Role of inorganic and organic soil amendments on immobilisation and phytoavailability of heavy metals: a review involving specific case studies.
Heavy metals (Research)
Soil pollution (Research)
Soil amendments (Research)
Bolan, N. S.
Duraisamy, V. P.
|Publication:||Name: Australian Journal of Soil Research Publisher: CSIRO Publishing Audience: Academic Format: Magazine/Journal Subject: Agricultural industry; Earth sciences Copyright: COPYRIGHT 2003 CSIRO Publishing ISSN: 0004-9573|
|Issue:||Date: May, 2003 Source Volume: 41 Source Issue: 3|
|Topic:||Event Code: 310 Science & research|
|Geographic:||Geographic Scope: New Zealand Geographic Code: 8NEWZ New Zealand|
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.
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.
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.
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.
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).
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.
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.
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.
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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.
(C) Corresponding author; email: N.S.Bolan@massey.ac.nz
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 al. (1995), lisation through Mandjiny et al. (1998), cation exchange, Xu et al. (1994), adsorption, surface Boisson et al. (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 and Schwab HP[O.sub.4] lisation through (1993), Pearson et phosphate-induced 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 and 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 and 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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