Bioavailability of heavy metals in soils: definitions and practical implementation--a critical review.

Environ Geochem Health DOI 10.1007/s10653-015-9695-y

ORIGINAL PAPER

Bioavailability of heavy metals in soils: definitions and practical implementation—a critical review Rog-Young Kim • Jeong-Ki Yoon • Tae-Seung Kim • Jae E. Yang • Gary Owens Kwon-Rae Kim

•

Received: 4 December 2014 / Accepted: 27 March 2015 Ó Springer Science+Business Media Dordrecht 2015

Abstract Worldwide regulatory frameworks for the assessment and remediation of contaminated soils have moved towards a risk-based approach, taking contaminant bioavailability into consideration. However, there is much debate on the precise definition of bioavailability and on the standardization of methods for the measurement of bioavailability so that it can be reliably applied as a tool for risk assessment. Therefore, in this paper, we reviewed the existing definitions of heavy metal bioavailability in relation to plant uptake (phytoavailability), in order to better understand both the conceptual and operational aspects of bioavailability. The related concepts of specific and

R.-Y. Kim J.-K. Yoon T.-S. Kim Soil and Groundwater Research Division, National Institute of Environmental Research, 42 Hwangyong-ro, Inchon 404-708, Republic of Korea J. E. Yang Department of Biological Environment, Kangwon National University, 1 Kangwondaehak-gil, Chuncheon 200-701, Republic of Korea

non-specific adsorption, as well as complex formation and organic ligand affinity were also intensively discussed to explain the variations of heavy metal solubility and mobility in soils. Further, the most frequently used methods to measure bioavailable metal soil fractions based on both chemical extractions and mechanistic geochemical models were reviewed. For relatively highly mobile metals (Cd, Ni, and Zn), a neutral salt solution such as 0.01 M CaCl2 or 1 M NH4NO3 was recommended, whereas a strong acid or chelating solution such as 0.43 M HNO3 or 0.05 M DTPA was recommended for strongly soil-adsorbed and less mobile metals (Cu, Cr, and Pb). While methods which assessed the free metal ion activity in the pore water such as DGT and DMT or WHAM/ Model VI, NICA-Donnan model, and TBLM are advantageous for providing a more direct measure of bioavailability, few of these models have to date been properly validated. Keywords Bioaccessibility Free metal ion activity Phytoavailability Mobility Specific adsorption Complex formation

G. Owens Environmental Contaminants Group, Mawson Institute, University of South Australia, Mawson Lakes, SA 5095, Australia

Introduction K.-R. Kim (&) Department of Agronomy and Medicinal Plant Resources, Gyeongnam National University of Science and Technology, Jinju 660-758, Republic of Korea e-mail: [email protected]

Since the industrial revolution, the earth’s ecosystem has been increasingly exposed to a variety of potentially hazardous elements, which may alone or in

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Environ Geochem Health

combination impair ecosystem and human health. In particular, accumulation of the non-degradable inorganic contaminants such as Cd, Cr, Cu, Ni, Pb, and Zn has occurred extensively in soil ecosystems (Adriano 2001; Desaules 2012; Desaules and Studer 1993; Herter and Kuelling 2001). For example, soils in North Rhine-Westphalia, an area having the highest population density in Germany, were contaminated with Cd up to 258 mg kg-1, Cr up to 25,245 mg kg-1, Cu up to 793 mg kg-1, Ni up to 770 mg kg-1, Pb up to 17,733 mg kg-1, and Zn up to 68,155 mg kg-1 (Liebe et al. 1997). The source of this contamination is primarily anthropogenic, arising from a combination of industrial discharges including mine wastes, dry or wet deposition of coal ashes, urban refuse, agricultural and animal wastes, fertilization with phosphate, compost, or sewage sludge, pesticide application such as fungicides, or irrigation with wastewater (Alloway and Ayres 1997; Kabata-Pendias 2011; Mousavi et al. 2013; Nriagu 1990; Papp 1994; Rieder and Schwertmann 1972; Sauerbeck and Rietz 1981). Since the 1990s, legislative authorities worldwide have established guideline values of potentially toxic elements in soils to manage contaminated sites and protect soils against harmful changes (Table 1; BMUB 1999; CCME 1999; EA 2004, 2009; FOEFL 1998a, b; ME 1995; MfE 1997, 2006; MOEJ 1991, 2002; NEPC 1999a, b, c; US EPA 2002, 2002, 2005; VROM 2007). Within the scope of these regulations, guideline values for the assessment and remediation of contaminated sites were mostly legislated on the basis of the total metal content in the soil. However, this may inadvertently overestimate the potential risk, causing unnecessary and expensive soil remediation efforts (Alexander 2000; Derz et al. 2012; McLaughlin et al. 2000; Meers et al. 2007; NRC Committee 2003). In addition, numerous studies have emphasized that an organism is adversely affected only by the biologically available (bioavailable) fraction of the total metal content in soil, and not by the fraction which is sequestered or irreversibly bounded to the soil matrix (non-available fraction) (Blume and Bruemmer 1991; Bruemmer et al. 1986; Kim et al. 2007; Lanno et al. 2004; Rieuwerts et al. 1998). The total content, in itself, is indicative of the saturation level of metals in the soil matrix and can be useful as a first impression of soil contamination or for the management of additional inputs of metals into soils. However, it is not

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enough to assess any actually occurring adverse effects on the soil ecosystem triggered by heavy metals. Recently, in many countries including Korea, regulatory frameworks for the assessment and remediation of contaminated soils have been moved or are being moved towards a risk-based approach taking bioavailability into account (Brand et al. 2009; Harmsen 2007; ISO 17402 2008; NEPC 1999b, c; US EPA 2003). However, there is much debate on the precise definition of bioavailability across scientific disciplines such as soil science, biology, pharmacology, and ecotoxicology, and methods for the measurement of bioavailability have yet to be standardized so that the concept can be reliably applied as a tool for risk assessment. Therefore, in this paper, we review the current existing definitions of heavy metal bioavailability in soils specifically in relation to plant uptake (phytoavailability), by considering metal–soil physicochemical interactions (e.g. pH dependence and organic matter content), plant uptake mechanism (e.g. active and passive uptake across root membranes), and biological endpoints (e.g. toxicity, growth or bioaccumulation). In order to understand the different behaviours of heavy metals in soils, the adsorption and desorption mechanisms resulting from different affinities to form complexes with hydroxides or organic ligands are intensively discussed. Existing methods for the measurement of metal phytoavailability based on both chemical extractions and mechanistic geochemical models are also briefly reviewed to provide an insight into how to develop such a measurement tool. The development of such tools which accurately predict metal bioavailability is the first step in providing a risk assessment of potential exposure for soil–crop–human transfer and can contribute to the improvement in both ecosystem and human health.

