Effects of Competing Anions and Iron Bioreduction on Arsenic ...

Water Air Soil Pollut (2012) 223:5707–5717 DOI 10.1007/s11270-012-1308-0

Effects of Competing Anions and Iron Bioreduction on Arsenic Desorption Juscimar Silva & Jaime Wilson Vargas de Mello & Massimo Gasparon & Walter Antônio Pereira Abrahão

Received: 1 February 2012 / Accepted: 29 August 2012 / Published online: 18 September 2012 # Springer Science+Business Media B.V. 2012

Abstract Dissimilatory iron-reducing bacteria play a fundamental role in catalysing the redox transformations that ultimately control the mobility of As in anoxic environments, a process also controlled by the presence of competing anions. In this study, we investigated the decoupling of As from loaded Al and Fe (hydr)oxides by competing anions in the presence of iron-reducing bacteria. Hematite, goethite, ferrihydrite, gibbsite and three aluminium-substituted goethites (AlGts) were synthesised and loaded with arsenate, followed by anaerobic incubation with different phosphate or carbonate-containing media in the presence of catalytic iron-reducing bacteria. Soluble Al, As, Fe and P contents were measured in aliquots J. Silva (*) Embrapa Vegetables, Rod. BR 060, km 09, Brasília/Anápolis, 70359-970, CP 218 Gama, Federal District, Brazil e-mail: [email protected] J. W. V. de Mello : W. A. P. Abrahão Soil Department, Federal University of Viçosa, Av. P.H. Rolfs, s/n, Viçosa, Minas Gerais, Brazil 36570-000

by inductively coupled plasma optical emission spectrometry following periodical sampling. Shewanella putrefaciens cells were able to utilise both noncrystalline and crystalline Fe (hydr)oxides as electron acceptors, releasing Fe and As into solution. Phosphate and carbonate affected the Fe bioreduction, probably due to the precipitation of metastable mineral phases and also to phosphate-induced stabilisation on the hydroxide surfaces. Phosphate precipitation acted as a sink for As, thus limiting its mobilisation. The highest fraction of desorbed As by phosphate was observed for gibbsite, followed by AlGts. Similarly, gibbsite showed significant amounts of arsenate displaced by carbonate. In spite of its low crystallinity, ferrihydrite was the most efficient compound in retaining arsenate, possibly due to As co-precipitation. This study provides new insight into the management of As-contaminated soils and sediments containing Algoethites and gibbsite, where the Fe activity may be too low to co-precipitate As-bearing vivianite. Thus, the dynamics of As(V) in flooded soils are significant in agriculture and environmental management. Keywords Arsenic contamination . Al-substituted goethites . Soil . Sediments . Redox stability

M. Gasparon School of Earth Sciences, The University of Queensland, St. Lucia, Brisbane, QLD 4072, Australia

1 Introduction

J. Silva : J. W. V. de Mello National Institute of Science and Technology, INCT Acqua, Belo Horizonte, Brazil

Coupling arsenic mobilisation with microbial reductive dissolution of Fe (hydr)oxide is one of the most

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important aspects of the arsenic biogeochemistry. The natural input of arsenic in the environment is mainly due to the weathering of As-bearing rocks and minerals. Anthropogenic sources, i.e. mining and smelting activities, pesticides and wood preservative are potential additions to the natural budget (Matschullat 2000; NHRMC and NRMMC 2004; Smedley and Kinniburgh 2002). An increase in the global cycling of As can be expected due to progressive industrialisation of developing nations (Leist et al. 2000) as well as to the increased exploitation of groundwater resources. Arsenic in nature can be found as inorganic and organic compounds, in several valence states, i.e. −3, −1, 0, +3 and +5. In terrestrial environments, As occurs mainly in inorganic forms as trivalent arsenite [As(III)] (as H 3AsO 3) or pentavalent arsenate [As(V)] (as H2AsO4− and HAsO42−). The lower is the valence state, the higher is the toxicity of As compounds, with the inorganic forms being more toxic than the organic ones. Despite the different degrees of toxicity, there is no distinction between As species in water quality standards (Kim et al. 2002). Redox potential (Eh) and pH control As distribution, and arsenite is expected to be the dominant species in anoxic conditions. However, due to relatively slow transformation following changes in redox conditions, both As(III) and As(V) can be often found in either redox environment. Aluminium and Fe (hydr)oxides are ubiquitous reactive constituents of soils, sediments and aquifers. In highly weathered tropical soils, these minerals are naturally abundant, and due to their high reactivity, they play a fundamental role on the biogeochemical cycle of many elements. The distribution of As under oxidising conditions is mostly controlled by these minerals, and for this reason, soil materials have been used as liners in remediation procedures at contaminated sites. Ladeira and Ciminelli (2004) working with an oxisoil sample composed primarily of gibbsite, goethite and Al silicate minerals verified that this material would be adequate as a protective liner in waste dams due to its remarkable As immobilisation capability. In the absence of oxygen, however, Fe(III) minerals can be used as electron acceptors during microbial respiration. Therefore, based on the association between many trace elements and Fe(III) (hydr)oxides, and the tendency of these minerals to be dissolved under suboxic conditions, biological Fe reduction can have a major impact on the persistence and mobility of toxic elements,

