soil remediation and plants || remediation of cd-contaminated soils

571Soil Remediation and Plants. http://dx.doi.org/10.1016/B978-0-12-799937-1.00020-6Copyright © 2015 Elsevier Inc. All rights reserved.

Remediation of Cd-Contaminated Soils: Perspectives and Advancements

Syed Hammad Raza,* Fahad Shafiq,* Umer Rashid,† Muhammad Ibrahim‡ and Muhammad Adrees‡

*Department of Botany, Government College University, Faisalabad, Pakistan, †Institute of Advanced Technology, Universiti Putra Malaysia, Serdang, Selangor, Malaysia, ‡Department of Environmental Sciences, Government College University, Faisalabad, Pakistan

BACKGROUND AND INTRODUCTION

Cadmium metal is a persistent environmental toxicant (di Toppi and Gabbrielli, 1999; Zulfiqar et al., 2012) and its inclusion in terrestrial communities is increas-ing at alarming rates. Sustainability of our ecosystem is highly questionable due to rapid induction of high quantities of this toxic heavy metal since the human industrial revolution 200 years or so. Natural resources are also the major con-tributors in the emission of cadmium into terrestrial commodities. However, release of this metal from unnatural sources is commonly more mobile than those from native natural origins (Chlopecka et al., 1996; Nordic Council of Ministers, 2003). As the most toxic heavy metal among all the others, it has no known biological function in aquatic and terrestrial living organisms, includ-ing humans (Chen et al., 2007). The global annual production of cadmium was 22,200 metric tons in 2011 (Tolcin, 2012); and for humans, the weekly toler-able limit of Cd intake established by WHO is 7 μg kg−1 of total body weight (FAO / WHO, 2003).

CADMIUM EMISSIONS

Soil cadmium contamination is principally derived from natural and anthro-pogenic sources (De Meeus et al., 2002; Silvera and Rohan, 2007). Elemental forms of Cd with its principal minerals like octavite (CdSe), monteponite (CdO) and greenockite (CdS) are rare in nature (Kabata-Pendias and Mukherjee, 2007). Associations of cadmium with the sulphide ores of lead, copper and mainly with zinc which is the major contributor of its release, are also found

Chapter 20

572 Soil Remediation and Plants

in nature (ECB, 2007; Atsdr, 2008; Cotuk, 2010). According to an estimate, 6 million tons of cadmium is released as a by-product of the zinc industry out of 1.9 billion tons of zinc resources worldwide holding 0.3% cadmium (Tol-cin, 2009). Weathering, volcanic eruptions and forest fires are the significant contributing factors for natural cadmium emission, which fluctuates between 10 and 50% of the total emission (Nordic Council of Ministers, 1992; OECD, 1994; Van Assche, 1998).

However, major causative factors to environmental cadmium contamination are anthropogenic in origin. The most prominent release of cadmium into the environment is from metal-mining industries (Miura, 2009). It is extensively released as a by-product from metal industries (Wu et al., 2004; Tolcin, 2009).

Moreover, the use of phosphate fertilizers is also responsible for Cd release in agro-ecosystems, which in turn entirely depends on the parental rock used for the manufacture of the fertilizer (Yu-Jing et al., 2003). Various reports regarding release of cadmium from inorganic phosphate fertilizers are evident (Grant and Sheppard 2008, 2011). Approximately 43 million tons of P2O5 fertilizer was produced during 2011–2012, out of which 99% derived from phosphate rocks (FAO, 2012). The igneous origin of phosphate rocks used for fertilizer manufac-turing contains 15 mg kg−1 of cadmium while sedimentary phosphate rocks have higher cadmium 20–245 mg kg−1 which is one of the major vehicles for depositing cadmium in our agro-ecosystems (Cotuk et al., 2010).

In addition to the above-mentioned sources, the excessive use of cadmium salts in different industries contributes to its release into the atmosphere. For instance, the use of cadmium chloride in electroplating, dyeing, photocopying, lubricants and as stabilizers; cadmium nitrate in coloured glass, porcelain and nuclear reactors; cadmium hydroxide in alkaline batteries (IARC, 1993; HSDB, 2009) is common. Furthermore, the deposition of industrial effluents and sew-age water in agricultural lands has also resulted in cadmium build up. Presently, the irrigation of agricultural lands with sewage water and effluents as a source of nutrients is a common practice in third world countries, which leads to soil contamination due to elevated levels of these heavy metals, especially cadmium (Wu et al., 2004).

In the same way, the manufacture and use of Ni–Cd batteries occupies a significant portion of Cd use and release. At present, the use of cadmium-coated materials and pigments is continuing to decrease while the use of Ni–Cd batteries is increasing (Cotuk et al., 2010). To an estimate, approxi-mately 85% of the global Ni–Cd batteries markets is located in Asia (Tolcin, 2008). About 80% of electric vehicles and a limited number of hybrid vehicles are powered by Ni–Cd batteries (Tolcin, 2008, 2009). Not only this, other sources of cadmium emission involve the use of cadmium oxide as ascari-cide and nematocide (HSDB, 2009) pesticides (Papafilippaki et al., 2007), combustion of fossil fuels (De Rosa et al., 2003), power stations, cement industries, heavy road traffic (Wu et al., 2004), electroplating and stabilizers (di Toppi and Gabrielli, 1999).

