Chlorinated effluent organic matter causes higher toxicity than chlorinated natural organic matter by inducing more intracellular reactive oxygen species

Science of the Total Environment 701 (2020) 134881

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Chlorinated effluent organic matter causes higher toxicity than chlorinated natural organic matter by inducing more intracellular reactive oxygen species Ye Du a, Wen-Long Wang b, Tao He c, Ying-Xue Sun d, Xiao-Tong Lv a, Qian-Yuan Wu a,⇑, Hong-Ying Hu b a Key Laboratory of Microorganism Application and Risk Control of Shenzhen, Guangdong Provincial Engineering Research Center for Urban Water Recycling and Environmental Safety, Tsinghua Shenzhen International Graduate School, Shenzhen 518055, China b 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, China c South China Institute of Environmental Sciences, Ministry of Ecology and Environment, Guangzhou, China d Department of Environmental Science and Engineering, Beijing Technology and Business University, Beijing 100048, China

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

 EOM formed more N-DBPs and TOX

than NOM during chlorination.  ROS damaged DNA, RNA and proteins

then led to cell cycle arrest, necrosis and apoptosis.  Chlorinated EOM caused higher oxidative stress, cytotoxicity and genotoxicity.  Oxidative stress especially ROS was the major cause of toxicity.

a r t i c l e

i n f o

Article history: Received 18 September 2019 Received in revised form 3 October 2019 Accepted 6 October 2019 Available online 1 November 2019 Editor: Daqiang Yin Keywords: Cytotoxicity Genotoxicity Oxidative stress Effluent organic matter Chlorination

a b s t r a c t During unplanned indirect potable reuse, treated wastewater that contains effluent organic matter (EOM) enters the drinking water source, resulting in different toxicity from natural organic matter (NOM) in surface water during chlorination. This study found that, during chlorination, EOM formed more total organic halogen (TOX) and highly toxic nitrogenous disinfection byproducts (DBPs) like dichloroacetonitrile and trichloronitromethane than NOM did. Oxidative stress including both reactive oxygen species (ROS) and reactive nitrogen species (RNS) in Chinese hamster ovary (CHO) cells substantially increased when exposed to chlorinated EOM and chlorinated NOM. The excessive ROS damaged biological macromolecules including DNA, RNA to form 8-hydroxy-(deoxy)guanosine and proteins to form protein carbonyls. Impaired macromolecule further triggered cell cycle arrest at the S and G2 phases, led to cell apoptosis and eventual necrosis. Cytotoxicity and genotoxicity of chlorinated EOM were both higher than those of chlorinated NOM. Adding the blocker L-buthionine-sulfoximine of intracellular antioxidant glutathione demonstrating that oxidative stress might be responsible for toxicity. ROS was further identified to be the main cause of toxicity induction. These findings highlight the risk from chlorinated EOM in the case of unplanned indirect potable reuse, because it showed higher level of cytotoxicity and genotoxicity than chlorinated NOM via inducing more ROS in mammalian cells. Ó 2019 Elsevier B.V. All rights reserved.

⇑ Corresponding author at: Room 1810, Division of Energy and Environment, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China. E-mail address: [email protected] (Q.-Y. Wu). https://doi.org/10.1016/j.scitotenv.2019.134881 0048-9697/Ó 2019 Elsevier B.V. All rights reserved.

