Insight into mechanism of arsanilic acid degradation in permanganate-sulfite system: Role of reactive species

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Insight into mechanism of arsanilic acid degradation in permanganatesulfite system: Role of reactive species ⁎

Zhenyu Shia,b, Can Jinc, Jing Zhangd, , Liang Zhua,

⁎

a

College of Environment, Hohai University, Nanjing 210098, PR China State Environmental Protection Key Laboratory of Monitoring and Analysis for Organic Pollutants in Surface Water, Environment Monitoring Center of Jiangsu Province, Nanjing 210036, PR China c Institute of Chemical Industry of Forest Products, Chinese Academy of Forestry, Key Laboratory of Biomass Energy and Material of Jiangsu Province, Nanjing 210042, PR China d Key Laboratory of the Three Gorges Reservoir Region’s Eco-Environment, Ministry of Education, School of Urban Construction and Environmental Engineering, Chongqing University, Chongqing 400045, PR China b

H I GH L IG H T S

G R A P H I C A L A B S T R A C T

acid was efficiently removed • Arsanilic in 15 s by Mn(VII)-S(IV) system. Reactive species in Mn -S process • were , OH, and supposed to be SO VII

4

• •

IV

%− %

Mn intermediates. Contribution from each reactive species was related to structures of organic contaminants. Background water matrices influence the performance of Mn(VII)-S(IV) process.

A R T I C LE I N FO

A B S T R A C T

Keywords: Permanganate Sulfite Sulfate radical Advanced oxidation process Arsanilic acid

This work investigated the fast degradation of ASA by a novel advanced oxidation process (permanganate (MnVII)-sulfite (SIV)). The results showed that the combination of Mn(VII) (50 μM) and S(IV) (250 μM) at pH 5.0 triggered near-instantaneous decomposition of ∼71% ASA (5 μM) within 15 s. The reaction parameters, which affected the degradation of ASA, such as pH (4.0–9.0), initial concentration of ASA (1–30 μM), and molar ratio of S(IV)/Mn(VII) (1–20), were systematically investigated. Specifically, sulfate radical (SO4%−), hydroxyl radical (%OH), and Mn intermediates were considered as the important reactive species in Mn(VII)-S(IV) process. Radical scavenging tests showed that both SO4%− and %OH contributed to the removal of ASA, with SO4%− playing a dominant role. The contributions of reactive radicals, which were excluded in Mn(VI)-S(IV) system by previous researchers, for the degradation of organic contaminants was clarified for the first time. Therefore, our study can be considered as a necessary complement to previous studies about Mn(VII)-S(IV) system. Halide ions inhibited the removal of ASA following a trend of F− < Cl− < Br− < I− due to their competition for reactive radicals with ASA. The degradation products of ASA were identified, and the plausible reaction pathways were proposed, and their toxicity was assessed using luminescent bacteria Vibrio fischeri. Finally, the satisfying performance of Mn(VII)-S(IV) system in a natural water sample indicated it was a promising method for degrading organic contaminants in real water.

⁎

Corresponding authors. E-mail addresses: [email protected] (J. Zhang), [email protected] (L. Zhu).

https://doi.org/10.1016/j.cej.2018.11.030 Received 18 August 2018; Received in revised form 20 October 2018; Accepted 3 November 2018 1385-8947/ © 2018 Elsevier B.V. All rights reserved.

Please cite this article as: Shi, Z., Chemical Engineering Journal, https://doi.org/10.1016/j.cej.2018.11.030

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1. Introduction

triclosan. Gao et al. [34] demonstrated that manganese intermediates (e.g. Mn(VI), Mn(V), and Mn(IV)) other than Mn(III) formed in triclosan oxidation by Mn(VII) was stabilized by various ligands. In Mn(VII)-S (IV) process, sulfite was also considered as a complexing ligand, which may stabilize manganese intermediates [35]. The ligands might competed with S(IV) for Mn intermediates and influence the removal of organics in Mn(VII)-S(IV) system. Additionally, several studies reported that halide ions quench not only SO4%− but also %OH [17,22,25]. Therefore, it is necessary to investigate the influence of the constituents of water on the degradation of contaminants by Mn(VII)-S(IV) system. Therefore, we performed series of batch experiments to (i) investigate the degradation of ASA in Mn(VII)-S(IV) system under various conditions (pH, molar ratio of S(IV) to Mn(VII), initial ASA concentration, dissolved oxygen concentration); (ii) identify the reactive species; (iii) explore the effect of co-existing solutes (humic acid, ligands, halide, and metal ions) on the degradation of ASA; and (iv) examine the transformation products and residual toxicity of ASA after treatment in Mn(VII)-S(IV) system.

