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Fe(II)–Induced Mineral Transformation of FerrihydriteOrganic Matter Adsorption and Coprecipitation Complexes in the Absence and Presence of As(III) Chunmei Chen, and Donald L. Sparks ACS Earth Space Chem., Just Accepted Manuscript • DOI: 10.1021/ acsearthspacechem.8b00041 • Publication Date (Web): 13 Sep 2018 Downloaded from http://pubs.acs.org on September 14, 2018
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ACS Earth and Space Chemistry
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Fe(II)–Induced Mineral Transformation of Ferrihydrite-Organic Matter Adsorption and
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Coprecipitation Complexes in the Absence and Presence of As(III)
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Chunmei Chen* and Donald L. Sparks
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Department of Plant and Soil Sciences
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Delaware Environmental Institute
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University of Delaware, Newark, DE, USA 19711
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Corresponding Author
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*Phone: (302)8318345. Fax: (302)8310605. E-mail:
[email protected] 21 22
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Abstract The poorly crystalline ferrihydrite (Fh) is often associated with organic matter (OM) and
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metal(loid)s like arsenic (As). The transformation of Fh to more stable and hence less reactive
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phases is catalyzed by surface reaction with Fe(II). However, little is known regarding the impact
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of various specific OM types and the co-existence of OM and As on the secondary
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mineralization of Fh. Accordingly, we explored the extent and the resulting secondary minerals
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of Fe(II)-induced transformation of (As(III)-adsorbed) OM-Fh adsorption and coprecipitation
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complexes, which were synthesized using two types of OM (DOM extracted from the O horizon
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of an Ultisol and polygalacturonic acid (PGA) as a proxy for polysaccharides. Regardless of OM
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type, increased contents of the coprecipitated or adsorbed OM led to a decrease in Fh
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transformation in both the absence and presence of As(III). Adsorbed As(III) caused a decrease
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in Fh conversion for both pure Fh and OM-Fh complexes. We observed only small differences in
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Fe(II)-induced transformation of OM-Fh coprecipitates vs. adsorptive complexes. However,
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without adsorbed As(III), OM types strongly influenced the secondary mineral products: DOM
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impeded goethite (Gt) and stimulated lepidocrocite (Lp) formation, while only Gt was formed for
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PGA. When As(III) and OM coexists, As (III) favored Lp over Gt formation for both DOM- and
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PGA-Fh complexes. These findings provide evidence that the mineral evolution of Fh largely
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depends on the potential additive/competing influence of coexisting constituents.
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Keywords: EXAFS; iron oxides; crystallization; dissolved organic matter; polysaccharides; arsenic
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Introduction Fe(III) (oxyhydr)oxides (hereafter termed Fe oxides) are ubiquitous in soils and
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sediments and can be strong sorbents for soil nutrients, contaminants and organic matter (OM).1-6
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Fe oxides are present in the environment as a wide range of minerals with different
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characteristics such as stability, specific surface area, and reactivity.1, 7-8 Fe oxides can undergo
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reductive dissolution in anoxic environments, releasing Fe(II)3, 9, 10. The adsorption of Fe(II) onto
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Fe(III) oxides induces electron transfer from Fe(II) to the host Fe oxides with accompanying
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recrystallization of the Fe oxides to thermodynamically more stable forms.11-16 This
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transformation process has wider implications than just the cycling of Fe alone, as the varying
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properties of Fe phases, such as surface area and crystallinity and their interaction with Fe(II),
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can affect the cycling of metal(loid)s and OM that are either adsorbed to or coprecipitated with
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Fe oxides17-19. Mineralogical changes are particularly dramatic when the initial Fe(III) oxide
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phase is ferrihydrite (Fh), a poorly-crystalline/amorphous Fe mineral, which readily undergoes
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Fe(II)-catalyzed crystallization to lepidocrocite (Lp), goethite (Gt) and magnetite11, 12-16.
