Oxidative degradation of chlorophenolic compounds with pyrite-Fenton process

Environmental Pollution 247 (2019) 349e361

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Oxidative degradation of chlorophenolic compounds with pyrite-Fenton process* Cetin Kantar a, *, Ozlem Oral a, Ozge Urken a, Nilgun Ayman Oz a, Selda Keskin b a b

Canakkale Onsekiz Mart University, Department of Environmental Engineering, 17100, Canakkale, Turkey Nano Magnetics Instruments Ltd., 06510, Ankara, Turkey

a r t i c l e i n f o

a b s t r a c t

Article history: Received 15 November 2018 Received in revised form 4 January 2019 Accepted 6 January 2019 Available online 22 January 2019

Batch experiments, in conjunction with chromatographic and spectroscopic measurements, were performed to comparatively investigate the degradation of various chlorophenolic (CP) compounds (e.g., 2CP, 4-CP, 2,3-DCP, 2,4-DCP, 2,4,6-TCP, 2,3,4,6-TeCP) by a modified Fenton process using pyrite as the catalyst. The batch results show that the CP removal by pyrite-Fenton process was highly dependent on chemical conditions (e.g., pH, CP and pyrite concentration), CP type, number and location of chlorine atoms on the aromatic ring. With the exception of 2,3,4,6-TeCP and 2,3-DCP, the CP removal decreased with increasing the number of chlorine constituents. While the main mechanism responsible for monochlorophenol removal (e.g., 2-CP and 4-CP) was the hydroxyl radical attack on aromatic rings, the CP removal for multichlorophenolic compounds (e.g., 2,3,4,6-TeCP) was driven by both: (1) hydroxyl radical attack on aromatic rings by both solution and surface-bound hydroxyl radicals and (2) adsorption onto pyrite surface sites. The adsorption affinity increased with increasing the number of Cl atoms on the aromatic ring due to enhanced hydrophobic effect. The TOC removal was not 100% complete for all CPs investigated due to formation of chemically less degradable chlorinated intermediate organic compounds as well as low molecular weight organic acids such as formic and acetic acid. Spectroscopic measurements with SEM-EDS, zeta potential and XPS provided evidence for the partial oxidation of pyrite surface Fe(II) and disulfide groups under acidic conditions. © 2019 Elsevier Ltd. All rights reserved.

Keywords: Oxidation Pyrite-Fenton Dechlorination Chlorophenols

1. Introduction The removal of chlorophenols (CP) from wastewater has become one of the most critical and urgent topics in environmental research since many of CPs are known to be carcinogenic and mutagenic (Pera-Titus et al., 2004; Karci et al., 2012). Fenton process, based on the generation of highly reactive species such as hydroxyl radical (OH*), offers cost effective and highly efficient solutions to the treatment of biologically non-degradable CPs as an alternative to biodegradation (Benitez et al., 2001; Pera-Titus et al., 2004; Hong et al., 2008; Poerschmann et al., 2009; Weerasooriya et al., 2010; Munoz et al., 2011; Ortiz de la Plata et al., 2012; Bae et al., 2013; Liu et al., 2015). However, classical homogeneous Fenton process, using ferrous sulfate (FeSO4) as the catalyst, suffers

*

This paper has been recommended for acceptance by Wen Chen. * Corresponding author. Canakkale Onsekiz Mart University, Faculty of Engineering, Department of Environmental Engineering, Canakkale, Turkey. E-mail address: [email protected] (C. Kantar). https://doi.org/10.1016/j.envpol.2019.01.017 0269-7491/© 2019 Elsevier Ltd. All rights reserved.

from major limitations in full-scale applications, including high chemical cost, high sludge formation and scavenging of OH* radicals by excess use of Fe2þ ions (Khabbaz and Entezari, 2017). Alternatively, a limited number of studies have, primarily, focused on new improvements and/or modifications to lower the operating cost and optimize the efficiency of Fenton process (Munoz et al., 2011). For example, natural cost effective and reactive catalysts such as pyrite (FeS2) have been tested to oxidatively degrade chlorinated organics (Weerasooriya et al., 2010; Che et al., 2011; Che and Lee, 2011; Ortiz de la Plata et al., 2012; Bae et al., 2013; Khabbaz and Entezari, 2017). Pyrite is one of the highly abundant sulfide minerals in the environment, and exhibits both hydrophobic and hydrophilic surface properties depending on solution pH (Khabbaz and Entezari, 2017). Bae et al. (2013), for instance, observed that pyrite-catalyzed Fenton process was very effective in treating diclofenac. Liu et al. (2015) showed that Fenton process using high dispersive FeS2 on graphene oxide as the catalyst effectively degraded 4-chlorophenol under a wide pH range. Similarly, Che et al. (2011) and Che and Lee (2011) reported on selective redox degradation of chlorinated organic compounds

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(e.g., trichloroethylene) by pyrite-Fenton reaction. Their results showed that the degradation of chlorinated organic compounds with pyrite-Fenton process was highly dependent on chemical factors such as pyrite loading, pH and hydrogen peroxide concentrations. The reaction mechanism of organic compound oxidation with pyrite-Fenton process is not clearly established in the literature. However, it is postulated that the Fenton reaction may take place in solution with Fe(II) ions directly dissolved from pyrite surface as well as on pyrite surface with adsorbed hydroxyl (OH*) and peroxy (HO*2) radicals (Che et al., 2011; Che and Lee, 2011; Liu et al., 2015). In solution, hydrogen peroxide may attack pyrite surface, thereby leading to the dissolution of Fe(III) into solution: þ 2FeS2 þ 15H2 O2 ¼ 2Fe3þ þ 4SO2 4 þ 2H þ 14H2 O

(1)

The ferric iron (Fe3þ) released into solution by H2O2 as a result of Reaction (1) may re-oxidize the pyrite surface, while releasing Fe(II) into solution by strongly adsorbing onto pyrite disulfide (S2 2 ) group through a strong s-bond (Fe(II)eSeSeFe(III), where Fe(II) eSeS represents the pyrite surface) (Kantar et al., 2015a, b): 3þ

FeS2 þ 14Fe

2þ

þ 8H2 O ¼ 15Fe

þ

2SO2 4

þ 16H

þ

(2)

followed by solution phase hydroxyl (OH*) and peroxy (HO*2) radical generation in solution:

Fe2þ þ H2 O2 ¼ Fe3þ þ OH þ OH

(3)