In the past decades, as a typical organoarsenic compound, 4-aminophenylarsenic acid (arsanilic acid, ASA) had been widely used as feed additives in poultry and swine production to promote growth, improve feed efficiency, and inhibit parasite infections [1,2]. Such additive is minimally metabolized within animals and is mostly excreted with no chemical structural change [3]. Specifically, they are inevitably introduced into the environment when poultry manure is applied to farmland as fertilizer. Although organoarsenic additives have relatively low toxicity, their transformation products involving inorganic arsenite (AsIII) and arsenate (AsV) significantly increased risk to the environment and human [4,5]. Therefore, the removal of organoarsenic prior to its entering environments is of crucial importance to reduce the potential risk. A variety of techniques such as Fenton process [6], adsorption [7–9], and photocatalytic degradation [10–12] have been developed to remove organoarsenic compounds in water. Recently, novel sulfite based advanced oxidation processes (S-AOPs) have attracted increasing attention, due to their cost-effectiveness and environmental-friendliness [13–16]. Sulfite, a common industrial contaminant in wastewaters or in exhaust gas, acts as an efficient source of reactive radicals. For instance, Chen et al. [17] demonstrated that 80% of Orange II was efficiently decolorized within 80 min in Fe(II)-sulfite system. Jiang et al. [18] reported that Cr(VI)-sulfite system could rapidly transform approximately 60% of As(III) to As(V) within 40 min at pH 3.5. Zhang et al. [19] found that Fe(VI)-sulfite system decomposed various of organic pollutants, including benzotriazole, phenol, and ciprofloxacin. Therefore, application of industrially available sulfite as a new source of sulfate radicals on the degradation of organoarsenic compounds seems promising. Generally, the reactive species (SO4%− and %OH) in the S-AOPs take place as follows: firstly, sulfite was oxidized to SO3%− by Fe(III) [20], Fe(VI) [21,22], and Cr(VI) [23–25], then SO3%− rapidly reacted with oxygen to form SO5%−, which could react with sulfite to generate SO4%− (Eqs. (1)–(3)) [26]. Meantime, SO4%− may be transformed to % OH through reaction with water (Eq. (4)) [22,26].

oxidant + HSO−3 /SO32 − → SO·3−

(1)

SO·3− + O2 → SO·5−

(2)

SO·5− + HSO−3 /SO32 − → SO·4− + SO24− (+H+)

(3)

SO·4− + H2 O→ SO24− +· OH + H+

(4)

2. Materials and methods 2.1. Chemical reagents Arsanilic acid (ASA), bisphenol A (BPA), 5,5-dimethyl-pyroline-Noxide (DMPO, 97%), and nitrobenzene (NB) were purchased from J&K Scientific Corporation. Potassium permanganate, anhydrous sodium sulfite (Na2SO3), hydroxylamine hydrochloride, sodium pyrophosphate (PP), ethylenediaminetetraacetic acid (EDTA), coumarin, and tert-butanol (TBA) were obtained from Sinopharm Chemical Reagent Co. Ltd. Manganese(III) acetate dihydrate was obtained from Sigma-Aldrich. These chemicals were of analytical grade, and used without further purification. Dichloromethane, methanol (MeOH), acetonitrile, and acetic acid of HPLC grade were supplied by Merck. All solutions were prepared with Milli-Q water. The stock solution of sodium sulfite (25 mM) was freshly prepared and quickly used. 2.2. Reaction procedures All experiments were conducted in a brown glass bottle with magnetic stirring at 20 ± 1 °C with a circulating water jacket. The pH of solution was adjusted to desired values with H2SO4 (0.1 M) and NaOH (0.1 M) solution. Reactions were initiated by adding Mn(VII) into solutions containing ASA and sulfite. Samples were withdrawn at specified time intervals and quenched by hydroxylamine (10 mM) to monitor the residual ASA. Since the degradation of ASA would finish in 15 s in Mn(VII)-S(IV) system, the collection of the first sample was done quickly. After the solution mixed for 10 s, the first sample was withdrawn with a syringe and at 15 s the sample in the syringe was push out rapidly into a beaker containing hydroxylamine to quench the reaction. Hydroxylamine was an effective quencher for eliminating all the reactive species in Mn(VII)-S(IV) system (Fig. S1). All batch experiments were conducted in duplicates or triplicates, and the average data and their standard deviations were obtained and shown in figures. To evaluate the effect of dissolved oxygen, we examined the performance of Mn(VII)-S(IV) system in the presence or absence of dissolved oxygen. The reaction solution containing ASA was purged with nitrogen gas (0.8 mL/min) for 60 min to remove the dissolved oxygen. Then, the reaction was started with adding S(IV) and Mn(VII) simultaneously. During the whole reaction, the solution mixture was continuously bubbled with nitrogen gas. The stock solution of Mn(III)-PP (pyrophosphate) was prepared following the reported literature [31]. In brief, PP was dissolved in water to reach a final concentration of 50 mM PP and pH was adjusted to 8.2, then manganese(III) acetate was slowly added to the PP solution during vigorous stirring to reach a final concentration of 10 mM as Mn (III). Finally, the pH of Mn(III)-PP solution was adjusted to 7.0.

However, Sun et al. [27–29] demonstrated that the rapid oxidation of various contaminants in the Mn(VII)-sulfite process was ascribed to the generation of highly reactive Mn(III). It is well known that the principal reaction pathway of organic pollutants oxidation by Mn(III) is controlled through one electron transfer mechanism [30–32]. Jiang et al. [33] proved that Mn(III) exhibited no reactivity towards carbamazepine for which two-electron oxygen donation may be the primary oxidation pathway. But Sun et al. [29] observed that approximately 90% of carbamazepine was rapidly removed in Mn(VII)-S(IV) process with Mn(III) as the main reactive specie. Thus, the contradictory results indicated the primary mechanism for Mn(VII)-S(IV) system requires further study, especially the real reactive species and their roles. Additionally, the transformation of ASA by Mn(VII)-S(IV) from aquatic environments may be affected by various environmental factors, such as pH, and co-existing constituents. The co-existing constituents include humic acid, ligands, halides, and metal ions, which are presenting extensively in aquatic ecosystems. Metal ions influence the oxidation of S(IV) by oxygen [36–38]. Furthermore, the ligands in water may impact the Mn species and its reactivity. For instance, Jiang et al. [33] reported several ligands, such as phosphate, pyrophosphate, EDTA, and humic acid, stabilized the reactive manganese intermediates formed upon Mn(VII) reduction, resulting in the enhanced removal of 2

