The Influence of Wastewater Irrigation on the

Transcription

C H A P T E R
F I V E
The Influence of Wastewater
Irrigation on the Transformation
and Bioavailability of Heavy Metal
(Loid)s in Soil
Anitha Kunhikrishnan,*,†,} Nanthi S. Bolan,*,† Karin Müller,‡
Seth Laurenson,§ Ravi Naidu,*,† and Won-Il Kim}
Contents
1. Introduction
2. Sources of Wastewater and Heavy Metal(Loid)s in Soils
2.1. Wastewater production and quality
2.2. Heavy metal(loid) sources
3. Effects of Wastewater Irrigation on Soil Properties Affecting Heavy
Metal(Loid) Interactions
3.1. Soil chemistry
3.2. Soil biology
3.3. Soil physics
4. Effect of Wastewater Irrigation on Heavy Metal(Loid) Dynamics
in Soils
4.1. Adsorption
4.2. Complexation
4.3. Redox reactions
4.4. Methylation/demethylation
4.5. Leaching and runoff
5. Bioavailability of Wastewater-Borne Heavy Metal(Loid)s in Soils
5.1. Chemical extraction
5.2. Bioassay
6. Conclusions and Research Needs
References
216
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* Centre for Environmental Risk Assessment and Remediation, University of South Australia, Mawson Lakes,
Australia
Cooperative Research Centre for Contamination Assessment and Remediation of the Environment,
Adelaide, Australia
{
Systems Modelling, The NZ Institute for Plant and Food Research Ltd., Hamilton, New Zealand
}
Land and Environment, AgResearch Ltd, Invermay, New Zealand
}
Chemical Safety Division, Department of Agro-Food Safety, National Academy of Agricultural Science,
Suwon-si, Gyeonggi-do, Republic of Korea
# 2012 Elsevier Inc.
Advances in Agronomy, Volume 115
ISSN 0065-2113, DOI: 10.1016/B978-0-12-394276-0.00005-6
All rights reserved.
{
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 organisms, 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 for
various activities including irrigation have led to unsustainable rates of water
consumption, which has not been assisted by declining rainfall and increased
rationing of water to the ecosystem (Brown, 2007; Seckler et al., 1998).
Considerable pressure is now being placed on communities, particularly
primary producers, to improve water-use efficiency and use alternative
water supplies including recycled wastewater sources for irrigation, in a
much better way. Although using wastewater for irrigation raises concerns
about public exposure to pathogens and contamination of soil, surface water,
and groundwater, under controlled management these water sources can be
employed safely and profitably for irrigation (Plate 1) (Qadir et al., 2007).
Wastewaters originate from a number of sources including domestic
sewage (municipal wastewater), agricultural, urban and industrial effluents,
and stormwater. Wastewater irrigation has many beneficial effects, including
groundwater recharging (Asano and Cotruvo, 2004) and nutrient supply to
plants (Anderson, 2003). There are, however, some detrimental effects, such as
build-up of salts, pesticides, and heavy metal(loid)s. At sites irrigated with
Wastewater and Bioavailability of Heavy Metal(Loid)s in Soil
217
A
B
Plate 1 Recycled water irrigation in horticultural crops (A) Carrots, (B) Olives (Bolan
et al., 2011a).
wastewater, mobilization and transport of pesticides and heavy metal(loid)s
into groundwater have been noted, as well as their enhanced bioavailability to
soil biota and higher plants. For example, dissolved organic matter (DOM)
present in wastewater and sewage sludge has been shown to facilitate the
transport of both pesticides and heavy metal(loid)s (Ashworth and Alloway,
2004; Bolan et al., 2011a; Müller et al., 2007; Sedlak et al., 1997; Tam and
Wong, 1996; Thevenot et al., 2009). Wastewater irrigation and sludge application have also been shown to act as a source of heavy metal(loid) input to
soils (Barman et al., 2001; Eriksson and Donner, 2009; Murtaza et al., 2008).
The term “heavy metal(loid)” in general includes elements (both metals and
metalloids) with an atomic density greater than 6gcm3 [with the exception
218
of arsenic (As), boron (B), and selenium (Se)]. This group includes both
biologically essential [e.g., cobalt (Co), copper (Cu), chromium (Cr), manganese (Mn), and zinc (Zn)] and nonessential [e.g., cadmium (Cd), lead (Pb), and
mercury (Hg)] elements (Sparks, 2003). Heavy metal(loid)s reach the soil
environment through both pedogenic (or geogenic) and anthropogenic processes. Anthropogenic activities, primarily associated with the disposal of
industrial 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 pesticides
has been reviewed recently (Müller et al., 2007), no comprehensive review
has focused on its role in the mobilization, transport, and bioavailability of
heavy metal(loid)s in soil. This review aims to classify the different sources
of wastewater irrigation and heavy metal(loid) input to soil. It describes the
influence of wastewater irrigation on soil properties affecting heavy metal
(loid) interactions and explains the role of wastewater irrigation on heavy
metal(loid) dynamics including adsorption and complexation, redox reactions and bioavailability (Fig. 1). Whilst some literature reviews have
examined metal(loid) input through inorganic fertilizers, sewage sludge,
Wastewater
irrigation
M+
Metal(loid) sink
Adsorption
Complexation
Precipitation
Redox reaction
M+
M+ M+ M+
Treated sewage
Stormwater
Farm dairy effluent
Piggery effluent
Winery effluent
Changes in soil
properties
+
M+ M
M+
M+
Metal(loid) source
Plant uptake
Microorganisms
Earthworms
Volatilization/demethylation
Leaching
pH, EC, CEC
TOC, DOC
Sodicity and salinity
Aggregate stability
BD and total porosity
HC and infiltration
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 and
by altering soil properties.
219
and atmospheric deposition (Adriano, 2001; Smith, 2009), most reviews on
wastewater irrigation have focused on environmental issues of nutrients and
salt accumulation (Bolan et al., 2009; Bond 1998; Carpenter, 1998). There
has been no comprehensive review on the input of metal(loid)s via effluent
and wastewater, and the subsequent transformation and bioavailability of
effluent-borne metal(loid)s in soils. Unlike wastewater irrigation, a number
of studies have assessed the environmental implications of metal(loid)s derived
from sewage sludge and manure slurry application to soils (Cornu et al., 2001;
McBride, 2002; McGrath et al., 1994). Since sewage sludge is derived during
wastewater treatment and there are major resemblances in the composition
and chemical properties between these two resources, some of the information on the distribution and bioavailability of metal(loid)s is inferred from
sewage sludge and manure research. An improved knowledge of wastewater
irrigation’s long-term effects on metal(loid) dynamics in soils can enhance the
development of strategies to mitigate environmental impacts and maximize
the 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 including
domestic sewage, agricultural and industrial effluents, and stormwater.
Recycled water is defined as wastewater that is treated and reused to supplement water supply (US EPA, 1992). The beneficial utilization of treated
wastewater for agriculture is the major water reuse application worldwide
(US EPA, 2004). This water source can have the advantage of being a
constant, reliable water source and furthermore reduces the amount of
water extracted from the environment (Toze, 2006). Approximately 70%
of the world’s water resources including all the water from underground and
redirected from rivers is used for agricultural irrigation, so reusing treated
wastewater for agricultural and landscape irrigation (Plate 1) reduces both the
amount of water that has to be extracted from natural water sources and the
uncontrolled discharge of wastewater to the environment (Pedrero et al.,
2010). Thus, treated wastewater is a valuable water source for recycling and
reuse, especially in the Mediterranean countries and other arid and semi-arid
regions including Australia with increasing water shortages (Pedrero et al.,
2010). Table 1 summarizes the amount of wastewater generated and reused
annually in selected countries. For example, 88% and 70% of the recycled
water in Spain and Israel, respectively, is used for agricultural purposes
(Kanarek and Michail, 1996; Lallana et al., 2001).
220
Table 1
Wastewater generated and reused annually in selected countries
Country
Wastewater generated (GL) Wastewater reused (GL) % Reuse
Argentina
*
Australia
Bahrain
Bolivia
Chile
Greece
Egypt
*
India
Jordan
Kuwait
Libya
Mexico
New Zealand
Oman
Peru
Saudi Arabia
Spain
Syria
Tunisia
UAE
US
200.3
1634
45
135.8
295.6
–
10,012
13,870
82
119
546
13,340
67
78
34.7
730
24,094
825
240
881
–
90.7
262.9
8
–
–
0.7
200
1460
64.9
52
40
280
16
8.6
18.6
122.6
1100
550
33.8
185.3
2271
45.28
16.09
17.77
–
–
–
1.998
10.53
79.15
43.69
7.332
2.104
23.88
11.03
53.60
16.79
4.574
66.67
14.08
21.03
–
Source: FAO AQUASTAT Database, *Mekala et al. (2008).
By 2020, it is expected that 65% of the irrigation water used in Israeli
agriculture will be sewage effluents (Assouline et al., 2002). Other arid and
semi-arid countries, such as Jordan and Tunisia, reclaim the vast majority of
municipal wastewater for agricultural irrigation. Wastewater has been
recycled in agriculture for centuries as a means of disposal in cities such as
Berlin, London, Milan, and Paris (AATSE, 2004). In Pakistan, 26% of
national vegetable production is irrigated with wastewater (Ensink et al.,
2004). In Hanoi, 80% of vegetable production is derived from urban and
peri-urban areas that receive a secured supply of recycled water (Lai, 2000).
In Ghana, irrigation involving diluted wastewater from rivers and streams
has been reported (Keraita and Drechsel, 2004). In Mexico, about 260,000
ha are irrigated with wastewater (Mexico CAN, 2004). Agriculture, being
the largest user of recycled water in Australia, accounts for approximately
66% (280 GL) of all recycled water used (ABS, 2006).
In many countries, municipal wastewater is not collected and treated but
discharged directly into surface water bodies or used in agriculture without
appropriate consent. In most developing countries, 90% of all wastewater is
221
discharged untreated into local waterways (Johnston, 2003). In the rest of
the world, most of the wastewater is collected and treated to remove solids,
pathogens, oils, and other contaminants. Two main sets of regulations exist
for 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 developing
countries. Permissible water quality criteria stipulated are less restrictive than
those described in the California Health Laws, which recommend waste
stabilization pond systems as the preferred treatment method as opposed to a
conventional 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 removal
and settling out of coarse material to the bottom of the tank as primary
sludge. In some treatment plants, flocculants such as aluminum sulfate
(i.e., alum) are added; (ii) secondary treatment, which aims to remove
soluble and colloidal biodegradable organic matter (OM) and suspended
solids. Secondary treatment generally consists of an aerobic biological process whereby microorganisms oxidize OM in the wastewater; (iii) tertiary
treatment or advanced treatment technologies, these referring to any physical, chemical, or biological treatment process used to accomplish a degree of
treatment greater than that achieved by secondary treatment, such as ozonization, rapid gravity filtration, and ultraviolet radiation. The conditions in
which wastewater is stored following treatment may further influence its
chemical composition (Droste, 1997; Yu et al., 1997).
Specific composition of a waste stream is dependent on its origin and the
degree of treatment it receives. Heterogeneity in influent waste streams may
include domestic, industrial sources (paper and printing manufacturing,
timber processing plants, leather, and textile industries) and agricultural
sources (dairy, poultry, meat, and vegetable processing operations). Wastewater quality defines certain biological, chemical, and physical characteristics that influence its suitability for a specific use (Ayers and Westcot, 1985;
WHO, 2006). Nutrient loading (N, P, K, and S), organic loading, dissolved
constituents, such as dissolved salts and solids, types and concentrations of
microorganisms, and heavy metal(loid)s, trace organic compounds including pharmaceuticals, and pH are all quality criteria. Wastewater characterization is further complicated by daily and seasonal variation. There is a
twofold risk associated with applying wastewaters to agricultural crops
with respect to anthropogenic contaminants including metal(loid)s. First,
wastewater-borne metal(loid)s can be assimilated by plants and subsequently
enter the food chain. Second, application of wastewater can also impact on
heavy metal(loid)s that have been applied to soil and crops prior to the
wastewater irrigation. There is, however, limited information in the literature
on both issues. In Tables 2 and 3, important wastewater types for agricultural
irrigation and their main organic and inorganic components are summarized.
Table 2
Metal
(loid)s
Heavy metal(loid) concentrations in various wastewater and waste sludge sources
Treated
sewage
(mgL1)
0.035 0.012
Cd
0.002 0.0003
Pb
0.003 0.002
Ni
0.011 0.002
Cu
0.002 0.001
Zn
0.059 0.021
As
–
Hg
–
Reference Antanaitis and
Antanaitis
(2004)
Cr
a
Storm
water
Dairy
effluent
Piggery
effluent
Pulp and
paper
secondary
sludge
0.04
-
–
20
0.04
–
–
0.073–
1.78
0.053
–
0.022–
7.033
0.056–
0.929
0.058
3.22
Barrett
et al.
(1993)
Dairy
cattle
slurry
Beef cattle Poultry
slurry
litter
Broiler
litter
Swine
slurry
Deep-pit
poultry
litter
Threshold
values
LTVa
(mgL1)
–
5.64
4.69
–
9.9
2.82
6
0.1
4.5
–
0.33
0.26
3
4.93
0.3
2
0.01
–
42
–
5.87
7.07
11
-
2.48
13
2
–
–
35
–
5.4
6.4
15
2.46
10.4
14
0.2
0.5–10.5
0.26
206
16.5
62.3
33.2
748
6.1
351
19
0.2
–
0.58
513
6480
209
133
718
743
575
252
2
–
–
Bolan et al.
(2003a)
–
0.17
–
0.3
Lowe
Hart and
(1993)
Speir
(1992),
Carnus
(1994)
1.44
–
Wallingford
et al.
(1975)
2.6
–
Nicholson
et al.
(1999)
43
–
Nicholson
et al.
(1999)
34.6
–
Moore
et al.
