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A Mechanistic Understanding of Hydrogen Peroxide Decomposition by Vanadium Minerals for Diethyl Phthalate Degradation Guodong Fang, Yamei Deng, Min Huang, Dionysios D. Dionysiou, Cun Liu, and Dong-Mei Zhou Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b05303 • Publication Date (Web): 29 Jan 2018 Downloaded from http://pubs.acs.org on January 29, 2018

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Environmental Science & Technology

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A Mechanistic Understanding of Hydrogen Peroxide

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Decomposition by Vanadium Minerals for Diethyl Phthalate

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Degradation

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Guodong Fanga, Yamei Denga, Min Huanga, Dionysios D. Dionysioub, Cun Liua, *,

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Dongmei Zhoua,*

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a

Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, P.R. China

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b

Environmental Engineering and Science Program, Department of Chemical and

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Environmental Engineering (ChEE), University of Cincinnati, Cincinnati, Ohio 45221-

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0071, USA

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*

Corresponding author. Tel.: +86 25 86881180; fax: +86 25 86881180.

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E-mail address: [email protected] (C. Liu)

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[email protected] (D.M. Zhou).

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ACS Paragon Plus Environment


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Abstract

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The interaction of naturally occurring minerals with H2O2 affects the remediation

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efficiency of polluted sites in in-situ chemical oxidation (ISCO) treatments. However,

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interactions between vanadium (V) minerals and H2O2 have rarely been explored. In this

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study, H2O2 decomposition by various vanadium-containing minerals including V(III),

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V(IV) and V(V) oxides was examined and the mechanism of hydroxyl radicals (•OH)

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generation for contaminant degradation was studied. Vanadium minerals were found to

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catalyze H2O2 decomposition efficiently to produce •OH for diethyl phthalate (DEP)

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degradation in both aqueous solutions with a wide pH range and in soil slurry. Electron

29

paramagnetic resonance (EPR), X-ray photoelectron spectroscopy (XPS), X-ray

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diffraction (XRD) analyses and free radical quenching studies suggested that •OH was

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produced via single electron transfer from V(III)/V(IV) to H2O2 followed a Fenton-like

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pathway on the surface of V2O3 and VO2 particles, while the oxygen vacancy (OV) was

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mainly responsible for •OH formation on the surface of V2O5 particles. This study

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provides a new insight into the mechanism of interactions between vanadium minerals

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and H2O2 during H2O2-based ISCO.

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TOC Art

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Introduction

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Hydrogen peroxide (H2O2)-based in-situ chemical oxidation (ISCO) has been

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frequently used for the remediation of contaminated soil and groundwater.1-4 In this

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practice, when H2O2 solution is directly injected to soil, it undergoes decomposition by

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natural occurring minerals to produce hydroxyl radical (•OH), a strong reactive species

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and capable of degrading a variety of pollutants.5-8 Consequently, minerals-based Fenton-

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like reactions are fundamental processes of ISCO and have been studied extensively; 9-13

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such reactions can overcome some problems of homogenous Fenton reactions in the

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degradation of contaminants. 14-17 The kinetics and reaction mechanisms of some Fe and

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Mn-containing minerals, such as goethite, hematite, ferrihydrite, and iron-containing

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montmorillonite, to catalyze H2O2 decomposition have been well established.10 However,

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the decomposition of H2O2 by other soil abundant metal oxides, such as vanadium

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minerals, are rarely explored.

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As an important transition metal, vanadium (V) is a ubiquitous element in soil, with

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concentrations ranging from 0 to 400 mg/kg worldwide.18, 19 Increasing industrial demand

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for V has led to a rapid increase in V concentration in soil due to leaching from sources

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such as petroleum coke and emissions from fossil fuel combustion.20 ,21 V(IV) and V(V)

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are the dominant V species present in soil and most biological systems, and can be

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converted into V(III) under a reducing environment. 22, 23 Owing to these multiple

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oxidation states, V is widely used as a homogeneous and heterogeneous (supported)

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catalyst for the oxidation of alkanes and allylic alcohols, and oxidative bromination or

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sulfoxidation of organic compounds in the presence of alkyl hydroperoxides, H2O2, or

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O2.24–26 However, these reactions were conducted homogeneously in acetonitrile solution,



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while the catalytic decomposition of H2O2 by V minerals for contaminant degradation has

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not been reported so far. Furthermore, similar to the cycles of Fe(II)/Fe(III) (0.77 V) in

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the Fe-mineral/H2O2 system,10 the cycles of V(IV)/V(V) by H2O2 would be

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thermodynamic feasible due to their similar reduction potential (0.99 V).

