Environ. Sci. Technol. 2004, 38, 1794-1798
Coupling Speciation and Isotope Dilution Techniques To Study Arsenic Mobilization in the Environment R E B E C C A E . H A M O N , * ,† E N Z O L O M B I , † PAOLO FORTUNATI,‡ A N N E T T E L . N O L A N , †,⊥ A N D M I K E J . M C L A U G H L I N †,§ CSIRO Land & Water, PMB 2, and School of Earth and Environmental Sciences, The University of Adelaide, Glen Osmond, South Australia 5064, Australia, and Soil Chemistry Section, Faculty of Agricultural Sciences, Agricultural and Environmental Chemistry Institute, Universita` Cattolica del Sacro Cuore, Via Emilia Parmense 84, I-29100 Piacenza, Italy
We have developed an approach to isolate mechanisms controlling mobility and speciation of As in soil-water systems. The approach uses a combination of isotopic exchange and chromatographic/mass spectrometric As speciation techniques. We used this approach to identify mechanisms responsible for changes in the concentration of soluble As in two contaminated soils (Eaglehawk and Tavistock) subjected to different redox conditions and microbial activity. A high proportion of the total As in both soils was present in a nonlabile form. Incubation of the soils under anaerobic conditions led to changes in the concentration of soluble As in each soil but did not change the As speciation or the proportion of total As in labile forms in the soils. Hence, a decrease in soluble As in the Eaglehawk soil was the result of an Eh-induced pH decrease, enhancing the solid-phase sorption of As(V). An increase in soluble As in the Tavistock soil was due to an Eh-induced pH increase, decreasing solid-phase sorption of As(V). Incubation of the soils under aerobic conditions with microbial activity stimulated by addition of glucose resulted in no change in the solution concentration or speciation of As in the Eaglehawk soil, but led to a large increase in the concentration of soluble As in the Tavistock soil. This increase was due to conversion of exchangeable forms of As(V) into less strongly sorbed As(III) species. Incubation under anaerobic conditions in the presence of glucose resulted in a large increase in the concentration of soluble As in both soils; however, different mechanisms were found to be responsible for the increase in each soil. In the Eaglehawk soil higher concentrations of As were again due to conversion of exchangeable forms of As(V) into less strongly sorbed As(III) species. In contrast in the Tavistock soil, the increased As in solution was the * Corresponding author phone: +61 8 8303 8489; fax: +61 8 8303 8565; e-mail:
[email protected]. † CSIRO Land & Water. ‡ Universita ` Cattolica del Sacro Cuore. § The University of Adelaide. ⊥ Present Address: National Analytical Reference Laboratory, NSW, Australia. 1794
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result of release of As(V) from the large reservoir of nonlabile soil As.
Introduction Elevated concentrations of arsenic (As) in waters and soils pose a significant threat to human health. This has been highlighted by recent findings that groundwaters used directly for human consumption as well as for irrigation of rice crops in large areas of Bangladesh and elsewhere are highly enriched in As, resulting in widespread As poisoning of the local populations (1, 2). The mobility and toxicity of As in water and soil systems is controlled by a complex array of biotic/abiotic sorption, precipitation, and redox processes which dictate its solid-phase partitioning and speciation. Due to this complexity, the relative contribution of these processes in governing As mobility in any given system remains to be unequivocally resolved. The concentration of As in waters and its bioavailability in soils are highly dependent on the ability of the associated solid phase (i.e. soil, sediment, aquifer material) to capture As, as well as its ability to retain As under changes in conditions such as fluctuating pH, redox, and oxidative/reductive microbial activity (3, 4). Under changing conditions, As concentrations in solution are initially buffered by reactive forms of As (labile As) associated with solid-phase exchange sites. The labile pool of As is therefore of fundamental importance because it represents the most chemically reactive and biologically relevant fraction of As. Under aerobic conditions, labile As is thought to occur predominantly as the anionic As(V) species (arsenate), which adsorbs strongly to positive charge sites associated with solid-phase iron and aluminum (hydr)oxides at acidic pH (5, 6). However, the presence of the more toxic As(III) species (arsenite) has also been observed in aerobic, but microbially active, systems (7). The As(III) species is largely uncharged under environmentally relevant pH conditions (pKa1 ) 9.2), and thus is thought to participate less substantively than As(V) in exchange reactions on many types of solid-phase surfaces (5). Hence, two mechanisms which could be responsible for mobilization of As from solid-phase exchange sites include a decrease in positively charged solid-phase exchange sites due to an increase in the system pH, or conversion of labile As(V) to As(III), either of which will result in liberation of adsorbed, labile forms of As. However, the solid phase can also host nonexchangeable (nonlabile) forms of As. Indeed recent studies have demonstrated that labile forms of As constitute only a small proportion of the total As in many contaminated soils, with >70% of the total As commonly observed to be present in nonlabile forms, probably occluded within iron (hydr)oxides (8, 9). This is likely to be similar for As-rich sediment and aquifer materials, with As-Fe-S precipitates proposed as a further nonlabile solid-phase sink for As in reducing systems (10). The nonlabile fraction of As is not in direct equilibrium with aqueous As species, and hence is not bioavailable per se. However, environmental perturbations may release nonlabile As into the labile pool as follows. Changing pH can solubilize solid-phase constituents. For example, the solubility of iron (hydr)oxides increases significantly below pH 4.5 (11). If nonlabile As is associated with iron (hydr)oxides, it can therefore be released into solution with decreases in pH, contrasting with the expected effect of pH on labile forms of As described above. Altered redox conditions and microbial activity can also change the solubility of the solid-phase constituents by causing reductive dissolution of iron (hydr)10.1021/es034931x CCC: $27.50