573Chapter | 20 Remediation of Cd-contaminated soils

SOIL DYNAMICS, RETENTION AND AVAILABILITY OF METALS

Metal are ubiquitous entities in soil modulating geochemistry by being impor-tant components of clay and mineral oxides of iron and manganese (Gadd, 2009). Commonly, sorption, ion-exchange capacities and binding energies of metals vary with respect to variation in soil components (Violante et al., 2010). The proportion of organic matter and clay in soil is one of the key fac-tors governing the availability of metals in soil medium. Higher clay contents, organic matter and hydrous oxides adsorb metal ions and tend to immobi-lize these due to their interaction, thereby reducing metal bioavailability in soil (Miller et al., 1976). Clay possess a significant feature associated with its particles, owing to higher specific surface area, which is responsible for entrapping higher concentrations of metals ions in comparison with coarse-textured soils (Ahmad et al., 2011). Considering the toxic ions’ entry into food chain, retention of metals on clayey soil minerals gives the advantage to sandy soils where these metals are far more easily available (Efremova and Izosimova, 2012). This retention of metals on clay particles can be a result of complexation, ion exchange, precipitation and adsorption (Zachara et al., 1993; Adriano et al., 2004). Conversely, organic matter decomposition may result in the recycling of organically bound metals (Tomáš et al., 2012).

Prime factors responsible for metal mobility and availability comprise pH, redox reactions, extent of organic and inorganic ligands, nature of the sor-bent, charged mineral particles, nutrients and organic acids from root exudates ( Efremova and Izosimova, 2012; Violante et al., 2010). To be more precise, the bioavailability and uptake of metals from soil is prominently dependent on soil physical and chemical characteristics along with climatic conditions (Tomáš et al., 2012).

The pH of the soil has a strong influence on solubility and speciation of metals from mineral sites, therefore it greatly affects the mobility and bioavail-ability of metals in soil medium (Zhao et al., 2010). Generally, solubility and availability of metallic species like Cr3+, Cd2+, Zn2+, Hg2+, Ni2+, Fe2+, Pb2+ and Mn+ decrease to a greater extent as the pH of the soil increases (McLaughlin, 2007). However, it is not the sole factor controlling all this movement of metal-lic ions within the soil. The bioavailability and solubility of heavy metals are also reported to be linearly correlated with redox potential (Patrick et al., 1990; Masscheleyn et al., 1991). It has been shown that, at identical pH, a decrease in redox potential may increase solubility of metals (Yaron et al., 1996).

Not surprisingly, free metals present in the soil also determine the bioavail-ability of a particular metal and its subsequent plant uptake (Moffett and Brand, 1996). Apart from this, application of sewage can increase metal mobility in recipient soils by lowering pH through the process of nitrification and microbial carbon dioxide production (Smith, 1996). Therefore, the characterization of the factors responsible for adsorption, leaching, availability and toxicity of metals in soils is of significant importance (Tomáš et al., 2012).


DYNAMICS OF CADMIUM IN SOILS

Likewise other metals, complexation, adsorption, ion exchange and precipita-tion influence cadmium bioavailability in soils (Christensen and Tjell, 1990). Clay fraction of the soil is a significant component mediating bioavailability and leaching of cadmium (Allen, 1993). The sorption of Cd in soil achieves its equilibrium state within 1 h, however, this is a reversible phenomenon (Christensen and Tjell, 1990). Similarly, soil pH is a key factor driving cadmium (Cd) mobil-ity and bioavailability (McBride, 1989; Cotuk et al., 2010). The mobility of Cd is greater in soils with low pH range, possessing fewer clay particles, low organic matter in soil, and low iron and manganese oxides, high Cd concentra-tions, lower cation-exchange capacities, zinc deficiency and at elevated tem-peratures (Kabata-Pendias, 2001). Generally, Cd mobility is greater at pH range 4.2–6.6, while it is moderate at pH range 6.7–8.8 (Schmitt and Sticher, 1991). Gray et al. (1999) showed that increasing the soil pH from 5.3 to 7.0 results in reduced Cd uptake in five different plant species. Thus, soil acidification phe-nomena like acid rain can positively affect Cd mobility and availability (Nigam et al., 2001). However, contrasting reports regarding pH effect on Cd mobility have also been documented (Eriksson, 1989).

It is not only the physicochemical properties of soil that modulate the Cd bioavailability, but it is also dependent on physiological attributes of roots, plant age and genetics of plant species (Harter and Naidu, 2001; Jung, 2008). The exchangeable, chelated and soluble Cd constituents are relatively more mobile in most of the soils with enhanced phytoavailability (Schmitt and Sticher, 1991). Conversely, carbonates present in soil can adsorb divalent cations like Pb2+, Ba2+ and Cd2+ on their reactive surfaces (Ming, 2002).