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The rate and extent of Fe(II)-catalyzed Fh transformation, as well as the mineralogy and
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physicochemical properties of the products, may be affected by the presence of surface-
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associated or structurally incorporated elements. In nature, Fe oxides like Fh are often associated
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with OM20-22, via adsorption and/or coprecipitation23-26. OM was shown to hinder Fe(II)-induced
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Fh transformation by surface-site blockage and/or organic Fe(II) complexation.19, 27, 28 With
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respect to OM adsorption on pre-existing Fe oxides, OM coprecipitation with Fh results in
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smaller crystal sizes and greater structural disorder19, 28-30, and may subsequently affect the
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reactivity and stability of Fh. Chen et al. showed that Fh coprecipitated with OM extracted from
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forest litter samples, displayed a linear decrease in mineral transformation rates with increasing
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OM contents following reaction with Fe(II), and favored Lp formation over Gt and magnetite.19
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Recently, it was observed that, at similar OM loadings, coprecipitated Fh was more reactive
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towards microbial reduction than Fh with adsorbed OM.31-32 However, the impact of adsorbed vs.
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copecipitated OM on abiotic Fh transformation induced by Fe(II) has not been directly compared.
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In addition, OM composition in natural environments is very complex, comprising a collection
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of simple and macromolecular organic groups33-35. However, the impact of varying specific
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organic compounds on Fh transformation and the nature of the resulting products is rarely
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investigated36, 37.
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OM-rich soil and sediments tend to show high affinities for trace metal(loid)s like arsenic
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(As).38-41 As(V) has a strong affinity for both Al and Fe oxides, while As(III) adsorption is
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largely limited to Fe oxides.42-45 The As content of naturally occurring iron oxides shows great
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variation ranging from an As/Fe molar ratio of 2.4 × 10−6 to 0.146-49, with As speciation primarily
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As(III) in some cases49. Although low As(V) concentrations (As/Fe 1.1), a complete suppression of Fh transformation was
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observed following 7 days of reaction (Figure 1; SI Figure S6 and S7). The impaired Fh
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transformation by OM, which is consistent with previous studies19, 27, 28, 60, could be attributable
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to the decreased Fe(II) adsorption (SI Table S3), the surface blockage as indicated by surface
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area measurement (SI Table S1), and organic complexation of Fe(II). This could reduce the
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direct contact between Fe(II) and the mineral surface, and hence inhibit/slowdown mineral
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transformation19, 28. Previous studies have also demonstrated that OM can retard transformation
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of Fh to more crystalline minerals by blocking dissolution sites or hindering nucleation of more
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stable minerals even in the absence of Fe(II).61, 62
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Apart from the amount of Fh-associated OM, we demonstrated for the first time that the
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type of OM is a another key factor in determining the products of the Fe(II)-catalyzed Fh
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transformation in the absence of As(III). Although the extent of Fh preservation was greater for
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DOM than for PGA at a C/Fe ratio of ~0.85, a similar degree of Fh preservation between PGA
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and DOM was observed at all the other C/Fe ratios (SI Figure S11). The reaction products of
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Fe(II)-catalyzed transformation of Fh are primarily a function of Fe(II)/Fh ratio, pH and ligand
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type. The Fe(II) concentration (~0.5 mmol/g Fh) used in this study was below the threshold
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required for magnetite precipitation (~1.0 mmol Fe(II)/g Fh) at circumneutral pH 11. Consistent
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with previous studies11, 16, 19, FeSO4 (~0.5 mmol/g Fh) favored the conversion of pure Fh to Gt at
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pH 7 (buffered with PIPES). In contrast to only Gt formation for pure Fh, DOM resulted in more
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Lp formation with less Gt (Figure 1a; SI Figure S6), as observed previously19. However,
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interestingly, only Gt formation was observed for PGA (Figure 1b; SI Figure S7). These results
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indicate a strong role of OM composition in Fh transformation products. Lp is
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thermodynamically unstable with respect to Gt, and thus Lp may serve as a precursor to Gt
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formation.1 Transformation of Lp to Gt was noticed in our previous study of reaction of pure Fh
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with Fe(II)19. However, mineral-associated DOM stabilized Lp from further transformation19. In
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contrast, only Gt formation with PGA may suggest that unlike DOM, polysaccharides may be
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unable to hinder the transformation of Lp to Gt. However, future studies on temporal mineral
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evolution are needed to unravel if Lp is formed as an intermediate product, or if Fh is directly
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converted to Gt during reaction of PGA-Fh with Fe(II). The amount of Fe(II) removed from
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solution following 1 hour of reaction is nearly equivalent for DOM- and PGA-Fh complexes (SI
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Table S3). This implies that the difference in the transformation products between DOM and
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PGA is unlikely due to Fe(II) sorption, although it is currently unknown if the amount of electron
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transfer from adsorbed Fe(II) to bulk Fe(III) differed between PGA- and DOM-Fh complexes. It
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was suggested that the effectiveness in suppressing crystallization depends on how strongly the
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organic compounds sorb onto Fe oxides36, 37. Although the solid-phase C content in all samples
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before and after reaction with Fe(II) indicates no significant loss of solid-phase OM (SI Table
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S2), DOM appeared to bind more strongly with Fe oxides than PGA (SI Figure S10). Thus, the
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displacement of DOM on oxides by the Fe(II) ions may be more difficult than that for PGA.
