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Oxidation of a dimethoxyhydroquinone by ferrihydrite and goethite nanoparticles: iron reduction versus surface catalysis Lelde Krumina, Gry Lyngsie, Anders Tunlid, and Per Persson Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b02292 • Publication Date (Web): 10 Jul 2017 Downloaded from http://pubs.acs.org on July 11, 2017
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Oxidation of a dimethoxyhydroquinone by ferrihydrite and goethite nanoparticles: iron
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reduction versus surface catalysis
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Lelde Krumina,1,2 Gry Lyngsie,1 Anders Tunlid,2 Per Persson1,2*
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Centre of Environmental and Climate Research, Lund University, SE-223 62, Lund, Sweden
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Department of Biology, Lund University, SE-223 62, Lund, Sweden
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*Corresponding author:
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E-mail:
[email protected] 12
Phone: +46 46-222 17 96
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Abstract
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Hydroquinones are important mediators of electron transfer reactions in soils with a capability
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to reduce Fe(III) minerals and molecular oxygen, and thereby generating Fenton chemistry
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reagents. This study focused on 2,6-dimethoxy hydroquinone (2,6-DMHQ), an analogue to a
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common fungal metabolite, and its reaction with ferrihydrite and goethite under variable pH
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and oxygen concentrations. Combined wet-chemical and spectroscopic analyses showed that
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both minerals effectively oxidized 2,6-DMHQ in the presence of oxygen. Under anaerobic
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conditions the first-order oxidation rate constants decreased by one to several orders of
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magnitude depending on pH and mineral. Comparison between aerobic and anaerobic results
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showed that ferrihydrite promoted 2,6-DMHQ oxidation both via reductive dissolution and
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heterogeneous catalysis while goethite mainly caused catalytic oxidation. These results were
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in agreement with changes in the reduction potential (EH) of the Fe(III) oxide/Fe(II)aq redox
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couple as a function of dissolved Fe(II) where EH of goethite was lower than ferrihydrite at
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any given Fe(II) concentration, which makes ferrihydrite more prone to reductive dissolution
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by the 2,6-DMBQ/2,6-DMHQ redox couple. This study showed that reactions between
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hydroquinones and iron oxides could produce favorable conditions for formation of reactive
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oxygen species, which are required for non-enzymatic Fenton-based decomposition of soil
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organic matter.
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Introduction
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Quinones are redox-active compounds that occur in three different oxidation states, which are
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coupled via one-electron transfer reactions. These redox states determine whether the quinone
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acts as an electron acceptor (quinone, Q), electron donor (hydroquinone, H2Q) or as an
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intermediate semiquinone radical (SQ-). These unique redox properties make quinones
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effective electron shuttles in a range of biological and soil processes.1–4 Quinones have been
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shown to be present in both soil and aquatic natural organic matter as a result of the
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decomposition of organic litter material.5 They are also biosynthesized by a number of
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microbes, plants and insects.6–8 The occurrence and redox chemistry of quinones have been
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thoroughly reviewed by Uchimiya and Stone.9,10
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Quinones take part in the degradation of wood via the so-called brown-rot mechanism
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mediated by several different fungi.11–14 This mechanism involves hydroxyl radicals (OH)
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generated by a reaction between Fe(II) and H2O2 (the Fenton reaction), and in this context the
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hydroquinone plays a dual role. It acts as a reductant towards Fe(III) producing Fe(II), and
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produces a superoxide radical (OOH) when oxidized by O2. This radical is also generated
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when the intermediate semiquinone or Fe(II) reacts with O2. The superoxide may
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subsequently dismutate and form H2O2,13,15 or react with Fe(III) to generate Fe(II).16,17 Hence,
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the hydroquinone has the potential to produce both Fenton reagents.
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The hydroxyl radicals generated by the Fenton reaction can degrade a wide range of organic
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compounds, including cellulose and partially also lignin. In brown-rot wood degradation the
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fungal metabolite identified for this purpose, so far, is the 2,5-dimethoxyhydroquinone (2,5-
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DMHQ). 2,5-DMHQ has been found in three distantly related species of brown-rot fungi, i.e.
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Gloeophyllum trabeum,12 Postia placenta13 and recently Serpula lacrymans.14 These findings
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suggest that the pathways for synthesizing 2,5-DMHQ were present in a common ancestor; in
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turn suggesting that 2,5-DMHQ is a rather widespread fungal metabolite. While the
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degradation of wood by OH has been thoroughly studied less is known about these processes
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in soil and their potential contribution to decomposition of soil organic matter (SOM). It has
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been shown that SOM is decomposed by an ectomycorrhizal fungus (Paxillus involutus) via
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non-enzymatic radical reactions.18 Since ectomycorrhizal fungi are common in boreal soils
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this suggests that radical-based reactions triggered by extracellular metabolites can have a
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substantial influence on the turnover of SOM. In order to assess the potential contribution
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from radical-based SOM degradation in complex soil environments, the reactions between
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extracellular metabolites and soil components have to be investigated and characterized. In
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the case of hydroquinones three main oxidation pathways has to be considered: (1)
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autoxidation to quinones by O2 present in the system; (2) catalytic oxidation by transition
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metal catalysts, similar to autoxidation but promoted by catalysts; (3) coupled hydroquinone
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oxidation and metal reduction. These are represented by the following overall reactions:
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(1) Autoxidation: H2Q + O2 → Q + H2O2
O2-dependent ௧௬௦௧
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(2) Redox metal as catalyst: H2Q + O2 ሱۛۛۛۛሮ Q + H2O2
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(3) Reductive dissolution: H2Q + 2FeOOH(s) → Q + 2Fe(II)aq + 4OH-
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Oxygen is the electron acceptor in the first two reactions hence these are O2-dependent while
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the third reaction involves a metal ion as electron acceptor.19 The latter reaction will thus be
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independent of the oxygen concentration.
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Earlier soil-relevant studies have primarily focused on the reactions between Fe(III)-bearing
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soil minerals and the simplest of hydroquinones, 1,4- and 1,2-hydroquinone (catechol).20–24 In
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buffered systems and anaerobic conditions these indicated a reaction stoichiometry of 1 H2Q:
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2 Fe(II): 1 Q; i.e. one molecule of H2Q produced two Fe(II) and one Q.20–23 The reaction
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mechanism proposed involves rapid hydroquinone adsorption to the mineral surface followed
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by electron transfer and formation of a semiquinone radical and an Fe(II) ion. A second
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reaction cycle is repeated with the semiquinone and another Fe(II) is released and the
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benzoquinone is formed. Comparison between ferrihydrite and goethite particle reactivity
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towards 1,4-hydroquinone showed that ferrihydrite was more reactive than goethite.