INFLUENCE OF THE ASSOCIATED CATIONS AND ANIONS ON CADMIUM BIOAVAILABILITY IN SOIL

The existence of metals in specific forms influences their relative behaviour within soil systems (Ge et al., 2005). The presence of various cations and anions is a significant force driving metal availability and dynamics in soil. Within soil systems, anions contribute to several reactions including the desorption process of heavy metals and are considered in the following.

Cadmium availability is affected by the presence of certain ions in the soil owing to their role in complexation (Degryse et al., 2004), ionic strength ( Gothberg et al., 2004) and competition of surface exchange sites (Tlustos, 2006). An inverse relationship between cadmium availability and ionic strength of the growing medium has been documented (Gothberg et al., 2004). Several anionic species have been reported to enhance the process of desorption of met-als from soil particles. It is reported that presence of acetate and chloride ions increase mobility of metal ions through complexation. Extractability of Cd was 11 times higher with magnesium chloride (1 M MgCl2) than magnesium nitrate [1 M Mg(NO3)2], as documented by Gommy et al. (1998). Likewise anions,


monovalent and divalent cationic species take part in soil dynamic reactions with reference to metal and soil components.

Monovalent cations, for instance Na+ and K+, are useful for increasing extraction of metals while divalent cations like Ca2+, Mg2+ can desorb met-als and, therefore, extraction rates are far greater than for monovalent cations. Divalent cations including Ca2+, Zn2+, Mn2+ and Mg2+ compete with Cd2+ for metal exchange sites (Degryse, 2004) and for uptake by plants (Ramachandran and D’Souza, 2002; Tlustos et al., 2006). The competence for metal exchange is reported as: Ba2+ > Ca2+ > Mg2+ > NH4

+ > K+ > Na+ with monovalent cations showing the least value (Gommy et al., 1998).

RESPONSE OF Cd TOWARDS NATURAL ELEMENTAL INORGANIC AMENDMENTS

Inorganic elements exhibit certain interactive effects in natural soil systems influencing Cd dynamics in soil. Many studies highlighted the effects of inorganic amendments on cadmium, either positive or negative, under dif-ferent textured soils. Hence, these can be utilized in cadmium-contaminated soils depending on objectives either for reducing cadmium bioavailability or to immobilize it through various interactive properties. As stated above, the interactive effect of inorganic amendments in immobilizing Cd is not always positive. Furthermore, these inorganic amendments can primarily be nutrients because of their tendency to mitigate metal stress (Jalloh et al., 2009). In order to comprehensively elucidate these mechanisms, the effect of diverse inorganic amendments with subsequent response towards cadmium has been taken into account.

Calcium (Ca)

If the purpose is to limit or immobilize cadmium in soil, the primary step can be targeting the soil pH through liming. Such a process increases soil pH and ultimately reduces Cd mobility while entrapping it in soil particles (Reeves and Chaney, 2004; Tsadilas et al., 2005). This practice can be achieved using inorganic salts of calcium such as CaCO3, Ca(OH)2, CaO and CaSO4•2H2O (Filius et al., 1998). In addition to this, applying base-rich fertilizers can also be helpful in this respect (Sarwar et al., 2010). Liming is a cost-effective pro-cess for reducing metal transport in plants (Efremova and Izosimova, 2012). Moreover, calcium ion competes with cadmium ions for absorption sites that are available at plant roots (Cataldo et al., 1983). In a study conducted by T lustos et al. (2006), application of calcium oxide and calcium carbon-ate resulted in a 50% reduction of Cd uptake in spring wheat, and this was explained on the basis of immobilization in response to soil liming. Further-more, chelation of cadmium into Cd–Ca crystals has also been reported in tobacco plants (Choi et al., 2001).


Nitrogen (N)

Addition of nitrogen to cadmium-contaminated soil can dramatically modu-late Cd availability. Nitrogen application in both nitrate (NO3

−) and ammo-nium (NH4

+) ion forms is effective for plant growth (Jalloh et al., 2009), but the effects of both the nitrogen forms on geochemistry and Cd dynamics in soil are through different mechanisms. Inorganic nitrogen applied in the form of ammonium ions can acidify soil (Landberg and Greger, 2003) either by a process of nitrification or by the release of protons (H+) that can seriously affect the bioavailability of Cd (Loosemore et al., 2004; Zaccheo et al., 2006). This is one of the major reasons why nitrogen fertilizers like mono-ammonium phos-phate, di-ammonium phosphate, urea and ammonium sulphate can result in cad-mium build-up, due to their ability to decrease soil pH (Zaccheo et al., 2006). In addition, ammonium-based amine complexes have been regarded as heavy-metal-mobilizing mechanisms (Lebourg et al., 1996; Pueyo et al., 2004). The dissociation of ammonium ion can ultimately result in the formation of soluble amine metal exchangeable complexes (Gryschko et al., 2005); however, con-trasting reports are also evident in the literature. In some recent studies, NH4

+ supplementation resulted in lower pH, but increased Cd contents were recorded in response to the application of NO3

− (Xie et al., 2009). Similarly, Jalloh et al. (2009) reported higher Cd content in response to NO3

− and attributed this to the synergistic interactions of nitrate on cadmium. Therefore, cadmium mobility and availability in response to nitrogen is dependent on the type of nitrogen source applied.