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Therefore, DOM could more effectively stabilize meta-stable oxides like Lp from
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recrystallization than extracellular organic compounds from plant root exudates or microbial EPS.
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In addition, DOM contains aromatic and phenolic C in addition to carboxyl groups, while PGA
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only has carboxylic C (SI Figure S1). The aromatic and phenolic C may hinder the nucleation of
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more stable minerals like Gt. Considering the common associations of OM and oxides in nature6,
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, the influence of varying OM types on mineral evolution needs to be fully explored. The FTIR spectra of the adsorbed and coprecipitated OM were similar for both PGA and
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DOM (SI Figure S2). In addition, we observed no difference in the extent or products of Fh
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transformation between coprecipitates and adsorption complexes at the lowest and highest OM
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contents for both DOM and PGA (Figure 1 and SI Figure S11). With intermediate OM contents
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(C/Fe=0.52-0.88), the coprecipitated OM resulted in slightly more Fh transformation (5-10% of
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total Fe) to Gt than the adsorbed OM for both DOM and PGA. This might be due to the smaller
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particle size and lower crystallinity of the coprecipitated Fh, based on Mössbauer analysis from
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previous studies19, 29. Overall, the reactivity of OM-Fh adsorption and coprecipitation complexes
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did not display large differences in the reactivity towards Fe(II). Similarly, a previous study,
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which compared microbial Fe(III) reduction of DOM-Fh complexes, showed Fe(III) reduction
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rates was only slightly higher for DOM-Fh coprepitates (0.038-0.058 mmol h-1) compared to the
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adsorption complexes (0.02-0.05 mmol h-1).32 In addition, it was previously reported that the
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differences in the degradability of the adsorbed and coprecipitated DOM and lignin were small 63.
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Collectively, we may not expect dramatic differences in the stability, transformation and
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composition of both organic and mineral components of OM-Fh complexes formed via
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coprecipitation vs. adsorption in natural environments.
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Fe(II)-induced transformation of As(III)-bearing (OM-)Fh
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Following reaction with Fe(II), the oxidation state of the adsorbed As(III) remained
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unchanged based on As K-edge XANES analysis (SI Section 8). Similar to As(III)-free systems,
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the extent of Fh preservation also increased with increasing OM contents when As(III) was
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adsorbed. Adsorbed As(III) resulted in a decrease in Fh conversion and hence the preservation of
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Fh, compared to the corresponding As-free treatment for both pure- and OM-Fh (Figure 2 and SI
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Figure S11), as determined by EXAFS LCF. Previous studies have found that adsorbed As(V)
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can inhibit Fe(II)-catalyzed Fh transformation.18, 51 We showed here that this is true for As(III)
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with an As(III)/Fe ratio of 0.03, and the co-existence of As(III) and OM resulted in less Fh
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transformation than OM-free and As-bearing systems following short-term (≤ 7 days) exposure
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to Fe(II). Such inhibition of phase transformation has been reasoned to result from As covalently
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bonded to the surface of Fh64, which can reduce the extent of electron exchange between Fe(II)
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and Fe(III) required for Fe(II)-catalyzed transformation 51, 65. Arsenic is also believed to retard
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Fh transformation in a similar way as hydroxyl-carboxylic acids 66. Cornell and Schwertmann
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suggested that the anion linkage of two or more units of Fh forms a network of particles resistant
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to dissolution 37.
