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Role of Inorganic and Organic Soil Amendments
5/1/2003
Australian Journal of Soil
Research
By V. P. Duraisamy
Abstract
Soil is not only considered as a 'source' of nutrients for plant growth, but
also as a 'sink' for the removal of contaminants from industrial and
agricultural waste materials. The origin of heavy metal contamination of soils
may be anthropogenic as well as geogenic. With greater public awareness of the
implications of contaminated soils on human and animal health, there has been
increasing interest amongst the scientific community in developing
cost-effective and community-acceptable remediation technologies for
contaminated sites. Unlike organic contaminants, most metals do not undergo
microbial or chemical degradation, thereby resulting in their accumulation in
soils. The mobilisation of metals in soils for plant uptake and leaching to
groundwater can, however, be minimised through chemical and biological
immobilisation. Recently there has been increasing interest in the immobilisation
of metals using a range of inorganic compounds, such as lime and phosphate (P)
compounds, and organic compounds, such as 'exceptional quality' biosolids.
In this review paper, the results from selected New Zealand studies on the
potential value of a range of soil amendments (phosphate compounds, liming
materials, and biosolids) in the immobilisation of cadmium (Cd), chromium (Ct),
and copper (Cu) is discussed in relation to remediation of contaminated soils.
These case studies have indicated that lime is effective in reducing the
phytoavailability of Cd and Cr(III), phosphate compounds are effective for Cd,
and organic amendments are effective for Cu and Ct(VI). The mechanisms proposed
for the immobilisation and consequent reduction in the phytoavailability of
metals by the soil amendments include: enhanced metal adsorption through
increased surface charge (e.g. phosphate-induced metal adsorption), increased
formation of organic and inorganic metal complexes (e.g. cadmium-phosphate
complex and copper-organic matter complex), precipitation of metals (e.g.
chromic hydroxide), and reduction of metals from higher valency mobile form to
lower valency immobile form [e.g. Cr(VI) to Cr(III)]. These case studies
indicated that since bioavailability is the key factor for remediation
technologies, chemical or biological immobilisation of metals may be a
preferred option.
Additional keywords: biosolid, cadmium, chromate reduction, copper, liming,
phosphate.
Introduction
The term 'heavy metal' in general includes elements (both metals and
metalloids) with an atomic density >6 g/cra3 [with the exception of arsenic
(As), boron (B), and selenium (Se)] (Adriano 2001). This group includes both
biologically essential [e.g. cobalt (Co), copper (Cu), chromium (Cr), manganese
(Mn), and zinc (Zn)] and non-essential [e.g. cadmium (Cd), lead (Pb), and
mercury (Hg)] elements. The essential elements (for plant, animal, or human
nutrition) are required in low concentrations, and hence are known as 'trace
elements' or 'micro nutrients'. The non-essential metals are phytotoxic and/or
zootoxic and are widely known as 'toxic elements'. Both groups are toxic to
plants, animals, and/or humans at extremely high concentrations. Nriagu (1988)
states that 'this very profound experiment, in which one billion (109) human
guinea pigs are being exposed to undue insults of toxic metals, has yet to
receive scientific attention that it clearly deserves'.
With increasing demand for safe disposal of wastes generated from
agricultural and industrial activities, soil is not only considered as a source
of nutrients for plant growth, but also used as a sink for the removal of
contaminants from these waste materials (Cameron et al. 1997; Edwards and
Someshwar 2000). As land treatment becomes an important waste management
practice, soil is increasingly being seen as a major source of metals reaching
the food chain, mainly through plant uptake and animal transfer. Such waste
disposals have led to significant build-up in soils of a wide range of metals,
such as Cd, Cr, Cu, Hg, Pb, and Zn, and metalloids, such as As, Cr, and Se.
Entry of soil-borne metals into the food chain depends on the amount and source
of metal input, the properties of the soil (especially soil pH, organic matter,
and clay content), the rate and magnitude of uptake by plants, and the extent
of absorption by grazing animals. The 'cleaning' action of soil is controlled
largely by the physico-chemical reactions of metals with soil components
carrying surface charge and the biochemical transformations involving soil
microorganisms (Bolan et al. 1999a; Adriano 2001).
Health authorities in many parts of the world are becoming increasingly
concerned about the effects of heavy metals on environmental and human health.
Historically, heavy metal toxicity to human health received attention primarily
as a result of two series of widespread poisoning. First, the many cases of
'Gasio-gas' poisoning, in which arsenic trioxide in wallpaper glue was
converted into volatile poisonous trimethyl arsine or 'Gasio-gas'
[[(C[H.sub.3]).sub.3]As]. Second, the hundreds of tragic cases of human
poisoning of Minamatas Bay and Niigata in Japan (Minamata
disease) in the late 1950s. These poisonings were believed to have occurred
from the ingestion of fish containing methylmercuric compounds probably derived
through biomethylation of mercuric salts by aquatic organisms. Other cases of
direct methyl mercury poisoning have occurred from the use of these compounds
as fungicides on seeds that were subsequently fed to swine as flour and
thereafter eaten by humans (Adriano 2001).
Recently, high concentrations of heavy metals such as As, Cd, Cu, Pb, and Zn
in soils have often been reported in number of countries. For example,
significant adverse impacts of As on human health have been recorded in Bangladesh, India,
and China,
and it is claimed that millions of people are potentially at risk from As
poisoning. Similarly, Cd accumulation in the offal (mainly kidney and liver) of
grazing animals in New Zealand
and Australia
made it unsuitable for human consumption and affected access of meat products
to overseas markets (Roberts et al. 1994). Furthermore, bioaccumulation of Cd
in potato, wheat, and rice crops has serious implications to local and
international commodity marketing (McLaughlin et al. 2000).
For diffuse distribution of metals (e.g. fertiliser-derived Cd input in
pasture soils), remediation options generally include amelioration of soils to
minimise the metal bioavailability. Bioavailability can be minimised through
chemical and biological immobilisation of metals using a range of inorganic
compounds, such as lime and phosphate (P) compounds (e.g. apatite rocks), and
organic compounds, such as 'exceptional quality' biosolid (Knox et al. 2000;
Basta et al. 2001). The more localised metal contamination found in urban
environments (e.g. Cr contamination in timber treatment plants) is remediated
by any combination of processes that include bioremediation (including
phytoremediation), chemical washing, electroremediation, landfills, and
immobilisation (Naidu et al. 1996). Removal of metals through phytoremediation
techniques and the subsequent recovery of the metals or their safe disposal are
attracting research and commercial interests (Cunningham and Lee 1995).
However, when it is not possible to remove the metals from the contaminated
sites by phytoremediation, other viable options, such as in situ
immobilisation, should be considered as an integral part of risk management.
Following a brief overview of the sources of specific soil amendments and
the major heavy metals in soils, this paper provides a synthesis of various
published data on the immobilisation of metals in soils. Selected New Zealand
case studies involving mostly laboratory incubation and glasshouse plant growth
experiments have been used to illustrate the potential value of three soil
amendments (phosphate compounds, liming materials, and biosolids) in the
immobilisation and phytoavailability of heavy metals (Cd, Ct, and Cu).
Sources of soil amendments
A number of studies have examined the potential value of various soil
amendments in the immobilisation of metals in soils, thereby reducing their
bioavailability (Table 1). It is important to understand the sources and
reactions of some of these amendments in soils in order to understand their
interactions with metals in soils.
Phosphate compounds
Phosphorus reaches soils through both pedogenic and anthropogenic sources.
Although most soil parent materials contain some P, the majority of the P is
introduced through fertiliser and manure additions. Phosphate compounds that
are used as a fertiliser source are broadly grouped into water-soluble
(fast-release) and water-insoluble (slow-release) fertilisers (Bolan et al.
