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Degradation of tris(2-chloroethyl) phosphate (TCEP) in aqueous solution by using pyrite activating persulfate to produce radicals
Ecotoxicology and Environmental Safety 174 (2019) 667–674
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Degradation of tris(2-chloroethyl) phosphate (TCEP) in aqueous solution by using pyrite activating persulfate to produce radicals
T
⁎
Wenjie Liana, Xiaoyun Yia,c, , Kaibo Huanga, Ting Tanga, Rui Wanga, Xueqin Taob, Zeli Zhengb, ⁎ Zhi Danga,c, Hua Yina,c, Guining Lua,c, a
School of Environment and Energy, South China University of Technology, Guangzhou 510006, China College of Environmental Science and Engineering, Zhongkai University of Agriculture and Engineering, Guangzhou 510225, China c The Key Lab of Pollution Control and Ecosystem Restoration in Industry Clusters, Ministry of Education, Guangzhou 510006, China b
A R T I C LE I N FO
A B S T R A C T
Keywords: Pyrite-persulfate Radicals Tris(2-chloroaniline) phosphate Product identification
Organophosphorus esters (OPEs), as one kind of emerging and toxic contaminant are ubiquitous in various environments. This study investigated the degradation of tris(2-chloroethyl) phosphate (TCEP) as a category OPEs by pyrite (FeS2)-activated persulfate (PS). The result shows that near-100% degradation of TCEP was achieved after 120 min in FeS2-PS system. The important role of Fe2+ in the activation mechanism was confirmed by the introduction of Fe2+ into the PS only system. Radical scavengers experiment and electron paramagnetic resonance (EPR) confirmed the presence of SO4·− and ·OH，which revealed that ·OH and SO4·− played major roles in TCEP degradation. The effect of various environmental factors, including pyrite and oxidant dosage, inorganic ions and pH were investigated. The result indicated that Fe3+ and Cl- can accelerate the degradation rate of TCEP and the reaction between TCEP and FeS2-PS favors acidic conditions (pH > 9). In addition, due to the acidification of pyrite, this system can be applied with a wide pH range. Besides, two oxidation products, C4H9Cl2O4P and C2H6ClO4P were identified, which suggest that hydroxylation was probably the main mechanism. This study greatly improves our understanding on TCEP removal in FeS2-PS system.
1. Introduction Organophosphate esters (OPEs) are one of the main kinds of organophosphorus flame retardants (OPFRs) which are widely existed as additives in floor polishes, lubricants, lacquers and hydraulic fluids (Marklund et al., 2003). There are two types of flame retardants that form covalent bonds with materials and dissolved in the product without chemical bonding. However, OPEs belong to the latter, which lead their easy release into the environment through various pathways, such as abrasion, leaching and volatilization during the process of production, usage and disposal (Wei et al., 2015; Wang et al., 2018). There have been a large number of studies reporting OPEs widespread existence in water (Regnery and Püttmann, 2010), atmosphere (Marklund et al., 2005), dust (Takimoto et al., 1999), soil (Mihajlovic et al., 2011) and sediment (Cao et al., 2012). Even worse, these OPEs have been detected in human body including urine, seminal fluid and breast milk, which was possible acquired by food, dust ingestion and inhalation (Moser et al., 2015). Previous studies have shown that the concentration of tris(2-chloroethyl) phosphate (TCEP) in drinking water, municipal effluents and river water reached the levels of ng L−1 ⁎
to μg L−1 (Moser et al., 2015; Wei et al., 2015). However, it has been confirmed that these compounds have carcinogenicity, neurotoxic effects, mutagenicity and genotoxicity (Moser et al., 2015; Wei et al., 2015), and can lead to epilepsy and memory impairment would occurred when exposed to high concentrations of TCEP as a type of OPEs (Tilson et al., 1990). Due to the weak absorb ultraviolet light capacity of the OPEs, direct photolysis was difficult to achieve, which increases the difficulty of natural degradation (Cristale et al., 2017). Moreover, the International Agency for Research on Cancer at the World Health Organization added TCEP on the list of three types of carcinogens on 2017. Therefore, it is desperate to develop an efficient degraded technology for TCEP removal. Advanced oxidation processes (AOPs) are effective and promising technologies for recalcitrant organic pollutants removal in aqueous solution, including ozonization, photolysis, photocatalysis and Fenton oxidation etc. (Gogate and Pandit, 2004; Diao et al., 2017). In recent years, activated persulfate through ultraviolet light, chemicals and thermal, as one of the classes of AOPs has gained more and more attention, because it can produce SO4·− which are highly reactive to degrade organic pollutants (Yang et al., 2010; Tang et al., 2019). In
Corresponding authors at: School of Environment and Energy, South China University of Technology, Guangzhou 510006, China. E-mail addresses: [email protected] (X. Yi), [email protected] (G. Lu).
https://doi.org/10.1016/j.ecoenv.2019.03.027 Received 21 January 2019; Received in revised form 5 March 2019; Accepted 7 March 2019 Available online 13 March 2019 0147-6513/ © 2019 Elsevier Inc. All rights reserved.
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to a 2 mL vial with 0.5 mL of Na2S2O3 (2 mol L−1) that can quench the reaction, and then was filtered with 0.45 µm polyethersulfone filter. Added 1 mL of chromatogram grade ethyl acetate into 0.5 mL of filtered reaction solution, sample extraction was performed with a vortex mixer that vigorously stirred for 20 min. After stratification, 0.5 mL supernatant was used for determination. In order to achieve the reaction effect better, a series of impact factor experiments have been conducted under different pH, dosages of pyrite and PS, and solution environment. When discussing the effects of pH, the pH values of the sample were adjusted with 0.01 mol L−1 NaOH and 0.01 mol L−1 HClO4 solution with digital pH-meter. No buffer solution was used to maintain a constant pH during the reaction process. In order to confirm the existence of radicals, the experiment was prepared by adding radical scavengers, which were methanol and tert-butyl alcohol. They preferentially react with radicals to achieve quenching. The DMPO was added to reaction solution for EPR analysis, and the solution pH of 3, 6 and 11 in FeS2-PS systems were implemented. Meanwhile, the oxidation products of TCEP were also analyzed during the reaction. In a predetermined time, 1 mL sample was taken out and filtered for product determination.
addition, SO4·− have better pH adaptability and higher redox potential (2.5–3.1 V) that exhibit high competence in degradation of hazardous contaminants (Xu et al., 2017). Matta et al. (2007) have investigated the activation of persulfate to degrade organics using various iron minerals as activators, and found that the activation of pyrite is the most prominent. The removal of p-chloroaniline, methyl tert-butyl ether and rhodamine B via generation of radicals through pyrite activated persulfate have explored in previous studies (Liang et al., 2010; Diao et al., 2017; Zhang et al., 2017). Pyrite, as one kind of widespread iron-rich minerals in the environment, has been extensively studied due to its low cost, high efficient and environment-friendly nature. Reactive oxygen species such as H2O2, O2·- and ·OH can be generated when only pyrite was present in aqueous solution (Zhang et al., 2015). The generated SO4·− by the pyrite added to the persulfate aqueous solution is significant in the degradation of pollutants. Therefore, pyrite is well deserved, greatly potential and promising activator. Currently, there is little effort on the degradation of TCEP. (Tang et al., 2018) have explored the degradation properties by TiO2-mediated photocatalysis under various solution conditions and found that common inorganic anions and humic acid inhibited the degradation of TCEP. Activated H2O2, persulfate (PS) and peroxymonosulfate (PMS) through UV produce radicals were also applied to the removal of TCEP with a high efficiency, and three products are detected in their system (Ou et al., 2017; Xu et al., 2017; Liu et al., 2018). In this study, we have investigated the reaction efficiency and oxidation products of TCEP by pyrite-activated PS oxidation. It also provides more experimental data for the mechanism of pyrite activation of persulfate. Besides, it not only finds a suitable path for the efficient degradation of TCEP, but also provides a theoretical basis for the application in the actual environment. On one hand, the activation mechanism was further demonstrated and the effect factors, including the PS and pyrite doses, co-existing inorganic ions and solution pH were investigated to evaluate their impacts on the removal of TCEP in deionized water. In addition, electron paramagnetic resonance (EPR) was used to verify the generation of radical species in this system. On the other hand, ultra-performance liquid chromatography-quadrupletime of flight-mass spectrometry (UPLC-Q-TOF) was applied to identify oxidation products and then the degradation pathways were obtained.
2.3. Analytical methods The quantitative determination of TCEP was analyzed by a ThermoTrace GC Ultra instrument that coupled with a Thermo-DSQ II mass spectrometer (Thermo-Electron Corporation, Waltham, USA). The electron impact ionization mode was set with an electron energy of 70 eV. In order to detect the possible intermediate products, the reaction solutions were conducted by an Agilent 1290 ultra-performance liquid chromatography-quadruple-time of flight-mass spectrometry (UPLC-Q-TOF), equipped with an electrospray ion (ESI) source. The total organic carbon (TOC) content was measured by Elementar Vario TOC. The scanning electron microscopy (SEM) is used to observe the surface morphology of the pyrite. Pre- and post-reaction pyrite were also characterized by X-ray diffraction (XRD) and X-ray fluorescence (XRF) to clarify the composition of pyrite. The Fe2+ and Fe3+ were measured by UV–Vis spectrophotometer. The absorbance of the system was recorded by o-phenanthroline spectrophotometry at λ = 510 nm to obtain the iron concentration.