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Fig. 1. Effect of molar ratio of sulfur(IV) to Mn(VII) (a) and initial ASA concentration (b) on the ASA removal in Mn(VII)-S(IV) system. Reaction conditions: (a) [ASA]0 = 5 μM, [Mn(VII)]0 = 50 μM, reaction time 15 s, and T = 20 °C; (b) [S(IV)]0 = 250 μM, pH = 5.0, the other reaction conditions are same as (a).

assess the toxicity of ASA solution (5 μM) by the combination of 50 μM Mn(VII) and 250 μM S(IV) during the reaction course. Freeze-dried Vibrio fischeri bacteria were purchased from Modern Water (USA). The acute toxicity test was performed according to ISO standard method 11348-3 [42]. Prior to toxicity assessment, Vibrio fischeri bacteria were reconstituted in 1.0 mL 3% NaCl solution and stored in the ice water bath. 0.2 mL of water sample withdrawn at specified time intervals, and 10 μL reactivated bacteria was added to 2 mL 3% NaCl solution and luminescence was measured by Microtox Model 500 (USA) after 15 min of incubation at 23 ± 1 °C. The inhibition of bioluminescence was calculated as follows (L represents the luminescence intensity) (Eq. (6)):

2.3. Analytical methods The concentration of ASA was determined by high-performance liquid chromatography (HPLC, Waters 2695, Waters, USA) equipped with a Waters XBridge C18 column (4.6 mm × 250 mm, 5 μm) and a Waters 2698 PDA detector at wavelength of 260 nm [39]. The mobile phase consisted of 1% acetic acid: methanol (97:3, v/v) with a flow rate of 1.0 mL/min. Electron spin resonance (ESR) experiments were performed using DMPO as a spin-trapping agent, and the detailed parameters and procedures are shown in Text S1. Coumarin (1 mM) was employed as a chemical probe for %OH (Eq. (5)) [40]. 7-Hydroxycoumarin was measured by monitoring the fluorescence emission at 460 nm under excitation at 332 nm using a spectrofluorometer (F-7000, Hitachi High Technologies, Japan). %

OH + coumarin → 7-hydroxycoumarin k = 2.0 × 109 M−1 s−1

Inhibition rate (%) =

Lblank − Lsample Lblank

× 100%

(6)

(5) 3. Results and discussion

Identification of intermediates was conducted by a 6890N gas chromatography instrument (Agilent, USA) equipped with a mass spectrometer detector MSD 5975C and a 7683B automatic liquid sampler. The pre-treatment process was as follows: 100 mL reaction solution (adjusted pH to 2.0 with 10% H2SO4) was extracted with 25 mL dichloromethane for three times, and the extracted solution was dehydrated using anhydrous sodium sulfate. The extract was further concentrated to 1.0 mL by rotary evaporation. Before injection, derivatization was carried out using 0.5 mL of BSTFA-TMCS (99:1) at 50 °C for 30 min. The final sample (1.0 μL) was automatically injected into GC with CP-8 MS capillary column (30 m, 0.25-mm inner diameter, 0.25μm film thickness, Varian, USA) using splitless mode. The MS was operated with 70 eV electron impact (EI) mode and positive ion mode. The different arsenic species, i.e., As(III) and As(V), in degradation products were quantified by HPLC coupled hydride generation atomic fluorescence spectrometry (HPLC-HG-AFS, SA-20, Beijing Titan Instruments, China) [41], and the principal operating conditions are given in Table S1. In the experiments to determine the concentrations of ammonia, ascorbic acid was used as quencher to avoid the interference from hydroxylamine. Ammonia was quantified by colorimetric method using salicylate procedure from Hach kit tests with 801 program from DR-2400 Hach equipment (Lognes, France). This method allows avoiding possible measurements interferences from SO42− and Mn2+ that present in reaction solution. The total organic carbon (TOC) content of reaction solution was determined with a TOC analyzer (TOCLCPH, Shimadzu, Japan).

3.1. Removal of ASA by Mn(VII)-S(IV) system The effect of different molar ratio of S(IV) to Mn(VII) on ASA removal was investigated at pH ranging from 4.0 to 9.0. As shown in Fig. 1a, a fast decomposition of ASA by Mn(VII)-S(IV) system was achieved in 15 s. No further degradation was observed in longer reaction time (Fig. S2), implying that the reaction of Mn(VII)-S(IV) ended within 15 s or less. In comparison, less than 29% of ASA (5 μM) was removed by Mn(VII) (50 μM) alone in 60 min (Fig. S3), and no degradation of ASA by S(IV) (250 μM) was observed (data not shown). The removal of ASA climbed significantly with the molar ratio of S(IV) to Mn(VII) increasing from 1 to 4 at pH 4.0, and reached a maximum value of 86% at the molar ratio of 4. Further increasing the molar ratio from 5 to 20 caused a decline of ASA removal from 84% to 65%. Similar phenomena were also observed at pH range of 5.0–7.0 with the maximum removal of ASA obtained at S(IV)/Mn(VII) molar ratio of 5, 8, and 10 at pH 5.0, 6.0, and 7.0, respectively. In the presence of 50 μM Mn(VII) and 250 μM S(IV), 71%, 61%, and 19% of ASA were removed within 15 s at pH 5.0, 6.0, and 7.0, respectively. With increasing pH from 8.0 to 9.0, only slight removal of ASA (∼5%) was observed. The inefficiency of Mn(VII)-S(IV) system for ASA degradation at pH 8.0–9.0 may be due to a shift of bisulfite species to sulfite (pKa = 7.0) which lowered the formation of reactive species during reaction under alkaline conditions [27]. This implied the removal of ASA in Mn(VII)-S(IV) system was more favored under acid condition. The influence of initial ASA concentration on the removal of ASA in Mn(VII)-S(IV) system was evaluated with a S(IV)/Mn(VII) molar ratio of 5 at pH 5.0. Complete removal of ASA was achieved in Mn(VII)-S(IV) system after 15 s when ASA concentration was 1 μM (Fig. 1b). The