Fe3þ þ H2 O2 ¼ Fe2þ þ HO2 þ Hþ

(4)

Alternatively, the Fenton reactions may also occur on pyrite surface with adsorbed OH* and peroxy (HO*2) radicals. A study by Bae et al. (2013) shows that pyrite-Fenton process was more effective in oxidizing diclofenac relative to conventional Fenton process. Similarly, Che et al. (2011) reported that while pyriteFenton process completely degraded trichloroethylene within less than 100 min of reaction time, classical Fenton process was much slower to oxidatively degrade TCE. Weerasooriya et al. (2010) observed that the pyrite oxidation occurred on both Fe(II) and disulfide (S2 2 ) sites, thereby resulting in oxidized surface Fe species (≡FeOOH) and sulfate via the formation of some reaction in* termediates such as S2O2 3 . The possible surface phase OH and peroxy (HO*2) radical formation may be postulated as follows (Zhang et al., 2014):

FeII S2 þ H2 O2 ¼ FeIII S2 þ ≡OHads þ OH

(5)

FeIII S2 þ H2 O2 ¼ FeII S2 þ ≡HO2ðadsÞ þ OH

(6)

Despite the fact that pyrite has been effectively used as the catalyst in Fenton or electro-Fenton treatment of a number of organic contaminants (e.g., ref. Che and Lee, 2011; Che et al., 2011; Zhang et al., 2014; Bae et al., 2013; Liu et al., 2015; Barhoumi et al., 2016), an extensive comparative study regarding the applicability of pyrite-Fenton process to the treatment of various chlorophenolic compounds is not available in the literature. A limited number of studies report that the treatability of CPs by chemical oxidation methods depends highly on the number and position of chlorine substituents on the phenolic ring (Huang et al., 1993; Pera-Titus et al., 2004; Wadley and Waite, 2005; Sharma et al., 2013). According to Weerasooriya et al. (2006), the presence of chlorine substituents on the aromatic ring enhanced CP adsorption onto pyrite surface. In an ozone oxidation study with chlorophenolic compounds, for instance, Trapido et al. (1997) found that the

degradation of CPs decreased in the order: 2,3,4,6-TeCP > 2,4,6TCP > 2,4-DCP > 2-CP > 4-CP. A study by Song-hu and Xiao-hua (2005), on the other hand, reported that the electro-Fenton process decomposed CP compounds in the decreasing order of: 2,4DCP > 2,4,6-TCP > 4-CP. Here, batch kinetic experiments, in conjunction with spectroscopic and chromatographic measurements, were performed to better understand and compare the performance of pyrite-Fenton process towards the degradation of various chlorophenolic compounds under variable chemical conditions (e.g., solid/liquid ratio, CP concentration, pH). The CPs used in the study include 2-chlorophenol (2-CP), 4-chlorophenol (4-CP), 2,3-dichlorophenol (2,3-DCP), 2,4-dichlorophenol (2,4-DCP), 2,4,6trichlorophenol (2,4,6-TCP) and 2,3,4,6-tetrachlorophenol (2,3,4,6TeCP). Some chemical properties of CPs used in the study are given in Table S1 (Supplementary Material). In addition, possible surface and solution phase reaction mechanisms involved in oxidative degradation of CPs by pyrite-Fenton process were also proposed. 2. Experimental 2.1. Materials Unless stated otherwise, all chlorophenolic (CP) compounds (2CP, 4-CP, 2,3-DCP, 2,4-DCP, 2,4,6-TCP, 2,3,4,6-TeCP), used in the experiments, were analytical grade, and purchased from SigmaAldrich. Hydrogen peroxide (30% by wt, H2O2), sulfuric acid, phenol, hydroquinone, p-benzoquinone, formic acid, acetic acid, maleic acid, oxalic acid, 1,1-phenantroline monohydrate, methanol, ammonium iron (II) sulfate hexahydrate were obtained from Merck (Germany). Tert-butanol (t-butanol), used as OH* scavenger, was purchased from Merck (Germany). n-Hexane and dichloromethane, purchased from Merck (Germany), were used as the extractants for the liquid-liquid extraction of chlorophenols and their oxidation products. The chemicals for chloride ion analysis (mercuric thiocyanate and ferric iron solutions) were purchased from Hach Lange, and directly used in the analysis. All solutions were prepared weekly, and stored in the refrigerator at 4  C before use. Natural pyrite samples were milled with a ceramic mortar and pestle, and sieved to <45 mm with no further purification. The BET surface area was 0.911 m2/g (Kantar et al., 2015a). 2.2. Degradation experiments Kinetic experiments were performed to determine CP removal rates by modified Fenton process under variable chemical conditions. The experiments were performed in 250 mL flasks containing 200 mL of solution at desired CP (100e500 mg/L), H2O2 (0.3 M) and pyrite (0.25e1 g/L) concentrations. The CP concentrations used in the current study are close to the range reported for CP concentrations in pharmaceutical and petrochemical wastewaters (Ramamoorthy and Ramamoorthy, 1997; Eslami et al., 2018). Similarly, the pyrite concentration lies in the range reported for pyrite-Fenton system by other researchers (e.g., Che et al., 2011). The initial pH of the suspensions was adjusted to a desired pH (pH 3, 4 or 5) using an appropriate amount of H2SO4. The suspensions were then placed on a shaker table in the dark at 275 rpm, and samples were taken at desired time intervals, filtered through 0.22 mm filters and analyzed for CP, total organic carbon (TOC), Fe2þ and total Fe concentrations. 0.1 mL of 1 M t-butanol was mixed with the samples to scavenge OH* radicals. Control experiments involving OH* and sulfate (SO*4) scavengers such as t-butanol, ethanol and chloride (Cl) were also conducted to identify which reactive oxygen species (OH* and SO*4) were involved in CP degradation by pyrite-Fenton system. While t-butanol was used as scavengers of both OH* and SO*4, the reaction rate of t-butanol with