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1. Introduction Unplanned indirect (de facto) potable reuse of wastewater occurs worldwide. Of 2056 drinking water treatment plants (DWTPs) in the United States of America, more than 50% were suffered from the treated wastewater (Rice and Westerhoff, 2014). Along the Yangtze River in China, the ratio of de facto reuse could reach 20% (Wang et al., 2017). In the case of de facto potable reuse, treated wastewater may be blended into or even fully supply as the drinking water source (Rice et al., 2013), subsequently going through the water purification processes in DWTPs. Although many technologies have been developed to wastewater purification (Zhang et al., 2015, 2017), contaminants in wastewater can go into drinking water source due to the improper or incomplete treatment. Chlorination is widely-used to disinfect waters in DWTPs but it produces toxic disinfection byproducts (DBPs) when the disinfectant chlorine reacts with the organic matter (Du et al., 2018a; Liu et al., 2017; Postigo et al., 2018; Richardson et al., 2007). Due to the unlike composition and property between effluent organic matter (EOM) in treated wastewater and the natural organic matter (NOM) in drinking water source (Hu et al., 2016), the formed DBPs and the associated toxicity during chlorination of EOM and NOM may differ a lot. Understanding the differences in toxicity of chlorinated EOM and chlorinated NOM helps know the potential risk of de facto potable reuse. Toxicity to mammalian cells is widely used to evaluate the water quality (Dong et al., 2017; Komaki et al., 2014). Cytotoxicity and genotoxicity are the two indicators that researchers concern most for the toxic effect of DBPs (Du et al., 2013; Yang et al., 2014). Most individual DBPs and DBP mixtures are found to be able to lower the cell viability, causing cytotoxicity (Han et al., 2018; Yang et al., 2014). Besides, DBPs may also cause genotoxicity. Micronuclei assay and alkaline single cell gel electrophoresis assay that characterize chromosome damage and DNA single-strand break, respectively, are previously used to determine the genotoxicity of DBPs (Dong et al., 2017; Maffei et al., 2009). However, there is no clear understanding of how chlorinated EOM and chlorinated NOM influence the genotoxicity of DNA double-strand breaks (DSBs) in mammalian cells. DSBs was considered the worst case of DNA damage, which can activate oncogenes, inactivate tumor suppressors, and change the modifier proteins that influence chemosensitivity and tumor progression, so that organisms are more susceptible to cancers and other diseases (Bassing and Alt, 2004; Wu et al., 2019; Kang et al., 2013). Evaluating the occurrence of DSBs in mammalian cells could give further insights into the risks that arise from chlorinated EOM and chlorinated NOM. Understanding the effects of cytotoxicity and genotoxicity from DBPs are not enough, their toxication mechanism needs to be explored further. Intracellular oxidative stress may rise in the presence of xenobiotic chemicals. Excessive superoxide anions (O2 ) firstly form in mammalian cells, which then may be converted to hydrogen peroxide (H2O2) and hydroxyl radicals (OH) to eliminate the xenobiotic chemicals (Boonstra and Post, 2004). Except for these reactive oxygen species (ROS), oxidative stress also increases when reactive nitrogen species (RNS) such as nitric oxide (NO), dinitrogen trioxide (N2O3), and peroxynitrite (ONOO ) are present _ zon _ at high concentrations (Du et al., 2018b; Drzezd et al., 2018; Maraldi et al., 2011). ROS and RNS in balanced amounts are important for intracellular signal transduction, but the excessive accumulation may lead to the reduced antioxidant protection, destruction of functional biological macromolecules into unusable byproducts, cause genomic instability or unregulated cellular growth stimulation (Halliwell and Whiteman, 2004; Shi et al., 2012; Wang et al., 2013). However, till now, the role of ROS and RNS in the cytotoxicity and genotoxicity induced by chlorinated EOM and chlorinated NOM remains unclear.

This work was designed to study the different toxic effects on mammalian cells during chlorination of EOM and NOM due to the unlike DBP formation. Different toxic impacts between chlorinated of EOM and chlorinated NOM on biological macromolecules and physiological processes were investigated. Furthermore, cytotoxicity and genotoxicity were evaluated and the mechanism of toxication was investigated. 2. Materials and methods 2.1. Chemicals and reagents Chemicals and reagents used for sample pretreatment, chemical analysis and biotoxicity assay were detailed in the Supporting Information (SI, Text S1). 2.2. Water sampling and water quality analysis The standard Suwannee River NOM (2R101N, International Humic Substances Society) was used to simulate the drinking water source according to previous studies (Jiang et al., 2017; Pan et al., 2016). Because standard EOM is commercially unavailable, a secondary effluent sample was collected from a municipal wastewater treatment plant in Southern China, which was treated by anaerobic, anoxic, and oxic (A2O) processes. The secondary effluent sample with low bromide concentration (below the quantitation limit of 10 lg/L) and low ammonia concentration (0.57 mg/L) was specially selected to avoid the interference of inorganic compounds on DBPs formation. NOM was dissolved in ultrapure water to obtain the NOM solution, of which the dissolved organic carbon (DOC) value was the same as the secondary effluent. Before analysis, both NOM solution and secondary effluents were filtered with the 0.45 lm filter. The water quality parameters, including DOC, dissolved organic nitrogen (DON), bromide, ammonia, pH, ultraviolet absorbance, were measured as described elsewhere (Du et al., 2017a). Water quality parameters were measured in triplicate and presented in Table S1. 2.3. Chlorination The 1000 mL of filtered treated wastewater or NOM solution was placed in a glass bottle, adjusted to pH 7 then buffered with a 1 mM phosphate solution at pH 7. Samples were chlorinated by adding 15 mg-Cl2/L NaClO for 1 h with no headspace. The free chlorine residual was quenched with 105% stoichiometric amounts of 0.1 M ascorbic acid. 2.4. Sample pretreatment, DBP analyses and toxicity index calculation The 50 mL chlorinated sample was used to analyze typical individual DBPs. Samples were pretreated by liquid-liquid extraction (LLE) into 5 mL MTBE containing 100 lg/L 1.2-dibromopropane as the internal standard. The method for LLE was detailed elsewhere (Du et al., 2017a). DBPs were analyzed by gas chromatography (7890B, Agilent Technologies, USA) equipped with an electron capture detector using the DB-5MS column. The detection limit of the DBPs was 0.1 lg/L. The 100 mL chlorinated sample was used to measure total organic halogen (TOX), including total organic chlorine (TOCl) and total organic bromine (TOBr), using the method of electrodialysis-photolysis. Detailed protocols were described by Zhang et al. (2018) and Bu et al. (2018). The 800 mL chlorinated sample was concentrated by solid phase extraction (SPE) for toxicity assay. Samples were concentrated with Oasis hydrophilic–lipophilic balance resin cartridges (500 mg, 6 mL, Waters, USA), then