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Therefore, this study aimed to (i) explore the underlying mechanism of H2O2

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decomposition by V minerals, and (ii) elucidate the pathways of •OH formation for

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contaminant degradation in the V/ H2O2 system. Diethyl phthalate (DEP) was chosen as a

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model pollutant because it is a typical phthalic acid ester (PAE) which is in the priority

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pollutant list of US Environmental Protection Agency (USEPA), and is frequently

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detected in soil and water.27 V2O3, VO2, and V2O5 particles were selected as model V

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minerals for elucidating the mechanism of the V-based Fenton-like process instead of

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naturally occurring V minerals because the latter, such as vanadinite, bravoite, and

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davidite, usually contain other transition metals, making it difficult to distinguish the role

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of V in H2O2 decomposition.28 Density functional theory (DFT) calculations and electron

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paramagnetic resonance (EPR) analysis were used to determine the detailed mechanism

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of DEP degradation and •OH formation in these processes.

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Materials and Methods

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Materials Chemicals used in this study and V mineral characterizations and related

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descriptions are presented in the Supporting Information (SI, Text S1). V mineral was

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characterized using X-ray diffraction (XRD), X-ray photoelectron spectroscopy (XPS),

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and EPR (Text S2).

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Experimental design All degradation experiments were carried out in the dark using a

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250-mL brown flask sealed with a plug and shielded by aluminum foil at 25 °C on an


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orbital shaker (120 rpm). The reaction solution pH was adjusted to desired values using

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concentrated NaOH and H2SO4 without buffer. The pH was monitored during the

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reaction, and adjusted to the initial pH when changes of more than 0.1 pH unit by an

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automatic pH control system used in our previous study.29 Briefly, the system consisted

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of automatic pH control equipment (CD-10, China), a peristaltic pump (HL-2, China),

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and a pH electrode (E201-C, China), NaOH or H2SO4 was automatically added when pH

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changed. All experiments were performed in triplicate to obtain average values with

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standard deviation.

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DEP degradation in aqueous solutions: Different amounts of V-mineral particles were

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added into a reaction solution (147 mL) containing DEP at pH 5.0. After mixing

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thoroughly, H2O2 solution (3 mL, 100 mM) was added to initiate the reaction. The final

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concentrations of DEP and H2O2 were 25 mg/L and 2.0 mM, respectively. Control

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experiments were also carried out using only V mineral particles or H2O2 under the

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identical conditions. Periodically, the reaction suspensions were filtered through a 0.22-

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µm PTFE membrane after adding quencher (ethanol) for HPLC analysis (Agilent 1200,

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USA). To determine DEP degradation intermediates, the reaction suspension was

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extracted with dichloromethane, derivatized with N,O-bis(trimethylsilyl)

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trifluoroacetamide (BSTFA)/trimethyl -chlorosilane (TMCS), and then analyzed by GC-

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MS (GC Varian CP3800/MS Saturn2200, USA). The H2O2 concentration was analyzed

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spectrophotometrically using the titanium sulfate method .12 V mineral particles were

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collected after reaction by freeze-drying and then analyzed using XRD, XPS, and EPR.

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DEP degradation in the soil slurry: To test the possibility that V mineral decomposed

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H2O2 for contaminant degradation in the realistic soil environment, we examined the



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effect of soil particles on DEP degradation in a heterogeneous soil-V2O3/H2O2 system.

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The properties of soil used in these experiments are described in the SI. To eliminate the

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effects of organic matter on H2O2 and •OH consumption, soil particles were first treated

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with H2O2 (30%) to remove organic matter.30 Environmentally relevant concentrations

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for DEP (60 mg/L), H2O2 (5.0 mM) and V2O3 (0.2 g/L) were chose in current study

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according to the concentration ranges that were commonly found in the actual

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contaminated soils and regular H2O2-based ISCO practices as reported in the previous

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studies.18, 31, 32 Soil particle loadings were 100, 200 and 500 g/L to examine the various

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soil to water ratios.

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EPR experiments: Hydroxyl radical formation in the V/H2O2 suspension was determined

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using the EPR method coupled with 5,5-dimethyl-1-pyrrolidine N-oxide (DMPO) as a

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spin-trapping agent. Briefly, the reaction suspension (V/H2O2) containing DMPO (0.1 M)

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was added to a quartz tube and determined by EPR (EMX 10/12, Bruker, Germany). EPR

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experiments were performed at room temperature, and parameters of EPR analysis for

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both aqueous and solid free radicals as well other analysis are presented in Text S3. The

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other analysis methods and DFT calculations used in this study are presented in Text S4.