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OH* was reported to be 3-orders of magnitude faster than that with SO*4 radical (Lian et al., 2017). Cl ion was shown to be scavenger of SO*4 radicals (Lian et al., 2017). As suggested by Diao et al. (2017), the scavenging effect of t-butanol for SO*4 radicals was tested with a molar ratio of t-butanol to CP (1000:1 and 500:1). Ethanol, on the other hand, was tested as OH* scavenger (Diao et al., 2017). 2.3. Surface characterization The pyrite samples used as the Fe catalyst in Fenton experiments were analyzed with X-Ray Photoelectron Spectroscopy (XPS) to determine oxidation products on pyrite surface during Fenton oxidation of CPs. Zeta potential measurements were taken with a Brookhaven ZetaPals zeta meter to measure surface zeta potentials under different chemical conditions. Scanning electron microscope (SEM) photographs of pyrite samples, in conjunction with energy dispersive spectroscopy (EDS), were taken with JEOL SEM- 7100EDX to provide data on surface morphology and elemental composition of pyrite surface. XRD (PANalytical Empyrean) analysis was conducted to identify the purity of pyrite used in the experiments and confirm surface oxidation products during Fenton oxidation of CPs. The analysis of natural pyrite samples with XRD suggested that pyrite had a very high purity (Fig. S1). The experimental procedure for surface analysis was provided by Kantar et al. (2015a, b). 2.4. Analytical procedures CPs and their aromatic (e.g., organic acids, catechol, hydroquinone, benzoquinone) and aliphatic (e.g., acetic acid, oxalic acid, formic acid) oxidation products were analyzed using HPLC (Shimadzu LC-20AT) equipped with SPD-M20A diode array detector and InterSustain C-18 column (4.6  150 mm). The operating conditions for HPLC analysis are provided in Table S2. The HPLC methods for the analysis of chlorophenolic compounds were established according to the manufacturer's instructions of InterSustain C-18 column. The mobile phases were 50% methanol/0.1% H3PO4 (Line A) and 100% methanol (Line B). The operating conditions for the gradient program were: 100% Line A for 10 min and a linear gradient evolution from 100% Line A to 20% Line A (80% Line B) in 20 min. The flowrate was 1 mL/min, and the injection volume was 40 mL. The column temperature was set to 30  C. The mobile phase used in the analysis of organic acids (e.g., formic acetic, oxalic acids) was 20 mm phosphate buffer adjusted to pH 2/acetonitrile (99:1, v/v). The measurement wavelength was 210 nm and the flow rate was 0.6 mL/min (Table S2). The reaction products of CP oxidation with pyrite-Fenton process were also analyzed with GC-MS using a procedure outlined by Karci et al. (2012) and Munoz et al. (2012). Prior to performing GCMS measurements, the reaction intermediates were extracted from samples using n-hexane. The samples were first acidified to pH 2 with concentrated H2SO4, and extracted three times with 15 mL nhexane. The extracted samples were then combined, and concentrated to 2 mL by evaporation at 40  C. The GC-MS analysis was performed with Termo Finnigan Trace GC Ultra equipped with an HP-5MS column (5% phenyl methyl silox) 30 m  0.25 mm x 0.25 mm capillary column using He as the carrier at a constant flow rate of 1 mL/min. The oven temperature program was 6 min at 80  C, 4  C/min to 180  C, and 10 min at 180  C. The temperatures for the injection port and ion source remained constant at 280 and 230  C, respectively. The MSD scan range of 50e700 amu was selected, and the WILEY 7 library was used to determine the structural assignment of the identified compounds. In addition, analytical standard samples were also run to confirm the compounds as suggested by the library. The reaction intermediates

351

were also qualitatively analyzed using LC-MS/MS (SHIMADZU LCMSMS-8040). The samples were directly injected into LC-MS/MS instrument. The mobile phases included 0.32 g formic acid, 2 mL acetic acid and 1 L water for mobile phase A, and 900 mL methanol with 100 mL mobile phase A for mobile phase B (isocratic elution mode). The flow rate was 0.5 mL/min, and the duration for sample analysis was approximately 1 min. The electrospray ionization mass spectrometer was used in the positive mode with full-scan mass spectra over the m/z range 10e400. Chloride ion was determined using spectrophotometric mercuric thiocyanate method (Hach Lange Method 20635e00). Hydrogen peroxide concentrations were measured with titanium sulfate method (APHA, 1995). The solution phase ferrous iron (Fe2þ) concentration was determined by means of spectrophotometric 1,10 phenanthroline method at 508 nm. The total Fe content of samples was analyzed with atomic absorption spectrophotometer (PG Instruments AA500F). The total organic carbon (TOC) measurements were done using HACH Lange IL 500 TOC/TN instrument. 2.5. Kinetic modeling The rate expression for the degradation of CPs by pyrite-Fenton system can be represented as follows:

d½CP ¼ k½CP dt

(7)

The integrated form of Eq. (7) is:

Ln½CP ¼ kt þ Ln½CPo 

(8)

where k is the pseudo-first order rate constant, and CPo is the initial CP concentration. 3. Results and discussion 3.1. Effect of chlorophenol type on CP removal by pyrite-Fenton process Fig. 1 shows a comparison of degradation kinetics of different CPs by Fenton system using pyrite as the catalyst at pH 3. These experiments contained a CP concentration of 100 mg/L, a pyrite concentration of 1 g/L and H2O2 concentration of 0.3 M. It is clear in

Fig. 1. The effect of CP type on CP removal by pyrite-Fenton process in systems containing 0.3 M H2O2, 1 g/L pyrite and 100 mg/L CP at pH 3. Solid lines represent model fit to experimental data using equation (8).