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Adsorbed As(III) also alters the secondary products of Fh transformation—for both pure-
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and OM-Fh (Figure 2). Unlike pure Fh, reaction of As(III)-bearing Fh with Fe(II) produced
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primarily Lp with much less Gt following 7 days of reaction (Figure 2; SI Figure S5b). The
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preferential formation of Lp during abiotic As(V)-bearing Fh transformation has also been noted
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previously 18, 51. A previous study showed that Lp could be slowly converted to Gt in the absence
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of other constituents at the longer reaction time scale (e.g. months), however no transformation
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of Lp was observed in the presence of DOM even following 3-month reaction with Fe(II).19
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Therefore similar to what was observed for DOM, As could inhibit or at least slow down the
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transformation of the relatively unstable Fe oxides like Lp to more stable Gt, possibly by
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blocking dissolution sites or by preventing polymerization of thermodynamically stable Fe(III)
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minerals 67, 68.
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An interesting and important finding is that adsorbed As(III) minimizes the impact of
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OM types on the secondary products of Fe(II)-catalyzed OM-Fh complex transformation with an
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As/Fe ratio of 0.03 (Figure 2). The co-existence of DOM and As(III) on Fh increased Lp
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formation by inhibiting Gt formation, relative to As-free DOM-Fh complexes (Figure 2; SI
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Figure S8). This favored Lp over Gt formation by adsorbed As(III), was much more dramatic
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for PGA-Fh complexes, compared to DOM-Fh complexes (Figure 2; SI Figure S9).
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Consequently, in the presence of adsorbed As(III), nearly identical secondary products were
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observed between DOM- and PGA-Fh complexes (Figure 2). While As(III) (As/Fe = 0.03) and
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OM (C/Fe = 0.3-0.6) co-occur in the systems with ~0.5 mmol FeSO4/g Fh and pH 7, Lp
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formation was much more pronounced relative to Gt, regardless of OM types. Therefore, with
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adsorbed As(III), the concentrations of Fh-associated OM and As(III) are critical in determining
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the mineral evolution. Owing to the significant co-occurrence of As and OM in contaminated
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sites, the observed less-crystalline (and hence more reactive) Fe oxides including Fh and Lp in
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the presence of As and OM, has to be considered when evaluating the reactivity of Fe minerals
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and the subsequent impact on the fate of OM and metal(oids) associated with these minerals in
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the environment.
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Environmental Implications In soils and sediments, where Fe oxides are impure, alteration of mineral transformation by impurities may influence the dynamics of OM, metals and nutrients associated with Fe oxides.
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Here, we note that the adsorbed or coprecipitated OM and the adsorbed As(III) decreased the
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extent of Fh transformation. These findings may provide an explanation for the preservation of
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Fh-like phases in natural environments 49, 69, 70. Transformation of Fh to presumably less-reactive
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Gt is expect to preferentially occur in environments lacking of DOM and As(III) (i.e.
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uncontaminated subsurface environments). In addition, the decreased Fh transformation, as well
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as the favored Lp over Gt formation in the presence of DOM and As(III), has important
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implications for predicting the fate of Fe minerals and their associated species such as As and
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OM. The persistence of Fh/Lp in OM-rich and As-contaminated fields may provide reactive
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surface area for OM, metals or nutrients, as well as electron transfer reactions. For example,
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rapid reductive dissolution of these presumably more bioavailable oxides such as Fh and Lp,
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may consequently drive OM 71, 72 and As mobilization 73, 74 to the aqueous phase under anoxic
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conditions.
