Role of Coupled Redox Transformations in the ... - ACS Publications

Chapter 21

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Role of Coupled Redox Transformations in the Mobilization and Sequestration of Arsenic Janet G. Hering,*,1 Stephan J. Hug,1 Claire Farnsworth,2 and Peggy A. O’Day3 1Eawag,

Swiss Federal Institute of Aquatic Science and Technology, CH-8600 Dübendorf, Switzerland 2Division of Engineering and Applied Sciences, California Institute of Technology, Pasadena CA 91125 3School of Natural Sciences, University of California, Merced, 95343 *[email protected]

Arsenic occurrence in groundwater, particularly in South and Southeast Asia, has had profoundly deleterious impacts on human health. To address this tragedy, extensive research has been conducted on the biogeochemical cycling of arsenic and its consequences for arsenic mobilization and sequestration. This research has elucidated a key role of microorganisms in redox transformations and the importance of iron and sulfur minerals as carrier phases for arsenic. Research gaps remain, particularly with regard to determining in situ rates of redox transformations and the coupled influence of hydrologic and biogeochemical processes on arsenic occurrence and mobility. Despite these gaps, the insights of this research can be applied to mitigate human exposure through improved water resources management as well as through treatment and remediation.

Introduction The identification of chronic arsenicosis in West Bengal, India in the early-tomid 1990’s (1) motivated extensive studies of human exposure to arsenic and its health effects as well as of arsenic biogeochemistry in recent decades (2, 3). The consumption of groundwater containing arsenic at levels 10 to 100-fold above drinking water standards has resulted in severe human health impacts including © 2011 American Chemical Society In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.


cancer mortality, and estimates of the exposed population throughout Asia number in the hundreds of millions. Chronic exposure to arsenic, primarily through drinking water, poses the greatest risk for the largest populations. About 90% of dissolved inorganic arsenic is passed from the gut into the bloodstream, but most is also rapidly excreted, on the order of days to weeks, with a relatively small fraction retained in organs and tissues (4). The inorganic forms of arsenic, As(III) and As(V), that occur in drinking water are substantially more toxic to humans than organically-bound forms, although specific toxicity associated with the large number of organo-arsenic compounds found in food sources is mostly unknown (5). The extent of exposure and health risk from different arsenic species in food sources, particularly from rice, is receiving more study, but quantitative models of food bioavailability based on sufficiently large populations are lacking (6). Large-scale campaigns to delineate the occurrence of elevated arsenic in groundwater – conducted initially in West Bengal, India and Bangladesh and later in Vietnam, Cambodia, and elsewhere in South and Southeast Asia – revealed some common features, specifically low arsenic concentrations in deep (Pleistocene) aquifers and substantial horizontal variability in arsenic concentrations in shallow (Holocene) aquifers (7). The spatial heterogeneity observed on a local scale poses a particular challenge to identify the geologic, biogeochemical and hydrologic controls on arsenic occurrence. The well-known geochemistry of arsenic was suggestive of several possible mechanisms for As mobilization, including the reductive dissolution of iron(III) carrier phases, oxidative dissolution of sulfide carrier phases, and competitive desorption due to phosphate (3). In these settings, the reductive release mechanism was found to be the most generally dominant. This has also been observed at landfills and biostimulation sites, where inputs of organic carbon drive the reductive release of iron and arsenic (8). In other settings, where sulfides are the primary carrier phase for arsenic, oxidative release can be the dominant mechanism; for example, elevated arsenic concentrations are often observed in acid mine drainage (9). Subsequent oxidative precipitation (i.e., of iron(III) oxyhydroxides) can sequester arsenic in such systems (10). Sequestration of arsenic can also be associated with authigenic sulfide precipitation, particularly in the presence of iron (11). The availability of iron and sulfur, as well as the prevailing redox and pH conditions, can shift the balance between arsenic mobilization and sequestration. Understanding these processes is crucial both to the predication of arsenic occurrence and mobility in groundwater and to the design of remediation strategies and treatment systems. The literature relevant to this topic is vast and cannot be comprehensively reviewed here. The reader is refered to other reviews (3, 7, 12, 13). Here, we seek to provide the background needed to discuss some unresolved questions and issues in this field and to comment on the implications for water resource management, remediation, and mitigation of human exposure to arsenic.