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Moreover, ferrihydrite reduction rates increased with decreasing particle size and increasing
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specific surface area.20,21
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In comparison with 1,4-hydroquinone and catechol the fungal metabolite 2,5-DMHQ has two
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additional –OCH3 groups, which increase the electron density of the aromatic system and
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thereby lowers the reduction potential. Indeed, fungi producing this extracellular metabolite
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have been shown to reduce soluble Fe(III) as well as initiating the Fenton reaction.11–13 In this
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study we explore 2,6-DMHQ, a commercially available isomer of 2,5-DMHQ, and the effect
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of the electron-donating substituents on hydroquinone reactivity towards the common Fe soil
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minerals ferrihydrite and goethite. The main objectives of the study were to determine how
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the 2,6-DMHQ oxidation pathways were affected by pH, oxygen concentrations and mineral
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property. This was accomplished by combining solution chemical analysis with in-situ IR
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spectroscopy, which probes the reactions at the water-mineral interfaces.
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Materials and methods
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Chemicals.
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dimethoxybenzoquinone, 97%) were purchased from Sigma-Aldrich (see S1.1 for quinone
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chemical properties). All quinone solutions were prepared fresh before each experiment using
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deionized water that was boiled and bubbled with N2 to remove CO2 from the starting
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solutions.
2,6-DMHQ
(2,6-dimethoxyhydroquinone,
97%),
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2,6-DMBQ
(2,6-
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Ferrihydrite and goethite were synthesized according to procedures previously described.25
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Typical spherical ferrihydrite and needle-shaped goethite particles were identified by
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transmission electron microscopy (Figure S1) and the structural purity was confirmed by X-
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ray diffraction. The particles used in the batch experiments were not dried at any stage but
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stored under nitrogen at pH 6-7 in a fridge at 4 °C as 1.9 g/L ferrihydrite and 10.6 g/L
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goethite stock suspensions. For characterization purposes only the iron oxide particles were
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freeze-dried. The surface areas of the goethite and the ferrihydrite were estimated to 66 and
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300 m2/g, respectively (see S1.1).
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Batch experiments. The 2,6-DMHQ oxidation and reductive iron oxide dissolution as a
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function of pH and dissolved oxygen concentrations were studied by means of batch
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experiments (see S1.2). These were performed at two different approximate oxygen
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concentrations denoted aerobic and anaerobic, respectively, in the following text. The aerobic
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condition corresponds to experiments at ambient atmospheric pressure and [O2] = 270-310
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µM whereas the anaerobic experiments were performed in a glove bag filled with N2 (g) with
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[O2] = 0-20 µM. These concentrations were measured in the stock solution and initial
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representative batch samples using either an oxygen optode system (UniSense) or a
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polarographic probe (Thermo scientific Orion Star A213); note that O2 was not monitored
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during the experiments. The total 2,6-DMHQ concentrations were normalized with respect to
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the iron oxide surface area in order to obtain the same concentrations in units of µmol/m2 for
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ferrihydrite and goethite. After reaction the quinone species were analyzed by means of
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HPLC and Fe(II) in solution was determined by the Ferrozine assay. The batch experiments
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were performed in triplicates and Fe(II) was measured in each replicate, while HPLC was
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performed on either duplicates or on one of these replicates only (see S1.2).
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IR spectroscopy. IR spectra as a function of time at fixed pH values were collected in-situ
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using a modified version of the simultaneous infrared and potentiometric method previously
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reported by Loring et al.26 and Krumina et al.25 (see S1.3). The IR spectral data sets were
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analyzed by means of a multivariate curve resolution with alternating least squares (MCR-
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ALS) Matlab script following the procedures described by Jaumot et al.27 This procedure
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decomposed the data sets into component spectra contributing to the spectral variation and
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their corresponding concentration profiles. The data sets were prepared for MCR-ALS
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analysis using the script of Felten et al.,28 which included asymmetric least squares smoothing
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baseline correction, and the final analysis was carried out with the script by Jaumot et al.27
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Density functional theory (DFT) calculations. Geometry optimization and frequency
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calculations were performed for 2,6-DMHQ and 2,6-DMBQ as well as the corresponding
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semiquinone both in a protonated and unprotonated state. We employed density functional
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theory (DFT) using the hybrid functionals B3LYP and the standard 6-31++G(d,p) basis set.
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Solvation effects were modeled using different numbers of explicit water molecules. The
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calculations and visualizations were performed with the program Spartan ‘14 by
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Wavefunction Inc.
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Results
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2,6-DMHQ oxidation by ferrihydrite and goethite. The batch experiments showed that
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ferrihydrite and goethite particles significantly enhanced the 2,6-DMHQ oxidation rates as
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compared to autoxidation under aerobic conditions in solution (Figure 1). Moreover, the pH
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dependence of the oxidation was similar for both minerals, and displayed increased aerobic
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oxidation rates when pH was raised from 4.5 to 7.0. The initial rates of 2,6-DMHQ oxidation
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approximately followed first-order kinetics (Figure S4). The first-order rate constants
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highlighted the large effects of the iron oxide surfaces at aerobic conditions, increasing the
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rate constants by ca. 1-2 orders of magnitude as compared to autoxidation (Table 1). The
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initial ferrihydrite rates were faster than the corresponding goethite ones at all conditions
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investigated. This indicated a higher reactivity of ferrihydrite, but at pH 4.0 and 4.5 there
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might also be an effect from a greater pH drift in presence of ferrihydrite (see Table S2) that
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will speed up the 2,6-DMHQ oxidation. The 2,6-DMHQ oxidation was significantly slower at
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anaerobic conditions, and this effect was most pronounced for goethite. At pH 4.5 both
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minerals accomplished complete aerobic oxidation after 4 h, whereas only ca. 80% or 20%
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was oxidized by ferrihydrite and goethite, respectively, during the same time at anaerobic
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conditions.
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We also observed reduced 2,6-DMHQ oxidation efficiency as a function of ferrihydrite age
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(Figure S5). This was consistent with previously detected aging and phase transformation
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effects in ferrihydrite suspensions.29 These effects were not observed in the goethite
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experiments (Figure S5), which is consistent with the thermodynamic stability of this phase.
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Figure 1. Oxidation of 2,6-DMHQ (520 µM, 1.5 µmol/m2, corresponding to 1.9 g/L and 10.6
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g/L for ferrihydrite and goethite, respectively) as a function of time in presence and absence
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of iron oxide particles. pH 7.0 aerobic,
pH 4.0 aerobic,
pH 4.5 aerobic,
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pH 4.5 anaerobic,
pH 7.0 anaerobic. At pH 4.0 and 4.5 in the presence ferrihydrite
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pH was continuously adjusted during the experiment with 40 mM HCl or 40 mM NaOH in
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0.1 M NaCl. No significant pH drift was observed in any experiment at pH 7.0 or in the
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presence of goethite. Ferrihydrite experiments were performed with particles aged for 2
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weeks. Experimental uncertainties are exemplified by duplicates analyzed in some
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experimental series.
