Chemical Reactivity Probes for Assessing Abiotic Natural Attenuation

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Chemical Reactivity Probes for Assessing Abiotic Natural Attenuation by Reducing Iron Minerals Dimin Fan, Miranda J. Bradley, Adrian W. Hinkle, Richard L. Johnson, and Paul G. Tratnyek Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.5b05800 • Publication Date (Web): 26 Jan 2016 Downloaded from http://pubs.acs.org on January 27, 2016

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Environmental Science & Technology

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Chemical Reactivity Probes for Assessing Abiotic Natural

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Attenuation by Reducing Iron Minerals

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Dimin Fan1, Miranda J. Bradley, Adrian W. Hinkle,

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Richard L. Johnson, and Paul G. Tratnyek*

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Institute of Environmental Health

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Oregon Health & Science University

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3181 SW Sam Jackson Park Road, Portland, OR 97239

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Current affiliation: U.S. Environmental Protection Agency, Office of Superfund Remediation and Technology Innovation, Arlington, VA 22202

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*Corresponding author:

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Email: [email protected], Phone: 503-346-3431, Fax: 503-346-3427

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Keywords: Chemical reactivity probes, indicators, mediators, shuttles; Oxidation-

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reduction potential; Iron oxides; Dechlorination, chlorinated solvents; In situ

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chemical reduction, ISCR

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Revised Version for Environmental Science & Technology

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24 January 2016

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ACS Paragon Plus Environment


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Abstract

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Increasing recognition that abiotic natural attenuation (NA) of chlorinated solvents can be

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important has created demand for improved methods to characterize the redox properties of the

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aquifer materials that are responsible for abiotic NA. This study explores one promising

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approach: using chemical reactivity probes (CRPs) to characterize the thermodynamic and

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kinetic aspects of contaminant reduction by reducing iron minerals. Assays of thermodynamic

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CRPs were developed to determine the reduction potentials (ECRP) of suspended minerals by

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spectrophotometric determination of equilibrium CRP speciation and calculations using the

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Nernst equation. ECRP varied as expected with mineral type, mineral loading, and Fe(II)

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concentration. Comparison of ECRP with reduction potentials measured potentiometrically using a

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Pt electrode (EPt) showed that ECRP was 100–150 mV more negative than EPt. When EPt was

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measured with small additions of CRPs, the systematic difference between EPt and ECRP was

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eliminated, suggesting that these CRPs are effective mediators of electron transfer between

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mineral and electrode surfaces. Model

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contaminants (4-chloronitrobenzene, 2-

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chloroacetophenone, and carbon tetrachloride)

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were used as kinetic CRPs. The reduction rate

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constants of kinetic CRPs correlated well with

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the ECRP for mineral suspensions. Using the

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rate constants compiled from literature for

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contaminants and relative mineral reduction

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potentials based on ECRP measurements,

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qualitatively consistent trends were obtained,

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suggesting that CRP-based assays may be

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useful for estimating abiotic NA rates of

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contaminants in groundwater.

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Introduction

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Abiotic reduction of contaminants by naturally-occurring reducing iron minerals (including iron

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oxides and sulfides) is an important component of natural attenuation (NA).1-4 The abiotic

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reduction of chlorinated ethenes (e.g., trichloroethene, TCE) is of particular interest, and has

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been described in many laboratory studies using reducing iron minerals prepared by chemical

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synthesis,5-8 or in-situ precipitation,9-11 and aquifer materials obtained from contaminated field

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sites.12-14 However, the rates of these reactions are usually slow, making it impractical to assess

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in-situ rates of abiotic NA. In contrast, much faster reaction rates with reducing minerals can be obtained with

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(model) contaminants, such as nitroaromatic compounds (NACs) or polyhalogenated alkanes

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(PHAs).15-19 In principle, easily reducible model contaminants could serve as chemical probes to

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characterize the reactivity of reducing aquifer materials with more recalcitrant contaminants if

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the two groups of compounds share similar reactivity patterns.20, 21 Using hexachloroethane

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(HCA) and 4-chloronitrobenzene (4Cl-NB) as probe compounds, Elsner et al.20 showed

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qualitatively similar trends in the reduction rates of these two compounds by a range of ferrous-

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iron containing minerals, even though the two probe compounds are reduced by different

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mechanisms and the absolute rates differ by almost 3 to 4 orders of magnitude. Rates of HCA

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reduction have further been shown (qualitatively) to parallel literature data for TCE reduction by

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magnetite6, green rust5, and iron sulfides6, suggesting that HCA and other surrogate chemical