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Results and Discussion

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Catalytic decomposition of H2O2 by V2O3 for DEP degradation

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Kinetics of DEP degradation in the V2O3/H2O2 system DEP degradation in the V2O3/

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H2O2 system as a function of H2O2 concentration was examined. Figure 1a shows that

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54.8% of DEP (25 mg/L) was degraded by 1.0 mM H2O2 in the presence of 0.1 g/L V2O3

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at pH 5.0 within 240 min, while a negligible amount of DEP was degraded with H2O2 or

130

V2O3 alone. By doubling H2O2 concentration to 2.0 mM, the DEP degradation efficiency


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increased to 64.0%, but decreased to 47.0% when the H2O2 concentration was further

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increased to 10 mM. The pseudo-first-order rate constants (kobs) of DEP degradation

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within 60 min were 0.0126 min–1 and 0.0117 min–1 for 1.0 mM and 2.0 mM H2O2,

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respectively, but decreased gradually to 0.0084 min–1 and 0.0054 min–1 when further

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increasing the H2O2 concentration to 5.0 and 10 mM (Figure S1a).

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Moreover, DEP was mainly degraded within 120 min, with 62.1% and 58.2% DEP

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degradation observed with 2.0 and 5.0 mM H2O2, respectively. Using 1.0 mM H2O2, DEP

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degradation occurred within 60 min, and then changed slightly when the reaction time

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was extended to 240 min, which was probably due to the rapid decomposition of H2O2

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within 60 min. Figure S1b shows that 1.0 and 2.0 mM H2O2 were almost completely

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decomposed by V2O3 within 60 and 120 min, respectively. However, the H2O2

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decomposition with initial concentrations at 5.0 and 10 mM, reached only up to 58.3%

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and 46.1%, respectively within 240 min. This behavior together was ascribed to the

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reaction between H2O2 and •OH to produce less reactive HOO• radical which was

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enhanced by excess of H2O2. This indicated that V2O3 activated H2O2 efficiently for DEP

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degradation, and that the rate of DEP degradation was H2O2-concentration dependent.

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Figure S2a shows that DEP was rapidly degraded by H2O2 (5.0 mM) with varied V2O3

148

loadings, and that the DEP degradation efficiency increased from 22.5% to 99.6%, and

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the corresponding kobs increased from 0.0045 to 0.0655 min–1 when V2O3 loading

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increased from 0.05 to 1.0 g/L (Figure S2b). In contrast, it was observed that only 6.3%

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of DEP (25 mg/L in total) was adsorbed by 1.0 g/L V2O3 in the absence of H2O2 (data not

152

shown). The combined results shown in Figures 1 and S2 suggested that the rate of DEP

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degradation was H2O2– and V2O3–concentration dependent, and not in direct proportion,



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and the optimum molar ratio of H2O2 and V2O3 was determined to be 1:0.6 in the

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degradation of 0.11 mM of DEP. Furthermore, changes in the total organic carbon (TOC)

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were monitored to evaluate the mineralization efficiency of DEP. As shown in Figure 1b,

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the TOC decreased markedly from 15.4 to 10.5 mgcarbon/L when increasing the V2O3

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loading from 0.05 to 1.0 g/L. The corresponding TOC removal increased from 3.1% to

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34.3% in the V2O3/H2O2 reaction suspension, while the TOC did not change in the

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reaction with DEP/H2O2 without V2O3. These results suggested that increasing the V2O3

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loading favored DEP degradation and mineralization in the V2O3/H2O2 system.

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GC-MS analysis of intermediates was conducted and the details are described in SI

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(Text S5, Figure S3). Briefly, the changes in the concentrations of the major

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intermediates, including MEP, PA, and m-OH-DEP followed a similar pattern with a

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marked increase within 60 min, and a swift decrease when the reaction extended to 240

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min (Figure 1c). These results were attributed to the accumulation of intermediates due to

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the initially rapid DEP degradation within 60 min (Figure 1a), and the delayed direct

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degradation of intermediates themselves after DEP was depleted. Furthermore, the mass

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balance of the total amount of products formed versus the amount of DEP degraded

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showed that the rate decreased from 81% to 11.2% by increasing the reaction time from

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30 to 240 min, which together with TOC analysis indicated that other ring-opening

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products were produced and further oxidized to CO2 and H2O (Figure 1b). In addition,

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the GC-MS analysis results were combined with quantum chemical calculations to

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identify the likely pathways of DEP degradation by •OH and the detailed discussion is

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presented in SI (Table S1, Text S5, Figures S3-S7).