Understanding bioavailability As indicated from the numerous definitions of bioavailability existing in the literature, to date, the term bioavailability has not been consistently defined. In addition, Harmsen (2007) argued that the definitions are often too comprehensive to be used in any practical measure of a quantity. For these reasons, over the last two decades, several groups of scientists have been engaged in developing a more conceptual

Environ Geochem Health Table 1 Summary of guideline values for potentially hazardous elements in soils from different countries (modified after MfE 2011) Country Australia Canada Germany

Japan

Name

Purpose

Basis data

Elements

References

Health investigation levels (HILs)

Site investigation

Hu

26

NEPC (1999a)

Ecological investigation levels (EILs)

Site investigation

Eco

11

NEPC (1999b, c)

Soil quality guidelines (SQGs)

Assessment and Remediation

Hu, Eco

29

CCME (1999) BMUB (1999)

Precautionary values

Sustainable soil quality

Eco

10

Trigger values

Site investigation

Hu, Eco

33

Action values

Remediation

Hu, Eco

33

Target values

Remediation

Hu, Eco

25

Soil leachate standards

Remediation

Hu

25

MOEJ (2002)

Soil content standard

Remediation

Hu

9

MOEJ (2002)

Republic of Korea

Worrisome levels

Remediation

Hu

21

ME (1995)

Countermeasure levels

Remediation

Hu

21

ME (1995)

New Zealand

Acceptance criteria (timber treatment)

Site investigation

Hu

7

MfE (1997)

Soil guideline values (sheep dip)

Site investigation

Hu

19

MfE (2006)

Guide values

Soil quality

Hu

11

FOEFL (1998a)

Trigger values

Site investigation

Hu

6

FOEFL (1998a)

Clean-up values

Remediation

Hu

7

FOEFL (1998a)

Concentration values

Remediation

Hu, Eco

18

FOEFL (1998b)

Remediation urgency assessment

Hu, Eco

83

VROM (2009)

Target values


Eco

80

VROM (2009)

Maximum values for soil quality classes


Hu, Eco

100

VROM (2007)

The UK

Soil guideline values (SGVs) Soil screening values (SSVs)

Site investigation Site investigation

Hu Eco

15 39

The USA

Soil screening levels (SSLs)

Site investigation

Hu

110

US EPA (2002)

Ecological soil screening levels (ecoSSLs)

Site investigation

Eco

24

US EPA (2005)

Switzerland

The Netherlands Intervention values

MOEJ (1991)

EA (2009) EA (2004)

Hu human, Eco ecosystem

definition combined with an operational aspect, which may allow the mechanistic understanding of bioavailability and the standardization of a measurement tool for bioavailability. This is of utmost importance as science moves from traditional discipline-specific research into more multidisciplinary research activities where researchers must speak the same language. For example, Peijnenburg et al. (1997) and Lanno et al. (2004) emphasized that bioavailability should be considered as a dynamic process comprising the

following three distinct steps: (1) a physicochemically controlled desorption process, referred to as environmental availability, (2) a physiologically controlled uptake process, referred to as environmental bioavailability, and (3) a physiologically induced effect or accumulation within the organism, referred to as toxicological bioavailability. Hund-Rinke and Koerdel (2003) also pointed out that bioavailability is a complex dynamic process strongly controlled by the type of organism, type of exposure, and metal speciation. The principle of dynamic processes was

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adopted in the US National Research Council (NRC) report (NRC 2003; see below) and later by the International Standards Organisation (ISO 17402 2008; see below). Some scientists have distinguished between external and internal bioavailability (Caussy et al. 2003). External bioavailability depends on the ability of metals to be dissolved and released from the soil matrix or other media, whereas internal bioavailability is determined by the ability of metals to be absorbed and exert subsequent toxicological effects in target tissues. Kramer and Ryan (2000) and Semple et al. (2004) proposed that the term bioaccessibility to be used synonymously with external bioavailability, which corresponds to environmental availability mentioned above. However, Reichenberg and Mayer (2006) stressed that bioaccessibility and chemical activity should also be distinguished in order to measure bioavailability. Bioaccessibility is an operationally defined quantity of a metal which can be mobilized into the pore water and become available for uptake within a given timescale and under given environmental conditions. This can be measured by using an extraction scheme. Chemical activity is related to the effective concentration of a free metal ion under thermodynamic equilibrium conditions and practically can be measured by using equilibrium sampling devices including ion-selective electrodes, sensors, and solid-phase microextraction (SPME). This concept, known as the free metal ion activity model (FIAM), assumed that plant roots took up metals solely in the form of freely dissolved metal ion in the pore water (Morel 1983). However, recently, several studies found that the complexed and adsorbed species also contributed to metal uptake (McLaughlin et al. 1997; Nolan et al. 2005; Zhao et al. 2004). The bioaccessible fraction is sometimes referred to as the mobile fraction as introduced by Herms and Bruemmer (1984) and Bruemmer et al. (1986). Zeien and Bruemmer (1989) defined the mobile fraction as the amount of a heavy metal which was dissolved or exchangeable adsorbed (non-specific adsorbed; see details below). Particular definitions are often stressed within particular disciplines, so, for example, while an environmental scientist concerned with site remediation is more likely to be engaged in determining the bioaccessible fraction of metals in the soil, a

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biologist is more likely to be concerned with the bioavailable fraction based on a particular biological response (i.e. internal bioavailability) for the same soil. Other scientists have proposed more practical definitions of bioavailability to reflect operational measurement. Warrington and Skogley (1997) defined bioavailability in terms of content (mol kg-1) as a quantity of heavy metal in the soil existing in forms and amounts that plant roots can take up during their lifetime. Alternatively, Taghon et al. (1999) and Shor and Kosson (2000) defined bioavailability in terms of flux, i.e. transport rate (mol m-2 s-1), at which metals can be transported to the organism through the soil. However, Harmsen (2007) suggested that the contentbased bioavailability was a more practical measure than the flux-based bioavailability simply due to better analytical feasibility. Only two large research authorities, the NRC and the ISO/TC 190 working group Bioavailability, have attempted to combine the conceptual definition of bioavailability with the operational definition. The NRC Committee (2003) first introduced the term bioavailability processes, defined as ‘‘the individual physical, chemical, and biological processes that determine the exposure of organisms to chemicals present in soils and sediments’’. They included the following five processes: (1) contaminant release from the solid phase, (2) transport of the released contaminants or (3) transport of the bound contaminants to the membrane of an organism, (4) passage across a physiological membrane, and (5) incorporation into a living system through metabolic processes. In order to design a standard measurement of bioavailability, a list of physical (e.g. spectroscopy), chemical (e.g. extraction), and biological (e.g. toxicity bioassay) methods has been evaluated based on several criteria with the key aspect of gaining a mechanistic understanding of bioavailability (e.g. form and associations of a heavy metal). According to the NRC, however, no single method can be universally used to measure bioavailability, and understanding the site-specific conditions is crucial to select an appropriate tool. The NRC encouraged an intensive effort to develop predictive models based on mechanistic understanding, preferred over conventional operational approaches such as extractions (Ehlers and Luthy 2003). Subsequently, the ISO/TC 190 working group Bioavailability published a guideline for the selection