2. Materials and methods 2.4. Statistical analysis
2.1. Materials
All the experiments were run in triplicate, and the results are presented as the means ± standard deviations of three replicates. Statistical analysis was conducted using Origin Pro 8.1 software.
Tris (2-chloroethyl) phosphate (TCEP, C6H12Cl3O4, purity > 98%) and sodium thiosulfate pentahydrate (Na2S2O3·5H2O, > 99%) were purchased from Sigma Chemical Co., Ltd. Analytical grade methanol and tertiary butanol, as well as sodium persulfate (PS, Na2S2O8, > 98%) were obtained from Tianjin Damao Chemical Reagent Factory (Tianjin, China). 5,5-dimethyl-1-pyrroline N-oxide (DMPO) was provided by ANPEL Laboratory Technologies (Shanghai) Inc. Nature pyrite was mined from Hunan, China. It was ground with an agate mortar and then sieved to obtain fractions smaller than 200 mesh. After that it was rinsed and filtrated with 50% HCl, ethanol and distilled water, respectively. Subsequently, it was dried at 60 °C in a drying vacuum oven under an inert gas, and stored in a drying vessel. Deionized distilled water produced by a Millipore milli-Q system (USA) was used in this study.
3. Results and discussion 3.1. Characterization of pyrite The morphology, structure and composition of pyrites were characterized by SEM, XRD and XRF. As shown in Fig. 1a, the surface of the pyrite was greatly rough with a large amount of particulates adhered. In addition, the surface of the pyrite tended to be smooth after the persulfate was added, indicating that some of particulates were dissolved. Fig. 1b shows the XRD patterns of pyrite before and after the reaction. Diffraction peaks at 33.1°, 37.1°, 40.8°, 56.3°, 59.0°, 61.7 and 64.3° for natural pyrite were well matched with standard pyrite according to ICDD PDF (2004) (Fig. S1). We can observe that FeS2 is the main component, along with other iron sulfide minerals. During the reaction, some of the water-soluble ingredient was gradually disappeared. The XRF was used to determine the chemical composition of pyrite. The result revealed that the mass fraction of pyrite reaches 99.55%, which indicated the high content of FeS2 (Table S1). In addition, it also contains small amounts of other metal oxides.
2.2. Experiments Unless otherwise specified, all of the degradation experiments were implemented in 250 mL of Erlenmeyer flasks containing 95 mL of 2 mg L−1 TCEP solution and 0.15 g L−1 pyrite. They were stirred on a multi-channel magnetic stirrer at 300 rpm and room temperature (30 ± 2 °C). After that, 5 mL PS (50 mmol L−1) was mixed into the reactor. In the specified time intervals, quantitative solution was added 668
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Fig. 1. SEM images of pre- and post-reaction pyrite (a). XRD patterns of pyrite before and after reaction (b).
TCEP, while addition of pyrite to persulfate solution allows TCEP to be more efficiently and rapidly degraded than the other two systems. In addition, the concentration of pyrite in the FeS2 system is more than 20 times that of the FeS2-PS system. Therefore, the adsorption of pyrite cannot be ignored for removal of TCEP in the FeS2 system, while due to low concentration of pyrite has weak adsorption, so physical adsorption is negligible in the FeS2-PS system. This result suggests that the pyrite is capable of activating persulfate for TCEP removal. Previous studies have reported that pyrite can spontaneously generate strong oxidant O·- 2, ·OH and H2O2 in aqueous solution, forming a
3.2. TCEP degradation in different systems The experiments were conducted to observe TCEP degradation in different systems (i.e. FeS2, PS and FeS2-PS systems). The results (Fig. 2) show that TCEP was almost completely degraded at 120 min in FeS2-PS system, and only 20% degradation in PS system, while in the FeS2 system, only 30% TCEP was degraded. The correlation coefficient (R2), kobs and t1/2 were displayed in Table S2. The degradation kinetics of TCEP in FeS2-PS systems fit the pseudo-first-order equation. This indicates that the persulfate or pyrite only cannot effectively degrade
Fig. 2. Degradation of TCEP in pyrite (a), and persulfate and pyrite-persulfate systems (b). Experimental conditions: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.15 g L−1, at ambient temperature. Note that the pyrite concentration of (a) was 4 g L−1. 669
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degrade pollutants. In order to further verify the role of Fe2+ in FeS2-PS system, we added FeSO4 into PS only system. As shown in Fig. 3b, TCEP degraded more rapidly in Fe2+-activated persulfate system than in PS only system, indicating that Fe2+ exert obvious influence on the TCEP degradation by PS. Therefore, the role of Fe2+ is to activate persulfate to produce SO4·− and ·OH. As a common radical scavenger, methanol (MA) and tert-butyl alcohol (TBA) were selected to verify the presence of SO4·− and ·OH. Among them, MA was used as scavenger of SO4·− and ·OH, while TBA was a scavenger of ·OH due to its low reactivity to SO4·− (Hazime et al., 2014; Bu et al., 2018). The volume ratio of scavenger to reaction solution was 1:100 in the present study. Fig. 3c manifested that the removal of TCEP were thoroughly suppressed and almost no degradation occurred in presence of scavengers. Only about 5% and 10% of degradation efficiency were achieved in presence of MA and TBA, suggesting that SO4·− and ·OH played a crucial role in degradation in FeS2-PS system. This results has also been confirmed in previous studies (Diao et al., 2017). In addition, the ·OH were not only generated in-situ by pyrite in water, but also generated via the reaction of SO4·− with water or hydroxide. To provide solid evidence of the existence of SO4·− and ·OH, electron paramagnetic resonance (EPR) was used to measure the presence of radicals in solution. The determination of the type of radicals was identified on the basis of hyperfine coupling constants, which in agreement with literature data (obtained by simulation, ·OH, αH = αN = 14.9 and SO4·−, αH = 9.5, 1.44, 0.79, αN = 13.8) (Feng et al., 2018). There are representative of SO4·− and ·OH added to DMPO (DMPO-SO4, DMPO-OH), suggesting that SO4·− and ·OH were predominant in FeS2PS system. As shown in Fig. 4, the peak of SO4·− and ·OH were really clear, and their intensity increased with decreased pH in FeS2-PS system. The result indicates that radicals prefer acidic environments to alkaline environments, which also implies that the FeS2-PS system is not suitable for application in an alkaline condition. A small amount of hydroxyl radicals had been founded in only pyrite system, which can explain why only 30% TCEP was degraded in FeS2 system after 7days. Based on above analysis, radicals are considered to dominate the reaction process in the FeS2-PS system.
Fenton-like system to oxidize organic pollutants (Eqs. (1)–(6)) (Pham et al., 2009; Wang et al., 2012; Zhang et al., 2015). Oxygen gains an electron from Fe2+ to form reactive oxygen species which plays an important role for the removal of contaminants in the aqueous solution. Therefore, both physical adsorption and chemisorption exist in pyrite system. When persulfate is mixed in aqueous solution of pyrite, it may be activated to generate radicals (Eqs. (7) and (8)) (Liang et al., 2010). 2FeS2 + 7O2 + 2H2O → 2Fe2+ + 4SO42-+ 4H+ Fe
2+
+ O2 → Fe
FeS2 + 14Fe
3+
3+
+
Fe
+ O2 + 2H2O → Fe
Fe
2+
+ H2O2 → Fe
·-
FeS2 +
15S2O82-
2S2O82-
3+
2+
+ 2SO4 + 16H
+
(3)
-
+ 2OH + H2O2
(4)
-
3+
+ 2H
+
+ 16H2O → 2Fe
→ Fe
2-
+ OH + ·OH
2FeS2 + 15H2O2 → 2Fe 2FeS2 +
(2)
+ 8H2O → 15Fe
2+
3+
(1)
O2·-
2+
+
2SO4·-
(5)
+ 3+
+
4SO42+
+ 14H2O
34SO42-
2SO42-
+ 2S
+ 32H
(6) +
(7) (8)
3.3. Active species identification According to Eqs. (1)–(5), we found that Fe2+ released from pyrite plays a significant role in the generation of reactive oxygen species. Fe2+ can transform O2 to O·- 2 in water, which would react with water to generate H2O2. When H2O2 and Fe2+ coexist, a Fenton system was formed. Therefore, the role of iron in this system was further investigated. As shown in Fig. 3a, the concentration of Fe3+ gradually increases while the concentration of Fe2+ remains constant, which were in accordance with the result from previous studies (Zhang et al., 2017). This is attributed to that the oxidation of persulfate causes the conversion of Fe2+ into Fe3+, while Fe2+ are continuously released from pyrite which result in an increase of the concentration of total dissolved iron. Therefore, it is advantageous to provide iron source with pyrite, which can continuously activate persulfate to persistently
Fig. 3. The variation of Fe2+, Fe3+ and total iron in the FeS2- PS system (a). Degradation of TCEP in Fe2+-activated persulfate and persulfate systems (b). Effect of scavenger on the degradation of TCEP in FeS2- PS system (c). Experimental conditions: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [Fe2+]= 10 mmol L−1, at ambient temperature. Note that the pyrite concentration of (a) was 0.5 g L−1, and (c) was 0.15 g L−1.