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Identification of dominant reactive species in the V2O3/H2O2 system: Free radical

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quenching studies with ethanol as a quencher of •OH were used to identify the dominant

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reactive species for DEP degradation. Figure S8a shows that the addition of ethanol (200

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mM) completely inhibited DEP degradation by 0.2 g/L V2O3 and 5.0 mM H2O2,

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suggesting that •OH was the dominant radical responsible for DEP degradation in the

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V2O3/H2O2 system. EPR coupled with spin trapping agent DMPO was used to further

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determine the types of free radical produced at different H2O2 concentrations and V2O3

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loadings. Figure 2a shows that a significant EPR signal containing four lines with

184

intensity ratios of 1:2:2:1 was formed in the V2O3/H2O2 suspension, while insignificant

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signal was found in V2O3 suspension or H2O2 solution. The EPR signal was attributed to

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DMPO-OH adducts with hyperfine splitting constants of aN = aH = 14.9 G, in accordance

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with a previous study,33-35 suggesting •OH formation in V2O3/H2O2. Additionally,

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DMPO-OH peak intensity increased gradually from 0.7 to 16.1 (104 a.u.) with increasing

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V2O3 dosage from 0.05 g/L to 1.0 g/L, which indicated that the •OH concentration

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increased by larger V2O3 loadings (Figure 2b). EPR spectra of the V2O3/H2O2 system

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with different H2O2 concentrations showed a decrease in DMPO-OH peak intensity from

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3.6 to 2.1 (104 a.u.) with increasing H2O2 concentration from 1.0 to 10 mM (Figures S8b

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and 8c). This indicated that excess H2O2 scavenged some of the •OH, which inhibited the

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trapping reaction between DMPO and •OH.

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Effects of pH on DEP degradation in the V2O3/H2O2 system

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Figure 3a shows that the DEP degradation efficiency decreased from 70.1% to 56.4%

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when increasing the pH from 3.0 to 9.0, and the corresponding kobs decreased from

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0.0165 to 0.0101 min–1, which showed that V2O3/H2O2 can degrade DEP effectively over



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a wide pH range, although the rate of DEP degradation decreased with increasing

200

reaction pH (Figure S9a). The concentration of V ion released from V2O3/H2O2 during

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DEP degradation was also examined. Figure S9b shows that the concentration of

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dissolved V ions from V2O3 during the reactions increased from 0.98 to 5.73 mg/L when

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the V2O3 loading was increased from 0.05 to 1.0 g/L in the presence of 2.0 mM H2O2 at

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pH 5.0 for 240 min. However, the corresponding ratio of V dissolution ([V]dissolved/

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[V]total×100%) decreased from 2.87% to 0.84%, suggesting that increasing V2O3 loading

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markedly reduced the rate of V ion dissolution. Figure 3b shows that the V ion

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concentration decreased from 3.12 to 0.13 mg/L when increasing the pH from 3.0 to 9.0

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(V2O3, 0.1 g/L), which indicated that the V ion dissolution was greatly reduced with

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increasing reaction pH in the V2O3/H2O2 system.

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Furthermore, it was observed that V(IV) was the dominant dissolved V species, with

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its concentration increasing from 0.62 to 4.64 mg/L as the V2O3 loading was increased

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from 0.05 to 1.0 g/L, which was markedly higher than the V(V) concentration (0.17 to

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0.49 mg/L) (Figure S9c). To further test the contribution of V(IV) ions to DEP

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degradation via homogeneous Fenton-like reactions, DEP degradation in the V(IV)/H2O2

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solution system was examined. The concentration of V(IV) used in this experiment

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ranged from 0.01 to 0.1 mM, which was identical to the concentration range of V(IV)

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dissolved from V2O3/H2O2. Figure S10 shows that less than 15% of DEP (25 mg/L) was

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degraded by 5.0 mM H2O2 in the presence of different V(IV) ion concentrations (0.01–

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0.1 mM), indicating that the homogenous Fenton reaction made a limited contribution to

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DEP degradation in the V2O3/H2O2 system.

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Catalytic decomposition of H2O2 by VO2 and V2O5 for DEP degradation


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As described above, the V2O3/H2O2 system was validated for efficient degradation of

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contaminant, however, V(V) and V(IV) are usually the dominant V species in soil.19

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Therefore, the catalytic decomposition of H2O2 by VO2 and V2O5 for DEP degradation

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were also examined. As shown in Figure S11, 48.8% and 54.1% of DEP was degraded by

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0.5 g/L VO2 in the presence of 1.0 and 2.0 mM H2O2, respectively, with the

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corresponding kobs increasing from 0.0033 to 0.0047 min–1 (Figure S11c). EPR results

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showed a strong DMPO-OH signal observed in the VO2/H2O2 system and the peak

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intensity gradually decreased as the H2O2 concentration increased from 0.5 to 10 mM,

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indicating that •OH was formed, but rapidly consumed by excess H2O2. The change in

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DMPO-OH intensity was consistent with DEP degradation kinetics (Figure S11d).