and application of methods to assess bioavailability of contaminants in soils, where bioavailability was described in three conceptual steps (Fig. 1): (1) availability of heavy metals in the soil (environmental availability), (2) uptake of heavy metals by the organism (environmental bioavailability), and (3) accumulation and toxic effect of heavy metals within the organism (toxicological bioavailability) (Harmsen 2007; ISO 17402 2008). The first bioavailability term, environmental availability, is determined by an available amount of the total content in the soil including both the actual available fraction dissolved in the pore water (e.g. heavy metal concentrations in the pore water) and the potential available fraction adsorbed to the soil matrix (e.g. specific and non-specific adsorbed, organically bound, surface-precipitated amounts). Potentially available fractions can be described as the maximum

amount that can be desorbed from the soil matrix under predefined worst-case conditions or within the whole exposure time, being in equilibrium with the dissolved fraction in the pore water (Bruemmer et al. 1986; ISO 17402 2008). On the other hand, Roemkens et al. (2009a) referred to the actual available fraction as the directly available pool of metals and the potential available fraction as the reactive pool. The ISO/TC 190 working group pointed out that the available fraction of total content in soil should not be considered as a fixed fraction but as a continuum, because the total exposure to an organism also depended on the time of exposure. Since time of exposure was an important consideration, several scientists (i.e. Derz et al. 2012; Fedotov et al. 2012) differentiated the potential available fraction into three sub-fractions on the basis of desorption kinetics: (1) rapidly desorbing fraction (from hours to days), (2)

Fig. 1 Three-step concept of heavy metal bioavailability in soils for plants (modified after Derz et al. 2012; Harmsen 2007; ISO 2008; Lanno et al. 2004; thick arrows indicate the most important factors affecting bioavailability)

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slowly desorbing fraction (from weeks to months), and (3) very slowly desorbing fraction (from months to years) (Fig. 1). The differentiation of these fractions is variable depending on both soil (e.g. texture, organic matter, and pH) and metal (e.g. binding strength) properties. The actual available fraction may include not only the amount of dissolved free ions and molecules but also the amount of species complexed with dissolved organic matter (e.g. humic and fulvic acids) or with inorganic anions (e.g. Cl-, OH-, SO42-, and HCO3-), particularly under slightly acidic to alkaline conditions (Blume et al. 2010; KabataPendias 2011). The most common concentrations and species of various dissolved metal ions in the pore water of agricultural and acidic forest soils are given in Table 2. The presence of dissolved organic matter (DOM) may result in a higher environmental availability of heavy metals in the pore water, in particular under moderately acidic up to alkaline soil conditions (Bruemmer et al. 1986). However, the complexation by DOM is generally believed to reduce the availability of metals for uptake by plant roots and therefore their toxic effects (i.e. environmental bioavailability), compared to that of the free metal ion, because the complexed species are sometimes too large to pass the root cell membranes (Brown et al. 2004; Paradelo et al. 2011; Smith 2009; Wang et al. 2010). The complexation of metal ions by inorganic anions such as Cl–, OH–, and HCO3– at intermediate up to alkaline soil pH is generally regarded as being potentially available for plant root membranes. Thus, environmental availability cannot be considered equal to environmental bioavailability and can only be used as an assessment tool if it is correlated well with toxicological bioavailability.

The second bioavailability term, environmental bioavailability, is defined as the fraction of dissolved metal species in the pore water which can be taken up by plant roots or other soil organisms (Fig. 1). This type of bioavailability is thus controlled by physiological process and is dependent on transfers from the pore water to plant roots and uptake mechanisms which may be specific to individual plants and heavy metals. The uptake of metal species from the pore water into the root cell can occur both by passive (diffusion or mass flow) and active (metabolic) processes. According to Morel (1997), in pore water, passive uptake dominates at high ion concentration ([0.1 mM), whereas active uptake using metabolic energy, e.g. H?-ATPase or phytosiderophores, dominates at lower ion concentrations against a chemical gradient (\0.5 lM). Other authors reported that Cd, Cr(III), Ni, and Pb are preferentially absorbed via passive uptake, while the absorption of Cu and Zn takes place preferentially via active uptake or via a combination of both active and passive uptake, because both Cu and Zn are essential plant nutrients (Degryse et al. 2006; Kabata-Pendias 2011; Klassen et al. 2000; Skeffington et al. 1976; Weiss et al. 2004). Plants can also modify the bioavailable fraction of heavy metals in the rhizosphere by changing physicochemical properties and therefore also biological compositions in the rhizosphere (Chen et al. 2013; Han et al. 2006; Kwon et al. 2013). Furthermore, antagonistic or synergetic effects due to the presence of mixtures of specific cations and anions in the pore water should also be considered. Among these diverse factors, pore water ion concentration is regarded as one of the most significant factors that control plant metal uptake (Kabata-Pendias 2011).

Table 2 Typical concentrations (soil saturation extraction and suction cup extraction) and important species of dissolved metal ions in the pore water from low or non-contaminated

acidic forest and agricultural soils (DOC dissolved organic carbon; Blume et al. 2010; Bradford et al. 1971; Campbell and Beckett 1988)

Pore water concentrations Acidic forest soil (lg L-1) Cd Cr

1–25 2–20

Ni Cu

5–30 1–50

Pb

2–100

Zn

80–2000

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Species

Agricultural soil (lg L-1)

Very strong acidic to moderately acidic soils

Slightly acidic to alkaline soils

\0.1–3

Cd2?, CdSO04, CdCl?

Cd2?, CdSO04, CdCl?

\1–15 1–30 3–60 \1–50 10–400

3?

?

Cr , CrSO4 , Cr-DOC 2?