Sulphur (S)

In relation to cadmium, the role of sulphur is very significant as sulphur amend-ments in the form of sulphated compounds can limit cadmium bioavailability in soils. The low solubility of metal sulphates is effective in limiting Cd ( Violante et al., 2010). Sulphur application can initiate the formation of insoluble CdS complexes through production of H2S, entrapping cadmium in soil (Hassan et al., 2005) by vulcanization (Bingham, 1979). Interestingly, higher Cd content has been reported in crop plants irrigated with water containing high sulphate content and this was attributed to sulphato complexes (McLaughlin et al., 2006). Sulphate reduction-based mobilization of metals like Pb, Cd and Cu through the formation of sulphide collide has also been reported (Violante et al., 2010).

Zinc (Zn)

Marked antagonistic interactions between Cd and Zn are documented in lit-erature (Cataldo et al., 1983; McKenna et al., 1993; Wei et al., 2003). Both of these elements posses close similarity in their ionic structure, electronegativity (Moustakas et al., 2011), electronic configuration and reactivity towards most of the ligands (Smilde et al., 1992). It is reported that Zn suppresses Cd uptake


(Cataldo et al., 1983); therefore, soil Zn / Cd ratio determines cadmium uptake in plant (Cotuk et al., 2010). Zinc availability in soils even at concentration 0.3 mg L−1 results in a marked competition with Cd for the sorption sites. The relationship between these two elements in soil media is significantly correlated (Eriksson, 1990) and various investigations reported inhibitory and antagonistic effects of Zn on cadmium uptake (McLaughlin and Singh, 1999; Long et al., 2003). Soil amendments with the application of Zn decreased Cd content of plants (Moustakas et al., 2011). It is documented that mass flow is also a factor contributing to plants’ zinc and cadmium uptake (Mullins et al., 1986). Higher Cd accumulations have been associated with Zn-deficient soils (Adiloglu, 2002).

Contrary to this, synergistic interactions between Zn and Cd were also evi-dent in studies where concentrations of both metals resulted in higher metal content of crops (Xue and Harrison, 1991; Nan et al., 2002; Kachenko and Singh, 2006).

Phosphorus (P)

Most agricultural soils have low concentration of phosphorus in comparison with nitrogen and potassium (Sarwar et al., 2010), that ranges to less than 0.15% (Havlin et al., 2007). Many studies have reported positive effects of phosphorus in reducing Cd (Haghiri, 1974; Smilde et al., 1992). Appliance of phosphate to the Cd-contaminated soils hinders the mobile form of Cd by making it immo-bile through the formation of insoluble cadmium phosphate (Dheri et al., 2007; Matusik et al., 2008). Application of phosphate fertilizers along with soil pH between 6.75 and 9.00 reduced the bioavailability of Cd up to 99% (Matusik et al., 2008). This prominent reduction in Cd bioavailability is dependent on sev-eral factors. Phosphate in soil media induces adsorption of Cd2+, its precipitation as Cd3 (PO4)2, a rise in the pH and surface charge of the soil medium and co-adsorption of P and Cd as an ion pair (Bolan et al., 2003a, b). Similarly, Dheri et al. (2007) attributed the phosphorus-mediated decrease in DTPA (diethyl-entriamene pentaacetate)-extractable Cd from soil to in situ immobilization.

Iron (Fe)

In most neutral and alkaline soils, iron exists as Fe3+ in the form of insoluble compounds (Sheng et al., 2008). Iron can regulate pH and redox potential of the soil and its scarcity in the soil can result in enhanced Cd accumulation within plants (Bao et al., 2009). In addition, iron also competes with Cd for membrane transporters in plant roots (Sarwar et al., 2010). Application of chelated iron fertilizer (EDTA – Na2Fe) to the Cd-contaminated soil resulted in decreased Cd content; however, FeSO4 application resulted in the opposite behaviour of Cd (Sheng et al., 2008). Similarly, it is reported that iron plaque formed from the oxidation of Fe2+ to Fe3+ can adsorb Cd on its surface, therefore iron


supplementation can mitigate Cd toxicity by reducing its bioavailability. Similar positive effects of iron supplementation on Cd immobilization in soil and its subsequent positive influence on rice were reported by Liu et al. (2008b).

Manganese, Silicon and Chloride

In the literature, contrasting reports are documented related to Mn and Cd inter-actions in soil medium. A synergistic effect between both these heavy metals has been cited by Chen et al. (2007), while these two heavy metals can also behave antagonistically in soil (Ramachandran and D’Souza, 2002). The oppo-site behaviour of Mn and Cd within soil was also affirmed by Baszynski et al. (1980) and was attributed to its competition with Cd in the soil.