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Table 1. First order rate constants, k (in min-1), of 2,6-DMHQ oxidation in presence and
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absence of iron oxides.* Ferrihydrite
Goethite
Autoxidation
kpH4, aerobic
1.5 ×10-1
2.0 ×10-2
6.8 ×10-4
kpH4.5, aerobic
1.2 ×10-1
3.0 ×10-2
7.2 ×10-4
kpH4.5, anaerobic
2.2 ×10-2
1.6 ×10-3
kpH7, aerobic
3.2 ×10-1**
1.8 ×10-1
kpH7, anaerobic
8.5 ×10-4
3.3 ×10-4
2.0 ×10-2
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*The rate constants were calculated from the initial points in Figure 1 obeying first-order kinetics.
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**Calculated from an experiment with ferrihydrite aged for 4 months in order to capture the fast 2,6-DMHQ oxidation (see Figure S4).
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Reaction products of 2,6-DMHQ oxidation. In presence of both 2,6-DMHQ and 2,6-
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DMBQ HPLC detected low concentrations of a third species that were assigned to the
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semiquinone (Figure S6). We base our assignment on the transient nature of this component,
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which is correlated to the co-existence of the hydroquinone and the quinone, and the presence
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in pure aqueous solution during autoxidation (Figure S6). This behavior of semiquinones is
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well documented in the literature and from previous results semiquinone concentrations in the
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micro-molar range are not unrealistic under our experimental conditions.30 When all
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hydroquinone was oxidized only the benzoquinone was detected. However, the complete
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aerobic oxidation of 2,6-DMHQ by the iron oxides (Figure 1) did not always result in a 100%
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recovery of 2,6-DMBQ (Figure 2). Separate IR experiments showed that adsorption of 2,6-
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DMBQ onto the iron oxide surfaces was negligible (data not shown), and accordingly the
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incomplete recovery of 2,6-DMBQ indicated side reactions. The fact that these products were
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not detected by HPLC suggested strong affinities for the iron oxide surfaces, and thus
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enrichment at the water-mineral interface. The side reactions were most significant at low pH
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whereas the 2,6-DMBQ recovery was almost complete at pH 7.0 in presence of both iron
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oxides (Figure 2). The results supported previous suggestions that a range of different reaction
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products, such as dimers, trimers and other polymers, aldehydes and even CO2, can form as a
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result of the reactions between hydroquinones or phenolics and mineral surfaces at aerobic
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reaction conditions.24,31–41
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Figure 2. Sum of the 2,6-DMHQ and 2,6-DMBQ concentrations after 240 minutes, expressed
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as the percentage of the total 2,6-DMHQ concentrations added in the batch experiments.
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Iron reduction. Under aerobic conditions Fe(II) concentrations generated by reductive
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dissolution were significantly lower than the amounts of 2,6-DMHQ oxidized (Figure 1 and
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3), and thus far from the theoretical 2 Fe:1 hydroquinone limit of iron oxide reduction by 2,6-
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DMHQ (Reaction 3 and Figure S7). Merely 30-40 µM Fe(II) was produced aerobically from
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goethite by 520 µM 2,6-DMHQ at pH 4.5, this was increased to ca. 70 µM at pH 4.0. A 70
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µM Fe(II) concentration was also obtained under anaerobic conditions at pH 4.5. This pattern
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of increasing Fe(II) concentration at low O2 was also observed when a lower total
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concentration of 2,6-DMHQ was employed (90 µM, 0.4 µmol/m2) (Figure 3). In general, the
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low Fe(II) concentrations together with the strong dependence of 2,6-DMHQ oxidation on O2
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concentration (Figure 1) indicated minor contribution from reductive dissolution (Reaction 3)
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and a predominance of catalytic oxidation (Reaction 2) in the presence of goethite. Note that
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the Fe(II) concentration may underestimate the reductive dissolution somewhat due to re-
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adsorption and/or re-oxidation (Figure S8). At pH 7.0 no Fe(II) was detected in solution in
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presence of goethite.
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Reduction of ferrihydrite at pH 4.0 and 4.5 yielded substantially higher Fe(II) concentrations
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as compared to goethite, but again these concentrations were close to or below the detection
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limit at pH 7.0 (Figure 3). The adsorption of Fe(II) onto ferrihydrite at pH 4.0 and 4.5 has
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been reported to be very low and correlated to the re-oxidation.42 Under our experimental
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conditions in absence and presence of ferrihydrite at pH 4.5 and absence of buffers the loss of
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Fe(II), either via adsorption or re-oxidation, was slow (Figure S8). Thus, at pH 4.0 and 4.5 the
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dissolved Fe(II) concentration was a quantitative measure of the extent of reductive
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dissolution. As shown by the ratios between Fe(II) produced and 2,6-DMHQ consumed
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(Figure S7) and the decrease in dissolved Fe(II) at increasing O2 (Figure 3), the reductive
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process was predominating at anaerobic conditions whereas the competition from catalytic
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oxidation of 2,6-DMHQ increased in the presence of O2. Furthermore, comparison between
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the rates of 2,6-DMHQ consumption and Fe(II) production showed that catalytic oxidation is
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faster than reductive dissolution (Figure S9).
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Figure 3. Fe reductive dissolution by 2,6-DMHQ in 0.1 M NaCl from ferrihydrite (left) and
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goethite (right).
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aerobic,
520 µM 2,6-DMHQ at pH 4.5 anaerobic,
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aerobic,
90 uM 2,6-DMHQ at pH 4.5 anaerobic,
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and anaerobic conditions. At pH 4.0 and 4.5 in the presence ferrihydrite pH was continuously
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adjusted during the experiment with 40 mM HCl or 40 mM NaOH in 0.1 M NaCl. No
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significant pH drift was observed in any experiment at pH 7.0 or in the presence of goethite.
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IR spectroscopy. The IR spectra of ferrihydrite reacted with 2,6-DMHQ at pH 4.5 and 7.0
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showed marked changes with time indicating at least two predominant surface species with
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different reaction kinetics (Figure S10). MCR analyses of these data sets resolved two main
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components (Figure 4). One component (C1), characterized by the bands at 1385 and 1532
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cm-1, displayed a steady increase and was completely dominating at the end of the
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experiment. At this point 2,6-DMBQ is the dominating quinone of the triad under aerobic
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conditions (Figure 1 and 2). Additional experiments performed with 2,6-DMBQ at identical
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total concentration produced IR data with no discernible bands implying that 2,6-DMBQ has
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very low affinity for the iron oxide surfaces. Therefore, the 1385 and 1532 cm-1 bands likely
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originated from the products of the side-reactions indicated by the incomplete 2,6-DMBQ
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recovery (Figure 2). Note that although the positions and relative intensities of these bands
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were very similar at all investigated conditions the C1 spectra contained bands that varied
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between the experiments e.g. at 1595 cm-1 and between 1000-1250 cm-1. This can either be
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due to incomplete separation from the C2 component discussed below or that range of
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additional side-products are formed depending on the experimental conditions. In any case,
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the strong contribution of the 1385 and 1532 cm-1 bands to the IR spectra showed that this
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side-product has a high affinity for the iron oxide surfaces.