2.4. Acute toxicity of ASA and its degradation products The bioluminescence of marine bacterium Vibrio fischeri was used to 3

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(7-hydroxycoumarin) via reacting with hydroxyl radical (Eq. (5)) [40]. Hence, coumarin was used as fluorescent probe to selectively detect % OH radical in Mn(VII)-S(IV) process, generating from the reaction of SO4%− with water (Eq. (4)). As shown in Fig. S7, the signal intensity of 7-hydroxycoumarin in Mn(VII)-S(IV) mixture exposing to air was obvious. With the introduction of 100 mM MeOH and TBA, the fluorescence intensity of 7-hydroxycoumarin decreased significantly due to the scavenging reactive radicals by alcohols. It should be note that 7-hydroxycoumarin can not be detected in the reaction of coumarin with S (IV), Mn(VII), or Mn(III), respectively. The formation of 7-hydroxycoumarin confirmed the production of %OH and SO4%− radicals in Mn (VII)-S(IV) system, since %OH was generated from SO4%− reacting with water. Previous studies demonstrated that Mn(III) would accept one electron from S(IV), resulting in the formation of chain propagating sulfite radical, SO3%− [38]. We observed the characteristic peak of Mn(III) at 258 nm in the presence of PP (Fig. S8), but its contribution to the degradation of ASA was negligible (Fig. S4a). It was still unknown whether Mn(III) would lead to the formation of relative radicals by reacting with S(IV). Thus, further exploration was made by carrying out alcohol quenching experiment in Mn(III)-PP and Mn(III)-PP/S(IV) mixtures. As displayed in Fig. S9, the addition of MeOH (250 mM) exhibited negligible influence on BPA degradation, suggesting that radicals were not formed in Mn(III)-PP process. However, the presence of both MeOH and TBA significantly inhibited the degradation of ASA, NB, and BPA by Mn (III)-PP/S(IV) process, and the inhibition from MeOH was remarkably stronger than that of TBA (Fig. S10). This suggested that reactive radicals, %OH and SO4%−, were formed in the reaction of Mn(III) with S (IV), which contributed to the degradation organic pollutants.

removal efficiency decreased from 93% to 41% with increasing initial ASA concentration increasing from 2 to 30 μM. This was attributed to that the competition of ASA for a certain amount of reactive species insitu generated in Mn(VII)-S(IV) system became more and more fierce with increasing ASA concentration. 3.2. Reactive species in Mn(VII)-S(IV) system Numbers of studies reported that the extraordinarily rapid degradation of organic pollutants in Mn(VII)-S(IV) system was attributed to the in situ formed Mn(III) species, which was a strong one-electron oxidizing agent, rather than hydroxyl and sulfate radicals [27,28]. Herein, we examined the role of Mn(III) in ASA degradation. For the reason that free Mn(III) ion is unstable, which will decompose to MnO2 and Mn(II) rapidly, the Mn(III)-PP complexes were prepared to replace the free Mn(III) according to the previous studies [31,33,43]. In the preliminary study, negligible degradation of ASA by Mn(III)-PP complexes was observed in Fig. S4a. However, 71% of ASA was decomposed within 15 s with S(IV)/Mn(VII) molar ratio of 5 in Mn(VII)-S(IV) system, as shown in Fig. S4b. This implied that there might be other reactive species besides Mn(III) in Mn(VII)-S(IV) process. In Fe(VI)-S (IV) [19,22] and Cr(VI)-S(IV) [18,24] systems, both sulfate- and hydroxyl-radicals have proved to be reactive species. Thus, Mn(VII)-S(IV) process might also be a radical-involved process. Therefore, a series of experiments, including electron spin resonance (ESR), free radicals quenching, and fluorescence spectroscopy analysis, were performed to gain direct evidences for the involvement of reactive radicals in Mn (VII)-S(IV) system. ESR spectra were collected using DMPO as a spin trap. As seen in Fig. S5a, the hyperfine coupling constants of measured spectra (∝N = 14.7 G, ∝H = 15.9 G) was consistent with DMPO-SO3%− adducts [44]. Other radicals were not observed perhaps because excess DMPO trapped all of the SO3%− and terminated any subsequent radical propagation reactions. In addition, we found that the spectra of DMPOSO3%− adducts was not observed (Fig. S5b), when adding DMPO immediately after Mn(VII) and S(IV) mixed, indicating a competitive sidereaction happened for SO3%− radical, which transformed SO3%− radical to other species. It was reported that SO3%− reacts with dissolved oxygen at near diffusion-limited rates (k = 2.5 × 109 M−1 s−1) and produces SO5%− [45], which is the precursor of SO4%− and %OH (Eqs. (2)–(4)) [21,22]. Sun et al. [46] recently demonstrated that in Mn(VII)S(VI) process, SO5%− was generated through the oxidation of SO3%− by dissolved oxygen. Therefore, some radicals, i.e., %OH or S-centered radicals (mainly SO3%−, SO4%−, and SO5%−), might be formed in Mn (VII)-S(IV) system [18,19]. Furthermore, a series of trap experiments for identification of reactive radicals were carried out by adding methanol (MeOH) and tertbutanol (TBA) as scavengers. Previous works have shown that MeOH can effectively scavenge %OH and SO4%− at a rate of (1.2–2.8) × 109 M−1 s−1 and (1.6–7.7) × 107 M−1 s−1, respectively [25]. Whereas, TBA shows high reactivity toward %OH (k = (3.8–7.6) × 108 M−1 s−1) and much lower reactivity to SO4%− (k = (4.0–9.1) × 105 M−1 s−1) [25]. Other radicals, e.g., SO3%− and SO5%−, are rather inert towards alcohols (k ≤ 103 M−1 s−1) [18]. As shown in Fig. 2, the removal of ASA was 71% at pH 5.0 in Mn(VII)-S(IV) system, while with the addition of 10 mM MeOH and TBA, it dropped to 43% and 68%, respectively. This suggested SO4%− contributed more than %OH to ASA removal. Additionally, since nitrobenzene (NB) could be oxidized by %OH (k = 3.9 × 109 M−1 s−1) [47] with a much greater rate constant than SO4%− (k < 106 M−1 s−1) [48], it was chosen as a probe compound to investigate the formation of %OH in Mn(VII)-S(IV) system at pH 5.0. As shown in Fig. S6, the removal of NB reached 13% within 15 s. After the introduction of 100 mM MeOH or TBA, the degradation of NB was negligible, indicating %OH was also formed in Mn (VII)-S(IV) process. It is well-known that coumarin, as a non-fluorescent test compound, can be transformed to a strongly fluorescent compound