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probes might help to predict relative rates of chloroethene reduction in abiotic NA. The fundamental validity and practical reliability of using model contaminants as probes

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to predict rates of abiotic NA requires identifying mineral properties that determine the rates of

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probe/contaminant reduction. This is, however, challenging because of the profound and

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complex effects on mineral redox chemistry that arise from the diversity of reducing iron

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minerals, including surface Fe(II), structural Fe(II), and FeS. For Fe(II) sorbed onto iron oxides,

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a number of studies have suggested that the mineral reactivity correlates with the density of the

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sorbed Fe(II).16, 20 However, the interpretation of Fe(II)ads data estimated with traditional

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extraction methods is complicated by the finding of electron transfer from adsorbed Fe(II) into

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the bulk materials and that aqueous Fe(II) appears to be necessary for maintaining mineral

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reactivity.22, 23 For mixed valent iron oxides, such as magnetite or green rust, the ratio of Fe(II) to

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Fe(III) has also been shown to affect the reduction kinetics of contaminants.9, 18 In homogeneous systems with aqueous Fe(II)-ligand complexes as the reductants, the one

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electron reduction potential of the reducing species has been shown to give a linear free energy

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relationship (LFER) with the reduction rate constants for oxyamyl, a carbamate pesticide.24

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LFERs structured in this way—one contaminant vs. many environmental reactants—are

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uncommon, compared with LFERs for many contaminants vs. one environmental reactant (e.g.,

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NACs vs. Fe(II)porpyrin25), but the two approaches are complementary and combining them has

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potential value for diagnostic and predictive purposes. Although it is more difficult to obtain the

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analogous LFERs for heterogeneous reductants, such as Fe(II) minerals, the mineral redox

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potential is still the most likely descriptor in such relationships because it dictates the

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thermodynamic driving force of these reactions with contaminants.26 The major impediment to development of LFERs for reactions with reducing minerals is

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the difficulty in obtaining reliable mineral redox potentials under environmental conditions.

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Conventional potentiometric measurements of oxidation-reduction potential (ORP) by electrodes

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are easilly performed on mineral suspensions, but these are mixed potentials and are highly

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sensitive to factors such as mass transport, aggregation, and secondary redox-active species.27-29

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The “formal” reduction potentials for iron oxide/Fe(II) couples that appear in some literature30

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are based on calculations using thermodynamic data from calorimetry, and therefore may not be

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representative of the mineral phases present in heterogenous environmental media (aquifers,

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sediments, and soils). Mineral reduction potentials of Fe(II) sorbed to iron oxides have also been

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estimated with LFERs developed for aqueous Fe(II)-ligand complexes, but the calculated

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potentials vary substantially with the contaminants used in the training data set for LFER

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development.31 Recent efforts to measure the reduction potentials of iron minerals have employed redox-

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mediator or “shuttle” compounds to provide an intermediate redox couple that rapidly and

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reversibly equilibrates between redox active minerals and the surface of a working electrode.32

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Alternatively, if the speciation of the mediator can be determined spectrophotometrically

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(provided negligible sorption of the mediator to mineral surfaces), these data can be used in the

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Nernst equation to calculate redox potentials of the medium.33, 34 Various redox indicators have

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been used to characterize the redox processes of biogeochemical systems, including iron

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oxides,35-37 soil suspensions,38 and anaerobic sediments.39, 40 However, the power of this

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approach for quantification of specific mineral redox properties (potentials, kinetics, and

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capacities) has not been systematically developed. In this study, spectrophotometric measurement of redox indicator speciation was used to

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characterize the redox properties of a range of reducing iron minerals containing Fe(II). The

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calculated mineral redox potentials were compared with those obtained by potentiometry, and

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used as a descriptor variable in correlation analysis with kinetic data for the reduction of (model)

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contaminants by the minerals. Based on their functions, the redox indicators and model

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contaminants are referred to as the thermodynamic and kinetic chemical reactivity probes

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(CRPs), respectively. The broader implications of these results for contaminant fate and

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remediation are explored by extending the correlation analysis to include literature data for a

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wider range of contaminants, reducing iron minerals, and conditions. The results provide a semi-

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quantitative framework for understanding the relative reactivity of important contaminants with

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common minerals.