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revealed that SO4·− can be consumed by the oxidant, like peroxymonosulfate, which will partially counteract the acceleration effect of the increasing PS dosage on the reaction. Li et al. (2013) considered that radical-radical reactions appears in front of radical-organic reactions, S2O82- can also quench SO4·−. Due to comprehensive consideration of cost savings and degradation efficiency, the reaction concentration was determined as 2.5 mmol L−1. SO4·− + S2O82- → SO42- + S2O8·− SO4·−
+
SO4·−
→ S2O8
2-
(9) (10)
3.4.3. Effect of pH To explore the effect of pH on the degradation of TCEP, a series of experiments were arranged with initial pH of 3, 5, 7, 9 and 11 (Fig. 5c). The result shows that the degradation of TCEP were retarded at pH 11 and slightly inhibited at pH 9, while rapidly degraded at pH 3–7. We may draw a conclusion from Table S5 that TCEP degradation is more favorable in acidic condition than alkaline condition. This is due to the fact that the acidic environment favors the presence of Fe2+ which plays a crucial role in the activation of persulfate. As a result, we assumed that when the pH value is less than 9, the degradation of TCEP was favorable in FeS2-PS system. As we all know, pH has a strong influence on Fenton reactions and even most chemical process. Fenton oxidations are highly dependent on the solution pH because of the formation of the precipitation of Fe (III) as ferric hydroxide in alkaline condition (Pignatello et al., 2006). For example, it found that the degradation efficiency of benzophenone-3 decreased with increasing pH which range is 3–9 (Pan et al., 2018). However, some prior studies also reported that acidic and slightly alkaline condition can achieve the purpose of removing organics (Fu et al., 2015). In order to further elucidate the influence of pH, the variation of pH during the reaction process was determined (Fig. 5d). With the exception of the initial pH of 11 whose pH value slowly declines, the remaining pH value drops dramatically in a shorter period of time and becomes an acidic aqueous solution. Pyrite is one of the major reasons that pyrite surface can be oxidized and weathered by oxygen and water (Eqs. (11)–(13)) for the formation of acidic mine wastewater (Bonnissel-Gissinger et al., 1998; Wang et al., 2012). As a consequence, the acidic solution is a favorable environment to remove TCEP by FeS2PS system. However, as the effect of pyrite, even strongly alkaline wastewater is still applicable to FeS2-PS system. Therefore, it is meaningful to apply in a wide pH range for FeS2-PS system. This also explains the reason why there is a large amount of Fe3+ in FeS2-PS system but no precipitation occurs. Unless otherwise stated, the pH of the solutions with an initial pH of 6 were not adjusted in the experiment, due to the pH is gradually reduced regardless of the initial pH value in the FeS2-PS system.
Fig. 4. ESR spectra recorded at ambient temperature and no pH adjustment (pH=6) for the DMPO adduct. ▽ represent for hydroxyl radical，• represent for sulfate radical.
3.4. Influence factors 3.4.1. Effect of initial pyrite concentration To optimize the reaction conditions, experiments were conducted by adding various pyrite dosages (0.03, 0.06, 0.09, 0.3, 0.5 g L−1) to solution. No pH adjustment was necessary because the reaction favors acidic condition, and the initial pH of deionized water is 6. The initial concentration of TCEP is 2 mg L−1 and persulfate is 2.5 mmol L−1. The result (Fig. 5a) shows the degradation rate increased with the increasing pyrite concentration. This is because the higher pyrite dosages can provide higher concentration of Fe2+. The identical conclusion was found in previous studies (Liang et al., 2010). The degradation kinetics of TCEP fit the pseudo-first-order equation. The correlation coefficient (R2), kobs and t1/2 were displayed in Table S3. We found a positive correlation between the rate constant and initial pyrite concentration. As a result, increasing pyrite concentration can improve both TCEP degradation efficiency and persulfate decomposition efficiency. 3.4.2. Effect of initial persulfate concentration For the purpose of exploring the influence of initial PS concentration on degradation, the concentration of 0.5, 1.5, 2.5, 3.5 and 5 mmol L−1 were selected. As Fig. 5b seen, when the PS dosages increased from 0.5 mmol L−1 to 2.5 mmol L−1, the increased oxidant dose can result in the increasing degradation efficiency of TCEP. However when the concentration of PS was greater than 2.5 mmol L−1, the removal rate was reduced despite the high degradation efficiency. The relevant data (R2, kobs and t1/2) were listed in Table S4. This phenomenon can be attributable to self-consumption of radicals and consumption of oxidants on the basis of Eqs. (9) and (10). Pan et al. (2018)
3.4.4. Effect of inorganic ion on the degradation of TCEP Since there are all kinds of ions in the natural water, it is necessary to consider the effect of coexisting inorganic ion on TCEP degradation. Therefore, eight common anions ions (HCO3-, Cl-, H2PO4-, and NO3-) and cations (Fe3+, NH4+, Ca2+, and Cu2+) are determined to estimate their impacts at the concentration of 1 mmol L− 1. The effect of anions ions (Fig. 5e) shows that obvious inhibition for TCEP degradation was observed in the presence of HCO3- and H2PO4-, while NO3- exert a slight inhibition. The suppression effect follows the order of HCO3- > H2PO4> NO3-. These inorganic anions are scavengers of SO4·− and ·OH (Yang et al., 2010; Xu et al., 2017). The hydrolysis of HCO3- and H2PO4- is 671
2FeS2 + 7O2 + 2H2O → 2Fe2+ + 4SO42- + 4H+
(11)
2Fe2+ + O2 + 4H+ → 2Fe3+ + 2H2O
(12)
15 2FeS2 + O2 2
(13)
+ 7H2 O → 2Fe(OH)3(S) + 4SO24− + 8H+
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Fig. 5. Effect of pyrite doses (a), PS concentrations (b) and initial pH (c). The variation of pH at various initial pH (d), anions (e) and cations (f) on the removal of TCEP in pyrite-persulfate system. Experimental conditions: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.015 g L−1 at ambient temperature. Besides, the pH of reaction solutions was adjusted by NaOH or HClO4.
persulfate to consume TCEP, which is shown in Eq. (3).
alkaline, which will increase the pH of the reaction solution to a certain extent, leading to negatively affect the degradation efficiency of TCEP. In addition, NO3- has less effect on the degradation of TCEP as the high redox potential of NO3·. However, the presence of Cl- accelerates the degradation of TCEP, which is consistent with the previous study (Bu et al., 2018). This may be attributed to that the Cl- could also react with SO4·− to produce the other reactive radicals such as Cl·, ClHO· and Cl2·− (Das, 2001; Anastasio and Matthew, 2006). Therefore, this result reflects that the generated active chlorine has the ability to oxidize TCEP, and there may be multiple ways to activate Cl- in the solution, which may be the reason for the enhancement in degradation efficiency. Fig. 5f depicted the degradation efficiency of TCEP in the presence of different inorganic cations solutions. The common cations such as NH4+, Ca2+ and Cu2+ have a slightly negative effect on the TCEP degradation. Among these three ions, Cu2+ exert the most significant inhibition, which may be attributed to the oxidation of Fe2+ by Cu2+ and oxidation of Cu2+ by O2 in acidic solution (Eqs. (14) and (15)) (HoungAloune et al., 2014). The reduction of Fe2+ in solution results in decrease of SO4·−. The influence of NH4+ and Ca2+ is not significant, and the cause of inhibition is not clear at present. Unlike the above cations, Fe3+ exhibited a promoting effect on the TCEP degradation. This is attributed to the fact that Fe3+ can accelerate the activation of
4Cu2+ + 4Fe2+ → 4Cu+ + 4Fe3+ 4Cu
+
+
+4H
+ O2 → 4Cu
2+
+ 2H2O
(14) (15)
3.5. Oxidation products and possible generation pathways TCEP molecule is composed of a phosphorus skeleton and three chloroethane branches. The oxidation products of TCEP by activating persulfate were detected by UPLC-Q-TOF. The total ion chromatograms were shown in Fig. 6a. By comparing samples at different times, new peaks were observed in the reaction samples with obvious changes in the relative intensity, suggesting the generation of oxidation products. With the increasing reaction time, the amount of TCEP gradually decreases, and the production of product 1 and product 2 continues to increase. Two stable oxidation products were confirmed by MS/MS spectrum. They were identified as C4H9Cl2O4P (product 1, m/z 222.9691) and C2H6ClO4P (product 2, m/z 160.9766). The details of the intermediates were shown in Table S6. Product 1 (Fig. 6b) is formed through the replacement an ethyl-chlorine arm of the phosphoric center by a 672
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Fig. 6. Total ion chromatogram (TIC) of TCEP reaction solutions sampled at different time (a). The MS2 spectrum and possible molecular structure of product 1 (222.9691) and product 2(160.9765), (b) and (c). Experimental condition: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.015 g L−1, at ambient temperature.
hydroxyl radical. This may be attributed to that SO4·− attacking the phosphorus backbone, resulting in cracking of one oxygen-ethylchlorine arm. Eventually, after an adding and a series of electron transport, products were generated and reaction reached equilibrium. Product 1 reacts further in the same mechanism to produce product 2 (Fig. 6c), which has a molecular weight of 160.49. After a chain of reactions, a TCEP molecule may be degraded to three chloride ions, one phosphate radical and other substances. In spite of only two intermediates were identified in this study, other products have been reported in other studies (Ye et al., 2017; Liu et al., 2018). However, because of the low concentration or being transformed instantaneously, others are difficult to measure. It is possible that as the degradation time increases TCEP and the intermediates may all be mineralized to CO2 and H2O. The concentration of TOC was measured to determine the degree of mineralization in the FeS2-PS system. As shown in Fig. 7, 60% of the TOC remained in the system after 120 min. Therefore, we speculate that the most primary reaction steps are TCEP → product 1 → product 2 → CO2. The detailed process is presented in Fig. 8.