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Supporting Information
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Additional details on Fe EXAFS analysis, XRD data, FTIR measurements, and As XANES
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analysis.
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Acknowledgements This research is a part of the Christina River Basin Critical Zone Observatory (CRB-CZO)
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project that was supported by the National Science Foundation (EAR 0724971). XAS analysis
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was carried out at the Stanford Synchrotron Radiation Lightsource, a Directorate of SLAC
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National Accelerator Laboratory and an Office of Science User Facility operated for the U.S.
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Department of Energy Office of Science by Stanford University. We are also grateful to the
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Shanghai Synchrotron Radiation Facility for use of the synchrotron radiation facilities at
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beamline 14W.
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19) Chen, C.; Kukkadapu, R.; Sparks, D. L. Influence of coprecipitated organic matter on Fe2+(aq) catalyzed transformation of ferrihydrite: implications for carbon dynamics. Environ. Sci. Technol. 2015, 49, 10927–10936. 20) Mcknight, D. M.; Bencala, K. E.; Zellweger, G. W.; Aiken, G. R.; Feder, G. L.; Thorn, K. A. Sorption of dissolved organic carbon by hydrous aluminum and iron oxides occurring at the confluence of Deer Creek with the Snake River, Summit County, Colorado. Environ. Sci. Technol. 1992, 26, 1388–1396. 21) Wagai, R. Mayer, L. M. Sorptive stabilization of organic matter in soils by hydrous iron oxides. Geochim. Cosmochim. Acta 2007, 71(1), 25–35. 22) Chen, C.; Dynes, J.J.; Wang, J.; Karunakaran, C.; Sparks, D.L. 2014. Soft X-ray Spectromicroscopy Study of Mineral-Organic Matter Associations in Pasture Soil Clay Fractions. Environ. Sci. Technol. 2014, 48(12), 6678–6686. 23) Riedel, T.; Zak, D.; Biester, H.; Dittmar, T. Iron traps terrestrially derived dissolved organic matter at redox interfaces. Proc. Natl. Acad. Sci. USA 2013, 110(25), 10101–10105. 24) Eusterhues, K.; Rennert, T.; Knicker, H.; Kögel-Knabner, I.; Totsche, K. U.; Schwertmann, U. Fractionation of Organic Matter Due to Reaction with Ferrihydrite: Coprecipitation versus Adsorption. Environ. Sci. Technol. 2011, 45, 527–533. 25) Chen, C.; Dynes J.; Wang, J.; Sparks, D.L. Properties of Fe-organic matter associations via coprecipitation versus adsorption. Environ. Sci. Technol. 2014, 48 (23), 13751–13759. 26) Mikutta, R.; Lorenz, D.;Guggenberger, G.; Haumaier, L.; Freund, A. Properties and reactivity of Fe-organic matter associations formed by coprecipitation versus adsorption: Clues from arsenate batch adsorption. Geochim. Cosmochim. Acta 2014, 144, 258–276. 27) Jones, A. M.; Collins, R. N.; Rose, J.; Waite, T. D. The effect of silica and natural organic matter on the Fe(II)-catalysed transformation and reactivity of Fe(III) minerals. Geochim. Cosmochim. Acta 2009, 73(15), 4409–4422. 28) ThomasArrigo, L. K.; Mikutta, C.; Byrne, J.; Kappler, A.; Kretzschmar, R. Iron(II)-Catalyzed Iron Atom Exchange and Mineralogical Changes in Iron-rich Organic Freshwater Flocs: An Iron Isotope Tracer Study. Environ. Sci. Technol. 2017, 51 (12), 6897–6907. 