderfer, 1979; Vogeler, 2009).Suspended solids of the wastewater can block water-conducting soil pores(Vinten et al., 1983b). The higher the concentration of TSS in the wastewater,the higher the probability of decreased hydraulic conductivity due to blockingof soil pores. A decrease in soil hydraulic conductivity can also be due tobiological (extracellular carbohydrates, cells, and microbiological waste pro-ducts) clogging of soil pores following the stimulation of microbial commu-nities by wastewater microbial growth. Magesan et al. (1999) indicated adecrease in the hydraulic conductivity of an allophanic soil after applyingsynthetic wastewater with a C:N ratio 50:1 for 14weeks in the laboratory.Yet no significant change was evident in field trials on the same soil irrigatedwith tertiary-treated wastewater with a C:N ratio 2:1 for 7 years. Wastewaterwith a high C:N led to net N immobilization, excess C and subsequently toan increase in microbial biomass and extracellular carbohydrates that blockedsoil pores and reduced the hydraulic conductivity. The plugging of soil poreswas shown to be more pronounced in fine textured soils due to the high initialmicroporosity (Vinten et al., 1983b). Further, wastewater irrigation can change


soil chemical properties, in particular the soil’s ESP, salinity, quantity, andquality of SOM and DOM, which impact on soil structure through clayswelling and dispersion, thereby affecting hydraulic properties (Jnad et al.,2000; Lado and Ben-Hur, 2009, 2010; Mandal et al., 2008).

Balks et al. (1996) found at an effluent plantation project in Australia thatthe soil’s dispersion tendency increased after 5 years, but this did not impacton the hydraulic conductivity. No significant change in permeability wasdetected after 5 years of irrigation with tertiary-treated wastewater in amajor Californian study (Seikh et al., 1998). In contrast, Magesan et al.(1996) reported that applying secondary-treated sewage effluents increasedthe macroporosity of a sandy loam soil from 11% to 19% and consequently,the hydraulic conductivity increased from 39 to 57mmh�1. The infiltrationrate influences the transport velocity of heavy metal(loid)s in soils. If theinfiltration rate is small, transport of heavy metal(loid)s will also be limitedor the transport time of heavy metal(loid)s to the groundwater will increase(Lu, 2005).

4. Effect of Wastewater Irrigation on Heavy

Metal(Loid) Dynamics in Soils

Heavy metal(loid)s introduced to soils undergo a number of reactionsthat include adsorption, complexation, precipitation, and reduction, thatcontrol their leaching and runoff losses, and bioavailability. In the case ofwastewater irrigation, these reactions are manifested predominantly by thepresence of high amounts of organic carbon (in particular DOM), solublesalt concentration (salinity), and acidification caused by the mineralizationof organic N.

4.1. Adsorption

The most important physicochemical process affecting the behavior ofmetal(loid)s in soils is its sorption from liquid to solid phase (Bolan et al.,1999; Li et al., 2006; Sparks, 2003). The retention and movement of heavymetal(loid)s in soils can be correlated with soil clays, surface area of particles,CEC/AEC, and soil pH (Kabata-Pendias and Pendias, 2001). For example,some studies have shown that the sorption of metal(loid)s by soils tends toincrease with increasing pH (Naidu et al., 1996; Violante et al., 2010), OM(Lair et al., 2007), CEC (Buchter et al., 1989; Kwon et al., 2010), and thecontents of Fe (Karpukhin and Ladonin, 2008) and Mn oxides (Brown andParks, 2001; Stahl and James, 1991).

It has often been observed that heavy metal(loid)s added through organicamendments, such as effluents, sewage sludge, and manures accumulate in


the surface layer, indicating a strong retention within surface soils enrichedwith carbon (Adriano, 2001). Table 6 shows selected references on theeffect of wastewater/sludge addition on heavy metal(loid) adsorption. Zhuet al. (1991) observed that the Cu adsorption maxima by two soils whichreceived 11 annual applications of Cu in the form of Cu-enriched swinemanure or CuSO4 were higher for the manure-treated soil. This was due toan increase in OM-induced CEC. Addition of organic amendments hasoften been shown to increase the CEC of soils, thereby resulting inincreased cation adsorption. The effect of effluent and manure additionon CEC-induced cation adsorption in soils is often inferred from literaturein which the effects of other organic amendments, such as biosolid havebeen examined. Bolan et al. (2003b), for example, observed that CEC perunit organic carbon was higher for soils than for biosolid, which wasattributed either to the difference in the nature of OM or to the significantcontribution of negative charge by the mineral components in soils.

Although a number of studies have shown general increases in metal(loid)adsorption with effluent and biosolid addition (Juste and Mench, 1992;Li et al., 2001), Zhou and Wong (2001) observed that with Cu, adsorptionby both acidic and calcareous soils decreased with the addition of sewagesludge, which they attributed to the formation of soluble Cu-DOM complex.There is much evidence in the literature stating that DOM can reduce metal(loid) adsorption onto soils (Ashworth and Alloway, 2008; Baham andSposito, 1994; Davis, 1984; Elliott and Denneny, 1982; Gove et al., 2001;Xu et al., 1989). A significant inverse relationship between the extent of Cuadsorption and the DOM in the soils treated with organic amendments wasobserved by Bolan et al. (2003c) and Hao et al. (2008). These Cu-DOMcomplexes are highly mobile in soils and may increase the leaching of Cu.

Wong et al. (2007) noted that the addition of DOM through anaerobicallydigested dewatered sludge significantly reduced theCd sorption capacitywith amaximum inhibition onmetal(loid) sorption occurring at pH 7–7.5. Al-Wabelet al. (2002) reported a positive correlation between increased soil DOMresulting from biosolid application and Pb and Cu concentrations indicatingformation of soluble metal(loid)-DOM complexes. Kunhikrishnan (2011)examined the effect of farm dairy, winery, and piggery wastewaters on theadsorption of Cd, Cu, and Pb using batch experiments and showed thatadsorption decreased in all the soils in the presence of wastewater sources.Results indicated that DOM in wastewater sources formed soluble metal(loid)complexes and consequently reduced the adsorption of Cu, Cd, and Pb.

4.2. Complexation

Heavy metal(loid)s form both inorganic and organic complexes with a rangeof solutes. As might be expected, the organic component in wastewater hasa high affinity for metal(loid)s due to the presence of ligands or functional

Table 6 Selected references on the effect of wastewater and waste sludge on heavy metal(loid) adsorption and complexation reactions

Metal(loid)s

Wastewater/

sludge Observations References

Cu, Fe, Mn,

Ni, Pb,

Zn

Sewage

effluent

Sewage irrigation for 20years resulted into significant build-up of

DTPA-extractable Zn (208%), Cu (170%), Fe (170%), Ni

(63%), and Pb (29%) in sewage-irrigated soils over adjacent tube

well water-irrigated soils, whereas Mn was depleted by 31%.

Rattan et al.

(2005)

Cd, Cr, Cu,

Ni, Pb,

Zn

Sewage water

or sludge

Concentrations of Cr in the sewage-irrigated soils exceeded the

permissible limits, the concentration of Zn in 55.6% of the

samples, and 44.4% for Cu were above the limits, while Pb and

Cd did not exhibit values beyond the allowable limits.

Chen et al.