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Fig. 1 that the CP removal by pyrite-Fenton reaction followed a firstorder rate law, with first order rate constants and model statistics given in Table S3. The reaction rates were highly dependent on the type of CP used, with reaction rate constants decreasing in the order 2,3,4,6-TeCP > 4-CP > 2,3-DCP > 2-CP > 2,4-DCP > 2,4,6-TCP. This difference observed in CP removal rates among CPs is indicative of presence of multiple different processes responsible for CP removal by pyrite-Fenton process depending on the type of CP. As given in Fig. 1, with the exception of 2,3,4,6-TeCP and 2,3-DCP, the CP removal decreased with increasing the number of chlorine constituents. Of all CPs investigated, 2,3,4,6-TeCP exhibited the highest removal rates. In a study involving homogenous Fenton process, Tang and Huang (1995) observed a decreasing trend on CP removal with increasing chlorine substituents on the aromatic ring for 2-CP, 2,4-DCP and 2,4,6-TCP. Similarly, Song-hu and Xiao-hua (2005) reported that CP removal decreased with increasing chlorine substituents, with the exception of 4-CP, in the following order: 2,4-DCP > 2,4,6-TCP > PCP > 4-CP. According to these authors, the lowest removal observed in 4-CP system was caused by the competitive effect of hydroxyl attack dechlorination and electrochemical reductive dechlorination. Fig. 2 shows a comparison of CP removal and Cl release as a result of Fenton reaction for the CPs studied. The complete release of chlorine atoms from the aromatic rings (C/Co ¼ 1) coincides well with the complete removal of monochlorophenols (2-CP, 4-CP) from solution (Fig. 2A and B), implying that hydroxyl radical dechlorination was the major process controlling CP removal by pyrite-Fenton process for monochlorophenols. Note that one mole of chlorine atom was released into solution per mole of CP removed. The control experiment performed in system containing ethanol as the OH* radical scavenger suggested (Fig. S2) that the CP removal was much lower relative to non-scavenger containing systems, implying that the generation of solution phase OH* radicals significantly contributed to CP removal (especially for monochlorophenols) in pyrite-Fenton system. Similarly, the control experiments conducted with 0.31 and 0.62 M t-butanol show that the 2-CP degradation was significantly inhibited with the addition of t-butanol, and increasing t-butanol concentration from 0.31 to 0.62 M had a slight adverse effect on 2-CP removal. The addition of Cl as the SO*4 scavenger did not significantly affect CP degradation relative to non-scavenger containing systems. These suggest that both OH* and SO*4 radicals formed in the system, but the degradation reaction was mainly controlled by OH* radicals. This result is good agreement with the results of Diao et al. (2017) that report that OH* radicals generated in pyrite-Fenton system were the major oxidants for the degradation of organic substances. On the other hand, while complete removal of multichlorophenols from solution was achieved within less than 40 min (Fig. 2C, D, E and F), the chloride release rates from the aromatic ring were much slower, and took much longer times (e,g., 2,3,4,6TeCP). This suggests that, in addition to hydroxyl radical attack on aromatic ring, other processes were also involved in CP removal in systems involving multichlorophenolic compounds, including the adsorption of CPs onto pyrite surface sites and/or the transformation of CPs into some reaction intermediates with much stronger CeCl bonds, which are more difficult to break. Song-hu and Xiao-hua (2005) reported that not only does the number of chlorine atoms but also the location of chlorine substituents on the aromatic ring significantly contribute to CP removal in Fenton reactions. Similarly, according to Tsyganok and Otsuka (1999), the chlorine atoms located at 3- and 4-positions of the aromatic ring of chlorinated aromatic organic compounds were degraded faster than those at 2-position due to steric effect which may hinder the cleavage of CeCl bond at 2-position. As presented in Fig. 1, when comparing the monochlorophenols, the removal rate for 4-CP was

much faster than that for 2-CP. Similarly, Fig. 2 C and D show that, for dichlorophenols, the chlorine atoms located at 2 and 3 positions were removed from the aromatic ring much faster than the chlorine atoms at 2 and 4 positions. This implies that the Cl atoms on 2 and 3 positions for 2,3-DCP were simultaneously attacked by OH* radicals, and thus rapidly released into solution. On the other hand, the Cl atoms located at 2 and 4 positions for 2,4-DCP were much more resistant to OH* degradation, and thus slowly released into solution. According to Tsyganok and Otsuka (1999), the dechlorination of 2,4-D was postulated to take place at 4-position first, then followed by stepwise cleavage of chlorine atom at 2-position. These authors also report that the difference in Cl release from different positions is mainly driven by steric effect, rather than the difference in electron density around chlorine atoms at positions of 2, 3, and 4. Song-hu and Xiao-hua (2005) reported that the steric hindering effect was much higher for 2,4,6-TCP than that for 2,4-DCP. Despite the evidence given here, additional experiments need to be performed to more accurately determine the effect of Cl location in the aromatic ring on CP removal by pyrite-Fenton system. For 2,4,6-TCP, the Cl release and CP removal rates overlapped at reaction time less than 20 min, suggesting simultaneous cleavage of multiple Cl atoms from the aromatic ring (Fig. 2E). Simultaneous cleavage of more than one chlorine atom from polychlorinated organic compounds was also reported by Tsyganok and Otsuka (1999). However, at reaction time greater than 20 min, the CP removal was much higher than the Cl release rate, implying that another mechanism such as the binding onto pyrite may be responsible for CP removal. Pyrite may also provide surfaces for the binding of CPs since it contains sites which exhibit both hydrophobic and hydrophilic properties (Weerasooriya et al., 2010). Here, additional control experiments were conducted to better understand the effects of evaporation, H2O2 and adsorption on CP removal from solution. Fig. S3 shows that while evaporation and H2O2 had negligible effect on CP removal for all CPs studied, the addition of pyrite to systems containing CPs resulted in a significant increase in CP removal. It is clear that CPs adsorbed onto pyrite surface, and the intensity of binding increased with increasing number of chlorine atoms on the aromatic ring. Note that, except for 2,3,4,6-TeCP, following CP binding onto pyrite surface, the surface-bound CPs were gradually desorbed back into solution, indicating a weak interaction between pyrite and CPs. Unlike other CPs studied, once adsorbed onto pyrite surface sites, 2,3,4,6-TeCP was retained on the surface with no release of CP back into solution, indicating a strong affinity of 2,3,4,6-TeCP for pyrite surface sites (Fig. S3F). Despite the fact that 2,3,4,6-TeCP remained on pyrite surface, the Cl concentration continued to increase gradually in solution (Fig. 2), indicating the formation of possible some surface-bound reactions such as CP dechlorinization by surface-bound OH* radicals. Weerasooriya et al. (2006) observed that the presence of chlorine atoms on the aromatic ring enhanced CP removal by pyrite. Similarly, Weerasooriya et al. (2010) show that pyrite exhibits both hydrophobic and hydrophilic behavior, and the 4-CP binding onto pyrite surface occurs due to a hydrophobic force, bringing CP molecules closer to the surface sites for chemical binding. The octanol-water coefficients (Log Kow) for CPs studied are presented in Table S1. The CP binding onto pyrite as a function of Log Kow is shown in Fig. 3. As can be seen, the CP removal is closely related to LogKow values, providing further evidence that the binding of CPs onto pyrite is mainly driven by hydrophobic bonding. 3.2. Effect of pH on CP removal by pyrite-Fenton process Fig. 4 shows the pH dependence of degradation kinetics of five different CPs by Fenton system using pyrite as the catalyst. These

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Fig. 2. Comparison of CP removal from solution and Cl release from aromatic ring as a function of time in systems containing 0.3 M H2O2 and 1 g/L pyrite at pH 3 in the presence of 100 mg/L:(A) 2-CP (B) 4-CP (C) 2,3-DCP (D) 2,4-DCP (E) 2,4,6-TCP and (F) 2,3,4,6-TeCP.