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Similar results were observed in the V2O5/H2O2 system. V2O5 can also catalyze the

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decomposition of H2O2 to form •OH and degrade DEP with higher efficiency (Figure

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S12).

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To further evaluate the efficiency of V mineral for catalytically decomposing H2O2,

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H2O2 (2.0 mM) decomposition rates by V2O3, VO2 and V2O5 (0.1 g/L) at pH 7.0 (50 mM

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borate buffer) were examined and the comparison was made with decomposition rates of

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H2O2 by Fe minerals reported in previous studies.12, 15 Figure S13 shows that H2O2 were

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decomposed efficiently by various V minerals. The rates of H2O2 decomposition (kd, H2O2)

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were 1.41, 0.2 and 0.07 h-1 for V2O3, VO2 and V2O5, respectively. Surface area is an

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important factor that influences heterogeneous H2O2 decomposition,15 and thus kd, H2O2

242

was normalized to the surface area of minerals in the reaction suspension, and denoted as

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kSA. Table S3 displays the kSA values of V minerals in this study and Fe minerals

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collected from previous studies.12, 15 The results suggest that kSA of V2O3 was 3–4 orders



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of magnitude larger than that of Fe minerals. For the soil dominant V species such as

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VO2 and V2O5, their kSA values were 136.5–2087, 9.8–150.5 times larger than those of Fe

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oxides such as goethite, hematite and FeOOH, respectively. These results suggest that V

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minerals were more efficient than Fe minerals in catalyzing H2O2 decomposition.

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Proposed •OH formation pathways in the V/H2O2 system

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Mechanism of H2O2 activation by V2O3 and VO2 particles: Similar to the Fe-mineral

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based Fenton-like process ,10 it was hypothesized that single electron transfer from

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surface V(III)( ≡V(III)) or V(IV) (≡V(IV)) to H2O2 was the dominant process for •OH

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generation (Eq. 1 and 2), and a chain reaction can be propagated as V(III) or V(IV) can

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be regenerated by H2O2 (Eq. 3 and 4).

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≡ V(III) + H 2 O 2 →≡ V(IV)+ • OH + - OH

(1)

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≡ V(IV) + H 2 O 2 →≡ V(V)+ •OH + - OH

(2)

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≡ V(IV) + H 2 O 2 →≡ V(III)+ HO•2

(3)

≡ V(V) + H 2 O 2 →≡ V(IV)+ HO•2

(4)

258 259

To clarify these processes, XPS, XRD, and EPR techniques were used to characterize

260

V-oxides before and after reaction in the V/H2O2 system. Figure S14 shows phase

261

transformations in the V minerals before and after reaction. For V2O3, only the V2O3

262

phase was observed before the reaction, but a new V6O13 phase, which was a mixture of

263

V(IV) and V(V) minerals, appeared after the reaction with H2O2, indicating that V(III)

264

transferred electrons to H2O2 to produce V(IV) and V(V) (Eq.1 and 2). Similarly, V6O13

265

was formed by VO2 after the reaction, which suggested that V6O13 was the dominant

266

mineral formed in the reaction of V2O3 or VO2 with H2O2.


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XPS analysis showed that V(IV) and V(V) were the dominant V species on the V2O3

268

surface before reaction, which was ascribed to the partial oxidation of V(III) by oxygen

269

(Figure S15). After reacting with H2O2, the ratio of V(IV) decreased from 22.2% to

270

15.4%, while the ratio of V(V) increased from 77.8% to 84.6% (Table S2). Similarly,

271

V(IV) and V(V) were the dominant V species on the VO2 surface, with the ratio of V(IV)

272

decreasing after the reaction with H2O2. These results further indicated that electron

273

transfer from surface V(III)/V(IV) to H2O2 was the dominant process in H2O2 activation,

274

which was consistent with XRD results. EPR analysis showed an eight-component

275

hyperfine structure for both VO2 and V2O3 bulk particles, which was characteristic of a

276

V(IV) center ion, according to a previous study (Figure S16).36 Moreover, the peak

277

intensities of V(IV) increased for V2O3, but decreased for VO2 after the reaction, which

278

further confirmed electron transfer from V(III) and V(IV) to H2O2 during the reaction and

279

was consistent with XRD and XPS results.