(2010)

Cu, Cr, Ni,

Zn

Reclaimed

wastewater

Irrigation with effluents also increased both the total and EDTA-

extractable metals in the fields. Highest levels of EDTA-

extractable elements were at top 20-cm layers, and available

fractions decreased with depth. Long-term irrigation (8 and 20

years) significantly increased EDTA-extractable Cu and Ni at

top 50-cm profiles, while only increased EDTA-extractable Cr

and Zn on top 30-cm soils.

Xu et al.

(2010)

Cr, Cu, Pb,

Zn

Municipal

wastewater

3years after discontinuation of wastewater application on organic

soils, heavy metals in soils were below the upper permissible

limits. Also the basic soil properties (OM, pH, BD, WHC, and

P2O5) were not changed.

Brzezinska

et al.

(2010)

Cd, Cu, Fe,

Mn Ni,

Pb, Zn

Untreated

sewage

effluent

Organic carbon content showed positive correlation with all heavy

metals except Zn. Degradation of sludge organic matter released

heavy metals in sewage sludge-amended soils.

Rana et al.

(2010)

Cu, Zn Sewage sludge Cu and Zn sorption capacity decreased in the presence of DOM.

The kd values for Cu without and with DOM were 121.20 and

36.88 and for Zn the values were 33.58 and 14.825 for Zn,

respectively.

Mesquita

and

Carranca

(2005)

Cu Sewage sludge Complexation of Cu by sewage sludge-derived dissolved organic

matter occurred due to reduced soil sorption and the

complexation was greatest at intermediate pH values.

Ashworth

and

Alloway

(2007)

Cd, Zn Anaerobically

digested

dewatered

sludge

The addition of DOM significantly reduced the Cd and Zn

sorption capacity with a maximum inhibition on metal sorption

occurring at pH 7–7.5.

The kd values for acidic sandy loam soil in the absence and

presence of DOM were 22.2 and 9.54 for Cd and 3.86 and 1.84

for Zn. The kd values for calcareous sandy loam soil in the

absence and presence of DOM were 329 and 107 for Cd and

212 and 40.2 for Zn.

Wong et al.

(2007)


groups that chelate metal(loid)s (Harter and Naidu, 1995). With increasingpH, the carboxyl, phenolic, alcoholic, and carbonyl functional groups inOM dissociate, thereby increasing the affinity of ligand ions for metal(loid)s.

It has been observed that addition of wastewater, sewage sludge, ormanure by-products increases the complexation of metal(loid)s in soils,the extent of this relates to the DOM concentration (Hesterberg et al.,1993) (Table 6). Complexation can result in the formation of both solubleand insoluble metal(loid)-DOM complexes, thus affecting both movementand bioavailability of heavy metal(loid)s. While insoluble complexes resultin the retardation of DOM and metal(loid) movement (Guggenberger andKaiser, 2003; Jansen et al., 2005; Martin and Goldblatt, 2007), soluble metal(loid)-DOM complexes enhance their movement. Accordingly, in soilscontaining large amounts of OM, such as pasture soils and organic manureor wastewater-amended soils, only a small proportion of metal(loid)s in soilsolution remains as free metal(loid) ion and a large portion are complexedwith DOM (Bolan et al., 2003a,c; Haruna et al., 2009; McLaren and Ritchie,1993). For example, del Castilho et al. (1993) observed that 30–70% of thedissolved Cu and all Cd in soils treated with cattle manure slurry was boundin relatively fast dissociating organic-metal(loid) complexes.

Although the formation of soluble metal(loid)-organic complexes reducesthe phytoavailability of heavy metal(loid)s, the mobility of the heavy metal(loid) may be greater in soils receiving alkaline-stabilized biosolid due to thereduction of metal(loid) adsorption and increased concentration of solublemetal(loid)-organic complex in solution (Brown et al., 1997; Gove et al.,2001). It has often been found that in manure- and effluent-amended soils,a large portion of Cd and Cu is complexed with DOM within soil solution(Buzier et al., 2006; del Castilho et al., 1993; van Veen et al., 2002). Similarly,Hyun et al. (1998) and Shan (2010) found a linear relationship betweenorganic carbon and soluble Cd in solution for sludge-treated soils, indicatingthat most of the Cd remained as metal(loid)-organic complex. As reported byBolan et al. (2003c), a decrease in Cu adsorption in the presence of DOM islikely to increase Cu mobility yet does not necessarily increase bioavailability.

Application of effluent and manure has been shown to increase the solublesalt concentration of soils, as measured by EC (del Castilho et al., 1993; Ranaet al., 2010; Rusan et al., 2007; Sutton et al., 1984). High concentrationof inorganic anions, such as Cl� and sulphate (SO4

2�) in effluents and manureproducts induces the formation of metal(loid)-inorganic complexes (e.g., Cd–Cl complex) that are considered to be even more phytoavailable (Japengaand Harmsen, 1990; Khoshgoftarmenesh et al., 2002; McLaughlin et al., 1998;Smolders et al., 1998). Although a wide variety of organic compounds inDOM contribute to the formation of soluble complexes with metal(loid)s,Daum and Newland (1982), del Castilho et al. (1993), and Zhou and Wong(2001) observed that the low-molecular-weight fractions, such as hydrophilicbases have strong affinity for forming soluble complexes with Cd, Cu, and Zn.


Thus, the formation of soluble aqueous metal(loid)-organic and to a lesserextent metal(loid)-inorganic complexes is expected to dominate the solutionchemistry ofmetal(loid)s inwastewater andmanure-amended soils (Hesterberg,1998; Hesterberg et al., 1993; Lipoth and Schoenau, 2007).

4.3. Redox reactions

As discussed above, adding biological waste materials, such as wastewater,livestock and poultry manures, and sewage sludge has often been shownto increase the amount of DOM in soils (Bolan et al., 2011b; Park et al.,2011; Schindler et al., 1992; Sleutel et al., 2006). These wastes of plant andanimal origin contain large amounts of DOM, and the addition of certainorganic manures, such as poultry manure increases the pH, thereby enhanc-ing the solubilization of SOM (Jackson and Miller, 2000; Jackson et al.,1999). Such an increase in DOMmay enhance microbial activity but lowerthe redox potential in the soil (Fig. 3; Bhandral et al., 2007; Luo et al., 2008;Redman et al., 2002).

A number of studies have shown that addition of OM-rich soil amend-ments enhances the reduction or biotransformation of certain heavy metal(loid)s, such as As, Cr, and Se (Alexander, 1999; Frankenberger and Losi,1995; Losi et al., 1994) (Table 7; Fig. 4). For example, Ajwa et al. (1998)noticed greater loss of Se from manure-borne Se than from inorganicfertilizer-borne Se, which they attributed to manure-facilitated volatiliza-tion due to the reduction of Se. Similarly, Banks et al. (2006), Cifuenteset al. (1996), Higgins et al. (1998), and Losi et al. (1994) reported a reductionof Cr(VI) to less toxic and less mobile Cr(III) in soils amended with cattlemanure. Various reasons could be attributed to the enhanced reductionof Cr(VI) in the presence of organic amendments, including the supply ofcarbon and protons and the stimulation of microorganisms that mediate andfacilitate the reduction of Cr(VI) to Cr(III) (Losi et al., 1994). Zhao et al.(2009) investigated the transport and fate of Cr(VI) and As(V) in soil zonesderived from moderately contaminated farmland irrigated with industrialwastewater for 30years. A column test showed that the concentration ofCr(III) and As(III) in the leachate increased by 6% and 5.6%, respectively,indicating DOM-induced reduction of these metal(loid)s (Fig. 5).