464 In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.


Direct and Indirect Dependence of Arsenic Mobility on Redox Conditions Redox conditions in aquatic systems generally reflect the dominant microbial respiratory processes, which are, in turn, influenced by the supply of electron donors (e.g., organic carbon) and electron acceptors (14). At the circumneutral pH values typical of aquatic systems, the range of redox conditions overlaps with the stability domains of multiple oxidation states of arsenic, iron and sulfur (as well as those of other elements). The redox transformations of arsenic itself (i.e., conversion between the +III and +V oxidation states common to aquatic systems) influence arsenic mobility directly, while those of other elements, particularly iron and sulfur, can influence arsenic mobility indirectly. These interactions are illustrated schematically in Fig. 1, which highlights sorption and precipitation in association with iron and sulfur species as the dominant processes that sequester arsenic.

Figure 1. Schematic illustration of interactions among oxidized and reduced forms of sulfer, arsenic, and iron. Shading indicates concomitant removal of chemical species through sorption and/or co-precipitation processes. Iron(II) minerals that may adsorb or co-precipitate with arsenic include phosphates (e.g. vivianite) or Fe(II,III)-oxides (e.g., green rust-type phases). Minerals containing As(III) or As(V) as a constituent ion have widely varying solubilities and their stabilities also depend on the availability of other constituent ions, which is influenced by prevailing redox and pH conditions (12). In addition, As(III) and As(V) exhibit different, pH-dependent sorption behavior, particularly with regard to sorption on aluminum carrier phases (15) and to the effects of competing sorbates such as phosphate (16). 465 In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.


Slow kinetics, either of redox transformations or of dissolution-precipitation reactions, can, however, skew expectations based on thermodynamics. For example, precipitation of As(V), as arsenate, with Fe(II) in an analog of vivianite (FeII3(PO4)2.8H2O) has been observed in microbial cultures that lack the capacity for As(V) reduction (17) even though concurrent reduction of As(V) and Fe(III) has been observed in other microbial systems and is consistent with the thermodynamics of these reactions (18). Despite these complexities, and bearing in mind the variability due to local conditions, it is a reasonable generalization that, when Fe(III) oxyhydroxides are present, they are likely to be a dominant carrier phase for both As(V) and As(III), while under sulfate-reducing conditions, reduced arsenic is sequestered in the form of sulfides. Maximum mobility of arsenic is thus expected under reducing conditions (where Fe(III) oxyhydroxides are subject to dissolution) with limited sulfide availability (19, 20).

Reductive Dissolution as a Mechanism for Arsenic Mobilization Despite the extensive studies of arsenic mobilization, questions remain regarding the factors that control the rate and extent of the reductive dissolution of Fe(III) oxyhydroxides and the role that As(V) reduction, per se, plays in arsenic mobilization. Reviews on this topic have reported conflicting observations that may reflect the influence of factors such as the mineralogy of the Fe(III) oxyhydroxide substrates, the capacity of microorganisms for Fe(III) and/or As(V) reduction, and the specific conditions in experimental or field systems (13, 21). There is no doubt, however, that microorganisms play a central role in arsenic mobilization, which will thus reflect the level of microbial activity and the inherent capabilities of the microbial community in a given system. Because both Fe(III) and As(V) reduction are dissimilatory processes, it is difficult to identify the specific microorganisms within a microbial community that are responsible for these transformations. Incorporation of C-13 into DNA from labeled acetate concurrent with As(V) reduction has been used as the basis for the provisional identification of the microbes responsible for arsenic mobilization (22). A mass balance approach has been used to attribute organic carbon oxidation to various terminal electron acceptors (including manganese and iron oxides) in the incubation of marine sediments (23). This study highlighted the importance of metal recycling (i.e., consecutive oxidation and reduction transformations). For arsenic, it has been shown that the same biofilm community can both reduce As(V) and oxidize As(III) with the relative importance of these processes being dependent on light conditions (24). Given the difficulty of determining in situ rates of Fe(III) and As(V) reduction directly, reactive transport modeling can be a useful approach to compare laboratory and field experiments and to assess transformation rates under field conditions. Although not yet widely applied, this approach has been used to elucidate the importance of arsenic recycling in lake sediments and its control by the supply of sulfate and organic matter (25). Application of reactive transport models requires better knowledge of site-specific rates associated with coupled 466 In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.


oxidation-reduction reactions, in particular the relationship between rates and the composition and bioavailability of organic matter, and the importance of non-carbon electron acceptors and donors (14). Additional uncertainties include the thermodynamic stabilities and dissolution/precipitation rates of metastable Fe(III)-oxide phases such as ferrihydrite, or Fe(II,III) oxides such as green rust-type minerals, which typically have higher solubilities and surface areas than stable iron oxide minerals. Ideally, reactive transport modeling could be combined with methods to assses transformation rates in situ and/or in laboratory incubations to constrain processes occurring in field systems.