520 µM 2,6-DMHQ at pH 4 aerobic,
520 µM 2,6-DMHQ at pH 4.5 90 µM 2,6-DMHQ at pH 4.5 520 µM 2,6-DMHQ at pH 7.0
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The second MCR component (C2) was the predominant surface species during the first 20-30
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minutes, and the C2 spectra varied as a function of the experimental conditions (Figure 4). At
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pH 7 and aerobic conditions the C2 spectrum closely resembled that of 2,6-DMHQ(aq) as well
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as the DFT calculated IR spectrum of hydrated 2,6-DMHQ (Figure S11). On the other hand,
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the C2 spectrum was distinctly different from the DFT-calculated IR spectra of the
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semiquinone and quinone species (cf. Figure S11 and S12). Accordingly, these results
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indicated adsorption of intact 2,6-DMHQ. The small band shifts as compared to 2,6-
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DMHQ(aq) mainly concerned complex vibrational motions involving hydrogens (Table S3)
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and we attribute these shifts to interactions between the OH groups of the neutral
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hydroquinone and the positively charged hydrated surface. At pH 4.5 and aerobic conditions
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the C2 spectrum lost the band at 1595 cm-1 and the main in this region band now appeared
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around 1500 cm-1 (Figure 4). The computed IR spectra of the quinone triad indicated that this
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shift was caused by the presence of semiquinones (Figure S12), and the overall spectral
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features of the C2 spectrum resembled those of the deprotonated semiquinone, or a mixture
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with the protonated form (Figure S13). The detection of the semiquinone was also in
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agreement with the HPLC results that suggested the existence of the semiquinone at pH 4.5
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and aerobic conditions during the first part of the experiment (Figure S6). Under anaerobic
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conditions at pH 4.5 the C2 spectrum was more similar to hydrated 2,6-DMHQ but with some
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contributions from the semiquinone. Hence, indicating a shift in distribution between the
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hydroquinone and the semiquinone induced by the low oxygen concentration.
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IR spectra collected at low 2,6-DMHQ concentration identical to batch conditions (1.5
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µmol/m2 ferrihydrite) yielded similar results but with different MCR concentration profiles
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(Figure S14). Under these conditions it was notable that C2 decayed faster (in agreement with
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Figure 1) whereas the behavior of C1, assigned to the side-reaction products, was in all
289
respects very similar to that observed at the high 2,6-DMHQ concentration.
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IR spectra of 2,6-DMHQ reacted with goethite were of poor quality and allowed only
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qualitative identification of the adsorbed hydroquinone and the bands at 1385 and 1532 cm-1
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originating from the oxidation side-products (Figure S15). The low signal-to-noise was most
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likely a result of the larger particle size of goethite as compared to ferrihydrite. This resulted
294
in a situation where a smaller number of adsorbed molecules in the ATR overlayers were
295
exposed to the IR light in the case of goethite at similar surface coverage.
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Figure 4. Multivariate curve resolution analysis of IR spectra of ferrihydrite during reaction
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with 2,6-DMHQ (total concentration = 9.9 µmol/m2). The IR data sets contained spectra were
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collected every minute for approximately 140 minutes. The estimated MCR concentration
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profiles and the corresponding spectra are coded using the same color and line style.
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Discussion
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Oxygen effects. A striking result from the present study was the strong oxygen-dependence
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of the iron oxide-mediated 2,6-DMHQ oxidation (Figure 1). The consumption of 2,6-DMHQ
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is due to oxidation by either Fe(III) or O2. Thus, under anaerobic conditions the Fe(III)-driven
306
reaction should predominate, and comparison between anaerobic and aerobic conditions
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facilitates distinction between oxidation via iron reductive dissolution and surface catalysis.
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Under all anaerobic conditions the first order oxidation rate constant decreased by 1 to 3
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orders of magnitude as compared to aerobic oxidation, with the largest relative effects at pH
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7.0 (Figure 1 and Table 1). It follows that catalytic oxidation is significant in the presence of
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oxygen. These results contrast those in a recent study on oxidation of 2-methoxyhydroquinone
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by low concentrations of soluble Fe(III) in aqueous solution that clearly showed similar and
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rapid oxidation rates both at aerobic and anaerobic conditions.43 This comparison emphasizes
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the importance of catalytic reactions in the presence of surfaces.
315
At pH 4.5 the difference in extent of 2,6-DMHQ aerobic and anaerobic oxidation was larger
316
for goethite than for ferrihydrite; both iron oxides caused complete aerobic oxidation after 4 h
317
while ca. 20% and 80% 2,6-DMHQ was remaining in presence of ferrihdyrite and goethite,
318
respectively, under aerobic conditions (Figure 1). Ferrihydrite mediated both Fe(III) reduction
319
and surface catalysis whereas Fe(III) reduction only played a minor role at goethite surfaces.
320
At pH 7.0 both iron oxides displayed an even more dramatic decrease in 2,6-DMHQ
321
oxidation rates when oxygen was excluded (Table 1). These results corroborated the minor
322
contribution at pH 7.0 from a mechanism involving Fe(III) reduction followed by rapid Fe(II)
323
adsorption and/or re-oxidation because in this case 2,6-DMHQ oxidation should have been
324
substantial also under anaerobic conditions. Moreover, these findings once more emphasized
325
the difference between oxidation in homogeneous solution and at iron oxide surfaces. Yuan et
326
al.43 showed that molecular oxygen did not contribute to hydroquinone oxidation in the
327
presence of aqueous Fe(III), but played an indirect role via the formation of OOH, by Fe(II)
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or semiquinone oxidation. Subsequently, the superoxide oxidized the hydroquinone. This
329
mechanism requires the formation Fe(II) and the semiquinone through Fe(III) reduction,
330
which was occurring only at a slow rate under our anaerobic conditions, if it occurred at all
331
(note that our experimental conditions were not strictly anaerobic and the slow oxidation
332
could be due to remaining low levels of dioxygen). Hence, Fe(III) reduction would be the
333
rate-limiting step also under aerobic conditions if oxidation is driven by the superoxide and
334
therefore cannot explain the very rapid reaction at pH 7.0 in the presence of both iron oxides.
335
This has to involve a heterogeneous catalytic reaction between 2,6-DMHQ and molecular
336
oxygen.