Under similar organic carbon loading, Bolan et al. (2003d) observed asignificant difference in the extent of Cr(VI) reduction between variousorganic manure composts. Reduction increased with increasing level ofDOM added through manure addition, which has been identified to facili-tate the reduction of Cr(VI) to Cr(III) in soils (Jardine et al., 1999; Nakayasuet al., 1999). For example, the hydroquinone groups in OM have beenidentified as the major source of electron donor for the reduction of Cr(VI)to Cr(III) in soils (Elovitz and Fish, 1995).

Table 7 Selected references on the effect of wastewater/manure addition on heavy metal(loid) reduction

Metal

(loid)s Wastewater/manure Observations References

As and Cr Industrial

wastewater

Cr(VI) and As(V) were reduced to Cr(III) and As(III)

indicating DOC-induced reduction

Zhao et al. (2009)

As Poultry litter Poultry litter increased the solubility of As by

complexation with DOC

Jackson et al. (2003)

Cr Cattle manure Enhanced the reduction of Cr(VI) to Cr(III) Cifuentes et al. (1996)

Cr Cattle manure Enhanced the reduction of Cr(VI) to Cr(III) Higgins et al. (1998)

Cr Composted cow

manure

Chromate leaching was reduced in soils in the

presence of elevated organic matter because of

reduction followed by retention on cation

exchange sites or precipitation

Banks et al. (2006)

As Sewage sludge and

poultry litter

The mixture of sewage sludge and poultry litter

reduced As(V) to more mobile and toxic As(III)

Jackson et al. (1999)

0 10 15 20

Time (weeks)

0

400

800

1200

Cr(

VI)

(mg

L–1)

0 Mg OM ha–1

50 Mg OM ha–1

5

Figure 4 Effect of organic matter addition on Cr(VI) reduction in soils (Losi et al.,1994).

Before0

20

40

60

80

100

120

After Before

Arsenic

After

Chromium

Met

al c

once

ntra

tion

(%)

Cr(III) As(III)As(V)Cr(VI)

Figure 5 Concentration of Cr(VI), Cr(III) and As (V), As(III) in wastewater-irrigatedsoil before and after column tests (Zhao et al., 2009).


The increase in Cr(VI) reduction in the presence of manure and effluentaddition may also result from enhanced microbial activity. Although Cr(VI)reduction can occur through both chemical and biological processes, thebioreduction is considered to be the dominant process in most arable soilsthat are low in Fe2þ ion. Losi et al. (1994) have reported that adding manure


compost generated a larger increase in the bioreduction than chemoreduc-tion of Cr(VI), indicating that the supply of microorganisms is moreimportant than the supply of organic carbon in enhancing the reductionof Cr(VI) when compost is added. It has often been reported that an increasein microbial activity will in turn increase the reduction of Cr(VI) to Cr(III)(Losi et al., 1994; Rajkumar et al., 2005; Sultan and Hasnain, 2006).

Protons are required for the reduction of Cr(VI) to Cr(III) (Eq. (3)).Wastewater and manure compost are generally rich in N, part of which isin the ammoniacal form. Oxidation of ammoniacal nitrogen to nitratenitrogen (nitrification) and ammonia volatilization result in the release ofprotons. It has often been observed that Cr(VI) reduction, being a protonconsumption (or hydroxyl release) reaction, increases with a decrease in soilpH (Cary et al., 1977; Eary and Rai, 1991).

2Cr2O7 þ 3C0 þ 16Hþ ! 4Cr3þ þ 3CO2 þ 8H2O ð3Þ

Increased concentration of Fe2þ and Mn2þ ions in drainage effluent frommanure- and effluent-amended soil is related to reducing conditions with theconsequent solubilization of these metal(loid)s in soils. Metal(loid)s, such asCo, are retained by Fe2þ and Mn2þ oxides under oxic conditions (McLarenet al., 1984) and the manure/effluent-induced reduction of these oxidesresults in the release of adsorbed metal(loid)s (L’Herroux et al., 1997; Siebeand Fischer, 1996). Wallingford et al. (1975) obtained a good correlationbetween Mn concentration in corn and cumulative level of feedlot manureapplication, which was attributed to enhanced solubilization of Mn due toreducing conditions in manure-treated soil.

4.4. Methylation/demethylation

Methylated derivatives of As, Hg, and Se can arise as a result of chemical andbiological processes that frequently alter their volatility, solubility, toxicity,and mobility. Biomethylation of these heavy metal(loid)s has emerged as amajor process for their removal during wastewater treatment using natural andconstructedwetlands (Kosolapov et al., 2004; Stasinakis andThomaidis, 2010).The major microbial methylating agents are methylcobalamin (CH3CoB12),involved in the methylation of Hg, and S-adenosylmethionine, involvedin the methylation of As and Se. Biomethylation may result in metal(loid)detoxification since methylated derivatives are excreted readily from cells, areoften volatile, and may be less toxic, for example, organoarsenicals.

Although methylation of heavy metal(loid)s occurs through both chemical(abiotic) and biological processes, biomethylation is considered to be thedominant process in soils and aquatic environments. At present there is substan-tial evidence for the biomethylation of As, Hg, and Se in soils and aquaticsystems (Gadd, 2004; Masscheleyn and Patrick, 1993; Nicholas et al., 2003;


Oiffer and Siciliano, 2009) and during wastewater treatment (Stasinakis andThomaidis, 2010). Microorganisms in soils and sediments act as biologicallyactive methylators and OM derived from wastewater application provides thesource of methyl donor for methylation (Kosolapov et al., 2004).

Benthic microbes are capable of methylating As under both aerobic andanaerobic conditions to produce methylarsines and methyl-arsenic com-pounds (Maher, 1988). Methylation may play a significant role in the mobili-zation of As by releasing it from the sediments to the aqueous environment(Anderson and Bruland, 1991). Similarly, methylated Hg species that arehighly toxic and biologically mobile have been observed in various wastewa-ter sources including dental wastewater (Gbondo-tugbawa et al., 2010; Gustinet al., 2006; Zhao et al., 2008).

When selenate [Se(VI)] and selenite [Se(IV)] are introduced into mod-erately reducing conditions they are quickly transformed through microbialprocesses to Se0 and/or organic Se compounds. Selenium biomethylation isof interest because it represents a potential mechanism for the removal of Sefrom contaminated environments, and it is believed that methylated com-pounds, such as dimethyl selenide are less toxic than dissolved Se oxyanions(Frankenberger and Losi, 1995). For example, biomethylation followed byvolatilization is considered as one pathway by which high Se concentrationsare dissipated from agricultural evaporation ponds in the San Joaquin Valleyof California (Gao and Tanji, 1995).