experiments were performed in systems containing 100 mg/L CP, 1 g/L pyrite and 0.3 M H2O2 at pHs 3, 4 and 5. As shown in Fig. 4, the pH dependence of CP removal by pyrite-Fenton process followed a first order rate law, with rate constants presented in Table S4. For all CPs studied, the reaction rates decreased with increasing solution pH. The decrease in CP removal with increasing solution pH may be attributed to: (1) the decrease in iron (III) solubility, and (2) change in pyrite surface characteristics. The Fe solubility for both Fe(II) and Fe(III) species decreases with increasing solution pH (Kantar, 2016), and thus Fe (III)-(oxy)-hydroxides which form as a result of Fenton

reaction (Reactions 1 and 3) may accumulate on pyrite surface. The XRD results indicated the presence of Fe-oxide coatings such as goethite on pyrite surface upon exposure to H2O2 at pH 4 (Fig. S1). As shown in Fig. S1, the intensity of peaks decreased significantly in pyrite-Fenton system, implying the dissolution of Fe from pyrite surface and/or the transformation of pyrite surface to Fe-oxides. Gil-Lozano et al. (2017) reported on the formation of ferrihydrite and goethite on pyrite surface once pyrite was exposed to H2O2 under anoxic conditions. According to Weerasooriya et al. (2010), the pyrite oxidation occurs at both Fe(II) and S2 2 surface sites,

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Fig. 3. CP binding onto pyrite plotted in terms of the amount of CP bound per g of pyrite vs. octanol-water partition coefficient (LogKow). Note that CP ¼ 100 mg/L; pyrite ¼ 1 g/L and pH ¼ 3.

leading to the formation of ≡FeOOH and S2O2 3 in the presence of oxygen. Hydrogen peroxide attacks the pyrite surface, leading to release of Fe3þ into solution (Reaction 1). The Fe3þ released can re-oxidize the pyrite surface to produce Fe2þ ions in solution (Reaction 2) (Kantar et al., 2015a, b; Kantar et al., 2017). Fig. S4 shows that the amount of Fe species dissolved from pyrite increased with increasing reaction time, and nearly more than 95% of Fe dissolved was in Fe(III) form. However, the accumulation of solid phase Fe (III)(oxy)hydroxides in solution and on pyrite surface with increasing solution pH may lead to: (1) a decrease in Fe concentration in solution, thus hindering the generation of solution phase hydroxyl radicals (Reactions 2 and 3), and (2) a shift in pyrite surface characteristics from hydrophobic to hydrophilic behavior (Kocabag et al., 1990). As shown in Fig. S3, the binding ability of CPs increased with increasing the number of Cl substitutes on the aromatic ring, implying that, in addition to hydroxyl radical attack on aromatic ring, the binding onto pyrite surface through weak hydrophobic interaction may also play a significant role in CP removal from solution, especially for multichlorophenolic compounds (Weerasooriya et al., 2006). However, as the solution pH increases, the hydrophobic effect decreases since the pyrite surface becomes hydrophilic due to accumulation of Fe-oxides, and thereby, leading to a decrease in CP removal, especially for highly hydrophobic compounds such as multichlorophenolic compounds. The binding of CPs onto oxidized pyrite surfaces has also been proposed by other researchers. For instance, in a spectroscopic study, Weerasooriya et al. (2010) observed that 4-CP chemisorbed onto oxidized pyrite surface via the formation of an inner-sphere complex with negatively charged > FeO surface sites. The pKa values of CPs investigated are given in Table S1, and are observed to increase in the order of: 2,3,4,6-TeCP < 2,4,6-TCP < 2,3-DCP < 2,4DCP < 2-CP < 4-CP. Kantar et al. (2015a) found that while pyrite itself had a pHpzc < 3, the pHpzc of oxidized pyrite surface was 6, suggesting that the oxidized pyrite surface becomes positively charged at pH < 6, and thus making it possible for deprotonated negatively charged chlorophenolic compounds (especially at pHs < pKa) to chemically interact with positively charged pyrite surfaces. However, considering the high pKa values for CPs (Table S1), the concentration of deprotonated CPs, especially for monochlorophenols, would be expected to be very low under the experimental conditions studied (i.e., pH < 5), and the interaction

of CPs with oxidized pyrite surface sites would be expected to be negligible relative to hydrophobic interactions between CPs and pyrite surface.

3.3. Effects of pyrite dose and CP concentration on CP removal by pyrite-Fenton process Fig. 5 shows the effect of pyrite dose on CP removal kinetics with pyrite-Fenton process at an initial pH of 3, an initial H2O2 concentration of 0.3 M and pyrite loading ranging from 0.25 to 1 g/L. The reaction kinetics are highly dependent on pyrite dose, with rate constants increasing with increasing pyrite dose (Table S5). Pyrite surface site concentration (SOH) was estimated to be 3.52 mmol/g based on pyrite surface area of 0.911 m2/g as follows (Kantar et al., 2015a):

SOHT ¼

SA ns 1018 NA

(9)

where SA is the surface area (m2/g), NA represents Avogadro's number and ns is the average site density (2.33 sites/nm2). The surface site concentration for 1 g/L pyrite is estimated to be 3.52 mM, and that for 0.25 g/L pyrite is 0.88 mM (Table S5). Note that the surface site concentrations are much less than the CP concentrations used in the experiments (>5  104 M), implying that processes involving both solution and surface phase reactions occur very fast during pyrite-Fenton removal of CPs. This agrees well with some literature data obtained for pyrite as the catalyst in the Fenton process (Che et al., 2011; Zhang et al., 2014; Ammar et al., 2015; Khabbaz and Entezari, 2017). According to these authors, the dissolution of Fe(II) from pyrite increased significantly with increasing pyrite loading, leading to an enhanced OH* generation and thus improvement in Fenton efficiency. As shown in Fig. S5, the pyrite concentration-independent rate constants (log k1) can be estimated using the following equation (Kantar et al., 2015a):

logk ¼ logk1 þ clog½FeS2 

(10)

where logk is the pyrite concentration-dependent rate constant (min1), FeS2 represents molar pyrite concentration, and c is the slope related to the reaction degree with respect to pyrite concentration (M). As shown in Table S6, the first-order rate constants