29) Mikutta, C.; Mikutta, R.; Bonneville, S.; Wagner, F.; Voegelin, A.; Christl, I.; Kretzschmar, R. Synthetic coprecipitates of exopolysaccharides and ferrihydrite. Part I: Characterization. Geochim. Cosmochim. Acta 2008, 72, 1111–1127. 30) Eusterhues, K.; Wagner, F. E.; Häusler, W.; Hanzlik, M.; Knicker, H.; Totsche, K. U.; KögelKnabner, I.; Schwertmann, U. Characterization of ferrihydrite-soil organic matter coprecipitates by X-ray diffraction and Mössbauer spectroscopy. Environ. Sci. Technol. 2008, 42, 7891–7897. 31) Eusterhues, K.; Hädrich, A.; Neidhardt, J.; Küsel, K.; Keller, T. F.; Jandt, K. D., Totsche K. U. Reduction of ferrihydrite with adsorbed and coprecipitated organic matter: microbial reduction by Geobacter bremensis vs. abiotic reduction by Na-dithionite. Biogeosci. Discuss. 2014, 11, 6039–6067. 32) Cooper, R. E.; Eusterhues, K.; Wegner, C. E.; Totsche, K. U.; Küsel, K. Ferrihydrite-associated organic matter (OM) stimulates reduction by Shewanella oneidensis MR-1 and a complex microbial consortia. Biogeosciences 2017, 14, 5171–5188. 33) Krull, E.S.; Baldock, J. A.; Skjemstad, J. O. Importance of mechanisms and processes of the stabilization of soil organic matter for modeling carbon turnover. Funct. Plant Biol. 2003, 30(2), 207–222. 34) Sollins, P.; Homman, P.; Caldwell, B. A. Stabilization and destabilization of soil organic matter: mechanisms and controls. Geoderma 1996, 74, 65–105. 35) Lehmann, J.; Kleber, M. The contentious nature of soil organic matter. Nature 2015, 528, 60– 68. 36) Cornell, R. M. Effect of simple sugars on the alkaline transformation of ferrihydrite into goethite and hematite. Clays Clay Miner. 1985, 33(3), 219–227.
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37) Cornell, R. M.; Schwertmann, U. Influence of organic anions on the crystallization of ferrihydrite. Clays Clay Miner. 1979, 27(6), 402–410. 38) Elliott, A. V. C.; Plach, J. M.; Droppo, I. G.; Warren, L. A. Comparative floc-bed sediment trace element partitioning across variably contaminated aquatic ecosystems. Environ. Sci. Technol. 2012, 46, 209–216. 39) ThomasArrigo, L. K.; Mikutta, C.; Byrne, J.; Barmettler, K.; Kappler, A.; Kretzschmar, R. Iron and arsenic speciation and distribution in organic flocs from streambeds of an arsenic-enriched peatland. Environ. Sci. Technol. 2014, 48, 13218–13228. 40) ThomasArrigo, L. K.; Mikutta, C.; Lohmayer, R.; Planer-Friedrich, B.; Kretzschmar, R. Sulfidization of organic freshwater flocs from a minerotrophic peatland: Speciation changes of iron, sulfur, and arsenic. Environ. Sci. Technol. 2016, 50, 3607–3616. 41) Shimizu, M.; Arai, Y.; Sparks, D. L. Multiscale Assessment of Methylarsenic Reactivity in Soil. 2. Distribution and Speciation in Soil. Environ. Sci. Technol. 2011, 45, 4300–4306. 42) Manning, B. A.; Goldberg, S. Arsenic(III) and arsenic(V) adsorption on three California soils. Soil Sci. 1997, 162, 886–895. 43) Arai, Y.; Elzinga, E. J.; Sparks, D. L. (2001) X-ray absorption spettroscopy investigation of arsenite and arsenate adsorption on the aluminium oxide-water interface. J Colloid Interf. Sci. 2001, 235, 80–88. 44) Dixit, S. I. Hering, J. G. Comparison of arsenic(V) and arsenic(III) sorption onto iron oxide minerals: implications for arsenic mobility. Environ Sci. Technol. 2003, 37(18), 4182–4189. 45) Masue, Y.; Loeppert, R. H.; Kramer, T. A. Arsenate and arsenite adsorption and desorption behavior on coprecipitated aluminum: iron hydroxides. Environ. Sci. Technol. 2007, 41 (3), 837–842. 