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eluted by 5 mL methanol, 2 mL acetone and 2 mL dichloromethane in sequence. The SPE method is reported in detail previously (Du et al., 2017b). The method for cytotoxicity index and genotoxicity index calculation was referenced to previous study (Chuang et al., 2019). Briefly, toxicity index of an individual DBP was firstly calculated by the measured concentration divided by the LC50 (cytotoxicity) or midpoint of tail moment (genotoxicity, performed by single cell gel electrophoresis assay) reported by the same group (Table S3) (Wagner and Plewa, 2017). The total toxicity index was then obtained by summing the toxicity index of all individual DBPs. 2.5. Cell culture The toxicity was tested on a Chinese hamster ovary cell line (CHO-K1), obtained from the American Type Culture Collection (USA). The cells were cultured following the protocol outlined in a previous study (Du et al., 2017b). The detailed methods are provided in the Text S2. 2.6. Cytotoxicity assay The cell viability was measured in an adenosine triphosphate (ATP) assay using CellTiter–GloÒ luminescent cell viability assay kits (Promega). The luminescent signal, which is proportional to the number of viable cells (Broyer et al., 2013), is recorded by a microplate reader (SpectraMax i3; Molecular Devices, USA). A phenol solution was used as the positive control to quantify the cytotoxicity equivalent, so the cytotoxicity equivalent in this study was expressed in mg-Phenol/L. The ATP assay protocol and cytotoxicity equivalent calculation are detailed in the Text S3. The cell viability tests were performed in triplicate. 2.7. DNA double-strand break measurement After DNA DSBs occur, ataxia telangiectasia mutated protein (ATM) and/or the DNA-dependent protein kinase (DNA-PK) phosphorylate histone H2AX at Ser139 to form pH2AX foci, which then gather further proteins to repair DSBs (Kang et al., 2013; Stiff et al., 2004). Thus, DNA DSBs can be measured by pH2AX foci because of the one to one correspondence. After being exposed to the samples for 24 h, the cells were fixed with paraformaldehyde (Sigma), permeated by Triton X-100 (Amresco), and blocked by bovine serum albumin (BSA, Amresco). The cells were incubated with the primary antibody phospho-Histone H2AX (Rabbit IgG, monoclonal antibody, Cell Signaling), and then were stained with the secondary antibody Alexa FluorÒ 647 conjugate (Anti-rabbit IgG, Cell Signaling) and Hoechst 33342 (Sigma). Images of the cells were obtained using a high content analysis (HCA) system (ImageXpressÒ Micro, Molecular Devices, USA) with a 40 object lens. The pH2AX foci were obtained from the Texas Red (TXR) channel and nucleus DNA were obtained from the 4,6-diamidino-2-phenylindole dihydrochloride (DAPI) channel. The range of DBP concentration that allows cells viability higher than 70% were used to plot the concentration-effect curves (Wagner and Plewa, 2017). The genotoxicity equivalent was calculated with 4-nitroquinoline N-oxide (4-NQO) as the positive control. The protocols for the pH2AX assay and genotoxicity equivalent quantification are detailed in the Text S4. The genotoxicity tests were performed in triplicate. 2.8. ROS and RNS measurement ROS was measured with an intracellular total ROS assay kit (KA4075, Abnova). After the cells were exposed to the samples for 24 h, the fluorescence probe in the kit was added into the cells. The probe specifically reacts with ROS and gives the fluorescent intensity, which was then recorded by the Tetramethylrhodamine