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radionuclides and organic contaminants under such conditions (Cummings et al. 1999; Zachara et al. 2001). Cummings et al. (1999) investigated the biological reduction of synthetic scorodite and noted that dissimilatory Fe reduction led to the release of As(V) and Fe(II) into solution. Conversely, Al is rather stable under low Eh conditions and its intrusion into the Fe(III) (hydr)oxides structure by isomorphic substitution reduces the rate of Fe reductive dissolution, as already reported (Bousserrhine et al. 1999; Jeanroy et al. 1991; Schwertmann 1984). Aluminium replacement of Fe under natural conditions is a very common process, and therefore most Fe minerals found in soils, particularly from tropical climates, show a certain degree of Al for Fe substitution. The use of such soils or precipitated mineral phases combining the higher binding affinity of Fe (hydr)oxides for As, and the higher stability of Al under anoxic conditions, may be advantageous for As immobilisation. Very few studies have focused on As sorption processes onto Al-substituted Fe (hydr)oxides or As release from these compounds. Masue et al. (2007) reported a decrease in both As(III) and As(V) adsorption onto Al/Fe hydroxides and an increase in As desorption by phosphate as the Al/Fe molar ratio increased. Recently, Silva et al. (2010) reported that the presence of Al substituting Fe in the goethite structure increased both the stability of goethite under low redox conditions and its As adsorption capacity, being about three times more efficient than pure goethite. Natural attenuation of As by sorption onto Al and Fe minerals may be also limited by other anions competing with arsenate for sorption sites (Hongshao and Stanforth 2001; Sahai et al. 2007; Zhang et al. 2008). Due to similar acid dissociation constants, phosphate (pKa1 0 2.1, pKa2 07.2 and pKa3 012.3) behaves much like arsenate (pKa1 02.2, pKa2 06.9 and pKa3 011.4). Besides phosphate, carbonate may also, to a lesser extent, affect the arsenate sorption reactions. The displacement of adsorbed As by carbonate was theoretically examined by Appelo et al. (2002) and this mechanism was proposed to be potentially one of the main reasons for high As concentrations in groundwater. Although investigations of competition between anions can provide insight into the reactions occurring on the mineral surface (Hongshao and Stanforth 2001), most of the As desorption studies have evaluated those reactions individually. Nevertheless, As mobilisation in natural environments may depend on


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the combination of different factors, such as the presence of competing anions and bacterial catalytic activities. Therefore, this work investigates the decoupling of As from Al-goethites and other synthetic Al and Fe (hydr)oxides by competing anions in the presence of dissimilatory iron-reducing bacteria.

2 Materials and Methods 2.1 Synthesis of Al and Fe (hydr)oxides Hematite (Hm), goethite (Gt) and two-line ferrihydrite (Fh) were synthesised according to Schwertmann and Cornell (2000). Goethites with 13, 20 and 23 cmolmol−1 of Al, henceforth referred to as AlGt13, AlGt20 and AlGt23, respectively, were prepared by precipitation following the reaction of ferrous and aluminium chloride solution with a potassium hydroxide solution and settling in plastic bottle for 90 days. Slow oxidation of Fe2+ to Fe3+ and incorporation of Al3+ in the goethite structure were achieved by opening the bottle daily and stirring the suspension vigorously for 5 min. To remove the excess of Al, the precipitates were washed twice with a 0.01 molL−1 KOH solution. Precipitates of (hydr)oxides were washed several times with Milli-Q water, centrifuged and air-dried at 50 °C in an oven, except for Fh, which was frieze-dried to prevent crystallisation. Gibbsite (Gb) was prepared after titrating an Al(NO3)3 solution with 4 molL−1 NaOH to a pH of 4.6± 0.2 (Kyle et al. 1975). The gelatinous precipitate was heated for 2 h at 40 °C, then washed twice, dialysed

against Milli-Q water for 36 days and air-dried at 50 °C. The mineral phases were characterised by X-ray diffraction, diffuse reflectance and Raman spectroscopy (data to be published elsewhere). Specific surface area was determined by N2 adsorption using the multiple point technique (Quantachrome model NOVA 1000, see Table 1). Samples were degassed at 110 °C for 2 h under a continuous N2 stream prior to surface area determination. These adsorbents were equilibrated with sufficient As(V) to achieve their maximum As adsorption capacity (Silva et al. 2010). Arsenate solutions were prepared by dissolving analytical reagent-grade disodium hydrogen arsenate heptahydrate (Na2HAsO4·7H2O; Ajax Finechem) in Milli-Q water. Solid samples (0.1000 g; Hm >