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Fig. 4. Effects of solution pH on CP removal by pyrite-Fenton in systems containing 1 g/L pyrite, 0.3 M H2O2 and 100 mg/L CP: (A) 2-CP (B) 4-CP (C) 2,3-DCP (D) 2,4-DCP (E) 2,4,6-TCP, and (F) 2,3,4,6-TeCP. Solid lines represent model fit to experimental data using equation (8).

are closely correlated to pyrite concentration, and for all chlorophenols studied, the slope is observed to be less than 1, implying multiple different reactions occurring concurrently during CP removal by pyrite-Fenton process (Kantar et al., 2015a). The effects of CP concentration on CP removal by pyrite-Fenton process are shown in Fig. S6. The corresponding kinetic constants and model statistics are given in Table S5. The CP degradation rates decreased with increasing CP concentration, indicative of limited number of hydroxyl radical generation both on pyrite surface and in solution available for interaction with chlorophenols. While the CP concentration change from 100 to 200 mg/L had a minor effect on reaction kinetics for 4-CP, the impact of concentration change from

100 to 200 mg/L on reaction rate was much greater for 2,4-DCP. This is not surprising due to the fact while the 4-CP removal was mainly controlled by hydroxyl radical attack on aromatic ring, the 2,4-DCP removal was influenced by multiple different reaction mechanisms including hydroxyl radical attack on aromatic ring as well as CP binding onto pyrite surface sites via hydrophobic interactions, which are highly concentration dependent (Che et al., 2011; Zhang et al., 2014). Duesterberg and Waite (2006) and Che et al. (2011) reported that, at high organic loading, hydroxyl radicals generated during Fenton process were mostly consumed by organic compound.

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Fig. 5. Effect of pyrite dose on CP removal by pyrite-Fenton in systems containing 0.3 M H2O2 at pH 3 in the presence of 100 mg/L: (A) 2-CP (B) 4-CP (C) 2,3-DCP (D) 2,4-DCP (E) 2,4,6-TCP. Solid lines represent model fit to experimental data using equation (8).

3.4. Reaction intermediates Fig. S7 shows total organic carbon (TOC) removal as a function of time for experiments performed at pH 3, 0.3 M H2O2 and 1 g/L pyrite. In spite of 100% CP removal for all CPs within less than 30e40 min (Fig. 2), the TOC analysis suggested that a significant portion of TOC remained in solution, indicating the formation of less chemically degradable reaction intermediate species during pyrite-Fenton oxidation of CPs. The degree of TOC removal exhibited significant dependence on number and location of

chloride atoms in the CP structure. As presented in Fig. S7, the TOC removal rates for monochlorophenolic compounds (e.g., 4-CP) were much faster than those of multichlorophenolic compounds (e.g., 2,4,6-TCP). The TOC removal for 2,3,4,6-TeCP was about 40% due to the fact the 2,3,4,6-TeCP removal was mainly controlled by binding onto pyrite surface through hydrophobic effect (Fig. S3F). The reaction by products of CP degradation by Fenton process have been extensively studied by other researchers in the literature (e.g., Munoz et al., 2011, 2012). These studies suggest that Fenton oxidation of CPs may lead to the formation of some intermediate

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species including chlorinated organic compounds, PCBs, dioxin, dichlorophenyl ethers, biphenyls, catechol, dibenzofurans, phenol, hydroquinone, chlorohydroquinones, benzenediols and low molecular weight organic acids such as acetic acid and formic acids (e.g., ref. Poerschmann et al., 2009; Munoz et al., 2011, 2012; Karci et al., 2012; Liu et al., 2015). As shown in Fig. 2, the mass balance on chlorine shows that 100% of chlorine in the structure for monochlorophenols was released into solution in parallel with CP degradation rates, suggesting that chlorine atoms were fully attacked by OH* radicals. On the other hand, the chlorine mass balance was closed at 60e80% depending on CP type for multichlorophenols, implying the existence of condensation products containing chlorine in their composition. For instance, a study by Munoz et al. (2011) shows that 4-CP was first oxidized to some intermediate condensation products including p-benzoquinone, and hydroquinone by hydroxyl radicals, followed by further oxidation to low molecular weight organic acids such as oxalic, formic and acetic acids. Similarly, Munoz et al. (2012) reported on the formation of chlorine containing aromatic organic compounds including diphenyl ethers and dibenzofurans during Fenton oxidation of multichlorophenols such as 2,4-DCP. Here, additional experiments were performed to determine the types of reaction intermediates during pyrite-Fenton oxidation of 2,4-DCP. Here, the reaction intermediate species of 2,4-DCP degradation by pyriteFenton system was monitored in order to compare the results with other studies previously reported in the literature for 2,4-DCP degradation by Fenton process (e.g., Karci et al., 2012). Fig. 6 shows the LC-MS/MS spectrum under positive ionization mode for a sample taken at a reaction time of 120 min for the experiment performed with 500 mg/L 2,4-DCP, pH 3, 1 g/L pyrite and 0.3 M H2O2. The analysis of peaks in the chromatogram showed peaks related to species such as acetic acid (m/z ¼ 60), 4-chlorophenol (m/z ¼ 129), and 2,4,5-TeCP (m/z ¼ 199). The chromatogram also includes peaks for high molecular weight organic substances, corresponding to several two-ring aromatic chlorinated compounds as demonstrated by Munoz et al. (2012). GC-MS and HPLC analyses provided further evidence for the existence of benzenediols, 2,4,5TCP, phenol, 4-CP, chloroacetic acid and low molecular weight