(1998)
1.68
–
0.1
–
–
–
0.002
Nicholson
Nicholson Bomke
Jackson
et al.
and
et al.
and
(1999)
Lowe
(1999)
Miller
(1991)
(2000)
Feedlot
manure
(mgkg1)
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
TDS
Suspended
solids
BOD5
COD
Total N
Total P
Fat
Na
K
Ca
Mg
Free
Chlorine
Nitrate
Phosphate
Sulfide
Reference
Meat
Milk
processing Raw
factory
secondary meat
wastewater effluent
effluent
Tannery
secondary Dairy
effluent
effluent
Piggery
effluent
Textile
effluent
Pulp and
Preliminary- Primarypaper
Untreated treated
treated
secondary
wastewater wastewater wastewater sludges
Biosolids
–
––
–
20–100
–
1155
–
120
–
–
–
–
1480
471
1152
132
844.8
121
780
105
–
–
–
65
1700
–
70
35
400
560
13
8
1
–
20–100
80–400
40–200
5–30
0–30
50–250
20–150
3–250
3–10
–
646
1544
–
–
110
–
–
–
–
–
30
410
130
1.6
–
2700
–
340
36
–
–
–
190
30
–
50
220
110
30
–
–
–
1300
600
8.3
–
500
–
–
–
645
2430
–
–
–
–
–
1.24
1.04
1.14
–
–
1415
6.4
–
205
60
55
48
–
–
–
1260
5.10
–
193
52
51
–
–
–
–
1124
5.14
–
154
42
47
–
–
–
–
32,000
8075
–
4586
2905
17,000
2000
–
–
–
8.8
2.8
–
–
1.8
4.9
1.7
–
–
–
–
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)
7.97
2.63
0.58
Yusuff and
Sonibare
(2004)
–
–
–
Yusuff and
Sonibare
(2004)
–
–
–
Yusuff and
Sonibare
(2004)
–
–
–
Yusuff and
Sonibare
(2004)
–
–
–
Hart and
Speir
(1992),
Carnus
(1994)
420
–
–
Nash et al.
(2011)
Units are mgL1 except for pulp and paper sludges and biosolids (mgkg1).
224
2.1.1. Municipal wastewater
In both developed and developing countries, land application of municipal
wastewater (both treated and untreated) is a common practice. Municipal
wastewater is composed of domestic and industrial wastewater (Hussain et al.,
2002; Pettygrove and Asano, 1984). Domestic wastewater consists of discharges from households, institutions, and commercial buildings. Where
country or state legislation permits, this wastewater is applied to land. However, this depends on the crop it is applied to and the level of treatment.
Secondary-treated wastewater typically contains low levels of contaminants as
these tend to settle under gravitation with solid fractions in the treatment
lagoons. Settling of suspended solids also lowers both the chemical and
biochemical oxygen demand. Municipal wastewater also contains high concentrations of nutrients, especially nitrogen (N) and phosphorus (P), trace
elements, such as iron (Fe) and Mn and dissolved salts, particularly sodium
(Na), chloride (Cl), and in some cases bicarbonates. These parameters are
critical when wastewater is reused in agriculture.
2.1.2. Farm wastewater
Farm effluents such as those emanating from dairy sheds and piggeries are
being increasingly employed as sources of irrigation water and nutrients
(Bolan et al., 2009; McDonald, 2007). For example, in New Zealand, dairy
and piggery effluents generate annually about 9000Mg of N, 1250Mg of P
and 14,000Mg of K (Bolan et al., 2004a). Effluents from farms differ in their
composition depending on the animal production system from which they
are derived (chicken, pigs, beef, dairy). Generally, farm wastewater is rich in
organic and inorganic components (Tables 2 and 3) (Wang et al., 2004).
Copper and Zn are commonly used as feed additives, growth promoters, for
disease prevention or treatment, and their concentration in the final wastewater can be significant (Bolan et al., 2004b; Sims and Wolf, 1994).
In many countries including Australia and New Zealand, farm effluents
have traditionally been treated biologically using two-pond systems and
then discharged to land or stream. Bolan et al. (2009) have suggested that land
application of farm effluent facilitates the recycling of valuable nutrients,
carbon (C), and water, and if managed well, helps to mitigate surface water
pollution. In many instances, this may be the cheapest and most socially/
culturally accepted form of final treatment. Application of farm effluents can
increase pasture yield due to the net loading of nutrients and water (Bolan
et 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 and
piggery effluents directly to land has become the most common method of
treatment in most parts of the world.
However, in many regions, the amount of farm effluents generated on a
per farm basis exceeds the quantity that can be safely accommodated by the
225
available agricultural land and repeated annual applications of large amounts
of effluent can cause soil nutritional side effects and environmental damage
(Balota et al., 2010). For example, Giacomini et al. (2009) observed that half
of the piggery effluent N is lost from soil–plant systems through leaching
and volatilization, resulting in environmental pollution. Balota et al. (2010)
and McDonald (2006) indicated that safe farm effluent application in agriculture is necessary to minimize the environmental damage.
2.1.3. Effluents from the agricultural industry
Recycling of water from agricultural industry is another common source
of wastewater. For example, in Australia, agricultural drainage effluent is
collected and reused as a source of irrigation water (Dillon, 2000). Similarly,
wastewater from farm animal treatment plants (abattoirs) is increasingly used
as a source of irrigation water (Luo et al., 2004). Wastewater from intensive
agricultural industries (fish processing plants) and rural industry (abattoirs) is
characterized by high chemical and biological oxygen demand and nutrients
relative to many other wastewaters (McLaren and Smith, 1996) (Tables 2
and 3). Mittal (2004) recently reviewed effluent wastewater from abattoirs
and concluded that water quality depended on animal and processing
type and water usage, that is, dilution. One concern associated with land
application of abattoir waste is the high level of pathogens that have the
potential to contaminate receiving water bodies either directly as point
discharge or indirectly in runoff. These effluents also contain elevated levels
of grease, blood, and organic chemicals added during processing and cleaning operations (Kretzschmar, 1990). In many countries including Australia
and New Zealand, abattoir effluent is usually disposed of to land due to high
costs associated with independent treatment systems and environmental
concerns over surface water discharge (Quinn and Fabiansson, 2001).
Afonso and Bórquez (2003) reported that the wastewaters generated during
fish meal production contain a high organic load, but unlike other industrial
effluents they do not contain any known toxic or carcinogenic materials.
In some regions, winery wastewater is also a viable water source that is
increasingly being recycled by grape growers and pastoralists for irrigation
(Stevens, 2009). Reuse is driven primarily through the obligations of the
winery to dispose of their wastewater, preferably in a sustainable and costeffective manner. Generally, winery wastewater is treated on-site through
systems of only small flow volume capacity and the chemical composition of
this wastewater source varies considerably between wineries and may
require a management approach that is site specific when irrigated to land
(Kumar and Christens, 2009). In general, winery wastewater contains high
salt concentrations thereby accounting for a considerably greater salt loading
relative 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
226
that imposed by salinity alone. A high concentration of either Na or K in
irrigated waters is undesirable and when continually applied to soils can
displace more desirable cations [i.e., calcium (Ca) and magnesium (Mg)]
from the soil exchange complex (Pils et al., 2007). This in turn raises the
potential for adverse changes to soil structure (Jayawardane et al., 2011). The
management of salts is imperative so water conservation benefits are not
compromised 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 mills
Pulp and paper mill effluent either from thermomechanical pulp mill or
chemi-thermomechanical pulp mill is often irrigated to land after primary
treatment (Smith et al., 2003; Wang et al., 1999). Pulp mill effluent has high
chemical and biochemical oxygen demand and some wood derived organic
compounds, metal(loid)s, fatty and resin acids, and relatively high C:N
ratios (Tables 2 and 3). Kookana and Rogers (1995) reviewed the effect of
pulp mill effluents on soil properties. An extensive investigation into the
different organic chemicals in pulp mill effluents and their behavior in soil
can be found in this excellent review paper. Effluent from pulp mills is a rich
source of OM, N, P, Ca, Mg, and trace elements (Kannan and Oblisami,
1990), and consequently the application of pulp mill effluents on land is
becoming a common method for recycling nutrients (Rubilar et al., 2008).
The presence of chlorinated organic compounds, most notably chlorine
substituted phenolic compounds, chlorinated lignins, dioxins, chlorobenzenes, and non-chlorinated organic compounds (Deriziotis, 2004; Gergov
et al., 1988), has raised concerns about land application (Kannan and Oblisami,
1990; Lavric et al., 2004).
2.1.5. Stormwater
Urban stormwater harvesting has emerged in recent years as a viable option to
reduce pressures on existing water sources and to alleviate adverse environmental impacts associated with stormwater runoff (Roy et al., 2008a). This is a
relatively abundant, local source of water, available throughout most urban
areas. In Australia, for instance, approximately 10,300 million liters of stormwater are generated annually (Laurenson et al., 2010). In many cases, urban
stormwater runoff contains a broad range of pollutants that are transported to
natural water systems (Aryal et al., 2010). Stormwater pollutants originate
from many sources and activities and can occur as either particulate or
dissolved forms. Many toxic chemicals, such as pesticides and herbicides are
found in stormwater, along with oil, grease, and heavy metal(loid)s such as
Cd, Cr, Cu, Ni, Pb, and Zn (Wong et al., 2000). Nutrients such as N and P
are also important pollutants in stormwater.
The harvesting of stormwater from industrial zones prior to its entry into
natural waterways is likely to reduce the subsequent impact of point source
227
discharge on surface waters by reducing pollutant loads (Davis et al., 2009;
Henderson et al., 2007). If suitably designed, a stormwater harvesting system
will also provide urban stream health benefits by mitigating frequent flows to
streams 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 ways
including biofiltration, porous pavement, rain garden, and groundwater
recharge (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 naturally in soil parent materials, chiefly in forms that are not readily available for
plant uptake (Adriano, 2001; Alloway, 2004; Bolan et al., 2008). Due to their
low solubility, the heavy metal(loid)s present in the parent materials are often
not bioavailable and have a minimum impact on soil organisms. Apart from
Se (Dhillon and Dhillon, 1990; Doblin et al., 2006) and As (Chakraborty
and Saha, 1987; Mahimairaja et al., 2005; Mukherjee et al., 2008; Naidu and
Skinner, 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 soil
ecosystem. Unlike pedogenic inputs, heavy metal(loid)s added through
anthropogenic activities typically have high bioavailability (Adriano et al.,
2004; Lamb et al., 2009; Naidu et al., 1996). Anthropogenic activities primarily associated with industrial processes, manufacturing, and the disposal of
domestic and industrial waste materials including wastewater are the major
source of metal(loid) enrichment in soils (Adriano, 2001). Fertilizer, manure,
effluents, and organic amendments addition to agricultural soils are considered to be the major sources of most minor elements including heavy metal
(loid)s that are essential for plant growth (Bolan et al., 2004b; Loganathan
et al., 2008; Park et al., 2011).
2.2.1. Fertilizer products
Of the heavy metal(loid)s present in fertilizers, the presence of elevated
concentrations of Cd is of greatest concern as it is highly toxic to humans
and 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 and
New Zealand due to the extensive use of high Cd-containing phosphate
fertilizers (Loganathan et al., 2003; McLaughlin et al., 1996; Naidu et al.,
1997). Increased concentrations of Cd in fertilizers that are applied to land
have 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
228
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 in
soils by using low Cd-containing P fertilizers. This is achieved by either
selective use of phosphate rocks (PRs) with low Cd or treating the PRs
during processing to remove Cd. Superphosphate fertilizer manufacturers in
many countries are introducing voluntary controls on the Cd content of P
fertilizers. A number of PRs with low Cd contents are available which can
be used for the manufacture of P fertilizers, but sources with higher Cd
contents continue to be used in many countries for practical and economic
reasons (Bolan and Duraisamy, 2003; Loganathan et al., 1995, 2003).
2.2.2. Biosolids
Organic amendments such as biosolids (e.g., Cd) and poultry manure (e.g.,
As) have been regarded as a major source of heavy metal(loid) accumulation
in soils, and a large volume of work has been carried out to examine the
mobilization and bioavailability of heavy metal(loid)s derived from these
sources (Bolan et al., 2004b; Haynes et al., 2009; McBride, 1995). The most
commonly detectable heavy metal(loid)s in biosolids, Pb, Ni, Cd, Cr, Cu,
and Zn originate primarily from the contamination of these wastes with
industrial wastewater (Haynes et al., 2009). Heavy metal(loid) concentrations
are governed by the nature and the intensity of the industrial activity, as well
as the type of process employed during the biosolid treatment (Mattigod and
Page, 1983; Oviasogie and Ndiokwere, 2008; Wang et al., 2003a).
Gove et al. (2001) reported that biosolid application (250kgN ha1
1
yr ) to a sandy or sandy loam soil resulted in loadings of approximately
6, 2, 5, and 0.2mgkg1 Zn, Cu, Pb, and Ni, respectively. Illera et al. (2000)
demonstrated that biosolid application to soil had little effect on the total
concentration of Ni and Cr but resulted in Cd, Cu, Pb, and Zn increasing
considerably as a consequence of the high content of these metal(loid)s in
biosolids. It is known that these heavy metal(loid)s are typically immobilized
in soils, but they can be toxic to soil micro flora and can accumulate in plants
and grazing animals (Haynes et al., 2009). Kao et al. (2006) reported that
the addition of biosolid accumulated Cu, Pb, and Zn but reduced microbial
biomass, indicating that microbial activities were disrupted by the heavy
metal(loid)s.
