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Influence of UV irradiation on the toxicity of chlorinated water to mammalian cells: Toxicity drivers, toxicity changes and toxicity surrogates
Water Research 165 (2019) 115024
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Influence of UV irradiation on the toxicity of chlorinated water to mammalian cells: Toxicity drivers, toxicity changes and toxicity surrogates Wen-Cheng Huang a, Ye Du b, Min Liu a, Hong-Ying Hu b, c, Qian-Yuan Wu d, *, Ying Chen a, ** a
College of Architecture and Environment, Sichuan University, Chengdu, 610065, PR China Shenzhen Environmental Science and New Energy Technology Engineering Laboratory, Tsinghua-Berkeley Shenzhen Institute, Shenzhen, 518055, PR China Environmental Simulation and Pollution Control State Key Joint Laboratory, State Environmental Protection Key Laboratory of Microorganism Application and Risk Control (SMARC), School of Environment, Tsinghua University, Beijing, 100084, PR China d Key Laboratory of Microorganism Application and Risk Control of Shenzhen, Guangdong Provincial Engineering Research Center for Urban Water Recycling and Environmental Safety, Graduate School at Shenzhen, Tsinghua University, Shenzhen, 518055, PR China b c
a r t i c l e i n f o
a b s t r a c t
Article history: Received 12 August 2019 Accepted 22 August 2019 Available online 23 August 2019
UV irradiation was reported to be able to degrade some kinds of DBPs, yet its influence on the toxicity of chlorinated water to mammalian cells remains unknown. This study systematically investigated the influence of low-pressure UV irradiation on the DBPs and toxicity of chlorinated drinking water (DW) and reclaimed water (RW). The apparent first-order rate constant (kobs) of degradation kinetics of known DBPs increased with the increased Br substitutions. Haloacetonitriles were identified as toxicity drivers among the detected DBPs, which even contributed more to the toxicity after UV irradiation, mainly due to the refractory bromochloroacetonitrile (BCAN) and dichloroacetonitrile (dCAN). Both total organic halogen, cytotoxicity and genotoxicity were significantly removed under UV irradiation, with the removal rate of 22.9%e41.7% for cytotoxicity and a higher rate of 33.1%e55.5% for genotoxicity under 2400 mJ/cm2 irradiation. UV irradiation significantly decreased the UV254, SUVA254 and fluorescence intensity (FLU) of chlorinated water. Results from high performance size exclusion chromatography revealed that chlorinated DW mainly contained high molecular weight (MW) compounds (>1000 Da) while chlorinated RW mainly contained lower MW compounds (100e500 Da). Chromophores and fluorophores in compounds of 100e500 Da increased in chlorinated DW while decreased in chlorinated RW under UV irradiation. Both the removal of UV254, SUVA254, FLU, MW-based UV254 (>1000 Da) and MWbased FLU (each fractions) were significantly correlated (p < 0.05) with the removal of toxicity under UV irradiation. The UV254 of chlorinated water was recommended as the optimal surrogate for toxicity removal. © 2019 Elsevier Ltd. All rights reserved.
Keywords: Chlorination UV irradiation Toxicity driver Cytotoxicity Genotoxicity Surrogate
1. Introduction Chlorination is a major disinfection method used widely both in drinking water and reclaimed water treatment processes. However, disinfection by-products (DBPs) would generate during the
* Corresponding author. Division of Energy and Environment, Graduate School at Shenzhen, Tsinghua University, Room 1810, Shenzhen, 518055, PR China. ** Corresponding author. College of Architecture and Environment, Sichuan University, Chengdu, 610065, PR China. E-mail addresses: [email protected] (Q.-Y. Wu), [email protected] (Y. Chen). https://doi.org/10.1016/j.watres.2019.115024 0043-1354/© 2019 Elsevier Ltd. All rights reserved.
chlorination process (Du et al., 2017a; Hua and Reckhow, 2007). To date, more than 700 kinds of DBP have been reported (Richardson and Postigo, 2011). Despite so many DBPs having been identified, only four kinds of trihalomethanes (THMs), five kinds of haloacetic acids (HAAs), chlorite and bromate are regulated (Guilherme and Rodriguez, 2014; Richardson et al., 2007). In recent years, many unregulated emerging DBPs have been reported and studied, such as haloacetonitriles (HANs), halonitromethanes (HNMs), haloacetamides (HAcAms) and halogenated aromatic DBPs (Richardson et al., 2007; Jiang et al., 2017), while the majority of DBPs still remain unidentified. More than 100 kinds of DBPs have proved to be cytotoxic and genotoxic through in vitro assay
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(Wagner and Plewa, 2017), and around 20 kinds of DBPs was found to be carcinogenic by in vivo assay (Richardson et al., 2007; Li and Mitch, 2018) The risk of DBPs is therefore a major concern during water chlorination. The risk posed by DBPs can be controlled by reducing the concentration. Generally, DBP concentrations can be controlled by precursor removal, disinfection process optimization, and elimination after formation (Doederer et al., 2014; Hua and Reckhow, 2007; Wang et al., 2013). In the supply system for drinking water or reclaimed water, secondary water supply tanks are usually needed to store and supply water to users. Recently, some studies showed that UV irradiation could degrade DBPs such as HAAs and THMs (Bu et al., 2018; Xiao et al., 2014; Zalazar et al., 2007). There is a possibility to eliminate the DBP level in secondary water supply tanks by UV irradiation right before use. However, previous studies mainly focused on several typical DBPs or DBPs in synthetic water, the degradation of overall DBPs generated in real water by UV photolysis needs further investigation. Among many kinds of DBPs, the regulated DBPs like THMs are most widely studied owing to their widespread presence and high concentrations in chlorinated water samples. However, only focusing on the high-concentration DBPs might be insufficient to guarantee the water safety. Because many emerging DBPs, especially nitrogenous DBPs (N-DBPs) (Muellner et al., 2007) are high toxic contaminants. For example, the emerging DBP bromochloroacetonitrile (BCAN) is 1000 times more cytotoxic than tCM (Wagner and Plewa, 2017). Identifying the toxicity drivers in known DBPs by both considering the concentration and toxic potency would be more scientifically significant in risk control. Moreover, due to the different degradation efficiency of DBPs under UV irradiation, it is therefore also necessary to understand toxicity drivers in chlorinated water after UV irradiation. Although more than 700 DBPs have been identified, known DBPs comprise only a small fraction of all DBPs, generally accounting for 30% of the total organic halogen (TOX) (Kristiana et al., 2015). Therefore, changes of TOX under UV irradiation would give a full description of DBP degradation. Furthermore, the concern lies behind the DBPs is their toxic effects. The incomplete dehalogenation of DBPs does not necessarily decrease the toxicity. For example, during the dehalogenation of trichloroacetic acid, monochloroacetic acid might form (Liu et al., 2017), which is even more toxic than trichloroacetic acid (Wagner and Plewa, 2017). Similarly, for haloacetonitriles and haloacetamides, less halogen substitution generally leads to higher toxicity (Wagner and Plewa, 2017). Toxicity elimination is the ultimate goal of DBP control. Cytotoxicity tests are mostly used to analyze the lethal effects of contaminants (Du et al., 2017b; Sayess et al., 2017). Some DBPs might not cause direct cell death, but impair the genetic materials at low concentration levels (McKie et al., 2015; Zheng et al., 2015). Results of cytotoxicity and genotoxicity would give a more comprehensive understanding about the effects of DBP control by UV irradiation, yet few studies clarified the toxicity changes of chlorinated water under UV irradiation. Cytotoxicity and genotoxicity are ideal indexes to characterize the water safety. However, the drawback is the cost of time and money during toxicity assay. When degrading DBPs, UV irradiation alters the water quality indexes at the same time. Finding surrogates for the toxicity changes would be more convenient for the toxicity regulation. In previous studies, some water parameters could be used as surrogates for toxicity formation during water chlorination. Such as UV254 absorbance can use as a surrogate for the increase of cytotoxicity (Du et al., 2017b); The formation of antiestrogenic activity can be predicted by fluorescence (FLU) intensity (Tang et al., 2014). However, little research clarified surrogates for the toxicity removal of chlorinated water, especially for cytotoxicity
removal and genotoxicity removal under UV irradiation. The main objective of this study is thusly to understand the effects of UV irradiation on toxicity removal of DBPs in chlorinated drinking water and reclaimed water. Degradation kinetics of DBPs, toxicity drivers in chlorinated water before and after UV irradiation were investigated. Changes in TOX, cytotoxicity, and genotoxicity of chlorinated water under UV irradiation were investigated. Furthermore, water quality parameters and molecular weight (MW) distribution were measured to analyze their correlation with toxicity and identify suitable surrogates for indicating toxicity removal. 2. Materials and methods 2.1. Chemical and reagents Organic solvents, including methanol (MeOH), acetone, dichloromethane, and methyl tert-butyl ether (MTBE), were all of HPLC grade and purchased from J.T. Baker (U.S.A.). NaOH (99%), NaClO (99%), Na2SO4 (99%), and NaS2O3 (99%) were obtained from Macklin (China). Eleven DBPs were detected in chlorinated water samples in this study, including trichloromethane (tCM), bromodichloromethane (BdCM), bromoform (tBM), dibromochloromethane (dBCM)), trichloronitromethane (tCNM), dibromoacetonitrile (dBAN), bromochloroacetonitrile (BCAN), dichloroacetonitrle (dCAN), trichloroacetamide (tCAcAm), chloral hydrate (CH), and 1,1,1-trichloropropanone (1,1,1-tCP). Suppliers and purities of all DBP standards are shown in Table S2. For the toxicity assay, dimethyl sulfoxide (DMSO) (99.7%) and fetal bovine serum (FBS) were purchased from Fisher Scientific (Pittsburgh, PA). Dulbecco's modified Eagle's medium/nutrient mixture F-12 (DMEM/F-12, 1:1) and penicillinestreptomycin were purchased from Gibco (USA). Cell Counting Kit-8 (CCK-8) for the cytotoxicity test was obtained from Dojindo Molecular Technologies (Japan). Paraformaldehyde was purchased from Sigma-Aldrich (USA). Triton X-100 and Albumin Bovine V were purchased from Solarbio (China). P-Histone H2A.X (S139) rabbit antibody and antirabbit IgG Fab2 Alexa Fluor 647 molecular probes were obtained from Cell Signaling Technology (USA). Hoechst 33342 was purchased from Biotium (USA). 2.2. Water sampling Simulated drinking water (DW) was prepared by dissolving Suwannee River natural organic matter (SRNOM 2R101; International Humic Substances Society) in ultrapure water. Reclaimed water (RW) samples A and B were collected from two different reclaimed water treatment plants. All water samples were filtered through 0.45-mm microfilters and stored at 4 C. Water quality parameters are shown in Table S1. Details of DOC and UV254 absorbance analyses are provided in Text S1. Threeedimensional excitation emission matrix of fluorescence spectra (FEEM) was measured using a fluorophotometer (F-7000, Shimadzu). The determination and data processing method for FEEM is described in detail elsewhere (Chen et al., 2003). 2.3. Chlorination The water samples were first adjusted to pH 7 ± 0.05 using H2SO4 or NaOH (2 M), stabilized with 1 mM phosphate buffer (pH 7), and then chlorinated with 10 mg/L NaClO (expressed as Cl2 concentration). After chlorination in the dark for 24 h, residual free chlorine was quenched with 105% of the stoichiometric amount of NaS2O3.