Literature reports that wastewater from some industries contains quitehigh concentrations of soluble Se, as high as 620mgL�1 from the Se com-pounds industry and 20–60mgL�1 from the Cu refining industry (Fujita et al.,2002). Moreover, in a recent study investigating the distribution of selenate,selenite, and selenocyanate (SeCN) in wastewater of an oil refinery plant(Miekeley et al., 2005), SeCN was by far the most abundant Se species,reaching concentrations of up to 90mgL�1. The authors reported that selenitewas detected only in one sample and selenate could not be identified in anyof the analyzed samples. In a study investigating Se removal from selenite-contaminated oil refinery wastewater wetland, Hansen et al. (1998) foundthat 89% of the Se was removed. They found that most of the Se wasimmobilized into the sediment and plant tissues, whereas biological volatili-zation could have accounted for 10–30% of its removal. Ye et al. (2003)reported that 79% of initial Se mass contained in coal gasification plantwastewater was removed in a constructed wetland. The primary sink for Seretention was the sediment, which accounted for 63%, whereas accumulationin plant tissues and biological volatilization to the atmosphere were of minorimportance (Ye et al., 2003).

Frankenberger and Arshad (2001) observed that microorganisms, partic-ularly Enterobacter cloacea, were very active in reduction of Se oxyanionspresent in irrigation drainage water, into insoluble Se0. Furthermore, bymonitoring various environmental conditions and addition of organicamendments, they confirmed that the process could be stimulated manifold.


4.5. Leaching and runoff

Long-term application of wastewater to soils can potentially affect thequality of groundwater resources by excess nutrient loading and heavymetal(loid) mobilization beyond the plant root zone (Abaidoo et al., 2009;Hamilton et al., 2007). The impact depends on a number of factors includ-ing the depth of the water table, soil properties, soil drainage, managementof wastewater irrigation, quality of groundwater, and the scale of wastewaterirrigation. The capacity for heavy metal(loid)s to contaminate groundwaterrelies on the mobility of the heavy metal(loid) concerned, and the amountsand proportions of complexed and free metal(loid) forms within the soilsolution. The leaching rate of heavy metal(loid)s is also influenced by thenatural OM content of the soil, the concentration and quality of DOM, andpH of the leaching solution (Antoniadis and McKinley, 2003; van Zomerenand Comans, 2004).

Many studies have examined the leaching behavior of heavy metal(loid)sfrom contaminated soils, industrial sludges, dredged sediments, and municipalsolid wastes (Dijkstra et al., 2004; Meima and Comans, 1997; Voegelin et al.,2003) (Table 8). The potential risk of heavy metal(loid)s in soils, with respectto their mobility and ecotoxicological significance, is determined by theirsolid-solution partitioning rather than the total heavy metal(loid) content(Dijkstra et al., 2004; Shi et al., 2009). The release of heavy metal(loid)s tosoil solution depends on their affinity to bind to reactive surfaces in the soilmatrix (Dijkstra et al., 2004). Downward migration of heavy metal(loid)s inwastewater is facilitated by forming soluble complexes with DOM (Zhou andWong, 2001).

L’Herroux et al. (1997) observed that repeated applications of swinemanure slurry increased the drainage water concentrations of Mn from 0.05to 14mgL�1, Co from 0.8 to 50mgL�1, and Zn from 17.3 to 100mgL�1.Studies on migration of metal(loid)s in soils after manure slurry applicationshave linked metal(loid) mobility with DOM (Amery et al., 2010; Japengaet al., 1992). Although the soluble organic metal(loid) fraction is not readilybioavailable to plants, it is relatively mobile and applying organic amend-ments including wastewater, biosolid, and animal manure has been shownto enhance the leaching of metal(loid)s in soils (Hsu and Lo, 2000). DelCastilho et al. (1993), for example, observed a positive relationship betweensoluble metal(loid) concentration and DOM in soils treated with cattlemanure slurry. Li and Shuman (1997) observed that leaching metal(loid)-contaminated soils with poultry litter extract increased the water-solublefractions of Cu and Zn, with a corresponding decrease in exchangeablefractions, indicating that poultry manure application enhances the solubili-zation and mobilization of metal(loid)s. Acidification caused by manureapplication due to nitrification also results in the release of soil metal(loid)s(del Castilho et al., 1993; Japenga et al., 1992).

Table 8 Selected references on the effect of wastewater and waste sludge on heavy metal(loid) leaching

Metal(loid)s Wastewater Observations References

Cd, Cu Treated

wastewater

Preferential flow and metal complexation with soluble organics

apparently allowed leaching of heavy metals.

Behbahaninia et al.

(2008)

Cu, Cd Untreated

wastewater

Water extractable Cu and Cd concentrations and the metal leachates

increase and correlate with DOC.

Herre et al. (2004)

Hg, Cd Untreated

wastewater

Groundwater was not contaminated through vertical infiltration-

induced leaching. However, substantial build-up of metals in river

sediments and wastewater-irrigated soils were observed.

Wu and Cao (2010)

Zn, Cu, Cd,

Cr

Treated sewage

effluent

Zn, Cu, and Cd mobility was observed due to acidic soil pH. Gwenzi and

Munondo (2008)

Cr, Zn, Cd,

Cu, Pb

Poultry litter Leaching of metals increased with increasing rates of poultry litter. Paramasivam et al.

(2009)

Cd, Ni, Zn Sewage sludge DOM applications significantly increased the extractability of metals. Antoniadis and

Alloway (2002)

Zn, Cd, Cu,

Ni, Cr

Sewage sludge The concentrations of Zn, Cd, Cu, Ni, and Cr in the saturation extract

closely correlated with the concentrations of DOM. Considerable

amounts of Zn and Cd from sewage sludge were found in the mobile

fractions of the soil with Cu, Ni, and Pb in organic particles.

Schaecke et al.

(2002)

Cu, Ni, and

Pb

Sewage sludge The solubility of the heavy metals showed a strong positive relationship

to the solubility of organic matter, particularly at high pH.

Ashworth and

Alloway (2008)

Cu, Cd, Pb,

and Zn

Pig manure

amendment in

mine soils

Pig manure amendment increased DOM in leachates, thereby

increasing the release of metals from mine soil.

Carmona et al.

(2008)

Cu, Zn Poultry and

livestock

manures

Total amounts of Cu and Zn eluted from the soil columns significantly

correlated with the extracted soil Cu and Zn concentrations.

Hao et al. (2008)


DOM plays an important role in facilitating the leaching of contaminantsin soil (Haberhauer et al., 2002; van Zomeren and Comans, 2004) by formingsoluble metal(loid) complexes (Bolan et al., 2011b; McCarthy and Zachara,1989; Weng et al., 2002). Herre et al. (2004) did a column experiment tostudy the effect of wastewater on the leaching of metal(loid)s (Cu and Cd) andDOM. They found that the amount of Cu leached correlated well with theDOM concentrations in the leachates (Fig. 6A, B). This agrees well withmany published reports emphasizing the importance of DOM for metal(loid)

0.40

100

200

300

400

500

Wa

ter

ex

tra

cta

ble

Cu

(mg

kg

-1)

Wa

ter

ex

tra

cta

ble

Cu

(mg

kg

-1)

Vertisol R2 = 0.71

Leptosol R2 = 0.54

0.20 0.2 0.6 0 0.1 0.3 0.4 0.5

DOC (mg g-1)DOC (mg g-1)

0

2

4

6

8

10

Vertisol R2 = 0.75

Leptosol R2 = 0.50

0 50 150100 200 250

DOC concentration (mg C kg-1 soil)

0

20

40

60

80

100

120

Cd

co

nc

en

tra

tio

n (

mg k

g- 1

)

C

Arthrosol R2 = 0.81

BA

Figure 6 Relationship between DOC and water extractable Cu (A) and Cd (B, C) indifferent soils [(A, B): Herre et al., 2004; (C): Shan, 2010].


mobility (Christensen et al., 1999; Ermakov et al., 2007; Kalbitz andWennrich, 1998; Zhu and Alva, 1993) (Fig. 6C). The presence of wastewaterDOMwould maintain metal(loid)s in solution and thus limit adsorption ontothe soil (Ashworth and Alloway, 2004; Harter and Naidu, 1995).