Silicon exits in soil solution naturally at concentrations ranging between 0.1 and 0.6 mmol (Epstein, 1999). It has been reported to be beneficial in mitiga-tion of Cd toxicity in several crops including wheat (Cocker et al., 1998), maize and rice (Liang et al., 2005). However, the Si-mediated reduction in cadmium uptake needs to be discussed with reference to soil dynamics. The primary effect mediated by Si application is the increase in soil pH resulting in immobilization of Cd (Liang et al., 2005). Reports related to cadmium complex formation with chloride in soil systems are evident. The Cd–Cl complex is highly soluble, and results in increased cadmium mobilization within soils (Degryse et al., 2004). That is why the application of KCl to Cd-contaminated soil resulted in increased Cd content in barley (Grant et al., 1999).

ORGANIC AMENDMENTS VERSUS CADMIUM-CONTAMINATED SOILS

In the early parts of this chapter, the significant influence of soil organic matter content to the bioavailability was stated. Organic matter (OM) content of soil and Cd availability are closely interlinked. In comparison to soils with high mineral contents, the soils higher in organic matter can retain metallic cations and exhibit about 30 times more Cd sorption affinity (Sarwar et al., 2010). Cer-tain reactive functional groups like carboxylic, hydroxyl and phenoxy groups control the adsorption of metals very effectively (Alloway, 1995). Therefore, the application of organic matter and related amendments like farmyard manure, poultry manure, biosolids, compost effectively immobilize Cd from contami-nated soils (Jamode et al., 2003; Puschenreiter et al., 2005; Sampanpanish and Pongpaladisai, 2011). In this way, it has been established that the addition of organic compounds to metal-contaminated soil reduces metal availability from the contaminated soils by complexation (Angelova et al., 2010) and solubility considerations (Ciecko et al., 2001).

The influence of some commonly used organic amendments, both natural and of synthetic origin, is grouped together for proper understanding in relation to soil Cd content.


NATURAL ORGANIC ADDITIVES

Commonly used organic additives include cow, farmyard and poultry manure, and partially decomposed matter termed humus. The organic amendments have been effective in improving soil physical characteristics (Bradshaw and Chadwick, 1980; Bouajila and Sanaa, 2011) and imparted positive influence on crops (Kaihura et al., 1999). Application of farmyard manure (FYM) can reduce the toxic effects of heavy metals on crops (Yassen et al., 2007). Increased soil organic fractions directly regulate its sorption and cation exchange for Cd affecting bioavailability (Alamgir et al., 2011). One unit pH decrease in soil is reported with the addition of 320 g kg−1 organic matter to soil but it resulted in 1.5- to 6-fold increased cation exchange capacity (CEC) depending on soil texture (He and Singh, 1993). Therefore, the decrease in plant cadmium content was attributed to organic-matter-mediated increment in CEC (He and Singh, 1993).

Similar reports regarding FYM-induced increase in CEC of soil (Alamgir et al., 2011), and in combination with lime to reduce cadmium in rice grain (Kibria et al., 2011) are evident. The above-mentioned FYM-mediated decrease in Cd availability was confirmed by Pearson’s correlation analysis. The concen-trations of Cd were found to be negatively correlated (= −0.841 and −0.869 for the shoot and root, respectively) in response to the FYM application (Alamgir et al., 2011).

In addition to FYM, application of chicken manure (Li et al., 2006) and compost (Chiu et al., 2006; Pitchel and Bradway, 2008) to the contaminated soil decrease the extent of phyto- or bio-available Cd by immobilizing it. Another organic amendment is humus that acquires negatively charged sites on phenol and carboxylic groups capable of binding metals (Stevenson and Fitch, 1994). Marked proportions of humified OM were found to be associated with cow manure which also resulted in metal immobilization (Tordoff et al., 2000). Furthermore, humus acts as a multi-ligand component system of soil due to pos-session of large numbers of complex sites (Buffle, 1988). Therefore, the organic portion of the soil medium not only regulates CEC of soil but also modulates its geochemistry. It augments soil buffer capacity, minimizes soil compact-ness and acts as a nutrient pool by recycling nutrients naturally (Stewart et al., 2000). Moreover, immobilized Cd–organic matter complexes are also reported ( Putwattana et al., 2010). However, a few contrasting reports are also available in the literature showing a positive relationship between OM application and Cd mobilization and its subsequent bio-sequestration (Narwal and Singh, 1998).

ROOT EXUDATES AND THE CONCEPT OF ORGANIC ACIDS AS NATURAL CHELATORS

Certain natural organic chelates including amino acids, polysaccharides, organic acids and related compounds are introduced into the rhizosphere by plants roots (Dong et al., 2007) to alter and modulate the solubility of


metal ions. Organic acids (anions) comprise the major proportion of these exudates that possess tendencies to strongly bind with metals despite their physical state (Jones and Darrah, 1994; Jones et al., 1996). Interaction of Cd with organic ligands upon supplementation with organic acids has been proposed (Nigam et al., 2001). In addition, organic acids can affect sorp-tion and desorption of Cd despite soil type and texture (Wang et al., 2013). Therefore, the detailed effect of various low-molecular-weight organic acids is considered below.