1993). The important water-soluble P fertilisers include single super-
phosphate (SSP), triple superphosphate (TSP), monoammonium phosphate (MAP), and
diammonium phosphate (DAP). The important water-insoluble P fertilisers include
phosphate rocks (PRs) and basic slag. Partially acidulated phosphate rocks (PAPR)
and superphosphate and reactive rock mixtures (e.g. Longlife super in New Zealand)
contain both water-soluble and water-insoluble P components. Monocalcium
phosphate (MCP) and ammonium phosphate (AMP) are the principal P components
present in superphosphates (SSP and TSP) and ammonium phosphates (MAP and DAP),
respectively. Apatite is the principal P mineral in PRs.
When water-soluble P compounds, such as superphosphate fertilisers, are
added to soils, the dissolution of MCP results in the formation of slowly
soluble dicalcium phosphate (DCP) with a release of phosphoric acid close to
the fertiliser granules (Eqn 1):
(1) Ca [([H.sub.2]P[O.sub.4]).sub.2] + [H.sub.2]O [right arrow]
CaHP[O.sub.4] * [H.sub.2]O + [H.sub.3]P[O.sub.4] (
Phosphoric acid subsequently dissociates into phosphate and hydrogen
ions
(protons--H+). The protons reduce the pH around the fertiliser granules
to a
very low level (<2 pH). When ammonium phosphate fertilisers are
added to soil, they dissociate into ammonium and phosphate ions. The
subsequent oxidation of N[H.sub.4.sup.+] to N[O.sub.3.sup.-] results in
the release of protons (Eqn 2): >
(2) N[H.sub.4.sup.+] + 2[O.sub.2] [right arrow]) N]O.sub.3.sup.-] +
2[H.sup.+] + [H.sub.2]O
The acidic solution around the fertiliser granules dissolves the iron (Fe)
and aluminium (A1) compounds in the soil, resulting in the adsorption and
precipitation of E The pH around the ammonium phosphate fertiliser granules,
however, is unlikely to be as low as that around superphosphate fertilisers,
causing less adsorption of phosphate ions. The acidity generated can have
important implications to the mobilisation of metals in soils.
When insoluble P fertilisers, such as PRs, are added, the phosphate mineral
apatite needs to be dissolved in soils for the P to become plant-available.
Dissolution of PRs is a prerequisite not only for the plant availability of P
(Rajan et al. 1996), but also for the immobilisation of metals through
precipitation as metal phosphates (Laperche and Traina 1998). In soils, PRs
dissolve by using the acid produced in the soils (Eqn 3):
(3) [Ca.sub.10][(P[O.sub.4]).sub.6][F.sub.2] + 12[H.sup.+] 10[Ca.sup.2+] +
6[H.sub.2]P[O.sub.4] + 2[F.sup.-]
This is a major reason why PRs are very effective as a nutrient
source
mainly in acid soils (pH <6.5) (Bolan et al. 1990) and as a metal
immobilising agent in acid environments (e.g. coal refuse and acid mine
drainage) (Evangelou and Zhang 1995). Once the PR is dissolved, the P
released undergoes similar adsorption and precipitation reactions as in
the case of soluble P fertilisers. >
Liming materials Although liming is primarily aimed at ameliorating soil
acidity, it is increasingly being accepted as an important management tool in
reducing the toxicity of heavy metals in soils (Brallier et al. 1996; Brown et
al. 1997; Bolan et al. 2003a). A range of liming materials is available, which
vary in their ability to neutralise the acidity. These include calcite
(CaC[O.sub.3]), burnt lime (CaO), slaked lime (Ca[(OH).sub.2]), dolomite
(CaMg[(C[O.sub.3]).sub.2]), and slag (CaSi[O.sub.3]). The acid-neutralising
value of liming materials is expressed in terms of calcium carbonate equivalent
(CCE), defined as the acid-neutralising capacity of a liming material expressed
as a weight percentage of pure CaC[O.sub.3]. A neutralising value >100
indicates greater efficiency of the material relative to pure CaC[O.sub.3]. The
amount of liming material required to rectify soil acidity depends on the
neutralising value of the liming material and pH buffering capacity of the
soil. Recently, the potential value of other Ca-containing compounds in
overcoming the problems associated with acidification has been evaluated (Dick
et al. 2000). Some of these materials include PRs, flue gas desulfurisation
(FGD) gypsum, fluidised bed boiler ash (FBA), fly ash, and lime-stabilised
organic composts.
Increasing amounts of PRs are added directly to soils mainly as a source of
P. Unlike soluble P fertilisers, such as superphosphates, PRs can also have a
liming value in addition to supplying P and Ca. The liming action of PRs can
occur through two processes. Firstly, most PRs contain some free CaC[O.sub.3],
which itself can act as a liming agent. Secondly, the dissolution process of
the P mineral component (i.e. apatite) in soils consumes [H.sup.+], thereby
reducing the soil acidity (Eqn 3). It is estimated that every 1 kg of P
dissolved from PRs generates a liming value equivalent to 3.2 kg CaC[O.sub.3]
(Bolan et al. 2003b). From the amounts of P and free CaC[O.sub.3] present in
the PR it may be possible to calculate its total liming value. For example, a
tonne of North Carolina Phosphate Rock (NCPR), which contains 13.1% P and 11.7%
free CaC[O.sub.3], can have a potential liming value of 536 kg CaC[O.sub.3]
(117 kg free CaC[O.sub.3] plus 3.2 x 131 = 419 kg CaC[O.sub.3] upon
dissolution). While the free CaC[O.sub.3] in PRs dissolves reasonably fast
providing a small amount of immediate liming value, the apatite dissolves at a
variable but generally slower rate providing liming value over a longer period
of time.
Biosolids
Conventionally the term 'biosolid' refers to the final product derived from
the biological treatment of municipal wastewaters. However, recently the
terminology connotates a more inclusive definition to also include livestock
waste. Traditionally biosolid is viewed as one of major sources of metal
accumulation in soils, and a large volume of work has been conducted to examine
the mobilisation and bioavailability of biosolid-borne metals in soil (Page
1974; Juste and Mench 1992; Adriano 2001). Advances in the treatment of sewage
water and isolation of industrial wastewater in the sewage treatment plants
have resulted in a steady decline in the metal content of biosolid.
Furthermore, stabilisation using alkaline materials has resulted in the
immobilisation of metals in biosolid. Recent studies have shown that these
alkaline-stabilised biosolids that are low in total and/or bioavailable metal
content (known as 'exceptional quality' biosolid or 'designer sludge') can be
used as an effective sink for reducing the bioavailability of metals in
contaminated soils and sediments (Brown et al. 1998; Basta et al. 2001).
The ability of biosolids to limit metal solubility was inadvertently realised.
Concerns over metal contamination of soils and the potential for adverse
effects on human health due to the transfer of metals from soils to food crops
formed the basis for initial concerns regarding the beneficial use of biosolids
on agricultural soils. Initial studies to determine maximum permissible metal
concentrations in biosolids were done using metal salts. When results of these
studies were compared with studies using median metal concentration biosolids,
it became clear that the behaviour of biosolid-borne metals followed a very
different pattern to metals added as salts. Further studies have shown that
biosolid addition to soil enhanced the ability of soil in adsorbing heavy
metals, thereby limiting their bioavailability (McGrath 1994)
Alkaline-stabilised biosolids are increasingly being used as an agricultural
lime substitute, soil amendment, and surrogate soil. Alkaline stabilisation of
biosolid utilises a combination of high pH, heat, and drying to kill pathogens
and stabilise organic matter. A range of alkaline materials are used for this
purpose, including, cement kiln dust, lime kiln dust, lime, limestone, alkaline
coal fly ash, FGD, FBA, other coal burning ashes, and wood ash (Basta 2000).