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Experimental

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Materials and Common Methods. All chemical reagents were ACS reagent grade and used as

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received without further purification. Resorufin (RSF) was acquired from Sigma Aldrich as

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sodium salt with 98% purity. Indigo-5,5'-disulfonate (I2S) and 9,10-anthraquinone-2-sulfonate

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(AQS) were purchased from TCI America (Portland, OR), as disodium salts with greater than

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95% purity. 9,10-anthraquinone-2,6-disulfonate (AQDS) was purchased as disodium salt with

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98% purity (Combi-Blocks, San Diego, CA). Anaerobic conditions were maintained for all

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operations involving Fe(II) and CRPs inside an anaerobic chamber (100% N2, O2 < 1 ppm),

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unless otherwise mentioned. Preparation of Mineral Suspensions. Goethite (GT), hematite (HT), and lepidocrocite

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(LP) were acquired from Bayferrox (Burgettstown, PA). These materials have been used in many

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prior studies of redox reactions involving iron oxide.20, 29, 41-44 A detailed analysis of the specific

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surface area of these samples is provided in the Supporting Information (SI, Table S1).

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Magnetite (MT) and green rust chloride (GR-Cl) were synthesized by co-precipitation methods

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adapted or modified from prior studies.18, 45 To prepare Fe(II) sorbed onto the three commercial 1/24/16


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iron oxides and synthetic magnetite, various quantities of Fe(II) stock solution (~100 mM) were

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added to the anoxic mineral suspensions with pH adjusted to 7.2 (or as noted) by 10 mM

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HEPES. The suspensions were transferred to amber bottles, which were then crimp sealed and

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placed on an end-over-end rotator for equilibration (~20 h). Additional details on the synthesis

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and preparation of suspensions of the iron oxides are described in the SI.

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CRP Assay for Mineral Reduction Potential. The thermodynamic CRPs were selected

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to (i) encompass the range of relevant redox potentials and (ii) be compatible with subsurface

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solid matrix (so that they can eventually be applied to characterize natural porous media). The

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latter criterion excluded a group of common redox indicators that are positively charged, because

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they tend to sorb strongly to negatively charged aquifer and sediment solids.40 The four

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thermodynamic CRPs selected are RSF, I2S, AQDS, and AQS. All CRPs except RSF are

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negatively charged due to one or more sulfonate groups, while RSF is neutral but relatively

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hydrophilic (e.g., log P = 1.77 predicted by chemicalize.org). The standard reduction potentials

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of these CRPs at pH 7 range from −46 mV to −229 mV (Table S2). Other properties of the

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CRPs, such as maximum absorption wavelength, molar absorptivity, and pKa are also given in

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Table S2. The reactions of the thermodynamic CRPs with the mineral suspensions were evaluated

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vs. mineral loading and dose of added Fe(II). For each combination of CRP and mineral, the

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reaction was carried out in a 10 mL amber serum vial to avoid photo reactions involving the CRP

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or aqueous Fe(II). Aliquots of deoxygenated CRP stock solutions (~1 mM) were added to 5 mL

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mineral suspensions, giving an initial concentration of ~20 µM for RSF and I2S and ~90 µM for

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AQDS and AQS. The reaction vial was crimp sealed with a 20 mm Viton septum (Sigma

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Aldrich) and placed on an end-over-end rotator overnight to reach equilibrium (45 rpm). Sample

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aliquots of 5 mL were then filtered by 0.2 µm nylon filters (Fisher). The first mL of filtrate was

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discarded and the rest was collected in a glass (for RSF and I2S) or quartz (for AQS and AQDS)

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cuvette (1 cm path length), which was then sealed by screw cap with a 4 mm Viton lining. The

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sealed cuvette was removed from the anaerobic chamber and the absorbance spectra of the

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reduced CRPs were immediately measured with a UV-vis spectrophotometer. The reduced CRP

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was then reoxidized by briefly exposing to ambient air, which served to determine how much of

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the CRP was adsorbed to the solids.

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Potentiometric Measurements. Open circuit potential (OCP) measurements were performed

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following the protocol developed in our previous study with suspensions of nano zerovalent iron

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(nZVI) 27. A two electrode system was used with a 3 mm Pt rotating disc electrode (RDE) as the

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working electrode and an Ag/AgCl electrode (3 M KCl filling) as the reference electrode. Before

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each set of experiments, the Pt electrode was polished with 0.05 µm alumina and sonicated to

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avoid electrode fouling or memory effects, which are common causes for inconsistent results

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with commercial ORP electrodes. The experiments were carried out in a 20-mL electrochemical

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cell that was purged with Ar gas for at least 10 minutes before the mineral suspension was

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introduced. Aliquots of the suspension were withdrawn with a 10-mL syringe, taken out of the

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anaerobic chamber, and immediately injected into the electrochemical cell. The Pt electrode was

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immersed into the suspension and rotated at 2000 rpm to maintain mixing. The OCP data was

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recorded using an Autolab PGSTAT30 (EcoChemie, Utrecht, The Netherlands) for 30 minutes,

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during which the headspace was continuously purged with Ar to maintain anoxic conditions. In

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selected experiments, the effects of CRPs on OCP were assessed by adding aliquots of

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appropriate CRPs into the cell during OCP measurement.