Ni , NiSO04, Ni-DOC Cu-DOC, Cu2?, CuSO04 Pb2?, Pb-DOC, PbSO04 Zn2?, ZnSO04

3Cr-DOC, CrCO3?, Cr(CO3)2 , Cr(CO3)3

NiCO03, NiHCO3?, NiB(OH)? 4 Cu-DOC, CuCO03, CuB(OH)? 4 PbCO03, PbHCO3?, Pb(CO3)22 ZnHCO3?, Zn2?, ZnSO04, ZnCO03


The third bioavailability term, toxicological bioavailability, is defined as a physiologically induced bioaccumulation or biological effect of heavy metal within plants or other organisms which depends on a number of complex processes including translocation, metabolism, and detoxification. Toxicological bioavailability can be determined by bioassays, e.g. accumulation in roots, shoots, and grains, growth inhibition, or toxic effects in plants at metabolic or individual levels (biological measurements in Fig. 1; ISO/FDIS 16198 2014). The bioaccumulation of metals in plants from soils can be predicted using a transfer factor (TF), which may differ considerably between plant, soil, and metal types under investigation. According to Alexander et al. (2006), the relative order of bioaccumulation in different vegetables was as follows: leafy vegetables (high) [ root vegetables (moderate) [ legumes (low). Kloke et al. (1984) reported that the TF of metals in contaminated soils decreased in the following order: Cd and Zn (1–10) [ Ni and Cu (0.1–1) [ Pb and Cr (0.01–0.1). In Table 3, common transfer factors from soils to plant shoots, common concentration ranges in plant leaves, and applicable safety limits in foodstuffs are given for Cd, Cu, Cr, Ni, Pb, and Zn compared to trigger values for arable soils. Within a particular plant species, translocation of Cd, Ni, and Zn is generally regarded as being moderately mobile, whereas translocation of Cu, Cr, and Pb is regarded as being strongly bound to

root cells and of relatively low mobility (KabataPendias 2011). Consequentially, understanding all these processes of metal interactions in each compartment from soils to biological response as a whole is essential to assess overall bioavailability.

Environmental availability of heavy metals in soils In order to measure the environmental availability, it is important to know both the concentrations and chemical forms (species) of the metals in the pore water, which are controlled by a myriad of complex interacting processes such as adsorption/desorption, complexation/dissociation, precipitation/dissolution, or very slow diffusion into the interior of clay minerals and oxides. These processes are in turn influenced by physiochemical parameters such as the total metal content, soil pH, redox conditions, and contents of organic matter, clay minerals, and oxides as well as contact time with the soil matrix (Blume et al. 2010; Gerth and Bruemmer 1981, 1983; Hund-Rinke and Korrdel 2012). As Bruemmer et al. (1986) and McBride (1994) reported, these processes are strongly metal-specific because of the different adsorption affinities of metals on clay mineral and oxide binding sites, the different propensities to form stable complexes with organic and inorganic ligands, and the different complex and mineral solubilities. Therefore,

Table 3 Approximate transfer factors (TF) of heavy metals from soils to plant shoots in contaminated soils, concentration ranges in mature leaf tissue, and safety limits in foodstuffs compared to the trigger values for arable soils Transfer factora

Concentration rangeb (mg kg-1 DW)

Safety limitc (mg kg-1 FW)

Deficient

Sufficient or normal

Leaf vegetables

0.05–0.2

Cd

1–10

–

Zn

1–10

\10–25

25–150 0.1–5

[5–10

0.10

0.20

–

–

–

150–600

–

–

50–230

–

[20–100

–

–

0.1–0.5

[1–2

–

–

1–5

[10–20

0.30

–

0.1–1

\2–5

Cr

0.01–0.1

Pb

0.01–0.1

Rice

0.20

[20–30

0.1–1

Cu

Stem/root vegetables

[150–400

5–20

Ni

a

Excessive or toxic

Trigger values for arable soilsd (mg kg-1 DW) 1–4

–

40–200

–

–

50–130

0.10

0.20

50–200

Transfer factor (TF) = plant content/soil total content; Kloke et al. (1984)

b

Adapted from Benton et al. (1991), Blume et al. (2010), and Kabata-Pendias (2011)

c

European Commission (2011)

d

Adapted from different countries in Table 1

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in order to design a more robust measurement tool, it is useful to firstly classify heavy metals into groups of similar environmental availability or mobility. Processes which involve adsorption/desorption of metals by soil minerals and oxides can be explained well using the concept of specific and non-specific adsorption, developed by Forbes et al. (1976) and Tiller et al. (1984). In a moderately acidic soil, up to a slightly alkaline soil pH, heavy metals exist mainly specifically adsorbed at hydroxyl surfaces of oxides or clay minerals (Blume et al. 2010; Bruemmer et al. 1986; Fischer et al. 2007). This specific adsorption is mainly related to the formation of inner sphere complexes of metal hydroxides (MeOH?) via covalent bonding. Heavy metals exhibiting a high level of specific adsorption are generally regarded as being poorly (environmental) available (Hodson et al. 2011). Tiller et al. (1984) and Herms and Bruemmer (1984) found that the specific adsorption of metals increased with the increasing hydroxide constant of the metal ions as follows, expressed as pKa (pKa = -log Ka; Ka = [MeOH?] 9 [H?]/[Me2?][H2O]): Cd2? (10.1) \ Ni2? (9.9) \ Zn2? (9.0) Cu2? (7.7) = Pb2? (7.7) Cr3? (4.0) (data from Baes and Mesmer 1976), indicating a decrease in metal mobility with the increasing hydroxide constant (see below). In contrast, at low pH, heavy metals are principally non-specifically adsorbed at the binding sites of cation exchange of clay minerals and oxides via electrostatic bonds, which are associated with the exchangeable fraction of metals and considered to be easily (environmental) available (Bruemmer et al. 1986; Hodson et al. 2011). For this reason, heavy metals are more mobile and (environmental) available at low pH than at moderately acidic up to slightly alkaline soil pH. In addition to specific and non-specific adsorption of metal ions, organic matter can also modify the adsorption and/or solubility of metal ions depending on the soil pH and the type of organic matter (Bolton and Thorose 1997; Bruemmer and Herms 1983; Han et al. 2006; Mahara et al. 2007; Minkina et al. 2006; Stevenson 1983; Wong et al. 2007; Zhou and Wong 2003). As shown in Table 4, in many cases, the stability constants of metal–organic complexes increase with increasing pH and are higher with humic acids (formed mainly at slightly acidic to neutral pH) than with fulvic acids (formed mainly at acidic soil pH) (Christensen and Christensen 2000; Elgala et al. 1976; Kitagishi and Yamane 1981; Norvell 1972;