LOW-MOLECULAR-WEIGHT ORGANIC ACIDS AND CADMIUM CHELATION

Citric, malic, fumaric, succinic, aspartic, glutamic and related low-molec-ular-weight organic acids (LMWOAs) are proposed as potential metal che-lators (Naidu and Harter, 1998). Most specifically, the carboxylic acids, malic and citric acid can result in stable complexes with divalent cations ( Cieslinski et al., 1998) and result in Cd chelation (Dong et al., 2007). In many investigations, the chelated Cd species in turn were translocated through the xylem and were accumulated by plants more rapidly. Applica-tion of citric acid (10 mmol kg−1) resulted in 1.3- to 3-fold increased uptake of Cd from Cd-contaminated growth medium in Brassica juncea (Quartacci et al., 2005). Similar positive beneficial effects have been documented by Duarte et al. (2007) and Qu et al. (2011). Generally, the use of citric acid is greatly preferred over other organic acids as it does not impair growth and biomass of plants, biodegrades effortlessly, is environmentally friendly and cost effective (Melo et al., 2008; Smolinska and Krol, 2010; Wuana, et al., 2010). Cd–citric acid chelation is attributed to transformation of exchange-able Cd into residual form (Mojiri, 2011). Conversely, the application of citric acid to soil for Cd removal from the soil is not always effective and can be attributed to rapid degradation (Ström et al., 2001). A degradation of approximately 90% in applied citric acid was recorded within the 25 days (Jia et al., 2009). In addition to this fact, citric acid can be mineralized by soil microorganisms (Lesage et al., 2005) which cause re-adsorption of metal contaminant on soil particles (Nascimento et al., 2006). Moreover, inefficacy of lower concentrations of citric acid is unable to desorb soil Cd (Elkhatib et al., 2001; Turgut et al., 2004), therefore it is regarded as a weak chelating agent (Kirpichtchikova et al., 2006).

The role of malic acid in mobilizing soil entrapped / sequestered Cd is also documented (Zhang et al., 1997; Naidu and Harter, 1998; Cieslinski et al., 1998) but citric acid application is preferred over malic acid (White et al., 1981). Besides, along with citric acid, many studies report the beneficial interaction of organic acids like malic and oxalic acids on mobilization of metals by increas-ing solubility (Peters, 1999; Nigam et al., 2001; Chen et al., 2003; Evangelou et al., 2006).


EFFICACY OF SYNTHETIC ORGANIC CHELATING AGENTS TOWARDS CADMIUM

The manufactured organic chelates are most effective in chelating metal con-taminants (Huang et al., 2005; Anwer et al., 2012). These include EDTA (ethyl-enediaminetetraacetic acid), DTPA, CDTA (1,2-cyclohexanediaminetetraacetic acid low sodium), EDDS (ethylenediaminedisuccinate), NTA (nitrilotetraacce-tic acid), EGTA (ethylene glycol-O,O-bis-[2-amino-ethyl]-N,N,N,N,- tetraacetic acid) and related compounds (Quartacci et al., 2007; Pastor et al., 2007; Anwer et al., 2012) that have a strong potential for mobilizing heavy metals from the soil (Kayser et al., 2000). The appliance of chelating agents results in metal complexation and ultimately increased uptake of the chelated metal species (Nascimento et al., 2006; Engelen et al., 2007), but this is not the case all the time. Despite the increasing metal solubility and its subsequent mobilization, these compounds do not always result in the enhancement of metal uptake in plants (Liu et al., 2008a; Khan et al., 2000). The effects of various synthetically derived organic chelates towards cadmium are discussed below.

Predominantly, EDTA is the most widely used synthetic chelant under Cd contamination due to its strong affinity for Cd and because it is slowly biodegrad-able (Means et al., 1980; Saifullah et al., 2009). The extremely high binding affin-ity of EDTA for various metals enables the release of metals from the insoluble phase to the soluble phase, therefore, plays a significant part in the soil metal mobilization (Nowack, 2002). Several studies resulted in Cd release from seques-tered portions in soil, affirming the positive effect of EDTA. In various studies, lower concentrations of EDTA resulted in Cd leachate from the soil (Römkens et al., 2002). Furthermore at elevated EDTA concentration, pH effect is masked showing not much influence on chelation (Ghestem and Bermond, 1998). Other than EDTA-assisted removal of Cd from soil, there are some limitations as well. Firstly, EDTA is a non-selective chelating agent with strong affinity for most of the divalent cations (Zeng et al., 2005). Secondly, it is not environmentally friendly as its application leads to a severe decline in soil nematodes (Wu et al., 2004). Most importantly, its application in soil results in marked reduction in soil essential micronutrients such as iron, magnesium and zinc (Wasay et al., 1998).