Many researchers have viewed DOM as an important contributor to theelevated mobility of heavy metal(loid)s in soils treated with wastewater, man-ures, and biosolids (Al-Wabel et al., 2002; Haynes et al., 2009; Kunhikrishnan,2011; Peckenham et al., 2008). For example, Kunhikrishnan (2011) examinedthe effect of piggery, farm dairy, and winery wastewaters on Cu leaching andnoticed that leaching increased with increasing levels of Cu and was higherin soils treated with wastewater sources than Milli-Q water. Kunhikrishnan(2011) suggested that the DOM in wastewater sources formed solubleCu-DOM complexes, thereby facilitating the movement of Cu in soils(Fig. 7). Thus, whilst free metal(loid) ions or readily dissociated inorganicallycomplexed metal(loid)s added to a soil would be expected to become quicklyadsorbed to soil solids (e.g., via cation exchange and complexation reactions),soluble organometal(loid) complexes may be maintained in the soil solution.

5. Bioavailability of Wastewater-Borne Heavy

Metal(Loid)s in Soils

Bioavailability of wastewater-, sludge-, and manure-borne metal(loid)sin soils can be examined using chemical extraction and bioassay tests. Chemicalextraction tests include single extraction and sequential fractionation (Basta and

0.2

0.4

0.6

0.8

00 4

PEWEFDEMQ

PEWEFDEMQ

8

Pore volumes

12

1.2

1.6

0

0.4

0.8

0 4

Pore volumes

8 12Cu

mu

lati

ve C

u c

on

cen

tra

tio

n (

mg

kg

-1)

A B

Figure 7 Effect of wastewater irrigation on cumulative Cu concentration of leachates ina silt loam soil, (A) 100mgkg�1 and (B) 500mgkg�1 (MQ, Milli-Q water; PE, piggeryeffluent; WE, winery effluent; FDE, farm dairy effluent; Kunhikrishnan, 2011).


Gradwohl, 2000; Ruby et al., 1996). Bioassay involves plants, animals, andmicroorganisms (Naidu et al., 2008; Yang et al., 1991).

5.1. Chemical extraction

5.1.1. Single extractionThe feasibility of predicting the bioavailability of heavy metal(loid)s tohigher plants and various organisms is assessed using selective chemicalextractants (Kelsey et al., 1997; Loibner et al., 2000). Both single extractions(Beckett, 1989) and sequential extractions (Tessier et al., 1979) are used toidentify those fractions of metal(loid)s in the soil that are more or less readilyavailable (Kennedy et al., 1997). Bioavailability is organism- and species-specific and a single chemical test is insufficient to precisely assess bioavailabil-ity accurately (Reid et al., 2000). However, extraction with non-exhaustiveselective extractants that mimics the bioavailability of pollutants is useful forproviding predictors of exposure.

Several methods have been used to evaluate the bioavailability of heavymetal(loid)s in soils which are based mainly on extractions by various solutions:(a) acids—mineral acids at various concentrations (e.g., 1N HCl), (b) chelatingagents (e.g., EDTA, DTPA), (c) buffered salts (e.g., 1MNH4OAc), (d) neutralsalts (CaCl2, NH4NO3), and (e) other extractants proposed for routine soiltesting. These extractants have been used to predict the bioavailability offertilizer-, wastewater-, manure-, and sludge-borne heavy metal(loid)s in soils(Gupta and Sinha, 2007; Marchi et al., 2009; Payne et al., 1988; van derWatt et al., 1994). Chelating agents such as EDTA and DTPA have oftenbeen found to bemore reliable in predicting the plant availability of sludge- andwastewater-borne heavy metal(loid)s (Gupta and Sinha, 2007; Sims andJohnson, 1991), since they are more effective in removing soluble metal(loid)-organic complexes that are potentially bioavailable. However, it shouldnot be readily assumed that these chelating agents actually measure availability(Beckett et al., 1983a, b; Peijnenburg et al., 2007).

Jagtap et al. (2010) conducted a study to ascertain the addition of heavymetal(loid)s, Cr, Cd, Cu, Ni, Pb, and Zn into agricultural fields throughmunicipal wastewater irrigation. They analyzed the available concentrationof heavy metal(loid)s using DTPA extraction and found a maximum of64.84% extraction in the case of Cr. They attributed the low extraction ofother metal(loid)s to the formation of high-affinity complexes of metal(loid)s and soil particles. They reported that the extraction of heavy metal(loid)s is dependent on pH, EC, CaCO3, organic carbon, type of soil, andmethod of extraction. They also noticed that EDTA was more suitable foracidic soils, whereas DTPA was considered more suitable for neutral andnear alkaline soils, as it buffered pH at 7.3 and therefore prevented CaCO3

from dissolution and release of occluded metal(loid)s (Chen et al., 2009; Linand Zhou, 2009). Luo et al. (2003) studied the accumulation, chemical


fractionation, and availability of Cu to rice in a paddy soil irrigated with Cu-enriched wastewater. They found that with irrigation the concentrations ofCu in the exchangeable (NH4OAc-extractable) and complexed (EDTA-extractable) fractions increased rapidly, from 0.33 to 6.30mgkg�1 and from14.1 to 98.0mgkg�1, respectively. EDTA-extractable Cu was much higherthan NH4OAc-extractable Cu in the soils.

Calcium chloride (CaCl2) soil extraction, although a neutral salt extrac-tion, is a widely used and internationally recognized technique (Harmsen,2007; Peijnenburg et al., 2007). van der Welle et al. (2007) reported thatunder aerobic conditions, plant heavy metal(loid) uptake was best predictedby the amount of CaCl2-extractable metal(loid)s. Metal(loid) extractionwith CaCl2 solution was found to be effective, with Sauve et al. (1996)and McBride (2001) reporting that Cu2þ in 0.01M CaCl2 correlatedstrongly with plant yields or tissue Cu concentrations of rye grass andcrop species such as maize, lettuce, and radish. Wightwick et al. (2010)determined the environmental availability of Cu in Australian vineyard soilscontaminated with fungicide derived Cu residues irrigated with treatedsewage. They reported that differences in Cu availability determined by0.01M CaCl2 extractable Cu concentrations were related to the total Cuconcentration and soil properties, including pH, clay, exchangeable K, silt,and CaCO3. Kunhikrishnan et al. (2011) compared the free Cu2þ concen-trations in CaCl2 extract and pore-water from soils in the presence of farmdairy and piggery wastewater sources. They reported that the free Cu2þ

concentrations were lower in soils incubated when wastewater sourceswere present. These results suggest the formation of Cu-DOM complexesdecreases the amount of free Cu2þ in the soil solution.