3.3. Role of dissolved oxygen In sulfite based-AOPs, the mild reactive species sulfite radical (SO3%−) is generated, further reacting with the dissolved oxygen in solution to produce SO5%−, which will decompose to sulfate radical (SO4%−) and hydroxyl radial (%OH) [22,24]. As SO5%−, SO4%−, and % OH are produced from the chain reaction of SO3%− with O2, we defined SO3%− as the first radical, and SO5%−, SO4%− and %OH as the secondary radicals. Sun et al. demonstrated that in Mn(VII)-S(IV) system, the dissolved oxygen involved in the oxidation of SO3%− to SO5%− [46], or merely played a minor role as oxidizing sulfite to sulfate [27], but the secondary radicals have negligible contribution to decompose organic contaminants. In our study, we observed the dissolved oxygen concentration dropped dramatically along with reaction course at various initial pH values (Fig. S11a), when the batch experiments were open to air. After 30 s, about 2.2 mg/L dissolved oxygen was depleted by the combination of 50 μM Mn(VII) and 250 μM S(IV) at pH 4.0–6.0. In contrast, the consumption of dissolved oxygen by sulfite alone was negligible under ambient environment within 30 min, as shown in Fig. S11b. This suggested that sulfite itself rarely consumed dissolved oxygen in the time scale of this research. Thus, the rapid consumption of dissolved oxygen in the Mn(VII)-S(IV) mixture indicated the oxygen molecule might take part in the radical propagation reaction for SO3%−. To further investigate the role of dissolved oxygen in Mn(VII)-S(IV) system, the degradation of ASA by Mn(VII)-S(IV) system at pH 5.0 was examined under deoxygenated condition by purging nitrogen. As presented in Fig. S12, the degradation of ASA was significantly inhibited under the deoxygenated condition. The similar phenomena were also observed for NB and BPA. Moreover, the formation of 7-hydroxycoumarin was completely inhibited under anoxic condition (Fig. S7). These results confirmed the participation of dissolved oxygen in Mn (VII)-S(IV) reaction. It is well known that the dissolved oxygen reacts with SO3%− to form SO5%−, which is the precursor of SO4%− [21,22], as shown in Eqs. (2)–(4). Under anoxic condition, the pathways for the formation of SO5%− and SO4%− would be cut off, resulting in little 4

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Fig. 2. Influences of (a) MeOH and (b) TBA on ASA degradation in Mn(VII)-S(IV) system. Reaction conditions: [ASA]0 = 5 μM, [Mn(VII)]0 = 50 μM, [S (IV)]0 = 250 μM, pH = 5.0, and T = 20 °C.

reactivity of Mn(III) than MnO2. This suggested that Mn intermediates played an important role for BPA degradation under anoxic condition. In contrast, Mn(III) has no activity to ASA and NB (Fig. S4a), and MnO2 has only limited reactivity towards ASA, and no activity to NB (Fig. S14). In addition, the enhanced degradation of ASA and NB by Mn(VII) was negligible in the presence of various ligands (HA, PP, and EDTA) which could stabilize the Mn intermediates (Fig. S15a and b). This implied that Mn intermediates gave a limited contribution to the removal of ASA and NB, consistent with the removal of ASA and NB by Mn(VII)-S(IV) system under anoxic condition. Thus, the fast decomposition of ASA and NB in Mn(VII)-S(IV) system can be safely attributed to the secondary radicals (SO4%− and %OH), while the removal of BPA might be due to the combination of the secondary radicals and the Mn intermediates. In sum, the performance of Mn(VII)-S(IV) system depends not only on the reaction conditions, i.e. dissolved oxygen, pH, molar ratio of S(IV) to Mn(VII), but also on the structures of organic compounds. The organic compounds containing electron rich moieties, such as phenol or aniline, can be degraded by Mn intermediates under the anoxic condition, while other organics retardant to Mn intermediates can only be decomposed by the secondary radicals under oxic condition. These results demonstrated that multi-oxidizing species (Mn intermediates, SO4%−, and %OH) might take part in the removal of organic compounds in Mn(VII)-S(IV) system.