337
Reaction mechanisms. At pH 7.0 IR spectra indicated a rapid initial adsorption of intact 2,6-
338
DMHQ onto ferrihydrite (Figure 4 and S11), and at the same time 2,6-DMHQ was aerobically
339
oxidized via a surface catalytic reaction. This catalyzed oxidation of 2,6-DMHQ is in many
340
ways similar to autoxidation involving a reaction between the hydroquinone and molecular
341
oxygen. In the case of autoxidation, the rate has been correlated to the pKa values of different
342
hydroquinones i.e. the rate of oxidation increases with decreasing pKa values, and from this
343
follows that oxidation rates also increases with increasing pH.19 This was also shown by our
344
results where the rate constant of autoxidation increased from 6.8×10-4 min-1 at pH 4 to
345
2.0×10-2 min-1 at pH 7.0 (Table 1). It follows that any interaction that lowers the pKa values
346
and thereby promotes deprotonation of the hydroquinone potentially will increase the rate of
347
oxidation. Previous studies have shown that deprotonated forms of organic acids adsorbed to
348
iron oxides are stabilized, and the pKa values of the interfacial species are lowered
349
substantially compared to bulk solution.44–47 Accordingly, we propose that a contributing
350
factor to the catalytic oxidation at the iron oxide surfaces is a downward shift of the 2,6-
351
DMHQ pKa values resulting from the hydroquinone accumulation at the positively charged
352
interface. The oxidation rate may be further increased by electronic structure changes of
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hydroquinone surface species and/or molecular oxygen induced by interactions with Fe(III).48
354
The predicted high pKa values of 2,6-DMHQ (10.80 and 12.79) implies that increased
355
oxidation rates are achieved without predominance of the deprotonated forms at pH 7.0; i.e.
356
pKa values can shift and oxidation rates increase and still the deprotonated forms will only be
357
small fractions of the total hydroquinone concentration. It is therefore not surprising that our
358
IR spectra indicated the predominance of protonated 2,6-DMHQ on ferrihydrite (Figure 4, C2
359
dotted green).
360
At pH 4.5 the aerobic oxidation rates on ferrihydrite and goethite were slower than at pH 7.0,
361
thus following the same pH trend as autoxidation, whereras the relative effect by the iron
362
oxide surfaces as compared to autoxidation was greater at pH 4.5 than at pH 7 (Table 1).
363
Under these mildly acidic conditions ferrihydrite caused both catalytic oxidation and
364
oxidation via reductive dissolution while goethite mainly acted as a catalyst (Figure 1). IR
365
spectra of ferrihydrite indicated initial predominance of an adsorbed semiquinone when both
366
catalytic and reductive dissolution processes were active (Figure 4, C2 dotted blue). Thus,
367
adsorption to the iron oxide surface could play a role in further stabilizing the semiquinone.
368
Under anaerobic conditions at pH 4.5 the surface speciation on ferrihydrite changed and intact
369
2,6-DMHQ predominated, similar to the observations at pH 7.0 (Figure 4, C2 dotted red).
370
This difference in surface speciation between aerobic and anaerobic conditions suggested that
371
surface-catalyzed oxidation of adsorbed 2,6-DMHQ is a rapid process maintaining the surface
372
concentration of 2,6-DMHQ at a low level in presence of oxygen. The fact that we detected
373
adsorbed semiquinone also indicated that oxidation of this species is comparatively slow.
374
Furthermore, the shift to predominance of adsorbed 2,6-DMHQ under anaerobic conditions is
375
in agreement with a slower 2,6-DMHQ oxidation via reductive dissolution than via surface
376
catalysis (Table 1). This explains why the initial oxidation of 2,6-DMHQ is faster than the
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release of dissolved Fe(II) from both ferrihydrite and goethite under aerobic conditions at pH
378
4.5 (Figure S9).
379
Generation of dissolved Fe(II). A consequence of the parallel catalytic and reductive
380
dissolution pathways of aerobic 2,6-DMHQ oxidation was that Fe(II) concentrations
381
generated via the latter process were far from the limiting value of one oxidized 2,6-DMHQ
382
per 2 Fe(II) produced, predicted by reaction (3) (Figure 1 and 3, S7). Under anaerobic
383
conditions and low pH the concentrations were closer to this ratio, but the reaction slowed
384
down or even stopped at substantial concentrations of 2,6-DMHQ remaining in the
385
suspensions (Figure 1). In order to reconcile these Fe(II) results as well as differences
386
between ferrihydrite and goethite we need to consider the reduction potentials (EH) of the
387
reactants and their variations with experimental conditions. In a recent study Gorski et al.49
388
convincingly showed how EH of the Fe(III) oxide/Fe(II)aq redox couple was decreased as a
389
function of increasing Fe(II)aq concentration. At the same time EH of the 2,6-DMBQ/2,6-
390
DMHQ redox couple will increase as a result of the oxidation according to the Nernst
391
equation, and at some point EH(2,6-DMBQ/2,6-DMHQ) will be higher than EH(Fe(III)
392
oxide/Fe(II)aq) and the reductive dissolution will stop. In Figure 5 we have compared the
393
change in EH(Fe(III) oxide/Fe(II)aq) to EH(2,6-DMBQ/2,6-DMHQ) calculated from the
394
concentrations of these quinone species after 240 minute reaction time under anaerobic
395
conditions (Figure 1). At pH 4.5 in the presence of goethite EH(2,6-DMBQ/2,6-DMHQ) is
396
indicated to cause reductive dissolution when the Fe(II) concentration is below ca. 100 µM.
397
This explains why the Fe(II) concentration reaches a limiting value around 80 µM, despite the
398
fact that ca. 400 µM 2,6-DMHQ still remains in solution, because the difference in EH values
399
between the redox couples does not produce the thermodynamic driving force needed for
400
reductive dissolution. In contrast, in presence of ferrihydrite EH(Fe(III) oxide/Fe(II)aq) is
401
sufficiently positive for reductive dissolution to occur also at milli-molar Fe(II)
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concentrations, which is indicated also by the continued increase of the Fe(II) concentration
403
(Figure 3). Here, the reaction rate is slowed down as a result of the low remaining 2,6-DMHQ
404
concentration, and the dissolution will eventually stop since this concentration will become
405
too low.
406
At pH 7.0 under anaerobic conditions the same general trends are observed, but in this case,
407
EH(Fe(III) oxide/Fe(II)aq) drops below that of EH(2,6-DMBQ/2,6-DMHQ) at lower Fe(II)
408
concentrations (Figure 5B) Therefore reductive dissolution is less efficient at pH 7.0 as
409
compared to pH 4.5 under anaerobic conditions, and it is likely that in the presence of oxygen
410
the contribution from iron oxide reduction to the overall 2,6-DMHQ oxidation is small. We
411
conclude from this discussion that the relative importance of the reductive dissolution and
412
catalytic oxidation pathways at a given 2,6-DMHQ concentration will largely be determined
413
by EH of the iron oxides and the O2 concentration.
414
415 416
417
Figure 5. Reduction potentials (EH) of the Fe(III) oxide/Fe(II)aq redox couple as a function of
418
dissolved Fe(II) concentrations in the presence of ferrihydrite or goethite at (A) pH 4.5 and
419
(B) pH 7.0. The solid lines were calculated according to Ref. 49. The dashed lines represent
420
EH of the 2,6-DMBQ/2,6-DMHQ redox couple at the 2,6-DMHQ and 2,6-DMBQ
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concentrations after 240 minutes under anaerobic conditions calculated using the Nernst
422
equation and the EH0 value from Huynh et. al.50 (see S2.3).