Logan and Harrison (1995) examined the value of a commercial
alkaline-stabilised biosolid product called 'N-Viro' soil as a soil substitute.
N-Viro is produced by heat treatment of a mixture of cement kiln dust and
municipal sewage sludge. Addition of this material improved the physical and
chemical properties of a degraded mine soil. Such alkaline materials are
effective in reducing the acidity produced during the nitrification of
N[H.sub.4.sup.+] in biosolids, thereby reducing the bioavailability of heavy
metals in biosolid-amended soils (Brown et al. 1997; Basta 2000).
To minimise metal mobility and bioavailability in biosolid-amended soils,
the USEPA recommends the application of alkaline-stabilised biosolids and other
liming agents to increase the soil pH to [greater than or equal to] 6.5. Although
a number of studies have examined the role of biosolid as a source of metal
contamination in soil (Page 1974; Juste and Mench 1992), only limited work has
been reported on the beneficial effect of organic amendments as a sink for the
immobilisation of metals in soils (Brown et al. 1997; Basta 2000).
Animal manure
With the continuous decline in the availability of land area for crop
production, the increase in food demand is likely to be met mainly through
intensive animal production. Confined animal agriculture (i.e. beef cattle,
dairy, poultry, and swine) is the major source of manure by-products in most
countries. For example, of about 900 million Mg organic and inorganic
agricultural recyclable by-products generated in US, approximately 45.4 million
Mg are dairy and beef cattle manure and 27 million Mg are poultry and swine
manure. These manure by-products generate annually about 7.5 million Mg of N
and 2.3 million Mg of P, compared with 9 million Mg of N and 1.6 million Mg of
P applied to agricultural land in the form of commercial fertilisers (Walker et
al. 1997). In addition to this, in Australia
and New Zealand,
where open grazing is followed, a large amount of manure is directly deposited
onto pasture land (Haynes and Williams 1993).
The manure byproducts have the potential for being recycled on agricultural
land and optimum use of these byproducts requires knowledge of their
composition not only in relation to beneficial use but also to environmental
implications. Maintaining the quality of the environment is the major
consideration when developing management practices to effectively use manure
by-products as a nutrient resource and soil conditioner in agricultural
production system.
Most manure products contain low levels of heavy metals (except Cu and Zn in
swine manure and As in poultry manure). Furthermore, recent advances in the
treatment of manure by-products have resulted in reduced bioavailability of
metals. For example, Westerman and Bicudo (2000) observed an 87% reduction in
Cu and Zn in the waste water from swine houses after treatment with lime
slurry, ferric chloride, or polymer. Similarly, Moore et al. (1998) observed
treatment of poultry manure with alum [A12(SO4)3] decreased the concentration
of water-soluble Zn, Cu, and Cd. Hence, unlike sewage sludge application, where
land application is limited based on allowable trace element loadings (USEPA
1999), regulations governing livestock and poultry manure by-products are
generally based on total N and P loading. Manure by-products that are low in
metal content can be used to immobilise metal contaminants in soils.
Sources of heavy metals
In terrestrial ecosystems, the soil is the main repository of contaminant
chemicals. Likewise in aquatic systems, the sediment serves as the ultimate
sink for these chemicals. Heavy metals reach the soil environment through both
pedogenic (or geogenic) and anthropogenic processes. Most metals occur
naturally in soil parent materials, chiefly in forms that are not available for
plant uptake. Because of their low bioavailability, the metals present in most
of the parent materials are often not available for plant uptake and cause
minimum impact to soil organisms. Often the concentrations of metals released
into the soil system by the natural pedogenic (or weathering) processes are
largely related to the origin and nature of the parent material. Apart from As
(Naidu and Skinner 1999), Cd (Singh et al. 1995), and Se (Dhillon and Dhillon
1990), other elements (e.g. Cr, Ni, Pb) derived via geogenic processes have
limited impact on soil. Unlike pedogenic inputs, metals added through
anthropogenic activities typically have high bioavailability (Naidu et al.
1996). Anthropogenic activities, primarily associated with industrial
processes, manufacturing, and the disposal of domestic and industrial waste
materials, are the major source of metal enrichment in soils (Adriano 2001)
(Table 2). Atmospheric pollution from Pb-based petrol is a major issue in many
developing countries where there is no constraint on the usage of leaded
gasoline. While sewage sludge is the major source of metal inputs in Europe and
North America, phosphate fertilisers are considered to be the major source of
heavy metal input, especially Cd, in Australia
and New Zealand.
Phosphate compounds contain a range of metals (McLaughlin et al. 1996;
Mortvedt 1996). According to Nriagu (1984), 'virtually every known element has
been found, at least in trace amounts, in a phosphate mineral'. Addition of P
compounds to soils not only helps to overcome the deficiency of some of the
essential trace elements, such as Mo, but also introduces toxic metals, such as
Cd (McLaughlin et al. 1996). In this regard Cd contamination of agricultural
soils is of particular concern because this metal reaches the food chain
through regular use of Cd-containing P fertilisers. This is one of the main
reasons why this element has been studied extensively in relation to soil and
plant factors affecting its bioavailability.
Accumulation of Cd in soils through regular fertiliser use has been observed
in many countries. For example, in New Zealand
and Australia,
most of the Cd accumulation in pasture soils has been derived from the use of P
fertilisers containing high Cd concentration (Roberts et al. 1994). The Cd in
most P fertilisers originates mainly from the PRs used for manufacturing P
fertilisers. It is important to stress that PRs deposits vary in their Cd
content, leading to the variation in Cd contents of manufactured P fertilisers.
The Cd in superphosphates is water-soluble and high analysis P fertilisers,
such as TSP, PAPR, and ammonium phosphates, generally contain lower Cd content
relative to P.
Comparison between native (i.e. unfertilised) and agricultural (i.e.
fertilised) soils has often been used to indicate contamination of soil through
agricultural practices. Roberts et al. (1994) conducted a survey of 398 sites
throughout New Zealand
with 312 farms sites and 86 native sites to a depth of 75 mm (Table 3). They
obtained evidence for the enrichment of Cd in pastoral soils and there was a
highly significant correlation between total soil P and total soil Cd across
all sites, confirming the role of P fertilisers in soil Cd enrichment. Similar
results were also obtained for a range of Australian soils (Table 3) and Norwegian
soils with different history of P fertilisation (He and Singh 1994). This is
not surprising considering the long history of use of superphosphates in both New Zealand and Australia, manufactured from Nauru
Island PR. Nauru superphosphates typically contain 34-69 mg Cd/kg (Rothbaum et
al. 1986).
Loganathan et al. (1996) examined the movement and distribution of Cd and P
in a pastoral soil amended annually for 10 years with 4 forms of P fertilisers
[SSP, DAP, NCPR, and Jordan PR (JPR)], which varied in their total Cd content.
Both total and plant-available Cd concentrations decreased with soil depth.
Single superphosphate and NCPR, which had higher Cd contents, produced higher
Cd concentration than DAP, JPR, and control treatments. Approximately 93% of the
applied Cd was recovered in the top 120 mm soil and plant recovery of applied
Cd ranged from 1.5 to 4.5%. Similarly, Gray et al. (1999b) found that the rate
of Cd accumulation in the 0-75 mm depth for an annual application of SSP at a
rate of 376 kg/ha for 44 years to irrigated pasture was estimated at 7.8 g
Cd/ha.year.
Although many countries have formulated threshold levels for Cd and other
heavy metal accumulation in soils due to the use of municipal sewage sludge,
such limits have not been established for fertiliser use. Based on the
threshold level for sewage application (3 mg Cd/kg soil), the number of years
required to exceed the threshold level in soil through addition of various
sources of P fertiliser is presented in Table 4. This indicates that although
fertiliser addition represents the major source of Cd input to soils, at the
normal annual rate of fertiliser input (40 kg P/ha) to pasture soils the rate
of Cd accumulation appears to be very slow.