2.2.3. Manure
Manures from intensive animal industries are a major source of organic
amendments for agricultural land. In Australia, beef and dairy cattle alone
produce approximately 4 million Mg of manure every year (Bünemann et al.,
2006). Similarly, in USA, of about 0.9 billion Mg organic and inorganic
agricultural recyclable by-products generated, approximately 45.4 million Mg
are dairy and beef cattle manure, and 27 million Mg are poultry and swine
229
manure (Camberato et al., 1997; Walker et al., 1997). As heavy metal(loid)s
are increasingly employed as a feed additive in intensive animal production
systems, manure application is likely to be an important source of certain metal
(loid) input to soils (Bolan et al., 2004b; Moscuzza and Fernández-Cirelli,
2009). Heavy metal(loid)s in manure by-products are also derived from
ingestion of contaminated soil by the animal, and during manure collection
and 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 animal
production systems is related to feed mineral content and the animal conversion efficiency (Nicholson et al., 1999). Increases in metal(loid) concentration
in animal feed have often resulted in corresponding increases in their concentration in manure by-products (Mohanna et al., 1999; Moscuzza and
Fernández-Cirelli, 2009; Nahm, 2002). A number of heavy metal(loid)s are
added to livestock and poultry feedstuff not only as essential nutrients but also
as supplement to improve health and feed efficiency. Diets of poultry and
livestock include heavy metal(loid)s (As, Co, Cu, Fe, Mn, Se, Zn) to prevent
diseases, improve weight gains, and increase egg production (Mondal et al.,
2007; Tufft and Nockels, 1991). Not all of the heavy metal(loid)s consumed
by animals are absorbed by their digestive tracts; consequently, the manure is
often metal(loid) enriched (Sistani and Novak, 2006). For example, swine can
excrete 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 coccidiosis in poultry has resulted in an increase in As level in poultry litter (Church
et al., 2010; Garbarino et al., 2003; Morrison, 1969; Sims and Wolf, 1994).
Regular application of manures and slurries has often been shown to
result in the accumulation of heavy metal(loid)s. Brink et al. (2003) reported
that application of swine effluent (annual mean of 10hacm) to Bermuda
grass pasture added annual averages of 0.6, 2.2, 0.3, and 0.86kgha1 Cu, Fe,
Mn, and Zn, respectively. Evers (2002) also reported application of 9Mg
ha1 broiler litter added averages of 5.85, 5.0, 9.4, and 6.55kgha1 Zn, Fe
and Cu, respectively. In another study, Jinadasa et al. (1997) reported that
high Cd levels in soils and vegetables throughout Sydney, Australia, were
due to repeated applications of poultry manures. In New Zealand, land
application of dairy pond effluent, based on a N loading of 150kgN ha1, is
likely to add a maximum of 31.5 and 73.7kg Cu ha1 through effluent and
manure sludge application, respectively (Bolan et al., 2003a).
2.2.4. Wastewater
Wastewaters 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 a
range of heavy metal(loid)s (Table 2) and continuous application is likely to
result in these heavy metal(loid)s accumulating in soils. Heavy metal(loid)s
are usually removed during common treatment processes and most of them
230
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 highly
treated wastewater, the concentration of heavy metal(loid)s is low and
considered safe when used for irrigation and other recreational purposes.
Consequently, heavy metal(loid)s are of little concern for irrigation of crops
when using treated effluents as a source of wastewater. Irrigation with
untreated or partially treated wastewaters has the potential to cause heavy
metal(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 an
industrial source or is less treated, then the effect of heavy metal(loid)s
would need to be inspected (Toze, 2006).
Wastewater also contains a range of components including dissolved
and particulate OM, soluble organic and inorganic anions which can interact with heavy metal(loid)s, thereby altering their mobility and subsequent
bioavailability. Wastewater often contains high levels of nutrients, which
can be beneficial to crop production (Liu et al., 2005) (Table 3). Bolan et al.
(2004c) have indicated that the application of dairy farm effluent irrigation
may provide an attractive means of increasing pasture growth through
increased nutrient loading and a cost-effective amelioration technique,
provided that the associated environmental risks of contamination of soils
and groundwater by heavy metal(loid)s are minimized.
Direct irrigation of untreated sewage effluents is a common practice
especially in Asian countries and many studies have reported the risks
associated with this system. Heavy metal(loid)s are easily accumulated
in 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 different
heavy metal(loid)s like Fe, Mn, Cu, and Zn, in vegetables irrigated with
water from different sources. The results indicated a substantial build-up of
heavy metal(loid)s in vegetables and the range was 116–378, 12–69, 5.2–
16.8, and 22–46mgkg1 for Fe, Mn, Cu, and Zn, respectively. They
reported that the highest mean levels of Fe and Mn were detected in mint
and spinach, whereas the levels of Cu and Zn were highest in carrot.
In another study, Latif et al. (2008) examined the heavy metal(loid) contamination of different water sources, soils, and vegetables and reported that
the concentrations of heavy metal(loid)s in sewage and industrial effluentsirrigated 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. Their
study demonstrated that the wastewater used for irrigation had high concentrations of Zn followed by Pb, Cr, Ni, Cu, and Cd and its continuous
application for more than 20years has led to accumulation of heavy metal
(loid)s in the soil. They observed that concentrations of Cd, Pb, and Ni
have crossed the safe limits for human consumption in all the vegetables.
231
They also noticed that the percentage contribution of fruit and vegetables
to daily human intake for Cu, Ni, Pb, and Cr was higher than that of leafy
vegetables, 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 influencing their reactions in soils and indirectly by altering soil properties
controlling their fate. Research shows that wastewater irrigation can result
in significant changes to soil physical, chemical, and biological properties
(Müller et al., 2007). In the following section, an overview on the impacts of
wastewater irrigation on soil properties relevant to the fate of metal(loid)s in
the soil environment is provided. The effect of wastewater irrigation on
metal(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 wastewater as a source of irrigation. On the one hand, the fertilization effect of
wastewater and the resultant independence from costly fertilizers has enhanced
the development of wastewater irrigation systems in many countries. On the
other hand, induced salinity and sodicity often limit its viability. Consequently, the impact of wastewater irrigation on soil chemical properties is
comparatively well investigated for a variety of soil types, climatic conditions,
and crops (Tables 4 and 5). A selection of soil chemical properties affected
by wastewater irrigation and relevant for the transformation, transport, and
complexation 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 pH
In the majority of studies, the soil pH significantly increased after long-term
irrigation with wastewater from different sources (Friedel et al., 2000; Hassanli
et 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, soil
pH was unaffected by long-term wastewater irrigation (Friedel et al., 2000;
Gwenzi and Munondo, 2008; Magesan et al., 1999), while others reported
decreased soil pH (Rattan et al., 2005; Rosabal et al., 2007; Xu et al., 2010).
For example, vinasse irrigation (pH 5.02) for 40years significantly decreased
the soil pH from 7.1 to 6.7 and from 6.2 to 5.9 in sampling depths of 0.1 and
1m, respectively (Rosabal et al., 2007). Angin et al. (2005) explained the
Table 4 Selected references on the effect of wastewater on soil properties
Site
Location
Soil type
Land-use
Wastewater (WW)
Experiment
Soil properties
Type (years)
(mm d1)
Depth
(m)
Bulk density
(gcm3)
TSS
SAR pH
(gm3)
Approach
Total
porosity
(%)
Ksat
(mm h1)
Suggested
mechanism
Different land
management of
control and
treated area
References
Vogeler
(2009)
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)
Levin, New
Zealand
Sand
Pasture
185
n.d.
6.2
0–0.05
Disc
infiltrometer
1.1* (1.2)
58 (54)
34* (13)*
Levin, New
Zealand
Sand
Pinus radiata
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
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
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
170
5.1
7.6
0–0.2
Intact core
1.1
(0.9)
n.d.
0.02
(0.06)
Physical blocking of Bhardwaj et al.
(2007)
pores, clay
swelling
Rotorua, New
Zealand
Sandy
loam
Pinus radiata
WW
22
9
Secondarytreated
WW
7
8
Tertiarytreated
WW
5
8
WW
0yr
2yr
5yr
15yr
n.s.
WW (drip
irrigation)
23
15
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
Vogeler
(2009)
Cook et al.
(1994)
Levin, New
Zealand
Sand
Pinus radiata
Primarytreated
WW
4
8
185
n.d.
6.2
0–0.1
Intact cores
0.8* (1.2)
70* (54)
185 (159)
Rotorua, New
Zealand
Sandy
loam
Pinus radiata
Tertiarytreated
WW
4
8
6
n.d.
7.2
0–0.1
Intact cores
0.7 (0.6)
72 (72)
114 (39)
Farm
WW
10
n.s.
WW
15
n.s.
WW
44
7
712
2.0
7.3
0–0.1b
Intact cores
1.2 (1.3)
38 (35)
135 (91)
712
2.0
7.3
0–0.1b
Intact cores
1.1 (1.3)
55 (35)
163 (91)
5
n.d.
7
0–0.2b
Landscape
approach
1.5*
n.d.
38* (116)
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)
Sandy
Madurai
Corporation,
loam
India
Sandy
Madurai
loam
Corporation,
India
Diverse:
Pennsylvania
silt
State
loam–
University,
silty
USA
clay
loam
Isfahan, Iran
Aridisol,
silty
clay
a
b
Farm
Farm
Sugar beet,
Secondarycorn,
treated
sunflower
WW
Increased
macroporosity
due to
stimulation of
microbial
communities
Increased
macroporosity
due to
stimulation of
microbial
communities
Input of OM
improved
structure
Input of OM
improved
structure
Excessive water,
impact of water
drops and
machinery, soil
transport across
landscape
Biological and
physical
clogging
Ponded infiltration rate.
Measured in increments for more than 1m depth in original paper. Values within brackets are from the control site (before irrigation of wastewater).
Magesan
(2001)
Magesan
(2001)
Mathan
(1994)
Mathan
(1994)
Walker and
Lin (2008)
Abedi-Koupai
et al.
(2006)
Table 5
Selected references on the effect of wastewater on soil properties (pH, SOC, and C-input)
Location
Soil type
Shiraz, Iran
–
Marvdasht,
Iran
Silty clay
Silty loam
Silty loam
Mezquital
Valley,
Mexico
Silt loam,
Mollic
Leptosol
Mezquital
Valley,
Mexico
Clay loam,
Eutric
Vertisol
Amman,
Jordan
Clay silt,
Vertisol
Wastewater Type
(years)
Untreated domestic
WW, (20yr
estimated)
Treated municipal
effluent, 25months
(39 ML ha1 yr1)
WW (11 ML ha1 yr1)
0
25
65
80
WW (11 ML ha1 yr1)
0
25
65
80
Municipal WW
0yr
2yr
5yr
15yr
Land-use
Depth (m)
pH
SOC (%)
C-input
(tha1 yr1)
Wheat
0–0.2
8.4 (7.3)
15.9 (0.4)
n.d.
Qishlaqi et al.
(2008)
Trees
0–0.3
0.3–0.6
0.6–0.9
Hassanli et al.
(2008)
0–0.15
0.0003
(<dL)
0.0006
(<dL)
0.0006
(<dL)
1.8
2.2
1.9
2.3
n.d.
Maize
8.8
(8.0)
8.8
(8.2)
8.8
(8.1)
7.5
7.9
7.6
7.7.
1.5
Friedel et al. (2000)
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)
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
References
Gharaibeh et al.
(2007)
7.4 (7.1)
4.5 (4.5)
2.2
Magesan et al.
(1999)
2.3
Magesan et al.
(1999)
n.d.
Rosabal et al. (2007)
5.8
(5.6)
2.7 (2.4)
0.9
(0.4)
2.9
(2.0)
2.7
(1.6)
2.3
(0.9)
1.2
(0.5)
0.4
(0.2)
0.2
(0.3)
þ164%
þ109%
þ118%
2.2
(0.6)
2.3
Filip et al. (2000)
0–0.15
7.5 (7.9)
0.65 (0.39)
n.d.
Rattan et al. (2005)
0–0.25
7.5 (7.2)
1.3 (0.6)
OM
n.d.
Lado et al. (2005)
0–0.25
7.8 (7.5)
3.9 (3.4)
OM
n.d.
Lado et al. (2005)
Sand; Typic
Udivitrand
WW; 5yr
(29 ML ha1 yr1)
Forest
plantation
0–0.1
0.1–0.2
Sand,
Psamment
WW, 7yr
(31 ML ha1 yr1)
Forest
plantation
0–0.1
0.1–0.2
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)
Erzurum,
turkey
n.s.
Raw WW; long-term
Cabbage,
potato
Decrease
Berlin,
Germany
Haplic Luvisol
Pasture
Delhi, India
RamatHacovech,
Israel
Mizra, Israel
Loamy sand;
sandy loam
Sand
Primary treated WW;
100yr
(20 ML ha1 yr1)
Sewage effluents
5, 10, 20yr
Effluents, 10yr
0–0.3
0.3–0.6
0.6–0.9
Topsoil
Rice, grain,
vegetables
Citrus
Clay
Effluents, 12yr
Citrus
Harare,
Zimbabwe
Loamy sand
Gleyic Lixisol
Rotorua,
New
Zealand
Levin, New
Zealand
Pasture
6.9 (5.9)
6.7
(6.0)
n.d.
n.d.
Angin et al. (2005)
n.d.
(Continued)
Table 5
(Continued)
Location
Soil type
Isfahan, Iran
Silty clay
Ebro Valley,
Spain
Xeric
Petrocalcid
Typic
Xerofluvent
Wastewater Type
(years)
Land-use
Treated WW
26yr
(15–37 ML ha1 yr1)
Municipal WW
(150, 300, 600t ha1
yr1)
Vegetable canning WW;
4yr
Annual
rotation
Values within brackets are from the control site (before irrigation of wastewater).
Depth (m)
pH
SOC (%)
0–0.3
0.3–0.6
0.6–0.9
0–0.3
150t ha1
300t ha1
600t ha1
0.3–0.6
150t ha1
300t ha1
600t ha1
0–0.3
5.0 (3.9)
5.0 (3.9)
5.4 (3.8)
(7.6)
7.4
7.3
7.2
(6.9)
6.6
6.5
6.3
n.d.
n.d.