423
Quinone decomposition, generation of radicals and implications to soil processes.
424
Previous studies have demonstrated that efficient degradation of aromatic molecules can be
425
accomplished by generating ROS through H2O2 decomposition using Fe-containing materials
426
as catalysts.51–53 Some of these catalytic systems are capable of ring cleavage thus converting
427
aromatic molecules into aliphatic products. A key species in these reactions is the OH, which
428
is formed via Haber-Weiss or Fenton-like processes. We conclude from this previous
429
literature that the 2,6-DMHQ oxidation side-products detected in our experiments (Figure 2)
430
probably are caused by reactions involving radicals. According to Reaction (2) the catalytic
431
oxidation of 2,6-DMHQ by the iron oxides has the potential of generating one H2O2 molecule
432
for every oxidized hydroquinone, and this reaction was most efficient in presence O2 at pH
433
7.0. It has been shown that Fe-containing solids can catalyze H2O2 decomposition and
434
generate hydroxyl radicals,54,55 but at the same time at pH 7.0 H2O2 is unstable and
435
disproportionate into O2 and H2O, a process which also is catalyzed by surfaces.54,56 Our
436
results from the pH 7.0 experiments showed that most 2,6-DMHQ was converted into 2,6-
437
DMBQ with little formation of side products, which indicated that generation of radicals at
438
this pH was inefficient. Still, the IR experiment performed at pH 7.0 and high 2,6-DMHQ
439
concentration showed small amounts of side products adsorbed to ferrihydrite indicating non-
440
negligible radical production.
441
At acidic pH values the thermodynamic stability of H2O2 increases. In addition, reductive
442
dissolution by 2,6-DMHQ is more efficient, which generates dissolved Fe(II) especially in
443
presence of ferrihydrite (Figure 3). This Fe(II) has a dual role. Together with H2O2 it will
444
generate OH via Fenton reaction and it may also contribute to the production of H2O2
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through re-oxidation and the formation of superoxide which subsequently dismutate to
446
peroxide:
447
(4) Fe(II)aq + O2 + H+ → Fe(III)s + HOO·
448
(5) 2HOO· → H2O2 + O2
449
Thus, at pH 4.5 under aerobic conditions the reactions between 2,6-DMHQ and the iron
450
oxide, produce all necessary ingredients to generate OH; H2O2 primarily via the catalytic
451
oxidation and Fe(II) via the reductive dissolution. Indeed, around 25% of 2,6-DMHQ was
452
converted into side products under these conditions in the presence of ferrihdyrite and
453
goethite (Figure 2). A decrease of O2 will lead to lower production of H2O2, and therefore the
454
lower production of oxidation side products in our anaerobic experiments (Figure 2) was
455
likely caused by H2O2 limitation.
456
Previous studies on soil remediation have demonstrated that belowground injection of H2O2
457
can effectively degrade organic contaminants, in particular in presence of iron
458
minerals.41,51,57,58 Similarly, the aggregated results from this study have shown that redox
459
reactions between dimethoxyhydroquinones, exuded by several brown-rot fungi, and iron
460
oxide minerals can produce conditions at the water-mineral interfaces that degrade organic
461
compounds. This implies that such non-enzymatic radical-based processes have the potential
462
to contribute to decomposition of soil organic matter without an external supply of H2O2. In
463
this respect, the possibility to generate Fenton chemistry at surfaces that also accumulate
464
organic matter will make the decomposition process efficient because the hydroxyl radical
465
operates at very short length-scales.
466
Acknowledgments
467
This work was supported by grants from the Swedish Research Council (VR) and the Knut
468
and Alice Wallenberg Foundation.
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Supporting Information. Additional information on chemicals, experimental methods,
470
results from quantitative chemical analysis and IR spectroscopy.
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References
472
(1)
and the O-Quinone Family of Cofactors. J. Nutr. 2000, 130, 719–727.
473
474
(2)
(3)
Newman, D. K.; Kolter, R. A Role for Excreted Quinones in Extracellular Electron Transfer. Nature 2000, 405, 94–97.
477
478
Misra, H. S.; Rajpurohit, Y. S.; Khairnar, N. P. Pyrroloquinoline-Quinone and its Versatile Roles in Biological Processes. J. Biosci. 2012, 37, 313–325.
475
476
Stites, T. E.; Mitchell, A. E.; Rucker, R. B. Physiological Importance of Quinoenzymes
(4)
Field, J. A.; Cervantes, F. J.; Van der Zee, F. P.; Lettinga, G. Role of Quinones in the
479
Biodegradation of Priority Pollutants: A Review. Water Sci. Technol. 2000, 42, 215–
480
222.
481
(5)
Cory, R. M.; McKnight, D. M. Fluorescence Spectroscopy Reveals Ubiquitous
482
Presence of Oxidized and Reduced Quinones in Dissolved Organic Matter. Environ.
483
Sci. Technol. 2005, 39, 8142–8149.
484
(6)
Biotechnol. Adv. 2009, 27, 185–194.
485
486
Sánchez, C. Lignocellulosic Residues: Biodegradation and Bioconversion by Fungi.
(7)
Tuor, U.; Winterhalter, K.; Fiechter, A. Enzymes of White-Rot Fungi Involved in
487
Lignin Degradation and Ecological Determinants for Wood Decay. J. Biotechnol.
488
1995, 41, 1–17.
489
(8)
Sclerotization of Insect Cuticle. Bioorg. Chem. 1987, 15, 194–211.
490
491 492
Sugumaran, M. Quinone Methide Sclerotization: A Revised Mechanism for β-
(9)
Uchimiya, M.; Stone, A. T. Redox Reactions between Iron and Quinones: Thermodynamic Constraints. Geochim. Cosmochim. Acta 2006, 70, 1388–1401.
ACS Paragon Plus Environment
25
Environmental Science & Technology
493
(10)
Uchimiya, M.; Stone, A. T. Reversible Redox Chemistry of Quinones: Impact on Biogeochemical Cycles. Chemosphere 2009, 77, 451–458.
494
495
(11)
Shimokawa, T.; Nakamura, M.; Hayashi, N.; Ishihara, M. Production of 2,5-
496
Dimethoxyhydroquinone by the Brown-Rot Fungus Serpula lacrymans to Drive
497
Extracellular Fenton Reaction. Holzforschung 2004, 58, 305–310.
498
(12)
Jensen, K. A.; Houtman, C. J.; Ryan, Z. C.; Hammel, K. E. Pathways for Extracellular
499
Fenton Chemistry in the Brown Rot Basidiomycete Gloeophyllum trabeum. Appl.