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organic acids such as acetic, formic and maleic acids (Table 1). The transformation of 2,4-DCP to highly chlorinated organic compounds such as 2,4,5-TCP provides further evidence for the explanation of slow release of Cl relative to 2,4-DCP degradation as illustrated in Fig. 2. Evolution of low molecular weight organic acids during pyriteFenton oxidation of 2,4-DCP was followed with HPLC as shown in Fig. S8. Note that the concentration of acetic acid initially reached a peak concentration at 5 min of reaction time, and then gradually decreased down to zero at higher reaction times, implying that acetic acid may be transformed to CO2. At reaction times greater than 50. min, while the concentration of maleic acid decreased, the concentration of acetic acid was observed to increase again. This suggests that one of intermediate reaction products, possibly maleic acid, was further oxidized to acetic acid during pyrite-Fenton oxidation (Munoz et al., 2011). In a study with H2O2/UV-C, Fenton and photo-Fenton treatment of 2,4-DCP, Karci et al. (2012) reported on the formation of low molecular weight organic acids such as formic and acetic acid, hydroquinone, phenol, 4-chlorophenol, 3,5dichloro-2-hydroxybenzaldehyde, and 2,5-dichlorohydroquinone as the intermediate reaction by products. Formation of low molecular weight organic acids (e.g., oxalic, formic and acetic acids) were also reported for the Fenton oxidation of other CPs such as 4CP (Munoz et al., 2011; Liu et al., 2015). 3.5. Surface characterization SEM-EDS and EDS mapping images of natural pyrite used in the experiments are shown in Figs. S9 and S10, respectively. SEM images of pyrite indicate that the particles were not uniformly shaped, and the EDS image showed that natural pyrite contained Fe and S with very little Al as the impurity (Fig. S9). EDS mapping of natural pyrite shows a uniform distribution of Fe and S throughout the pyrite samples (Fig. S10). On the other hand, SEM-EDS images of pyrite exposed to 0.3 M H2O2 and 100 mg/L 2,4,6-TCP at pH 3 present peaks related to Fe, C, O and S, confirming the oxidation of pyrite surface as well the binding of CP and/or reaction intermediates onto pyrite surface (Fig. 7). The existence of O in the

Fig. 6. LC-MS/MS spectrum under positive ionization mode for a sample taken at a reaction time of 120 min from batch experiment performed with 1 g/L pyrite, 500 mg/L 2,4-DCP and 0.3 M H2O2 at pH 3.

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Table 1 Products of 2,4-DCP oxidation by pyrite-Fenton analyzed with GC-MS and HPLC. Intermediate

Analytical technique

Retention time (min)

Molecular weight

m/z fragments

4-chloro-1,3-benzenendiol 2,4,5-trichlorophenol Chloroacetic acid Phenol, 2,4-bis (1,1-dimethylethyl) 4-chlorophenol Acetic acid Formic acid Maleic acid

GC-MS GC-MS GC-MS GC-MS GC-MS HPLC HPLC HPLC

5.41 15.04 19.29 20.49 9.72 6.72 4.09 8.96

144 196 95 206 128 60 46 116

144,142,146,114,116 196,198,200,132,134,160,97 95,96,97 206,191, 128,130,100

Fig. 7. SEM-EDS images of pyrite exposed to 0.3 M H2O2 in systems containing 100 mg/L 2,4,6-TCP and 1 g/L pyrite at pH 3.

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spectra is also indicative of accumulation of surface oxidation products (e.g., Fe-oxides) on pyrite surface, as also suggested by XRD measurements given in Fig. S1. EDS mapping image (Fig. S11) shows an uneven distribution of elements (e.g., Fe, C and O) on pyrite surface. Zeta potential measurements of pyrite samples prepared in 200 mL solution at pH 3 showed a zeta potential value of 21.87 mV (Fig. S12). However, in systems containing pyrite and 0.3 M H2O2,

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the surface became positively charged, confirming the oxidation and formation of Fe-oxides on pyrite surface. In system containing 0.3 M H2O2 and 100 mg/L 2,4-DCP, the surface potential slightly decreased from þ14 mV to þ11 mV, suggesting the binding of CP and/or oxidation products on pyrite surface. The XPS Fe 2p spectra of pyrite surfaces exposed to 0.3 M H2O2 at pH 3 in the presence of 100 mg/L CP are given in Fig. 8A, C and E for 2,4-DCP, 2,3,4,6-TeCP and 4-CP, respectively. The Fe spectra for

Fig. 8. Fitted Fe 2p and S 2p X-ray photoelectron spectra (XPS) of pyrite (1 g/L) exposed to 0.3 M H2O2 at pH 3 in systems containing (A, B) 100 mg/L 2,4-DCP (C, D) 100 mg/L 2,3,4,6TeCP and (E, F) 100 mg/L 4-CP.

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all CPs analyzed show similar peaks at 703.76e707.88 eV for Fe(II) eS and at 708.08e709.31 eV for Fe(II)eO. The presence of Fe(II) sites on pyrite surface is indicative of partial oxidation of Fe sites and/or continuous generation of new surface sites during Fenton process (Kantar et al., 2015a, b; Kantar et al., 2017). Che et al. (2011), Zhang et al. (2014) and Kantar et al. (2015a, b) reported on the importance of Fe redox cycling for the generation of new sites and sustainable Fenton operations (Reaction 2). As given in Fig. 8, however, no peaks for Fe(III)eO were observed in any sample analyzed due to the fact that the sample preparation was done at pH 3 at which Fe(III) was readily released into solution due to very high solubility of Fe(III) species under acidic conditions, as consistent with the batch experiments given in Fig. S4. Similarly, the S 2p spectra of pyrite samples exposed to 0.3 M H2O2 and 100 mg/L CP at pH 3 are given in Fig. 8B, D and F for 2,4-DCP, 2,3,4,6-TeCP and 4-CP, respectively. The samples exhibit disulfide peaks (S2 2 ) at 162e163.7 eV, polysulfides (S2 n ) at 164.1e165.15 eV, monosulfide (S2) at 161.5e162 eV and sulfate (SO2 4 ) at 168.2e169.43 eV, indicating that the disulfide groups (S2 2 ), initially present on pyrite surface, were partially oxidized to sulfate by Fenton process via the formation of some unstable intermediates such as polysulfides (Kantar et al., 2015b). Zhao et al. (2017) reported that S(-II) had a pronounced impact on Fenton process by accelerating Fe(III)/Fe(II) redox cycling and generating hydroxyl radicals via activation of H2O2. The analysis of pyrite surface for C 1s shows peaks for CeC bond at 284e285.2 eV, CeO bond at 285.5e286.8 eV and OeC]O bond at 288.2e289 eV for CPs (e.g., 2,4-DCP, 2,3,4,6-TeCP and 4-CP) (Fig. S13). This implies that CPs and/or reaction intermediates interacted with pyrite surface sites via various functional groups such as carboxylic groups. The presence of CeO bond may be explained through the attachment of aliphatic carbon with oxygen as ether group. The OeC]O bonds are indicative of carboxylic groups, which may come from the binding of intermediate reaction products such as low molecular weight organic acids (e.g., acetic acid) onto pyrite surface sites. Fig. S14, shows that, of all the CPs investigated, only 2,3,4,6-TeCP showed FeeCl peak at 200.2 eV and CeCl peak at 201.9 eV in the Cl 2p spectra, confirming the high binding affinity of 2,3,4,6-TeCP for pyrite surface sites, as consistent with the results of batch experiments given in Fig. S3F. 4. Conclusions Oxidative dechlorination has been shown to be a viable and cost effective process for dealing with a particularly persistent class of contaminants (e.g., CPs) often found in industrial wastewaters such as pharmaceuticals. Here, laboratory batch experiments were performed to comparatively understand the degradation of various chlorophenolic (CP) compounds (e.g., 2-CP, 4-CP, 2,3-DCP, 2,4-DCP, 2,4,6-TCP, 2,3,4,6-TeCP) using a Fenton process with pyrite as the catalyst. The batch results show that the CP removal by pyriteFenton process was highly affected by chemical conditions (e.g., pH, pyrite and CP concentration), CP type, number and location of chlorine atoms on the aromatic ring. CP removal decreased with increasing solution pH due to accumulation of surface oxidation products (e.g., Fe-oxides) on pyrite surface. Decreasing pyrite dose led to a decrease in CP removal since pyrite surface had limited number of surface sites, available for interaction with H2O2 and CPs. While the main mechanism responsible for the removal of monochlorophenolic compounds by pyrite-Fenton system was the hydroxyl radical attack on aromatic rings, the multichlorophenolic compounds were degraded by both hydroxyl radical attack on aromatic ring as well as adsorption onto pyrite surface sites. The binding affinity of CPs onto pyrite surface increased with increasing the number of Cl atoms on the aromatic ring due to enhanced