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(TRITC) channel and calculated by MetaXpressÒ software in the HCA system. The ROS equivalent was calculated with hydrogen peroxide (H2O2) as the positive control. The methods for ROS equivalent quantification was similar to the genotoxicity equivalent quantification. Similarly, RNS was measured with an intracellular RNS assay kit (STA-800, OxiselectTM) using a specific NO probe. The fluorescent intensity was recorded by the FITC channel of the HCA system. The cell nucleuses were stained by Hoechst 33342 at the same time so that the cell numbers could be counted. The fluorescence of ROS/RNS per cell was calculated using the total fluorescent intensity divided by the cell number. The ROS and RNS level of each sample were measured in triplicate. 2.9. 8-hydroxy-(deoxy)guanosine (8-OH(d)G) measurement After the cells were exposed to the samples for 24 h, the cells were fixed with paraformaldehyde, permeated with Triton-100, and blocked with bovine serum albumin. The cells were then incubated with the primary antibody (DNA/RNA oxidative damage markers, Thermo), and stained with the secondary antibody Alexa FluorÒ plus 555 (Thermo), and Hoechst 33342. Cell images were obtained with the HCA system with a 10 object lens. The fluorescent intensity of 8-OH(d)G was recorded by the TRITC channel and calculated using MetaXpressÒ software in the HCA. The cell nucleuses were obtained with the DAPI channel. The 8-OH(d)G level per cell was calculated using the total fluorescent intensity divided by the cell number. The 8-OH(d)G level of each sample was measured in triplicate. 2.10. Protein carbonylation measurement After the cells were exposed to the samples for 24 h, they were collected and lysed by freezing and thawing in liquid nitrogen. The protein contents of the lysates were determined using the bicinchoninic acid method, as outlined in the manual of the protein quantification kit (Nanjing Jiancheng Bioengineering Institute, China). The protein carbonyl was determined using the OxiSelectTM protein carbonyl ELISA kit (Cell Biolabs), following the manufacturer’s instructions. The method was also detailed in a previous study (Du et al., 2013). The fully-reduced BSA standards were used as a negative control. The fully-oxidized BSA standards were used to derive the standard curve. The absorbance was recorded by a microplate reader (SpectraMax i3, Molecular Devices, USA) at 450 nm. The protein carbonyl contents in each sample were measured in triplicate and were determined by comparing the absorbance with the standard curve. 2.11. Malondialdehyde (MDA) adducts measurement After the cells were exposed to the samples for 24 h, the protein contents of the cell lysates were determined as outlined above. The malondialdehyde (MDA) adducts were determined using the OxiSelectTM MDA adducts competitive ELISA kit (Cell Biolabs) following the manufacturer’s instructions. The method was detailed in a previous study (Du et al., 2013). The standard curve was derived from the MDA-BSA standards provided by the manufacturer. The absorbance was recorded by a microplate reader at 450 nm. The MDA adduct contents in each sample were measured in triplicate and were determined by comparing the absorbance of the sample with the standard curve. 2.12. Cell cycle assay After the cells were exposed to the samples for 24 h, they were fixed overnight with precooled alcohol at 4 °C. Then the fixed cells were incubated with a working solution that comprised propidium

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iodide (PI, 50 lg/mL PBS), ribonuclease (100 lg/mL PBS), and Triton X-100 (0.2% v/v). DNA would be stained by PI and the fluorescent intensity was detected by the DAPI channel of HCA system. Fluorescence intensity would vary with the phase of cell cycle. Cells at different phases were distinguished by the MetaXpressÒ software in the HCA system.

2.13. Cell apoptosis/necrosis assay Cell apoptosis and necrosis were determined using a FITC Annexin V cell apoptosis kit (BD Pharmingen). Annexin V detects apoptotic cells sensitively by combining phosphatidylserine on the cytomembrane. The kit also included PI, which was used to detect necrotic cells. The assay was conducted according to the manufacturer’s protocol. Briefly, after the cells were exposed to samples for 24 h, they were digested and resuspended in tubes by a binding buffer. Then Annexin V and PI were added to the tubes to stain the cells. The fluorescent intensity was detected by flow cytometry (FACSCalibur, BD, USA). For each sample, 10,000 cells were collected and analyzed by the flow cytometry.

2.14. Statistical analysis One-way analysis of variance (ANOVA) tests were conducted to determine if a sample at a certain concentration factor caused a significant level (p < 0.05) of toxicity against the negative control. If so, a post hoc test of Holm-Sidak multiple comparison versus the negative control analysis was then performed to identify the lowest toxicity concentration. The power of the test statistic (1 b) was set to 0.8 at a = 0.05 (Wu et al., 2019). Regression analysis was used to analyze the concentration-effect curves of each toxicity endpoint. For the cytotoxicity assay, the sample concentration at which the relative cell viability is 50% is defined as the LC50. For the genotoxicity assay, the sample concentration at which the pH2AX induction ratio is 1.5 (over the negative control) is defined as the IR1.5. Cytotoxicity and genotoxicity equivalents are obtained as described in the SI. One-way ANOVA tests were also performed to determine significant difference between toxicity equivalents.

Fig. 1. DBP formation during the chlorination of EOM and NOM (TCM: trichloromethane; BDCM: bromodichloromethane; DCP: dichloropropanone; TCP: trichloropropanone; CH: chlorohydrate; TCAcAm: trichloroacetamide; DCAN: dichloroacetonitrile; TCAN: trichloroacetonitrile; TCNM: trichloronitromethane; TOCl: total organic chlorine. * represents the significant difference between DBPs generated from EOM and NOM; # means the DBP was below detection limit).