500
Environ. Microbiol. 2001, 67, 2705–2711.
501
(13)
Wei, D.; Houtman, C. J.; Kapich, A. N.; Hunt, C. G.; Cullen, D.; Hammel, K. E.
502
Laccase and Its Role in Production of Extracellular Reactive Oxygen Species during
503
Wood Decay by the Brown Rot Basidiomycete Postia Placenta. Appl. Environ.
504
Microbiol. 2010, 76, 2091–2097.
505
(14)
Korripally, P.; Timokhin, V. I.; Houtman, C. J.; Mozuch, M. D.; Hammel, K. E.
506
Evidence from Serpula lacrymans that 2,5-Dimethoxyhydroquinone is a
507
Lignocellulolytic Agent of Divergent Brown Rot Basidiomycetes. Appl. Environ.
508
Microbiol. 2013, 79, 2377–2383.
509
Page 26 of 31
(15)
Kerem, Z.; Jensen, K. A.; Hammel, K. E. Biodegradative Mechanism of the Brown Rot
510
Basidiomycete Gloeophyllum Trabeum: Evidence for an Extracellular Hydroquinone-
511
Driven Fenton Reaction. FEBS Lett. 1999, 446, 49–54.
512
(16)
Kinetics, Mechanism, and Implications. Environ. Sci. Technol. 1998, 32, 1417–23.
513
514 515
Lin, S.-S.; Gurol, M. D. Catalytic Decomposition of Hydrogen Peroxide on Iron Oxide:
(17)
Kwan, W. P.; Voelker, B. M. Decomposition of Hydrogen Peroxide and Organic Compounds in the Presence of Dissolved Iron and Ferrihydrite. Environ. Sci. Technol.
ACS Paragon Plus Environment
26
Page 27 of 31
Environmental Science & Technology
2002, 36, 1467–1476.
516
517
(18)
Rineau, F.; Roth, D.; Shah, F.; Smits, M.; Johansson, T.; Canbäck, B.; Olsen, P. B.;
518
Persson, P.; Grell, M. N.; Lindquist, E.; et al. The Ectomycorrhizal Fungus Paxillus
519
Involutus Converts Organic Matter in Plant Litter Using a Trimmed Brown-Rot
520
Mechanism Involving Fenton Chemistry. Environ. Microbiol. 2012, 14, 1477–1487.
521
(19)
Song, Y.; Buettner, G. R. Thermodynamic and Kinetic Considerations for the Reaction
522
of Semiquinone Radicals to Form Superoxide and Hydrogen Peroxide. Free Radic.
523
Biol. Med. 2010, 49, 919–962.
524
(20)
Line Ferrihydrite. J. Phys. Chem. C 2008, 112, 12127–12133.
525
526
(21)
Anschutz, A. J.; Penn, R. L. Reduction of Crystalline Iron(III) Oxyhydroxides Using Hydroquinone: Influence of Phase and Particle Size. Geochem. Trans. 2005, 6, 60–66.
527
528
Erbs, J. J.; Gilbert, B.; Penn, R. L. Influence of Size on Reductive Dissolution of Six-
(22)
Stack, A. G.; Eggleston, C. M.; Engelhard, M. H. Reaction of Hydroquinone with
529
Hematite: I. Study of Adsorption by Electrochemical-Scanning Tunneling Microscopy
530
and X-Ray Photoelectron Spectroscopy. J. Colloid Interface Sci. 2004, 274, 433–441.
531
(23)
Stack, A. G.; Rosso, K. M.; Smith, D. M. A.; Eggleston, C. M. Reaction of
532
Hydroquinone with Hematite: II. Calculated Electron-Transfer Rates and Comparison
533
to the Reductive Dissolution Rate. J. Colloid Interface Sci. 2004, 274, 442–450.
534
(24)
Iron Oxides. Clays Clay Miner. 1988, 36, 303–309.
535
536
(25)
Krumina, L.; Kenney, J. P. L.; Loring, J. S.; Persson, P. Desorption Mechanisms of Phosphate from Ferrihydrite and Goethite Surfaces. Chem. Geol. 2016, 427, 54–64.
537
538
Kung, K.-H.; McBride, M. B. Electron Transfer Processes Between Hydroquinone and
(26)
Loring, J. S.; Sandström, M. H.; Norén, K.; Persson, P. Rethinking Arsenate
ACS Paragon Plus Environment
27
Environmental Science & Technology
Coordination at the Surface of Goethite. Chem. - A Eur. J. 2009, 15, 5063–5072.
539
540
Page 28 of 31
(27)
Jaumot, J.; Gargallo, R.; De Juan, A.; Tauler, R. A Graphical User-Friendly Interface
541
for MCR-ALS: A New Tool for Multivariate Curve Resolution in MATLAB. Chemom.
542
Intell. Lab. Syst. 2005, 76, 101–110.
543
(28)
Felten, J.; Hall, H.; Jaumot, J.; Tauler, R.; de Juan, A.; Gorzsás, A. Vibrational
544
Spectroscopic Image Analysis of Biological Material Using Multivariate Curve
545
Resolution–Alternating Least Squares (MCR-ALS). Nat. Protoc. 2015, 10, 217–240.
546
(29)
Rejuvenation Cycle. Elements 2011, 7, 101–106.
547
548
Raiswell, R. Iron Transport from the Continents to the Open Ocean: The Aging-
(30)
Roginsky, V. A.; Pisarenko, L. M.; Bors, W.; Michel, C. The Kinetics and
549
Thermodynamics of Quinone-Semiquinone- Hydroquinone Systems Under
550
Physiological Conditions. J. Chem. Soc. Trans. 2 1999, 2, 871–876.
551
(31)
Geochim. Cosmochim. Acta 1989, 53, 961–971.
552
553
(32)
Sawhney, B.; Kozloski, R.; Isaacson, P.; Gent, M. Polymerization of 2,6Dimethylphenol on Smectite Surfaces. Clays Clay Miner. 1984, 32, 108–114.
554
555
LaKind, J. S.; Stone, A. T. Reductive Dissolution of Goethite by Phenolic Reductants.
(33)
Polubesova, T.; Eldad, S.; Chefetz, B. Adsorption and Oxidative Transformation of
556
Phenolic Acids by Fe(III)-Montmorillonite. Environ. Sci. Technol. 2010, 44, 4203–
557
4209.
558
(34)
Yong, R. N.; Desjardins, S.; Farant, J. P.; Simon, P. Influence of pH and Exchangeable
559
Cation on Oxidation of Methylphenols by a Montmorillonite Clay. Appl. Clay Sci.
560
1997, 12, 93–110.
561
(35)
Colarieti, M. L.; Toscano, G.; Ardi, M. R.; Greco, G. Abiotic Oxidation of Catechol by
ACS Paragon Plus Environment
28
Page 29 of 31
Environmental Science & Technology
Soil Metal Oxides. J. Hazard. Mater. 2006, 134, 161–168.