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2.4. UV irradiation The chlorinated samples were exposed to a quasi-collimated beam with a wavelength of 254 nm, produced by two 41 W lowpressure Hg lamps (LightSources, Orange, CT, USA). The 250 mL cylindrical reactor (diameter, 10 cm) was placed vertically below the lamps. The UV dose applied to the test solution during photodegradation experiments was determined to be 1.20 ± 0.05 mW/ cm2 using iodideeiodate chemical actinometry, as described by Bolton and Linden (2003). The reactor containing the test solution was stirred at 350 rpm throughout the experiment. The irradiation time under the highest UV dose (2400 mJ/cm2) condition was around 20 min. The temperature of water samples under UV irradiation was maintained at 27.5 ± 1.0 C. After various UV irradiation doses, water samples were collected for analysis. 2.5. Solid phase extraction (SPE) SPE was used as the pretreatment for TOX and toxicity assays. The water samples were first adjusted to pH 2 ± 0.05 using H2SO4 (2 M) and then concentrated using 6 mL Oasis hydrophilicelipophilic balance resin cartridges (Waters, USA) that had been activated with methanol (10 mL) and ultrapure water (10 mL). The water samples were passed through the cartridges at 3 mL/min. After adsorption was finished, the resin was eluted with methanol (5 mL), acetone (2 mL), and dichloromethane (2 mL). Finally, the extracts were fully dried under nitrogen flow. 2.6. Determination of DBPs and TOX The DBPs were enriched (ten-fold) by liquideliquid extraction (LLE) of the treated water samples, which had already been adjusted to pH 2 ± 0.05. MTBE was used as an extraction agent, containing 1,2-dibromopropane (100 mg/L) as the internal standard. The samples were then shaken for 5 min and allowed to stand for 10 min. GC-ECD (Agilent, USA) equipped with a DB-5MS capillary column (30 m 0.25 mm 0.25 mm, Agilent, USA) was used to measure the DBPs, using nitrogen as the carrier gas. The injector and detector temperatures were 190 and 290 C, respectively. The initial column temperature was 35 C (maintained for 9 min), which was increased to 40 C at a rate of 2 C/min (maintained for 1 min), then 160 C at a rate of 10 C/min (maintained for 2 min), and finally 220 C at a rate of 40 C/min (maintained for 2 min). The TOX content, including total organic chlorine (TOCl) and total organic bromine (TOBr) contents, was analyzed as the index of the overall halogenated DBP content. After SPE, the extracts were dissolved in ultrapure water and irradiated by vacuum UV (VUV) lamps (185 nm, 12 W, UV Tec, Co., China) according to the method of Bu et al. (2018) to convert organic halogens into inorganic halides. The Cl and Br concentrations were determined by ion chromatography (ICS-900, ThermoFisher Scientific, USA) after irradiation. The TOX content was calculated from the difference between inorganic halide contents before and after VUV irradiation. 2.7. Toxicity index calculation Cytotoxicity index and genotoxicity index were calculated using the toxic potency data reported by Wagner and Plewa (2017). In the literature, the cytotoxicity index (CTI) of the detected DBPs was calculated based on the concentration at which the relative cell viability was 50% (LC50) using a crystal violet cytotoxicity assay. The genotoxicity index (GTI) of the detected DBPs was calculated using the SCGE 50% Tail DNA (TDNA) value obtained using a single cell gel electrophoresis (SCGE) genotoxicity assay (Wagner and Plewa, 2017). Both cytotoxicity and genotoxicity assays used Chinese
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Hamster Ovary (CHO) cell line as the subject. The LC50 and 50% TDNA values of individual DBPs are shown in Table S3, as reported by Wagner and Plewa (2017). Assuming that the toxicity of individual DBPs was additive, the CTI and GTI were weighted for toxic potency by dividing the DBP concentrations by the LC50 or 50% TDNA values (Chuang et al., 2019). 2.8. Molecular weight (MW) distribution transformation Transformations in the molecular weight distributions of chlorinated water samples after UV irradiation were characterized by high-performance size exclusive chromatography (HPSEC) (LC20, Shimadzu, Japan) using two tandem columns (Tsk-Gel G3000PWXL and Tsk-Gel G2500PWXL, Japan) with an UV array detector, fluorescence (FLU) detectors, and a modified total organic carbon detector (Sievers900, GE, USA). Details of this procedure can be found in previous studies (Wang et al., 2019; Wu et al., 2016). 2.9. Cell culture The Chinese Hamster Ovary (CHO-k1) cell line, which is widely used as a mammalian cell subject in toxicity assays (Du et al., 2018a; Plewa et al., 2010), was obtained from the American Type Culture Collection to perform the toxicity assay. Cells were cultured with DMEM/F-12 (1:1) supplemented with 10% fetal bovine serum, 100 U/mL penicillin, and 100 mg/L streptomycin. The cell dishes were placed in an incubator with 5% CO2 gas at a high humidity and 37 C. Cells passaging was performed every two days. Cells with 2e5 passages were used for the toxicity assay. 2.10. Cytotoxicity assay CCK-8 was used to detect cytotoxicity in this experiment, using highly water-soluble tetrazolium salt 2-(2-methoxy-4nitrophenyl)-3-(4-nitrophenyl)-5-(2,4-disulfophenyl)-2H-tetrazolium monosodium salt (WST-8) to produce a water-soluble formazan dye after reduction with intracellular dehydrogenase. The 450 nm absorbance of the dye showed a good linear relationship with the number of living cells (Fig. S1), as measured using a SpectraMax i3 system (Molecular Devices, USA). After SPE, the extracts of water samples were dissolved in DMEM/ F-12 mixed with 0.5% DMSO as test samples. The cytotoxicity of the phenol solution was measured simultaneously and used as a reference compound to quantify the cytotoxicity equivalents. Cells were preincubated for 12 h for adherent growth, then exposed to contaminants for 48 h. Each test was performed with 3e4 replicates. The detailed cytotoxicity assay procedure is provided in Text. S2. 2.11. Genotoxicity test The phosphorylated H2AX (pH2AX) was used as a genotoxicity endpoint, representing the DNA double-strand breaks (DSBs) (Bonner et al., 2008), which is widely used in genotoxicity assays (Audebert et al., 2010; Redon et al., 2011; Watters et al., 2009). For the genotoxicity assay, cells were seeded at 5 103 per well. After preincubation for 12 h and exposure for 24 h, the cells were fixed, permeated, and blocked. The cells were then treated with primary antibody P-Histone H2A.X (S139) rabbit antibody and stained with secondary antibody anti-rabbit IgG Fab2 Alexa Fluor 647 molecular probe and Hoechst 33258 together. Finally, cell images were obtained using an HCA system (ImageXpress Micro; Molecular Devices, USA). Each test was performed with 4e6 replicates. The pH2AX values per cell for which the cell viability was higher than 70% were used to obtain concentration-response curves. Details can be found in Wu et al. (2019).
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2.12. Statistical analysis DBP degradation under UV irradiation was fitted by first-order reaction kinetics. For cytotoxicity assay, the relative cell viability (against the negative control) v.s. different concentration factors were used to obtain concentration-response curves by regression analysis. LC50 values were used to calculate the cytotoxicity equivalents (phenol). For genotoxicity, the pH2AX induction rate (against the negative control) v.s. different concentration factors were used to obtain concentration-response curves. The concentration of the 1.5-fold induction rate (IR1.5) of pH2AX per well was used to calculate the genotoxicity equivalents (4-NQO). One-way analysis of variance (ANOVA) was used to determine whether there was a significant decrease in sample toxicity after UV irradiation (p < 0.05). To find the surrogates for toxicity removal under UV irradiation, linear fitting was applied to the removal rate of toxicity and water quality parameter and the significant correlation was considered when p < 0.05.
3. Results and discussion 3.1. Degradation of detected DBPs under UV irradiation Concentration of the detected DBPs after chlorination and their changes during UV irradiation are shown in Figs. S2ae2d. The formation of THMs during chlorination were highest in both RW and DW compared with other DBPs, especially in DW, which was similar to the previous result (Krasner et al., 2016). TCP, CH, dCAN, and tCAcAm were also found in chlorinated DW, at concentrations
of 10.8, 8.0, 5.2, and 1.7 mg/L, respectively. Concentrations of nitrogenous DBPs such as haloacetonitriles (HANs) and haloacetamide in chlorinated RW were higher than chlorinated DW, which might be because RW contain more organic nitrogen (Hu et al., 2016). Owing to the presence of Br, brominated DBPs (BrDBPs) were found in RW, such as BdCM and BCAN. Except tCP and CH, all DBPs showed significant degradation under UV irradiation. In DW, almost all of tCAcAm was removed, tCM and dCAN reduced 85.8% and 30% in 2000 s (2400 mJ/cm2). The removal rates of tCM in RW were around 58%, which was a little lower compared with in DW. The tCNM, which only found in RW, was also significantly removed by 91.1% and 80.1% in RW A and RW B. The removal rate of Br-DBPs were higher than there chlorinated counterparts. For example, BCAN reduced by 55% in RW while DCAN only reduced by 20%e30%. Fig. 1d shows the apparent first-order rate constants of degradation (kobs) of the detected DBPs in DW and RW (kinetics curves of DBPs in each sample are shown in Fig. 1aec. Br-DBPs degraded faster than their chlorinated counterparts under UV irradiation. The kobs values increased with the number of Br substitutions increasing. For example, the degradation order was tBM > CdBM > tCM and dBAN > BCAN > dCAN. Other DBPs, such as haloacetic acids, THMs and halophenolic also followed this degradation rule, which was reported previously (Wang et al., 2017a,b; Xiao et al., 2014; Liu et al., 2019). The carbonehalogen bonds in DBPs, such as CeBr and CeCl, can be broken by sufficient internal energy through absorption of the photon of UV light to an excited state (Pandit et al., 2016). The photon absorption capacity of DBPs depends on their molar
Fig. 1. Degradation kinetics of DBPs in different water samples by UV irradiation. Degradation kinetics of different DBPs in (a) DW, (b) RW A and (c) RW B. (d) Comparison of apparent first order rate constant (kobs) of degradation of DBPs in different water samples. (#: Data not available).
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extinction coefficient (ε). Higher ε values lead to a faster degradation in DBPs through direct UV photodegradation (Fang et al., 2013). Each substitution of a Br can increase the ε value of trihalogenated DBPs by 240 ± 74 M1cm1 (Chuang et al., 2016), indicating that more Br substitution leads to an increase in the photodegradation rate of DBPs, as confirmed by our results. In a previous study, dCAN was found to hardly degrade in ultrapure water under UV direct photolysis with a UV intensity of 6.23 mW/cm2 (Hou et al., 2017), which was much higher than that in this study (1.2 mW/cm2). However, the kobs values of dCAN in DW and RW in this study were 0.132e0.170 103 s1, indicating that direct UV photolysis was not the only degradation pathway for DBPs in DW and RW. Hydroxyl radical (OH$) and other radicals have been found to play an important role in the degradation of some DBPs, such as THMs and 2,4-bromophenol (Luo et al., 2019; Xiao et al., 2014). In RW and DW, triplet dissolved organic matter (3DOM*) and reactive oxygen species (ROS) can be generated through UV irradiation of dissolved organic matter (DOM) (Jacobs et al., 2011; Vione et al., 2006). These radicals can react with eBr, eCN, CeH, and other bonds, eventually leading to bond cleavage and reconstruction (Wang et al., 2018), which contributes to DBP
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degradation. Compared with Cl substitution, Br substitution improved the reactivity of DBPs with OH$ (Chuang et al., 2016). This also explained why the kobs values of DBPs were higher with the increased Br substitution.
3.2. Toxicity drivers in detected DBPs before and after UV irradiation To determine which DBPs posed the greatest risk among the detected DBPs before and after UV irradiation, by using the toxic potency data of individual DBPs reported by the same group (Wagner and Plewa, 2017), cytotoxicity index (CTI) and genotoxicity index (GTI) of the detected DBPs to CHO cells were calculated. The CTI of detected DBPs in chlorinated RW and chlorinated DW were significantly decreased by UV irradiation, with 32.1%, 42.2% and 51.0% removal from DW, RW A and RW B, respectively, after UV irradiation at 2400 mJ/cm2. Clearly, before UV irradiation, HANs contributed most of the CTI in all three water samples (Fig. 2a and b). In DW, dCAN was the main cytotoxicity driver, with its contribution ratio to the CTI increasing from 86.0% to 89.4% after UV irradiation. The remaining CTI of less than 15% was contributed by
Fig. 2. Changes in the toxicity index under UV irradiation: (a) the decrease of CTI, (b) changes in the contribution of DBPs to CTI, (c) the decrease of GTI and (d) changes in the contribution of DBPs to GTI.
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tCM, CH, and tCAcAm. The regulated tCM only contributed 8.6% of the CTI initially, decreasing to 3.1% after UV irradiation (2400 mJ/ cm2). Before UV irradiation, HANs had a greater effect on the CTI in RW (more than 97%), which contained dCAN, BCAN, and dBAN, contributed 38.8%, 52.6% and 5.8% of CTI in chlorinated RW A, 25.4% 55.1% and 17.6% of CTI in chlorinated RW B, respectively. When the UV dose reached 2400 mJ/cm2, 97% of the CTI was contributed by dCAN and BCAN. Furthermore, the contributions of tCM to the CTI were less than 1% in RW. The results showed a decrease in the GTI of detected DBPs similar to that of the CTI (Fig. 2c and d). UV irradiation significantly decreased the GTIs of DW, RW A, and RW B, by 35.5%, 61.2%, and 65.4%, respectively. The initial contribution ratio of the GTI of HANs to DW, RW A, and RW B was 91.4%, 69.7%, and 83.6%, respectively, all of which increased to more than 90% after irradiation with a UV dose of 2400 mJ/cm2. BCAN and dCAN remained the major toxicity drivers during UV photolysis. TCNM made a different contribution to the GTI of untreated RW A and RW B, at 27.9% and 15.2%, respectively. However, owing to the rapid degradation of TCNM, its contribution to the GTI was less than 10% after UV irradiation. HANs were the major toxicity drivers among the detected DBPs, especially dCAN and BCAN. Although dBCN was more toxic than the other two (Plewa et al., 2002), the amount of dBCN generated during chlorination was small and it degraded much faster under UV irradiation. BCAN and dCAN were not only highly toxic, but also had slow degradation rates under UV irradiation. Therefore, the contribution ratios of BCAN and dCAN to the toxicity index increased during UV irradiation. THMs, the regulated DBPs, had the highest concentrations among the detected DBPs after chlorination, and did make some contribution to the CTI in DW before irradiation (no more than 10%). However, owing to the much lower toxicity and higher degradation rates of THMs compared with those of BCAN and dCAN, their contribution ratio to the toxicity index decreased during UV irradiation. In RW, the concentration ratio of HANs among the detected DBPs was much higher than that in DW, such that the toxicity index contribution of THMs was negligible. Previous study also implied that HANs were the toxicity drivers of known DBPs in drinking water (Plewa et al., 2017). However, after UV irradiation, this study for the first time revealed that HANs even contributed more to the toxicity index, mainly resulted from the recalcitrant BCAN and dCAN under UV irradiation. These results suggested that except for UV irradiation, other treatments might be needed to further eliminate the HANs to guarantee the water safety.
suggested that brominated DBPs were more easily degraded than chlorinated DBPs under UV irradiation. 3.4. Removal of cytotoxicity and genotoxicity As shown above, although it was found that HANs were toxicity drivers among the detected DBPs, the toxicity index from all the detected DBPs remained very low (only at 103 level, the level of 1 means LC50). Only focusing on the toxicity of a few known DBPs is not enough. Evaluation of changes in cytotoxicity and genotoxicity of all the DBPs under UV irradiation are therefore necessary. The concentration-response curves of cytotoxicity for each sample are shown in Fig. 4aec. The curves shifted to the right as the UV dose increased, leading to an increased LC50, which indicated decreased cytotoxicity. The cytotoxicity equivalent was significantly reduced by UV irradiation (Fig. 4d), by 22.9%, 29.8%, and 41.7% (p < 0.05) for chlorinated DW, RW A, and RW B, respectively. The rate of cytotoxicity decrease was fast in the first 600 mJ/cm2, but gradually slowed thereafter. The cytotoxicity assay reflected the macroscopic hazards of DBPs on CHO cells, while the genotoxicity assay evaluated the effects of DBPs on cellular genes. In this study, pH2AX was measured to characterize the DNA double-strand breaks (DSBs), which is the most severe form of DNA damage (Khanna and Jackson, 2001; Mills
3.3. Removal of TOX by UV irradiation The known DBPs generally only accounted for 30% of the total halogenated DBPs (Krasner et al., 2006). The TOX content is widely used to indicate the total halogenated DBP content (Du et al., 2018b; Jiang et al., 2017; Wu et al., 2019). Therefore, to comprehensively understand the removal of DBPs under UV irradiation, changes in TOX contents were investigated. Fig. 3 shows the effect of UV irradiation on the TOX contents in each water sample. The initial TOCl concentration in DW, RW A, and RW B was 315.8, 212.3, and 116.7 mg/L respectively. After UV irradiation, the TOCl contents of all chlorinated water samples were significantly decreased, by 25.9%, 35.1% and 53.6% for RW A, RW B and DW, respectively. TOBr was only found in RW, at about 10 mg/L, with more than 60% degradation under UV irradiation, the removal rate of which was higher than that of TOCl. As the TOX comprises a mixture of different DBPs, it might be improper to fit the degradation kinetics. However, by plotting the C/C0 vs. UV dose, it was found that TOBr clearly degraded faster than the TOCl, which was similar to the degradation of detected DBPs. All these results
Fig. 3. Effect of UV irradiation on the total organic halogen (TOX). (a) TOX concentration, (b) TOX degradation ratio.