Role of Organic Substrates It is to be expected that the availability and quality of organic substrates would significantly influence the rates of microbial Fe(III) and As(V) reduction. Many, if not most, incubation experiments documenting arsenic mobilization from sediments have used organic carbon amendments (usually acetate or lactate) to stimulate microbial processes. At sites with localized, intense inputs of anthropogenic organic carbon, such as sanitary landfills (26) and biostimulation sites (8), it is clear that these organic inputs are the driver for release of naturally-occurring arsenic. In such systems, arsenic associated with aquifer sediments is immobile in the absence of anthropogenic perturbation. The question of the source and quality of organic carbon supporting arsenic mobilization in Asian aquifer sediments has, however, evoked considerable controversy. Investigations linked arsenic mobilization initially with the mineralization of sedimentary organic carbon (27) or petroleum (28). However, isotopic studies have demonstrated the predominant influence of recent sources of organic carbon (29). Studies suggest that fresh organic carbon delivered to the subsurface by infiltration from surface ponds is important in stimulating arsenic mobilization (30). It has also been suggested that traditional agricultural waste management practices may supply organic carbon from fields (31).

Factors Affecting Arsenic Sequestration Although arsenic sequestration in natural groundwater systems has been less of a focus of research than mobilization, it is clear that processes that lead to mobilization under reducing conditions are at least to some degree reversible under oxidizing conditions. Thus observed net mobilization and sequestration rates often reflect some degree of recycling. Processes in both directions are important in sediments close to the oxic-anoxic interface and particularly in soils that are subject to changing redox conditions (e.g., caused by seasonal or more frequent fluctuations of the water table).



An Example of the Spatial Variability in Arsenic Speciation at a Contaminated Site The variability in arsenic speciation in solid phases reflects the spatial heterogeneity of the environment as a function of sediment composition and texture, water-table level and groundwater flow and, at contaminated sites, the source of arsenic contamination. In tidally-influenced sediments near a former pesticide manufacturing facility, arsenic oxidation state and sequestration in shallow sediments, as As-sulfide, As(III)-oxide, or As(V)-oxide, were related to lithologic horizons and depth to the water table on a cm-to-m scale (20). Arsenic sulfide and minor iron sulfide phases were identified in the upper unsaturated, organic-rich clay layer, while the lower sand-dominated, oxic zone contained only As(V) associated with iron oxide phases (Fig. 2). The zone of transition between the presence and absence of arsenic and iron sulfides corresponded to the approximate seasonal water table level associated with shallow groundwater, and also corresponded to a minimum concentration in sediment arsenic. This example demonstrates, on a relatively small scale, how vertical infiltration of surface salt water promotes sulfate reduction and sulfide precipitation in the upper zone, while the influx of oxygenated water in the aquifer sands leads to arsenic oxidation and sorption. In this setting, arsenic mobility is maximized in groundwater zones that fluctuate between reduced and oxidized conditions. Arsenic Sequestration during Remediation and Treatment Arsenic sequestration has received considerable attention in the context of treatment and remediation. A number of treatments in engineered or augmented systems take advantage of the same processes by which arsenic is sequestered under natural conditions. Reductive sequestration has been proposed for in situ remediation (32, 33). This is likely to require amendment with sulfate to overcome the limitation by sulfate availability commonly observed in freshwater systems (19, 20). Formation of sulfide minerals has also been invoked as a mechanism for arsenic sequestration in subsurface permeable reactive barriers employing zerovalent iron, ZVI (34). While iron corrosion provides Fe(II)- and Fe(III)-phases for arsenic adsorption or co-precipitation, the availability of dissolved sulfate and organic carbon in field settings determines the extent to which sequestration in sulfide minerals can contribute to arsenic immobilization and hence to the effectiveness and longevity of the barrier. In (ex situ) treatment systems, however, arsenic removal is more commonly based on sequestration under oxic conditions. For example, since groundwaters in South Asia often contain elevated concentrations of both Fe(II) and arsenic, the oxidative precipitation of Fe(III)-phases offers the possibility of concomitant removal of arsenic. The composition, mineralogy and reactivity of Fe(III)-phases formed by oxidation of dissolved Fe(II) by dissolved O2 are, however, influenced by the solution composition. In waters with low phosphate and silicate concentrations, Fe(III)(hydr)oxides such as lepidocrocite, ferrihydrite, and goethtite are typically precipitated, while mixed Fe(III)phosphate-hydroxides and 468 In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.