280

Mechanism of H2O2 activation by V2O5 particles: Catalytic decomposition of H2O2 by

281

V2O5 has been proposed to exhibit a similar mechanism to other Fenton-like processes,15

282

with V(IV) regenerated on V2O5 surface (Eq. 4) activating H2O2 to produce •OH. XRD

283

results showed that the V2O5 phase changed slightly before and after the reaction,

284

suggesting that V2O5 was a stable catalyst toward H2O2 activation for DEP degradation.

285

However, interestingly, 7.5% V(IV) was formed on the surface of V2O5 particles before

286

the reaction according to XPS analysis (Table S2). This behavior was due to vanadyl

287

oxygen being easily desorbed from the V2O5 lattice during V2O5 particle synthesis, which

288

resulted in the formation of oxygen vacancies and V(IV) center ions.37 As each oxygen

289

vacancy can leave an electron at the vanadium ion site, reducing V(V) to V(IV),38 both



290

oxygen vacancies (OVs) and related V(IV) species play an important role in the catalytic

291

activity of V2O5 particles. Furthermore, an OV is generally considered a basic defect in

292

the V2O5 lattice and, as such, defects usually contain an unpaired electron and can be

293

detected by EPR. However, V(IV) has a wide range of EPR signals (3200–3700 G)

294

overlapping with the OV EPR signal at 3400 G. Consequently, the existence of OV was

295

not detected by EPR in current study. To determine the role of V(IV) regeneration on the particle surface in H2O2 activation,

296 297

the kobs of DEP degradation was normalized to V2O5 and VO2 loadings, giving

298

normalized values of 0.02 and 0.009 min–1·L g-1, respectively. Unexpectedly, the results

299

indicated the V2O5 was more efficient than VO2 toward H2O2 activation for DEP

300

degradation. Theoretically, the catalytic ability of VO2 should be significantly higher than

301

that of V2O5, because VO2 contains more V(IV) and easily transfers electrons to H2O2.

302

However, XPS and EPR analysis showed that the content of V(IV) on the V2O5 surface

303

and in bulk V2O5 particles was increased instead of decreased due to the consumption by

304

H2O2. These combined experimental results showed that the catalytic decomposition of

305

H2O2 by the regenerated V(IV) on the surface of V2O5 was not the only factor controlling

306

•

OH formation for DEP degradation.

307

OVs have been reported to play an important role in •OH formation in other systems,

308

such as BiOCl and CeO2 nanoparticle-based Fenton-like processes.39, 40 Therefore, it was

309

hypothesized that electron transfer from the OV to H2O2, producing bound-OH, was

310

another important source of •OH for DEP degradation. To verify OV-induced •OH

311

formation on the V2O5 surface, we added F– to the V2O5/H2O2 suspension to desorb OV-

312

bound •OH into solution.41 Figure 4a shows that the addition of F– significantly enhanced


Page 14 of 28

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313

the DMPO-OH EPR signal in the V2O5/H2O2 system, with the peak intensities increasing

314

2.3-fold, which suggested significant amount of surface-bound •OH was released into

315

solution. In contrast, the peak intensities of DMPO-OH were only increased by about 30%

316

in the VO2/H2O2 system with F–, suggesting that surface-bound OH accounted for only a

317

small proportion of total •OH on VO2 surface (Figure S17). These results indicated that

318

surface-bound OH produced from ≡V(IV) (bound metal) and H2O2 was easily trapped by

319

DMPO, leading to a significant EPR signal in aqueous solution. However, OV-bound

320

•

321

not be desorbed into aqueous solution or even trapped by DMPO, suggesting that OV-

322

bound-OH was the dominant radical species in the V2O5/H2O2 system.

323

OH formed from OV and H2O2 was bound so strongly to the V2O5 surface that it could

V2O5 single crystals or nanoparticles usually contain a large amount of OVs that can be

324

detected by EPR.42 To further testify the role of OVs in H2O2 activation, V2O5

325

nanoparticles (V2O5-N) were synthesized. XRD and TEM analysis showed that the

326

prepared V2O5 particles were crystalline with average diameter of 90 nm (Figure S17).

327

EPR results showed a strong singlet peak for V2O5-N, which was characteristic of an OV

328

with a g-factor of 2.0031, and confirmed that OVs were formed on the V2O5-N surface

329

(Figure 4b). Meanwhile, the EPR signal for V(IV) was also observed, although its peak

330

intensity was relatively low, which suggested that V(IV) formation was accompanied by

331

OV production (Figure S18). Furthermore, the OV peak intensities decreased markedly in

332

the presence of H2O2, decreasing by 32.1% when increasing the H2O2 concentration to 10

333

mM within 60 min. Formation of a DMPO-OH signal was also observed, with its peak

334

intensities increasing 7.43-fold in the presence of F– (Figure 4c), which was significantly

335

higher (2.3-fold) than that of the purchased micrometer sized V2O5 particles used in



336

current study. These results indicated that accompanied by OVs disappeared, electron

337

was transferred from OV sites to H2O2, and confirmed that OVs played a key role in the

338

formation of surface-bound OH in the V2O5/H2O2 system.