46) Pedersen, H. D.; Postma, D.; Jakobsen, R. Release of arsenic associated with the reduction and transformation of iron oxides. Geochim. Cosmochim. Acta 2006, 70, 4116–4129. 47) Bowell, R. J. Sorption of arsenic by iron oxides and oxyhydroxides in soils. Appl. Geochem. 1994, 9, 279–286. 48) Pichler, T.; Veizer, J.; Hall, G. E. M. Natural input of arsenic into a coral-reef ecosystem by hydrothermal fluids and its removal by Fe(III) oxyhydroxides. Environ. Sci. Technol. 1999, 33, 1373–1378. 49) LeMonte, J. J.; Stuckey, J. W.; Sanchez, J. Z. Tappero, R. V.; Rinklebe, J.; Sparks, D. L. Sea level rise induced arsenic release from historically contaminated coastal soils. Environ. Sci. Technol. 2017, 51(11), 5913–5922. 50) Ford, R. G. Rates of hydrous ferric oxide crystallization and the influence on coprecipitated arsenate. Environ. Sci. Technol. 2002, 36, 2459–2463. 51) Gomez, M. A.; Hendry, M. J.; Hossain, A.; Das, S.; Elouatik, S. Abiotic reduction of 2-line ferrihydrite: effects on adsorbed arsenate, molybdate, and nickel. RSC Adv. 2013, 3, 25812– 25822. 52) Kocar, B. D.; Fendorf, S. Thermodynamic constraints on reductive reactions influencing the biogeochemistry of arsenic in soils and sediments. Environ. Sci. Technol. 2009, 43 (13), 4871– 4877. 53) Stuckey, J. W.; Schaefer, M. V.; Benner, S. G.; Fendorf, S. Reactivity and speciation of mineral-associated arsenic in seasonal and permanent wetlands of the Mekong Delta. Geochim. Cosmochim. Acta 2015, 171, 143–155. 54) Knee, E. M.; Gong, F. C.; Gao, M.; Teplitski, M.; Jones, A. R.; Foxworthy, A.; Mort, A. J.; Bauer, W. D. Root mucilage from pea and its utilization by rhizosphere bacteria as a sole carbon source. Mol. Plant Microbe Interact. 2001, 14, 775–784. 55) Cheshire, M. V. Nature and Origin of Carbohydrates in Soils Academic Press, London, 1979. 56) Brunauer, S.; Emmett, P. H.; Teller, E. Adsorption of gases in multimolecular layers. J. Am. Chem. Soc. 1938, 60, 309–319.
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57) Stookey, L. L. Ferrozine-A new spectrophotometric reagent for iron. Anal. Chem. 1970, 42(7), 779–781. 58) Webb, S. M. SIXPack a graphical user interface for XAS analysis using IFEFFIT. Phys. Scr. 2005, T115, 1011–1014. 59) Hansel, C. M.; Benner, S. G.; Neiss, J.; Dohnalkova, A.; Kukkadapu, R. K.; Fendorf S. Secondary mineralization pathways induced by dissimilatory iron reduction of ferrihydrite under advective flow. Geochim. Cosmochim. Acta 2003, 67, 2977–2992. 60) Henneberry, Y. K.; Kraus, T. E. C.; Nico, P. S.; Horwath, W. R. Structural stability of coprecipitated natural organic matter and ferric iron under reducing conditions. Org. Geochem. 2012, 48, 81–89. 61) Cornell, R. M.; Schwertmann, U. Influence of organic anions on the crystallization of ferrihydrite. Clays Clay Miner.1979, 27, 402−410. 62) Xiao, W.; Jones, A. M.; Li, X.; Collins, R. N.; Waite, T. D. Effect of Shewanella oneidensis on the kinetics of Fe(II)-catalyzed transformation of ferrihydrite to crystalline iron oxides. Environ. Sci. Technol. 2018, 52 (1), 114–123. 63) Eusterhues, K.; Neidhardt, J.; Hädrich, A.; Küsel, K; Totsche K. U. Biodegradation of ferrihydrite-associated organic matter. Biogeochem. 2014, 119, 45–50. 64) Waychunas, A.; Rea, B.A.; Fuller, C. C.; Davis, J. A. Surface chemistry of ferrihydrite: 1. EXAFS studies of the geometry of coprecipitated and adsorbed arsenate. Geochim. Cosmochim. Acta 1993, 57, 2251–2269. 65) Katz, J. E.; Zhang, X.; Attenkofer, K.; Chapman, K. W.; Frandsen, C.; Zarzycki, P.; Rosso, K. M.; Falcone, R. W.; Waychunas, G. A.; Gilbert, B. Electron small polarons and their mobility in iron (oxyhydr)oxide nanoparticles. Science 2012, 337, 1200–1203. 