17.2 (7.1)
4.3 (3.8)
3.3 (2.9)
(0.6)
0.9
0.9
1.1
(0.6)
0.7
0.8
0.8
1.2
(1.2)
0.8
(1.0)
C-input
(tha1 yr1)
References
58
116
173
Gwenzi and
Munondo
(2008)
Khoshgoftarmanesh
and Kalbasi
(2002)
0.06
0.26
Virto et al. (2006)
237
lower soil pH of a long-term wastewater-irrigated soil by the increased
mineralization of OM, while Xu et al. (2010) ascribed the effect to the applied
wastewater’s acidity. In general, the effect of wastewater irrigation on soil
pH depends on the pH of the wastewater source and the pH buffering
capacity of soil.
pH is an important factor that controls the accumulation, mobility, and
bioavailability of heavy metal(loid)s in wastewater-irrigated soils. The pH is
often 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 soils
and crops and reported that pH increased by 2–3 units and heavy metal
(loid)s (notably Pb and Ni) accumulated in topsoil above maximum permissible limits. Although they found a positive relationship (P<0.01)
between pH and total contents of Pb and Ni in soils, they reported only
1.3–7.7% 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 with
paper mill wastewater was observed, the data did not show a significant
positive correlation with the metal(loid) ions.
3.1.2. Soil organic matter
Depending 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 adds
more OM to a soil than freshwater irrigation or rain. Hence, various studies
conducted in long-term farm dairy/sewage effluent irrigated areas reported
significantly increased SOM contents in the topsoil (Barkle et al., 2000;
Bhandral et al., 2007; Filip et al., 2000; Friedel et al., 2000; Gwenzi and
Munondo, 2008; Marecos do Monte, 1998; Qishlaqi et al., 2008; Rattan
et 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 also
been attributed to indirect positive effect on biomass production through
the nutritional benefit of wastewater irrigation leading to more residues in
the soil (Ramı́rez-Fuentes et al., 2002). However, our attempts to correlate
the C inputs associated with wastewater irrigation with the observed net
C-increases failed because of: (i) the frequent lack of information on the
C-input through wastewater application; and (ii) the difference in soil types,
irrigation periods, effluent characteristics, and climate of the studies considered (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 with
graywater. Falkiner and Smith (1997) even observed that total C in the top
0.1m of a sandy loam under forest plantation after 4years of irrigation with
secondary-treated municipal wastewater was significantly reduced. This
SOM loss was explained by accelerated decomposition rates caused by the
238
frequent wetting and drying cycles at the site. In three Luvisols and three
Vertisols, long-term municipal wastewater irrigation led to SOM losses in
1m depth ranging between 0.6 and 4.9Mgha1 compared with freshwaterirrigated soils, while the impacts in the topsoils were inconsistent (Jueschke
et al., 2008). Jueschke et al. (2008) postulated that the observed losses
resulted from increased microbial activity stimulated through the application of easily degradable and complex organic substances with wastewater.
In another study, the SOM content in 0–0.75m depth of a dairy factory
effluent-irrigated allophanic soil was the same as in the non-irrigated control
site after 22years, but a redistribution of SOM from the top 0.1m down to
0.50m depth was observed. This was partly attributed to leaching and the
modified 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 of
long-term wastewater irrigation (90years) on the quality of SOM in two soil
types, Leptosols and Vertisols in the Mezquital valley in Mexico. They
found that the quality of SOM (i.e., differences in carbon mineralization)
had changed and carbon mineralization in the irrigated soils significantly
increased. Consequently, the DOM concentrations in the irrigated soils also
increased. This effect was more pronounced in the Leptosols than the
Vertisols, 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-irrigated
sites have often been observed and explained by the direct input of wastewaterborne DOM and the indirect solubilization of SOM resulting from increased
pH (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 irrigation
to a pasture soil, which varied with the type of effluent (Fig. 2). In another
study, the DOM concentrations in wastewater-irrigated soils not only
increased significantly but the aromaticity of the DOM in soil solutions
decreased at the same time (Jueschke et al., 2008). Increased DOM concentrations in soil solutions may affect soil physical properties, such as soil
aggregate stability and water binding potential (Frenkel et al., 1992). It also
provides organic substrate for soil microorganisms and mobile sorbents to
the system.
OM content is also one of the most important factors that control the
accumulation, mobility, and bioavailability of heavy metal(loid)s in wastewater-irrigated soils. Increase in SOM content can lead to increased soil
adsorption capacity by which accumulation of heavy metal(loid)s will be
enhanced. Qishlaqi and Moore (2007) carried out statistical analysis of the
sources and accumulation of heavy metal(loid)s in agricultural soils and
noticed that SOM was the most important factor controlling the distribution of heavy metal(loid)s. It was revealed that soil samples with high SOM
content accumulated significantly higher concentrations of heavy metal
239
Concentration (mg kg–1 soil)
175
TFDE
TPFE
Water
150
UFDE
TME
Control
125
100
75
50
0
10
20
30
40
Days after treatment application
50
60
Figure 2 Soil DOM concentration at 10cm depth following the winter application of
water and a range of effluents types to sheep grazed pasture (TFDE, treated farm dairy
effluent; 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 as
a result of wastewater irrigation controls the mobility and bioavailability of
metal(loid)s (Bolan et al., 2011b; Jackson et al., 2006; Khan et al., 2006).
3.1.3. Cation exchange capacity
A long-term increase in SOM content resulting from wastewater irrigation,
which is sometimes accompanied by an increased soil pH, can result in
an increase of the CEC (Angin et al., 2005; Falkiner and Smith, 1997). This
has been observed, for example, in a 5-year study conducted in Portugal
comparing the impact of potable water, primary effluent, and secondary
effluent on various chemical parameters including CEC (Marecos do
Monte, 1998). Qishlaqi et al. (2008) reported that the CEC of a sandy
topsoil that has been irrigated with raw wastewater for about 20years
increased by about 880%. Others, however, did not observe a significant
increase in CEC in spite of significantly increased SOM contents through
wastewater irrigation (Gharaibeh et al., 2007). Madyiwa et al. (2002) studied
the effects of combined sewage sludge and effluent application on soil
properties of a sandy soil under pasture. The relatively high metal(loid)
(Cu, Ni, Pb, and Zn) concentrations within the top 10cm compared to the
lower horizons in the irrigated area confirmed the immobility of most heavy
metal(loid)s. They argued that considering the lower clay content in top
20cm, the high CEC resulting from high OM content of these layers
attributed to metal(loid) immobilization. They confirmed that the four
240
metal(loid)s in their study were strongly correlated to CEC (R2 ¼0.94–0.99)
and OM (R2 ¼0.88–0.99) in the sewage effluent irrigated soil.
3.1.4. Salinity
Salinity is the most restricting factor for using wastewater as an irrigation
source, especially in Australia’s arid climate conditions. It refers to the total
concentration of all salts in the irrigation water or soil solution and is
determined by measuring the electrical conductivity (EC) and/or the total
dissolved solid (TDS) content in the water. Long-term wastewater irrigation adds large amounts of salts to a soil system (e.g., Bond, 1998; Falkiner
and Smith, 1997; Gharaibeh et al., 2007; Gwenzi and Munondo, 2008;
Menneer et al., 2001; Xu et al., 2010) as typical TDS concentrations in raw
municipal sewage and tertiary-treated wastewater range from 200 to 3000
mgL1 (Feigin et al., 1991). Rana et al. (2010) indicated that long-term
addition of sewage water to agricultural lands enhanced EC values from
0.99 dS m1 for well irrigated to 1.65 dS m1 for sewage irrigated soil.
Salinity is likely to be at a minimum immediately after an irrigation event
when the soil water content is maximal. Water removal through evapotranspiration can lead to salt accumulation in the topsoil (increased salinity),
which may harm the crop depending on its salt tolerance (Mass and
Hoffman, 1977). Salts in wastewater can also reduce water availability to
the crop by changing the osmotic potential between plant and soil to the
extent that the plant’s water, nutrient, and metal(loid) uptake and yield are
affected. The annual variation of the water balance has to be taken into
account when designing wastewater irrigation systems, and sufficient leaching for removal of excessive salts from the root zone has to be warranted
(Smith et al., 1996). In addition to decreasing plant available water, salinity
can also impact on soil structure through flocculation/deflocculation processes (Shainberg and Letey, 1984).
McLaughlin et al. (1994) studied the causes of elevated Cd concentrations in field-grown potato tubers. They noticed that tuber Cd
concentrations were positively related to soil EC and extractable Cl (R2 ¼
0.62, P<0.001) in the topsoil, with extractable Cl accounting for more
variation than EC. They observed that the tuber Cd was unrelated to tuber
concentrations of P or sulfur (S) but was positively related to concentrations
of Na. They concluded that the cause of elevated Cd concentrations in
tubers was due to the effect of Cl mobilizing Cd within the soil and
increasing the availability to plants irrigated with saline waters.
3.1.5. Sodicity
Municipal wastewater, farm effluents, and effluents from agricultural industries usually have high Na concentrations (e.g., secondary municipal wastewater, 50–250mgL1; Feigin et al., 1991). Sodium has the opposite effect on
soils to that of elevated salt concentrations (Sumner, 1993). While high levels
241
of salinity promote flocculation, elevated Na levels enhance clay swelling,
clay dispersion, and aggregate slaking. Clay dispersion can lead to structural
breakdown of a soil and can have adverse effects on soil physical properties,
such as soil porosity and permeability (Bond, 1998). The sodicity of irrigation
water can be quantified with the sodium adsorption ratio (SAR), which is the
level of Na relative to other cations in the irrigation water:
ðNaþ Þ
SAR ¼ qffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi
2þ
ðCa ÞþðMg2þ Þ
2
ð1Þ
with concentrations of Na, Ca, and Mg in meqL1. The SAR of wastewater can vary considerably, often within the range of 4.5–7.9 for secondary
municipal wastewater (Arienzo et al., 2009; Feigin et al., 1991; Lado and
Ben-Hur, 2009). The threshold level of SAR in relation to dispersion varies
between soil types (Sumner, 1995). A measure of a soil’s sodicity is the
exchangeable sodium percentage (ESP), the prevalence of exchangeable Na
compared to other exchangeable cations, mainly Ca, Mg, K, hydrogen (H),
and Aluminum (Al):
ESP % ¼
100ðexchangeable NaÞ
CEC
ð2Þ
where the Na concentration and CEC are in cmolc kg1. Many studies
demonstrated a positive correlation between a soil’s ESP and the SAR of the
irrigation water (Harron et al., 1983; Jayawardane et al., 2011; Rengasamy and
Marchuk, 2011). In Australia, soils that have more than 6% ESP are considered
to have structural stability problems (Sumner, 1995). This threshold is 15% ESP
under American conditions due principally to the differences in clay mineralogy (Halliwell et al., 2001), indicating that the thresholds are not absolute
figures. The impact of increased ESP on soil physical properties is very complex
and dependent on many other factors, such as clay content and mineralogy, EC
of the soil solution, SOM, and DOM content, pH, and thus, cannot be readily
predicted (Sumner, 1993). The opposing effects of salinity and sodicity of
irrigation water on soil dispersion mean that while the likelihood of clay
dispersion increases with high SAR-values, this may be mitigated by the
increased flocculation due to high salt concentrations, an increased EC.
Increases in EC and SAR in soil solutions have been observed with
different 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 study
in Iran that the soil SAR decreased significantly under effluent irrigation
compared with borehole water irrigation. Curiously, the quality of the
borehole water was often found to be inferior to the effluent quality (SAR,
242
15 and 8, respectively). Stewart et al. (1990) reported an increase in ESP from
3.2% to 9.8% in 0.15–0.35m depth after effluents with an SAR of 5.4 have
been irrigated to a plantation for 4years. Amongst several treatments, Falkiner
and Smith (1997) observed a maximum increase in ESP from <2% to 25% in
0.3–0.4m depth under a plantation after 4years of weekly secondary-treated
effluent irrigation (SAR of effluent, 4.8). Menneer et al. (2001) irrigated two
different soil types with sodium-rich dairy factory wastewater over a period
of 5 years and reported an increased ESP of 31% compared to 0.4% at the
soil surface of unirrigated soils. A significant increase was also measured in an
Iranian trial after 1-year irrigation of municipal waste leachate
(Khoshgoftarmanesh and Kalbasi, 2002).
The interrelationship between salinity and sodicity affects soil structure
and thus, transport of heavy metal(loid)s. For example, Usman et al. (2005)
investigated the effect of immobilizing substances (three clay minerals, iron
oxides, and phosphate fertilizers) and NaCl salinity on the availability of
heavy metal(loid)s Zn, Cd, Cu, Ni, and Pb to wheat grown in sewage
sludge-amended soil. The plants were irrigated either with deionized or
saline water containing 1600mgL1 NaCl. They reported that the addition
of metal(loid) immobilizing substances—specifically bentonite clay—
significantly decreased metal(loid) availability to wheat. They noticed that
irrigation with saline water resulted in a significant increase in metal(loid)
chloride species (MClþ and MCl02) with the highest concentration observed
for Cd, which was about 53% of its total soil solution concentration. They
concluded that saline water increased the availability of Cd and Pb to wheat
and decreased the efficiency of bentonite to immobilize soluble Cd.
3.2. Soil biology
Soil biological properties as affected by wastewater application have been
investigated with variable results, depending on the experimental design
and measurements monitored. For example, traditionally microbiological
counts have been reported, whereas in more recent studies molecular
biological approaches concentrating on gene expression and enzymatic
activities, are employed. Controversial results can also be explained by the
different nature of wastewater from different sources and the length of
wastewater irrigation. For example, while wastewater irrigation is generally
considered as a stimulant of microbial activities, long-term wastewater
irrigation can lead to the accumulation of metal(loid)s, salts, and organic
compounds such as pesticides in soils which might be toxic to soil fauna and
flora (Müller et al., 2007). Antibiotics are bioactive compounds and can
reach soils through wastewater irrigation, thereby affecting soil biological
activity (Kinney et al., 2006). Moreover, wastewater-borne microorganisms
might compete with indigenous microbial communities (Sidhu et al., 2001),
thereby affecting the biotransformation of metal(loid)s.