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Schnitzer and Khan 1978; Takamatsu and Yoshida 1978). In general, for a typical soil pH range, the stability of metal–organic complexes can be summarized as follows: with Cd, Zn, and Ni being of low stability and Pb, Cu, and Cr being of high stability (Bruemmer et al. 1986; Herms 1982; Smith 2009: Stevenson 1982). Regarding the solubility of metal– organic complexes, the mobility of metal ions is generally reduced at acidic pH and is enhanced at slightly acidic to alkaline pH, due to the dominant negative charge of most metal–organic complexes (Bloomfield et al. 1976; Blume et al. 2010; Schnitzer and Kerndorff 1981; Sholkovitz and Copland 1981). At pH [ 6, metal–organic complexes are the dominant metal species in the pore water (Bloomfield et al. 1976; Bruemmer and Herms 1983; Bruemmer et al. 1986; Jeffery and Uren 1983; McBride and Bouldin 1984). These results indicated that the solubility of Cu, Cr and Pb, at moderately acidic up to slightly alkaline soil pH, can be significantly enhanced by dissolved organic matter, associated with increased environmental availability. In Figs. 2 and 3, the total metal content, as estimated from aqua regia dissolution, and the 1 M NH4NO3extractable content expressed as the percentage of the total content for 335 soil samples from North RhineWestphalia, Germany, are plotted against soil pH (CaCl2) and land use (data from Liebe et al. (1997) with permission). The total metal content is strongly related to the land use. In general, metal content was largest at pH 6–8 in contaminated industrial sites, while for Pb and Cu, metal content was also large at pH \ 4 in organic layers of forest soils in emission regions (Fig. 2). In comparison, the NH4NO3-extractable contents expressed as a percentage of the total had minima at pH near 6, which increased with decreasing pH, reaching up to 90 % (Cd) [ 78 % (Zn) [ 30 % (Pb) [ 24 % (Ni) [ 6.4 % (Cu) [ 1.2 % (Cr) at pH 2.7 (Fig. 3). The partly enhanced mobility of Cu in the neutral to alkaline pH range, which showed the maximum percentage of 7.1 % at pH 8.1, probably resulted from the formation of soluble metal–organic complexes and soluble carbonate complexes at these pHs (Table 2). On the other hand, the enhanced mobility of Cr at pH 8.1, which had the maximum percentage of 8.3 %, can be explained by the presence of soluble Cr(VI) (validated by Kim (2009)). Other studies have also been conducted to investigate the mobility of heavy metals, measured by the

Environ Geochem Health Table 4 Review of stability constants of metal–organic complexes in different pH Organic substances

pH

Stability constants (log K)

References

Fulvic acids

3.0

Zn2? (2.3) \ Pb2? (2.6) \ Ni2? (3.2) \ Cu2? (3.3)

Schnitzer and Hansen (1970)

3.5

Zn2? (1.7) \ Pb2? (3.1) \ Ni2? (3.5) \ Cu2? (5.8)

Norvell (1972)

5.0 3.5

Zn2? (2.3) \ Ni2? (4.1) \ Pb2? (6.2) \ Cu2? (8.7) Zn2? (2.7) \ Cd2? (2.8) \ Ni2? (3.2) \ Pb2? (3.7) \ Cu2? (5.3)

Norvell (1972) Pandey et al. (2000)

5.0

Cd2? (6.3) \ Pb2? (8.4) \ Cu2? (8.7)

Takamatsu and Yoshida (1978)

–

Cd2? (6.9) \ Pb2? (8.7) \ Cu2? (8.9)

Stevenson (1976)

5.8

Ni2? \ Zn2? \ Cd2? \ Pb2? \ Cu2? \ Cr3?

Gamble (1986)

7.0

Cd2? (8.9) \ Ni2? (9.6) \ Zn2? (10.3) \ Cu2? (12.3)

Kitagishi and Yamane (1981)

Humic acids

2?

2?

(3.8) \ Zn

2?

(5.0) \ Cu

Citric acid

6.0

Cd

DOC

5.0

Zn2? (4.0) \ Cd2? (4.3) \ Ni2? (4.4)

7.0

Zn2? (4.6) \ Cd2? (4.8) = Ni2? (4.8)

Christensen and Christensen (2000)

–

Cd2? (16.4) = Zn2? (16.4) \ Pb2? (17.9) \ Ni2? (18.5) \ Cr3? (24.0)

Lindsay (1979)

EDTA

(5.9)

ratio of dissolved to bound metal ions, in various minerals and soils with different pH, organic matter content, and contamination (e.g. Aryal et al. 2007; Dong et al. 2009; Elliott et al. 1986; Fonseca et al. 2011; Irha et al. 2009; Labanowski et al. 2007; Lei et al. 2010; Liu et al. 2014; Pueyo et al. 2004). Their relative mobilities are listed in Table 5 and can be generally summarized as follows: Cr(VI) [ Cd [ Ni [ Zn [ Cu [ Pb = Cr(III). This sequence of mobility corresponds to both the sequence of pKa values of metal ion hydroxides and that of stability constant of metal–organic complexes. Based on these results, heavy metals can be grouped into two classes: Cd, Ni, and Zn of relatively high mobility and Cu, Cr, and Pb of relatively low mobility in the common soil pH range. Understanding this metal-specific environmental availability facilitates the selection of the appropriate measurement tool.

Methods to measure metal bioavailability in plants For the measurement of bioavailability, bioassays are usually preferred over traditional chemical assays because bioassays give a more direct measurement of biological response and bioaccumulation without the need for significant interpolations of extraction concentration to an effect. However, bioassays are often time-consuming, laborious, and costly due to the different endpoints, exposure time, and uptake mechanisms involved. In some cases, bioassays are

Keller and Roemer (2000) Christensen and Christensen (2000)

excessively complicated and chemical measurements can replace biological assays, if and when an accurate correlation between a chemical and a biological assay is found. For more than thirty years, a number of studies have been conducted to develop, optimize, and validate methods to measure or predict bioavailable fraction of metals for plants, based on either chemical extraction or mechanistic modelling (e.g. Brand et al. 2009; Degryse et al. 2009; Derz et al. 2012; Fedotov et al. 2012; Groenenberg et al. 2010; Lindsay and Norvell 1978; Meers et al. 2007; Meeussen 2003; Parkhurst and Appelo 2013; Peijnenburg et al. 2007; Pueyo et al. 2004; Tessier et al. 1979; Zeien and Bruemmer 1989). It is not within the scope of this paper to evaluate all of these different procedures in detail, and thankfully, an extensive assessment of existing chemical methods has already been previously provided (e.g. Brand et al. 2009; Fedotov et al. 2012; Houba et al. 1996; ISO 17402 2008; Macholz et al. 2011). In Table 6, the most frequently evaluated methods are reviewed both for actual and potential available metal fractions, depending on the type of tools (chemical extraction or mechanistic calculation), strength of extraction media (weak or strong), mechanism of extraction (exchange, chelation, diffusion, or equilibrium sampling), or models involved (free metal ion activity model (FIAM) or surface complexing model). The extraction-based methods can be divided into two groups: (1) those which measure the actual available concentrations in the pore