3.2. Oxidative stress induced by chlorinated EOM and chlorinated NOM Intracellular oxidative stress could arise when mammalian cells were exposed to xenobiotics. After being exposed to chlorinated EOM and chlorinated NOM, significant higher level of oxidative stress in CHO cells, including intracellular ROS and RNS, was observed than the negative control (Fig. 2). The concentrationeffect curve and ROS equivalents of samples were shown in Fig. S1. This was in accordance with previous studies that some individual DBPs, like haloacetic acid and halobenzoquinones, could increase the oxidative stress (Li et al., 2016; Pals et al., 2013). Besides, among the samples, chlorinated EOM showed stronger ability to both induce the ROS and RNS. 3.3. Toxication pathway of chlorinated EOM and chlorinated NOM

3. Results and discussion 3.1. DBP formation during chlorination of EOM and NOM Thirteen kinds of typical individual DBPs, including trihalomethanes, haloacetones, haloacetonitriles, halonitromethanes, haloacetamides and the overall index TOX were measured. Most brominated DBPs and TOBr were not detected due to the low concentration of bromide in NOM solution and EOM samples (Table S1). During chlorination, although NOM generated more trichloromethane (TCM) and dichloropropanone (DCP), EOM formed higher concentration of chlorohydrate (CH) and all kinds of nitrogenous DBPs (N-DBPs), including dichloroacetonitrile (DCAN), trichloroacetonitrile (TCAN), trichloronitromethane (TCNM) and trichloroacetamide (TCAcAm) (Fig. 1). It was reported that NDBPs are much toxic than carbonaceous DBPs (Muellner et al., 2007). DON serves as the important precursor of N-DBPs (Bond et al., 2012) and higher DON concentration was observed in EOM (0.35 mg-N/L) than that in NOM (0.09 mg-N/L). This was in accordance with previous studies that EOM generally contains more DON and forms more N-DBPs during chlorination (Hu et al., 2016).

3.3.1. Oxidative damage to biological macromolecules After being accumulated excessively, how oxidative stress leads to the toxicity was further studied. Biological macromolecules either serve as important components of organisms or play critical roles in metabolism activities (Butterfield et al., 2010). After excessive ROS weakening the antioxidant system, it may attack biological macromolecules like DNA, RNA, proteins and lipids (Boonstra and Post, 2004; Du et al., 2013). Oxidative damage to these macromolecules in CHO cells induced by oxidative stress was investigated. The 8-hydroxy-deoxyguanosine (8-OHdG) and 8-hydroxyguanosine (8-OHG) are base modification products that form when ROS attacks the C-8 position of guanine in deoxyguanosine and guanosine, causing DNA and RNA damage, respectively (Col et al., 2010; Shi et al., 2012). Typical images obtained by HCA system for 8-OH(d)G analysis are given in Fig. S2 and the statistical results are shown in Fig. 3a. All the samples enhanced 8-OH(d)G formation against the negative control (p < 0.05). Chlorinated EOM showed the strongest ability to induce 8-OH(d)G (20-fold concentrated sample induced 2 times of the fluorescence intensity over negative control). As well as matching with cytosine normally, 8-OHdG mismatches with adenine, resulting in gene mutation (Col et al., 2010).

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Fig. 2. Increase in oxidative stress when CHO cells were exposed to chlorinated EOM and chlorinated NOM. (a) Reactive oxygen species (ROS); (b) Reactive nitrogen species (RNS). (* represents the significant difference against the negative control.)

Fig. 3. Formation of oxidative damage markers when CHO cells were exposed to chlorinated EOM and chlorinated NOM. (a) DNA and RNA damage marker 8-OH(d) G; (b) protein damage marker protein carbonyls; (c) lipid damage marker MDA adduct (* represents the significant difference against the negative control).

The damaged 8-OHG in RNA also interferes with the normal expression of proteins (Shi et al., 2012). The high levels of 8-OH (d)G were reported to correlate with cancer and cardiovascular disease (Kryston et al., 2011). Therefore, chlorinated EOM might pose serious risks to health during unplanned potable reuse because it showed the strongest ability to induce 8-OH(d)G. Carbonylation of proteins can form at breakage sites where peptides were attacked by ROS, or on the residues of lysine, arginine, and threonine that are oxidized by ROS (Berlett and Stadtman, 1997; Suzuki et al., 2010). Protein carbonyls only form under severe oxidative damage and seriously affect protein function (Brooks and Gu, 2006). All the samples led to the protein carbonyls in CHO cells increased as the concentration factor increased (Fig. 3b). It showed that chlorinated EOM and chlorinated NOM were similarly able to induce protein carbonyls. When ROS is present at excessive levels, it might react with the polyunsaturated fatty acids in membrane lipids and causes lipid peroxidation, impair membrane permeability, modify lipid-protein interactions and form bioactive degradation products (Boonstra and Post, 2004; Ye et al., 2010). MDA adducts can be used as an indicator of lipid oxidative damage. However, no obvious concentrationeffect relationships between samples and MDA adducts were found (Fig. 3c). These results are similar to those of Du et al. (2013), who investigated the MDA adduct contents in mammalian cells exposed to an emerging class of DBP halobenzoquinones. The results suggested that membrane lipid may not be attacked by ROS resulted from chlorinated EOM and chlorinated NOM.