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et al., 2003). Fig. 5a and Fig. S4 shows the images of pH2AX foci formed in cell nucleus when cells were exposed to chlorinated water (before and after UV irradiation, UV dose ¼ 2400 mJ/cm2). There was a significant decrease of pH2AX foci numbers after the UV irradiation of chlorinated water. The IR1.5 value increased with increasing UV dose, representing the decrease in genotoxicity under UV irradiation (Fig. 5b and Fig. S5). Compared with the reduction in cytotoxicity, UV irradiation had a greater influence on the removal of genotoxicity (Fig. 5c). The initial genotoxicity of chlorinated RW was higher than that of chlorinated DW. UV irradiation significantly reduced the genotoxicity by 33.1%, 46.5%, and 55.5% (p < 0.05) for chlorinated DW, RW A and RW B, respectively. The trend in the reduction rate of genotoxicity was similar to that of cytotoxicity, showing gradual slowing as the UV dose increased. In conclusion, UV irradiation could significantly reduce the cytotoxicity and genotoxicity of chlorinated water. 3.5. Effect of UV irradiation on the quality of chlorinated water Changes in DOC content, UV254 absorbance and FLU values of chlorinated water were measured. The DOC contents of the three chlorinated water samples did not change much during UV irradiation (Fig. S6), showing that UV irradiation was difficult to mineralize the organic matters. UV irradiation significantly reduced the UV254 and specific UV absorbance at 254 nm (SUVA254) of chlorinated DW and chlorinated RW (Fig. S7 and Fig. S8), which represented the reduction in aromaticity, double bonds and other unsaturated bonds (Singer, 1999; Uyguner and Bekbolet, 2005). FEEM was divided into five regions according to a previous study (Chen et al., 2003). DW contained more fulvic
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humic substances (Region III) while RW contained more tryptophan-like proteins (Region II) (Figs. S9eS11). After UV irradiation, the FLU values of all chlorinated water samples first increased slightly and then decreased (Fig. S12). The organic matters in water have been substantially oxidized during chlorination, which led to the reduction in FLU values. When water samples were irradiated with UV, the generated hydrated electron (e(aq)) are generated, which can partially reduce the oxidized organic matters, causing the FLU value to increase (Neta et al., 1988). Then, with increasing UV dose, the FLU value decreased due to further degradation of organic matters. 3.6. Changes of MW distribution under UV irradiation Figs. 6, S13, and S14 show the spectra of MW distribution of chlorinated DW and chlorinated RW under UV irradiation, as determined by HPSEC coupled with three tandem FLU, UV and TOC detectors. To quantitatively determine the changes in the different MW fractions of the organic matter, the chromatograms were deconvoluted into Gaussian peaks according to a previous study (L. Wang et al., 2017). Chromatograms were divided into four fractions based on the apparent MW, namely F1 (<100 Da), F2 (100e500 Da), F3 (500e1000 Da) and F4 (>1000 Da). Signals from all the three detectors suggested that DW mainly contained high MW F4 (>1000 Da) while RW mainly contained lower MW F2 (100e500 Da). This was in accordance with previous study that organic matters in RW consists of more low MW compounds (Hu et al., 2016). The HPSEC-FLU showed that, changes in the FLU (Ex/Em ¼ 240/ 360 nm) of chlorinated DW first increased and then decreased, with
Fig. 4. (a)e(c): Concentration-response curves for cytotoxicity assay of sample under UV irradiation (a) chlorinated DW, (b) chlorinated RW A and (c) chlorinated RW B; (d) Changes in cytotoxicity equivalents of chlorinated water under UV irradiation.
8
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this change mainly attributed to 100e500 Da compounds (Fig. S13b). HPSEC-FLU tended to decrease with increasing UV dose in chlorinated RW, especially for F2 (100e500 Da) (Fig. 6b). Therefore, it seems that fluorophores in F2 (100e500 Da) were most susceptible to UV irradiation. The HPSEC-TOC showed that, TOC content in all three samples did not change much during UV irradiation, but showed significant transformations between different MW fractions. High MW F4 (>1000 Da) decreased while F3 (500e1000 Da) increased, which mainly occurred under the first 300 mJ/cm2 UV irradiation (Fig .6c). The HPSEC-UV showed that, UV254 absorbance of all three water samples showed a downward trend under UV irradiation. In chlorinated DW, the decrease in UV254 absorbance was mainly due to high MW F4 (>1000 Da) (Fig. S13d). Simultaneously, the UV254 absorbance of the low MW F2 (100e500 Da) increased, which might be attributed to the transformation of high MW compounds into low MW compounds. For chlorinated RW, the decreased UV254 mainly resulted from the fractions with MW > 100 Da (Fig. 6d and Fig. S14d), suggesting that chromophores in F2eF4 were more susceptible to UV irradiation (Uyguner and Bekbolet, 2005). 3.7. Correlation between removal of toxicity and water quality parameters To identify surrogates of toxicity removal during UV irradiation, when there was a significant linear correlation (p < 0.05) between toxicity removal and changes in water quality parameters, the
coefficient of determination (R2) was used to evaluate the degree of relevance (Taylor, 1990). The data are summarized in Table 1. No significant correlation was observed between removal of DOC content and the toxicity, because UV irradiation cause little mineralization of the organic matters. Removal of UV254 and SUVA254 showed good correlation with removal of cytotoxicity (p ¼ 0.003, R2 ¼ 0.602 and p ¼ 0.002, R2 ¼ 0.619, respectively) and genotoxicity (p < 0.001, R2 ¼ 0.757 and p < 0.001, R2 ¼ 0.744, respectively) (Figs. S15aeS15d). The FLU values from five regions were significantly correlated with the removal of cytotoxicity and genotoxicity, especially the total FLU value (p < 0.001, R2 ¼ 0.816 and p < 0.001, R2 ¼ 0.793, respectively) (Figs. S15e and S15f). The strong correlation between toxicity removal and the changes in FLU value, UV254 and SUVA254 indicated that toxicity removal from chlorinated water during UV irradiation occurred simultaneously with the oxidation of unsaturated bonds. By evaluating the relationship between the MW distribution and the toxicity removal rate, the UV254 absorbance of F4 (>1000 Da) was found to be strongly correlated (p < 0.001) with the toxicity removal rate, with R2 values of 0.761 and 0.844 for cytotoxicity and genotoxicity, respectively. No strong correlation between the UV254 absorbance of fractions with MW < 1000 Da and toxicity removal was found. For MW-based FLU, except F1 (<100 Da), all fractions were significantly correlated with toxicity removal. The MW-based TOC cannot serve as surrogates for toxicity removal as no significant correlation was found in all the fractions. Considering the total FLU value would first increase, it may be
Fig. 5. (a) Images of pH2AX foci in cell nucleus when cells were exposed to chlorinated RW B (before and after UV irradiation, UV dose ¼ 2400 mJ/cm2, 30 folds concentration). (b) Concentration-response curves for genotoxicity assay of chlorinated RW B under UV irradiation. (c) Changes in genotoxicity equivalents of chlorinated water under UV irradiation.
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9
Fig. 6. (a) Effects of UV irradiation on MW distribution spectra of chlorinated RW B. (b)e(c): Transformation of each MW fractions of chlorinated RW B: (b) FLU, (c) TOC and (d) UV254.
Table 1 Correlation between removal of water quality parameters and toxicity (n ¼ 12). Cytotoxicity Slope
Genotoxicity Intercept
R2
P
Slope
Intercept
R2
P
TOC
N/A
>0.05
N/A
UV254
0.712
0.047
0.602
0.003
1.028
0.069
0.757
< 0.001
SUVA
0.609
0.089
0.619
0.002
0.861
0.116
0.744
< 0.001
I II III IV V
0.658 0.564 1.117 0.401 1.056 0.861
0.204 0.173 0.266 0.247 0.369 0.228
0.641 0.427 0.615 0.545 0.476 0.816
0.002 0.021 0.003 0.006 0.011 < 0.001
0.89 0.682 1.51 0.545 1.372 1.094
0.298 0.265 0.381 0.358 0.515 0.332
0.708 0.376 0.654 0.608 0.484 0.793
< 0.001 0.034 0.001 0.003 0.012 < 0.001
> 1000 500e1000 100e500 <100 Total
0.637 N/A N/A N/A 1.068
0.051
0.761
< 0.001 >0.05 >0.05 >0.05 < 0.001
0.865 N/A 0.480 N/A 1.354
0.092
0.844
0.289
0.376
< 0.001 >0.05 0.036 >0.05 < 0.001
FLU
> 1000 500e1000 100e500 <100 Total
0.306 0.317 1.117 N/A 0.289
0.01 0.034 0.003 >0.05 0.039
0.433 0.409 0.312 N/A 0.404
TOC
>1000 500e1000 100e500 <100 Total
N/A N/A N/A N/A N/A
>0.05 >0.05 >0.05 >0.05 >0.05
N/A N/A N/A N/A N/A
FLU volume
MW
Region Region Region Region Region Total UV254
*N/A: p > 0.05, not applicable.
0.073
0.797
0.142 0.195 0.266
0.498 0.377 0.615
0.201
0.359
>0.05
0.115
0.772
0.209 0.288 0.334
0.6 0.379 0.377
0.294
0.422
0.003 0.033 0.034 >0.05 0.022 >0.05 >0.05 >0.05 >0.05 >0.05
10
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improper to indicate the toxicity removal when UV dose was less than 300 mJ/cm2. Besides, the availability of SUVA254 also requires the measurement of UV254. Therefore, changes in UV254 was the best surrogate for toxicity removal during UV irradiation of chlorinated water. UV254 of F4 (>1000 Da) and MW-based FLU were not recommended to be surrogates due to the inconvenient availability and relatively low R2, although they were significantly correlated with toxicity removal. 4. Conclusions Influences of UV irradiation on chlorinated water were systematically investigated in this study. Degradation kinetics of known DBPs and their toxicity indexes were evaluated. Changes in TOX contents, cytotoxicity and genotoxicity of chlorinated water were studied. The correlation between removal of toxicity and water quality parameters were also evaluated to identify surrogates for toxicity removal. The main conclusions were as follows: (1) For the same class of DBPs, increase in the number of Br substitutions increased the degradation kobs under UV irradiation. HANs were the toxicity drivers in the detected DBPs in chlorinated water, which even contributed more to the toxicity index after UV irradiation, with the refractory BCAN and dCAN being the main contributor. (2) UV irradiation significantly decreased the TOX in chlorinated water. Under the UV dose of 2400 mJ/cm2, TOCl was eliminated by 25.9%e53.6% while TOBr was eliminated with a higher ratio of around 60%. Both cytotoxicity and genotoxicity were significantly removed under UV irradiation, with the removal rate of 22.9%e41.7% for cytotoxicity equivalent and a higher rate of 33.1%e55.5% for genotoxicity equivalents. (3) UV irradiation showed little impact on DOC of chlorinated water, but significantly reduced the UV254 and SUVA254. FLU values first increased then followed by a decrease under UV irradiation. Results of HPSEC revealed that chlorinated DW mainly contained high MW compound (>1000 Da) while chlorinated RW mainly contained lower MW compound (100e500 Da). Chromophores and fluorophores in F2 (100e500 Da) increased in chlorinated DW while decreased in chlorinated RW under UV irradiation. (4) Both the removal of UV254, SUVA254, FLU (total and five regions), MW-based UV254 (>1000 Da) and MW-based FLU were significantly correlated with the removal of toxicity in chlorinated water under UV irradiation. The UV254 was recommended as the optimal surrogate because of the highest R2 value. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgments This study was supported by National Natural Science Foundation of China (No. 51738005/51678332), the Shenzhen Science, Technology and Innovation Commission (No. JCYJ20170818091859147), special support program for high-level personnel recruitment in Guangdong Province (2016TQ03Z384), and the Development and Reform Commission of Shenzhen Municipality.
Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.watres.2019.115024. References Audebert, M., Riu, A., Jacques, C., Hillenweck, A., Jamin, E.L., Zalko, D., Cravedi, J.P., 2010. Use of the gH2AX assay for assessing the genotoxicity of polycyclic aromatic hydrocarbons in human cell lines. Toxicol. Lett. 199, 182e192. Bolton, J.R., Linden, K.G., 2003. Standardization of methods for fluence (UV Dose) determination in bench-scale UV experiments. J. Environ. Eng. 129, 209e215. Bonner, W.M., Redon, C.E., Dickey, J.S., Nakamura, A.J., Sedelnikova, O.A., Solier, S., Pommier, Y., 2008. gH2AX and cancer. Nat. Rev. Cancer 8, 957e967. Bu, Y., Wang, L., Chen, B., Niu, R., Chen, Y., 2018. Effects of typical water components on the UV254 photodegradation kinetics of haloacetic acids in water. Separ. Purif. Technol. 200, 255e265. Chen, W., Westerhoff, P., Leenheer, J.A., Booksh, K., 2003. Fluorescence excitationemission matrix regional iIntegration to quantify spectra for dissolved organic matter. Environ. Sci. Technol. 37, 5701e5710. Chuang, Y.-H., Parker, K.M., Mitch, W.A., 2016. Development of predictive models for the degradation of halogenated disinfection byproducts during the UV/H2O2 advanced oxidation process. Environ. Sci. Technol. 50, 11209e11217. Chuang, Y.H., Szczuka, A., Mitch, W.A., 2019. Comparison of toxicity-weighted disinfection byproduct concentrations in potable reuse waters and conventional drinking waters as a new approach to assessing the quality of advanced treatment train waters. Environ. Sci. Technol. 53, 3729e3738. , M.J., Pidou, M., Weinberg, H.S., Gernjak, W., 2014. Rejection of Doederer, K., Farre disinfection by-products by RO and NF membranes: influence of solute properties and operational parameters. J. Membr. Sci. 467, 195e205. Du, Y., Lv, X.T., Wu, Q.Y., Zhang, D.Y., Zhou, Y.T., Peng, L., Hu, H.Y., 2017a. Formation and control of disinfection byproducts and toxicity during reclaimed water chlorination: a review. J. Environ. Sci. 58, 51e63. Du, Y., Wu, Q.Y., Lu, Y., Hu, H.Y., Yang, Y., Liu, R., Liu, F., 2017b. Increase of cytotoxicity during wastewater chlorination: impact factors and surrogates. J. Hazard Mater. 324, 681e690. Du, Y., Wu, Q.Y., Lv, X.T., Wang, Q.P., Lu, Y., Hu, H.Y., 2018a. Exposure to solar light reduces cytotoxicity of sewage effluents to mammalian cells: roles of reactive oxygen and nitrogen species. Water Res. 143, 570e578. Du, Y., Wu, Q.Y., Lv, X.T., Ye, B., Zhan, X.M., Lu, Y., Hu, H.Y., 2018b. Electron donating capacity reduction of dissolved organic matter by solar irradiation reduces the cytotoxicity formation potential during wastewater chlorination. Water Res. 145, 94e102. Fang, J.Y., Ling, L., Shang, C., 2013. Kinetics and mechanisms of pH-dependent degradation of halonitromethanes by UV photolysis. Water Res. 47, 1257e1266. Guilherme, S., Rodriguez, M.J., 2014. Occurrence of regulated and non-regulated disinfection by-products in small drinking water systems. Chemosphere 117, 425e432. Hou, S., Ling, L., Shang, C., Guan, Y., Fang, J., 2017. Degradation kinetics and pathways of haloacetonitriles by the UV/persulfate process. Chem. Eng. J. 320, 478e484. Hu, H.Y., Du, Y., Wu, Q.Y., Zhao, X., Tang, X., Chen, Z., 2016. Differences in dissolved organic matter between reclaimed water source and drinking water source. Sci. Total Environ. 551e552, 133e142. Hua, G., Reckhow, D.A., 2007. Comparison of disinfection byproduct formation from chlorine and alternative disinfectants. Water Res. 41, 1667e1678. Jacobs, L.E., Fimmen, R.L., Chin, Y., Mash, H.E., Weavers, L.K., 2011. Fulvic acid mediated photolysis of ibuprofen in water. Water Res. 45, 4449e4458. Jiang, J., Zhang, X., Zhu, X., Li, Y., 2017. Removal of intermediate aromatic halogenated DBPs by activated carbon adsorption: a new approach to controlling halogenated DBPs in chlorinated drinking water. Environ. Sci. Technol. 51, 3435e3444. Khanna, K.K., Jackson, S.P., 2001. DNA double-strand breaks: Signaling, repair and the cancer connection. Nat. Genet. 27, 247e254. Krasner, S.W., Weinberg, H.S., Richardson, S.D., Pastor, S.J., Chinn, R., Sclimenti, M.J., Onstad, G.D., Thruston, A.D., 2006. Occurrence of a new generation of disinfection byproducts. Environ. Sci. Technol. 40, 7175e7185. Krasner, S.W., Kostopoulou, M., Toledano, M.B., Wright, J., Patelarou, E., Kogevinas, M., Villanueva, C.M., Carrasco-Turigas, G., Marina, L.S., Fern andezSomoano, A., Ballester, F., Tardon, A., Grazuleviciene, R., Danileviciute, A., Cordier, S., Costet, N., Righi, E., Aggazzotti, G., Stephanou, E.G., Kargaki, S., Nieuwenhuijsen, M.J., 2016. Occurrence of DBPs in drinking water of European regions for epidemiology studies. Am. Water Work. Assoc. 108, E501eE512. Kristiana, I., McDonald, S., Tan, J., Joll, C., Heitz, A., 2015. Analysis of halogen-specific TOX revisited: Method improvement and application. Talanta 139, 104e110. Li, X.F., Mitch, W.A., 2018. Drinking water disinfection byproducts (DBPs) and human health effects: multidisciplinary challenges and opportunities. Environ. Sci. Technol. 52, 1681e1689. Liu, Y., Mao, R., Tong, Y., Lan, H., Zhang, G., Liu, H., Qu, J., 2017. Reductive dechlorination of trichloroacetic acid (TCAA) by electrochemical process over Pd-In/ Al2O3 catalyst. Electrochim. Acta 232, 13e21. Liu, J., Zhang, X., Li, Y., Li, W., Hang, C., Sharma, V.K., 2019. Phototransformation of halophenolic disinfection byproducts in receiving seawater : kinetics , products , and toxicity. Water Res. 150, 68e76.
W.-C. Huang et al. / Water Research 165 (2019) 115024 Luo, C., Gao, J., Wu, D., Jiang, J., Liu, Y., Zhou, W., Ma, J., 2019. Oxidation of 2,4bromophenol by UV/PDS and formation of bromate and brominated products: a comparison to UV/H2O2. Chem. Eng. J. 358, 1342e1350. McKie, M.J., Taylor-Edmonds, L., Andrews, S.A., Andrews, R.C., 2015. Engineered biofiltration for the removal of disinfection by-product precursors and genotoxicity. Water Res. 81, 196e207. Mills, K.D., Ferguson, D.O., Alt, F.W., 2003. The role of DNA breaks in genomic instability and tumorigenesis. Immunol. Rev. 194, 77e95. Muellner, M.G., Wagner, E.D., Mccalla, K., Richardson, S.D., Woo, Y.T., Plewa, M.J., 2007. Haloacetonitriles vs. regulated haloacetic acids: are nitrogen-containing DBFs more toxic? Environ. Sci. Technol. 41, 645e651. Neta, P., Huie, R.E., Ross, A.B., 1988. Rate constants for reactions of inorganic radicals in aqueous solution. J. Phys. Chem. Ref. Data 17, 1027e1284. Pandit, S., Preston, T.J., King, S.J., Vallance, C., Orr-Ewing, A.J., 2016. Evidence for concerted ring opening and C-Br bond breaking in UV-excited bromocyclopropane. J. Chem. Phys. 144. Plewa, M.J., Kargalioglu, Y., Vankerk, D., Minear, R.A., Wagner, E.D., 2002. Mammalian cell cytotoxicity and genotoxicity analysis of drinking water disinfection by-products. Environ. Mol. Mutagen. 40, 134e142. Plewa, M.J., Simmons, J.E., Richardson, S.D., Wagner, E.D., 2010. Mammalian cell cytotoxicity and Genotoxicity of the haloacetic acids, a major class of drinking water disinfection by-products. Environ. Mol. Mutagen. 51, 871e878. Plewa, M.J., Wagner, E.D., Richardson, S.D., 2017. TIC-Tox: a preliminary discussion on identifying the forcing agents of DBP-mediated toxicity of disinfected water. J. Environ. Sci. 58, 208e216. Redon, C.E., Nakamura, A.J., Martin, O.A., Parekh, P.R., Weyemi, U.S., Bonner, W.M., 2011. Recent Developments in the Use of g-H2AX as a Quantitative DNA Double-Strand Break Biomarker, vol. 3, pp. 168e174. Richardson, S.D., Postigo, C., 2011. Drinking Water Disinfection By-Products. Springer-Verlag. Emerging Organic Contaminants and Human Health. Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., DeMarini, D.M., 2007. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection by-products in drinking water: a review and roadmap for research. Mutat. Res. Rev. Mutat. Res. 636, 178e242. Sayess, R., Khalil, A., Shah, M., Reckhow, D.A., Godri Pollitt, K.J., 2017. Comparative cytotoxicity of six iodinated disinfection byproducts on nontransformed epithelial human colon cells. Environ. Sci. Technol. Lett. 4, 143e148. Singer, P.C., 1999. Humic substances as precursors for potentially harmful disinfection by-products. Water Sci. Technol. 40, 25e30. Tang, X., Wu, Q.Y., Du, Y., Yang, Y., Hu, H.Y., 2014. Anti-estrogenic activity formation potential assessment and precursor analysis in reclaimed water during chlorination. Water Res. 48, 490e497. Taylor, R. , 1990. Interpretation of the correlation coefficient: a basic review. J. Diagn.
11
Med. Sonogr. 6 (1), 35e39. Uyguner, C.S., Bekbolet, M., 2005. A comparative study on the photocatalytic degradation of humic substances of various origins. Desalination 176, 167e176. Vione, D., Falletti, G., Maurino, V., Minero, C., Pelizzetti, E., Malandrino, M., Ajassa, R., Olariu, R.I., Arsene, C., 2006. Sources and sinks of hydroxyl radicals upon irradiation of natural water samples. Environ. Sci. Technol. 40, 3775e3781. Wagner, E.D., Plewa, M.J., 2017. CHO cell cytotoxicity and genotoxicity analyses of disinfection by-products: an updated review. J. Environ. Sci. (China) 58, 64e76. Wang, D.S., Zhao, Y.M., Yan, M.Q., Chow, C.W.K., 2013. Removal of DBP precursors in micro-polluted source waters: a comparative study on the enhanced coagulation behavior. Separ. Purif. Technol. 118, 271e278. Wang, L., Niu, R., Chen, B., Wang, Liwei, Zhang, G., 2017a. A comparison of photodegradation kinetics, mechanisms, and products between chlorinated and brominated/iodinated haloacetic acids in water. Chem. Eng. J. 330, 1326e1333. Wang, W.L., Zhang, X., Wu, Q.Y., Du, Y., Hu, H.Y., 2017b. Degradation of natural organic matter by UV/chlorine oxidation: molecular decomposition, formation of oxidation byproducts and cytotoxicity. Water Res. 124, 251e258. Wang, W.L., Wu, Q.Y., Huang, N., Xu, Z. Bin, Lee, M.Y., Hu, H.Y., 2018. Potential risks from UV/H2O2oxidation and UV photocatalysis: a review of toxic, assimilable, and sensory-unpleasant transformation products. Water Res. 141, 109e125. Wang, W.L., Cai, Y.Z., Hu, H.Y., Chen, J., Wang, J., Xue, G., Wu, Q.Y., 2019. Advanced treatment of bio-treated dyeing and finishing wastewater using ozonebiological activated carbon: a study on the synergistic effects. Chem. Eng. J. 359, 168e175. Watters, G.P., Smart, D.J., Harvey, J.S., Austin, C.A., 2009. H2AX phosphorylation as a genotoxicity endpoint. Mutat. Res. Genet. Toxicol. Environ. Mutagen. 679, 50e58. Wu, Q., Li, C., Wang, W., He, T., Hu, H., Du, Y., Wang, T., 2016. Removal of fluorescence and ultraviolet absorbance of dissolved organic matter in reclaimed water by solar light. J. Environ. Sci. (China) 43, 118e127. Wu, Q.Y., Zhou, Y.T., Li, W., Zhang, X., Du, Y., Hu, H.Y., 2019. Underestimated risk from ozonation of wastewater containing bromide: both organic byproducts and bromate contributed to the toxicity increase. Water Res. 162, 43e52. Xiao, Y., Fan, R., Zhang, L., Yue, J., Webster, R.D., Lim, T.T., 2014. Photodegradation of iodinated trihalomethanes in aqueous solution by UV254 irradiation. Water Res. 49, 275e285. Zalazar, C.S., Labas, M.D., Brandi, R.J., Cassano, A.E., 2007. Dichloroacetic acid degradation employing hydrogen peroxide and UV radiation. Chemosphere 66, 808e815. Zheng, D., Andrews, R.C., Andrews, S.A., Taylor-Edmonds, L., 2015. Effects of coagulation on the removal of natural organic matter, genotoxicity, and precursors to halogenated furanones. Water Res. 70, 118e129.