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Reduction of Kinetic CRPs. Three kinetic CRPs were selected for batch experiments: carbon

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tetrachloride (CT), 4-chloronitrobenzene (4Cl-NB), and 2-chloroacetophenone (2-CAP). These

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compounds represent three classes of model organic contaminants that are relatively labile to

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abiotic reduction: perchlorinated alkanes,21, 46 nitro aromatics,25, 47 and α–haloketones.48, 49 For

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CT, an aliquot of stock solution was added into a ~19 mL mineral suspension in a 20 mL serum

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vial to an initial concentration of ~10 µM with minimum headspace. The vial was immediately

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crimp sealed with a Viton septum, transferred out of the anaerobic chamber, and placed on an

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end-over-end rotator at 125 rpm in the dark at 23ºC. At each sampling point, 0.25 mL mineral

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suspension was withdrawn after injecting 0.3 mL N2 into the vial using a gas-tight glass syringe.

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The mineral suspension was injected into a headspace vial containing 2 mL deionized water.

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After brief vortex mixing, the CT concentration in headspace was measured by gas

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chromatography with an electron capture detector. For 4Cl-NB and 2-CAP, the reactions were carried out in 60 mL amber serum bottles,

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crimp sealed with Viton septa, containing an initial concentration of ~10 µM and ~50 µM,

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respectively. The reactors were mixed on the end-over-end rotator at 125 rpm at 23ºC. At each

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sampling time point, 2.5 mL aliquot was withdrawn and filtered by a 0.20 µm filter. For the 1/24/16


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combination of goethite and 4Cl-NB, the reactions were carried out in 10 mL syringes inside the

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anaerobic chamber due to rapid reaction. The samples were analyzed for disappearance of parent

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compounds by a high performance liquid chromatography (HPLC) with a photodiode detector.

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Results and Discussion

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Reduction of Thermodynamic CRPs by Fe(II) containing minerals. Preliminary experiments

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showed that 1 mM aqueous Fe(II) reduced RSF completely and I2S partially, but did not reduce

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AQDS or AQS (data not shown). This result is consistent with aqueous Fe(II) being a weak

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reductant that only reduces the CRPs with relatively high reduction potentials (Table S2). The

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same trend in CRP reduction was observed with Fe(II) in suspensions of the minerals. Using

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lepidocrocite and 0.5 mM Fe(II) as an example, Figure S2 shows complete reduction of RSF,

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significant reduction of I2S, little reduction of AQDS, and no reduction of AQS. The extent of

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CRP reduction was used to select appropriate CRPs for calculating mineral redox potentials

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(described later). The reduction of the CRPs varied with mineral type and loading and Fe(II) concentration

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(Figure 1). The absorbance spectra in Figure 1A show that the degree of AQDS reduction after

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20 hr of exposure to suspensions of ~50 m2/L iron oxide and 1.0 mM Fe(II) ranged from no

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reduction for hematite, to partial reduction for lepidocrocite and magnetite, and to complete

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reduction for goethite. Lowering the goethite loading from 2.2 g/L to 0.3 g/L (in Figure 1B)—

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with constant Fe(II) dose—decreased the extent of AQDS reduction about 10-fold. Figure 1C

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and 1D show increasing reduction of AQS with increasing Fe(II) concentration for magnetite and

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goethite, respectively. The observed effects of these variables on the redox speciation of CRPs

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has not been reported previously (in the limited number of prior studies that used this approach),

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but it resembles the effects on contaminant reduction by sorbed Fe(II) that have been identified

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in many prior studies. These studies have shown that reduction rates of various contaminants

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vary with mineral type, loading, and Fe(II) concentrations.16, 20, 31, 50 The similarity between this

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work and previous work formed the basis for our attempts to correlate the results obtained with

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thermodynamic and kinetic CRPs, which is discussed later.

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Figure 1. The effects of (A) oxide type, (B) goethite loading, (C) Fe(II) loading with magnetite, and (D) Fe(II) loading with goethite on reduction of CRPs by Fe(II) + iron oxide suspensions. The CRP was AQDS for (A) and (B) and AQS for (C) and (D). All spectra were taken after 20 hr of exposure to the mineral suspensions.