removal of organic compounds. As shown in Fig. S12, in the absence of O2, the removal of ASA and NB were almost completely inhibited in the Mn(VII)-S(IV) mixture, while approximately 45% of BPA removal was still achieved. Compared with BPA degradation by Mn(VII) alone, BPA removal in Mn(VII)-S(IV) mixture under anoxic condition was higher, but still lower than that under oxic condition (Fig. S13). These results indicated that the degradation of ASA and NB in the presence of O2 were mainly attributed to the secondary radicals such as SO4%−, and %OH generated in Mn (VII)-S(IV) system, and there might be some other reactive species responsible for the degradation of BPA besides the secondary radicals. The first radical SO3%− was reported to be a mild oxidant, and Feng et al. [21] recently reported that the involvement of SO3%− to oxidize trimethoprim (TMP) by Fe(VI)-S(IV) was minimal under anoxic condition. Fig. S4a demonstrates that the Mn(III)-PP is highly effective to BPA, but has no activity to ASA and NB. Since the secondary radicals (SO5%−, SO4%− and %OH) were hardly formed in the anoxic condition, Mn intermediates produced in Mn(VII)-S(IV) were more likely responsible for oxidation of BPA under anoxic condition. Although a variety of manganese intermediates might be formed during the reduction of Mn(VII), such as Mn(VI, V, IV, and III), Sun et al. [29] demonstrated Mn(III) was the major intermediate in Mn(VII)-S(IV) system, which would auto-decompose to MnO2 and Mn(II) quickly. Gao et al. [34] recently reported that in the presence of ligands, Mn (VI, V, and IV) other than Mn(III) participated in enhancement oxidation of organic compounds by Mn(VII). Since Mn (VI and V) is not stable and can not be ex-situ prepared, Mn(III) and MnO2 were chosen to investigate the role of Mn intermediates. As depicted in Figs. S4a and S14, both Mn(III) and MnO2 degraded BPA efficiently, with a higher

3.4. Influence of ligands Various ligands (e.g., PP, EDTA, oxalate, and citrate) are potentially released into the environment due to their wide applications in industry, pharmaceutical and agriculture [49,50]. In this study, PP and

Fig. 3. Influences of pyrophosphate (PP), EDTA and HA on ASA degradation in Mn(VII)-S(IV) system: (a) PP, (b) EDTA, and (c) HA. Reaction conditions: [ASA]0 = 5 μM, [Mn(VII)]0 = 50 μM, [S(IV)]0 = 250 μM, pH = 5.0, reaction time 15 s, and T = 20 °C. 5

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Cl− < Br− < I− and the inhibition effects climbed with increasing the initial concentrations of halide ions. This can be attributed to that halide ions reacted with not only %OH, but also SO4%− [52]. The second-order rate constants for the reactions between halide ions and % OH or SO4%− were listed in Table S2. The reaction of halide ions with % OH and SO4%− produced less reactive halogen radicals [55], leading to the less removal of ASA. Although previous study reported that the reaction of reactive radicals (SO4%− and %OH) with F− is insufficient to generate fluorine radical due to its the highest electronegativity (3.98) in the periodic table [26], we observed ASA removal was inhibitedwith increasing F− concentration , which might be attributed to the strong complexing capacities of F− [26]. No extra attempts were made to identify the exact mechanism of the inhibited effect of F− on Mn(III)-S (IV) system, which will be further studied by our group.

EDTA were selected to investigate the influence of ligands on the degradation of ASA by Mn(VII)-S(IV) system. PP is a non-redox active ligand, while EDTA is an organic ligand. Fig. 3a shows that PP concentration has no effect on ASA transformation in the Mn(VII)-S(IV) system. The similar phenomenon was also observed for NB degradation (Fig. S16a). In the absence of PP, MnO2 was found to be the final product of Mn(VII) in the Mn(VII)-S(IV) system. After the addition of PP, a new characteristic absorbance peak at 258 nm appeared, indicating that Mn(III) was stabilized by PP with the formation of Mn(III)PP complexes. Fig. S4a shows that the Mn(III)-PP complexes did not involve in the degradation of ASA and NB. But for BPA, a different effect of PP was observed. As shown in Fig. S16b, when the PP dosage increased from 0 μM to 250 μM, BPA degradation increased markedly from 71% to 98%. While further increasing the PP concentration to 1000 μM, BPA degradation slightly dropped to 87%. Our result is consistent with a recent study [33], which demonstrated that the PP at lower concentration mainly stabilized the in-situ formed Mn(III) and enhanced the performance of Mn(III) in the overall oxidation of organics, but further increased PP concentration reduced the reactivity of Mn(III) due to the over high stability of Mn(III)-PP complexes. In contrast, EDTA inhibited ASA degradation in Mn(VII)-S(IV) system. As illustrated in Fig. 3b, the removal of ASA decreased from 71% to 42% with EDTA concentration increasing from 0 μM to 1000 μM. Although Jiang et al. [33] reported that EDTA promoted the activity of Mn(III) by forming complexes, in this study the Mn(III) did not react with ASA, and only the reactive radicals (SO4%− and %OH) were responsible for ASA degradation. Thus, the negative effect caused by EDTA might be related to the competition for reactive radicals between ASA and EDTA [50,51], which was further confirmed by the degradation of EDTA in Mn(VII)-S(IV) process. As shown in Fig. S17, the removal of EDTA alone in Mn(VII)-S(IV) process decreased from 83% to 21% with increasing initial EDTA concentration from 5 to 500 μM. This implied that EDTA consumed a portion of reactive radicals (SO4%− and %OH).