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400

Total Cr content (mg kg-1)

Total Cd content (mg kg-1)


200

15 10 5 0

2

4

6

8

10

40000 20000

300

200

100

0

12

2

4

1000

200 150 100 50

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4

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8

10

12

10

12

10

12

200

100

0

12

2

4

6

8

pH(CaCl 2) 80000

Total Zn content (mg kg-1)

30000

Total Pb content (mg kg-1 )

10

500

pH(CaCl 2)

15000

1000 750 500 250 0

8

1000

500

0

6

pH(CaCl 2)

Total Ni content (mg kg-1)

Total Cu content (mg kg-1)

pH(CaCl 2)

2

4

6

8

10

12

40000

2000 1500 1000 500 0

pH(CaCl 2)

2

4

6

8

pH(CaCl 2) Arable land Forest Grassland Allotment Industrial area

Fig. 2 Variation of total heavy metal content (aqua regia dissolution) in 335 soil samples from North Rhine-Westphalia, Germany, with soil pH (CaCl2) and land use (data from Liebe et al. 1997 with permission)

water, either the total dissolved concentrations extracted by unbuffered neutral salts solution such as 0.001–0.01 M CaCl2, 0.1 M NaNO3, or 1.0 M

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NH4NO3 or the free metal ion concentrations extracted by diffusion gradient in thin films technique (DGT) or Donnan membrane technique (DMT) and (2) those

100

NH4NO3-Cr/total-Cr*100 (%)

NH4NO3-Cd/total-Cd*100 (%)


80 60 40 20 0

10.0 8.0 1.0

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6

8

pH(CaCl 2)

30

NH4NO3-Zn/total-Zn*100 (%)

NH4NO3-Pb/total-Pb*100 (%)

6

pH(CaCl 2)

NH4NO3-Ni/total-Ni*100 (%)

NH4NO3-Cu/total-Cu*100 (%)

pH(CaCl 2)

25 20 15 10 5 0

100 80 60 40 20 0

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pH(CaCl 2) Arable land Forest Grassland Allotment Industrial area

Fig. 3 Variation of 1 M NH4NO3-extractable metal content (% of total content) in soils from North Rhine-Westphalia, Germany, with soil pH (CaCl2) and land use (data from Liebe et al. 1997 with permission)

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Environ Geochem Health Table 5 Review of relative mobility of heavy metals in different soils with different metal contaminations Soils

Sequences of mobility

References

Montmorillonite and kaolinite

Cr(VI) [ Cd [ Zn [ Pb [ Cu [ Cr

Griffin and Shimp (1978)

Soil clay minerals

Cd(II) [ Ni(II) [ Zn(II)

Tiller et al. (1984)

Acidic soils Mineral soils (pH 5.0)

Cd(II) [ Ni(II) [ Zn(II) Cu(II) [ Pb(II) Cd(II) [ Zn(II) [ Cu(II) [ Pb(II)

Herms (1982) Elliott et al. (1986)

Contaminated garden/paddy soils (pH 4.8–5.5)

Cd(II) [ Cu(II) [ Zn(II) [ Pb(II)

Lei et al. (2010)

Contaminated agricultural soils (pH 4.8–6.3)

Cd(II) Pb(II) [ Cu(II) = Zn(II)

Liu et al. (2014)

Contaminated soils (pH 3.6–7.9)

Cd(II) [ Zn(II) [ Cu(II) [ Pb(II)

Pueyo et al. (2004)

Subsoil of a Podzoluvisol (pH 7.1)

Cd(II) [ Cu(II) [ Cr(III) [ Pb(II)

Irha et al. (2009)

Mineral soils containing 20–40 g kg-1 OM (pH 5.0)

Zn(II) [ Cd(II) [ Cu(II) [ Pb(II)

Elliott et al. (1986)

Zn-contaminated agricultural soils (pH 6.6)

Zn(II) = Cd(II) [ Cu(II) [ Pb(II)

Labanowski et al. (2007)

Loamy sand soil (pH 5.5–7.0)

Zn(II) [ Cd(II) [ Pb(II) [ Cu(II) [ Cr(III)

Fonseca et al. (2011)

Drainage sediments of residential areas (pH 7.5)

Zn(II) [ Ni(II) [ Cu(II) [ Pb(II) = Cr(III)

Aryal et al. (2007)

Contaminated farmland soils (pH 7.0–8.2)

Cr(VI) [ Cd(II) [ Pb(II)

Dong et al. (2009)

Under oxidation conditions of soil

Cd(II) [ Ni(II) = Pb(II) [ Zn(II) = Cu(II) [ Cr(III)

Kabata-Pendias (2011)

which measure the potential available contents in the soil solid phase, extracted either by exchange with diluted strong acids such as 0.1 M HCl and 0.43 M HNO3 or by chelation with strong complexing organic agents such as 0.05 M EDTA and 0.5 M DTPA. The geochemical process-based models, which are developed to calculate the dissolved free metal ion concentration in the pore water and the actual available metal fraction using the thermodynamic reactions and constants, can be also grouped into several types: (1) models which calculate the free metal ion activity, for instance MINTEQ 3.0, PHREEQ 3 (pH-REdox EQuilibrium), and ORCHESTRA (objects representing chemical speciation and transport models), (2) models which calculate metal ion binding to geocolloids such as humic and fulvic acid, for instance NICA-Donnan model (non-ideal competitive adsorption) and WHAM/Model VI (Windermere humic aqueous model) which are incorporated as sub-models such as in Visual MINTEQ, and (3) models which additionally predict the metal interaction at the site of toxic action (i.e. biotic ligands) and can calculate the toxic effect levels, for instance, BLM (biotic ligand model) and TBLM (terrestrial biotic ligand model). Benefits and limitations of each method are described in Table 6. Table 7 reveals that three of the extraction methods have already been adopted in a regulatory framework, i.e. 0.01 M CaCl2 (the Netherlands; Houba et al.