silicate-rich Fe(III)-phases are formed in water with higher dissolved phosphate and silicate concentrations (35, 36). Oxidative precipitation of Fe(III)-phases is also significant in that the oxidation of Fe(II) by O2 can promote co-oxidation of As(III). The reactive oxygen intermediates, such as O2-, H2O2 and OH-radicals and possibly higher-valent (e.g., +IV) iron species that are formed during the reaction of Fe(II) with O2, are also able to oxidize As(III) (37). The oxidation of As(III) to As(V) concomitant with Fe(II) oxidation tends to promote arsenic uptake into the Fe(III)-phases formed at circumneutral pH in the presence of phosphate and silicate (38).

Figure 2. Results from a sediment core showing total As and Fe concentrations (a), As XANES spectra from corresponding depths (b), and example scanning electron micrographs (c) (at 61-76 cm from the same core and at 259-274 cm from a similar core ~15 m away). XANES results show the transition from sulfide-associated As in the upper unsaturated zone to a lower oxidized zone with only As(V), which corresponds to the approximate level of the seasonally-averaged groundwater table (horizonal line in (a)). The upper reduced sediments contain rare cubic Fe- and S-rich crystals, while the lower aquifer sediments contain abundant surface coatings on quartz grains. Modified from (20).



These processes are successfully exploited in the removal of arsenic using sand filters in Vietnam, where Fe(II) concentrations of up to 40 mg/L occur naturally in the groundwater (39). In addition to co-oxidation with Fe(II), As(III) oxidation in the top layer of sand is most likely accelerated by a population of microorganism that is typically established within days. Manganese (III,IV) oxides, formed by oxidative precipitation of Mn(II) within the filters, can also oxidize As(III). In contrast, naturally-occurring Fe(II) in groundwater in Bangladesh is insufficient to achieve effective arsenic removal, particularly in the presence of elevated phosphate and silica concentrations (40). In the SONO-filter, a successful and widely-used household filter, corrosion of metallic iron serves as a continuous source of Fe(II) for several years (41). The continuous Fe(II) release supports co-oxidation of As(III) and arsenic removal during the oxidative precipitation of Fe(III)-phases. It is likely, however, that these filters are populated by (as yet uncharacterized) microorganisms that also promote microbial As(III) and Mn(II) oxidation. Furthermore, iron corrosion processes form magnetite, maghemite, hematite and other solid phases that can also sorb arsenic (42). Transport limitations at corroding iron surfaces resulting in anoxic regions with locally and/or temporally high Fe(II) concentrations can lead to transformations of As-containing Fe(III)-phases. Reduction of As(III)- or As(V)-containing ferrihydrite by microbially produced Fe(II), for example, forms magnetite with adsorbed As(III) or incorporated As(V), respectively (43). Transformation of Fe(III)-phases to more dense phases such as magnetite might prevent clogging of filters. The strong affinity of arsenic both to sorb to and coprecipitate with iron oxides has also been exploited in the development of technologies for removing arsenic from soil or stabilizing it in situ. A variety of treatments with different forms of iron oxides, iron sulfates, and recycled iron metal have been used to immobilize arsenic in contaminated geomedia (44). The similarity in chemical behavior between arsenate and oxyanions such as sulfate and phosphate has also been exploited as a means to lower arsenic mobility under oxidizing conditions through addition of Portland cement and sulfate treatments to stabilize contaminated soils (45). Lime or organic amendments combined with stabilization treatments for arsenic have been used to promote plant growth for phytostabilization in addition to chemical stabilization.