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Influence of UV irradiation on the toxicity of chlorinated water to mammalian cells: Toxicity drivers, toxicity changes and toxicity surrogates Wen-Cheng Huang a, Ye Du b, Min Liu a, Hong-Ying Hu b, c, Qian-Yuan Wu d, *, Ying Chen a, ** a
College of Architecture and Environment, Sichuan University, Chengdu, 610065, PR China Shenzhen Environmental Science and New Energy Technology Engineering Laboratory, Tsinghua-Berkeley Shenzhen Institute, Shenzhen, 518055, PR China Environmental Simulation and Pollution Control State Key Joint Laboratory, State Environmental Protection Key Laboratory of Microorganism Application and Risk Control (SMARC), School of Environment, Tsinghua University, Beijing, 100084, PR China d Key Laboratory of Microorganism Application and Risk Control of Shenzhen, Guangdong Provincial Engineering Research Center for Urban Water Recycling and Environmental Safety, Graduate School at Shenzhen, Tsinghua University, Shenzhen, 518055, PR China b c
a r t i c l e i n f o
a b s t r a c t
Article history: Received 12 August 2019 Accepted 22 August 2019 Available online 23 August 2019
UV irradiation was reported to be able to degrade some kinds of DBPs, yet its influence on the toxicity of chlorinated water to mammalian cells remains unknown. This study systematically investigated the influence of low-pressure UV irradiation on the DBPs and toxicity of chlorinated drinking water (DW) and reclaimed water (RW). The apparent first-order rate constant (kobs) of degradation kinetics of known DBPs increased with the increased Br substitutions. Haloacetonitriles were identified as toxicity drivers among the detected DBPs, which even contributed more to the toxicity after UV irradiation, mainly due to the refractory bromochloroacetonitrile (BCAN) and dichloroacetonitrile (dCAN). Both total organic halogen, cytotoxicity and genotoxicity were significantly removed under UV irradiation, with the removal rate of 22.9%e41.7% for cytotoxicity and a higher rate of 33.1%e55.5% for genotoxicity under 2400 mJ/cm2 irradiation. UV irradiation significantly decreased the UV254, SUVA254 and fluorescence intensity (FLU) of chlorinated water. Results from high performance size exclusion chromatography revealed that chlorinated DW mainly contained high molecular weight (MW) compounds (>1000 Da) while chlorinated RW mainly contained lower MW compounds (100e500 Da). Chromophores and fluorophores in compounds of 100e500 Da increased in chlorinated DW while decreased in chlorinated RW under UV irradiation. Both the removal of UV254, SUVA254, FLU, MW-based UV254 (>1000 Da) and MWbased FLU (each fractions) were significantly correlated (p < 0.05) with the removal of toxicity under UV irradiation. The UV254 of chlorinated water was recommended as the optimal surrogate for toxicity removal. © 2019 Elsevier Ltd. All rights reserved.
Keywords: Chlorination UV irradiation Toxicity driver Cytotoxicity Genotoxicity Surrogate
1. Introduction Chlorination is a major disinfection method used widely both in drinking water and reclaimed water treatment processes. However, disinfection by-products (DBPs) would generate during the
* Corresponding author. Division of Energy and Environment, Graduate School at Shenzhen, Tsinghua University, Room 1810, Shenzhen, 518055, PR China. ** Corresponding author. College of Architecture and Environment, Sichuan University, Chengdu, 610065, PR China. E-mail addresses: [email protected] (Q.-Y. Wu), [email protected] (Y. Chen). https://doi.org/10.1016/j.watres.2019.115024 0043-1354/© 2019 Elsevier Ltd. All rights reserved.
chlorination process (Du et al., 2017a; Hua and Reckhow, 2007). To date, more than 700 kinds of DBP have been reported (Richardson and Postigo, 2011). Despite so many DBPs having been identified, only four kinds of trihalomethanes (THMs), five kinds of haloacetic acids (HAAs), chlorite and bromate are regulated (Guilherme and Rodriguez, 2014; Richardson et al., 2007). In recent years, many unregulated emerging DBPs have been reported and studied, such as haloacetonitriles (HANs), halonitromethanes (HNMs), haloacetamides (HAcAms) and halogenated aromatic DBPs (Richardson et al., 2007; Jiang et al., 2017), while the majority of DBPs still remain unidentified. More than 100 kinds of DBPs have proved to be cytotoxic and genotoxic through in vitro assay
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(Wagner and Plewa, 2017), and around 20 kinds of DBPs was found to be carcinogenic by in vivo assay (Richardson et al., 2007; Li and Mitch, 2018) The risk of DBPs is therefore a major concern during water chlorination. The risk posed by DBPs can be controlled by reducing the concentration. Generally, DBP concentrations can be controlled by precursor removal, disinfection process optimization, and elimination after formation (Doederer et al., 2014; Hua and Reckhow, 2007; Wang et al., 2013). In the supply system for drinking water or reclaimed water, secondary water supply tanks are usually needed to store and supply water to users. Recently, some studies showed that UV irradiation could degrade DBPs such as HAAs and THMs (Bu et al., 2018; Xiao et al., 2014; Zalazar et al., 2007). There is a possibility to eliminate the DBP level in secondary water supply tanks by UV irradiation right before use. However, previous studies mainly focused on several typical DBPs or DBPs in synthetic water, the degradation of overall DBPs generated in real water by UV photolysis needs further investigation. Among many kinds of DBPs, the regulated DBPs like THMs are most widely studied owing to their widespread presence and high concentrations in chlorinated water samples. However, only focusing on the high-concentration DBPs might be insufficient to guarantee the water safety. Because many emerging DBPs, especially nitrogenous DBPs (N-DBPs) (Muellner et al., 2007) are high toxic contaminants. For example, the emerging DBP bromochloroacetonitrile (BCAN) is 1000 times more cytotoxic than tCM (Wagner and Plewa, 2017). Identifying the toxicity drivers in known DBPs by both considering the concentration and toxic potency would be more scientifically significant in risk control. Moreover, due to the different degradation efficiency of DBPs under UV irradiation, it is therefore also necessary to understand toxicity drivers in chlorinated water after UV irradiation. Although more than 700 DBPs have been identified, known DBPs comprise only a small fraction of all DBPs, generally accounting for 30% of the total organic halogen (TOX) (Kristiana et al., 2015). Therefore, changes of TOX under UV irradiation would give a full description of DBP degradation. Furthermore, the concern lies behind the DBPs is their toxic effects. The incomplete dehalogenation of DBPs does not necessarily decrease the toxicity. For example, during the dehalogenation of trichloroacetic acid, monochloroacetic acid might form (Liu et al., 2017), which is even more toxic than trichloroacetic acid (Wagner and Plewa, 2017). Similarly, for haloacetonitriles and haloacetamides, less halogen substitution generally leads to higher toxicity (Wagner and Plewa, 2017). Toxicity elimination is the ultimate goal of DBP control. Cytotoxicity tests are mostly used to analyze the lethal effects of contaminants (Du et al., 2017b; Sayess et al., 2017). Some DBPs might not cause direct cell death, but impair the genetic materials at low concentration levels (McKie et al., 2015; Zheng et al., 2015). Results of cytotoxicity and genotoxicity would give a more comprehensive understanding about the effects of DBP control by UV irradiation, yet few studies clarified the toxicity changes of chlorinated water under UV irradiation. Cytotoxicity and genotoxicity are ideal indexes to characterize the water safety. However, the drawback is the cost of time and money during toxicity assay. When degrading DBPs, UV irradiation alters the water quality indexes at the same time. Finding surrogates for the toxicity changes would be more convenient for the toxicity regulation. In previous studies, some water parameters could be used as surrogates for toxicity formation during water chlorination. Such as UV254 absorbance can use as a surrogate for the increase of cytotoxicity (Du et al., 2017b); The formation of antiestrogenic activity can be predicted by fluorescence (FLU) intensity (Tang et al., 2014). However, little research clarified surrogates for the toxicity removal of chlorinated water, especially for cytotoxicity
removal and genotoxicity removal under UV irradiation. The main objective of this study is thusly to understand the effects of UV irradiation on toxicity removal of DBPs in chlorinated drinking water and reclaimed water. Degradation kinetics of DBPs, toxicity drivers in chlorinated water before and after UV irradiation were investigated. Changes in TOX, cytotoxicity, and genotoxicity of chlorinated water under UV irradiation were investigated. Furthermore, water quality parameters and molecular weight (MW) distribution were measured to analyze their correlation with toxicity and identify suitable surrogates for indicating toxicity removal. 2. Materials and methods 2.1. Chemical and reagents Organic solvents, including methanol (MeOH), acetone, dichloromethane, and methyl tert-butyl ether (MTBE), were all of HPLC grade and purchased from J.T. Baker (U.S.A.). NaOH (99%), NaClO (99%), Na2SO4 (99%), and NaS2O3 (99%) were obtained from Macklin (China). Eleven DBPs were detected in chlorinated water samples in this study, including trichloromethane (tCM), bromodichloromethane (BdCM), bromoform (tBM), dibromochloromethane (dBCM)), trichloronitromethane (tCNM), dibromoacetonitrile (dBAN), bromochloroacetonitrile (BCAN), dichloroacetonitrle (dCAN), trichloroacetamide (tCAcAm), chloral hydrate (CH), and 1,1,1-trichloropropanone (1,1,1-tCP). Suppliers and purities of all DBP standards are shown in Table S2. For the toxicity assay, dimethyl sulfoxide (DMSO) (99.7%) and fetal bovine serum (FBS) were purchased from Fisher Scientific (Pittsburgh, PA). Dulbecco's modified Eagle's medium/nutrient mixture F-12 (DMEM/F-12, 1:1) and penicillinestreptomycin were purchased from Gibco (USA). Cell Counting Kit-8 (CCK-8) for the cytotoxicity test was obtained from Dojindo Molecular Technologies (Japan). Paraformaldehyde was purchased from Sigma-Aldrich (USA). Triton X-100 and Albumin Bovine V were purchased from Solarbio (China). P-Histone H2A.X (S139) rabbit antibody and antirabbit IgG Fab2 Alexa Fluor 647 molecular probes were obtained from Cell Signaling Technology (USA). Hoechst 33342 was purchased from Biotium (USA). 2.2. Water sampling Simulated drinking water (DW) was prepared by dissolving Suwannee River natural organic matter (SRNOM 2R101; International Humic Substances Society) in ultrapure water. Reclaimed water (RW) samples A and B were collected from two different reclaimed water treatment plants. All water samples were filtered through 0.45-mm microfilters and stored at 4 C. Water quality parameters are shown in Table S1. Details of DOC and UV254 absorbance analyses are provided in Text S1. Threeedimensional excitation emission matrix of fluorescence spectra (FEEM) was measured using a fluorophotometer (F-7000, Shimadzu). The determination and data processing method for FEEM is described in detail elsewhere (Chen et al., 2003). 2.3. Chlorination The water samples were first adjusted to pH 7 ± 0.05 using H2SO4 or NaOH (2 M), stabilized with 1 mM phosphate buffer (pH 7), and then chlorinated with 10 mg/L NaClO (expressed as Cl2 concentration). After chlorination in the dark for 24 h, residual free chlorine was quenched with 105% of the stoichiometric amount of NaS2O3.