There have been increasing efforts in reducing the accumulation of Cd in
soils through the use of low Cd-containing P fertilisers. This is achieved by
either selective use of PRs with low Cd or treating the PRs during processing
to remove Cd. Superphosphate fertiliser manufacturers in many countries are
introducing voluntary controls on the Cd content of P fertilisers. For example,
the fertiliser industry in New
Zealand has achieved its objective of
lowering the Cd content in P fertilisers from 340 mg Cd/kg P in the 1990s to
280 mg Cd/kg P by the year 2000. A number of PRs with low Cd contents are
available that can be used for the manufacture of P fertilisers, but sources
with higher Cd contents continue to be used in many countries for practical and
economic reasons. Alternatively, since Cd has a low boiling point (BP =
789[degrees]C) it can be removed by calcining the PRs. Phosphoric acid used in
the food industry is manufactured mostly only after the removal of Cd through
calcination of the PRs. Calcination of PRs may not become a likely option in the
fertiliser industry because it is expensive and calcination decreases the
reactivity of PRs, making them less suitable for direct application as a source
of P (Ando 1987).
Case studies for immobilisation of metals using soil amendments
In this section, 6 previously published case studies, illustrating the
potential value of phosphate anion, liming materials, and biosolids in the
immobilisation and the consequent reduction in the phytoavailability of Cd, Cr,
and Cu are presented. The properties of the soils used in these studies are
presented in Table 5. The experimental details, major observations obtained in
these case studies, and the changes in dry matter yield, plant tissue metal
concentration, and soil metal fractions resulting from soil amendments are
given in Tables 6, 7, and 8, respectively. The mechanisms for the
immobilisation of heavy metals in these case studies are discussed in relation
to other published data.
Case study 1: Phosphate-induced cadmium immobilisation
In this study, the effect of phosphate on the surface charge and Cd
adsorption was examined in 7 soils (Table 5) that varied in their
variable-charge components (Bolan et al. 2003c). The effect of phosphate on
immobilisation and phytoavailability of Cd from one of the soils, treated with
various levels of Cd as Cd(NO3)2, was evaluated using mustard (Brassica juncea
L.) plants.
Results indicated that phosphate immobilised Cd, thereby effectively
reducing the phytotoxicity of Cd (Tables 7 and 8). Phosphate-induced
immobilisation of Cd in soils could be explained by: (i) phosphate-induced
[Cd.sup.2+] adsorption; and (ii) precipitation of Cd as Cd[(OH).sub.2] and
[Cd.sub.3][(P[O.sib.4]).sub.2]. Several mechanisms can be advanced for
phosphate-induced [Cd.sup.2+] adsorption observed in this study. These include:
(i) increase in pH; (ii) increase in surface charge; (iii) co-adsorption of
phosphate and Cd as an ion pair; and (iv) surface complex formation of Cd on
the phosphate compound.
Levi-Minzi and Petruzzelli (1984) observed that phosphate-induced variation
in soil pH influenced the solubility of Cd in soils. They noticed that while
the effect of phosphate on pH and Cd solubility was less evident in an organic
soil with high pH buffering capacity, the addition of DAP increased soil pH,
thereby reducing the solubility of Cd in a mineral soil with low pH buffering
capacity. A number of studies have shown that specific adsorption of anions
increases the net negative charge of variable charge surfaces (Bolland et al.
1977; Naidu et al. 1996), thereby increasing the retention of metal cations,
such as C[d.sup.2+], C[u.sup.2+], and Z[n.sup.2+] (Bolland et al. 1977; Bolan
et al. 1999b). Specifically sorbed anions, such as phosphate, form complexes
with the soil surface so that cations are adsorbed onto the adsorbed anions
(Helyar et al. 1976; Bolland et al. 1977). Hence, surface complexation has also
been suggested as a mechanism for the immobilisation of metals by
hydroxyapatite (Xu et al. 1994) (Eqn 4):
[Ca.sub.10][(P[O.sub.4]).sub.6][(OH).sub.2+] [right arrow]
(C[d.sub.x],[Ca.sub.10][(P[O.sub.4])].sub.6][(OH)].sub.2] + x[Ca..sup.2+] (4)
Precipitation as metal phosphates has also been proven to be a main
mechanism in immobilising metals, such as Pb and Zn, by phosphate compounds
(Street et al. 1978; Pierzynski and Schwab 1993). The formation of the solid
phase (i.e. precipitates) occurs when the ion activity product in the solution
exceeds the solubility product of that phase. In typical soils, precipitation
of metals is unlikely, but in highly metal-contaminated soils this process can
play a major role in metal immobilisation.
Although there was no direct evidence for [Cd.sub.3][(P[O.sub.4]).sub.2]
formation even at the highest level of phosphate and Cd addition in the soil
samples used in the case study, it did not preclude the formation of mixed
Ca-Cd phosphate or the amorphous Cd phosphate compounds with different
solubility product. It is important to note that the allophanic soil used in
the glasshouse experiment adsorbed a very high amount of phosphate, thereby
maintaining a very low concentration of phosphate in soil solution. In general,
the solubility of [Cd.sub.3][(P[O.sub.4]).sub.2] has been shown to be too high
to control the concentration of Cd in soils (Soon 1981). However, McGowen et
al. (2001) observed that application of DAP at high level (2300 mg P/kg) was
found to be very effective in immobilising Cd, Pb and Zn from a
smelter-contaminated soil. Others have also shown that
[Cd.sub.3][(P[O.sub.4]).sub.2] can control Cd solubility in phosphate-enriched
soils (Street et al. 1978).
Many investigators have provided conclusive evidence for the ability of
phosphate to immobilise dissolved Pb in contaminated soils through
precipitation as fluoropyromorphite, pyromorphite, hydroxypyromorphite, and
chloropyromorphite, and as hopeite in the case of Zn (Bolan et al. 2003b).
Similarly, Seaman et al. (2001) indicated that the decrease in the solubility
of a range of metals in the presence of hydroxyapatite is caused by the
formation of secondary metal phosphate precipitates rather than metal
adsorption by weathered apatite crystals.
Case studies 2 and 3: Lime-induced cadmium and chromium immobilisation
The effect of 3 liming materials [FBA, Ca[(OH).sub.2], and CaC[O.sub.3]] on
the immobilisation and phytoavailability of Cd and Cr(III) was evaluated using
mustard (Bolan et al. 2003d) and sunflower (Helianthus annuus) plants (Bolan
and Thiagarajan 2001), respectively. Results indicated that although the
addition of Ca[(OH).sub.2] effectively reduced Cd phytotoxicity (Tables 7 and
8), Cd uptake increased at the highest Ca[(OH).sub.2] level, probably due to
decreased [Cd.sup.2+] adsorption resulting from increased [Ca.sup.2+]
competition (Naidu et al. 1996). FBA and CaC[O.sub.3] were found to be
effective in immobilising Cr(III), thereby reducing phytotoxicity (Tables 7 and
8).
Liming as part of the normal cultural practices has often been shown to
reduce the concentration of Cd and other metals in the edible parts of a number
of crops. Addition of alkaline waste materials, such as coal fly ash, has also
been shown to decrease Cd content of plants (Knox et al. 2000). The effect of
liming materials in decreasing Cd uptake has been attributed to both decreased
mobility of Cd in soils and competition between [Ca.sup.2+] and [Cd.sup.2+]
ions on the root surface. In general, Cd uptake by plants decreases with
increasing pH. For example, higher Cd concentrations were obtained for lettuce
and Swiss chard on acid soils (pH 4.8-5.7) than on calcareous soils (pH
7.4-7.8) (Mahler et al. 1978). Consequently, it is recommended that soil pH be
maintained at pH [greater than or equal to]6.5 in land receiving biosolids
containing Cd (Adriano 2001). However, it is also possible that in alkaline
soils, mobilisation of Cd can be enhanced due to facilitated complexation of Cd
with humic or organic acids (Harter and Naidu 1995).