Another chelating agent purposely used for Cd complexation is DTPA. In many studies, applications of DTPA resulted in increased Cd solubility (Mehmood et al., 2012), therefore increased its uptake in plants (Kirkham, 2006; Engelen et al., 2007). Concentrations of Cd in the lechate were directly correlated with the DTPA dose applied (Wu et al., 2004) and this effect was prominent on Cd solubility even after the course of 90 days depicting high persistence of DTPA in soil (Wenzel et al., 2003). However in contrast, some severe limitations are associated with DTPA application in soil. It is less efficient in Cd solubilization than EDTA and is light sensitive (Metsärinne et al., 2004; Engelen et al., 2007; Evangelou et al., 2007). Furthermore, its application is toxic to plants which makes it unsuitable for phytoextraction (Nascimento et al., 2006).


Other than EDTA and DTPA, there are certain other synthetic chelating com-pounds that modulate metal mobility but extensive investigations are required to properly elucidate their role in terms of Cd dynamics in soil. NTA is one such compound that can bind with several metals and increase their solubility (Bolton et al., 1996) but it is reported to be very toxic to plants. Application of NTA at 10 mmol kg−1 resulted in death of mustard plants within 48 h (Quartacci et al., 2006). Similarly, CDTA and EGTA can significantly form complexes with higher proportions of Cd from the soil (Mehmood et al., 2012) but investiga-tions are limited.

Therefore, variation is evident considering the influence of various amendments on cadmium in soil. A generalized effect of these amendments is presented in Figure 20.1.

RECENT PRESENTED REPORTS REGARDING GRAIN CROPS

Potentiality and feasibility of cereal crops to be used for remediating cadmium-contaminated soils cannot be neglected. Generally, Cd translocation from soil to grain component is largely dependent on soil and plant type along with agricul-tural practices (Rodda et al., 2011). The trend of Cd bioaccumulation in cereal crops from various studies is discussed in the following paragraph (see also Table 20.1).

Cadmium stress resulted in reduced growth rates of barley seedlings (Wu et al., 2007). In a study conducted by Wu et al. (2003), four barley genotypes responded differently to Cd in terms of biochemical attributes. The grain Cd con-centration depends on root and shoot Cd content (Wu et al., 2007). Interestingly, the cadmium content of developing barley was maximum (51%) in grain as com-pared with the other parts which was in following order awn > stem > grain > rachis > glume (Chen et al., 2007). However, the Cd content of wheat grain was least in grain followed by straw and root, respectively (Wang et al., 2012). Cd con-centration in wheat grain grown in non-contaminated soils in the United States ranged from 0.002 to 0.207 mg Cd kg−1 dry weight. In comparison with bread wheat genotypes, durum wheat tends to accumulate more Cd content in the grains (Li et al., 1997) but variations in Cd bioaccumulation in grains among durum wheat genotypes is also evident (Clarke et al., 2002). Sorghum is also an impor-tant cereal and a fodder crop but its biomass and productivity is affected by soil Cd contamination. Application of organic matter led to an increase in sorghum biomass under Cd toxicity; however, increased growth indirectly contributed to higher Cd accumulation (Pinto et al., 2005). The effect of nitrogen fertilizer in increasing the wheat grain Cd content has been documented (Perilli et al., 2010). High soil salinity resulted in increased Cd bioaccumulation in maize was attrib-uted to Cd–Cl soluble complexes (Chuken, 2012). Many other contrasting reports among genotypes of a single species are documented in literature. Consequently, from a soil and plant perspective, exploration of the physiological mechanisms of Cd accumulation in grain crops need further exploration.

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Amendments

Organic

LMOWA Organic Matter Synthetic Chelates

CDTA

Inorganic

CitricAcid

MalicAcid

OxalicAcid

Tartaric Acid

CowManure

PoultryManure Compost EGTAEDTA DTPAFarmyard

Manure

Mobile Immobile Mobile MobileImmobile Immobile / Mobile

CadmiumCadmium

Calcium Phosphorus Silicon Nitrogen Chloride Sulphur Zinc Iron Manganese

FIGURE 20.1 Schematic diagram displaying the influence of various amendments (organic and inorganic) on cadmium dynamics in soil. LMOWA, low-molecular-weight organic acids.