3.3.2. Disturbance of physiological processes Physiological processes in mammalian cells are closely regulated by various macromolecules, damages to DNA, RNA and proteins induced by ROS might further interfere with physiological processes. Cell cycles are regulated by proteins like cyclins and cyclin-dependent kinases (CDKs) (Uhlmann et al., 2011). A typical cycle of eukaryotic cells includes G1 (protein and lipid synthesis), S (DNA synthesis), G2 (ensuring DNA replication is completed), and mitotic (M) (cell division) phases. Some cells, when temporarily outside the cycle, may be in the G0 phase (Graña and Reddy, 1995). Distribution of CHO cells in different phases of the cell cycle is shown in Fig. 4a. Ratio of cells in the G2 phase almost increased in all the samples except the non-chlorinated NOM, meaning that cell cycle arrest mainly occurred in the G2 phase. The G2 phase ratio was highest for chlorinated EOM, changing from 19.3% to 53.6% when being exposed to 60-fold chlorinated EOM. Ratio of the S phase also increased, for example, from 9.8% to 14.1% when being exposed to 60-fold chlorinated EOM. The results that cell cycle arrested in the S and G2 phases indicates that chlorinated EOM and NOM hindered the normal replication of DNA, which is consistent with the 8-OHdG formation. Apoptosis is the last line of defense against oxidative stress (Boonstra and Post, 2004). ROS could trigger the release of cytochrome C and damage mitochondrial DNA, resulting in apoptosis (Circu and Aw, 2010). Cells will start the apoptosis program to eliminate the damaged cells and will even undergo necrosis in severe cases. Fig. S3 shows the images for apoptosis/necrosis analysis

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Fig. 5. (a) Cytotoxicity index and genotoxicity index based on the DBP concentration; (b) Cytotoxicity equivalents and genotoxicity equivalents of chlorinated EOM and chlorinated NOM based on the cell viability assay and DSBs assay. (* represents the significant difference against the sample without chlorination.) Fig. 4. Disturbance to the physiological process when CHO cells were exposed to chlorinated EOM and chlorinated NOM. (a) Cell cycle arrest; (b) Cell apoptosis and necrosis.

obtained by flow cytometry. The ratio of normal, early apoptotic, late apoptotic, and necrotic cells is shown in Fig. 4b. Apoptosis, including early apoptosis and late apoptosis, was induced in all the samples. Early apoptotic cells may return to normal cells when stimulations disappear but late apoptotic cells will inevitably die (Elmore, 2007). At each concentration, the apoptosis ratio was generally higher for chlorinated EOM than chlorinated NOM. Further, necrosis was observed when cells were exposed to all the samples, indicating that severe and irreversible cell damage occurred. It was also found that chlorinated EOM resulted in the highest ratio of necrosis. 3.4. Cytotoxicity and genotoxicity of chlorinated EOM and NOM Based on the measured DBP concentrations, cytotoxicity index and genotoxicity index were calculated according to the LC50 and midpoint of Tail moment (single cell gel electrophoresis assay) (Table S2) reported by Jeong et al. (2015), Muellner et al. (2007), Plewa et al. (2008, 2004), , Plewa and Wagner (2009). Chlorinated EOM showed higher cytotoxicity index and genotoxicity index than chlorinated NOM (Fig. 5a). DCAN contributed most to the

cytotoxicity while both DCAN and TCNM contributed much to the genotoxicity, indicating N-DBPs might play important role in inducing higher toxicity of chlorinated EOM. The above calculated toxicity index from the limited individual DBPs represent only a small part of the overall toxicity, as many DBPs are not identified. The actual toxicity of the organic extracts was thusly more important and measured in this study. Fig. S4 shows concentration-effect curves of samples and phenol in the cytotoxicity assay. The LC50 values were listed in Table S3. Cytotoxicity equivalents of EOM, chlorinated EOM, NOM and chlorinated NOM were 7.8, 20.2, 3.0 and 12.2 mg-Phenol/L, respectively (Fig. 5b). Concentration-effect curves of the pH2AX foci in the DSB assay are shown in Fig. S5 (see typical images of pH2AX foci in Fig. S6). Genotoxicity equivalents of EOM, chlorinated EOM, NOM and chlorinated NOM were 1.9, 8.2, 0.6 and 5.0 lg-4-NQO/ L, respectively. Both cytotoxicity and genotoxicity of EOM and NOM increased after chlorination, mainly due to the DBP formation. Chlorinated EOM was more cytotoxic, and showed higher ability to induce DSBs in CHO cells than chlorinated NOM. This was in accordance with the results of above discussed toxicity index. The difference in toxicity might be explained by the TOX concentration (877 lgCl/L for chlorinated EOM and 675 lg-Cl/L for chlorinated NOM, Fig. 1). Previous studies found that water samples contained higher