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Reduction Potentials of Minerals Determined with Thermodynamic CRPs. For each mineral

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suspension, CRPs were selected that did not give too much (e.g., Figure S2A) or too little (e.g.,

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Figure S2D) reduction (defined as 5%, respectively). The matches between CRP

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and reductant that we deemed suitable are identified in Table S3. Using the selected CRP, the

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amount of CRP reduction was quantified using the difference in absorbance at λmax of the

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oxidized form (Table S2) before and after reoxidation, a protocol adapted from Orsetti et al.33 In

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most cases, adsorption of the CRP was found to be negligible (because the absorbance after

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reoxidation was similar to the control). Interference from Fe(III) (due to oxidation of aqueous

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Fe(II)) was also minimal, because reoxidation of the reduced CRPs is much faster (complete in a

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few seconds) than Fe(II) oxidation at neutral pH (t1/2 = ~3 min). Assuming that equilibrium was

230 231 232

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reached between the CRP and the mineral suspension, the mineral reduction potentials were

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calculated from the CRP data (ECRP), using the Nernst equation for a two-electron redox couple

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with two relevant pKa’s for the reduced CRP (eq 1):

ECRP = E 0 +

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2 RT RT ⎛ [CRPox ] ⎞ ln ⎡⎣H+ ⎤⎦ + K r1 ⎡⎣H+ ⎤⎦ + K r1K r2 + ln ⎜ ⎟ 2F 2F ⎜⎝ [CRPred ] ⎟⎠

)

(

(1)

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where E0 is the standard reduction potential of the CRP (Table S2), R is the gas constant, T is

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temperature, F is the Faraday constant, and [CRPox] and [CRPred] are the concentrations of

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oxidized and reduced forms of the CRP. Additional pH effects due to other pKa’s are minimal

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near neutral pH 33, 51 and therefore were neglected. For the conditions and assumptions that apply

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to this work, eq 1 can be further simplified to eq 2: 0 ECRP = EpH +

253

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RT ⎛ [CRPox ] ⎞ ln ⎜ ⎟ 2F ⎜⎝ [CRPred ] ⎟⎠

(2)

0 where EpH is the formal reduction potential of the CRP at pH 7.2 used in the assay.

The mineral reduction potentials (ECRP) calculated using eq 2 are given in Table S3 and

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summarized in Figure 2. In Figure 2, each type of mineral tested is represented as a category

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(column), which contains several sub columns. Each sub column includes a series of ECRP values

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(horizontal markers) determined as a function of the variable (e.g., Fe(II) dose, mineral loading)

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that is indicated by the bold label above the corresponding sub column. At the lowest Fe(II)

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concentration tested (0.2 mM), reduction potentials were not obtained for hematite,

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lepidocrocite, and goethite because no CRP reduction occurred, suggesting that the reduction

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potentials of these oxides under these conditions are out of the range of potentials that are

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accessible by the CRPs used in this study. With increasing Fe(II), the calculated reduction

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potentials became more negative, as expected, for all oxides (Figure 2), but goethite gave a

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much wider range of potentials than hematite and lepidocrocite. Magnetite (0.75 g/L) without

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Fe(II) gave a negative potential around −180 mV because of structural Fe(II), but only dropped

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~30 mV as aqueous Fe(II) increased from 0.2 to 1.0 mM. The different sensitivities of ECRP to

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the Fe(II) concentration for different minerals is likely attributed to the intrinsic difference of

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mineral properties (e.g., bulk crystal conductivity52).

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Other notable trends shown in Figure 2 include (i) increasing ECRP with increasing

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addition of natural organic matter (NOM) for goethite, which is consistent with the consensus

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that NOM decreases the reductant reactivity,53, 54 (ii) decreasing ECRP with increasing mineral

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loading for MT (with no added Fe(II)), which simply shows the correlation between ECRP and the

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amount of structural Fe(II), and (iii) a bimodal pattern of ECRP with increasing mineral loading at

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1 mM Fe(II) for GT and LP, which is also shown in Figure S3. The latter result—which is

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discussed in more detail in SI—may prove to be analogous to the effect reported in several recent

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studies of Fe(II) adsorption onto iron oxides, wherein removal of excess aqueous Fe(II) was

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found to significantly lower rates of contaminant reduction by Fe(III) oxides with adsorbed

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Fe(II).16, 22

280 281 282 283 284 285 286 287

Figure 2. Summary of the ECRP values determined for different reducing iron minerals under various conditions. Each marker is labeled with the specific value of the varied parameter (bold labels). The plain text annotation above each subcolumn indicates the experimental parameter that was constant in that series of experiments. For reference, reduction potentials of the four CRPs at pH = 7 (EHʹ) are plotted in the far left column for reference. For comparison, formal reduction potentials reported in prior studies are plotted in the far right column. Tabulated values are from Stumm55 and measured values are based on Orsetti et al.33 and Gorski et al.34.