5.1.2. Sequential fractionationFractionation studies are often used to examine the influence of amend-ments, such as wastewater, CaCO3, P compounds, and biosolid on theimmobilization of heavy metal(loid)s. Following adsorption, irrespectiveof the nature of interaction between heavy metal(loid)s and soil colloidalparticles, metal(loid) ions are redistributed amongst organic and mineral soilconstituents (Bolan et al., 2003e; Fedotov and Miro, 2008). Factors affect-ing the distribution of heavy metal(loid)s among different forms includepH, ionic strength of the soil solution, solid and solution components aswell as their relative concentration and affinities for heavy metal(loid), andreaction time (Bolan et al., 2003e; Shuman, 1991). The various formsof the heavy metal(loid)s that are sequentially extracted can be classifiedas soluble, adsorbed/exchangeable, carbonate-bound, organic-bound,amorphous ferromanganese hydrous oxide-bound, crystalline ferroman-ganese hydrous oxide-bound, and residual or lattice mineral-bound.The phytoavailability of the different forms of the solid phase speciesgenerally decreases in the following order: soluble>adsorbed/exchangeable


>organic-bound>carbonate-bound>ferromanganese hydrous oxide-bound>residual or refractory (i.e., fixed in mineral lattice) (Tessier et al., 1979).

Studies suggest treatment of soils with organic amendments such assludge or wastewater shifts the solid phases of the heavy metal(loid)s awayfrom immobile fractions to forms that are potentially more mobile, labile,and bioavailable. For example, Dudka and Chlopecka (1990) found withsewage sludge application the residual forms of Cd2þ, Cu2þ, and Zn2þ insoil decreased from 34–43% to 6–34%, with a corresponding increase inthe readily phytoavailable forms. Through sequential extraction of Cu,Cd, Pb in four soils irrigated with wastewater, Flores et al. (1997) discov-ered metal(loid)s were predominantly associated with organic soil frac-tions. Bashir et al. (2007) studied the fractionation of heavy metal(loid)s (Cd, Mg, and Zn) in soils irrigated with untreated sewage effluent for along period of time. Extraction procedure showed that most of the heavymetal(loid)s (>50%) was bound to residual fraction. Among nonresidualfractions, Cd and Mn were present in reducible fraction while Zn waspresent in oxidizable fraction. Luo et al. (2003) analyzed the fractionationof Cu in a paddy soil irrigated with Cu-enriched wastewater. Theyreported marked increases in the weak acid-soluble (HOAc-extractable),reducible Fe and Mn oxide-bound (NH2OH�HCl-extractable), oxidiz-able OM-bound (H2O2-extractable), and residual fractions of Cu in thewastewater-irrigated soils, indicating an increase in mobility and bioavail-ability of Cu leading to Cu toxicity in the plants.

Physiologically based in vitro chemical fractionation schemes are becom-ing increasingly popular for examining the bioavailability of heavy metal(loid)s (Basta and Gradwohl, 2000; Juhasz et al., 2009; Ruby et al., 1996).These schemes include physiologically based extraction tests (PBET),potentially bioavailable sequential extraction (PBASE), simplified bioacces-sibility extraction test (SBET), Deutsches Institut fur Normung (DIN), andgastrointestinal (GI) test. These improved tests make it possible to predictthe bioavailability of heavy metal(loid)s in soil and sediments or wheningested by animals and humans (Bolan et al., 2008; Juhasz et al., 2009).The PBET and GI tests are in vitro screening-level assays used for predictingthe bioaccessibility of contaminants from a soil matrix. While the PBETmethod has been applied to both organic and inorganic contaminants, it ismore commonly recognized as an assay for assessing heavy metal(loid)bioaccessibility (Bolan et al., 2008). Assadian and Margez (2006) studiedthe bioaccessibility of heavy metal(loid)s (Cd, Cr, Ni, and Pb) using chemi-cal fractionation and in vitro GI and PBET methods in soils blended withuntreated effluent and biosolids. The results indicated that chemical frac-tionation of selected heavy metal(loid)s in soil did not reflect metal(loid)accumulation in oat forage or in sheep kidney, liver, or muscle tissue.However, PBET method was close to predicting Cd and Cr concentrationsmeasured in sheep tissues.


5.2. Bioassay

5.2.1. PhytoavailabilityIt can be expected that in wastewater-treated soils just like in sewage sludge-and manure-treated soils, plants may exhibit more tolerance to heavy metal(loid)s (Fig. 8; Table 9). Chang et al. (1992) and Logan et al. (1997), forexample, have demonstrated that when maize and other crops were grownon Cu-contaminated sludge-amended soils, inconsequential changes in planttissue Cu concentrations in response to substantial increases in total Culoading in soils occurred. Organic amendments may also alleviate oxyanionphytotoxicity; for example, the uptake of Se was less in the presence oforganic amendments (seleniferous plant tissues and manure) than that frominorganic sources (selenite) (Ajwa et al., 1998; Sharma et al., 2011). Singh et al.(2010b) examined the role of fertilizers (organic fertilizer as farmyard manure(FYM), commercial inorganic NPK, and a combination of FYMþNPK) inreducing the heavy metal(loid) availability in the soil, and subsequent uptakein Beta vulgaris L. (var. All green). They observed that phytoavailability of Cd,Cu, Pb, Zn, Mn, Ni, and Cr determined by bioconcentration factor (BF¼Plant concentration/soil concentration) was lowest in FYM and highest inNPK treated soil, compared to the untreated control (Fig. 8). They alsonoticed that the yield of B. vulgaris was also highest in FYM-treated soil andsuggested that application of FYM alone and in combination with NPK may

0

0.5

1

1.5

2

2.5

Cd Cu

Bio

conc

entr

atio

n fa

ctor

Pb Zn Mn Ni Cr

Control

Farmyard manure

NPK fertilizer

Farmyard manure+NPK

Figure 8 Effect of organic and inorganic fertilizers on heavy metal(loid) uptake in Betavulgaris L. grown in wastewater-irrigated soils (Singh et al., 2010b).

Table 9 Selected references on the effect of wastewater irrigation on heavy metal(loid) phytoavailability

Metal(loid)s Wastewater Plant species Observations References

Cu, Cd, Pb, Zn,

Fe, Mn

Treated municipal

wastewater

Barley Plant Cu, Zn, Fe, Mn increased with 2

years of wastewater irrigation, and

then reduced with longer period.

Plant Pb and Cd increased with longer

periods of irrigation.

Rusan et al.

(2007)

Al, Cr, Mn, Fe

Co, Ni, Cu,

Zn, Cd, Pb

Treated municipal

wastewater

Sunflower,

Sorghum

Sorghum accumulated higher

concentrations of Mn and Zn,

whereas sunflower accumulated

higher concentrations of Cr.

Ahmed and

Al-Hajri

(2009)

Zn, Cu, B, Mn,

Fe, Mo

Raw wastewater Cabbage It increased the metal content of cabbage

plants.

Kiziloglu

et al.

(2007)

Cu, Zn, Mn, Fe Raw wastewater Radish, spinach,

turnip, brinjal,

cauliflower, mint,

coriander, carrot,

lotus stem

Increased build-up of metals in plants,

high levels of Fe and Mn detected in

mint and spinach, whereas Cu and Zn

were highest in carrot.

Arora et al.

(2008)

Mn, Zn, Cu, Pb,

Ni, Cr, Cd

Municipal

wastewater

Leafy vegetable,

palak

Mn showed maximum uptake followed

by other metals.

Singh and

Agrawal

(2010)

Zn, Pb, Cr, Ni Water

contaminated by

industrial and

domestic effluent

Melilotus officinalis Pb, Cr, and Ni exceeded their permitted

limits in roots of plants.

Amiri et al.