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TOX concentration generally showed higher toxicity during chlorination (Du et al., 2017b; Li et al., 2017). These results suggested that, in the case of unplanned indirect potable reuse, EOM might pose higher risk during chlorination. 3.5. Roles of ROS and RNS in toxicity induction To explore the influence of oxidative stress on the toxicity induced by chlorinated EOM and chlorinated NOM, a blocker of the key antioxidant glutathione (GSH) in mammalian cells, Lbuthionine-sulfoximine (BSO), was added to destroy the intracellular antioxidant system. GSH is a tripeptide that consists of c-Lglutamine, L-cysteine, and glycine, and is synthesized by cglutamylcysteine ligase (c-GCL) and GSH synthetase, with c-GCL as the rate-limiting enzyme (Lee et al., 2008). BSO can selectively and irreversibly inhibit c-GCL (Lee et al., 2008), thus blocking GSH synthesis and damaging the oxidation resistance. The influences of BSO on the cytotoxicity and genotoxicity induced by chlorinated EOM and NOM are shown in Fig. 6a and Fig. 6b, respectively. For cytotoxicity test, the added BSO concentration was 100 lM. For genotoxicity test, to ensure the cell viability >70%, only 10 lM was added. Adding BSO alone showed little

Fig. 6. (a) Influence of L-buthionine-sulfoximine (BSO) on the cytotoxicity induced by chlorinated EOM and chlorinated NOM (BSO 100 lM); (b) Influence of BSO on the genotoxicity induced by chlorinated EOM and chlorinated NOM (BSO 10 lM). # means that data were not presented because cell viabilities were <70% after adding BSO; * represents the significant difference against the sample without BSO addition).

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impact on the toxicity to CHO cells (Figs. S7 and S8). The CHO cell viability was 84% when exposed to 10-fold chlorinated EOM, but substantially decreased to 9% when BSO (100 lM) was added. The number of pH2AX foci was 9.3/cell when exposed to 9-fold chlorinated EOM, but significantly increased to 16.9/cell after being co-exposed to 10 lM BSO. These results suggested that, destroying the antioxidant GSH in cells substantially aggravated the cytotoxicity and genotoxicity, demonstrating the toxicity might mainly resulted from the elevated oxidative stress induced by chlorinated samples. The oxidative stress, including both ROS and RNS level were elevated in CHO cells when exposed to the samples. To further distinguish the roles of ROS and RNS in inducing cytotoxicity and genotoxicity, N-acetylcysteine (L-NAC), a ROS scavenger, and nitroarginine (L-NNA), a RNS inhibitor, were added to cell cultures at the same time when adding chlorinated EOM and chlorinated NOM. L-NAC is an antioxidant that is frequently used to scavenge intracellular ROS generated by xenobiotic pollutants (Du et al., 2013; Lin et al., 2013). The reductive L-NAC reacts directly with ROS (Samuni et al., 2013). RNS (NO) forms when nitric oxide synthase catalyzes the substrate L-arginine. As a competitor of Larginine, the substrate L-NNA will not produce NO (Cherng et al., 2011). The influences of L-NAC (500 lM) and L-NNA (500 lM) on the cytotoxicity and genotoxicity are shown in Fig. 7a and Fig. 7b, respectively. The added concentration of L-NAC and L-NNA themselves showed little toxic effect on CHO cells (Figs. S7 and S8). When L-NAC was added, the number of pH2AX foci decreased significantly and cell viability also increased considerably (p < 0.05). For example, when L-NAC was added, the cell viability

Fig. 7. Influence of L-NAC and L-NNA on the (a) cytotoxicity and (b) genotoxicity induced by chlorinated EOM and chlorinated NOM (L-NAC 500 lM, L-NNA 500 lM, * represents the significant difference against the sample without inhibitor addition).

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increased from 11% to 80% when exposed to 50-fold chlorinated EOM. However, the L-NNA did not attenuate the cytotoxicity and genotoxicity (p > 0.05), which suggests that RNS at excessive levels may have little impact on the toxicity induced by chlorinated EOM and NOM. When L-NAC and L-NNA were both added, the cell viability and pH2AX foci numbers were almost the same as those when only L-NAC was added, demonstrating that ROS was most important in cytotoxicity and genotoxicity induced by chlorinated EOM and chlorinated NOM. Among the samples, chlorinated EOM showed the strongest ability to induce ROS (fluorescence intensity divided by the concentration factor), again emphasizing the potential risk from chlorinated EOM. 4. Conclusion In cases of unplanned indirect potable reuse, there may be higher risks from EOM in treated wastewater than NOM in drinking water source during chlorination. The DBP formation, cytotoxicity, genotoxicity and toxication mechanism of chlorinated EOM and chlorinated NOM were investigated in this study. The main conclusions are drawn below: (1) During chlorination, EOM formed more N-DBPs and TOX than NOM did. Toxicity index results showed typical DBPs from EOM chlorination were more cytotoxic and genotoxic than those from NOM chlorination. DCAN contributed most to cytotoxicity while both DCAN and TCNM contributed much to genotoxicity. (2) Oxidative stress in CHO cells substantially increased when exposed to chlorinated EOM and chlorinated NOM. The elevated ROS damaged biological macromolecules including DNA, RNA and proteins, causing increases in both 8-OH(d) G and protein carbonyls. Impaired macromolecule further triggered cell cycle arrest at the S and G2 phases, led to cell apoptosis and eventual necrosis. (3) Chlorination increased the cytotoxicity and genotoxicity of both EOM and NOM. The cytotoxicity and genotoxicity of chlorinated EOM were higher than those of chlorinated NOM, indicating the potential risk in the case of unplanned indirect potable reuse. (4) Oxidative stress might be responsible for the cytotoxicity and genotoxicity as demonstrated by adding the GSH blocker BSO. Among the reactive species that increased the oxidative stress, ROS was further identified as the main cause of toxicity induction.