243
3.2.1. Microbial communities
A field trial using tertiary-treated domestic wastewater in a Pinus radiata
forest on allophanic soils showed no significant differences in microbial
biomass, basal respiration, and sulfatase activity relative to mains water
irrigation possibly due to the low nutrient and carbon contents of the
wastewater (Schipper et al., 1996). However, many other studies reported
a positive impact of long-term wastewater irrigation on total microbial
biomass 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; Friedel
et al., 2000; Goyal et al., 1995; Monnett et al., 1995; Ramı́rez-Fuentes et al.,
2002). This phenomenon was ascribed to: first, the enrichment of the soils
with microbial available carbon and nutrient sources stimulating the soil
microbial populations; and second, to favorable pH and moisture conditions
(Filip et al., 2000). Shapir et al. (2000) reported that wastewater irrigation
mainly affected the soil microbial communities of the topsoil layers of a
sandy soil. They emphasized that the observed increase in bacterial counts
did not always correlate with similar changes in bacterial activity.
Blume and Horn (1982) reported a shift in the microbiological population from aerobic to anaerobic microorganisms due to short-term oxygen
depletion of the topsoil resulting from wastewater irrigation, as seen by a
decrease in oxygen diffusion rate (Fig. 3; Bhandral et al., 2007). They also
noticed a higher proportion of nitrifying and ammonifying microorganisms
than in the control. The stimulation of copiotrophic bacteria was observed
in the same long-term wastewater irrigation area (Filip et al., 1999). Similarly, increased denitrification rates under wastewater irrigation were
Concentration (mg cm–2 min–1)
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
TFDE
UFDE
TPFE
TME
Water
Control
Figure 3 Oxygen diffusion rate (ODR) values (mgcm2 min1) from the 10cm soil
depth following the winter application of water and a range of effluent types to sheep
grazed pasture (Bhandral et al., 2007).
244
reported and explained by higher substrate availability (Schipper et al., 1996)
or denitrification enhancing surfactants (Friedel et al., 2000). Adenylate
energy charge ratios were reduced and attributed to the addition of Na and
salts with the wastewater irrigation (Friedel et al., 2000). Ramı́rez-Fuentes
et al. (2002) assumed that the observed inhibition of N2O-oxidation in a
long-term wastewater-irrigated soil was due to the accumulation of salts,
heavy metal(loid)s, and toxic organic compounds. Others also observed
changes in the structure and function of microbial communities due to
wastewater irrigation (Faryal et al., 2007; Oved et al., 2001). These examples
show that changed environmental conditions in the topsoil of long-term
wastewater-irrigated sites can have a selective impact on the composition
of microbial populations and soil functional diversity. It is noteworthy that
the increase in microbial biomass in a long-term wastewater irrigation area
in Berlin, Germany, remained detectable 20years after wastewater irrigation
ceased (Filip et al., 1999).
Additions of heavy metal(loid) salts to soils usually cause an immediate
decrease in respiration rates, but long-term responses are determined by the
properties 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 soil
respiration rates in clay soils but may decrease respiration rates in sandy
soils that may be attributed to the difference in bioavailability of Pb between
soil types. It has been reported that a neutral soil may contain high levels of
Mn, Al, or Pb without any sign of toxicity to microorganisms whereas
toxicity 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 example
Cu, Ni, and Zn, are essential trace elements for plants and microorganisms
(Alloway, 1995). Even these trace elements, however, may become toxic at
higher concentrations (Kosolapov et al., 2004). Copper at high concentration has a detrimental effect on soil microorganisms and modification to the
population structure of microbial communities has been reported (Ranjard
et al., 2006; Tom-Petersen et al., 2003).
DOM is considered the most dynamic C fraction in soils and it represents a major source of energy and cellular C for the soil microbial community. Therefore, a close relationship exists between DOM and soil microbial
activity and this C fraction contributes substantially to the total CO2 flux
from soils (van Hees et al., 2005). Liu and Haynes (2010) investigated the
microbial activity of soils that had received dairy factory wastewater irrigation for greater than 60years. Soil organic C content was unaffected by
irrigation but the size (microbial biomass C and N) and activity (basal
respiration) of the soil microbial community were increased. They concluded 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
245
which has been irrigated for 15years with lagooned municipal wastewater.
They noticed that MBC, soluble C, cumulative respiration, and enzymatic
activity were significantly higher in the soils irrigated with wastewater than
soils irrigated with “clear” water; they also found that the metabolic quotient (qCO2) was significantly lower in wastewater-irrigated soil, indicating
that the microbial biomass used the energy sources more efficiently.
3.2.2. Earthworms
In a field trial, about 500mm dairy shed effluent applied during 270days to a
silt 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 soil
leading to higher dry matter production and lower earthworm mortality
during summer as a result of desiccation. The irrigation with dairy factory
effluents for 22years to pasture on an allophanic soil led to lower earthworm
numbers but a higher biomass of earthworms than in the control, which was
accompanied by a modified abundance of the five species present (Degens
et al., 2000). Increased earthworm numbers were also recorded by Yeates
(1995) in a 7-year experiment of sewage application to a 17-year-old
P. radiata plantation on a sandy soil. However, Blume and Horn (1982)
warned that high wastewater irrigation rates have a detrimental impact on
earthworms due to anaerobic soil conditions as observed under long-term
wastewater irrigation “Rieselfelder” around Berlin.
3.3. Soil physics
Long-term wastewater irrigation can affect soil physical and hydraulic
properties (Daniel and Bouma, 1974; Jnad et al., 2000; Lado and BenHur, 2010; Mathan, 1994; Vinten et al., 1983a,b; Vogeler, 2009). Changes
in soil physical properties are primarily caused by the impact of wastewater
irrigation on soil chemical properties including soil pH, SOM content and
quality, salinity, and sodicity. In this section, we considered potential risks
of wastewater irrigation on soil structure, including changes in aggregate
stability, bulk density, and hydraulic properties that in turn influence the
retention and transport of heavy metal(loid)s.
3.3.1. Aggregate stability
Aggregate stability is an important soil property because it affects water
infiltration and flow through soils. Wastewater irrigation impacts on soil
aggregate stability through the continuous addition of DOM and salts to the
system (Assouline et al., 2002; Gharaibeh et al., 2007; Menneer et al., 2001;
Vogeler, 2009). DOM stabilizes aggregates through its binding action
and increases in microbiological activity (Vogeler, 2009), while Na accumulation through wastewater irrigation can lead to aggregate dispersion
246
(Misra and Sivongxay, 2009). Flocculation of fine soil particles under saline
conditions has also been observed, highlighting the complex interactions
between sodicity and salinity (Ghadiri et al., 2007). Others found no
significant effect of wastewater irrigation on aggregates stability (Bhardwaj
et al., 2007; Levy et al., 2003).
Sodium in wastewater below critical coagulation concentration can
cause a reduction in aggregate stability, decrease in infiltration rate, and
an increase in the risk of runoff. Alvarez-Bernal et al. (2006) studied the
effect of tannery wastewater on chemical and biological soil properties and
observed that aggregate stability and infiltration properties were adversely
affected by increased Na content in the effluent.
3.3.2. Bulk density and total porosity
Changes in soil bulk density and porosity induced by wastewater irrigation
are dependent on the wastewater quality, in particular, the concentration
of dissolved and particulate constituents of the irrigation water. High concentrations of total suspended solids (TSS) tend to increase soil bulk density,
while wastewater with lower TSS contents has no significant impact on
bulk density (Magesan, 2001; Sopper and Richenderfer, 1979; Vogeler,
2009). Improvements in bulk density and soil porosity have been ascribed to
the addition of DOM (Vogeler, 2009).
Mathan (1994) reported significantly lower bulk density up to a depth of
0.6m after 15years of wastewater irrigation to a sandy loam. Total porosity was
increased by 67% in the topsoil of the wastewater-irrigated soil. An improvement in total porosity was measurable up to a depth of 1.2m. Similarly, bulk
density was significantly decreased and total porosity increased after irrigation
of primary-treated wastewater to a sandy soil for 7 years at a rate of 55mm per
week (Magesan, 2001). In contrast, the study by Jnad et al. (2000) demonstrated
a decrease in the volume of large soil pores under subsurface drip irrigation with
effluents. They attributed this shift in pore size distribution to an accumulation
of suspended solids and an increased salt concentration leading to clay particle
dispersion. Amongst 11 soil physical properties including bulk density, field
water holding capacity, total porosity, clay content, and saturated hydraulic
conductivity measured in the topsoil, Wang et al. (2003b) found that only total
porosity was affected by long-term wastewater irrigation. Shahalam et al.
(1998), however, demonstrated that the impact of short-term wastewater
irrigation on porosity of a silt loam soil was not significant. Wastewater irrigation led to compaction and reduced total porosity, suggesting that total porosity
might be a better soil quality indicator than bulk density (Wang et al., 2003b).
3.3.3. Soil hydraulic conductivity and infiltration rate
The impact of wastewater irrigation on the hydraulic conductivity and
infiltration rate is variable and depends on the soil type, clay content, presence
of CaCO3, antecedent moisture content, the quality of the wastewater and
247
the irrigation technique (Lado and Ben-Hur, 2009; Lado et al., 2005).
A compactive force, for instance the kinetic energy of water drops hitting
the soil, may cause sealing of surface pores due to physical disintegration of
soil aggregates (Mamedov et al., 2000), aggregate slaking controlled by the
wetting rate of the soil (Mandal et al., 2008), TSS of the wastewater
(de Vries, 1972) and/or by clay dispersion and the subsequent blocking of
soil 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 soil
solution is reduced during rain events which can lead to clay dispersion at
the soil surface, seal formation and subsequently to decreased infiltration rates
(Lado et al., 2005; Mandal et al., 2008). In addition, DOM of wastewater can
enhance soil water repellency (Assouline et al., 2002; Tarchitzky et al., 2007;
Travis et al., 2010; Vogeler, 2009; Wallach et al., 2005). Soil water repellency
inhibits water infiltration (Müller et al., 2010). Reduced infiltration rates have
also been attributed to the collapse of soil structure caused by the dissolution
of SOM (Jnad et al., 2000; Lieffering and McLay, 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 increased
macroporosity, which was explained by the increased biological activity
following the wastewater applications. Gharaibeh et al. (2007) observed that
the length of wastewater irrigation might play a role as well. In their study, up
to 5years of wastewater irrigation significantly decreased the infiltration rate
of Vertisol, but 15years of wastewater irrigation increased the infiltration rate
due to the formation of large cracks.
Similarly, it has been suggested that various mechanisms affect hydraulic
conductivity as a result of wastewater irrigation (Table 4). Most studies reported
reduced 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 blocking
of soil pores. A decrease in soil hydraulic conductivity can also be due to
biological (extracellular carbohydrates, cells, and microbiological waste products) clogging of soil pores following the stimulation of microbial communities by wastewater microbial growth. Magesan et al. (1999) indicated a
decrease in the hydraulic conductivity of an allophanic soil after applying
synthetic 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 irrigated
with tertiary-treated wastewater with a C:N ratio 2:1 for 7 years. Wastewater
with a high C:N led to net N immobilization, excess C and subsequently to
an increase in microbial biomass and extracellular carbohydrates that blocked
soil pores and reduced the hydraulic conductivity. The plugging of soil pores
was shown to be more pronounced in fine textured soils due to the high initial
microporosity (Vinten et al., 1983b). Further, wastewater irrigation can change
248
soil chemical properties, in particular the soil’s ESP, salinity, quantity, and
quality of SOM and DOM, which impact on soil structure through clay
swelling 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 that
the soil’s dispersion tendency increased after 5 years, but this did not impact
on the hydraulic conductivity. No significant change in permeability was
detected after 5 years of irrigation with tertiary-treated wastewater in a
major Californian study (Seikh et al., 1998). In contrast, Magesan et al.
(1996) reported that applying secondary-treated sewage effluents increased
the macroporosity of a sandy loam soil from 11% to 19% and consequently,
the hydraulic conductivity increased from 39 to 57mmh1. The infiltration
rate influences the transport velocity of heavy metal(loid)s in soils. If the
infiltration rate is small, transport of heavy metal(loid)s will also be limited
or 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 reactions
that include adsorption, complexation, precipitation, and reduction, that
control their leaching and runoff losses, and bioavailability. In the case of
wastewater irrigation, these reactions are manifested predominantly by the
presence of high amounts of organic carbon (in particular DOM), soluble
salt concentration (salinity), and acidification caused by the mineralization
of organic N.
4.1. Adsorption
The most important physicochemical process affecting the behavior of
metal(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 heavy
metal(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 to
increase 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 the
contents of Fe (Karpukhin and Ladonin, 2008) and Mn oxides (Brown and
Parks, 2001; Stahl and James, 1991).
It has often been observed that heavy metal(loid)s added through organic
amendments, such as effluents, sewage sludge, and manures accumulate in
249
the surface layer, indicating a strong retention within surface soils enriched
with carbon (Adriano, 2001). Table 6 shows selected references on the
effect of wastewater/sludge addition on heavy metal(loid) adsorption. Zhu
et al. (1991) observed that the Cu adsorption maxima by two soils which
received 11 annual applications of Cu in the form of Cu-enriched swine
manure or CuSO4 were higher for the manure-treated soil. This was due to
an increase in OM-induced CEC. Addition of organic amendments has
often been shown to increase the CEC of soils, thereby resulting in
increased cation adsorption. The effect of effluent and manure addition
on CEC-induced cation adsorption in soils is often inferred from literature
in which the effects of other organic amendments, such as biosolid have
been examined. Bolan et al. (2003b), for example, observed that CEC per
unit organic carbon was higher for soils than for biosolid, which was
attributed either to the difference in the nature of OM or to the significant
contribution 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, adsorption
by both acidic and calcareous soils decreased with the addition of sewage
sludge, 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 and
Sposito, 1994; Davis, 1984; Elliott and Denneny, 1982; Gove et al., 2001;
Xu et al., 1989). A significant inverse relationship between the extent of Cu
adsorption and the DOM in the soils treated with organic amendments was
observed by Bolan et al. (2003c) and Hao et al. (2008). These Cu-DOM
complexes are highly mobile in soils and may increase the leaching of Cu.