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1996), 0.1 M NaNO3 (Switzerland; FOEFL 1998a), and 1 M NH4NO3 (Germany; BMUB 1999). Six extraction methods have also been standardized or are under development by ISO, i.e. 0.001 M CaCl2 (ISO/ TS 21268-1, 21268-2, and 21268-3 2007a), 1 M NH4NO3 (ISO 19730 2008), buffered DTPA (ISO 14870 2001), 0.43 M HNO3 (ISO/DIS 17586 2013). In general, when measuring extractant concentrations, centrifugation and filtration procedures are strongly recommended to separate all particles and colloids that would not pass the biological membranes of plant roots (ISO 17402 2008). According to several authors (Birke and Werner 1991; Gryschko et al. 2005; Koster et al. 2005; Meers et al. 2007; Menzies et al. 2007; Tokalioglu et al. 2004), a batch or leaching test employing a neutral salt solution such as CaCl2 or NH4NO3 solutions with a solution/soil ratio of 2:1 is sufficient to measure the bioavailable fraction of relatively highly mobile metals such as Cd, Ni, and Zn. This results from the fact that plants generally uptake these mobile metals from pore water as actual dissolved metal species. The soil contents of these metals extracted with 0.01 M CaCl2 or 1 M NH4NO3 solution are substantially better correlated with the biological contents in plants than those extracted with DTPA or aqua regia (e.g. Birke and Werner 1991). Several authors (Ettler et al. 2007; Houba et al. 1996; Meers et al. 2007; Pueyo et al. 2004) have proposed a 0.01 M CaCl2 solution as

Method

Results stimulate

Donnan membrane technique (DMT)

Diffusive gradient in thin films technique (DGT)

1 M NH4NO3

1/0.1 M NaNO3

0.01/0.0025/0.001 M CaCl2 (batch or percolation test)

Free ion concentration and weak complexes with organic and inorganic ligands (DGT); free ion concentration in pore water (DMT); proved for Cd, Cu, Ni, Pb, Zn

Total concentration in pore water; adequate for highly mobile metals, e.g. Cd, Ni, Zn

0.43 M HNO3

Exchange by strong acid

Potential soluble total concentration in pore water; for Cd, Cr, Cu, Ni, Pb, and Zn

Potential soluble total concentration in pore water; for Cd, Cr, Cu, Ni, Pb, and Zn, particularly for Cr, Cu, Pb

Worst-case scenario within a specific timescale; similarity between 0.05 M EDTA and 0.43 M HNO3; 0.43 M HNO3 can be used to predict phytoavailability by means of soil parameters

Assumes equilibrium condition; requires specialized equipment; timeconsuming; not suitable for routine analysis

0.01 M CaCl2 reduces DOC concentration in soils with low Ca; 1 M NH4NO3 reduces pH in low buffered soils

Benefits and limitations

Predicts effective toxic levels of metals; validated only for barley

Allen et al. (2008), Antunes and Kreager (2009), Lock et al. (2007), Thakali et al. (2006)

Free ion activity, metal-biotic ligand, validated only for Cu and Ni

Benedetti et al. (1996), Bonten et al. (2008), Dijkstra et al. (2004, 2009), Kinniburgh et al. (1996), Weng et al. (2001)

TBLM (terrestrial biotic ligand model)

Assumes that metals in soil solid and in the solution are in equilibrium

Allison et al. (1991), Meeussen (2003), Parkhurst and Appelo (2013)

Ge et al. (2005), Groenenberg et al. (2010), Pampura et al. (2006), Tipping et al. (2003)

Free ion activity and ion binding to dissolved organic matter; validated only for Cd, Cu, Zn, and Pb

NICA-Donnan model (non-ideal competitive adsorption)

Calculate equilibria for inorganic ions

Brun et al. (2001), Davies (1992), De Vries et al. (2005), Echevarria et al. (2006), Groenenberg et al. (2002), Hseu (2006), Madrid et al. (2007), Manouchehri et al. (2006), Mendoza et al. (2006), Menzies et al. (2007), Roemkens et al. (2009b), Remon et al. (2005), Tokalioglu et al. (2004), Zhang et al. (2006)

Koopmans et al. (2008), Pampura et al. (2006), Weng et al. (2005)

Agbenin and Welp (2012), Davison and Zhang (1994), Degryse et al. (2009), Koster et al. (2005)

Birke and Werner (1991), Chojnacka et al. (2005), Davies (1992), Degryse et al. (2003), Ettler et al. (2007), Houba et al. (1996), Meers et al. (2007), Menzies et al. (2007), Pueyo et al. (2004), Roemkens et al. (2009a)

References

WHAM/Model VI (Windermere humic aqueous model)

Free ion activity

MINTEQ 3.0, PHREEQC3, and ORCHESTRA

Mechanistic geochemical models to calculate metal speciation and toxic effect levels (EC50)

0.1/0.5 M HCl

0.05 M EDTA 0.005 M DTPA

Desorption by chelating agents

Chemical methods to measure potentially available metal contents

Separation by diffusion

Exchange by unbuffered salts

Chemical methods to measure actually available metal contents

Mechanism

Table 6 Summary of the most commonly used methods for measuring the bioavailable fraction for plants from soils


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Environ Geochem Health Table 7 Existing legislations and international standards for measurement of phytoavailability of heavy metals in soils Legislation/ISO standards

Extractant

Stimulates

VBBo (Swiss FOEFL 1998a)

0.1 M NaNO3 (2.5 L/kg, 2 h, batch test)

Soil fertility and plant quality

BBodSchV (German BMUB 1999)

1 M NH4NO3 (2.5 L/kg, 2 h, batch test)

Plant quality and toxicity

ISO/TS 21268-1 (2007a) ISO/TS 21268-2 (2007b)

1 mM CaCl2 (2 L/kg, 24 h, batch test) 1 mM CaCl2 (10 L/kg, 24 h, batch test)

Actually available content

ISO/TS 21268-3 (2007c)

1 mM CaCl2 (until 10 L/kg, percolation test)

ISO 19730 (2008)

1 M NH4NO3 (2.5 L/kg, 2 h, batch test)

ISO 14870 (2001)

5 mM DTPA (2 L/kg, 2 h, batch test, pH 7.3)

ISO/DIS 17586 (2013)