66) Sun, T.; Paige, C. R.; Snodgrass, W. J. Combined effect of arsenic and cadmium on the transformation of ferrihydrite onto crystalline products. J. Univ. Sci. Technol. B 1999, 3, 168– 173 67) Schwertmann, U.; Cornell, R. M. Iron Oxides in the Laboratory – Preparation and Characterization. VCH Verlagsgesellschaft mbH, Weinheim, Germany 1991. 68) Schwertmann, U.; Taylor, R. M. Natural and synthetic poorly crystallised lepidocrocite. Clay Miner. 1979, 14, 85–293. 69) Schwertmann, U.; Murad, E. The nature of an iron oxide: Organic iron association in a peaty environment. Clay Miner. 1988, 23 (3) 291–299 70) Chen, C.; Kukkadapu, R. K.; Lazareva, O.; Sparks, D. L. Solid-phase Fe speciation along the vertical redox gradients in floodplains using XAS and Mössbauer spectroscopies. Environ. Sci. Technol. 2017, 51 (14), 7903–7912. 71) Adhikari,D.; Zhao, Q.; Das, K.; Mejia, J.; Huang, R.; Wang, X.; Poulson, S. R.; Tang, Y.; Roden, E. E.; Yang, Y. Dynamics of ferrihydrite-bound organic carbon during microbial reduction. Geochim. Cosmochim. Acta 2017, 212, 221–223. 72) Pan, W.; Kan, J.; Inamdar, S.; Chen, C.; Sparks, D. L. Dissimilatory microbial iron reduction release DOC (dissolved organic carbon) from carbon-ferrihydrite association. Soil Biol. Biochem. 2016, 103, 232–240. 73) Horneman, A.; van Geen, A.; Kent, D.; Mathe, P. E.; Zheng, Y.; Dhar, R. K.; O’Connell, S.; Hoque, M.; Aziz, Z.; Shamsudduha, M.; Seddique, A.; Ahmed, K. M. Decoupling of As and Fe release to Bangladesh groundwater under reducing conditions. Part I: Evidence from sediment profiles. Geochim. Cosmochim. Acta 2004, 68, 3459–3473. 74) Tufano, K. J.; Fendorf, S. Confounding impacts of iron reduction on arsenic retention. Environ. Sci. Technol. 2008, 42(13), 4777–4783.
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Figure 1 Secondary minerals formed following 7 days of reaction of 1 mM Fe(II) with
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ferrihydrite (C/Fe = 0) as well as (a) DOM- and (b) PGA-ferrihydrite adsorption vs.
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coprecipitation complexes, as a function of C/Fe molar ratios. Mineral percentages were obtained
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via linear combination fitting of k3-weighted Fe EXAFS spectra with reference minerals. The k3-
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weighted EXAFS spectra and linear combination fits are shown in Supporting Information (SI
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Figure S13 and S14).
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Figure 2 Secondary minerals formed following 7 days of reaction of 1 mM Fe(II) with As(III)-
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bearing ferrihydrite, as well as As(III)-bearing (a) DOM- and (b) PGA-ferrihydrite adsorption vs.
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coprecipitation complexes, as a function of C/Fe molar ratios. Mineral percentages were obtained
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via linear combination fitting of k3-weighted Fe EXAFS spectra with reference minerals. The k3-
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weighted EXAFS spectra and linear combination fits are shown in Supporting Information (SI
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Figure S15 and S16).
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