Wong et al. (2007) noted that the addition of DOM through anaerobically
digested dewatered sludge significantly reduced the Cd sorption capacity with a
maximum inhibition on metal(loid) sorption occurring at pH 7–7.5. Al-Wabel
et al. (2002) reported a positive correlation between increased soil DOM
resulting from biosolid application and Pb and Cu concentrations indicating
formation of soluble metal(loid)-DOM complexes. Kunhikrishnan (2011)
examined the effect of farm dairy, winery, and piggery wastewaters on the
adsorption of Cd, Cu, and Pb using batch experiments and showed that
adsorption 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 range
of solutes. As might be expected, the organic component in wastewater has
a 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
Cu, Fe, Mn,
Ni, Pb,
Zn
Sewage
effluent
Cd, Cr, Cu,
Ni, Pb,
Zn
Sewage water
or sludge
Cu, Cr, Ni,
Zn
Reclaimed
wastewater
Cr, Cu, Pb,
Zn
Municipal
wastewater
Cd, Cu, Fe,
Mn Ni,
Pb, Zn
Untreated
sewage
effluent
Observations
References
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%.
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.
Irrigation with effluents also increased both the total and EDTAextractable metals in the fields. Highest levels of EDTAextractable 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.
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.
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.
Rattan et al.
(2005)
Chen et al.
(2010)
Xu et al.
(2010)
Brzezińska
et al.
(2010)
Rana et al.
(2010)
Cu, Zn
Sewage sludge
Cu
Sewage sludge
Cd, Zn
Anaerobically
digested
dewatered
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.
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.
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.
Mesquita
and
Carranca
(2005)
Ashworth
and
Alloway
(2007)
Wong et al.
(2007)
252
groups that chelate metal(loid)s (Harter and Naidu, 1995). With increasing
pH, the carboxyl, phenolic, alcoholic, and carbonyl functional groups in
OM dissociate, thereby increasing the affinity of ligand ions for metal(loid)s.
It has been observed that addition of wastewater, sewage sludge, or
manure 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 soluble
and insoluble metal(loid)-DOM complexes, thus affecting both movement
and bioavailability of heavy metal(loid)s. While insoluble complexes result
in the retardation of DOM and metal(loid) movement (Guggenberger and
Kaiser, 2003; Jansen et al., 2005; Martin and Goldblatt, 2007), soluble metal
(loid)-DOM complexes enhance their movement. Accordingly, in soils
containing large amounts of OM, such as pasture soils and organic manure
or wastewater-amended soils, only a small proportion of metal(loid)s in soil
solution remains as free metal(loid) ion and a large portion are complexed
with 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 the
dissolved Cu and all Cd in soils treated with cattle manure slurry was bound
in relatively fast dissociating organic-metal(loid) complexes.
Although the formation of soluble metal(loid)-organic complexes reduces
the phytoavailability of heavy metal(loid)s, the mobility of the heavy metal
(loid) may be greater in soils receiving alkaline-stabilized biosolid due to the
reduction of metal(loid) adsorption and increased concentration of soluble
metal(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 between
organic carbon and soluble Cd in solution for sludge-treated soils, indicating
that most of the Cd remained as metal(loid)-organic complex. As reported by
Bolan et al. (2003c), a decrease in Cu adsorption in the presence of DOM is
likely to increase Cu mobility yet does not necessarily increase bioavailability.
Application of effluent and manure has been shown to increase the soluble
salt concentration of soils, as measured by EC (del Castilho et al., 1993; Rana
et al., 2010; Rusan et al., 2007; Sutton et al., 1984). High concentration
of inorganic anions, such as Cl and sulphate (SO2
4 ) in effluents and manure
products induces the formation of metal(loid)-inorganic complexes (e.g., Cd–
Cl complex) that are considered to be even more phytoavailable (Japenga
and Harmsen, 1990; Khoshgoftarmenesh et al., 2002; McLaughlin et al., 1998;
Smolders et al., 1998). Although a wide variety of organic compounds in
DOM 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 hydrophilic
bases have strong affinity for forming soluble complexes with Cd, Cu, and Zn.
253
Thus, the formation of soluble aqueous metal(loid)-organic and to a lesser
extent metal(loid)-inorganic complexes is expected to dominate the solution
chemistry of metal(loid)s in wastewater and manure-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 shown
to 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 and
animal origin contain large amounts of DOM, and the addition of certain
organic manures, such as poultry manure increases the pH, thereby enhancing the solubilization of SOM (Jackson and Miller, 2000; Jackson et al.,
1999). Such an increase in DOM may enhance microbial activity but lower
the 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 amendments 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 inorganic
fertilizer-borne Se, which they attributed to manure-facilitated volatilization due to the reduction of Se. Similarly, Banks et al. (2006), Cifuentes
et al. (1996), Higgins et al. (1998), and Losi et al. (1994) reported a reduction
of Cr(VI) to less toxic and less mobile Cr(III) in soils amended with cattle
manure. Various reasons could be attributed to the enhanced reduction
of Cr(VI) in the presence of organic amendments, including the supply of
carbon and protons and the stimulation of microorganisms that mediate and
facilitate 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 zones
derived from moderately contaminated farmland irrigated with industrial
wastewater for 30years. A column test showed that the concentration of
Cr(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 a
significant difference in the extent of Cr(VI) reduction between various
organic manure composts. Reduction increased with increasing level of
DOM added through manure addition, which has been identified to facilitate the reduction of Cr(VI) to Cr(III) in soils (Jardine et al., 1999; Nakayasu
et al., 1999). For example, the hydroquinone groups in OM have been
identified 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
As and Cr
As
Wastewater/manure
Observations
References
Industrial
wastewater
Poultry litter
Cr(VI) and As(V) were reduced to Cr(III) and As(III)
indicating DOC-induced reduction
Poultry litter increased the solubility of As by
complexation with DOC
Enhanced the reduction of Cr(VI) to Cr(III)
Enhanced the reduction of Cr(VI) to Cr(III)
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
The mixture of sewage sludge and poultry litter
reduced As(V) to more mobile and toxic As(III)
Zhao et al. (2009)
Cr
Cr
Cr
Cattle manure
Cattle manure
Composted cow
manure
As
Sewage sludge and
poultry litter
Jackson et al. (2003)
Cifuentes et al. (1996)
Higgins et al. (1998)
Banks et al. (2006)
Jackson et al. (1999)
255
1200
Cr(VI) (mg L–1)
0 Mg OM ha–1
50 Mg OM ha–1
800
400
0
0
5
10
15
20
Time (weeks)
Figure 4
1994).
Effect of organic matter addition on Cr(VI) reduction in soils (Losi et al.,
Metal concentration (%)
120
100
80
60
40
20
0
After
Before
Chromium
Cr(VI)
Before
After
Arsenic
Cr(III)
As(V)
As(III)
Figure 5 Concentration of Cr(VI), Cr(III) and As (V), As(III) in wastewater-irrigated
soil before and after column tests (Zhao et al., 2009).
The increase in Cr(VI) reduction in the presence of manure and effluent
addition may also result from enhanced microbial activity. Although Cr(VI)
reduction can occur through both chemical and biological processes, the
bioreduction is considered to be the dominant process in most arable soils
that are low in Fe2þ ion. Losi et al. (1994) have reported that adding manure
256
compost generated a larger increase in the bioreduction than chemoreduction of Cr(VI), indicating that the supply of microorganisms is more
important than the supply of organic carbon in enhancing the reduction
of Cr(VI) when compost is added. It has often been reported that an increase
in 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 is
in the ammoniacal form. Oxidation of ammoniacal nitrogen to nitrate
nitrogen (nitrification) and ammonia volatilization result in the release of
protons. It has often been observed that Cr(VI) reduction, being a proton
consumption (or hydroxyl release) reaction, increases with a decrease in soil
pH (Cary et al., 1977; Eary and Rai, 1991).
2Cr2 O7 þ 3C0 þ 16Hþ ! 4Cr3þ þ 3CO2 þ 8H2 O
ð3Þ
Increased concentration of Fe2þ and Mn2þ ions in drainage effluent from
manure- and effluent-amended soil is related to reducing conditions with the
consequent solubilization of these metal(loid)s in soils. Metal(loid)s, such as
Co, are retained by Fe2þ and Mn2þ oxides under oxic conditions (McLaren
et al., 1984) and the manure/effluent-induced reduction of these oxides
results in the release of adsorbed metal(loid)s (L’Herroux et al., 1997; Siebe
and Fischer, 1996). Wallingford et al. (1975) obtained a good correlation
between Mn concentration in corn and cumulative level of feedlot manure
application, which was attributed to enhanced solubilization of Mn due to
reducing conditions in manure-treated soil.
4.4. Methylation/demethylation
Methylated derivatives of As, Hg, and Se can arise as a result of chemical and
biological processes that frequently alter their volatility, solubility, toxicity,
and mobility. Biomethylation of these heavy metal(loid)s has emerged as a
major process for their removal during wastewater treatment using natural and
constructed wetlands (Kosolapov et al., 2004; Stasinakis and Thomaidis, 2010).
The major microbial methylating agents are methylcobalamin (CH3CoB12),
involved in the methylation of Hg, and S-adenosylmethionine, involved
in the methylation of As and Se. Biomethylation may result in metal(loid)
detoxification since methylated derivatives are excreted readily from cells, are
often 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 the
dominant process in soils and aquatic environments. At present there is substantial evidence for the biomethylation of As, Hg, and Se in soils and aquatic
systems (Gadd, 2004; Masscheleyn and Patrick, 1993; Nicholas et al., 2003;
257
Oiffer and Siciliano, 2009) and during wastewater treatment (Stasinakis and
Thomaidis, 2010). Microorganisms in soils and sediments act as biologically
active methylators and OM derived from wastewater application provides the
source of methyl donor for methylation (Kosolapov et al., 2004).
Benthic microbes are capable of methylating As under both aerobic and
anaerobic conditions to produce methylarsines and methyl-arsenic compounds (Maher, 1988). Methylation may play a significant role in the mobilization of As by releasing it from the sediments to the aqueous environment
(Anderson and Bruland, 1991). Similarly, methylated Hg species that are
highly toxic and biologically mobile have been observed in various wastewater sources including dental wastewater (Gbondo-tugbawa et al., 2010; Gustin
et al., 2006; Zhao et al., 2008).
When selenate [Se(VI)] and selenite [Se(IV)] are introduced into moderately reducing conditions they are quickly transformed through microbial
processes to Se0 and/or organic Se compounds. Selenium biomethylation is
of interest because it represents a potential mechanism for the removal of Se
from contaminated environments, and it is believed that methylated compounds, such as dimethyl selenide are less toxic than dissolved Se oxyanions
(Frankenberger and Losi, 1995). For example, biomethylation followed by
volatilization is considered as one pathway by which high Se concentrations
are dissipated from agricultural evaporation ponds in the San Joaquin Valley
of California (Gao and Tanji, 1995).
Literature reports that wastewater from some industries contains quite
high concentrations of soluble Se, as high as 620mgL1 from the Se compounds industry and 20–60mgL1 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 90mgL1. The authors reported that selenite
was detected only in one sample and selenate could not be identified in any
of the analyzed samples. In a study investigating Se removal from selenitecontaminated oil refinery wastewater wetland, Hansen et al. (1998) found
that 89% of the Se was removed. They found that most of the Se was
immobilized into the sediment and plant tissues, whereas biological volatilization could have accounted for 10–30% of its removal. Ye et al. (2003)
reported that 79% of initial Se mass contained in coal gasification plant
wastewater was removed in a constructed wetland. The primary sink for Se
retention was the sediment, which accounted for 63%, whereas accumulation
in plant tissues and biological volatilization to the atmosphere were of minor
importance (Ye et al., 2003).
Frankenberger and Arshad (2001) observed that microorganisms, particularly Enterobacter cloacea, were very active in reduction of Se oxyanions
present in irrigation drainage water, into insoluble Se0. Furthermore, by
monitoring various environmental conditions and addition of organic
amendments, they confirmed that the process could be stimulated manifold.
258
4.5. Leaching and runoff
Long-term application of wastewater to soils can potentially affect the
quality of groundwater resources by excess nutrient loading and heavy
metal(loid) mobilization beyond the plant root zone (Abaidoo et al., 2009;
Hamilton et al., 2007). The impact depends on a number of factors including the depth of the water table, soil properties, soil drainage, management
of wastewater irrigation, quality of groundwater, and the scale of wastewater
irrigation. The capacity for heavy metal(loid)s to contaminate groundwater
relies on the mobility of the heavy metal(loid) concerned, and the amounts
and proportions of complexed and free metal(loid) forms within the soil
solution. The leaching rate of heavy metal(loid)s is also influenced by the
natural OM content of the soil, the concentration and quality of DOM, and
pH of the leaching solution (Antoniadis and McKinley, 2003; van Zomeren
and Comans, 2004).
Many studies have examined the leaching behavior of heavy metal(loid)s
from contaminated soils, industrial sludges, dredged sediments, and municipal
solid 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 respect
to their mobility and ecotoxicological significance, is determined by their
solid-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 to
soil solution depends on their affinity to bind to reactive surfaces in the soil
matrix (Dijkstra et al., 2004). Downward migration of heavy metal(loid)s in
wastewater is facilitated by forming soluble complexes with DOM (Zhou and
Wong, 2001).
L’Herroux et al. (1997) observed that repeated applications of swine
manure slurry increased the drainage water concentrations of Mn from 0.05
to 14mgL1, Co from 0.8 to 50mgL1, and Zn from 17.3 to 100mgL1.