(2008)

Cu Sewage sludge Fescue The effects of sewage sludge (SS) on Cu

in solution and plants depended on the

degree of weathering. In tailings with

a low degree of sulfide oxidation, SS

application resulted in increased

solubility and shoot accumulation of

Cu compared with NPK treated

tailings, probably due to the DOC

forming soluble complexes with Cu.

Forsberg

et al.

(2009)

Pb, Cd, As, Zn,

Hg

Fermented pig

slurry

Tomato All tomato samples were within the

legislation limits of tested metals.

Kourimska

et al.

(2009)


be considered as an easy and cost-effective technique for reducing the levels ofcontamination in food crops.

Addition of DOM to soils through wastewater irrigation and sludgeaddition can influence phytotoxic effectiveness of ions in at least twodifferent ways. On the soil side, an increase in DOM will shift metal(loid)partitioning toward the soil solution and hence increase the content ofsoluble metal(loid) in solution. On the solution side, although the solublemetal(loid) increases, the free metal(loid) ion is decreased due to DOMcomplexation. While wastewater can act as a sink to reduce the heavy metal(loid) uptake it can also act as a source of heavy metal(loid)s (Table 9).Although metal(loid)-DOM complexes are more mobile in soils, poten-tially leading to groundwater contamination, these complexes have beenshown to be less available for plant uptake, thereby alleviating phytotoxicitythat may otherwise result from excessive metal(loid) accumulation in soils(Ashworth and Alloway, 2007; Bolan et al., 2003c; Han et al., 2001).

In soils treated with wastewater and manure, only a small proportion ofmetal(loid)s dissolved in pore-water is likely to be available for plant uptake,the remainder is complexed with DOM (Huynh et al., 2008; Kunhikrishnanet al., 2011). Bolan et al. (2003c) studied Cu uptake in using mustard plantsamended with biosolids plus various levels of Cu (0–400mgkg�1 soil). Theyobserved that adding manure compost increased the adsorption and com-plexation of Cu in soil, noting a significant inverse relationship between theextent of Cu adsorption and DOM in the manure-amended samples. Thisindicated that DOM formed soluble complexes with Cu. They reportedthat although soluble DOM complexes were formed, addition of biosolidswas effective in reducing the phytotoxicity of Cu, especially at high levelsof Cu addition. In mustard plants amended with farm dairy and piggerywastewaters, Kunhikrishnan et al. (2011) observed an increase in Cu uptakewith increasing Cu input. However, at the same level of Cu application,plants took up less Cu from wastewater-amended soils than from Milli-Qwater amended soils. They concluded that the presence of DOM in thewastewater sources was effective in reducing the phytotoxicity of Cu at highlevels of Cu addition, indicating that the Cu-DOM complexes decreasedthe plant availability of Cu (Fig. 9a).

Qishlaqi et al. (2008) assessed the negative impacts of wastewater irriga-tion on soils and crops collected from two wastewater-irrigated sites anda reference site where bore water was irrigated. The results showed thatamong the five heavy metal(loid)s (Ni, Pb, Cd, Zn, and Cr) studied, usinguntreated wastewater caused contamination of spinach and lettuce with Cddue to its high phytoavailability in topsoil and excessive accumulation of Niand Pb in wheat. This scenario was due to the continual addition of heavymetal(loid)s through long-term wastewater application. They reported thataccumulation of metal(loid)s strongly depended on the crop’s physiologicalproperties (Liu et al., 2005) and the soil properties (Sharma et al., 2007).

0 200 400 600

Cu level (mg kg-1) Cu level (mg kg-1)

0

20

40

60

80

100T

issu

e C

u c

on

cen

tra

tio

n (

mg

kg

-1)

MQFDE

PE

0 400 800 12000

10

20

30

40

mg C

-CO

2 r

ele

ased

g-1

so

il h

-1 MQFDEWEPE

A B

Figure 9 Effect of Cu levels on (A) Cu concentration in plant tissue and (B) substrate-induced respiration in a silt loam soil in the presence of wastewater sources (PE, piggeryeffluent; WE, winery effluent; FDE, farm dairy effluent) and MQ water ((A):Kunhikrishnan et al., 2011; (B): Kunhikrishnan, 2011).


Rusan et al. (2007) noticed that Cu, Zn, Fe, Mn increased with 2years oftreated municipal wastewater irrigation in barley, and then declined withlonger irrigation periods. However, Pb and Cd continued to increase withlonger periods of irrigation. Several other studies report an increased uptakeof heavy metal(loid)s by plants due to continuous loading of metal(loid)s tosoil via irrigation (Abbas et al., 2007; Amiri et al., 2008; Kiziloglu et al., 2007;Moyo and Chimbira, 2009) (Table 9). Kiziloglu et al. (2008) comparedthe accumulation of heavy metal(loid)s (Fe, Cu, Mn, Zn, Pb, Ni, Cd) incauliflower and cabbage species irrigated with either untreated or primary-treated wastewater. Metal(loid)s in vegetables irrigated with untreated waste-water were higher than those irrigated with primary-treated wastewater.

5.2.2. Microbial and earthworm availabilityAs in the case of phytoavailability, the microbial availability of metal(loid)s islargely controlled by the activity of free ionic species in soil solution. Bolanet al. (2003a) observed the concentration of total Cu required to cause 50%reduction in basal respiration (microbial toxicity—MT50) was lower forCuSO4 (297mg Cu kg�1) than for the dairy pond sludge-Cu (783mg Cukg�1), inferring that sludge-borne Cu was less detrimental to microbialactivity than inorganic CuSO4. However, when the respiration value wasplotted against the concentration of free ionic Cu2þ, a single smooth curvewas obtained for both CuSO4 and sludge-Cu, and the MT50 value wasfound to be 18.37mgkg�1. This indicates that the difference in the effect on


respiration between the two Cu sources (i.e., organic vs. inorganic) is due tothe difference in bioavailable Cu content in soil.

Soil microbial activity as measured by respiration and microbial biomasscarbon was monitored by Kunhikrishnan (2011) following application ofvarious levels of Cu (0–1000mgkg�1), added as copper nitrate-spikedMilli-Q water, farm dairy, piggery, and winery wastewaters. The effect ofCu on soil microbial activity varied between Milli-Q water and wastewatersources and was attributed to the difference in the concentration of DOM.Metabolic quotient values were lower in soils in the presence of wastewaterthan in the Milli-Q water. The results indicated that wastewater sourcesdecreased the inhibitory effect of Cu on microbial activity and suggestedthat it could be attributed to the formation of Cu-DOM complexes (Fig. 9B).

Earthworms are also negatively influenced by the presence of heavy metal(loid)s in wastewater-irrigated soils. The bioaccumulation, however, dependson factors such as type and formofmetal(loid) and concentration (Heikens et al.,2001; Hobbelen et al., 2006; Nahmani et al., 2007; Spurgeon et al., 2006), soiltype and characteristics (Hendrickx et al., 2004; Hobbelen et al., 2006; Janssenet al., 1997; Kizilkaya, 2005; Spurgeon et al., 2006), test species (Heikens et al.,2001; Hendrickx et al., 2004; Nahmani et al., 2007), temperature (Olchawaet al., 2006), and exposure duration (Nahmani et al., 2007). Field observationshave demonstrated that Cu is detrimental to lumbricid earthworms (Niklas andKennel, 1978; van Rhee, 1975). This is supported by laboratory studies, whichshowed that the toxicity of Cu to earthworms is influenced by the pH andOMof the soil (Ma, 1984; Streit, 1984; Streit and Jaeggy, 1983). Aporrectodeatuberculata (Beyer et al., 1987) and Aporrectodea caliginosa (Peramaki et al., 1992)have been found to accumulate high Cd in acidic soils. This is largely related tothe fact that most of the heavy metal(loid)s that accumulate in an earthworm’sbody originate from pools of dissolvedmetal(loid)s which are bioavailable in thesoil pores of acidic soils (Herms and Brummer, 1984).