Various reasons have been advanced for pH-induced immobilisation of metals
in soils. Firstly, an increase in pH in variable-charge soils causes an
increase in surface negative charge, resulting in an increase in cation
adsorption (James and Bartlett 1983; Naidu et al. 1994). Secondly, an increase
in soil pH is likely to result in the formation of hydroxy species of metal
cations (e.g. CdO[H.sup.+]) that have a greater affinity for adsorption sites
than just the metal cation (Naidu et al. 1994). And thirdly, precipitation of
Cd as Cd[(OH).sub.2] is likely to result in greater retention at pH > 10
(Naidu et al. 1994). It is important to stress that liming is unlikely to raise
the soil pH >8.3, whereas Cr is likely to precipitate at pH >5.5 as
Cr[(OH).sub.3] (Rai et al. 1987). The pH of the FBA- and lime-treated soils
ranged from 7.18 to 8.04, which coincides with the effective precipitation
range for Cr(III) as Cr[(OH).sub.3].
Soil solution pH is one of the major factors controlling surface properties
of variable charge components (Barrow 1985). An increase in pH increases the net
negative charge, which is attributed to the dissociation of H+ from weakly
acidic functional groups of organic matter and some clay minerals (Thomas and
Hargrove 1984). The amount of surface charge acquired through an increase in pH
depends on the amount and nature of variable charge components (Bolan et al.
1999b). The surface charge of the soil mineral component is generally far less
pH-dependent than that of soil organic matter. However, the pH-dependence of
mineral surface charge can vary considerably depending on the nature of the
component minerals (Thomas and Hargrove 1984).
Attempts have been made to relate the pH-induced increases in surface charge
to [Cd.sup.2+] adsorption by variable charge soils (Naidu et al. 1994; Bolan et
al. 1999b). For example, Bolan et al. (1999b) observed that approximately 50%
of the pH-induced increase in surface negative charge in variable charge soils
was occupied by Cd. The remaining surface negative charge was presumed to be
occupied by the [H.sup.+] and [K.sup.+] ions, added in acid and alkali
solutions to alter the soil pH. Similarly, Naidu et al. (1994) demonstrated
that the effects of ionic strength and specifically adsorbed anions on
[Cd.sup.2+] adsorption operate partly through their effects on surface charge.
In limed soil, the activities of free [Cd.sup.2+] and OH ions, and
C[O.sub.2] partial pressure, control the precipitation of Cd as CdC[O.sub.3]
(octavite) and Cd[(OH).sub.2] (Street et al. 1978). Street et al. (1978)
obtained evidence for precipitation of Cd as CdC[O.sub.3] only in a sandy soil
having low organic matter and low CEC. In another instance, Soon (1981)
examined the effect on the solubility of Cd in two soils of a number of sewage
sludges that had been treated with Ca[(OH).sub.2] [Al.sub.2][(S[O.sub.4).sub.3],
or Fe[Cl.sub.3] to precipitate phosphate from effluent water. At low levels of
Cd addition, the solubility of Cd was controlled by adsorption that was
enhanced by increasing pH resulting from the sludge addition. At high levels of
Cd addition, however, there was evidence for the precipitation of Cd as
[Cd.sub.3][(P[O.sub.4]).sub.2] and CdC[O.sub.3], which controlled the
solubility.
Case studies 4 and 5: Organic amendment-induced cadmium and copper
immobilisation
The effect of a number of organic amendments on the adsorption and
complexation of Cd (Bolan et al. 2003e) and Cu (Bolan et al. 2003f) was
examined. The effect of one of the amendments (i.e. biosolid) on the uptake of
Cd and Cu was also examined using mustard plants.
Addition of organic amendments increased the complexation of Cd and Cu by
soils. Dissolved organic carbon in the organic amendments formed soluble
complexes with these metals. Addition of biosolid was effective in reducing the
phytotoxicity at all levels of Cd addition but only at high levels of Cu
addition (Tables 7 and 8).
It has often been observed that plants exhibit greater tolerance to metals
introduced through sewage sludge addition than when they are added as inorganic
salt. For example, Chang et al. (1992) and Logan et al. (1997) presented data
for maize and other crops, grown on metal-contaminated sludge-amended soils,
which revealed inconsequential change in tissue Cd concentrations in response
to substantial increases in total Cd loading in soils. The decrease in the
phytoavailability of metals in the presence of organic amendments is often
attributed to increased complexation of the metal by the organic constituents
(Adriano 2001). However, the presence of phosphates, aluminium compounds, and
other inorganic minerals in typical municipal sewage sludge is also believed to
be responsible for inducing the 'plateau effect' in Cd uptake by crops, thereby
preventing the increased Cd availability suggested in the 'time bomb'
hypothesis (Brown et al. 1998).
It has often been found that in soils containing large amounts of organic
matter, such as pasture soils and organic manure-amended soils, a large portion
of soil solution Cd is complexed with dissolved organic carbon (DOC) (del
Castilho et al. 1993; Sauve et al. 2000). Similarly, Hyun et al. (1998)
obtained linear relationship between organic carbon and soluble Cd in solution
for sludge-treated soils, indicating that majority of the Cd remained as
metal-organic complex. Although a wide variety of organic compounds in DOC are
involved in the formation of soluble complex with metals, Zhou and Wong (2001)
and del Castilho et al. (1993) observed that low molecular fractions, such as
hydrophilic bases, have strong affinity to form soluble Cd complexes.
Similarly, Riffaldi et al. (1983) obtained significant correlations between Cd
sorption and phenolic hydroxyl groups and carboxyl groups of fulvic acids.
It has often been observed that in soils treated with organic amendments Cu
is associated more with organic fractions than with other fractions. For
example, Keefer et al. (1984) fractionated the metal organic components
extracted from a sludge-amended soil and found that the strongly bound Cu is
associated with hydrophobic acids (phenols) and hydrophobic neutrals (hydrocarbons).
On the other hand, the weakly bound Cu was complexed with hydrophilic neutrals
(carbohydrates). Although the DOC-induced decrease in the adsorption of Cu by
the soils in their study may lead to increased mobility of Cu, it does not
necessarily increase the bioavailability of Cu.
Case study 6: Organic amendment-induced chromium reduction
In this study, 7 organic amendments were investigated for their effects on
the reduction of chromate [Cr(VI)] in a mineral soil low in organic matter
content (Bolan et al. 2003g). The effect of biosolid compost on the uptake of
Cr(VI) from the soil, treated with various levels of Cr(VI), was examined using
mustard plants.
Addition of organic amendments enhanced the rate of reduction of Cr(VI) to
Ct(III) in the soil (Table 9), thereby reducing phytotoxicity (Tables 7 and 8).
Various reasons could be presented for the increase in the reduction of Cr(VI)
in the presence of organic manure compost. These include the supply of carbon,
protons, and microorganisms that are considered to be the major factors
enhancing the reduction of Cr(VI) to Cr(III) (Losi et al. 1994). The extent of
Cr(VI) reduction increased with increasing level of easily oxidisable carbon
and DOC added through manure addition, and there was a significant linear
relationship between the extent of Cr(VI) reduction and DOC (Fig. 1). Dissolved
organic carbon has been identified to have facilitated the reduction of Cr(VI)
to Cr(III) in soils (Jardine et al. 1999).
[FIGURE 1 OMITTED]
Based on the reaction between organic carbon and Cr(VI) reduction (Eqn 5),
it is estimated that 1.00 mg of organic carbon causes a reduction of 5.78 mg
Cr(VI) (Adriano 2001).