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The calculated mineral reduction potentials were further compared to previously

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measured or tabulated values for Fe(II) containing minerals, which are summarized in the far

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right column of Figure 2. The measured potentials for Fe(II) sorbed onto goethite reported by

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Orsetti et al.33 and for Fe(II) sorbed to synthetic hematite and magnetite reported by Gorski34 are

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of particular relevance because (i) both studies used thermodynamic CRPs (i.e., redox indicators)

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to calculate the mineral reduction potentials and (ii) the origins of minerals used in those studies

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are the same as some of the minerals used here. In general, the ECRP values agree well with the literature data. The reduction potential for

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10 g/L goethite suspension with 1.4 mM Fe(II) determined by AQDS at pH 6.7 ranged from

297

−180 to −160 mV,33 which agrees well with −181 mV, measured in this study for 10 g/L goethite

298

with 1 mM Fe(II) at pH 7.2 (Figure 2). The measured reduction potential of stoichiometric

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magnetite at pH 7.434 also falls into the range of magnetite reported here. In contrast, the

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synthetic hematite used by Gorski showed a much lower reduction potential than the Bayferrox

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hematite used in this study.34 The reason for this discrepancy is unclear, but it is consistent with

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other work showing that synthetic hematite is generally more reactive with contaminants than the

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Bayferrox hematite.31

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Comparison with Electrochemically Measured Potentials. EPt (i.e., OCP referenced to

305

the standard hydrogen electrode (SHE)) was measured on suspensions of hematite, lepidocrocite,

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goethite, and magnetite with a range of initial Fe(II) concentrations. In selected experiments,

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CRPs were added during the EPt measurement to serve as electron transfer mediators to improve

308

the electrode response to the redox active minerals.32 Sample chronopotentiograms (CPs) are

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shown in Figure 3A for goethite and hematite suspensions with 1 mM aqueous Fe(II) (after 20 h

310

pre-equilibrium). Initially, the EPt for the goethite suspension was only ~20 mV below the EPt for

311

the hematite suspension (Figure 3A), which is much less than their difference in ECRP shown in

312

Figure 2. However, adding small concentrations (~12 µM) of a CRP during the CP measurement

313 314

resulted in a sharp drop in EPt. The decrease of EPt in the presence of CRP is evidence that the

315

CRPs served as mediators of electron transfer between the mineral surfaces and the electrode

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surface, thereby improving electrode response to minerals. However, this effect was only seen

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with AQS on the goethite suspension and I2S on the hematite suspension, and not with AQDS on

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the hematite suspension (Figure 3A). The absence of an EPt decrease with AQDS in the hematite

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suspension is because the standard reduction potential of AQDS is substantially lower than the

320

mineral reduction potential (greater than 60 mV), which also explains the lack of AQDS

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reduction by hematite with 1 mM Fe(II) shown in Figure 1A. Based on constraints defined by

322

the Nernst equation, Sander et al. concluded that for an one-electron mediator to be effective in

323

transfering electrons between mineral and electrode, the difference between the standard

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reduction potentials of mineral and mediator should be less than 120 mV32. For two-electron

325

mediators used in this study, the difference should be less than 60 mV, which is consistent with

326

our data. CP experiments such as those shown in Figure 3A were performed with other oxides and

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oxide loadings, and the plateau values of EPt are shown in Figure 3B as a correlation with the

329

corresponding ECRP data (from Figure 2 and Table S3). All of the EPt data obtained without

330

added CRPs were significantly more positive than their ECRP counterparts, but addition of

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appropriate CRPs resulted in lower values of EPt that agree well with ECRP. These results

332

demonstrate that the potentials obtained by equilibration of CRPs with mineral suspensions are

333

consistent regardless of whether the potential is determined potentiometrically or

334

spectrophotometrically.

335 336 337 338 339

Figure 3. (A) The EPt of mineral suspension as a function of time measured by a 3 mm Pt electrode (rotated at 2000 rpm) and a Ag/AgCl reference electrode under Ar headspace (arrows indicate the time aliquots of thermodynamic CRPs were added); (B) Potentiometrically measured EPt vs. ECRP shown in Figure 2.