Kunhikrishnan (2011) examined the bioavailability of Cu to earthwormsin the presence of farm dairy, piggery, and winery wastewaters varying inDOM. Bioavailability of Cu to earthworms as measured by mortality andavoidance test was monitored at various levels of Cu (0–1000mgkg�1),added as copper nitrate-spiked Milli-Q water and wastewater sources. Theresults indicated that the wastewater sources decreased the inhibitory effectof Cu on the earthworm toxicity due to the formation of Cu-DOMcomplexes which are not readily available for uptake (Fig. 10). Metal(loid)concentrations of earthworms depended on CaCl2-extractable free Cu2þ

concentrations in the soil. Kunhikrishnan (2011) also observed that theearthworms clearly avoided soils with high levels of Cu concentrations.

Although DOM plays a protective role in reducing metal(loid) toxicityto earthworms, evidence suggests that earthworms play a humifying role inthe soil because humic acids were detected in earthworm-worked soil thatwere not present in the non-humified starting material (Businelli et al.,

0

10

20

30

40

50

60

70

MQ

Cu

conc

entr

atio

n in

ear

thw

orm

s (m

g kg

–1)

0 100 500 0 100 500 0 100 500 0 100 500

FDE WE PE

Figure 10 Effect of wastewater irrigation on Cu concentration in earthworms in a siltloam soil (PE, piggery effluent; WE, winery effluent; FDE, farm dairy effluent;Kunhikrishnan, 2011).


1984). Humic acids are known to transform the availability of metal(loid)sto plants by forming organometal(loid) complexes (Evangelou et al., 2004;Halim et al., 2003). Some authors suggest that increased uptake of heavymetal(loid)s by plants due to earthworm activity may be a direct resultof metal(loid)-chelating organic materials released by earthworms formingorganometal(loid) complexes (Currie et al., 2005; Udovic et al., 2007;Wanget al., 2006). A significant link has been found between the DOM increase insoils by earthworms Eisenia fetida and the concentration of water extractablemetal(loid)s (Zn,Cu,Cr, Cd,Co,Ni, and Pb) (Wen et al., 2004). An increasein DOM in soil has also been noted in one study reporting that the presenceof earthworms Metaphire guillelmi increased the availability of Cu to plants(Dandan et al., 2007).

6. Conclusions and Research Needs

Growing population, increased urbanization, improved living condi-tions, and economic development have led to a considerable increase in thevolume of wastewater generated by domestic, industrial, and commercialpractices (Asano et al., 2007; Lazarova and Bahri, 2005; Qadir et al., 2010).Although water quality management is a high priority and a major concernfor developing countries, most do not have sufficient resources to treatwastewater. Therefore, wastewater in a partially treated, diluted, or untreated


form is diverted and used by urban and peri-urban farmers to grow a range ofcrops (Ensink et al., 2002; Murtaza et al., 2010). Farmers consider wastewaterto be a reliable or sometimes the only water source available for irrigationthroughout the year and it often negates the need for fertilizer applicationsince it provides a source of nutrients. Similarly, the use of treated wastewaterfor both agricultural production and environmental protection has increasedin recent years in several continents including Australia, Europe, and NorthAmerica (Qadir et al., 2007; US EPA, 2004).

Like the supply of nutrients and OM through wastewater irrigation, italso contains different types and levels of undesirable constituents dependingon the source and level of its treatment. On the positive side, OM addedthrough wastewater improves soil structure, enhances charge characteristicsof irrigated soils, such as CEC, which may retain undesirable metal(loid)ions rendering them less available for plants, and acts as a storehouse ofessential nutrients for crop growth. On the negative side, heavy metal(loid)inputs to soils via wastewater irrigation are incommodious because, onceaccumulated, it is difficult to remove them. This situation may subsequentlylead to toxicity-related issues in plants grown on contaminated soils, posepotential harm to people and animals who may consume contaminatedcrops and they can be transported from soils to groundwater or surfacewater, thereby rendering the water hazardous for other uses (Murtaza et al.,2010). Most wastewater sources are rich in DOM which influences thebiological transformation processes of heavy metal(loid)s including theirmobility and bioavailability. The transport and bioavailability of heavymetal(loid)s can be strongly influenced by forming soluble and insolublecomplexes with DOM. Such interactions can alter the chemical speciationof the heavy metal(loid)s modifying their affinity for sorptive surfaces in thesoil matrix or their uptake, accumulation, and eventual toxicity to organ-isms (Arnold et al., 2010; Boyd et al., 2005). While the insoluble complexesare not available to plants and other soil organisms, the question ariseswhether the soluble heavy metal(loid) complexes in wastewater becomebioavailable or not. Anodic stripping voltammetry measurements and otherspeciation techniques have indicated that only a small percentage of the totaldissolved heavy metal(loid)s exist as free ions and the remainder appearsto be complexed with DOM. More research into this area is required tounravel the stability of such complexes, how they affect soil organisms andplants and long-term effects of application of DOM-enriched wastewater.

Given the current knowledge on the influence of wastewater in the (im)mobilization and bioavailability of metal(loid)s in contaminated soils, thefollowing research areas could be pursued:

� Long-term stability and biogeochemistry of metal(loid)s immobilized bywastewater sources.

� Influence of wastewater sources on rhizosphere biochemistry in relationto metal(loid) dynamics.


� Long-term leaching studies examining groundwater contaminationthrough the movement of a wide range of chemical pollutants in waste-water, especially in the case of untreated industrial effluents.

� Influence of wastewater on the redox reactions of metal(loid)s such as Cr,As, and Hg in relation to their speciation, mobility, and bioavailability.

� Variation in wastewater-derived DOM composition and concentrationcan have a diverse effect on metal(loid) speciation in soil. Therefore,characterization of DOM, employing molecular weight determination,and fractionation is necessary in order to understand the influence ofmetal(loid)-DOM complexation on bioavailability and toxicity of heavymetal(loid)s.

� Although wastewater-borne metal(loid)s are reported to be less toxic tosoil microorganisms, long-term studies are required to understand theirdynamics in soils.

� In addition to the accretion of salts and nutrients, under certain conditionswastewater irrigation has the potential to translocate pathogenic bacteriaand viruses to groundwater. It is therefore essential that other controloptions should be continued in parallel with ongoing efforts to identifykey wastewater pollutants and suitable techniques for their treatment.

Pragmatic approaches are required to protect water quality and ensurethat wastewater is used in a sustainable way. Risk assessment conductedprior to wastewater irrigation is highly recommended to enable the safe useof wastewater for landscape and agricultural irrigation. There are severalother opportunities for improving wastewater management through guide-lines and policies, which would reduce potential environment and publichealth risk. For instance, governments should implement an integratedwater management approach, promote public participation, disseminateexisting knowledge, generate new knowledge, and monitor and administerimposed standards.

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