Fig. 7. Temporal changes of TOC content. [TCEP]0 = 10 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.5 g L−1 at ambient temperature.
system. The acidic environment favors the production of radicals, but the acidification of pyrite enables the system to be applied with a wide range of initial pH range. Inorganic cations have little effect on the degradation process, whereas HCO3- and H2PO4- in anions have significant inhibition. The mechanism of oxidation of SO4·− and ·OH was verified by EPR, both them played a crucial role in TCEP degradation. Furthermore, two intermediates were detected by and a simple degradation path was proposed.
4. Conclusion Pyrite activated persulfate to produce SO4·− and ·OH has been demonstrated to be an effective oxidant process for TCEP removal in this study. Under certain conditions, the degradation rate increases with the increasing pyrite dosage, and there is a threshold for oxidant in FeS2-PS 673
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Fig. 8. Proposed possible degradation pathways of TCEP in pyrite-PS system.
Acknowledgements
190, 225–233. Marklund, A., Andersson, B., Haglund, P., 2003. Screening of organophosphorus compounds and their distribution in various indoor environments. Chemosphere 53, 1137–1146. Marklund, A., Andersson, B., Haglund, P., 2005. Organophosphorus flame retardants and plasticizers in Swedish sewage treatment plants. Environ. Sci. Technol. 39, 7423–7429. Matta, R., Hanna, K., Chiron, S., 2007. Fenton-like oxidation of 2,4,6-trinitrotoluene using different iron minerals. Sci. Total Environ. 385, 242–251. Mihajlovic, I., Miloradov, M.V., Fries, E., 2011. Application of twisselmann extraction, SPME, and GC-MS to assess input sources for organophosphate esters into soil. Environ. Sci. Technol. 45, 2264–2269. Moser, V.C., Phillips, P.M., Hedge, J.M., et al., 2015. Neurotoxicological and thyroid evaluations of rats developmentally exposed to tris(1,3-dichloro-2-propyl)phosphate (TDCIPP) and tris(2-chloro-2-ethyl)phosphate (TCEP). Neurotoxicol. Teratol. 52, 236–247. Ou, H.-s., Liu, J., Ye, J.-s., et al., 2017. Degradation of tris(2-chloroethyl) phosphate by ultraviolet-persulfate: kinetics, pathway and intermediate impact on proteome of Escherichia coli. Chem. Eng. J. 308, 386–395. Pan, X.X., Yan, L.Q., Qu, R.J., et al., 2018. Degradation of the UV-filter benzophenone-3 in aqueous solution using persulfate activated by heat, metal ions and light. Chemosphere 196, 95–104. Pham, H.T., Suto, K., Inoue, C., 2009. Trichloroethylene transformation in aerobic pyrite suspension: pathways and kinetic modeling. Environ. Sci. Technol. 43, 6744–6749. Pignatello, J.J., Oliveros, E., MacKay, A., 2006. Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry. critical reviews in environ. Sci. Technol. 36, 1–84. Regnery, J., Püttmann, W., 2010. Seasonal fluctuations of organophosphate concentrations in precipitation and storm water runoff. Chemosphere 78, 958–964. Takimoto, K., Hirakawa, T., Ito, K., et al., 1999. Source and transport of tricresyl phosphate (TCP) isomers in Kurose river basin. Atmos. Environ. 33, 3191–3200. Tang, T., Lu, G., Wang, R., et al., 2019. Rate constants for the reaction of hydroxyl and sulfate radicals with organophosphorus esters (OPEs) determined by competition method. Ecotoxicol. Environ. Saf. 170, 300–305. Tang, T., Lu, G., Wang, W., et al., 2018. Photocatalytic removal of organic phosphate esters by TiO2: effect of inorganic ions and humic acid. Chemosphere 206, 26–32. Tilson, H.A., Veronesi, B., McLamb, R.L., et al., 1990. Acute exposure to tris(2-chloroethyl) phosphate produces hippocampal neuronal loss and impairs learning in rats. Toxicol. Appl. Pharmacol. 106, 254–269. Wang, R., Tang, T., Lu, G., et al., 2018. Rapid debromination of polybrominated diphenyl ethers (PBDEs) by zero valent metal and bimetals: mechanisms and pathways assisted by density function theory calculation. Environ. Pollut. 240, 745–753. Wang, W., Qu, Y., Yang, B., et al., 2012. Lactate oxidation in pyrite suspension: a Fentonlike process in situ generating H2O2. Chemosphere 86, 376–382. Wei, G.-L., Li, D.-Q., Zhuo, M.-N., et al., 2015. Organophosphorus flame retardants and plasticizers: sources, occurrence, toxicity and human exposure. Environ. Pollut. 196, 29–46. Xu, X.X., Chen, J., Qu, R.J., et al., 2017. Oxidation of tris (2-chloroethyl) phosphate in aqueous solution by UV-activated peroxymonosulfate: kinetics, water matrix effects, degradation products and reaction pathways. Chemosphere 185, 833–843. Yang, S., Wang, P., Yang, X., et al., 2010. Degradation efficiencies of azo dye acid orange 7 by the interaction of heat, UV and anions with common oxidants: persulfate, peroxymonosulfate and hydrogen peroxide. J. Hazard. Mater. 179, 552–558. Ye, J.S., Liu, J., Li, C.S., et al., 2017. Heterogeneous photocatalysis of tris(2-chloroethyl) phosphate by UV/TiO2: degradation products and impacts on bacterial proteome. Water Res. 124, 29–38. Zhang, Y., Tran, H.P., Du, X., et al., 2017. Efficient pyrite activating persulfate process for degradation of p-chloroaniline in aqueous systems: a mechanistic study. Chem. Eng. J. 308, 1112–1119. Zhang, Y., Tran, H.P., Hussain, I., et al., 2015. Degradation of p-chloroaniline by pyrite in aqueous solutions. Chem. Eng. J. 279, 396–401.
The work was financially supported by the National Natural Science Foundation of China (Nos. 41573091, 41673090 and 41771346), the Guangdong Natural Science Funds for Distinguished Young Scholar (No. 2015A030306005), and the Tip-top Scientific and Technical Innovative Youth Talents of Guangdong Special Support Program (No. 2015TQ01Z233). Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2019.03.027. References Anastasio, C., Matthew, B.M., 2006. A chemical probe technique for the determination of reactive halogen species in aqueous solution: Part 2-chloride solutions and mixed bromide/chloride solutions. Atmos. Chem. Phys. 6, 2439–2451. Bonnissel-Gissinger, P., Alnot, M., Ehrhardt, J.J., et al., 1998. Surface oxidation of pyrite as a function of pH. Environ. Sci. Technol. 32, 2839–2845. Bu, L., Zhu, S., Zhou, S., 2018. Degradation of atrazine by electrochemically activated persulfate using BDD anode: role of radicals and influencing factors. Chemosphere 195, 236–244. Cao, S., Zeng, X., Song, H., et al., 2012. Levels and distributions of organophosphate flame retardants and plasticizers in sediment from Taihu lake, China. Environ. Toxicol. Chem. 31, 1478–1484. Cristale, J., Dantas, R.F., De Luca, A., et al., 2017. Role of oxygen and DOM in sunlight induced photodegradation of organophosphorous flame retardants in river water. J. Hazard. Mater. 323, 242–249. Das, T.N., 2001. Reactivity and role of SO5•- radical in aqueous medium Chain oxidation of sulfite to sulfate and atmospheric sulfuric acid generation. J. Phys. Chem. A 105, 9142–9155. Diao, Z.-H., Liu, J.-J., Hu, Y.-X., et al., 2017. Comparative study of rhodamine B degradation by the systems pyrite/H2O2 and pyrite/persulfate: reactivity, stability, products and mechanism. Sep. Purif. Technol. 184, 374–383. Feng, Y., Li, H.L., Lin, L., et al., 2018. Degradation of 1,4-dioxane via controlled generation of radicals by pyrite-activated oxidants: synergistic effects, role of disulfides, and activation sites. Chem. Eng. J. 336, 416–426. Fu, X.R., Gu, X.G., Lu, S.G., et al., 2015. Benzene depletion by Fe2+-catalyzed sodium percarbonate in aqueous solution. Chem. Eng. J. 267, 25–33. Gogate, P.R., Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment II: hybrid methods. Adv. Environ. Res. 8, 553–597. Hazime, R., Nguyen, Q.H., Ferronato, C., et al., 2014. Comparative study of imazalil degradation in three systems: UV/TiO2, UV/K2S2O8 and UV/TiO2/K2S2O8. Appl. Catal. B Environ. 144, 286–291. HoungAloune, S., Kawaai, T., Hiroyoshi, N., et al., 2014. Study on schwertmannite production from copper heap leach solutions and its efficiency in arsenic removal from acidic sulfate solutions. Hydrometallurgy 147, 30–40. Li, B.Z., Li, L., Lin, K.F., et al., 2013. Removal of 1,1,1-trichloroethane from aqueous solution by a sono-activated persulfate process. Ultrason. Sonochem. 20, 855–863. Liang, C., Guo, Y.-Y., Chien, Y.-C., et al., 2010. Oxidative degradation of MTBE by pyriteactivated persulfate: proposed reaction pathways. Ind. Eng. Chem. Res. 49, 8858–8864. Liu, J., Ye, J.S., Chen, Y.F., et al., 2018. UV-driven hydroxyl radical oxidation of tris(2chloroethyl) phosphate: intermediate products and residual toxicity. Chemosphere
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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Degradation of tris(2-chloroethyl) phosphate (TCEP) in aqueous solution by using pyrite activating persulfate to produce radicals
T
⁎
Wenjie Liana, Xiaoyun Yia,c, , Kaibo Huanga, Ting Tanga, Rui Wanga, Xueqin Taob, Zeli Zhengb, ⁎ Zhi Danga,c, Hua Yina,c, Guining Lua,c, a
School of Environment and Energy, South China University of Technology, Guangzhou 510006, China College of Environmental Science and Engineering, Zhongkai University of Agriculture and Engineering, Guangzhou 510225, China c The Key Lab of Pollution Control and Ecosystem Restoration in Industry Clusters, Ministry of Education, Guangzhou 510006, China b
A R T I C LE I N FO
A B S T R A C T
Keywords: Pyrite-persulfate Radicals Tris(2-chloroaniline) phosphate Product identification
Organophosphorus esters (OPEs), as one kind of emerging and toxic contaminant are ubiquitous in various environments. This study investigated the degradation of tris(2-chloroethyl) phosphate (TCEP) as a category OPEs by pyrite (FeS2)-activated persulfate (PS). The result shows that near-100% degradation of TCEP was achieved after 120 min in FeS2-PS system. The important role of Fe2+ in the activation mechanism was confirmed by the introduction of Fe2+ into the PS only system. Radical scavengers experiment and electron paramagnetic resonance (EPR) confirmed the presence of SO4·− and ·OH，which revealed that ·OH and SO4·− played major roles in TCEP degradation. The effect of various environmental factors, including pyrite and oxidant dosage, inorganic ions and pH were investigated. The result indicated that Fe3+ and Cl- can accelerate the degradation rate of TCEP and the reaction between TCEP and FeS2-PS favors acidic conditions (pH > 9). In addition, due to the acidification of pyrite, this system can be applied with a wide pH range. Besides, two oxidation products, C4H9Cl2O4P and C2H6ClO4P were identified, which suggest that hydroxylation was probably the main mechanism. This study greatly improves our understanding on TCEP removal in FeS2-PS system.