2004 American Chemical Society Published on Web 02/14/2004
TABLE 1. Selected Properties of the Soils soil property
Eaglehawk
Tavistock
soil property
Eaglehawk
Tavistock
pH amt of organic C (g/kg) amt of CaCO3 (g/kg) amt of clay (g/kg) amt of silt (g/kg)
8.3 5 11 10 40
4.9 47 0 100 350
amt of sand (g/kg) total As amt (mg/kg) dissolved As concn (aerobic, mg/L) labile As concn (aerobic, % of total)
940 725 1.4 12
550 4770 0.29 2.6
oxides or oxidative dissolution of sulfide materials (7, 10). These types of mechanisms could be responsible for potentially catastrophic release into solution of contaminants such as As which are stored in large quantities in nonlabile forms in the solid phase (12). Simple measurements of concentration changes in solution following system perturbation cannot differentiate between elements captured or released by the labile versus the nonlabile solid-phase pool (12-14), nor can they quantify accurately changes which may have occurred in the oxidation state of labile As. Therefore, this type of assessment cannot provide a mechanistic explanation of the processes involved in As mobilization/immobilization. Other techniques which have been employed to examine As speciation are similarly unable to reveal the mechanisms underlying As mobilization/immobilization. For example, while the stability of different solid-phase As associations has been assessed through examination of the coordination environment of As using X-ray spectrometric methods (e.g., refs 10 and 15), these techniques cannot discriminate between nonlabile and labile fractions of elements (16) and are unable to speciate the chemically and biologically reactive (labile) As fraction. Hence, though there has been much reasoned speculation as to the mechanisms controlling As speciation and partitioning in different materials, to date the actual mechanisms have yet to be confirmed analytically (4). One approach for quantifying element partitioning between labile and nonlabile pools is the isotopic dilution methodology. The isotopic exchange E value method has been applied successfully to investigate the stability of solid-phase associations of various elements (e.g., refs 9, 12, and 17-20). However, determination of the isotopically labile pools of two different oxidation states of an element has not been attempted until now. The aim of this study was therefore to unravel and identify the mechanisms involved in mobilization of As under conditions of changing redox status and microbial activity through the development and use of an isotopic dilution method coupled with speciation of aqueous stable As (by HPLC-ICP-MS) and the radioisotope 73As (by HPLC and γ spectrometry).
Materials and Methods Soil Treatments. An acidic soil (Tavistock, U.K.) and an alkaline soil (Eaglehawk, Australia), both contaminated with As as a result of historical mining activities, were sampled for use in this study (Table 1). The soils were sieved ( 0.05) irrespective of the original oxidation state of isotope introduced (Figure 1), thus validating our hypothesis. Moreover, the fact that both oxidation states of the isotope gave rise to the same E values demonstrates that the low amount of NH2OH‚HCl that was added with the As(III) form of the isotope had no effect on the equilibrium of the As species in the soils. It is important to note that there is the potential to incur large errors in E value determinations for elements exhibiting more than one oxidation state, unless the existence of each oxidation state is recognized and accounted for, or unless each oxidation state of the element has the same Kd. This can be demonstrated by examining the calculations that would be used to assess the E value in the case where the oxidation states are accounted for, and comparing this to the incorrect case where they are not. The correct equation for calculating the total E value (Etot) for an element having two oxidation states can be written as follows: 9
Etot ) ((Sox1/S*ox1)(R1) + (Sox2/S*ox2)(R2))(v/m)
(3)
where R1 + R2 ) R, i.e., the total radioactivity introduced to the system, and the other symbols are as defined above. However, if the system is not speciated, and hence the concentration and activity of the different species are not accounted for, the Etot would be incorrectly calculated (IEtot) as
IEtot ) ((Sox1 + Sox2)/(S*ox1 + S*ox2))(R1 + R2)(v/m) (4) which can be rearranged as
(1)
Eox2 ) (Sox2/S*ox2)(R - (Eox1(S*ox1/Sox1)(m/v)))(v/m) (2)
1796
FIGURE 1. E values of various treatments obtained using 73As(V) radioactive spike compared to E values of the same treatments obtained using 73As(III) radioactive spike. Filled and hollow circles indicate E values (mg/kg) for the As(III) and As(V) species in the soil, respectively. The line indicates the regression fit of the relationship.
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 38, NO. 6, 2004
IEtot ) ((Sox1/(S*ox1 + S*ox2))(R1 + R2) + (Sox2/(S*ox1 + S*ox2))(R1 + R2))(v/m) (5) The function Etot will approach IEtot when Eox2 approaches 0, but in soils containing appreciable quantities of element in both oxidation states, comparison of eqs 3 and 5 shows that Etot will only be equal to IEtot when Kd,ox1 ) Kd,ox2. The Kd for each As species is unlikely to be the same in most soils (5). Indeed, calculation of IEtot and Etot values for the experiments described in this paper showed that IEtot overestimated Etot by more than an order of magnitude in samples which contained high amounts of As(III) relative to As(V) (data not shown). Hence, accurate determination of the E value for As in soils containing both species requires use of the coupled speciation/isotope dilution procedure as described here. As shown below, application of the isotopic dilution method can provide valuable information about the location of elements in soil S water systems, allowing mechanistic insights into processes underlying element mobilization and immobilization that would not otherwise be possible. However, the above example and others (e.g., refs 19 and 23) demonstrate that care needs to be exercised to ensure that the method is appropriately applied. Mechanisms Controlling As Mobility. The labile fraction of As was found to be a small proportion of the total As in both of the soils (i.e., both soils contained a large amount of As stored in nonlabile forms) (Table 1). The values reported here for labile As are close to values found previously for soils contaminated by mine activity, where