339

Proposed framework for H2O2 decomposition and •OH formation in the V/H2O2 system

340

According to the results above, the mechanism of catalytic decomposition of H2O2 with

341

different V minerals is discussed. The pathways of •OH formation in the V/H2O2 system

342

are proposed in Figure 5. For V2O3 and VO2 particles, single-electron transfer from

343

≡V(III)/≡V(IV) to H2O2 produces •OH on the particle surface, and the formed ≡V(V) is

344

further reduced by H2O2 to regenerate ≡V(IV). The cycle of V(IV)/V(V) by H2O2 (Eq. 4)

345

is thermodynamic feasible due to the relatively high potential (0.99 V), which was easier

346

than Fe(II)/Fe(II) cycled by H2O2 as verified in the previous study.10 However, for V2O5,

347

the OV on the particle surface transfers an electron to H2O2, inducing the formation of

348

OV-bound •OH, which was a dominant species for contaminant degradation.

349

Decomposition of H2O2 by V2O3 for DEP degradation in soil slurry

350

Because V minerals are ubiquitously distributed in soil environments in considerable

351

concentration; it is anticipated they can decompose H2O2 for contaminant degradation in

352

the soil. To test this hypothesis, the effect of soil particles on the degradation of DEP in

353

the soil-V2O3/H2O2 system was examined. Figure 6a shows that DEP degradation was

354

inhibited in the presence of soil particles, and its degradation efficiency decreased from

355

60.1% to 44.8% in V2O3/H2O2 with 100 g/L of soil particles. When increasing soil

356

particle loading from 100 to 500 g/L (soil/water ratio, 1:2), the DEP degradation

357

efficiency decreased from 44.8% to 24.2%. Adsorption on soil particles contributed to

358

less than 10% of DEP loss in the solution phase. Figure 6b shows that the concentration


Page 16 of 28

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359

of V ion dissolution from V2O3 decreased dramatically in the presence of soil particles.

360

The V ion concentration decreased from 2.7 to 0.2 mg/L when increasing the soil particle

361

loading from 0 to 200 g/L, and decreased below the detection limit when the soil particle

362

loading was further increased to 500 g/L, suggesting that the V ion leaching was greatly

363

reduced by soil particles. The reduction of V ion concentration in the solution was

364

attributed to the strong adsorption of V ion by Fe oxides or clay minerals in the soil. This

365

would result in the formation new V minerals on the surface of soil particles.19 These

366

results together suggest that DEP was efficiently degraded in the soil slurry, although its

367

degradation rates were noticeably reduced. More importantly, the adsorption of V ion by

368

soil particles, and subsequently formation of new V minerals would greatly reduce the

369

potential risk of V ion leaching to ground and surface water.

370

Environmental Implications Despite the fact that geochemical parameters, such as

371

minerals and humic substances in the subsurface environment, have important effects on

372

the efficiency of soil and groundwater remediation by H2O2-ISCO processes, interactions

373

between the soil abundant vanadium minerals and H2O2 have rarely been investigated.

374

This study provides a detailed mechanism of H2O2 decomposition and •OH formation in

375

the V/H2O2 system. In V2O3 and VO2-based Fenton-like processes, •OH was produced

376

via single electron transfer from V(III)/V(IV) to H2O2. However, in the V2O5/H2O2

377

system, OVs played an important role in •OH formation, with transfer of an OV electron

378

to H2O2 yielding OV-bound •OH. This process was testified by reacting the synthesized

379

OV-rich V2O5 nanoparticles with H2O2 to further verify the role of OVs, with EPR

380

analysis showing that OV was consumed during the reaction, accompanied by •OH

381

formation. This study provided a new insight into the mechanism of interactions between



382

vanadium minerals and H2O2 when using H2O2 for ISCO remediation of contaminated

383

soil and groundwater.

384

Additionally, this study suggested that V mineral can efficiently catalyze the

385

decomposition of H2O2 for contaminant degradation under environmentally relevant

386

concentration. For example, the soil slurry experiments were conducted with 0.2 g/L

387

V2O3 (127 mg/kg V in soil), which was in the range of V background content (0-400

388

g/kg18). Hence, H2O2 may be decomposed by V minerals naturally for contaminant

389

degradation in the H2O2-based ISCO process. Particularly for soil near vanadium

390

titanomagnetite mining sites, the V concentration is usually high (∼940 mg/kg in the

391

Panzhihua region of China), and coexists with other organic contaminants .21 Therefore,

392

these organic contaminants are degraded by adding H2O2 without adding other catalysts.