1. Introduction Organophosphate esters (OPEs) are one of the main kinds of organophosphorus flame retardants (OPFRs) which are widely existed as additives in floor polishes, lubricants, lacquers and hydraulic fluids (Marklund et al., 2003). There are two types of flame retardants that form covalent bonds with materials and dissolved in the product without chemical bonding. However, OPEs belong to the latter, which lead their easy release into the environment through various pathways, such as abrasion, leaching and volatilization during the process of production, usage and disposal (Wei et al., 2015; Wang et al., 2018). There have been a large number of studies reporting OPEs widespread existence in water (Regnery and Püttmann, 2010), atmosphere (Marklund et al., 2005), dust (Takimoto et al., 1999), soil (Mihajlovic et al., 2011) and sediment (Cao et al., 2012). Even worse, these OPEs have been detected in human body including urine, seminal fluid and breast milk, which was possible acquired by food, dust ingestion and inhalation (Moser et al., 2015). Previous studies have shown that the concentration of tris(2-chloroethyl) phosphate (TCEP) in drinking water, municipal effluents and river water reached the levels of ng L−1 ⁎
to μg L−1 (Moser et al., 2015; Wei et al., 2015). However, it has been confirmed that these compounds have carcinogenicity, neurotoxic effects, mutagenicity and genotoxicity (Moser et al., 2015; Wei et al., 2015), and can lead to epilepsy and memory impairment would occurred when exposed to high concentrations of TCEP as a type of OPEs (Tilson et al., 1990). Due to the weak absorb ultraviolet light capacity of the OPEs, direct photolysis was difficult to achieve, which increases the difficulty of natural degradation (Cristale et al., 2017). Moreover, the International Agency for Research on Cancer at the World Health Organization added TCEP on the list of three types of carcinogens on 2017. Therefore, it is desperate to develop an efficient degraded technology for TCEP removal. Advanced oxidation processes (AOPs) are effective and promising technologies for recalcitrant organic pollutants removal in aqueous solution, including ozonization, photolysis, photocatalysis and Fenton oxidation etc. (Gogate and Pandit, 2004; Diao et al., 2017). In recent years, activated persulfate through ultraviolet light, chemicals and thermal, as one of the classes of AOPs has gained more and more attention, because it can produce SO4·− which are highly reactive to degrade organic pollutants (Yang et al., 2010; Tang et al., 2019). In
Corresponding authors at: School of Environment and Energy, South China University of Technology, Guangzhou 510006, China. E-mail addresses: [email protected] (X. Yi), [email protected] (G. Lu).
https://doi.org/10.1016/j.ecoenv.2019.03.027 Received 21 January 2019; Received in revised form 5 March 2019; Accepted 7 March 2019 Available online 13 March 2019 0147-6513/ © 2019 Elsevier Inc. All rights reserved.
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to a 2 mL vial with 0.5 mL of Na2S2O3 (2 mol L−1) that can quench the reaction, and then was filtered with 0.45 µm polyethersulfone filter. Added 1 mL of chromatogram grade ethyl acetate into 0.5 mL of filtered reaction solution, sample extraction was performed with a vortex mixer that vigorously stirred for 20 min. After stratification, 0.5 mL supernatant was used for determination. In order to achieve the reaction effect better, a series of impact factor experiments have been conducted under different pH, dosages of pyrite and PS, and solution environment. When discussing the effects of pH, the pH values of the sample were adjusted with 0.01 mol L−1 NaOH and 0.01 mol L−1 HClO4 solution with digital pH-meter. No buffer solution was used to maintain a constant pH during the reaction process. In order to confirm the existence of radicals, the experiment was prepared by adding radical scavengers, which were methanol and tert-butyl alcohol. They preferentially react with radicals to achieve quenching. The DMPO was added to reaction solution for EPR analysis, and the solution pH of 3, 6 and 11 in FeS2-PS systems were implemented. Meanwhile, the oxidation products of TCEP were also analyzed during the reaction. In a predetermined time, 1 mL sample was taken out and filtered for product determination.
addition, SO4·− have better pH adaptability and higher redox potential (2.5–3.1 V) that exhibit high competence in degradation of hazardous contaminants (Xu et al., 2017). Matta et al. (2007) have investigated the activation of persulfate to degrade organics using various iron minerals as activators, and found that the activation of pyrite is the most prominent. The removal of p-chloroaniline, methyl tert-butyl ether and rhodamine B via generation of radicals through pyrite activated persulfate have explored in previous studies (Liang et al., 2010; Diao et al., 2017; Zhang et al., 2017). Pyrite, as one kind of widespread iron-rich minerals in the environment, has been extensively studied due to its low cost, high efficient and environment-friendly nature. Reactive oxygen species such as H2O2, O2·- and ·OH can be generated when only pyrite was present in aqueous solution (Zhang et al., 2015). The generated SO4·− by the pyrite added to the persulfate aqueous solution is significant in the degradation of pollutants. Therefore, pyrite is well deserved, greatly potential and promising activator. Currently, there is little effort on the degradation of TCEP. (Tang et al., 2018) have explored the degradation properties by TiO2-mediated photocatalysis under various solution conditions and found that common inorganic anions and humic acid inhibited the degradation of TCEP. Activated H2O2, persulfate (PS) and peroxymonosulfate (PMS) through UV produce radicals were also applied to the removal of TCEP with a high efficiency, and three products are detected in their system (Ou et al., 2017; Xu et al., 2017; Liu et al., 2018). In this study, we have investigated the reaction efficiency and oxidation products of TCEP by pyrite-activated PS oxidation. It also provides more experimental data for the mechanism of pyrite activation of persulfate. Besides, it not only finds a suitable path for the efficient degradation of TCEP, but also provides a theoretical basis for the application in the actual environment. On one hand, the activation mechanism was further demonstrated and the effect factors, including the PS and pyrite doses, co-existing inorganic ions and solution pH were investigated to evaluate their impacts on the removal of TCEP in deionized water. In addition, electron paramagnetic resonance (EPR) was used to verify the generation of radical species in this system. On the other hand, ultra-performance liquid chromatography-quadrupletime of flight-mass spectrometry (UPLC-Q-TOF) was applied to identify oxidation products and then the degradation pathways were obtained.
2.3. Analytical methods The quantitative determination of TCEP was analyzed by a ThermoTrace GC Ultra instrument that coupled with a Thermo-DSQ II mass spectrometer (Thermo-Electron Corporation, Waltham, USA). The electron impact ionization mode was set with an electron energy of 70 eV. In order to detect the possible intermediate products, the reaction solutions were conducted by an Agilent 1290 ultra-performance liquid chromatography-quadruple-time of flight-mass spectrometry (UPLC-Q-TOF), equipped with an electrospray ion (ESI) source. The total organic carbon (TOC) content was measured by Elementar Vario TOC. The scanning electron microscopy (SEM) is used to observe the surface morphology of the pyrite. Pre- and post-reaction pyrite were also characterized by X-ray diffraction (XRD) and X-ray fluorescence (XRF) to clarify the composition of pyrite. The Fe2+ and Fe3+ were measured by UV–Vis spectrophotometer. The absorbance of the system was recorded by o-phenanthroline spectrophotometry at λ = 510 nm to obtain the iron concentration.
2. Materials and methods 2.4. Statistical analysis
2.1. Materials
All the experiments were run in triplicate, and the results are presented as the means ± standard deviations of three replicates. Statistical analysis was conducted using Origin Pro 8.1 software.