Studies on migration of metal(loid)s in soils after manure slurry applications
have linked metal(loid) mobility with DOM (Amery et al., 2010; Japenga
et al., 1992). Although the soluble organic metal(loid) fraction is not readily
bioavailable to plants, it is relatively mobile and applying organic amendments including wastewater, biosolid, and animal manure has been shown
to enhance the leaching of metal(loid)s in soils (Hsu and Lo, 2000). Del
Castilho et al. (1993), for example, observed a positive relationship between
soluble metal(loid) concentration and DOM in soils treated with cattle
manure slurry. Li and Shuman (1997) observed that leaching metal(loid)contaminated soils with poultry litter extract increased the water-soluble
fractions of Cu and Zn, with a corresponding decrease in exchangeable
fractions, indicating that poultry manure application enhances the solubilization and mobilization of metal(loid)s. Acidification caused by manure
application 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
Untreated
wastewater
Untreated
wastewater
Preferential flow and metal complexation with soluble organics
apparently allowed leaching of heavy metals.
Water extractable Cu and Cd concentrations and the metal leachates
increase and correlate with DOC.
Groundwater was not contaminated through vertical infiltrationinduced leaching. However, substantial build-up of metals in river
sediments and wastewater-irrigated soils were observed.
Zn, Cu, and Cd mobility was observed due to acidic soil pH.
Behbahaninia et al.
(2008)
Herre et al. (2004)
Cu, Cd
Hg, Cd
Zn, Cu, Cd, Treated sewage
Cr
effluent
Cr, Zn, Cd, Poultry litter
Cu, Pb
Cd, Ni, Zn Sewage sludge
Zn, Cd, Cu, Sewage sludge
Ni, Cr
Cu, Ni, and Sewage sludge
Pb
Cu, Cd, Pb, Pig manure
and Zn
amendment in
mine soils
Cu, Zn
Poultry and
livestock
manures
Wu and Cao (2010)
Gwenzi and
Munondo (2008)
Leaching of metals increased with increasing rates of poultry litter.
Paramasivam et al.
(2009)
DOM applications significantly increased the extractability of metals. Antoniadis and
Alloway (2002)
The concentrations of Zn, Cd, Cu, Ni, and Cr in the saturation extract Schaecke et al.
(2002)
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.
The solubility of the heavy metals showed a strong positive relationship Ashworth and
to the solubility of organic matter, particularly at high pH.
Alloway (2008)
Pig manure amendment increased DOM in leachates, thereby
Carmona et al.
increasing the release of metals from mine soil.
(2008)
Total amounts of Cu and Zn eluted from the soil columns significantly Hao et al. (2008)
correlated with the extracted soil Cu and Zn concentrations.
260
DOM plays an important role in facilitating the leaching of contaminants
in soil (Haberhauer et al., 2002; van Zomeren and Comans, 2004) by forming
soluble metal(loid) complexes (Bolan et al., 2011b; McCarthy and Zachara,
1989; Weng et al., 2002). Herre et al. (2004) did a column experiment to
study the effect of wastewater on the leaching of metal(loid)s (Cu and Cd) and
DOM. They found that the amount of Cu leached correlated well with the
DOM concentrations in the leachates (Fig. 6A, B). This agrees well with
many published reports emphasizing the importance of DOM for metal(loid)
A
B
10
Water extractable Cu (mg kg-1)
Water extractable Cu (mg kg-1)
500
400
300
200
100
0
0
0.2
0.4
DOC (mg
g-1)
0.6
8
6
4
2
0
0
0.1
0.2
0.3
0.4
0.5
DOC (mg g-1)
Vertisol R 2 = 0.71
Vertisol R 2 = 0.75
Leptosol R = 0.54
Leptosol R 2 = 0.50
2
C
Cd concentration (mg kg-1)
120
Arthrosol R 2 = 0.81
100
80
60
40
20
0
0
50
100
200
150
250
DOC concentration (mg C kg-1 soil)
Figure 6 Relationship between DOC and water extractable Cu (A) and Cd (B, C) in
different soils [(A, B): Herre et al., 2004; (C): Shan, 2010].
261
mobility (Christensen et al., 1999; Ermakov et al., 2007; Kalbitz and
Wennrich, 1998; Zhu and Alva, 1993) (Fig. 6C). The presence of wastewater
DOM would maintain metal(loid)s in solution and thus limit adsorption onto
the soil (Ashworth and Alloway, 2004; Harter and Naidu, 1995).
Many researchers have viewed DOM as an important contributor to the
elevated mobility of heavy metal(loid)s in soils treated with wastewater, manures, and biosolids (Al-Wabel et al., 2002; Haynes et al., 2009; Kunhikrishnan,
2011; Peckenham et al., 2008). For example, Kunhikrishnan (2011) examined
the effect of piggery, farm dairy, and winery wastewaters on Cu leaching and
noticed that leaching increased with increasing levels of Cu and was higher
in soils treated with wastewater sources than Milli-Q water. Kunhikrishnan
(2011) suggested that the DOM in wastewater sources formed soluble
Cu-DOM complexes, thereby facilitating the movement of Cu in soils
(Fig. 7). Thus, whilst free metal(loid) ions or readily dissociated inorganically
complexed metal(loid)s added to a soil would be expected to become quickly
adsorbed 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)s
in soils can be examined using chemical extraction and bioassay tests. Chemical
extraction tests include single extraction and sequential fractionation (Basta and
Cumulative Cu concentration (mg kg-1)
A
B
0.8
1.6
PE
WE
FDE
MQ
0.6
1.2
0.4
0.8
0.2
0.4
0
0
4
8
Pore volumes
PE
WE
FDE
MQ
12
0
0
4
8
Pore volumes
12
Figure 7 Effect of wastewater irrigation on cumulative Cu concentration of leachates in
a silt loam soil, (A) 100mgkg1 and (B) 500mgkg1 (MQ, Milli-Q water; PE, piggery
effluent; WE, winery effluent; FDE, farm dairy effluent; Kunhikrishnan, 2011).
262
Gradwohl, 2000; Ruby et al., 1996). Bioassay involves plants, animals, and
microorganisms (Naidu et al., 2008; Yang et al., 1991).
5.1. Chemical extraction
5.1.1. Single extraction
The feasibility of predicting the bioavailability of heavy metal(loid)s to
higher plants and various organisms is assessed using selective chemical
extractants (Kelsey et al., 1997; Loibner et al., 2000). Both single extractions
(Beckett, 1989) and sequential extractions (Tessier et al., 1979) are used to
identify those fractions of metal(loid)s in the soil that are more or less readily
available (Kennedy et al., 1997). Bioavailability is organism- and speciesspecific and a single chemical test is insufficient to precisely assess bioavailability accurately (Reid et al., 2000). However, extraction with non-exhaustive
selective extractants that mimics the bioavailability of pollutants is useful for
providing predictors of exposure.
Several methods have been used to evaluate the bioavailability of heavy
metal(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) chelating
agents (e.g., EDTA, DTPA), (c) buffered salts (e.g., 1M NH4OAc), (d) neutral
salts (CaCl2, NH4NO3), and (e) other extractants proposed for routine soil
testing. These extractants have been used to predict the bioavailability of
fertilizer-, wastewater-, manure-, and sludge-borne heavy metal(loid)s in soils
(Gupta and Sinha, 2007; Marchi et al., 2009; Payne et al., 1988; van der
Watt et al., 1994). Chelating agents such as EDTA and DTPA have often
been found to be more reliable in predicting the plant availability of sludge- and
wastewater-borne heavy metal(loid)s (Gupta and Sinha, 2007; Sims and
Johnson, 1991), since they are more effective in removing soluble metal
(loid)-organic complexes that are potentially bioavailable. However, it should
not 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 heavy
metal(loid)s, Cr, Cd, Cu, Ni, Pb, and Zn into agricultural fields through
municipal wastewater irrigation. They analyzed the available concentration
of heavy metal(loid)s using DTPA extraction and found a maximum of
64.84% extraction in the case of Cr. They attributed the low extraction of
other 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, and
method of extraction. They also noticed that EDTA was more suitable for
acidic soils, whereas DTPA was considered more suitable for neutral and
near 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; Lin
and Zhou, 2009). Luo et al. (2003) studied the accumulation, chemical
263
fractionation, and availability of Cu to rice in a paddy soil irrigated with Cuenriched wastewater. They found that with irrigation the concentrations of
Cu in the exchangeable (NH4OAc-extractable) and complexed (EDTAextractable) fractions increased rapidly, from 0.33 to 6.30mgkg1 and from
14.1 to 98.0mgkg1, respectively. EDTA-extractable Cu was much higher
than NH4OAc-extractable Cu in the soils.
Calcium chloride (CaCl2) soil extraction, although a neutral salt extraction, is a widely used and internationally recognized technique (Harmsen,
2007; Peijnenburg et al., 2007). van der Welle et al. (2007) reported that
under aerobic conditions, plant heavy metal(loid) uptake was best predicted
by the amount of CaCl2-extractable metal(loid)s. Metal(loid) extraction
with CaCl2 solution was found to be effective, with Sauvé et al. (1996)
and McBride (2001) reporting that Cu2þ in 0.01M CaCl2 correlated
strongly with plant yields or tissue Cu concentrations of rye grass and
crop species such as maize, lettuce, and radish. Wightwick et al. (2010)
determined the environmental availability of Cu in Australian vineyard soils
contaminated with fungicide derived Cu residues irrigated with treated
sewage. They reported that differences in Cu availability determined by
0.01M CaCl2 extractable Cu concentrations were related to the total Cu
concentration and soil properties, including pH, clay, exchangeable K, silt,
and CaCO3. Kunhikrishnan et al. (2011) compared the free Cu2þ concentrations in CaCl2 extract and pore-water from soils in the presence of farm
dairy and piggery wastewater sources. They reported that the free Cu2þ
concentrations were lower in soils incubated when wastewater sources
were present. These results suggest the formation of Cu-DOM complexes
decreases the amount of free Cu2þ in the soil solution.
5.1.2. Sequential fractionation
Fractionation studies are often used to examine the influence of amendments, such as wastewater, CaCO3, P compounds, and biosolid on the
immobilization of heavy metal(loid)s. Following adsorption, irrespective
of the nature of interaction between heavy metal(loid)s and soil colloidal
particles, metal(loid) ions are redistributed amongst organic and mineral soil
constituents (Bolan et al., 2003e; Fedotov and Mirò, 2008). Factors affecting the distribution of heavy metal(loid)s among different forms include
pH, ionic strength of the soil solution, solid and solution components as
well as their relative concentration and affinities for heavy metal(loid), and
reaction time (Bolan et al., 2003e; Shuman, 1991). The various forms
of the heavy metal(loid)s that are sequentially extracted can be classified
as soluble, adsorbed/exchangeable, carbonate-bound, organic-bound,
amorphous ferromanganese hydrous oxide-bound, crystalline ferromanganese hydrous oxide-bound, and residual or lattice mineral-bound.
The phytoavailability of the different forms of the solid phase species
generally decreases in the following order: soluble>adsorbed/exchangeable
264
>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 as
sludge or wastewater shifts the solid phases of the heavy metal(loid)s away
from immobile fractions to forms that are potentially more mobile, labile,
and bioavailable. For example, Dudka and Chlopecka (1990) found with
sewage sludge application the residual forms of Cd2þ, Cu2þ, and Zn2þ in
soil decreased from 34–43% to 6–34%, with a corresponding increase in
the readily phytoavailable forms. Through sequential extraction of Cu,
Cd, Pb in four soils irrigated with wastewater, Flores et al. (1997) discovered metal(loid)s were predominantly associated with organic soil fractions. Bashir et al. (2007) studied the fractionation of heavy metal(loid)
s (Cd, Mg, and Zn) in soils irrigated with untreated sewage effluent for a
long period of time. Extraction procedure showed that most of the heavy
metal(loid)s (>50%) was bound to residual fraction. Among nonresidual
fractions, Cd and Mn were present in reducible fraction while Zn was
present in oxidizable fraction. Luo et al. (2003) analyzed the fractionation
of Cu in a paddy soil irrigated with Cu-enriched wastewater. They
reported marked increases in the weak acid-soluble (HOAc-extractable),
reducible Fe and Mn oxide-bound (NH2OHHCl-extractable), oxidizable OM-bound (H2O2-extractable), and residual fractions of Cu in the
wastewater-irrigated soils, indicating an increase in mobility and bioavailability of Cu leading to Cu toxicity in the plants.
Physiologically based in vitro chemical fractionation schemes are becoming 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 bioaccessibility extraction test (SBET), Deutsches Institut für Normung (DIN), and
gastrointestinal (GI) test. These improved tests make it possible to predict
the bioavailability of heavy metal(loid)s in soil and sediments or when
ingested 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 predicting
the bioaccessibility of contaminants from a soil matrix. While the PBET
method has been applied to both organic and inorganic contaminants, it is
more commonly recognized as an assay for assessing heavy metal(loid)
bioaccessibility (Bolan et al., 2008). Assadian and Margez (2006) studied
the bioaccessibility of heavy metal(loid)s (Cd, Cr, Ni, and Pb) using chemical fractionation and in vitro GI and PBET methods in soils blended with
untreated effluent and biosolids. The results indicated that chemical fractionation 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 concentrations
measured in sheep tissues.
265
5.2. Bioassay
5.2.1. Phytoavailability
It can be expected that in wastewater-treated soils just like in sewage sludgeand 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), for
example, have demonstrated that when maize and other crops were grown
on Cu-contaminated sludge-amended soils, inconsequential changes in plant
tissue Cu concentrations in response to substantial increases in total Cu
loading in soils occurred. Organic amendments may also alleviate oxyanion
phytotoxicity; for example, the uptake of Se was less in the presence of
organic amendments (seleniferous plant tissues and manure) than that from
inorganic 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) in
reducing the heavy metal(loid) availability in the soil, and subsequent uptake
in Beta vulgaris L. (var. All green). They observed that phytoavailability of Cd,
Cu, Pb, Zn, Mn, Ni, and Cr determined by bioconcentration factor (BF¼
Plant concentration/soil concentration) was lowest in FYM and highest in
NPK treated soil, compared to the untreated control (Fig. 8). They also
noticed that the yield of B. vulgaris was also highest in FYM-treated soil and
suggested that application of FYM alone and in combination with NPK may
Bioconcentration factor
2.5
2
1.5
1
0.5
0
Cd
Cu
Pb
Zn
Mn
Ni
Control
NPK fertilizer
Farmyard manure
Farmyard manure+
NPK
Cr
Figure 8 Effect of organic and inorganic fertilizers on heavy metal(loid) uptake in Beta
vulgaris 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
Rusan et al.