3.7. Influence of coexisting cations Various cations presenting in groundwater and wastewater may influence the ASA degradation in Mn(VII)-S(IV) system. Fig. 5a and b show that no influences were observed for Ca2+ and Mg2+, because neither Ca2+ or Mg2+ involved into oxidation of S(IV) to SO3%−. While the transition metals, such as Fe2+, Fe3+, and Cu2+, enhanced ASA degradation slightly. Previous studies have shown that SO3%− would be generated during the redox reaction between transition metal cations and S(IV) [37,38,56]. Thus, the synergistic effect of these metals on ASA degradation might be caused by a higher production of reactive radicals. Moreover, it should be noted that Co(II) can activate S(IV) at the pH range of 6.0–11.0 [57,58], but in our study, Co(II) have no significant influence on the degradation of ASA due to the less reactivity of Co(II) to S(IV) at pH 5.0 [58]. 3.8. Transformation pathways and mineralization of ASA Nearly 71% of ASA was degraded by Mn(VI)-S(IV) system at pH 5.0. However, little mineralization was observed (Fig. 6a), suggesting that the intermediates from ASA were difficult to be further decomposed to CO2 under the tested conditions. Hence, it is essential to identify the byproducts of ASA after Mn(VI)-S(IV) process. The concentration of NH4+-N generated from ASA (5 μM) after the treatment by 50 μM Mn (VII) and 250 μM S(IV) at pH 5.0 was determined to be 0.66 mg/L (∼3.7 μM) (Fig. 6a), while no NO2-N and NO3-N was detected. The formed NH4+-N comprises 74% of total N in the feedstock. The production of NH4+-N and the decomposition of ASA were comparable to each other, indicating the cleavage of –NH2 group from ASA. Figs. 6b and S20 show the arsenite concentration increased to its maximum of 1.7 μM at 15 s, and then dropped to 0.16 μM, while the arsenate concentration increased continuously during the whole reaction. Moreover, the total inorganic arsenic (∼3.6 μM) was detected in the solution after 15 s, suggesting that about 72% of organic arsenic was transformed to inorganic arsenic. Thus, it can be concluded that the arsenite was firstly cleaved from aromatic ring of ASA, and subsequently oxidized to arsenate by the reactive species, i.e., Mn(III), SO4%−, and %OH. The organic degradation products of ASA after Mn(VII)-S(IV) treatment were identified by GC–MS. The total ion chromatograms of the GC–MS analysis were provided in Fig. S21, and the major peaks were subjected to qualitative MS analysis. As listed in Table S3, six intermediates were identified, including aniline, aminophenol, hydroquinone, and p-benzoquinone. Similar to %OH, the reaction pathways of SO4%− with organic pollutants contain H-abstraction, addition, and electron transfer. The possible pathways of ASA in Mn(VII)-S(IV) process were, therefore, proposed based on the formation of intermediate products (Fig. 7). The carbon–arsenic bond in ASA may be attacked by SO4%− and %OH to generate aniline and –AsO(OH) group, namely arsenite. Then, the arsenite was subsequently oxidized to arsenate. The aniline was then hydroxylated to aminophenol (orto, meta, and para isomers), which followed a mineralization pathway where phenol was

3.5. Influence of humic acid Natural organic matters were ubiquitous in natural environment, which affect various environmental processes. Humic acid (HA) is one of the main components of natural organic matters. Therefore, the effect of HA on the performance of Mn(VII)-S(IV) system was worthy to be examined. Fig. 3c shows that the removal of ASA decreased with increasing HA concentration from 0 mg C/L to 5 mg C/L, due to that HA consumes SO4%− and %OH generated in Mn(VII)-S(IV) system [52]. It is well known that quinone moieties are considered as the most important redox-active functional groups in HA [53,54], thus the impact of benzoquinone (BQ) on ASA degradation was further investigated. As expected, the degradation of ASA was greatly inhibited with increasing BQ concentration from 5 to 250 μM (Fig. S18). We also observed that the both BQ and HA consumed S(IV) (Fig. S19a and b), which might lead to a reduced production of SO4%− and %OH. Therefore, the influence of HA on ASA degradation in Mn(VII)-S(IV) system was mainly attributed to the consumption of not only radicals, but also S(IV). 3.6. Influence of halide ions The effects of halide ions (F−, Cl−, Br−, and I−) on the removal of ASA by Mn(VII)-S(IV) system were shown in Fig. 4. The halide ions inhibited the degradation of ASA markedly. For example, 71% of ASA was degraded in the absence of halide ions, whereas the degradation efficiency decreased to 53%, 30%, 28% and 15% in the presence of F− (500 mM), Cl− (1000 mM), Br− (100 mM), and I− (100 μM), respectively. The inhibitive capacity generally followed the order of I− > Br− > Cl−. Similar phenomenon was also observed in the oxidation of Orange II by Fe(II)/S(IV) system [17], where halide ions tended to slow down the discoloration of Orange II in the order of 6

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Fig. 4. Influences of halide ions on ASA degradation in the Mn(VII)-S(IV) system. Reaction conditions: [ASA]0 = 5 μM, [Mn(VII)]0 = 50 μM, [S(IV)]0 = 250 μM, pH = 5.0, reaction time 15 s, and T = 20 °C.