562
563
(36)
Chloroanilines at Metal Oxide Surfaces. J. Agric. Food Chem. 1998, 46, 2049–2054.
564
565
(37)
Shindo, H.; Higashi, T. Polymerization of Hydroquinone as Influenced by Selected Inorganic Soil Components. Soil Sci. Plant Nutr. 1986, 32, 305–309.
566
567
Pizzigallo, M. D. R.; Ruggiero, P.; Crecchio, C.; Mascolo, G. Oxidation of
(38)
Boyd, S. A.; Mortland, M. M. Radical Formation and Polymerization of Chlorophenols
568
and Chloroanisole on Copper(II)-Smectite. Environ. Sci. Technol. 1986, 20, 1056–
569
1058.
570
(39)
Synthesis of Hydroquinone-Derived Humic Polymers. Appl. Clay Sci. 1985, 1, 71–81.
571
572
Shindo, H.; Huang, P. M. The Catalytic Power of Inorganic Components in the Abiotic
(40)
Nishino, N.; Arey, J.; Atkinson, R. Yields of Glyoxal and Ring-Cleavage Co-Products
573
from the OH Radical-Initiated Reactions of Naphthalene and Selected
574
Alkylnaphthalenes. Environ. Sci. Technol. 2009, 43, 8554–8560.
575
(41)
Liu, H.; Bruton, T. A.; Li, W.; Buren, J. Van; Prasse, C.; Doyle, F. M.; Sedlak, D. L.
576
Oxidation of Benzene by Persulfate in the Presence of Fe(III)- and Mn(IV)-Containing
577
Oxides: Stoichiometric Efficiency and Transformation Products. Environ. Sci. Technol.
578
2016, 50, 890–898.
579
(42)
Jones, A. M.; Griffin, P. J.; Collins, R. N.; Waite, T. D. Ferrous Iron Oxidation under
580
Acidic Conditions - The Effect of Ferric Oxide Surfaces. Geochim. Cosmochim. Acta
581
2014, 145, 1–12.
582
(43)
Yuan, X.; Davis, J. A.; Nico, P. S. Iron-Mediated Oxidation of Methoxyhydroquinone
583
Under Dark Conditions: Kinetic and Mechanistic Insights. Environ. Sci. Technol. 2016,
584
50, 1731–1740.
ACS Paragon Plus Environment
29
Environmental Science & Technology
585
(44)
Boily, J. F.; Nilsson, N.; Persson, P.; Sjöberg, S. Benzenecarboxylate Surface
586
Complexation at the Goethite (α-FeOOH)/Water Interface: I. A Mechanistic
587
Description of Pyromellitate Surface Complexes from the Combined Evidence of
588
Infrared Spectroscopy, Potentiometry, Adsorption Data, and Surface Complexatio.
589
Langmuir 2000, 16, 5719–5729.
590
(45)
Page 30 of 31
Johnson, B. B.; Sjöberg, S.; Persson, P. Surface Complexation of Mellitic Acid to
591
Goethite: an Attenuated Total Reflection Fourier Transform Infrared Study. Langmuir
592
2004, 20, 823–828.
593
(46)
Tejedor-Tejedor, M. I.; Yost, E. C.; Anderson, M. A. Characterization of Benzoic and
594
Phenolic Complexes at the Goethite/Aqueous Solution Interface Using Cylindrical
595
Internal Reflection Fourier Transform Infrared Spectroscopy. Part 1. Methodology.
596
Langmuir 1990, 6, 979–987.
597
(47)
Lindegren, M.; Persson, P. Competitive Adsorption between Phosphate and Carboxylic
598
Acids: Quantitative Effects and Molecular Mechanisms. Eur. J. Soil Sci. 2009, 60,
599
982–993.
600
(48)
Bertini, I.; Gray, H. B.; Lippard, S. J.; Valentine, J. S. Dioxygen Reactions. In
601
Bioinorganic Chemistry; University Science Books, Mill Valley, CA, 1994; pp 253–
602
313.
603
(49)
Gorski, C. A.; Edwards, R.; Sander, M.; Hofstetter, T. B.; Stewart, S. M.
604
Thermodynamic Characterization of Iron Oxide–Aqueous Fe 2+ Redox Couples.
605
Environ. Sci. Technol. 2016, 50, 8538–8547.
606
(50)
Huynh, M. T.; Anson, C. W.; Cavell, A. C.; Stahl, S. S.; Hammes-Schiffer, S. Quinone
607
1 e – and 2 e – /2 H + Reduction Potentials: Identification and Analysis of Deviations
608
from Systematic Scaling Relationships. J. Am. Chem. Soc. 2016, 138, 15903–15910.
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Page 31 of 31
609
Environmental Science & Technology
(51)
Degradation in the Presence of Aquifer Material. Water Res. 1995, 29, 2353–2359.
610
611
Miller, C.; Valentine, R. L. Hydrogen Peroxide Decomposition and Quinoline
(52)
He, J.; Yang, X.; Men, B.; Bi, Z.; Pu, Y.; Wang, D. Heterogeneous Fenton Oxidation
612
of Catechol and 4-Chlorocatechol Catalyzed by Nano-Fe3O4: Role of the Interface.
613
Chem. Eng. J. 2014, 258, 433–441.
614
(53)
Pera-Titus, M.; Garcı́a-Molina, V.; Baños, M. A.; Giménez, J.; Esplugas, S.
615
Degradation of Chlorophenols by Means of Advanced Oxidation Processes: a General
616
Review. Appl. Catal. B Environ. 2004, 47, 219–256.
617
(54)
Pignatello, J. J.; Oliveros, E.; MacKay, A. Advanced Oxidation Processes for Organic
618
Contaminant Destruction Based on the Fenton Reaction and Related Chemistry. Crit.
619
Rev. Environ. Sci. Technol. 2006, 36, 1–84.
620
(55)
He, J.; Yang, X.; Men, B.; Wang, D. Interfacial Mechanisms of Heterogeneous Fenton
621
Reactions Catalyzed by Iron-Based Materials: A Review. J. Environ. Sci. 2016, 39, 97–
622
109.
623
(56)
in Model Subsurface Systems. J. Hazard. Mater. 1999, 69, 229–243.
624
625
(57)
Valentine, R. L.; Wang, H. C. A. Iron Oxide Surface Catalyzed Oxidation of Quinoline by Hydrogen Peroxide. J. Environ. Eng. 1998, 124, 31–38.
626
627
Watts, R. J.; Foget, M. K.; Kong, S. H.; Teel, A. L. Hydrogen Peroxide Decomposition
(58)
Kwan, W. P.; Voelker, B. M. Rates of Hydroxyl Radical Generation and Organic
628
Compound Oxidation in Mineral-Catalyzed Fenton-Like Systems. Environ. Sci.
629
Technol. 2003, 37, 1150–1158.
630
ACS Paragon Plus Environment
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