(2007)
Al, Cr, Mn, Fe
Co, Ni, Cu,
Zn, Cd, Pb
Treated municipal
wastewater
Sunflower,
Sorghum
Zn, Cu, B, Mn,
Fe, Mo
Raw wastewater
Cabbage
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.
Sorghum accumulated higher
concentrations of Mn and Zn,
whereas sunflower accumulated
higher concentrations of Cr.
It increased the metal content of cabbage
plants.
Cu, Zn, Mn, Fe
Raw wastewater
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.
Mn, Zn, Cu, Pb,
Ni, Cr, Cd
Municipal
wastewater
Radish, spinach,
turnip, brinjal,
cauliflower, mint,
coriander, carrot,
lotus stem
Leafy vegetable,
palak
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.
Mn showed maximum uptake followed
by other metals.
Ahmed and
Al-Hajri
(2009)
Kiziloglu
et al.
(2007)
Arora et al.
(2008)
Singh and
Agrawal
(2010)
Amiri et al.
(2008)
Cu
Sewage sludge
Fescue
Pb, Cd, As, Zn,
Hg
Fermented pig
slurry
Tomato
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.
All tomato samples were within the
legislation limits of tested metals.
Forsberg
et al.
(2009)
Kouřimská
et al.
(2009)
268
be considered as an easy and cost-effective technique for reducing the levels of
contamination in food crops.
Addition of DOM to soils through wastewater irrigation and sludge
addition can influence phytotoxic effectiveness of ions in at least two
different ways. On the soil side, an increase in DOM will shift metal(loid)
partitioning toward the soil solution and hence increase the content of
soluble metal(loid) in solution. On the solution side, although the soluble
metal(loid) increases, the free metal(loid) ion is decreased due to DOM
complexation. 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, potentially leading to groundwater contamination, these complexes have been
shown to be less available for plant uptake, thereby alleviating phytotoxicity
that 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 of
metal(loid)s dissolved in pore-water is likely to be available for plant uptake,
the remainder is complexed with DOM (Huynh et al., 2008; Kunhikrishnan
et al., 2011). Bolan et al. (2003c) studied Cu uptake in using mustard plants
amended with biosolids plus various levels of Cu (0–400mgkg1 soil). They
observed that adding manure compost increased the adsorption and complexation of Cu in soil, noting a significant inverse relationship between the
extent of Cu adsorption and DOM in the manure-amended samples. This
indicated that DOM formed soluble complexes with Cu. They reported
that although soluble DOM complexes were formed, addition of biosolids
was effective in reducing the phytotoxicity of Cu, especially at high levels
of Cu addition. In mustard plants amended with farm dairy and piggery
wastewaters, Kunhikrishnan et al. (2011) observed an increase in Cu uptake
with increasing Cu input. However, at the same level of Cu application,
plants took up less Cu from wastewater-amended soils than from Milli-Q
water amended soils. They concluded that the presence of DOM in the
wastewater sources was effective in reducing the phytotoxicity of Cu at high
levels of Cu addition, indicating that the Cu-DOM complexes decreased
the plant availability of Cu (Fig. 9a).
Qishlaqi et al. (2008) assessed the negative impacts of wastewater irrigation on soils and crops collected from two wastewater-irrigated sites and
a reference site where bore water was irrigated. The results showed that
among the five heavy metal(loid)s (Ni, Pb, Cd, Zn, and Cr) studied, using
untreated wastewater caused contamination of spinach and lettuce with Cd
due to its high phytoavailability in topsoil and excessive accumulation of Ni
and Pb in wheat. This scenario was due to the continual addition of heavy
metal(loid)s through long-term wastewater application. They reported that
accumulation of metal(loid)s strongly depended on the crop’s physiological
properties (Liu et al., 2005) and the soil properties (Sharma et al., 2007).
269
A
B
40
MQ
FDE
PE
80
mg C-CO2 released g-1 soil h-1
Tissue Cu concentration (mg kg-1)
100
60
40
20
0
0
200
400
-1
Cu level (mg kg )
600
MQ
FDE
WE
PE
30
20
10
0
0
400
800
1200
Cu level (mg kg-1)
Figure 9 Effect of Cu levels on (A) Cu concentration in plant tissue and (B) substrateinduced respiration in a silt loam soil in the presence of wastewater sources (PE, piggery
effluent; 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 of
treated municipal wastewater irrigation in barley, and then declined with
longer irrigation periods. However, Pb and Cd continued to increase with
longer periods of irrigation. Several other studies report an increased uptake
of heavy metal(loid)s by plants due to continuous loading of metal(loid)s to
soil via irrigation (Abbas et al., 2007; Amiri et al., 2008; Kiziloglu et al., 2007;
Moyo and Chimbira, 2009) (Table 9). Kiziloglu et al. (2008) compared
the accumulation of heavy metal(loid)s (Fe, Cu, Mn, Zn, Pb, Ni, Cd) in
cauliflower and cabbage species irrigated with either untreated or primarytreated wastewater. Metal(loid)s in vegetables irrigated with untreated wastewater were higher than those irrigated with primary-treated wastewater.
5.2.2. Microbial and earthworm availability
As in the case of phytoavailability, the microbial availability of metal(loid)s is
largely controlled by the activity of free ionic species in soil solution. Bolan
et al. (2003a) observed the concentration of total Cu required to cause 50%
reduction in basal respiration (microbial toxicity—MT50) was lower for
CuSO4 (297mg Cu kg1) than for the dairy pond sludge-Cu (783mg Cu
kg1), inferring that sludge-borne Cu was less detrimental to microbial
activity than inorganic CuSO4. However, when the respiration value was
plotted against the concentration of free ionic Cu2þ, a single smooth curve
was obtained for both CuSO4 and sludge-Cu, and the MT50 value was
found to be 18.37mgkg1. This indicates that the difference in the effect on
270
respiration between the two Cu sources (i.e., organic vs. inorganic) is due to
the difference in bioavailable Cu content in soil.
Soil microbial activity as measured by respiration and microbial biomass
carbon was monitored by Kunhikrishnan (2011) following application of
various levels of Cu (0–1000mgkg1), added as copper nitrate-spiked
Milli-Q water, farm dairy, piggery, and winery wastewaters. The effect of
Cu on soil microbial activity varied between Milli-Q water and wastewater
sources and was attributed to the difference in the concentration of DOM.
Metabolic quotient values were lower in soils in the presence of wastewater
than in the Milli-Q water. The results indicated that wastewater sources
decreased the inhibitory effect of Cu on microbial activity and suggested
that 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, depends
on factors such as type and form of metal(loid) and concentration (Heikens et al.,
2001; Hobbelen et al., 2006; Nahmani et al., 2007; Spurgeon et al., 2006), soil
type and characteristics (Hendrickx et al., 2004; Hobbelen et al., 2006; Janssen
et al., 1997; Kizilkaya, 2005; Spurgeon et al., 2006), test species (Heikens et al.,
2001; Hendrickx et al., 2004; Nahmani et al., 2007), temperature (Olchawa
et al., 2006), and exposure duration (Nahmani et al., 2007). Field observations
have demonstrated that Cu is detrimental to lumbricid earthworms (Niklas and
Kennel, 1978; van Rhee, 1975). This is supported by laboratory studies, which
showed that the toxicity of Cu to earthworms is influenced by the pH and OM
of the soil (Ma, 1984; Streit, 1984; Streit and Jaeggy, 1983). Aporrectodea
tuberculata (Beyer et al., 1987) and Aporrectodea caliginosa (Perämäki et al., 1992)
have been found to accumulate high Cd in acidic soils. This is largely related to
the fact that most of the heavy metal(loid)s that accumulate in an earthworm’s
body originate from pools of dissolved metal(loid)s which are bioavailable in the
soil pores of acidic soils (Herms and Brummer, 1984).
Kunhikrishnan (2011) examined the bioavailability of Cu to earthworms
in the presence of farm dairy, piggery, and winery wastewaters varying in
DOM. Bioavailability of Cu to earthworms as measured by mortality and
avoidance test was monitored at various levels of Cu (0–1000mgkg1),
added as copper nitrate-spiked Milli-Q water and wastewater sources. The
results indicated that the wastewater sources decreased the inhibitory effect
of Cu on the earthworm toxicity due to the formation of Cu-DOM
complexes 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 the
earthworms clearly avoided soils with high levels of Cu concentrations.
Although DOM plays a protective role in reducing metal(loid) toxicity
to earthworms, evidence suggests that earthworms play a humifying role in
the soil because humic acids were detected in earthworm-worked soil that
were not present in the non-humified starting material (Businelli et al.,
271
Cu concentration in earthworms (mg kg–1)
70
60
50
40
30
20
10
0
0
100 500
MQ
0
100 500
FDE
0
100 500
WE
0
100 500
PE
Figure 10 Effect of wastewater irrigation on Cu concentration in earthworms in a silt
loam 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)s
to plants by forming organometal(loid) complexes (Evangelou et al., 2004;
Halim et al., 2003). Some authors suggest that increased uptake of heavy
metal(loid)s by plants due to earthworm activity may be a direct result
of metal(loid)-chelating organic materials released by earthworms forming
organometal(loid) complexes (Currie et al., 2005; Udovic et al., 2007; Wang
et al., 2006). A significant link has been found between the DOM increase in
soils by earthworms Eisenia fetida and the concentration of water extractable
metal(loid)s (Zn, Cu, Cr, Cd, Co, Ni, and Pb) (Wen et al., 2004). An increase
in DOM in soil has also been noted in one study reporting that the presence
of 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 conditions, and economic development have led to a considerable increase in the
volume of wastewater generated by domestic, industrial, and commercial
practices (Asano et al., 2007; Lazarova and Bahri, 2005; Qadir et al., 2010).
Although water quality management is a high priority and a major concern
for developing countries, most do not have sufficient resources to treat
wastewater. Therefore, wastewater in a partially treated, diluted, or untreated
272
form is diverted and used by urban and peri-urban farmers to grow a range of
crops (Ensink et al., 2002; Murtaza et al., 2010). Farmers consider wastewater
to be a reliable or sometimes the only water source available for irrigation
throughout the year and it often negates the need for fertilizer application
since it provides a source of nutrients. Similarly, the use of treated wastewater
for both agricultural production and environmental protection has increased
in recent years in several continents including Australia, Europe, and North
America (Qadir et al., 2007; US EPA, 2004).
Like the supply of nutrients and OM through wastewater irrigation, it
also contains different types and levels of undesirable constituents depending
on the source and level of its treatment. On the positive side, OM added
through wastewater improves soil structure, enhances charge characteristics
of irrigated soils, such as CEC, which may retain undesirable metal(loid)
ions rendering them less available for plants, and acts as a storehouse of
essential nutrients for crop growth. On the negative side, heavy metal(loid)
inputs to soils via wastewater irrigation are incommodious because, once
accumulated, it is difficult to remove them. This situation may subsequently
lead to toxicity-related issues in plants grown on contaminated soils, pose
potential harm to people and animals who may consume contaminated
crops and they can be transported from soils to groundwater or surface
water, thereby rendering the water hazardous for other uses (Murtaza et al.,
2010). Most wastewater sources are rich in DOM which influences the
biological transformation processes of heavy metal(loid)s including their
mobility and bioavailability. The transport and bioavailability of heavy
metal(loid)s can be strongly influenced by forming soluble and insoluble
complexes with DOM. Such interactions can alter the chemical speciation
of the heavy metal(loid)s modifying their affinity for sorptive surfaces in the
soil matrix or their uptake, accumulation, and eventual toxicity to organisms (Arnold et al., 2010; Boyd et al., 2005). While the insoluble complexes
are not available to plants and other soil organisms, the question arises
whether the soluble heavy metal(loid) complexes in wastewater become
bioavailable or not. Anodic stripping voltammetry measurements and other
speciation techniques have indicated that only a small percentage of the total
dissolved heavy metal(loid)s exist as free ions and the remainder appears
to be complexed with DOM. More research into this area is required to
unravel the stability of such complexes, how they affect soil organisms and
plants 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, the
following research areas could be pursued:
Long-term stability and biogeochemistry of metal(loid)s immobilized by
wastewater sources.
Influence of wastewater sources on rhizosphere biochemistry in relation
to metal(loid) dynamics.
273
Long-term leaching studies examining groundwater contamination
through the movement of a wide range of chemical pollutants in wastewater, 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 concentration
can 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 of
metal(loid)-DOM complexation on bioavailability and toxicity of heavy
metal(loid)s.
Although wastewater-borne metal(loid)s are reported to be less toxic to
soil microorganisms, long-term studies are required to understand their
dynamics in soils.
In addition to the accretion of salts and nutrients, under certain conditions
wastewater irrigation has the potential to translocate pathogenic bacteria
and viruses to groundwater. It is therefore essential that other control
options should be continued in parallel with ongoing efforts to identify
key wastewater pollutants and suitable techniques for their treatment.
Pragmatic approaches are required to protect water quality and ensure
that wastewater is used in a sustainable way. Risk assessment conducted
prior to wastewater irrigation is highly recommended to enable the safe use
of wastewater for landscape and agricultural irrigation. There are several
other opportunities for improving wastewater management through guidelines and policies, which would reduce potential environment and public
health risk. For instance, governments should implement an integrated
water management approach, promote public participation, disseminate
existing knowledge, generate new knowledge, and monitor and administer
imposed standards.
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The Influence of Wastewater Irrigation on the

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