hydrophobic effect. LC-MS/MS, GC-MS and HPLC measurements suggested that the CPs were oxidized to some chlorinated intermediate organic compounds as well as low molecular weight organic acids such as formic and acetic acid. Spectroscopic measurements performed with SEM-EDS, zeta potential and XPS provided further evidence that the pyrite surface Fe(II) and disulfide groups were partially oxidized upon exposure to H2O2 under the experimental conditions of batch experiments (e.g. pH ¼ 3). Overall, our results suggest that pyrite-Fenton system can be used as a viable process for the degradation of biologically non-degradable organic substances as a pretreatment process prior to biological degradation. Acknowledgments This study was fully funded by the Scientific and Technological Research Council of Turkey (TUBITAK) under a grant # 115Y329. We'd also like to acknowledge Canakkale Onsekiz Mart University, Center for Environmental Research and Application for HPLC analysis. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.envpol.2019.01.017. References Ammar, S., Oturan, M.A., Labiadh, L., Guersalli, A., Abdelhedi, R., Oturan, N., Brillas, E., 2015. Degradation of tyrosol by a novel electro-Fenton process using pyrite as heterogeneous source of iron catalyst. Water Res. 74, 77e87. APHA, 1995. Standard Methods for the Examinations of Water and Wastewater, nineteenth ed. American Public Health Association, Washington, DC. Bae, S., Dongwook, K., Lee, W., 2013. Degradation of diclofenac by pyrite catalyzed Fenton oxidation. Appl. Catal. B Environ. 134e135, 93e102. Barhoumi, N., Oturan, N., Olvera-Vargas, H., Brillas, E., Gadri, A., Ammar, S., Oturan, M.A., 2016. Pyrite as a sustainable catalyst in electro-Fenton process for improving oxidation of sulfamethazine. Kinetics, mechanism and toxicity. Water Res. 94, 52e61. Benitez, F.J., Beltran-Heredia, J., Acero, J.L., Rubio, F.J., 2001. Oxidation of several chlorophenolic derivatives by UV irradiation and hydroxyl radicals. J. Chem. Technol. Biotechnol. 76, 312e320. Che, H., Lee, W., 2011. Selective redox degradation of chlorinated aliphatic compounds by Fenton reaction in pyrite suspension. Chemosphere 82, 1103e1108. Che, H., Bae, S., Lee, W., 2011. Degradation of trichloroethylene by Fenton reaction in pyrite suspension. J. Hazard Mater. 185, 1355e1361. Diao, Z.-H., Liu, J.-J., Hu, Y.-X., Kong, L.-J., Jiang, D., Xu, X.-R., 2017. Comparative study of Rhodamine B degradation by the systems pyrite/H2O2 and pyrite/persulfate: reactivity, stability, products and mechanism. Separ. Purif. Technol. 184, 374e383. Duesterberg, C.K., Waite, T.D., 2006. Process optimization of Fenton oxidation using kinetic modeling. Environ. Sci. Technol. 40, 4189e4195. Eslami, A., Hashemi, M., Ghanbari, F., 2018. Degradation of 4-chlorophenol using catalyzed peroxymonosulfate with nano-MnO2/UV radiation: toxicity assessment and evaluation for industrial wastewater treatment. J. Clean. Prod. 195, 1389e1397. Gil-Lozano, C., Davila, A.F., Losa-Adams, E., Fairen, A.G., Gago-Duport, L., 2017. Quantifying Fenton reaction pathways driven by self-generated H2O2 on pyrite surfaces. Sci. Rep. 7 (43703), 1e11. Hong, S.-H., Kwon, B.-H., Lee, J.-K., Kim, H.-K., 2008. Degradation of 2-chlorophenol by Fenton and photo- Fenton processes. Kor. J. Chem. Eng. 25 (1), 46e52. Huang, C.P., Dong, C., Tang, Z., 1993. Advanced chemical oxidation: its present role and potential future in hazardous waste treatment. Waste Manag. 13, 361e377. Kantar, C., 2016. Role of low molecular weight organic acids on pyrite dissolution in aqueous systems: implications for catalytic chromium (VI) treatment. Water Sci. Technol. 74 (1), 99e109. Kantar, C., Ari, C., Keskin, S., Dagaroglu, Z.G., Karadeniz, A., Alten, A., 2015a. Cr(VI) removal from aqueous systems using pyrite as the reducing agent: batch, spectroscopic and column experiments. J. Contam. Hydrol. 174, 28e38. Kantar, C., Arı, C., Keskin, S., 2015b. Comparison of different chelating agents to enhance reductive Cr(VI) removal by pyrite treatment procedure. Water Res. 76, 66e75. Kantar, C., Bulbul, M.S., Keskin, S., 2017. Role of humic substances on Cr(VI) removal from groundwater with pyrite. Water Air Soil Pollut. 228 (48), 1e11. Karci, A., Alaton, I.A., Hanci, T.O., Bekbolet, M., 2012. Transformation of 2,4dichlorophenol by H2O2/UV-C, Fenton and photo-Fenton processes: oxidation

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