Tris (2-chloroethyl) phosphate (TCEP, C6H12Cl3O4, purity > 98%) and sodium thiosulfate pentahydrate (Na2S2O3·5H2O, > 99%) were purchased from Sigma Chemical Co., Ltd. Analytical grade methanol and tertiary butanol, as well as sodium persulfate (PS, Na2S2O8, > 98%) were obtained from Tianjin Damao Chemical Reagent Factory (Tianjin, China). 5,5-dimethyl-1-pyrroline N-oxide (DMPO) was provided by ANPEL Laboratory Technologies (Shanghai) Inc. Nature pyrite was mined from Hunan, China. It was ground with an agate mortar and then sieved to obtain fractions smaller than 200 mesh. After that it was rinsed and filtrated with 50% HCl, ethanol and distilled water, respectively. Subsequently, it was dried at 60 °C in a drying vacuum oven under an inert gas, and stored in a drying vessel. Deionized distilled water produced by a Millipore milli-Q system (USA) was used in this study.
3. Results and discussion 3.1. Characterization of pyrite The morphology, structure and composition of pyrites were characterized by SEM, XRD and XRF. As shown in Fig. 1a, the surface of the pyrite was greatly rough with a large amount of particulates adhered. In addition, the surface of the pyrite tended to be smooth after the persulfate was added, indicating that some of particulates were dissolved. Fig. 1b shows the XRD patterns of pyrite before and after the reaction. Diffraction peaks at 33.1°, 37.1°, 40.8°, 56.3°, 59.0°, 61.7 and 64.3° for natural pyrite were well matched with standard pyrite according to ICDD PDF (2004) (Fig. S1). We can observe that FeS2 is the main component, along with other iron sulfide minerals. During the reaction, some of the water-soluble ingredient was gradually disappeared. The XRF was used to determine the chemical composition of pyrite. The result revealed that the mass fraction of pyrite reaches 99.55%, which indicated the high content of FeS2 (Table S1). In addition, it also contains small amounts of other metal oxides.
2.2. Experiments Unless otherwise specified, all of the degradation experiments were implemented in 250 mL of Erlenmeyer flasks containing 95 mL of 2 mg L−1 TCEP solution and 0.15 g L−1 pyrite. They were stirred on a multi-channel magnetic stirrer at 300 rpm and room temperature (30 ± 2 °C). After that, 5 mL PS (50 mmol L−1) was mixed into the reactor. In the specified time intervals, quantitative solution was added 668
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Fig. 1. SEM images of pre- and post-reaction pyrite (a). XRD patterns of pyrite before and after reaction (b).
TCEP, while addition of pyrite to persulfate solution allows TCEP to be more efficiently and rapidly degraded than the other two systems. In addition, the concentration of pyrite in the FeS2 system is more than 20 times that of the FeS2-PS system. Therefore, the adsorption of pyrite cannot be ignored for removal of TCEP in the FeS2 system, while due to low concentration of pyrite has weak adsorption, so physical adsorption is negligible in the FeS2-PS system. This result suggests that the pyrite is capable of activating persulfate for TCEP removal. Previous studies have reported that pyrite can spontaneously generate strong oxidant O·- 2, ·OH and H2O2 in aqueous solution, forming a
3.2. TCEP degradation in different systems The experiments were conducted to observe TCEP degradation in different systems (i.e. FeS2, PS and FeS2-PS systems). The results (Fig. 2) show that TCEP was almost completely degraded at 120 min in FeS2-PS system, and only 20% degradation in PS system, while in the FeS2 system, only 30% TCEP was degraded. The correlation coefficient (R2), kobs and t1/2 were displayed in Table S2. The degradation kinetics of TCEP in FeS2-PS systems fit the pseudo-first-order equation. This indicates that the persulfate or pyrite only cannot effectively degrade
Fig. 2. Degradation of TCEP in pyrite (a), and persulfate and pyrite-persulfate systems (b). Experimental conditions: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.15 g L−1, at ambient temperature. Note that the pyrite concentration of (a) was 4 g L−1. 669
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degrade pollutants. In order to further verify the role of Fe2+ in FeS2-PS system, we added FeSO4 into PS only system. As shown in Fig. 3b, TCEP degraded more rapidly in Fe2+-activated persulfate system than in PS only system, indicating that Fe2+ exert obvious influence on the TCEP degradation by PS. Therefore, the role of Fe2+ is to activate persulfate to produce SO4·− and ·OH. As a common radical scavenger, methanol (MA) and tert-butyl alcohol (TBA) were selected to verify the presence of SO4·− and ·OH. Among them, MA was used as scavenger of SO4·− and ·OH, while TBA was a scavenger of ·OH due to its low reactivity to SO4·− (Hazime et al., 2014; Bu et al., 2018). The volume ratio of scavenger to reaction solution was 1:100 in the present study. Fig. 3c manifested that the removal of TCEP were thoroughly suppressed and almost no degradation occurred in presence of scavengers. Only about 5% and 10% of degradation efficiency were achieved in presence of MA and TBA, suggesting that SO4·− and ·OH played a crucial role in degradation in FeS2-PS system. This results has also been confirmed in previous studies (Diao et al., 2017). In addition, the ·OH were not only generated in-situ by pyrite in water, but also generated via the reaction of SO4·− with water or hydroxide. To provide solid evidence of the existence of SO4·− and ·OH, electron paramagnetic resonance (EPR) was used to measure the presence of radicals in solution. The determination of the type of radicals was identified on the basis of hyperfine coupling constants, which in agreement with literature data (obtained by simulation, ·OH, αH = αN = 14.9 and SO4·−, αH = 9.5, 1.44, 0.79, αN = 13.8) (Feng et al., 2018). There are representative of SO4·− and ·OH added to DMPO (DMPO-SO4, DMPO-OH), suggesting that SO4·− and ·OH were predominant in FeS2PS system. As shown in Fig. 4, the peak of SO4·− and ·OH were really clear, and their intensity increased with decreased pH in FeS2-PS system. The result indicates that radicals prefer acidic environments to alkaline environments, which also implies that the FeS2-PS system is not suitable for application in an alkaline condition. A small amount of hydroxyl radicals had been founded in only pyrite system, which can explain why only 30% TCEP was degraded in FeS2 system after 7days. Based on above analysis, radicals are considered to dominate the reaction process in the FeS2-PS system.
Fenton-like system to oxidize organic pollutants (Eqs. (1)–(6)) (Pham et al., 2009; Wang et al., 2012; Zhang et al., 2015). Oxygen gains an electron from Fe2+ to form reactive oxygen species which plays an important role for the removal of contaminants in the aqueous solution. Therefore, both physical adsorption and chemisorption exist in pyrite system. When persulfate is mixed in aqueous solution of pyrite, it may be activated to generate radicals (Eqs. (7) and (8)) (Liang et al., 2010). 2FeS2 + 7O2 + 2H2O → 2Fe2+ + 4SO42-+ 4H+ Fe
2+
+ O2 → Fe
FeS2 + 14Fe
3+
3+
+
Fe
+ O2 + 2H2O → Fe
Fe
2+
+ H2O2 → Fe
·-
FeS2 +
15S2O82-
2S2O82-
3+
2+
+ 2SO4 + 16H
+
(3)
-
+ 2OH + H2O2
(4)
-
3+
+ 2H
+
+ 16H2O → 2Fe
→ Fe
2-
+ OH + ·OH
2FeS2 + 15H2O2 → 2Fe 2FeS2 +
(2)
+ 8H2O → 15Fe
2+
3+
(1)
O2·-
2+
+
2SO4·-
(5)
+ 3+
+
4SO42+
+ 14H2O
34SO42-
2SO42-
+ 2S
+ 32H
(6) +
(7) (8)
3.3. Active species identification According to Eqs. (1)–(5), we found that Fe2+ released from pyrite plays a significant role in the generation of reactive oxygen species. Fe2+ can transform O2 to O·- 2 in water, which would react with water to generate H2O2. When H2O2 and Fe2+ coexist, a Fenton system was formed. Therefore, the role of iron in this system was further investigated. As shown in Fig. 3a, the concentration of Fe3+ gradually increases while the concentration of Fe2+ remains constant, which were in accordance with the result from previous studies (Zhang et al., 2017). This is attributed to that the oxidation of persulfate causes the conversion of Fe2+ into Fe3+, while Fe2+ are continuously released from pyrite which result in an increase of the concentration of total dissolved iron. Therefore, it is advantageous to provide iron source with pyrite, which can continuously activate persulfate to persistently
Fig. 3. The variation of Fe2+, Fe3+ and total iron in the FeS2- PS system (a). Degradation of TCEP in Fe2+-activated persulfate and persulfate systems (b). Effect of scavenger on the degradation of TCEP in FeS2- PS system (c). Experimental conditions: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [Fe2+]= 10 mmol L−1, at ambient temperature. Note that the pyrite concentration of (a) was 0.5 g L−1, and (c) was 0.15 g L−1.