Fh. Arsenic release commenced readily after phosphate addition (∼4 h), increasing quickly up to 48 h and decreasing slowly thereafter (Fig. 1). For example, nearly 70 % of the total As content was released from Gt within 48 h. Considering that no soluble Fe was detected during the initial interval (4–48 h), the As displacement may be attributed exclusively to ligand exchange reactions. More than 50 % of the As was released from Gb and Al-goethites within the time frame of the experiment. For Gb, this can be attributed entirely to exchange by phosphate on the Gb surface, but for Algoethites, part of the As released resulted from Fe bioreduction. The fraction of As mobilised from Algoethites increased with structural Al content (Fig. 1, AlGt13, AlGt20 and AlGt23), suggesting that the released As was preferentially associated with Al sites. Such weak As retention to Al sites suggests the formation of unstable surface complexes likely related to surface precipitation of As on these adsorbents. This hypothesis is not fully supported by Ladeira et al. (2001), who showed that As forms an inner-sphere bidentate binuclear complex on the surface of Al oxyhydroxyl octahedra. Detailed spectroscopic investigation is warranted to further elucidate the type of surface complexes formed between As and those matrices. Our data are in agreement with the report of Masue et al. (2007) of increasing As(V) desorption by phosphate from poorly crystalline Al-Fe hydroxides as the Al/Fe ratio increased, reaching almost total arsenate desorption from pure Al hydroxide after 24 h. The much higher P/As (7500:1) ratio used by Masue et al. (2007) in their experiment may account for the faster and higher amount of As desorbed. The relatively low concentration of As desorbed by phosphate from crystalline Fe (hydr)oxides (26 % for Fh and Hm and 37 % for Gt) is consistent with the findings of Hongshao and Stanforth (2001), who suggested that arsenate is adsorbed mainly as a nonexchangeable ion, although a lesser amount is exchangeable by phosphate (Fig. 1). Indeed, As(V) or phosphate adsorption kinetics are considered as a twophase reaction, with a rapid initial step followed by a much slower reaction (Hongshao and Stanforth 2001; Torrent et al. 1992; Zhang and Selim 2008). A different behaviour was observed for Fh, where the amount of As released reached its maximum by 192 h and then suddenly dropped to almost 0. This depletion of soluble As coincided with P immobilisation, which is


likely to be due to biogenic vivianite formation, as stated earlier. Unlike for P, there was no expectation for As immobilisation due to Fe(II)-As phase precipitation; however, arsenate can be associated with As-bearing vivianite by co-precipitation reactions, as arsenate behaves much like phosphate. This hypothesis is realistic considering the structure of arsenate and phosphate minerals, where typical tetrahedral anions [XO4]3− are bound to octahedrally coordinated transition metal ions. Variations and multiples of these bonding patterns tend to create relatively open structures that allow for extensive substitution of cations, anions, anions groups and water (O'Day 2006). Consequently, the precipitation of Fe(II)-P phases could provide a sink for As, giving rise to vivianite-like minerals [Fe3(II)[(As, P)O4]2·8H2O]. The adsorption of As onto vivianite surface may to a lesser extent explain the depletion in soluble As noted by Thinnappan et al. (2008), who showed that sorption of arsenate onto natural well-crystallised vivianite in the pH range 3–11 varied between 25 and 40 % from a starting concentration of 100 μmolL−1. In addition, the adsorption was complete for lower As(V) concentrations (1 and 10 μmolL−1), and the sorption reactions were little affected by pH variations. These observations have significant environmental implications and provide useful information on the fate of As in eutrophicreducing aquatic systems. Adsorbents not loaded with As(V) showed similar P immobilisation patterns, with P precipitation/sorption reactions taking place in the first hours, and equilibrium attained over the course of the experiment (Fig. 2). The amount of P immobilised was, however, greater than that observed in the As-loaded adsorbents, due mostly to the higher availability of adsorptive sites. The largest difference was noticed in Fh, where P immobilisation varied from 30 to 50 % in the presence and absence of adsorbed As(V), respectively. For Al-goethites, the differences were smaller than for the well-crystallised Gb, Gt and Hm and corresponded roughly to the amount of available adsorption sites, as measured by maximum As adsorption capacity (Table 1). As previously addressed, the presence of adsorbed oxyanions decreased the efficiency of bacteria in using Fe as an electron acceptor due to phosphate coating. This effect is due to electrostatic repulsion between negatively charged mineral surface and bacteria cells, leading to phosphate-induced stabilisation. Similar results were observed by Borch et al. (2007) for Fh,

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where phosphate induced alterations in surface reactivity, by decreasing Fe(III) reduction nearly linearly for increasing P surface coverage. The amount of bioreduced Fe was quite low and similar to that measured in the presence of P and As(V), especially for Al-goethites that showed even lower Fe concentrations (