0.43 M HNO3 (10 L/kg, 2 h, batch test)

the most preferred extraction medium over 1 M NH4NO3 solution, because (1) the ionic strength of 0.01 M CaCl2 is similar to that of pore water, (2) the low salt concentration reduces the analytical interferences using ICP techniques, and (3) Ca2? is better able to displace metals such as Cd and Zn from exchange sites than NH4?. On the other hand, Schroeder et al. (2005) argued that 0.0025 M CaCl2 solution would be a better choice to mimic pore water, because the ionic strength of clayey soils is usually around 0.0025 M, indicating that 0.01 M CaCl2 overestimated the metal concentrations in pore water. For strongly adsorbed metals, such as Cu, Cr, and Pb, the potentially available contents extracted with 0.05 M DTPA or 0.43 M HNO3 show better correlations with actual uptake (Macholz et al. 2011). This is simply because the neutral salts solutions are too weak to extract the phytoavailable fractions of these less mobile metals. The neutral salts are supposed to displace the readily soluble metal fraction from the exchangeable site into soil solution stimulating the natural pore water, whereas the organic chelating agents such as 0.05 M DTPA are believed to mimic the organic exudates produced by plants, capable of removing metals more aggressively from exchange sites into soil solution (Fang et al. 2007; Han et al. 2006). The contents extracted by 0.05 M DTPA are similar to those of 0.43 M HNO3. Ettler et al. (2007) showed that DTPA procedure was unsuitable for highly organic acid-rich forest soils, because anionic metal–DTPA complexes were re-adsorbed on the positively charged surfaces of soil organic matter and oxides. In addition, Brand et al. (2009) asserted that the 0.43 M HNO3 extraction procedure can also be used to calculate and predict the free metal ion concentration in pore water by modelling using

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Potentially available content

parameters such as pH, DOC, and transfer functions. For determination of strongly chelating metals such as Cu, it is recommended to use ion-selective electrodes, DGT, and DMT or WHAM/Model VI, NICA-Donnan model, and TBLM (Dijkstra et al. 2009; McLaughlin et al. 1997). However, although DGT and DMT provided a significant advantage to measure bioavailability, these methods are not yet considered suitable for routine analysis due to the poor detection limit, time-consuming procedures, and a lack of significant data with which to validate the methods properly (Brand et al. 2009). Mechanistic models also show great promise, but until now, only a few studies illustrating the applicability of these models to a large range of soil types have been published, and consequently, validation for a wide variety of metal concentration, plant, and soil types is still required (see Table 7).

Summary In this paper, we have presented the most recent definitions, both conceptual and operational, of heavy metal bioavailability pertinent to plant uptake from soils. Consequentially, bioavailability can be considered as a complex dynamic process comprising three steps of (1) environmental availability, i.e. the total amount of heavy metal in the soil, including both actual and potential fractions which can be dissolved from the soil matrix into the pore water, (2) environmental bioavailability, i.e. the amount of dissolved fraction in the pore water which can be taken up by plant roots or other soil organisms, and (3) toxicological bioavailability, i.e. the amount of heavy metal which can physiologically induce bioaccumulation or


other effect within plants depending on translocation, metabolism, and detoxification. For each heavy metal, differences in solubility and hence mobility in the soil as well as translocation within the plants are dominated by specific and nonspecific adsorption as well as different complexation affinities for inorganic and organic ligands which to a large degree depend on soil pH. As a result, at common soil pH ranges, heavy metals can be divided into two categories: Cd, Ni, and Zn which exhibit relatively high mobility and Cu, Cr, and Pb which exhibit relatively low mobility. Moreover, we evaluated the most frequently employed methods, either chemical extractions or mechanistic geochemical models, currently used to measure bioavailable metal fraction in soils. Depending on the type of tools used (chemical extraction or mechanistic calculation), strength of extraction media (weak or strong), or mechanism of extraction (exchange, chelation, diffusion, or ion-selective electrode), variable results could be obtained for specific metals. For relatively highly mobile metals (Cd, Ni, and Zn), a neutral salt solution such as 0.01 M CaCl2 or 1 M NH4NO3 is recommended for the measurement of bioavailable fractions. In contrast, for strongly adsorbed, relatively low mobile metals (Cu, Cr, and Pb), extractions with 0.05 M DTPA or 0.43 M HNO3 are recommended. For strongly chelating metals such as Cu, it may also be desirable to employ specific diffusion separation techniques, such as DGT, DMT, and ion-selective electrode, or mechanistic geochemical models such as WHAM/Model VI, NICADonnan model, and TBLM. However, more research still needs to be conducted in the future to properly validate such methods. Overall, for simplicity and ease of operation, to measure bioavailability, in particular in relation to plant quality and toxicity, simple extractions with either 0.01 M CaCl2 and/or 0.43 M HNO3 solutions are presently proposed for routine analysis.

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Effects of freshwater leaching on potential bioavailability of heavy metals in tidal flat soils.

Predicting bioavailability of metals from sludge-amended soils.

The behaviour of heavy metals in sewage sludge-amended soils.

Remediation of heavy metals contaminated saline soils: a halophyte choice?

Bioavailability assessment of phosphorus and metals in soils and sediments: a review of diffusive gradients in thin films (DGT).

Analysis of spatial variations and sources of heavy metals in farmland soils of Beijing suburbs.

Ecological Risk of Heavy Metals and a Metalloid in Agricultural Soils in Tarkwa, Ghana.

The mobility of heavy metals in soils in the Krušné Hory and Slavkovský Les mountains.

Evaluation of heavy metals, cytotoxicity, and antioxidant activity of tomatoes grown in toxic muddy soils.

Lixiviation of natural radionuclides and heavy metals in tropical soils amended with phosphogypsum.

Heavy metals in vegetables and respective soils irrigated by canal, municipal waste and tube well waters.

Multivariate geostatistical analyses of heavy metals in soils: spatial multi-scale variations in Wulian, Eastern China.

Inhibition of the bioavailability of heavy metals in sewage sludge biochar by adding two stabilizers.

Partitioning the heavy metals in soils contaminated by past mining and smelting activities.

Distribution of Heavy Metals in the Soils Associated with the Commonly Used Pesticides in Cotton Fields.

Perspectives in endocrine toxicity of heavy metals--a review.

16S rDNA pyrosequencing analysis of bacterial community in heavy metals polluted soils.

Evaluation of phytoavailability of heavy metals to Chinese cabbage (Brassica chinensis L.) in rural soils.

Health Risk Assessment of Heavy Metals in Soils from Witwatersrand Gold Mining Basin, South Africa.

Pollution of Flooded Arable Soils with Heavy Metals and Polycyclic Aromatic Hydrocarbons (PAHs).

Multivariate-statistical assessment of heavy metals for agricultural soils in northern China.

Ecological risk assessment of heavy metals in soils surrounding oil waste disposal areas.

Mycoremediation potential of Pleurotus species for heavy metals: a review.

Effect of heavy metals on fish larvae deformities: A review.