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revealed that SO4·− can be consumed by the oxidant, like peroxymonosulfate, which will partially counteract the acceleration effect of the increasing PS dosage on the reaction. Li et al. (2013) considered that radical-radical reactions appears in front of radical-organic reactions, S2O82- can also quench SO4·−. Due to comprehensive consideration of cost savings and degradation efficiency, the reaction concentration was determined as 2.5 mmol L−1. SO4·− + S2O82- → SO42- + S2O8·− SO4·−
+
SO4·−
→ S2O8
2-
(9) (10)
3.4.3. Effect of pH To explore the effect of pH on the degradation of TCEP, a series of experiments were arranged with initial pH of 3, 5, 7, 9 and 11 (Fig. 5c). The result shows that the degradation of TCEP were retarded at pH 11 and slightly inhibited at pH 9, while rapidly degraded at pH 3–7. We may draw a conclusion from Table S5 that TCEP degradation is more favorable in acidic condition than alkaline condition. This is due to the fact that the acidic environment favors the presence of Fe2+ which plays a crucial role in the activation of persulfate. As a result, we assumed that when the pH value is less than 9, the degradation of TCEP was favorable in FeS2-PS system. As we all know, pH has a strong influence on Fenton reactions and even most chemical process. Fenton oxidations are highly dependent on the solution pH because of the formation of the precipitation of Fe (III) as ferric hydroxide in alkaline condition (Pignatello et al., 2006). For example, it found that the degradation efficiency of benzophenone-3 decreased with increasing pH which range is 3–9 (Pan et al., 2018). However, some prior studies also reported that acidic and slightly alkaline condition can achieve the purpose of removing organics (Fu et al., 2015). In order to further elucidate the influence of pH, the variation of pH during the reaction process was determined (Fig. 5d). With the exception of the initial pH of 11 whose pH value slowly declines, the remaining pH value drops dramatically in a shorter period of time and becomes an acidic aqueous solution. Pyrite is one of the major reasons that pyrite surface can be oxidized and weathered by oxygen and water (Eqs. (11)–(13)) for the formation of acidic mine wastewater (Bonnissel-Gissinger et al., 1998; Wang et al., 2012). As a consequence, the acidic solution is a favorable environment to remove TCEP by FeS2PS system. However, as the effect of pyrite, even strongly alkaline wastewater is still applicable to FeS2-PS system. Therefore, it is meaningful to apply in a wide pH range for FeS2-PS system. This also explains the reason why there is a large amount of Fe3+ in FeS2-PS system but no precipitation occurs. Unless otherwise stated, the pH of the solutions with an initial pH of 6 were not adjusted in the experiment, due to the pH is gradually reduced regardless of the initial pH value in the FeS2-PS system.
Fig. 4. ESR spectra recorded at ambient temperature and no pH adjustment (pH=6) for the DMPO adduct. ▽ represent for hydroxyl radical，• represent for sulfate radical.
3.4. Influence factors 3.4.1. Effect of initial pyrite concentration To optimize the reaction conditions, experiments were conducted by adding various pyrite dosages (0.03, 0.06, 0.09, 0.3, 0.5 g L−1) to solution. No pH adjustment was necessary because the reaction favors acidic condition, and the initial pH of deionized water is 6. The initial concentration of TCEP is 2 mg L−1 and persulfate is 2.5 mmol L−1. The result (Fig. 5a) shows the degradation rate increased with the increasing pyrite concentration. This is because the higher pyrite dosages can provide higher concentration of Fe2+. The identical conclusion was found in previous studies (Liang et al., 2010). The degradation kinetics of TCEP fit the pseudo-first-order equation. The correlation coefficient (R2), kobs and t1/2 were displayed in Table S3. We found a positive correlation between the rate constant and initial pyrite concentration. As a result, increasing pyrite concentration can improve both TCEP degradation efficiency and persulfate decomposition efficiency. 3.4.2. Effect of initial persulfate concentration For the purpose of exploring the influence of initial PS concentration on degradation, the concentration of 0.5, 1.5, 2.5, 3.5 and 5 mmol L−1 were selected. As Fig. 5b seen, when the PS dosages increased from 0.5 mmol L−1 to 2.5 mmol L−1, the increased oxidant dose can result in the increasing degradation efficiency of TCEP. However when the concentration of PS was greater than 2.5 mmol L−1, the removal rate was reduced despite the high degradation efficiency. The relevant data (R2, kobs and t1/2) were listed in Table S4. This phenomenon can be attributable to self-consumption of radicals and consumption of oxidants on the basis of Eqs. (9) and (10). Pan et al. (2018)
3.4.4. Effect of inorganic ion on the degradation of TCEP Since there are all kinds of ions in the natural water, it is necessary to consider the effect of coexisting inorganic ion on TCEP degradation. Therefore, eight common anions ions (HCO3-, Cl-, H2PO4-, and NO3-) and cations (Fe3+, NH4+, Ca2+, and Cu2+) are determined to estimate their impacts at the concentration of 1 mmol L− 1. The effect of anions ions (Fig. 5e) shows that obvious inhibition for TCEP degradation was observed in the presence of HCO3- and H2PO4-, while NO3- exert a slight inhibition. The suppression effect follows the order of HCO3- > H2PO4> NO3-. These inorganic anions are scavengers of SO4·− and ·OH (Yang et al., 2010; Xu et al., 2017). The hydrolysis of HCO3- and H2PO4- is 671
2FeS2 + 7O2 + 2H2O → 2Fe2+ + 4SO42- + 4H+
(11)
2Fe2+ + O2 + 4H+ → 2Fe3+ + 2H2O
(12)
15 2FeS2 + O2 2
(13)
+ 7H2 O → 2Fe(OH)3(S) + 4SO24− + 8H+
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Fig. 5. Effect of pyrite doses (a), PS concentrations (b) and initial pH (c). The variation of pH at various initial pH (d), anions (e) and cations (f) on the removal of TCEP in pyrite-persulfate system. Experimental conditions: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.015 g L−1 at ambient temperature. Besides, the pH of reaction solutions was adjusted by NaOH or HClO4.
persulfate to consume TCEP, which is shown in Eq. (3).
alkaline, which will increase the pH of the reaction solution to a certain extent, leading to negatively affect the degradation efficiency of TCEP. In addition, NO3- has less effect on the degradation of TCEP as the high redox potential of NO3·. However, the presence of Cl- accelerates the degradation of TCEP, which is consistent with the previous study (Bu et al., 2018). This may be attributed to that the Cl- could also react with SO4·− to produce the other reactive radicals such as Cl·, ClHO· and Cl2·− (Das, 2001; Anastasio and Matthew, 2006). Therefore, this result reflects that the generated active chlorine has the ability to oxidize TCEP, and there may be multiple ways to activate Cl- in the solution, which may be the reason for the enhancement in degradation efficiency. Fig. 5f depicted the degradation efficiency of TCEP in the presence of different inorganic cations solutions. The common cations such as NH4+, Ca2+ and Cu2+ have a slightly negative effect on the TCEP degradation. Among these three ions, Cu2+ exert the most significant inhibition, which may be attributed to the oxidation of Fe2+ by Cu2+ and oxidation of Cu2+ by O2 in acidic solution (Eqs. (14) and (15)) (HoungAloune et al., 2014). The reduction of Fe2+ in solution results in decrease of SO4·−. The influence of NH4+ and Ca2+ is not significant, and the cause of inhibition is not clear at present. Unlike the above cations, Fe3+ exhibited a promoting effect on the TCEP degradation. This is attributed to the fact that Fe3+ can accelerate the activation of
4Cu2+ + 4Fe2+ → 4Cu+ + 4Fe3+ 4Cu
+
+
+4H
+ O2 → 4Cu
2+
+ 2H2O
(14) (15)
3.5. Oxidation products and possible generation pathways TCEP molecule is composed of a phosphorus skeleton and three chloroethane branches. The oxidation products of TCEP by activating persulfate were detected by UPLC-Q-TOF. The total ion chromatograms were shown in Fig. 6a. By comparing samples at different times, new peaks were observed in the reaction samples with obvious changes in the relative intensity, suggesting the generation of oxidation products. With the increasing reaction time, the amount of TCEP gradually decreases, and the production of product 1 and product 2 continues to increase. Two stable oxidation products were confirmed by MS/MS spectrum. They were identified as C4H9Cl2O4P (product 1, m/z 222.9691) and C2H6ClO4P (product 2, m/z 160.9766). The details of the intermediates were shown in Table S6. Product 1 (Fig. 6b) is formed through the replacement an ethyl-chlorine arm of the phosphoric center by a 672
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Fig. 6. Total ion chromatogram (TIC) of TCEP reaction solutions sampled at different time (a). The MS2 spectrum and possible molecular structure of product 1 (222.9691) and product 2(160.9765), (b) and (c). Experimental condition: [TCEP]0 = 2 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.015 g L−1, at ambient temperature.
hydroxyl radical. This may be attributed to that SO4·− attacking the phosphorus backbone, resulting in cracking of one oxygen-ethylchlorine arm. Eventually, after an adding and a series of electron transport, products were generated and reaction reached equilibrium. Product 1 reacts further in the same mechanism to produce product 2 (Fig. 6c), which has a molecular weight of 160.49. After a chain of reactions, a TCEP molecule may be degraded to three chloride ions, one phosphate radical and other substances. In spite of only two intermediates were identified in this study, other products have been reported in other studies (Ye et al., 2017; Liu et al., 2018). However, because of the low concentration or being transformed instantaneously, others are difficult to measure. It is possible that as the degradation time increases TCEP and the intermediates may all be mineralized to CO2 and H2O. The concentration of TOC was measured to determine the degree of mineralization in the FeS2-PS system. As shown in Fig. 7, 60% of the TOC remained in the system after 120 min. Therefore, we speculate that the most primary reaction steps are TCEP → product 1 → product 2 → CO2. The detailed process is presented in Fig. 8.