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Kinetic CRP Measurements and Links to ECRP. To explore the kinetic aspect of

340 341

reduction by minerals, we used three model contaminants (4Cl-NB, 2-CAP, and CT) that are

342

labile to abiotic degradation as kinetic CRPs in batch experiments with various doses of iron

343

oxide and aqueous Fe(II). No effort was made to quantify the reaction products as they have

344

been investigated in detail by prior studies.17, 20, 48 All of the data fit pseudo first order kinetics

345

well, and the resulting values of knobs were used to calculate the mass- and surface-area

346

normalized rate constants (kM and kSA, respectively) (Table S4). All three probes gave good

347

correlations to values of ECRP determined under equivalent conditions, but only the correlation to

348

kSA is shown (Figure 4) because it gave fewer outliers and should have the most generality. For

349

all three kinetic CRPs, the increase in kSA with decreasing ECRP of the mineral suspension appears

350

to be linear, so the trends were fit by linear regression. The slopes of these correlations—which

351

reflect the sensitivity of the reaction to mineral reduction potential—are similar for the two

352

dechlorination reactions (CT and 2-CAP) and roughly 2-fold less than for the nitro reduction

353

reaction (2-Cl-NB). Overall, the rate constants for nitro reduction are consistently greater than

354

for dechlorination (by 1-4 orders of magnitude). The linear correlations between log kSA and ECRP (Figure 4) resemble the (L)FERs

355 356

developed in previous studies formed between second-order rate constants for contaminant

357

reduction and one electron-reduction potentials of aqueous Fe(II)–ligand complexes.31, 56

358

However, the mechanistic significance of the LFERs reported here is less clear because both the

359

response (kSA) and descriptor (ECRP) variables used are confounded with effects for which there

360

still is no practical method of normalization. For kSA, the main issue is that normalization to total

361

surface area may not reflect variations in reactive surface area (e.g., surface coverage of reactive

362

sites).57 For ECRP, the apparent dependence on dose of mineral and Fe(II) (Figure 2) suggests

363

that this is an operationally-defined measure of effective reduction potential (rather than an ideal

364

thermodynamic property of the materials). This is expected because there are many factors that

365

make reduction potentials difficult to define for sorbed Fe(II) as Fe(II) sorption is proved to be a

366

dynamic and complex process.22, 23, 58, 59 However, the relative success of the LFERs based on

367

ECRP in Figure 4 could also be evidence that the value of this descriptor variable reflects actual

368

differences in the material properties of each preparation.

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369 370 371 372 373 374 375

Figure 4. Pseudo-first order rate constants (log kSA) for reduction of 4Cl-NB, 2-CAP, and CT by iron minerals versus mineral reduction potentials determined by chemical reactivity probes (ECRP) Fitting results (slope, intercept, r2) are −0.029±0.005, −6±1, 0.839; −0.008±0.001, −5.8±0.3, 0.885; and −0.011±0.003, −6.1±0.6, 0.738 for 4-Cl-NB, 2-CAP, and CT, respectively. (Oxide loading (g/L), type, and Fe(II) concentration (mM) for each data point are annotated next to the marker).

To explore the generality of the correlations between mineral reduction potential and

376 377

contaminant reduction kinetics, we compiled kinetic data from previous studies on the abiotic

378

reduction of three classes of contaminants by Fe(II) containing oxides and sulfides (summarized

379

in Tables S5–S8). In addition to nitroaromatics (represented by NB and 4Cl-NB), and

380

chlorinated alkanes (represented by CT and HCA), we also included chlorinated ethenes

381

(represented by TCE and PCE) in order to examine whether the approach developed in this study

382

can be extended to those environmentally-relevant and highly-recalcitrant contaminants. The

383

literature data (kobs) were normalized to mineral mass loading (kM) instead of kSA, as surface area

384

data were not provided in all studies. The results are summarized in Figure 5 by plotting kM

385

(from the literature and this study) versus a semi-quantitative ranking of the reduction potential

386

of abiotic iron oxide and sulfide based minerals. The position of each mineral included in this

387

study (magnetite, goethite, lepidocrocite, and hematite) on the x-axis was assigned to the

388

corresponding value of ECRP measured at 1 mM Fe(II) in this study. For green rust and iron

389

sulfide, their reduction potentials were more negative than could be characterized with the

390

thermodynamic CRPs that we have tested so far (data not shown), so they were assigned 1/24/16


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Page 16 of 23

391

potentials slightly below the limit of the most strongly reducing CRP tested (AQS). FeS was

392

assigned a lower potential than green rust as FeS is generally considered to be more reactive, but

393

the difference between these two potentials was arbitrary.