Coupling of Hydrologic and Biogeochemical Processes The influence of transport processes has often been neglected in biogeochemical studies of arsenic mobilization and sequestration. Investigations targeting the supply of organic carbon, sulfate, and dissolved oxygen have, however, drawn attention to the importance of transport processes and constraints. As already mentioned, arsenic mobilization is particularly sensitive to the supply of organic carbon (46, 47); the localization of arsenic sequestration reflects the supply of sulfate (19) and dissolved oxygen (20, 48, 49). Groundwater seepage into (oxygenated) surface water creates a gradient in redox conditions conducive 470 In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.

to arsenic sequestration (48). Water table fluctuations associated with irrigated agriculture (29, 49) or intermittent pumping in riverbank filtration systems (50) can significantly perturb redox conditions and promote redox cycling.


An Example of Redox Cycling in Rice Fields Successive mobilization and sequestration of arsenic under varying redox conditions has been observed in irrigated rice fields in Bangladesh. In yearly cycles, arsenic accumulates in top soils during irrigation and is partly remobilized and redistributed during monsoon flooding (49). On a shorter time scale, arsenic is also mobilized and sequestered during irrigation (51). After the rice is planted, the fields are kept flooded for several weeks, mainly to limit weed growth and to establish a robust population of rice plants. During this time, anoxic conditions develop in the topsoil (0-25 cm depth) and arsenic concentrations of up to 500 µg/L are measured in the soil porewater with >70% of the arsenic occurring as As(III); elevated concentrations of Fe(II) and Mn(II), up to 10 and 2 mg/L respectively, are also observed. Once the rice plants are established, the fields are irrigated less frequently and the soils become dry at the surface. Partly oxic conditions develop down to a depth of 20 cm; only low concentrations of dissolved Fe and Mn were observed with arsenic, predominantly (>70%) as As(V), occurring at concentrations below 50 µg/L. Although the limited time resolution of pore water samplers precludes the determination of rates of arsenic-mobilization and sequestration, the timescales for these processes may be constrained to days or less. Thus the practice of irrigation with alternating wet-dry cycles may significantly reduce the exposure of plant roots to arsenic and the subsequent uptake of arsenic by the rice plants.

Open Issues Related to Human Exposure, Treatment, Remediation and Water Resources Management The extensive and multi-faceted studies of arsenic in recent decades were motivated, in large part, by the human tragedy in South and Southeast Asia. While much has been learned, the problem of human exposure has not been adequately addressed and open questions remain regarding the future consequences of management, remediation, and treatment strategies. Quantitative predictions of arsenic mobility, attenuation, or remediation effectiveness are limited by insufficient knowledge of the range of reaction rates associated with multiple redox processes, and of the coupling between biogeochemical and hydrologic factors at specific sites. Minimizing the current human exposure to elevated concentrations of arsenic in drinking water is clearly the most pressing need. This can be accomplished either by treating the contaminated water to achieve an acceptable quality or substituting an alternate, uncontaminated water source. Well-switching and drilling of deep wells are the most common practices to avoid the arsenic contamination in shallow aquifers in South Asia (52), but the effectiveness 471 In Aquatic Redox Chemistry; Tratnyek, P., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2011.


of these strategies may be compromised by lack of convenience and public acceptance (53). There is no shortage of effective technologies for the removal of arsenic from drinking water and many of these can also be used at the household scale (3). In developing countries, cost is a major barrier to implementation (52), and it must also be recognized that arsenic occurrence in drinking water is only one of many pressing public health concerns. Even in developed countries, the option of drinking water treatment presents the issue of appropriately handling treatment residuals. This concern has been raised in regard to the use of Fe(III)-based coagulants or sorbents for arsenic removal; disposal of arsenic-rich treatment residuals in municipal landfills could result in arsenic mobilization under reducing conditions (54). Preventing (further) degradation of both shallow and deep aquifers is an important concern for the future. In the long-term, use of deep wells may also lead to their contamination with arsenic (and potentially with other metals or organic contaminants) derived from the surface or shallow aquifers (4, 55). Intensive use of deep aquifers for irrigation is likely to lead to future contamination of deep wells and thus should be avoided (7). The question of whether the modification of current agricultural practices could influence the amount or composition of organic carbon that reaches shallow aquifers and thus arsenic mobilization deserves further investigation.

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