2[Cr.sub.2][O.sub.7] + 3[C.sup.0] + 16[H.sup.+] [right arrow] 4[Cr.sup.3+] +
3C[O.sub.2] + 8[H.sub.2]O
However, the linear regression indicated that only a small fraction of DOC
is used as an energy source for the reduction of Cr(VI) to Cr(III). This
suggests that only certain components of the organic carbon act as electron
donor for the reduction of Cr(VI) to Cr(III). For example, in natural organic
matter the hydroquinone groups 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).
It has often been observed that Cr(VI) reduction, being a proton consumption
(or hydroxyl release) reaction (Eqn 5), increases with a decrease in soil pH
(Eary and Rai 1991). The organic amendments are rich in ammoniacal nitrogen,
which is likely to result in the release of protons during subsequent
nitrification and ammonia volatilisation. The increase in Cr(VI) reduction in
the presence of organic amendments may also result from an increase in
microbial activity. Losi et al. (1994) have shown that the addition of a manure
compost caused a larger increase in the biological reduction than the chemical
reduction of Cr(VI), which suggests that the supply of microorganisms is more
important than the supply of organic carbon in enhancing the reduction of
Cr(VI) with the addition of organic compost. Addition of organic manure compost
has often been shown to increase the microbial activity of soil, as measured by
increased respiration (Kanazawa et al. 1988). An increase in microbial activity
has often been reported to increase the reduction of Cr(VI) to Cr(III) (Losi et
al. 1994). Although Cr(VI) reduction can occur through both chemical and
biological processes, the biological reduction is considered to be the dominant
process in most agricultural soils, which are low in ferrous ([Fe.sup.2+]) ion.
Conclusions
Results from these case studies indicated that the addition of soil
amendments, such as phosphate compounds, liming materials, and biosolid
amendments to metal-contaminated sites, reduced the phytoavailability of
metals. Since bioavailability is the key point for remediation technologies,
immobilisation may be a preferred option (Mench et al. 1994; James 1996). A
major inherent problem associated with immobilisation techniques is that
although the heavy metals become less bioavailable, their total concentration
in soils remains unchanged. The immobilised heavy metal may become
plant-available with time through natural weathering process or through
breakdown of high molecular weight organic-metal complexes. For example, Stacey
et al. (2001) have observed that the rate of release of Cd and Zn from a range
of biosolids during the decomposition of organic matter in the biosolids
depends, to a large extent, on the chemical composition of the biosolids.
However, Hyun et al. (1998) obtained no evidence for increased
phytoavailability of Cd with the breakdown of organic matter in sludge-treated
soils. Furthermore, recently Li et al. (2001) observed evidence for greater
affinity for Cd adsorption by the inorganic components of the biosolid-amended
soils indicating that the increased adsorption of Cd is independent of the
added organic matter and of a persistent nature.
Although the formation of a soluble metal-organic complex reduces the
phytoavailability of metals, the mobility of the metal may be facilitated greatly
in soils receiving alkaline-stabilised biosolid because of the reduction of
metal adsorption and increased concentration of soluble metal-organic complex
in solution (Brown et al. 1997; Gove et al. 2001).
Metal phytotoxicity in soils is determined by the fraction of the metal that
is bioavailable. This has implications for our current regulatory policies,
which are generally based on total metal content (McLaughlin et al. 2000;
Adriano 2001). It is important to emphasise that there is a dynamic equilibrium
amongst various fractions in soils and any depletion of the available pool
(soluble and exchangeable fractions) due to immobilisation, plant uptake, or
leaching losses will result in the continuous release from other fractions to
replenish the available pool. This is one of the main reasons why there is some
reluctance towards using bioavailable pool in soils for regulatory purposes by
environmental agencies in monitoring contaminated sites. In addition, the
bioavailable pool is sensitive to edaphic and environmental conditions as
solubilisation of metals from sparingly soluble compounds responds to soil pH,
redox potential, temperature, etc. However, use of the isotopic dilution
technique to estimate the exchangeable pool (E value) and labile pool (L value)
has enabled a relatively realistic determination of metal bioavailability in
soils compared with methods using chemical extractants (Stacey et al. 2001).
Numerous heavy metal contaminated sites have been reported in New Zealand and
other countries. These include the cadmium (Cd) contamination in pasture soils
resulting from continuous use of phosphate fertilisers, chromium (Cr)
contamination in timber treatment plants, and Cu contamination in orchards.
Field trials need to be set up in these sites to examine the potential value of
compost and other soil amendments in sequestering and mitigating the phytotoxic
effect of these toxic heavy metals so that more diverse land use can be
facilitated.
Acknowledgments
We would like to thank Drs DC Adriano (University
of Georgia) and R Naidu (University of South Australia) for their valuable
comments on the paper. We would also like to thank Drs P Mani, S Mani, A
Arulmozhiselvan, and R Natesan for their contributions to the glasshouse
experiments discussed in the case studies.
Table 1. Selected references on the immobilisation and bioavailability of cadmium by various soil amendments
LSB, lime stabilised biosolid; AADB, anaerobically digested biosolid; ADB, aerobically digested biosolid; BS, biosolid; SS, sewage sludge;
CM, cattle manure; PM, poultry manure; PMS, paper mill sludge; SSDS, secondary digested sewage sludge
Amendments Observations on References
immobilisation and
bioavailability
Hydroxyapatite Increased immobi- Jeanjean et a(1995),
lisation through Mandjiny etal(1998),
cation exchange, Xu et al. (1994),
adsorption, surface Boisson eta(1999),
complexation, Seamanet al. (2001)
precipitation and
co-precipitation
Rock phosphate Increased immobi- Basta et al. (2001),
lisation through Chen et al. (1997)
adsorption and
precipitation
[K.sub.2] Increased immobi- Pierzynski & Schwab
HP[O.sub.4] lisation through (1993)Pearson
phosphate-induced et al. (2000)
adsorption and
precipitation
K[H.sub.2] Increased immobi-Bolan et al.(1999b),
P[O.sub.4] lisation through Naidu et al.(1994)
phosphate-induced
adsorption
Ca[([H.sub.2] Increased immobi- Bolan et al.(1999b)
P[O.sub.4]) lisation through
.sub.2] phosphate-induced
adsorption
[(N[H.sub.4])Increased adsorption Pierzynski & Schwab
.sub.2] due to an increase (1993), Levi-Minzi
HP[O.sub.4]in pH, precipi- and Petruzzelli (1984),
tation of McGowen et al. (2001)
[Cd.sub.3][(P
[O.sub.4]).sub.2]
CaC[O.sub.3]Increased immobi- Andersson and Siman
lisation through (1991), Bingham et al.
adsorption and (1979), Brown et al.
precipitation; (1997), Han and Lee
decreased plant (1996), Chaney et al.
uptake (1977), He and Singh
(1994), Hooda and
Alloway (1996), John
and van Laerhoven
(1976), John et al.
(1972), Lehoczky et al.