394 395 396 397 398

Figure 5. Mass-normalized reduction rate constants (kM) of kinetic CRPs (including nitroaromatics and chlorinated alkanes) and chlorinated ethenes with various reducing iron minerals (open symbols: values reported in literature; solid symbols: value determined in this study). Literature data used in this plot are summarized in Tables S5–S8.

Figure 5 shows variable, sometimes substantial, degrees of scatter, some of which is

399 400

undoubtedly because the values of ECRP used to define the descriptor axis do not fully represent

401

all of the operational factors that vary within and among the experiments summarized (Tables

402

S5-S8). For example, the variation in kM for CT reduction by magnetite appears to be related to

403

differences in synthesis methods (Table S5), and kM for reduction of chlorinated compounds by

404

FeS appears to be affected by whether freeze drying is used during mineral preparation (Table

405

S8). However, despite the variability in kM that is not described by the representative values of

406

ECRP used to make Figure 5, the kinetic data for the three classes of contaminants do form

407

clusters and trends that offer potentially useful insights into aspects of abiotic contaminant

408

degradation by reducing minerals. The relative rates of the three classes of contaminant reduction

409

processes (nitro aromatics > chlorinated alkanes > chlorinate alkenes) is consistent with other

410

evidence (e.g., Figure 4 and Elsner20). Similarly, the decrease in reduction rates with less 1/24/16


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411

strongly reducing minerals that is evident for nitro aromatics and chlorinated alkanes is

412

consistent with expectations (e.g., Figure 4), but the available data are insufficient to

413

demonstrate if this trend holds for reduction of chlorinated ethenes by weakly reducing mineral

414

phases. Nevertheless, assuming a slope similar to the other two classes of contaminants,

415

extrapolation from the available data for chlorinated ethenes to the less reactive oxides suggests

416

rate constants on the order of 10−6 hr−1 (t1/2 ≈ 8 yr). Such slow rates would be very difficult to

417

measure in laboratory batch experiments (which undoubtedly is part of the reason for the absence

418

of literature data), but they could still be environmentally-significant in the context of abiotic NA

419

in the field.

420

Implications for Assessing Abiotic NA. Characterization of abiotic NA of recalcitrant

421

contaminants, such as TCE, in the subsurface is challenging mainly because the absolute

422

degradation rates are slow under most field conditions. These challenges might be overcome by

423

characterization using kinetic and thermodynamic CRP methods developed in this study.

424

However, extension of this approach to previously reported kinetic data from laboratory studies

425

of abiotic contaminant reduction by reducing minerals confirmed that there is considerable

426

variability in reactivity within classes of minerals (presumably due to operational factors ranging

427

from accessibility of reactive surface area to electronic effects of mineral stoichiometry). Much

428

of this variability may be reduced by in situ measurements with thermodynamic CRPs, but the

429

overall behavior of the thermodynamic CRPs when applied to real environmental samples may

430

be affected by factors that were not addressed in this study: sorption to natural organic matter or

431

non-reactive mineral surfaces, non-uniform permeability and transport, etc. Controlled

432

characterization of these effects probably is best done with column experiments (e.g. with

433

aquifer material collected from field sites), but the ultimate goal of this study is to develop CRP-

434

based assays that can be used as reactive tracers in push-pull tests for in situ characterization of

435

field conditions. Preliminary results suggest that the thermodynamic CRP assays developed in

436

this study are effective in push-pull tests performed on laboratory columns, and these CRPs are

437

compounds that should be suitable for field-scale applications, in terms of cost, toxicity, etc.

438

Acknowledgements

439

The work was supported by a grant from the Strategic Environmental Research and

440

Development Program of the U.S. Department of Defense (ER-2308). Surface areas of the

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441

oxides were measured by Maria Fuentes-Deonate, while on an internship funded by the NSF

442

Center for Coastal Margin Observation and Research.

443

Supporting Information Available:

444

Details on the mineral preparation, mineral surface area, properties of thermodynamic CRPs,

445

reduction kinetics of kinetic CRPs, and compiled literature data of contaminant reduction rate

446

constants can be found in the Supporting Information. This material is available free of charge

447

via the Internet at http://pubs.acs.org.

448

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Chemical Reactivity Probes for Assessing Abiotic Natural Attenuation

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