584Soil R

emediation and Plants

TABLE 20.1 Concentration of Cadmium in Grain of Some Cereal Crops from Recent Literature

Sr. No Crop

Grain Cadmium Concentration Experimental Soil / Growing Medium Amendment Used ReferenceMin Max

1 Maize 3.5 mg kg−1 12.3 mg kg−1 Sandy Loam Zinc Akay and Koleli, 2007

2 Barley 0.01 mg kg−1 0.07 mg kg−1 — — Nakayama et al., 2009

3 Maize 26 ng g−1 37 ng g−1 — Nitrogen Kui et al., 2009

4 Sorghum 0.14 mg kg−1 0.36 mg kg−1 — — Faruruwa et al., 2013

5 Millet 0.11 mg kg−1 0.22 mg kg−1 — — Faruruwa et al., 2013

6 Wheat 0.015 mg kg−1 0.040 mg kg−1 Acidic (pH 5.1–5.8) — Wieczorek et al., 2005

7 Barley 0.020 mg kg−1 0.064 mg kg−1 Acidic (pH 5.1–5.8) — Wieczorek et al., 2005

8 Wheat 0.034 mg kg−1 0.099 mg kg−1 Loamy Soil Inorganic Nitrogen Li et al., 2011

9 Wheat 1.7 nmol plant−1 5.1 nmol plant−1 Loamy soil Zinc Herren and Feller, 1997

10 Rice 1.0 mg kg−1 6.3 mg kg−1 — Farmyard manure Mathew et al., 2002

11 Cow pea 1.2 mg kg−1 2.8 mg kg−1 — Farmyard manure Mathew et al., 2002

12 Sesame 3.2 mg kg−1 13.7 mg kg−1 — Farmyard manure Mathew et al., 2002

13 Rice 0.03 μg g−1 0.12 μg g−1 Paddy soils — Machiwa, 2010

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TABLE 20.1 Concentration of Cadmium in Grain of Some Cereal Crops from Recent Literature—cont’d

14 Durum wheat 0.08 μg g−1 0.32 μg g−1 Five different soil types

— Cifuentes et al., 2012

15 Rice 0.0519 mg kg−1 0.0857 mg kg−1 Polluted water irrigated soil

— Bakhtiarian et al., 2001

16 Wheat 2.99 mg kg−1 4.87 mg kg−1 Hydroponics — Zhang et al., 2002

17 Durum wheat 0.025 mg kg−1 0.359 mg kg−1 Saline soil — Norvell et al., 2000

18 Rice (brown) 0.1 mg kg−1 0.8 mg kg−1 Hydroponics — Kukier and Chaney, 2002

19 Rice 0.02 mg kg−1 5 mg kg−1 Zn / Cd contaminated

— Simmons et al., 2003

20 Soybean 1.081 mg kg−1 1.71 mg kg−1 Zn / Cd contaminated

— Simmons et al., 2003

21 Wheat 0.020 mg kg−1 0.049 mg kg−1 Calcareous soils Sewage Sludge Qiong et al., 2012

22 Maize 0.0019 mg kg−1 0.023 mg kg−1 Calcareous soils Sewage Sludge Qiong et al., 2012


CONCLUSIONS AND THE CONCEPT OF COUPLED PHYTOREMEDIATION AS A FUTURE PERSPECTIVE

So far we have considered the efficacy of a diverse group of chelating agents vide inorganic and organic origins. Though most of these agents exhibited promising characteristics in affecting cadmium availability by altering its mobility in the soil system, the efficacy in terms of complete immobilization or mobilization is scarce.

Phytoremediation is an emergent and advanced methodology selected and executed recently for removing metal contaminants from the soil by growing plant species. Many studies emphasized the use of plants for reme-diation of Cd-contaminated soils but still certain limitations are associated with phytoremediation technology. This includes the inability of biomass production by plants or sometimes inefficient transfer of metal contaminant from the soil to aerial plant parts. The adverse effects of cadmium toxic-ity on biomass production ability and biochemical functioning (Raza and Shafiq, 2013; Raza et al., 2013) are important factors governing the selec-tion of suitable plant species for phytoremediation technology. To tackle these limitations, coupled phytoremediation can be a comprehensive solu-tion involving the use of two successive approaches. Coupled phytoremedia-tion involves the selection of an efficient chelating agent along with suitable plant species. For instance, consider the chelating efficiency of ammonium sulphate. Ammonium ion can enhance Cd mobility while, in contrast, the acidic radical (sulphate) sometimes reduces its mobility within the soil sys-tem. In accordance with the concept of coupled phytoremediation, the use of ammonium nitrate will be far better for mobilizing Cd through the soil than ammonium sulphate as it liberates NH4

+ and NO−3 ions. Both nitrate and ammonium ions (acidic and basic radicals) are pronounced in enhancing Cd mobility in the soils either by chelating processes or by affecting pH-related attributes. Furthermore the application of ammonium nitrate can be coupled with the cultivation of high-biomass-producing, metal-accumulating plant species. To be more precise, the removal of cadmium will be targeted by using two simultaneous approaches acting together. The use of inorganic salts having potential of mobilizing Cd or enhancing its bioavailability from the soil both through acidic and basic radicals can therefore be effective. Likewise, natural organic acids and synthetic organic chelates can also be coupled for phytoextraction of cadmium metal. However in contrast, if the purpose is to immobilize Cd in soil to limit its uptake by edible crop plant species, the use of inorganic and organic amendments effective in immobi-lizing soil Cd can be simultaneously coupled with growing genotypes exhib-iting potentiality to exclude Cd and vice versa. In the latter context, the use of CaSO4 or ZnSO4 with the addition of organic matter will be beneficial in limiting Cd in soils.


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