393

Supporting Information Available: XRD, XPS and EPR analysis of vanadium minerals

394

before and after reaction, DFT calculations and detailed discussion on the pathways of

395

DEP degradation can be found in this section (Text S1-S5, Tables S1-3, and Figures S1-

396

18). This material is available free of charge via the Internet at http://pubs.acs.org/.

397

Acknowledgments This work was supported by grants from the National Key Research

398

and Development Program of China (2017YFA0207001, 2016YFD0800204), the

399

National Natural Science Foundation of China (41671478, 41671239), the Natural

400

Science Foundation of Jiangsu Province of China (BK20170050), the 135 Program of

401

Institute of Soil Science (ISSASIP1660) and Youth Innovation Promotion Association of

402

CAS (2014270).


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526 527



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Figure 1

530

(b)

531

(a) 532 533 534 535 536 537 538 539

(c)

540 541 542 543 544 545 546 547 548 549 550 551

Figure 1 Kinetics of DEP degradation as a function of H2O2 concentration and V2O3 loading in the V2O3/H2O2 system: (a) DEP degradation; (b) changes in TOC and TOC removal during reaction; (c) concentration profiles of DEP degradation intermediates and mass balance of products (2.0 mM H2O2 and 0.1 g/L V2O3). Reaction conditions: V2O3 loading = 0.05–1.0 g/L; [H2O2]0 = 1.0–10 mM; [DEP]0 = 25 mg/L; pH 5.0; and 25 °C.

552 553 554 555


Page 23 of 28

556


Figure 2

557 558

(a) 559 560 561 562 563 564 565 566 567

(b)

568 569 570 571 572 573 574 575 576 577

Figure 2 Identification of reactive species in the V2O3/H2O2 system using EPR methods: (a) EPR spectrum; (b) changes in peak intensities of DMPO-OH signal as a function of V2O3 loadings at 5 min. Reaction conditions: V2O3 loading = 0.05–1.0 g/L; [H2O2]0 = 2.0 mM; [DMPO]0 = 100 mM and pH 5.0.

578 579 580 581



582

Figure 3

583 584

(a) 585 586 587 588 589 590 591 592 593

(b) 594 595 596 597 598 599 600 601 602

Figure 3 Effect of pH on DEP degradation and V ion dissolution in the V2O3/H2O2 system: (a) degradation kinetics; (b) concentration of V ion dissolution at various pH values. Reaction conditions: V2O3 loading = 0.1 g/L; [H2O2]0 = 2.0 mM; [DEP]0 = 25 mg/L; and 25 °C for 4 h.

603 604 605 606 607


Page 24 of 28

Page 25 of 28

608


Figure 4

609 610

(a)

(b)

611 612 613 614 615 616 617 618 619

(c)

620 621 622 623 624 625 626 627 628 629 630 631 632

Figure 4 EPR spectrum of DMPO-OH formed from V/H2O2 in the presence and absence of NaF at 5 min: (a) V2O5/H2O2; (b) EPR spectrum of oxygen vacancies in V2O5 nanoparticles (nano-V2O5) before and after reaction; (c) EPR spectrum of DMPO-OH formed from nano-V2O5/H2O2 (2.0 mM) in the presence and absence of NaF at 5 min. Reaction conditions: V2O5/nano-V2O5 loadings = 0.1 g/L; [H2O2]0 = 2.0–10 mM; [DMPO]0 = 100 mM; [NaF]0 = 2.0 mM, and pH 5.0.

633 634 635



636

Figure 5

637 638 639 640

Figure 5 Proposed pathways for H2O2 decomposition and hydroxyl radical generation in the V/H2O2 system.

641 642 643 644 645 646 647 648 649 650 651 652 653 654


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655


Figure 6

656 657

(a)

658 659 660 661 662 663 664 665

(b) 666 667 668 669 670 671 672 673 674 675 676

Figure 6 Effect of soil particles on DEP degradation in the V2O3/H2O2 system: (a) DEP degradation kinetics in the soil slurry; (b) concentration of V ion dissolution with different particle loadings. Reaction conditions: V2O3 loading = 0.2 g/L; soil particle loading = 100–500 g/L; [H2O2]0 = 5.0 mM; [DEP]0 = 60 mg/L; pH 5.0; 25 °C; and 4 h.

677




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