Fig. 7. Temporal changes of TOC content. [TCEP]0 = 10 mg L−1, [PS]0 = 2.5 mmol L−1, [pyrite]0 = 0.5 g L−1 at ambient temperature.
system. The acidic environment favors the production of radicals, but the acidification of pyrite enables the system to be applied with a wide range of initial pH range. Inorganic cations have little effect on the degradation process, whereas HCO3- and H2PO4- in anions have significant inhibition. The mechanism of oxidation of SO4·− and ·OH was verified by EPR, both them played a crucial role in TCEP degradation. Furthermore, two intermediates were detected by and a simple degradation path was proposed.
4. Conclusion Pyrite activated persulfate to produce SO4·− and ·OH has been demonstrated to be an effective oxidant process for TCEP removal in this study. Under certain conditions, the degradation rate increases with the increasing pyrite dosage, and there is a threshold for oxidant in FeS2-PS 673
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Fig. 8. Proposed possible degradation pathways of TCEP in pyrite-PS system.
Acknowledgements
190, 225–233. Marklund, A., Andersson, B., Haglund, P., 2003. Screening of organophosphorus compounds and their distribution in various indoor environments. Chemosphere 53, 1137–1146. Marklund, A., Andersson, B., Haglund, P., 2005. Organophosphorus flame retardants and plasticizers in Swedish sewage treatment plants. Environ. Sci. Technol. 39, 7423–7429. Matta, R., Hanna, K., Chiron, S., 2007. Fenton-like oxidation of 2,4,6-trinitrotoluene using different iron minerals. Sci. Total Environ. 385, 242–251. Mihajlovic, I., Miloradov, M.V., Fries, E., 2011. Application of twisselmann extraction, SPME, and GC-MS to assess input sources for organophosphate esters into soil. Environ. Sci. Technol. 45, 2264–2269. Moser, V.C., Phillips, P.M., Hedge, J.M., et al., 2015. Neurotoxicological and thyroid evaluations of rats developmentally exposed to tris(1,3-dichloro-2-propyl)phosphate (TDCIPP) and tris(2-chloro-2-ethyl)phosphate (TCEP). Neurotoxicol. Teratol. 52, 236–247. Ou, H.-s., Liu, J., Ye, J.-s., et al., 2017. Degradation of tris(2-chloroethyl) phosphate by ultraviolet-persulfate: kinetics, pathway and intermediate impact on proteome of Escherichia coli. Chem. Eng. J. 308, 386–395. Pan, X.X., Yan, L.Q., Qu, R.J., et al., 2018. Degradation of the UV-filter benzophenone-3 in aqueous solution using persulfate activated by heat, metal ions and light. Chemosphere 196, 95–104. Pham, H.T., Suto, K., Inoue, C., 2009. Trichloroethylene transformation in aerobic pyrite suspension: pathways and kinetic modeling. Environ. Sci. Technol. 43, 6744–6749. Pignatello, J.J., Oliveros, E., MacKay, A., 2006. Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry. critical reviews in environ. Sci. Technol. 36, 1–84. Regnery, J., Püttmann, W., 2010. Seasonal fluctuations of organophosphate concentrations in precipitation and storm water runoff. Chemosphere 78, 958–964. Takimoto, K., Hirakawa, T., Ito, K., et al., 1999. Source and transport of tricresyl phosphate (TCP) isomers in Kurose river basin. Atmos. Environ. 33, 3191–3200. Tang, T., Lu, G., Wang, R., et al., 2019. Rate constants for the reaction of hydroxyl and sulfate radicals with organophosphorus esters (OPEs) determined by competition method. Ecotoxicol. Environ. Saf. 170, 300–305. Tang, T., Lu, G., Wang, W., et al., 2018. Photocatalytic removal of organic phosphate esters by TiO2: effect of inorganic ions and humic acid. Chemosphere 206, 26–32. Tilson, H.A., Veronesi, B., McLamb, R.L., et al., 1990. Acute exposure to tris(2-chloroethyl) phosphate produces hippocampal neuronal loss and impairs learning in rats. Toxicol. Appl. Pharmacol. 106, 254–269. Wang, R., Tang, T., Lu, G., et al., 2018. Rapid debromination of polybrominated diphenyl ethers (PBDEs) by zero valent metal and bimetals: mechanisms and pathways assisted by density function theory calculation. Environ. Pollut. 240, 745–753. Wang, W., Qu, Y., Yang, B., et al., 2012. Lactate oxidation in pyrite suspension: a Fentonlike process in situ generating H2O2. Chemosphere 86, 376–382. Wei, G.-L., Li, D.-Q., Zhuo, M.-N., et al., 2015. Organophosphorus flame retardants and plasticizers: sources, occurrence, toxicity and human exposure. Environ. Pollut. 196, 29–46. Xu, X.X., Chen, J., Qu, R.J., et al., 2017. Oxidation of tris (2-chloroethyl) phosphate in aqueous solution by UV-activated peroxymonosulfate: kinetics, water matrix effects, degradation products and reaction pathways. Chemosphere 185, 833–843. Yang, S., Wang, P., Yang, X., et al., 2010. Degradation efficiencies of azo dye acid orange 7 by the interaction of heat, UV and anions with common oxidants: persulfate, peroxymonosulfate and hydrogen peroxide. J. Hazard. Mater. 179, 552–558. Ye, J.S., Liu, J., Li, C.S., et al., 2017. Heterogeneous photocatalysis of tris(2-chloroethyl) phosphate by UV/TiO2: degradation products and impacts on bacterial proteome. Water Res. 124, 29–38. Zhang, Y., Tran, H.P., Du, X., et al., 2017. Efficient pyrite activating persulfate process for degradation of p-chloroaniline in aqueous systems: a mechanistic study. Chem. Eng. J. 308, 1112–1119. Zhang, Y., Tran, H.P., Hussain, I., et al., 2015. Degradation of p-chloroaniline by pyrite in aqueous solutions. Chem. Eng. J. 279, 396–401.
The work was financially supported by the National Natural Science Foundation of China (Nos. 41573091, 41673090 and 41771346), the Guangdong Natural Science Funds for Distinguished Young Scholar (No. 2015A030306005), and the Tip-top Scientific and Technical Innovative Youth Talents of Guangdong Special Support Program (No. 2015TQ01Z233). Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2019.03.027. References Anastasio, C., Matthew, B.M., 2006. A chemical probe technique for the determination of reactive halogen species in aqueous solution: Part 2-chloride solutions and mixed bromide/chloride solutions. Atmos. Chem. Phys. 6, 2439–2451. Bonnissel-Gissinger, P., Alnot, M., Ehrhardt, J.J., et al., 1998. Surface oxidation of pyrite as a function of pH. Environ. Sci. Technol. 32, 2839–2845. Bu, L., Zhu, S., Zhou, S., 2018. Degradation of atrazine by electrochemically activated persulfate using BDD anode: role of radicals and influencing factors. Chemosphere 195, 236–244. Cao, S., Zeng, X., Song, H., et al., 2012. Levels and distributions of organophosphate flame retardants and plasticizers in sediment from Taihu lake, China. Environ. Toxicol. Chem. 31, 1478–1484. Cristale, J., Dantas, R.F., De Luca, A., et al., 2017. Role of oxygen and DOM in sunlight induced photodegradation of organophosphorous flame retardants in river water. J. Hazard. Mater. 323, 242–249. Das, T.N., 2001. Reactivity and role of SO5•- radical in aqueous medium Chain oxidation of sulfite to sulfate and atmospheric sulfuric acid generation. J. Phys. Chem. A 105, 9142–9155. Diao, Z.-H., Liu, J.-J., Hu, Y.-X., et al., 2017. Comparative study of rhodamine B degradation by the systems pyrite/H2O2 and pyrite/persulfate: reactivity, stability, products and mechanism. Sep. Purif. Technol. 184, 374–383. Feng, Y., Li, H.L., Lin, L., et al., 2018. Degradation of 1,4-dioxane via controlled generation of radicals by pyrite-activated oxidants: synergistic effects, role of disulfides, and activation sites. Chem. Eng. J. 336, 416–426. Fu, X.R., Gu, X.G., Lu, S.G., et al., 2015. Benzene depletion by Fe2+-catalyzed sodium percarbonate in aqueous solution. Chem. Eng. J. 267, 25–33. Gogate, P.R., Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment II: hybrid methods. Adv. Environ. Res. 8, 553–597. Hazime, R., Nguyen, Q.H., Ferronato, C., et al., 2014. Comparative study of imazalil degradation in three systems: UV/TiO2, UV/K2S2O8 and UV/TiO2/K2S2O8. Appl. Catal. B Environ. 144, 286–291. HoungAloune, S., Kawaai, T., Hiroyoshi, N., et al., 2014. Study on schwertmannite production from copper heap leach solutions and its efficiency in arsenic removal from acidic sulfate solutions. Hydrometallurgy 147, 30–40. Li, B.Z., Li, L., Lin, K.F., et al., 2013. Removal of 1,1,1-trichloroethane from aqueous solution by a sono-activated persulfate process. Ultrason. Sonochem. 20, 855–863. Liang, C., Guo, Y.-Y., Chien, Y.-C., et al., 2010. Oxidative degradation of MTBE by pyriteactivated persulfate: proposed reaction pathways. Ind. Eng. Chem. Res. 49, 8858–8864. Liu, J., Ye, J.S., Chen, Y.F., et al., 2018. UV-driven hydroxyl radical oxidation of tris(2chloroethyl) phosphate: intermediate products and residual toxicity. Chemosphere
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