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2.4. UV irradiation The chlorinated samples were exposed to a quasi-collimated beam with a wavelength of 254 nm, produced by two 41 W lowpressure Hg lamps (LightSources, Orange, CT, USA). The 250 mL cylindrical reactor (diameter, 10 cm) was placed vertically below the lamps. The UV dose applied to the test solution during photodegradation experiments was determined to be 1.20 ± 0.05 mW/ cm2 using iodideeiodate chemical actinometry, as described by Bolton and Linden (2003). The reactor containing the test solution was stirred at 350 rpm throughout the experiment. The irradiation time under the highest UV dose (2400 mJ/cm2) condition was around 20 min. The temperature of water samples under UV irradiation was maintained at 27.5 ± 1.0 C. After various UV irradiation doses, water samples were collected for analysis. 2.5. Solid phase extraction (SPE) SPE was used as the pretreatment for TOX and toxicity assays. The water samples were first adjusted to pH 2 ± 0.05 using H2SO4 (2 M) and then concentrated using 6 mL Oasis hydrophilicelipophilic balance resin cartridges (Waters, USA) that had been activated with methanol (10 mL) and ultrapure water (10 mL). The water samples were passed through the cartridges at 3 mL/min. After adsorption was finished, the resin was eluted with methanol (5 mL), acetone (2 mL), and dichloromethane (2 mL). Finally, the extracts were fully dried under nitrogen flow. 2.6. Determination of DBPs and TOX The DBPs were enriched (ten-fold) by liquideliquid extraction (LLE) of the treated water samples, which had already been adjusted to pH 2 ± 0.05. MTBE was used as an extraction agent, containing 1,2-dibromopropane (100 mg/L) as the internal standard. The samples were then shaken for 5 min and allowed to stand for 10 min. GC-ECD (Agilent, USA) equipped with a DB-5MS capillary column (30 m 0.25 mm 0.25 mm, Agilent, USA) was used to measure the DBPs, using nitrogen as the carrier gas. The injector and detector temperatures were 190 and 290 C, respectively. The initial column temperature was 35 C (maintained for 9 min), which was increased to 40 C at a rate of 2 C/min (maintained for 1 min), then 160 C at a rate of 10 C/min (maintained for 2 min), and finally 220 C at a rate of 40 C/min (maintained for 2 min). The TOX content, including total organic chlorine (TOCl) and total organic bromine (TOBr) contents, was analyzed as the index of the overall halogenated DBP content. After SPE, the extracts were dissolved in ultrapure water and irradiated by vacuum UV (VUV) lamps (185 nm, 12 W, UV Tec, Co., China) according to the method of Bu et al. (2018) to convert organic halogens into inorganic halides. The Cl and Br concentrations were determined by ion chromatography (ICS-900, ThermoFisher Scientific, USA) after irradiation. The TOX content was calculated from the difference between inorganic halide contents before and after VUV irradiation. 2.7. Toxicity index calculation Cytotoxicity index and genotoxicity index were calculated using the toxic potency data reported by Wagner and Plewa (2017). In the literature, the cytotoxicity index (CTI) of the detected DBPs was calculated based on the concentration at which the relative cell viability was 50% (LC50) using a crystal violet cytotoxicity assay. The genotoxicity index (GTI) of the detected DBPs was calculated using the SCGE 50% Tail DNA (TDNA) value obtained using a single cell gel electrophoresis (SCGE) genotoxicity assay (Wagner and Plewa, 2017). Both cytotoxicity and genotoxicity assays used Chinese
3
Hamster Ovary (CHO) cell line as the subject. The LC50 and 50% TDNA values of individual DBPs are shown in Table S3, as reported by Wagner and Plewa (2017). Assuming that the toxicity of individual DBPs was additive, the CTI and GTI were weighted for toxic potency by dividing the DBP concentrations by the LC50 or 50% TDNA values (Chuang et al., 2019). 2.8. Molecular weight (MW) distribution transformation Transformations in the molecular weight distributions of chlorinated water samples after UV irradiation were characterized by high-performance size exclusive chromatography (HPSEC) (LC20, Shimadzu, Japan) using two tandem columns (Tsk-Gel G3000PWXL and Tsk-Gel G2500PWXL, Japan) with an UV array detector, fluorescence (FLU) detectors, and a modified total organic carbon detector (Sievers900, GE, USA). Details of this procedure can be found in previous studies (Wang et al., 2019; Wu et al., 2016). 2.9. Cell culture The Chinese Hamster Ovary (CHO-k1) cell line, which is widely used as a mammalian cell subject in toxicity assays (Du et al., 2018a; Plewa et al., 2010), was obtained from the American Type Culture Collection to perform the toxicity assay. Cells were cultured with DMEM/F-12 (1:1) supplemented with 10% fetal bovine serum, 100 U/mL penicillin, and 100 mg/L streptomycin. The cell dishes were placed in an incubator with 5% CO2 gas at a high humidity and 37 C. Cells passaging was performed every two days. Cells with 2e5 passages were used for the toxicity assay. 2.10. Cytotoxicity assay CCK-8 was used to detect cytotoxicity in this experiment, using highly water-soluble tetrazolium salt 2-(2-methoxy-4nitrophenyl)-3-(4-nitrophenyl)-5-(2,4-disulfophenyl)-2H-tetrazolium monosodium salt (WST-8) to produce a water-soluble formazan dye after reduction with intracellular dehydrogenase. The 450 nm absorbance of the dye showed a good linear relationship with the number of living cells (Fig. S1), as measured using a SpectraMax i3 system (Molecular Devices, USA). After SPE, the extracts of water samples were dissolved in DMEM/ F-12 mixed with 0.5% DMSO as test samples. The cytotoxicity of the phenol solution was measured simultaneously and used as a reference compound to quantify the cytotoxicity equivalents. Cells were preincubated for 12 h for adherent growth, then exposed to contaminants for 48 h. Each test was performed with 3e4 replicates. The detailed cytotoxicity assay procedure is provided in Text. S2. 2.11. Genotoxicity test The phosphorylated H2AX (pH2AX) was used as a genotoxicity endpoint, representing the DNA double-strand breaks (DSBs) (Bonner et al., 2008), which is widely used in genotoxicity assays (Audebert et al., 2010; Redon et al., 2011; Watters et al., 2009). For the genotoxicity assay, cells were seeded at 5 103 per well. After preincubation for 12 h and exposure for 24 h, the cells were fixed, permeated, and blocked. The cells were then treated with primary antibody P-Histone H2A.X (S139) rabbit antibody and stained with secondary antibody anti-rabbit IgG Fab2 Alexa Fluor 647 molecular probe and Hoechst 33258 together. Finally, cell images were obtained using an HCA system (ImageXpress Micro; Molecular Devices, USA). Each test was performed with 4e6 replicates. The pH2AX values per cell for which the cell viability was higher than 70% were used to obtain concentration-response curves. Details can be found in Wu et al. (2019).
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2.12. Statistical analysis DBP degradation under UV irradiation was fitted by first-order reaction kinetics. For cytotoxicity assay, the relative cell viability (against the negative control) v.s. different concentration factors were used to obtain concentration-response curves by regression analysis. LC50 values were used to calculate the cytotoxicity equivalents (phenol). For genotoxicity, the pH2AX induction rate (against the negative control) v.s. different concentration factors were used to obtain concentration-response curves. The concentration of the 1.5-fold induction rate (IR1.5) of pH2AX per well was used to calculate the genotoxicity equivalents (4-NQO). One-way analysis of variance (ANOVA) was used to determine whether there was a significant decrease in sample toxicity after UV irradiation (p < 0.05). To find the surrogates for toxicity removal under UV irradiation, linear fitting was applied to the removal rate of toxicity and water quality parameter and the significant correlation was considered when p < 0.05.
3. Results and discussion 3.1. Degradation of detected DBPs under UV irradiation Concentration of the detected DBPs after chlorination and their changes during UV irradiation are shown in Figs. S2ae2d. The formation of THMs during chlorination were highest in both RW and DW compared with other DBPs, especially in DW, which was similar to the previous result (Krasner et al., 2016). TCP, CH, dCAN, and tCAcAm were also found in chlorinated DW, at concentrations
of 10.8, 8.0, 5.2, and 1.7 mg/L, respectively. Concentrations of nitrogenous DBPs such as haloacetonitriles (HANs) and haloacetamide in chlorinated RW were higher than chlorinated DW, which might be because RW contain more organic nitrogen (Hu et al., 2016). Owing to the presence of Br, brominated DBPs (BrDBPs) were found in RW, such as BdCM and BCAN. Except tCP and CH, all DBPs showed significant degradation under UV irradiation. In DW, almost all of tCAcAm was removed, tCM and dCAN reduced 85.8% and 30% in 2000 s (2400 mJ/cm2). The removal rates of tCM in RW were around 58%, which was a little lower compared with in DW. The tCNM, which only found in RW, was also significantly removed by 91.1% and 80.1% in RW A and RW B. The removal rate of Br-DBPs were higher than there chlorinated counterparts. For example, BCAN reduced by 55% in RW while DCAN only reduced by 20%e30%. Fig. 1d shows the apparent first-order rate constants of degradation (kobs) of the detected DBPs in DW and RW (kinetics curves of DBPs in each sample are shown in Fig. 1aec. Br-DBPs degraded faster than their chlorinated counterparts under UV irradiation. The kobs values increased with the number of Br substitutions increasing. For example, the degradation order was tBM > CdBM > tCM and dBAN > BCAN > dCAN. Other DBPs, such as haloacetic acids, THMs and halophenolic also followed this degradation rule, which was reported previously (Wang et al., 2017a,b; Xiao et al., 2014; Liu et al., 2019). The carbonehalogen bonds in DBPs, such as CeBr and CeCl, can be broken by sufficient internal energy through absorption of the photon of UV light to an excited state (Pandit et al., 2016). The photon absorption capacity of DBPs depends on their molar
Fig. 1. Degradation kinetics of DBPs in different water samples by UV irradiation. Degradation kinetics of different DBPs in (a) DW, (b) RW A and (c) RW B. (d) Comparison of apparent first order rate constant (kobs) of degradation of DBPs in different water samples. (#: Data not available).
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extinction coefficient (ε). Higher ε values lead to a faster degradation in DBPs through direct UV photodegradation (Fang et al., 2013). Each substitution of a Br can increase the ε value of trihalogenated DBPs by 240 ± 74 M1cm1 (Chuang et al., 2016), indicating that more Br substitution leads to an increase in the photodegradation rate of DBPs, as confirmed by our results. In a previous study, dCAN was found to hardly degrade in ultrapure water under UV direct photolysis with a UV intensity of 6.23 mW/cm2 (Hou et al., 2017), which was much higher than that in this study (1.2 mW/cm2). However, the kobs values of dCAN in DW and RW in this study were 0.132e0.170 103 s1, indicating that direct UV photolysis was not the only degradation pathway for DBPs in DW and RW. Hydroxyl radical (OH$) and other radicals have been found to play an important role in the degradation of some DBPs, such as THMs and 2,4-bromophenol (Luo et al., 2019; Xiao et al., 2014). In RW and DW, triplet dissolved organic matter (3DOM*) and reactive oxygen species (ROS) can be generated through UV irradiation of dissolved organic matter (DOM) (Jacobs et al., 2011; Vione et al., 2006). These radicals can react with eBr, eCN, CeH, and other bonds, eventually leading to bond cleavage and reconstruction (Wang et al., 2018), which contributes to DBP
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degradation. Compared with Cl substitution, Br substitution improved the reactivity of DBPs with OH$ (Chuang et al., 2016). This also explained why the kobs values of DBPs were higher with the increased Br substitution.
3.2. Toxicity drivers in detected DBPs before and after UV irradiation To determine which DBPs posed the greatest risk among the detected DBPs before and after UV irradiation, by using the toxic potency data of individual DBPs reported by the same group (Wagner and Plewa, 2017), cytotoxicity index (CTI) and genotoxicity index (GTI) of the detected DBPs to CHO cells were calculated. The CTI of detected DBPs in chlorinated RW and chlorinated DW were significantly decreased by UV irradiation, with 32.1%, 42.2% and 51.0% removal from DW, RW A and RW B, respectively, after UV irradiation at 2400 mJ/cm2. Clearly, before UV irradiation, HANs contributed most of the CTI in all three water samples (Fig. 2a and b). In DW, dCAN was the main cytotoxicity driver, with its contribution ratio to the CTI increasing from 86.0% to 89.4% after UV irradiation. The remaining CTI of less than 15% was contributed by
Fig. 2. Changes in the toxicity index under UV irradiation: (a) the decrease of CTI, (b) changes in the contribution of DBPs to CTI, (c) the decrease of GTI and (d) changes in the contribution of DBPs to GTI.
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tCM, CH, and tCAcAm. The regulated tCM only contributed 8.6% of the CTI initially, decreasing to 3.1% after UV irradiation (2400 mJ/ cm2). Before UV irradiation, HANs had a greater effect on the CTI in RW (more than 97%), which contained dCAN, BCAN, and dBAN, contributed 38.8%, 52.6% and 5.8% of CTI in chlorinated RW A, 25.4% 55.1% and 17.6% of CTI in chlorinated RW B, respectively. When the UV dose reached 2400 mJ/cm2, 97% of the CTI was contributed by dCAN and BCAN. Furthermore, the contributions of tCM to the CTI were less than 1% in RW. The results showed a decrease in the GTI of detected DBPs similar to that of the CTI (Fig. 2c and d). UV irradiation significantly decreased the GTIs of DW, RW A, and RW B, by 35.5%, 61.2%, and 65.4%, respectively. The initial contribution ratio of the GTI of HANs to DW, RW A, and RW B was 91.4%, 69.7%, and 83.6%, respectively, all of which increased to more than 90% after irradiation with a UV dose of 2400 mJ/cm2. BCAN and dCAN remained the major toxicity drivers during UV photolysis. TCNM made a different contribution to the GTI of untreated RW A and RW B, at 27.9% and 15.2%, respectively. However, owing to the rapid degradation of TCNM, its contribution to the GTI was less than 10% after UV irradiation. HANs were the major toxicity drivers among the detected DBPs, especially dCAN and BCAN. Although dBCN was more toxic than the other two (Plewa et al., 2002), the amount of dBCN generated during chlorination was small and it degraded much faster under UV irradiation. BCAN and dCAN were not only highly toxic, but also had slow degradation rates under UV irradiation. Therefore, the contribution ratios of BCAN and dCAN to the toxicity index increased during UV irradiation. THMs, the regulated DBPs, had the highest concentrations among the detected DBPs after chlorination, and did make some contribution to the CTI in DW before irradiation (no more than 10%). However, owing to the much lower toxicity and higher degradation rates of THMs compared with those of BCAN and dCAN, their contribution ratio to the toxicity index decreased during UV irradiation. In RW, the concentration ratio of HANs among the detected DBPs was much higher than that in DW, such that the toxicity index contribution of THMs was negligible. Previous study also implied that HANs were the toxicity drivers of known DBPs in drinking water (Plewa et al., 2017). However, after UV irradiation, this study for the first time revealed that HANs even contributed more to the toxicity index, mainly resulted from the recalcitrant BCAN and dCAN under UV irradiation. These results suggested that except for UV irradiation, other treatments might be needed to further eliminate the HANs to guarantee the water safety.
suggested that brominated DBPs were more easily degraded than chlorinated DBPs under UV irradiation. 3.4. Removal of cytotoxicity and genotoxicity As shown above, although it was found that HANs were toxicity drivers among the detected DBPs, the toxicity index from all the detected DBPs remained very low (only at 103 level, the level of 1 means LC50). Only focusing on the toxicity of a few known DBPs is not enough. Evaluation of changes in cytotoxicity and genotoxicity of all the DBPs under UV irradiation are therefore necessary. The concentration-response curves of cytotoxicity for each sample are shown in Fig. 4aec. The curves shifted to the right as the UV dose increased, leading to an increased LC50, which indicated decreased cytotoxicity. The cytotoxicity equivalent was significantly reduced by UV irradiation (Fig. 4d), by 22.9%, 29.8%, and 41.7% (p < 0.05) for chlorinated DW, RW A, and RW B, respectively. The rate of cytotoxicity decrease was fast in the first 600 mJ/cm2, but gradually slowed thereafter. The cytotoxicity assay reflected the macroscopic hazards of DBPs on CHO cells, while the genotoxicity assay evaluated the effects of DBPs on cellular genes. In this study, pH2AX was measured to characterize the DNA double-strand breaks (DSBs), which is the most severe form of DNA damage (Khanna and Jackson, 2001; Mills
3.3. Removal of TOX by UV irradiation The known DBPs generally only accounted for 30% of the total halogenated DBPs (Krasner et al., 2006). The TOX content is widely used to indicate the total halogenated DBP content (Du et al., 2018b; Jiang et al., 2017; Wu et al., 2019). Therefore, to comprehensively understand the removal of DBPs under UV irradiation, changes in TOX contents were investigated. Fig. 3 shows the effect of UV irradiation on the TOX contents in each water sample. The initial TOCl concentration in DW, RW A, and RW B was 315.8, 212.3, and 116.7 mg/L respectively. After UV irradiation, the TOCl contents of all chlorinated water samples were significantly decreased, by 25.9%, 35.1% and 53.6% for RW A, RW B and DW, respectively. TOBr was only found in RW, at about 10 mg/L, with more than 60% degradation under UV irradiation, the removal rate of which was higher than that of TOCl. As the TOX comprises a mixture of different DBPs, it might be improper to fit the degradation kinetics. However, by plotting the C/C0 vs. UV dose, it was found that TOBr clearly degraded faster than the TOCl, which was similar to the degradation of detected DBPs. All these results
Fig. 3. Effect of UV irradiation on the total organic halogen (TOX). (a) TOX concentration, (b) TOX degradation ratio.