(2000), Maclean (1976),
Oliver et al. (1996),
Singh and Myhr (1998),
Singh et al. (1995),
Tyler and Olsson (2001)
Ca[(OH).sub.2]Decreased bio- Basta and Sloan (1999),
availability Chaney et al. (1977),
Brallier et al. (1996),
Gray et al. (1999a)
CaO Decreased phyto- Vasseur et al. (1998)
availability
MgC[O.sub.3]Decreased phyto-Williams and David (1976)
availability
CaMgC[O.sub.3]Decreased phyto- Kreutzer (1995)
availability
Milorganite Decreased phyto- John and van
availability Laerhoven (1976)
LSB, N-Viro;Adsorption by Basta et al. (2001),
ADB, AADB inorganic compo- Basta and Sloan (1999),
nents, metal- Keefer et al. (1984),
organic matter Pietz etal. (1983),
complex formation Soon (1981)
SS Increased adsorption Hyun et al. (1998),John
and complexation and van Laerhoven(1976), Street et al. (1978)
BS Increased the Brown et al. (1998),
affinity of Li et al. (2001)
inorganic fraction
of BS treated soil
for Cd adsorption
CM Cd in soil solution del Castilho et al.
was bound in fast (1993)
dissociating metal
complexes
PMS and BS PMS and BS decreased Merrington and
Cd adsorption, but Madden (2000)
increased Cd in
ryegrass
Composted Increased bio- Pearson et
leaves accumulation of Cd al. (2000)
by earthworm
CM, PM,Increased Cd in the Pierzynski & Schwab (1993)
N-Viro organic fraction
SDSS Increased adsorption Riffaldi et al. (1983)
Table 2. Sources of heavy metals in soils and their expected ionic species in soil solution
Source: Adriano (2001)
Metal Density Ionic species in soil solution
(g/[cm.sup.3])
Arsenic (As) 5.73 As(III): As[(OH).sub.3],
As[O.sub.3.sup.3-];
As(V): [H.sub.2]
[As.sub.4.sup.-],
HAs[O.sub.4.sup.2-]
Cadmium (Cd) 8.64 [Cd.sup.2+], CdO[H.sup.+],
Cd[Cl.sup.-],
CdHC[O.sub.3+],
CdS[O.sub.4.sup.0]
Chromium (Cr) 7.81 Cr(III): [Cr.sup.3+],
Cr[O.sub.2.sup.-],
CrO[H.sup.2+],
Cr[(OH).sub.4.sup.-];
Cr(VI): [Cr.sub.2]
[0.sub.7.sup.2-],
Cr[O.sub.4.sup.2-]
Copper (Cu) 8.96 [Cu.sub.2+] (II),
[Cu.sub.2+] (III)
Lead (Pb) 11.35 [Pb.sup.2+], PbO[H.sup.+],
Pb[Cl.sub.-],
PbHC[O.sub.3.sup.+],
PbS[O.sub.4.sup.0]
Manganese 7.21 [Mn.sup.2+], MnO[H.sup.+],
(Mn) Mn[C1.sup.-],
MnC[O.sub.3.sup.0],
MnHC[O.sub.3.sup.+],
MnS[O.sub.4.sub.0]
Mercury (Hg) 13.55 [Hg.sub.2+], HgO[H.sup.+],
Hg[Cl.sub.2.sup.0],
C[H.sub.3][Hg.sup.+],
Hg[(OH).sub.2.sub.0]
Molybdenum 10.2 Mo[O.sub.4.sup.2-],
(Mo) HMo[O.sub.4.sup.-],
[H.sub.2]Mo[O.sub.4.sup.0]
Nickel (Ni) 8.90 [Ni.sub.2+], NiS[O.sub.4.
sup.0], NiHC[O.sub.3.sup.+]
NiC[O.sub.3.sup.0]
Zinc (Zn) 7.13 [Zn.sub.2+], ZnS[O.sub.4.
sup.0], Zn[Cl.sup.+],
ZnHC[O.sub.3.sup.+],
ZnC[O.sub.3.sup.0]
Metal Contaminant sources Toxicity (A)
Arsenic (As) Timber treatment, paints, Toxic to plants, humans,
pesticides, geothermal and animals
Cadmium (Cd) Electroplating, batteries, Toxic to plants, humans,
fertilisers and animals
Chromium (Cr) Timber treatment, leather Cr (VI) toxic to plants,
tanning, pesticides, humans, and animals (B)
dyes
Copper (Cu) Fungicides, electrical, Toxic to plants, humans,
paints, pigments, and animals
timber treatment,
fertilisers, mine
tailings
Lead (Pb) Batteries, metal products, Toxic to plants, humans,
preservatives, petrol and animals
additives
Manganese Fertiliser Toxic to plants
(Mn)
Mercury (Hg) Instruments, fumigants, Toxic to humans and
geothermal animals
Molybdenum Fertiliser Toxic to animals
(Mo)
Nickel (Ni) Alloys, batteries, mine Toxic to plants, humans,
tailings and animals
Zinc (Zn) Galvanising, dyes, paints, Toxic to plants
timber treatment,
fertilisers, mine
tailings
(A) Most likely to observe at elevated concentrations in soils and
water.
(B) Though Cr(VI) is very mobile and highly toxic, Cr(III) is
essential in animal and human nutrition and
generally immobile in the environment.
Table 3. Cadmium concentration (mg/kg) in unfertilised and fertilised
surface soils in Australia and New Zealand
Source: Williams and David (1976); Roberts et al. (1994)
Soil type Unfertilised Fertilised
Australia
Red brown earth 0.055 0.12
Red podzolic 0.024 0.085
Krasnozem 0.030 0.30
Alluvial 0.14 0.27
Podzol 0.033 0.34
New Zealand
Alluvial 0.13 0.16
Brown grey loam 0.19 0.49
Gley 0.24 0.42
Peat 0.22 0.69
Yellow brown earth 0.16 0.22
Yellow brown loam 0.23 0.70
Yellow brown peat 0.31 0.75
Yellow grey earth 0.13 0.12
Table 4. Phosphorus and cadmium concentrations (g/kg) in various
phosphate fertilisers and the calculated number of years required to
exceed the threshold concentration of Cd (3 mg Cd/kg) in soils due to
fertiliser application
Source: Bolan et al. (2003b)
Phosphate fertiliser Phosphorus Cadmium Years required
to exceed
the threshold
limit (A)
Single superphosphate 98 0.032 166
Triple superphosphate 190 0.070 152
Diammonium phosphate 200 0.010 1125
North Carolina phosphate rock 132 0.054 135
Sechura phosphate rock 131 0.012 614
Egyptian phosphate rock 130 0.010 732
Gafsa phosphate rock 134 0.070 107
(A) At an annual fertiliser application rate of 40 kg P/ha.
Table 5. Characteristics of soils used in the case studies
Source: Bolan et al. (2003c)
Soil Soil Classification Organic carbon pH
locations (g/kg)
Ballantrae Typic Dystrandept 58.5 5.62
Egmont Typic Dystrandept 78.5 5.85
Foxton Dystric Fluventic Eutrochrept 23.1 5.85
Manawatu Dystric Fluventic Eutrochrept 29.1 6.01
Patua Typic Dystrandept 89.7 6.12
Tokomaru Typic Fragiaqualf 34.3 5.67
Ramiha Typic Dystrandept 56.2 5.75
Soil P retention CEC
locations (%) (cmol/kg)
Ballantrae 42 18.2
Egmont 83 26.2
Foxton 23 3.52
Manawatu 33 7.6
Patua 95 32.5
Tokomaru 51 11.2
Ramiha 77 22.4
Soil Dominant clay minerals
locations
Ballantrae Mica/illite, chlorite, kaolinite, smectite, vermiculite
Egmont Allophane, volcanic glass, chlorite, kaolinite,
halloysite
Foxton Mica/illite, chlorite
Manawatu Mica/illite, chlorite, smectite, kandite
Patua Allophane, volcanic glass, kandite
Tokomaru Mica/illite, chlorite, kandite, kaolinite, smectite,
vermiculite
Ramiha Allophane, volcanic glass, chlorite, kandite, halloysite
Table 6. Experimental details for the case studies
Case Heavy Amendments Soils
study no. metal
1 Cd K[H.sub.2]P[O.sub.4] Ballantrae,
Egmont
Foxton,
Manawatu,
Patua,
Ramiha
and
Tokomaru,
2 Cd Ca[(OH).sub.2] Egmont and
Tokomaru
3 Cr(III) CaC[O.sub.3], fluidised Egmont and
bed boiler ash (FBA) Tokomaru
4 Cd Biosolid Egmont and
Tokomaru
5 Cu Biosolid, farm yard manure, Manawatu
pig manure, poultry manure Tokomaru
and spent mushroom
6 Cr(VI) Biosolid, farm yard manure, Manawatu
pig manure, poultry manure
and spent mushroom
Case Treatments for plant growth studies
study no.