Fig. 5. Influences of different concentration of coexisting metal ions on ASA degradation in Mn(VII)-S(IV) system. Reaction conditions: [ASA]0 = 5 μM, [Mn (VII)]0 = 50 μM, [S(IV)]0 = 250 μM, pH = 5.0, reaction time 15 s, and T = 20 °C. (a) [Fe(II)]=[Fe(III)]=[Cu(II)]=[Co(II)] = 5 μM, [Ca(II)]=[Mg(II)] = 50 μM; (b) [Fe(II)]=[Fe(III)]=[Cu(II)]=[Co(II)] = 50 μM, [Ca(II)]=[Mg(II)] = 500 μM.

a constant and continuous production of reactive radicals.

formed and ammonia was released. Some studies reported the ammonia was further oxidized to nitrate [59,60], but in our study no nitrate was detected. Finally, phenol might be oxidized to hydroquinone and pbenzoquinone. The degradation products (hydroquinone, p-benzoquinone, and aminophenol) of ASA were supposed to be decomposed easily by reactive radicals, but negligible TOC removal was observed. The little mineralization in our study should be due to the termination of radicals’ chain reaction in a very short time (within 15 s). In order to confirm our hypothesis, we retreated the reaction solution collected after the oxidation by Mn(VII)-S(IV) system for a second cycle under the same conditions. With adding 50 μM Mn(VII) and 250 μM S(IV), 15% of TOC removal was achieved in the second cycle (Fig. S22). Therefore, in order to achieve the simultaneous degradation and mineralization of ASA, slow release capsules for Mn(VII) or S(IV) can be developed to maintain

3.9. Toxicity evaluation of ASA and its products The luminescence inhibition rate of Vibrio fischeri bacteria was conducted to evaluate the toxicity of reaction solution containing ASA and its products in Mn(VII)-S(IV) system. Fig. S23 shows that the inhibition rate of the initial solution containing 5 μM ASA was about 3% and the inhibition rate increased to 35% after 5 min. The increased toxicity may be mainly caused by the degradation products of ASA. In order to confirm our assumptions, we measured the toxicity of each product alone respectively with a fixed concentration. Even though the concentration of each product was not totally the same as that observed in the experiment, the obtained value still had a significant meaning for indicating the relative toxicity difference between each product and the 7

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Fig. 6. Change of the concentrations of (a) TOC, (a) NH4+-N, and (b) arsenic species as a function of time. Reaction conditions: [ASA]0 = 5 μM, [Mn(VII)]0 = 50 μM, [S(IV)]0 = 250 μM, pH = 5.0, and T = 20 °C.

satisfying, suggesting that Mn(VII) coupled S(IV) was a promising method for degrading contaminants in real water. 4. Conclusions In this study, the extraordinarily fast oxidation of ASA by Mn(VII)-S (IV) was achieved in 15 s. The degradation of ASA depended on pH, initial ASA concentration, molar ratio of S(IV) to Mn(VII), dissolved oxygen, ligands, and co-existing solutes. The results of radical scavenger experiments demonstrated that both SO4%− and %OH involved in ASA degradation with SO4%− as the principal reactive species. Dissolved oxygen played a crucial role by reacting with SO3%− to form SO5%−, which is the precursor of SO4%− and %OH. Thus, under anoxic conditions, the degradation of ASA was totally inhibited due to cutting off the formation of SO4%− and %OH and less reactivity between ASA and Mn intermediates. Compared with ASA, approximately 45% of BPA removal under anoxic condition in Mn(VII)-S(IV) system was attributed to the generation of Mn intermediates. These results indicated that the participation of multi-oxidizing species (Mn intermediates, SO4%−, and % OH) in the removal of organic compounds in Mn(VII)-S(IV) system. Halide ions tended to slow down the removal of ASA in the order of Cl− < Br− < I− and the inhibition effects increased with increasing halide ions concentration. Finally, a detailed pathway for ASA degradation was proposed based on the transformation products and the toxicity of reaction solution was evaluated. This finding improves the fundamental understanding of a new class of advanced oxidation processes based on the combination of Mn(VII) and S(IV) for the treatment of wastewater.

Fig. 7. Proposed degradation pathways of ASA in Mn(VII)-S(IV) system.

parent compound ASA. As shown in Fig. S24, the inhibition rates of Vibrio fischeri caused by inorganic arsenics, i.e., arsenite and arsenate, and the main organic products, except for aniline, were obviously higher than that of ASA. Phenol groups were reported to be more liable to penetrate into the cell of luminescent bacteria due to the lower steric resistance [61,62], which would leading to a higher toxicity. This explained why the organic products (hydroquinone, p-benzoquinone, and aminophenol) processed higher toxicity than ASA. Since the products of ASA contributed to the increased toxicity, other advanced processes, such as adsorption (i.e., activated carbon and Fe3O4), are necessary following the Mn(VII)-S(IV) system in order to reduce residual toxicity.

Acknowledgements This work was supported by the National Natural Science Foundation of China (51508152 and 51878095), the Natural Science Foundation of Jiangsu Province, China (BK20150812), the Foundation of the State Key Laboratory of Pollution Control and Resource Reuse, China (PCRRF16019).

3.10. Degradation of ASA in natural river water To assess the performance of Mn(VII)-S(IV) system in real water, ASA (5 μM) was spiked into natural water samples (the characteristics were listed in Table S4) collected from a river near a poultry factory, Yixing, China, and then treated by the combination of 50 μM Mn(VII) and 250 μM S(IV). As shown in Fig. S25, at pH 5.0 and 6.0, 64% and 51% of ASA were degraded by Mn(VII)-S(IV) system within 15 s in natural river water, respectively. Comparatively, the removal of ASA was slightly lower than that in Milli-Q water. The slight inhibition might be resulted from that organic matter and co-existing solutes in natural river water competed for the reactive radicals with ASA. Generally, the performance of Mn(VII)-S(IV) system in natural water was

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