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et al., 2003). Fig. 5a and Fig. S4 shows the images of pH2AX foci formed in cell nucleus when cells were exposed to chlorinated water (before and after UV irradiation, UV dose ¼ 2400 mJ/cm2). There was a significant decrease of pH2AX foci numbers after the UV irradiation of chlorinated water. The IR1.5 value increased with increasing UV dose, representing the decrease in genotoxicity under UV irradiation (Fig. 5b and Fig. S5). Compared with the reduction in cytotoxicity, UV irradiation had a greater influence on the removal of genotoxicity (Fig. 5c). The initial genotoxicity of chlorinated RW was higher than that of chlorinated DW. UV irradiation significantly reduced the genotoxicity by 33.1%, 46.5%, and 55.5% (p < 0.05) for chlorinated DW, RW A and RW B, respectively. The trend in the reduction rate of genotoxicity was similar to that of cytotoxicity, showing gradual slowing as the UV dose increased. In conclusion, UV irradiation could significantly reduce the cytotoxicity and genotoxicity of chlorinated water. 3.5. Effect of UV irradiation on the quality of chlorinated water Changes in DOC content, UV254 absorbance and FLU values of chlorinated water were measured. The DOC contents of the three chlorinated water samples did not change much during UV irradiation (Fig. S6), showing that UV irradiation was difficult to mineralize the organic matters. UV irradiation significantly reduced the UV254 and specific UV absorbance at 254 nm (SUVA254) of chlorinated DW and chlorinated RW (Fig. S7 and Fig. S8), which represented the reduction in aromaticity, double bonds and other unsaturated bonds (Singer, 1999; Uyguner and Bekbolet, 2005). FEEM was divided into five regions according to a previous study (Chen et al., 2003). DW contained more fulvic
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humic substances (Region III) while RW contained more tryptophan-like proteins (Region II) (Figs. S9eS11). After UV irradiation, the FLU values of all chlorinated water samples first increased slightly and then decreased (Fig. S12). The organic matters in water have been substantially oxidized during chlorination, which led to the reduction in FLU values. When water samples were irradiated with UV, the generated hydrated electron (e(aq)) are generated, which can partially reduce the oxidized organic matters, causing the FLU value to increase (Neta et al., 1988). Then, with increasing UV dose, the FLU value decreased due to further degradation of organic matters. 3.6. Changes of MW distribution under UV irradiation Figs. 6, S13, and S14 show the spectra of MW distribution of chlorinated DW and chlorinated RW under UV irradiation, as determined by HPSEC coupled with three tandem FLU, UV and TOC detectors. To quantitatively determine the changes in the different MW fractions of the organic matter, the chromatograms were deconvoluted into Gaussian peaks according to a previous study (L. Wang et al., 2017). Chromatograms were divided into four fractions based on the apparent MW, namely F1 (<100 Da), F2 (100e500 Da), F3 (500e1000 Da) and F4 (>1000 Da). Signals from all the three detectors suggested that DW mainly contained high MW F4 (>1000 Da) while RW mainly contained lower MW F2 (100e500 Da). This was in accordance with previous study that organic matters in RW consists of more low MW compounds (Hu et al., 2016). The HPSEC-FLU showed that, changes in the FLU (Ex/Em ¼ 240/ 360 nm) of chlorinated DW first increased and then decreased, with
Fig. 4. (a)e(c): Concentration-response curves for cytotoxicity assay of sample under UV irradiation (a) chlorinated DW, (b) chlorinated RW A and (c) chlorinated RW B; (d) Changes in cytotoxicity equivalents of chlorinated water under UV irradiation.
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this change mainly attributed to 100e500 Da compounds (Fig. S13b). HPSEC-FLU tended to decrease with increasing UV dose in chlorinated RW, especially for F2 (100e500 Da) (Fig. 6b). Therefore, it seems that fluorophores in F2 (100e500 Da) were most susceptible to UV irradiation. The HPSEC-TOC showed that, TOC content in all three samples did not change much during UV irradiation, but showed significant transformations between different MW fractions. High MW F4 (>1000 Da) decreased while F3 (500e1000 Da) increased, which mainly occurred under the first 300 mJ/cm2 UV irradiation (Fig .6c). The HPSEC-UV showed that, UV254 absorbance of all three water samples showed a downward trend under UV irradiation. In chlorinated DW, the decrease in UV254 absorbance was mainly due to high MW F4 (>1000 Da) (Fig. S13d). Simultaneously, the UV254 absorbance of the low MW F2 (100e500 Da) increased, which might be attributed to the transformation of high MW compounds into low MW compounds. For chlorinated RW, the decreased UV254 mainly resulted from the fractions with MW > 100 Da (Fig. 6d and Fig. S14d), suggesting that chromophores in F2eF4 were more susceptible to UV irradiation (Uyguner and Bekbolet, 2005). 3.7. Correlation between removal of toxicity and water quality parameters To identify surrogates of toxicity removal during UV irradiation, when there was a significant linear correlation (p < 0.05) between toxicity removal and changes in water quality parameters, the
coefficient of determination (R2) was used to evaluate the degree of relevance (Taylor, 1990). The data are summarized in Table 1. No significant correlation was observed between removal of DOC content and the toxicity, because UV irradiation cause little mineralization of the organic matters. Removal of UV254 and SUVA254 showed good correlation with removal of cytotoxicity (p ¼ 0.003, R2 ¼ 0.602 and p ¼ 0.002, R2 ¼ 0.619, respectively) and genotoxicity (p < 0.001, R2 ¼ 0.757 and p < 0.001, R2 ¼ 0.744, respectively) (Figs. S15aeS15d). The FLU values from five regions were significantly correlated with the removal of cytotoxicity and genotoxicity, especially the total FLU value (p < 0.001, R2 ¼ 0.816 and p < 0.001, R2 ¼ 0.793, respectively) (Figs. S15e and S15f). The strong correlation between toxicity removal and the changes in FLU value, UV254 and SUVA254 indicated that toxicity removal from chlorinated water during UV irradiation occurred simultaneously with the oxidation of unsaturated bonds. By evaluating the relationship between the MW distribution and the toxicity removal rate, the UV254 absorbance of F4 (>1000 Da) was found to be strongly correlated (p < 0.001) with the toxicity removal rate, with R2 values of 0.761 and 0.844 for cytotoxicity and genotoxicity, respectively. No strong correlation between the UV254 absorbance of fractions with MW < 1000 Da and toxicity removal was found. For MW-based FLU, except F1 (<100 Da), all fractions were significantly correlated with toxicity removal. The MW-based TOC cannot serve as surrogates for toxicity removal as no significant correlation was found in all the fractions. Considering the total FLU value would first increase, it may be
Fig. 5. (a) Images of pH2AX foci in cell nucleus when cells were exposed to chlorinated RW B (before and after UV irradiation, UV dose ¼ 2400 mJ/cm2, 30 folds concentration). (b) Concentration-response curves for genotoxicity assay of chlorinated RW B under UV irradiation. (c) Changes in genotoxicity equivalents of chlorinated water under UV irradiation.
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Fig. 6. (a) Effects of UV irradiation on MW distribution spectra of chlorinated RW B. (b)e(c): Transformation of each MW fractions of chlorinated RW B: (b) FLU, (c) TOC and (d) UV254.
Table 1 Correlation between removal of water quality parameters and toxicity (n ¼ 12). Cytotoxicity Slope
Genotoxicity Intercept
R2
P
Slope
Intercept
R2
P
TOC
N/A
>0.05
N/A
UV254
0.712
0.047
0.602
0.003
1.028
0.069
0.757
< 0.001
SUVA
0.609
0.089
0.619
0.002
0.861
0.116
0.744
< 0.001
I II III IV V
0.658 0.564 1.117 0.401 1.056 0.861
0.204 0.173 0.266 0.247 0.369 0.228
0.641 0.427 0.615 0.545 0.476 0.816
0.002 0.021 0.003 0.006 0.011 < 0.001
0.89 0.682 1.51 0.545 1.372 1.094
0.298 0.265 0.381 0.358 0.515 0.332
0.708 0.376 0.654 0.608 0.484 0.793
< 0.001 0.034 0.001 0.003 0.012 < 0.001
> 1000 500e1000 100e500 <100 Total
0.637 N/A N/A N/A 1.068
0.051
0.761
< 0.001 >0.05 >0.05 >0.05 < 0.001
0.865 N/A 0.480 N/A 1.354
0.092
0.844
0.289
0.376
< 0.001 >0.05 0.036 >0.05 < 0.001
FLU
> 1000 500e1000 100e500 <100 Total
0.306 0.317 1.117 N/A 0.289
0.01 0.034 0.003 >0.05 0.039
0.433 0.409 0.312 N/A 0.404
TOC
>1000 500e1000 100e500 <100 Total
N/A N/A N/A N/A N/A
>0.05 >0.05 >0.05 >0.05 >0.05
N/A N/A N/A N/A N/A
FLU volume
MW
Region Region Region Region Region Total UV254
*N/A: p > 0.05, not applicable.
0.073
0.797
0.142 0.195 0.266
0.498 0.377 0.615
0.201
0.359
>0.05
0.115
0.772
0.209 0.288 0.334
0.6 0.379 0.377
0.294
0.422
0.003 0.033 0.034 >0.05 0.022 >0.05 >0.05 >0.05 >0.05 >0.05
10
W.-C. Huang et al. / Water Research 165 (2019) 115024
improper to indicate the toxicity removal when UV dose was less than 300 mJ/cm2. Besides, the availability of SUVA254 also requires the measurement of UV254. Therefore, changes in UV254 was the best surrogate for toxicity removal during UV irradiation of chlorinated water. UV254 of F4 (>1000 Da) and MW-based FLU were not recommended to be surrogates due to the inconvenient availability and relatively low R2, although they were significantly correlated with toxicity removal. 4. Conclusions Influences of UV irradiation on chlorinated water were systematically investigated in this study. Degradation kinetics of known DBPs and their toxicity indexes were evaluated. Changes in TOX contents, cytotoxicity and genotoxicity of chlorinated water were studied. The correlation between removal of toxicity and water quality parameters were also evaluated to identify surrogates for toxicity removal. The main conclusions were as follows: (1) For the same class of DBPs, increase in the number of Br substitutions increased the degradation kobs under UV irradiation. HANs were the toxicity drivers in the detected DBPs in chlorinated water, which even contributed more to the toxicity index after UV irradiation, with the refractory BCAN and dCAN being the main contributor. (2) UV irradiation significantly decreased the TOX in chlorinated water. Under the UV dose of 2400 mJ/cm2, TOCl was eliminated by 25.9%e53.6% while TOBr was eliminated with a higher ratio of around 60%. Both cytotoxicity and genotoxicity were significantly removed under UV irradiation, with the removal rate of 22.9%e41.7% for cytotoxicity equivalent and a higher rate of 33.1%e55.5% for genotoxicity equivalents. (3) UV irradiation showed little impact on DOC of chlorinated water, but significantly reduced the UV254 and SUVA254. FLU values first increased then followed by a decrease under UV irradiation. Results of HPSEC revealed that chlorinated DW mainly contained high MW compound (>1000 Da) while chlorinated RW mainly contained lower MW compound (100e500 Da). Chromophores and fluorophores in F2 (100e500 Da) increased in chlorinated DW while decreased in chlorinated RW under UV irradiation. (4) Both the removal of UV254, SUVA254, FLU (total and five regions), MW-based UV254 (>1000 Da) and MW-based FLU were significantly correlated with the removal of toxicity in chlorinated water under UV irradiation. The UV254 was recommended as the optimal surrogate because of the highest R2 value. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgments This study was supported by National Natural Science Foundation of China (No. 51738005/51678332), the Shenzhen Science, Technology and Innovation Commission (No. JCYJ20170818091859147), special support program for high-level personnel recruitment in Guangdong Province (2016TQ03Z384), and the Development and Reform Commission of Shenzhen Municipality.
Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.watres.2019.115024. References Audebert, M., Riu, A., Jacques, C., Hillenweck, A., Jamin, E.L., Zalko, D., Cravedi, J.P., 2010. Use of the gH2AX assay for assessing the genotoxicity of polycyclic aromatic hydrocarbons in human cell lines. Toxicol. Lett. 199, 182e192. Bolton, J.R., Linden, K.G., 2003. Standardization of methods for fluence (UV Dose) determination in bench-scale UV experiments. J. Environ. Eng. 129, 209e215. Bonner, W.M., Redon, C.E., Dickey, J.S., Nakamura, A.J., Sedelnikova, O.A., Solier, S., Pommier, Y., 2008. gH2AX and cancer. Nat. Rev. Cancer 8, 957e967. Bu, Y., Wang, L., Chen, B., Niu, R., Chen, Y., 2018. Effects of typical water components on the UV254 photodegradation kinetics of haloacetic acids in water. Separ. Purif. Technol. 200, 255e265. Chen, W., Westerhoff, P., Leenheer, J.A., Booksh, K., 2003. Fluorescence excitationemission matrix regional iIntegration to quantify spectra for dissolved organic matter. Environ. Sci. Technol. 37, 5701e5710. Chuang, Y.-H., Parker, K.M., Mitch, W.A., 2016. Development of predictive models for the degradation of halogenated disinfection byproducts during the UV/H2O2 advanced oxidation process. Environ. Sci. Technol. 50, 11209e11217. Chuang, Y.H., Szczuka, A., Mitch, W.A., 2019. Comparison of toxicity-weighted disinfection byproduct concentrations in potable reuse waters and conventional drinking waters as a new approach to assessing the quality of advanced treatment train waters. Environ. Sci. Technol. 53, 3729e3738. , M.J., Pidou, M., Weinberg, H.S., Gernjak, W., 2014. Rejection of Doederer, K., Farre disinfection by-products by RO and NF membranes: influence of solute properties and operational parameters. J. Membr. Sci. 467, 195e205. Du, Y., Lv, X.T., Wu, Q.Y., Zhang, D.Y., Zhou, Y.T., Peng, L., Hu, H.Y., 2017a. Formation and control of disinfection byproducts and toxicity during reclaimed water chlorination: a review. J. Environ. Sci. 58, 51e63. Du, Y., Wu, Q.Y., Lu, Y., Hu, H.Y., Yang, Y., Liu, R., Liu, F., 2017b. Increase of cytotoxicity during wastewater chlorination: impact factors and surrogates. J. Hazard Mater. 324, 681e690. Du, Y., Wu, Q.Y., Lv, X.T., Wang, Q.P., Lu, Y., Hu, H.Y., 2018a. Exposure to solar light reduces cytotoxicity of sewage effluents to mammalian cells: roles of reactive oxygen and nitrogen species. Water Res. 143, 570e578. Du, Y., Wu, Q.Y., Lv, X.T., Ye, B., Zhan, X.M., Lu, Y., Hu, H.Y., 2018b. Electron donating capacity reduction of dissolved organic matter by solar irradiation reduces the cytotoxicity formation potential during wastewater chlorination. Water Res. 145, 94e102. Fang, J.Y., Ling, L., Shang, C., 2013. Kinetics and mechanisms of pH-dependent degradation of halonitromethanes by UV photolysis. Water Res. 47, 1257e1266. Guilherme, S., Rodriguez, M.J., 2014. Occurrence of regulated and non-regulated disinfection by-products in small drinking water systems. Chemosphere 117, 425e432. Hou, S., Ling, L., Shang, C., Guan, Y., Fang, J., 2017. Degradation kinetics and pathways of haloacetonitriles by the UV/persulfate process. Chem. Eng. J. 320, 478e484. Hu, H.Y., Du, Y., Wu, Q.Y., Zhao, X., Tang, X., Chen, Z., 2016. Differences in dissolved organic matter between reclaimed water source and drinking water source. Sci. Total Environ. 551e552, 133e142. Hua, G., Reckhow, D.A., 2007. Comparison of disinfection byproduct formation from chlorine and alternative disinfectants. Water Res. 41, 1667e1678. Jacobs, L.E., Fimmen, R.L., Chin, Y., Mash, H.E., Weavers, L.K., 2011. Fulvic acid mediated photolysis of ibuprofen in water. Water Res. 45, 4449e4458. Jiang, J., Zhang, X., Zhu, X., Li, Y., 2017. Removal of intermediate aromatic halogenated DBPs by activated carbon adsorption: a new approach to controlling halogenated DBPs in chlorinated drinking water. Environ. Sci. Technol. 51, 3435e3444. Khanna, K.K., Jackson, S.P., 2001. DNA double-strand breaks: Signaling, repair and the cancer connection. Nat. Genet. 27, 247e254. Krasner, S.W., Weinberg, H.S., Richardson, S.D., Pastor, S.J., Chinn, R., Sclimenti, M.J., Onstad, G.D., Thruston, A.D., 2006. Occurrence of a new generation of disinfection byproducts. Environ. Sci. Technol. 40, 7175e7185. Krasner, S.W., Kostopoulou, M., Toledano, M.B., Wright, J., Patelarou, E., Kogevinas, M., Villanueva, C.M., Carrasco-Turigas, G., Marina, L.S., Fern andezSomoano, A., Ballester, F., Tardon, A., Grazuleviciene, R., Danileviciute, A., Cordier, S., Costet, N., Righi, E., Aggazzotti, G., Stephanou, E.G., Kargaki, S., Nieuwenhuijsen, M.J., 2016. Occurrence of DBPs in drinking water of European regions for epidemiology studies. Am. Water Work. Assoc. 108, E501eE512. Kristiana, I., McDonald, S., Tan, J., Joll, C., Heitz, A., 2015. Analysis of halogen-specific TOX revisited: Method improvement and application. Talanta 139, 104e110. Li, X.F., Mitch, W.A., 2018. Drinking water disinfection byproducts (DBPs) and human health effects: multidisciplinary challenges and opportunities. Environ. Sci. Technol. 52, 1681e1689. Liu, Y., Mao, R., Tong, Y., Lan, H., Zhang, G., Liu, H., Qu, J., 2017. Reductive dechlorination of trichloroacetic acid (TCAA) by electrochemical process over Pd-In/ Al2O3 catalyst. Electrochim. Acta 232, 13e21. Liu, J., Zhang, X., Li, Y., Li, W., Hang, C., Sharma, V.K., 2019. Phototransformation of halophenolic disinfection byproducts in receiving seawater : kinetics , products , and toxicity. Water Res. 150, 68e76.
W.-C. Huang et al. / Water Research 165 (2019) 115024 Luo, C., Gao, J., Wu, D., Jiang, J., Liu, Y., Zhou, W., Ma, J., 2019. Oxidation of 2,4bromophenol by UV/PDS and formation of bromate and brominated products: a comparison to UV/H2O2. Chem. Eng. J. 358, 1342e1350. McKie, M.J., Taylor-Edmonds, L., Andrews, S.A., Andrews, R.C., 2015. Engineered biofiltration for the removal of disinfection by-product precursors and genotoxicity. Water Res. 81, 196e207. Mills, K.D., Ferguson, D.O., Alt, F.W., 2003. The role of DNA breaks in genomic instability and tumorigenesis. Immunol. Rev. 194, 77e95. Muellner, M.G., Wagner, E.D., Mccalla, K., Richardson, S.D., Woo, Y.T., Plewa, M.J., 2007. Haloacetonitriles vs. regulated haloacetic acids: are nitrogen-containing DBFs more toxic? Environ. Sci. Technol. 41, 645e651. Neta, P., Huie, R.E., Ross, A.B., 1988. Rate constants for reactions of inorganic radicals in aqueous solution. J. Phys. Chem. Ref. Data 17, 1027e1284. Pandit, S., Preston, T.J., King, S.J., Vallance, C., Orr-Ewing, A.J., 2016. Evidence for concerted ring opening and C-Br bond breaking in UV-excited bromocyclopropane. J. Chem. Phys. 144. Plewa, M.J., Kargalioglu, Y., Vankerk, D., Minear, R.A., Wagner, E.D., 2002. Mammalian cell cytotoxicity and genotoxicity analysis of drinking water disinfection by-products. Environ. Mol. Mutagen. 40, 134e142. Plewa, M.J., Simmons, J.E., Richardson, S.D., Wagner, E.D., 2010. Mammalian cell cytotoxicity and Genotoxicity of the haloacetic acids, a major class of drinking water disinfection by-products. Environ. Mol. Mutagen. 51, 871e878. Plewa, M.J., Wagner, E.D., Richardson, S.D., 2017. TIC-Tox: a preliminary discussion on identifying the forcing agents of DBP-mediated toxicity of disinfected water. J. Environ. Sci. 58, 208e216. Redon, C.E., Nakamura, A.J., Martin, O.A., Parekh, P.R., Weyemi, U.S., Bonner, W.M., 2011. Recent Developments in the Use of g-H2AX as a Quantitative DNA Double-Strand Break Biomarker, vol. 3, pp. 168e174. Richardson, S.D., Postigo, C., 2011. Drinking Water Disinfection By-Products. Springer-Verlag. Emerging Organic Contaminants and Human Health. Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., DeMarini, D.M., 2007. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection by-products in drinking water: a review and roadmap for research. Mutat. Res. Rev. Mutat. Res. 636, 178e242. Sayess, R., Khalil, A., Shah, M., Reckhow, D.A., Godri Pollitt, K.J., 2017. Comparative cytotoxicity of six iodinated disinfection byproducts on nontransformed epithelial human colon cells. Environ. Sci. Technol. Lett. 4, 143e148. Singer, P.C., 1999. Humic substances as precursors for potentially harmful disinfection by-products. Water Sci. Technol. 40, 25e30. Tang, X., Wu, Q.Y., Du, Y., Yang, Y., Hu, H.Y., 2014. Anti-estrogenic activity formation potential assessment and precursor analysis in reclaimed water during chlorination. Water Res. 48, 490e497. Taylor, R. , 1990. Interpretation of the correlation coefficient: a basic review. J. Diagn.
11
Med. Sonogr. 6 (1), 35e39. Uyguner, C.S., Bekbolet, M., 2005. A comparative study on the photocatalytic degradation of humic substances of various origins. Desalination 176, 167e176. Vione, D., Falletti, G., Maurino, V., Minero, C., Pelizzetti, E., Malandrino, M., Ajassa, R., Olariu, R.I., Arsene, C., 2006. Sources and sinks of hydroxyl radicals upon irradiation of natural water samples. Environ. Sci. Technol. 40, 3775e3781. Wagner, E.D., Plewa, M.J., 2017. CHO cell cytotoxicity and genotoxicity analyses of disinfection by-products: an updated review. J. Environ. Sci. (China) 58, 64e76. Wang, D.S., Zhao, Y.M., Yan, M.Q., Chow, C.W.K., 2013. Removal of DBP precursors in micro-polluted source waters: a comparative study on the enhanced coagulation behavior. Separ. Purif. Technol. 118, 271e278. Wang, L., Niu, R., Chen, B., Wang, Liwei, Zhang, G., 2017a. A comparison of photodegradation kinetics, mechanisms, and products between chlorinated and brominated/iodinated haloacetic acids in water. Chem. Eng. J. 330, 1326e1333. Wang, W.L., Zhang, X., Wu, Q.Y., Du, Y., Hu, H.Y., 2017b. Degradation of natural organic matter by UV/chlorine oxidation: molecular decomposition, formation of oxidation byproducts and cytotoxicity. Water Res. 124, 251e258. Wang, W.L., Wu, Q.Y., Huang, N., Xu, Z. Bin, Lee, M.Y., Hu, H.Y., 2018. Potential risks from UV/H2O2oxidation and UV photocatalysis: a review of toxic, assimilable, and sensory-unpleasant transformation products. Water Res. 141, 109e125. Wang, W.L., Cai, Y.Z., Hu, H.Y., Chen, J., Wang, J., Xue, G., Wu, Q.Y., 2019. Advanced treatment of bio-treated dyeing and finishing wastewater using ozonebiological activated carbon: a study on the synergistic effects. Chem. Eng. J. 359, 168e175. Watters, G.P., Smart, D.J., Harvey, J.S., Austin, C.A., 2009. H2AX phosphorylation as a genotoxicity endpoint. Mutat. Res. Genet. Toxicol. Environ. Mutagen. 679, 50e58. Wu, Q., Li, C., Wang, W., He, T., Hu, H., Du, Y., Wang, T., 2016. Removal of fluorescence and ultraviolet absorbance of dissolved organic matter in reclaimed water by solar light. J. Environ. Sci. (China) 43, 118e127. Wu, Q.Y., Zhou, Y.T., Li, W., Zhang, X., Du, Y., Hu, H.Y., 2019. Underestimated risk from ozonation of wastewater containing bromide: both organic byproducts and bromate contributed to the toxicity increase. Water Res. 162, 43e52. Xiao, Y., Fan, R., Zhang, L., Yue, J., Webster, R.D., Lim, T.T., 2014. Photodegradation of iodinated trihalomethanes in aqueous solution by UV254 irradiation. Water Res. 49, 275e285. Zalazar, C.S., Labas, M.D., Brandi, R.J., Cassano, A.E., 2007. Dichloroacetic acid degradation employing hydrogen peroxide and UV radiation. Chemosphere 66, 808e815. Zheng, D., Andrews, R.C., Andrews, S.A., Taylor-Edmonds, L., 2015. Effects of coagulation on the removal of natural organic matter, genotoxicity, and precursors to halogenated furanones. Water Res. 70, 118e129.