1 K[H.sub.2]P[O.sub.4]:0-1000 mg P/kg soil
Cd: 0-10 mg/kg soil
2 Ca[(OH).sub.2]:0-180 mmol O[H.sup.-]/kg soil
Cd: 0-10 ms/kg soil
3 CaC[O.sub.3] and FBA: 66 and 200 mmol
O[H.sup.-]/kg soil. Cr(III): 0-3200 ms/kg soil
4 Biosolid: 0-100 g organic carbon/kg soil
Cd: 0-10 mg/kg soil
5 Cu: 0400 mg/kg soil
6 Biosolid: 0-100 g organic carbon/kg soil
Cr(VI): 0-1200 mg/kg
Case Plant species Reference
study no.
1 Brassica juncea Bolan et al. (2003c)
2 Brassica juncea Bolan et al. (2003d)
3 Helianthus Bolan and
annuus Thiyagarajan (2001)
4 Brassica juncea Bolan et al. (2003e)
5 Brassica juncea Bolan et al. (2003f)
6 Brassica juncea Bolan et al. (2003g)
Table 7. Major observations obtained in the case studies
Case Soil reactions and metal fractionation
study
no.
1 Phosphate increased pH, negative charge and
Cd adsorption; the phosphate-induced
effects were more pronounced in the
allophanic than non-allophanic soils,
Phosphate decreased the concentration of
the soluble and exchangeable Cd fraction
but increased the concentration of
inorganic-bound Cd fraction in soil
2 Ca[(OH).sub.2] increased soil pH, thereby
increasing the adsorption of Cd. Ca[(OH).sub.2]
decreased the concentration of the soluble
and exchangeable Cd fraction but
increased the concentration of
inorganic-bound Cd fraction in soil
3 Both CaC[O.sub.3] and FBA increased the retention
of Cr(III). CaC[O.sub.3] and FBA decreased the
concentration of the soluble and
exchangeable Cr fraction but increased the
concentration of inorganic-bound Cr
fraction in soil
4 Organic amendments increased negative
charge and Cd complexation. Biosolid
decreased the concentration of the soluble
and exchangeable Cd fraction but
increased the concentration of
organic-bound Cd fraction in soil
5 Organic amendments increased the
adsorption and complexation of Cu.
Dissolved organic carbon formed soluble
organic Cu complexes. Biosolid decreased
the concentration of the soluble and
exchangeable Cu fraction but increased the
concentration of organic-bound Cu
fraction in soil
6 Organic amendments enhanced the rate of
reduction of Cr(VI) to Cr(III). Biosolid
decreased the concentration of the soluble
and exchangeable Cr fraction but increased
the concentration of organic-bound Cr
fraction in soil
Case Plant growth experiment
study
no.
1 Plant growth decreased with increasing Cd level.
Plant growth at all levels of Cd increased with
increasing level of P. Cd addition increased Cd
concentration in plants from 1.2 to 263 mg Cd/kg
dry matter. Phosphate decreased Cd
concentration in plants
2 Plant growth decreased with increasing Cd level.
Plant growth at all levels of Cd increased with
increasing level of Ca[(OH).sub.2]. There was a slight
decrease in plant growth at the highest level of
Ca[(OH).sub.2]. Cd addition increased Cd concentration
in plants from 2.1 to 275 mg Cd/kg dry matter.
Low levels of Ca(OH)2 decreased Cd
concentration in plants. The highest level of
Ca[(OH).sub.2] caused a slight increase in plant Cd
concentration
3 Plant growth decreased with increasing Cr level.
Plant growth at all levels of Cr increased with
increasing level of CaC[O.sup.3] or FBA. Cr addition
increased Cr concentration in plants from 0.85 to
10.2 mg Ct/kg dry matter. CaC[O.sup.3] and FBA
decreased Cr concentration in plants
4 Plant growth decreased with increasing Cd level.
Plant growth at all levels of Cd increased with
increasing level of biosolid. Cd addition
increased Cd concentration in plants from 1.09 to
285 mg Cd/kg dry matter. Biosolid decreased Cd
concentration in plants
5 Plant growth decreased at high levels of Cu. Plant
growth at high levels of Cu increased with
increasing level ofbiosolid. Cu addition
increased Cu concentration in plants from 3.2 to
187 mg Cu/kg dry matter. Biosolid decreased Cu
concentration in plants
6 Plant growth decreased with increasing Cr level.
Plant growth at all levels of Cr increased with
increasing level ofbiosolid. Cr addition increased
Cr concentration in plants from 0.14 to 24.3 mg
Cr/kg dry matter. Biosolid decreased Cr
concentration in plants
Table 8. Changes in dry matter yield, tissue metal concentration and
soil metal fractions (-, decrease; +, increase over nil amendment)
Metal loading: case study 1, 2, and 4, 10 mg Cd/kg soil; case study 3,
1600 mg Cr/kg soil; case study 5,400 mg Cu/kg soil; case study 6,
600 mg Cr/kg soil. Amendment level: case study 1, 1000 mg P/kg soil;
case study 2, 180 mmol O[H.sup.-]/kg soil; case study 3,200 mmol
O[H.sup.-]/kg soil; case study 4-6, 100 g organic carbon/kg soil
Case Metal Amendment Dry matter Tissue
study yield metal
no. (g/pot) concen-
tration
(mg/kg)
1 Cd K[H.sub.2] +15.4 -228
P[O.sub.4]
2 Cd Ca[(OH).sub.2] +18.7 -254
3 Cr(III) CaC[O.sub.3] +5.7 -6.2
FBA +5.3 -4.8
4 Cd Biosolid +25.6 -253
5 Cu Biosolid +19.8 -157
6 Cr(VI) Biosolid +11.2 -12.4
Soil
fractions
(mg/kg)
Soluble Organic Oxide Residual
plus
exchange-
able
1 Cd -2.02 -0.21 +1.19 +0.17
2 Cd -2.75 +0.50 +1.20 +1.20
3 Cr(III) -77.0 -26.0 +45.0 +131
-85.3 -21.0 +85.0 +24.0
4 Cd -0.38 +1.11 +0.19 +0.03
5 Cu -69.8 +98.0 -17.5 +34.6
6 Cr(VI) -121.8 +163 -18.0 -22.0
Table 9. Parameters of the equation describing the rate of reduction of
Cr(VI) in soils
Y = [Y.sub.m] (1- [Exp.sup.-rx]), where Y is amount of Cr(VI) reduced
(mg/kg), [Y.sub.m] is maximum amount of Cr(VI) reduction (mg/kg), r is
rate constant, and x is incubation period (days). Source: Bolan et al.
(2003g)
Treatment [Y.sub.m] r (rate Relative
(maximum factor) rate of Cr
reduction) reduction
Soil 125.3 0.201 1.00
Soil + biosolid 470.5 0.410 2.04
Soil + farm yard manure 210.7 0.221 1.10
Soil + fish manure 250.6 0.252 1.25
Soil + horse manure 175.5 0.202 1.00
Soil + spent mushroom 310.6 0.280 1.39
Soil + pig manure 320.3 0.305 1.52
Soil + poultry manure 380.6 0.351 1.75
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Manuscript
received I October 2002, accepted 24 February 2003
N.
S. Bolan (A,C) and V. P. Duraisamy (B)
(A)
Institute of Natural Resources, Massey University, Palmerston North, New
Zealand.
(B)
Tamil Nadu Agricultural University, Coimbatore,
Tamil Nadu, India.
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