Acknowledgements This study was supported by the National Natural Science Foundation of China (No. 51678332/51738005), the Fundamental Research Program of Shenzhen Science and Technology innovation Committee (JCYJ20170818091859147), the Development and Reform Commission of Shenzhen Municipality, the Fundamental Research Program for the State Level Public Welfare Research Institutes (PM-zx097-201602-059) and Science and Technology Program of Guangdong, China (2017A020216003), the Special support program for high-level personnel recruitment in Guangdong Province (2016TQ03Z384). Appendix A. Supplementary material Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2019.134881.

References Bassing, C.H., Alt, F.W., 2004. The cellular response to general and programmed DNA double strand breaks. DNA Repair 3 (8), 781–796. Berlett, B.S., Stadtman, E.R., 1997. Protein oxidation in aging, disease, and oxidative stress. J. Biol. Chem. 272 (33), 20313–20316. Bond, T., Templeton, M.R., Graham, N., 2012. Precursors of nitrogenous disinfection by-products in drinking water––a critical review and analysis. J. Hazard. Mater. 235, 1–16. Boonstra, J., Post, J.A., 2004. Molecular events associated with reactive oxygen species and cell cycle progression in mammalian cells. Gene 337, 1–13. Brooks, C.L., Gu, W., 2006. p53 ubiquitination: Mdm2 and beyond. Mol. Cell 21 (3), 307–315. Broyer, L., Goetsch, L., Broussas, M., 2013. Evaluation of complement–dependent cytotoxicity using ATP measurement and C1q/C4b binding. In: Glycosylation Engineering of Biopharmaceuticals. Humana Press, Totowa, NJ, pp. 319–329. Bu, Y., Song, M., Han, J., Zhang, Z., Chen, B., Zhang, X., Yang, M., 2018. A facile and green pretreatment method for nonionic total organic halogen (NTOX) analysis in water–Step II. Using photolysis to convert NTOX completely into halides. Water Res. 145, 579–587. Butterfield, D.A., Galvan, V., Lange, M.B., Tang, H., Sowell, R.A., Spilman, P., Fombonne, J., Gorostiza, O., Zhang, J., Sultana, R., Bredesen, D.E., 2010. In vivo oxidative stress in brain of Alzheimer disease transgenic mice: requirement for methionine 35 in amyloid b-peptide of APP. Free Radical Bio. Med. 48 (1), 136–144. Cherng, T.W., Paffett, M.L., Jackson-Weaver, O., Campen, M.J., Walker, B.R., Kanagy, N.L., 2011. Mechanisms of diesel-induced endothelial nitric oxide synthase dysfunction in coronary arterioles. Environ. Health. Persp. 119 (1), 98–103. Chuang, Y.H., Szczuka, A., Mitch, W.A., 2019. Comparison of toxicity-weighted disinfection byproduct concentrations in potable reuse waters to conventional drinking waters as a new approach to assess the quality of advanced treatment train waters. Environ. Sci. Technol. 2019 (53), 3729–3738. Circu, M.L., Aw, T.Y., 2010. Reactive oxygen species, cellular redox systems, and apoptosis. Free Radical Bio. Med. 48 (6), 749–762. Col, C., Dinler, K., Hasdemir, O., Buyukasik, O., Bugdayci, G., 2010. Oxidative stress and lipid peroxidation products: effect of pinealectomy or exogenous melatonin injections on biomarkers of tissue damage during acute pancreatitis. Hepatob. Pancreat. Dis. 9 (1), 78–82. Dong, S., Nguyen, T.H., Plewa, M.J., 2017. Comparative mammalian cell cytotoxicity of wastewater with elevated bromide and iodide after chlorination, chloramination, or ozonation. J. Environ. Sci. 58, 296–301. _ zon, _ Drzezd J., Jacewicz, D., Chmurzyn´ski, L., 2018. The impact of environmental contamination on the generation of reactive oxygen and nitrogen species – consequences for plants and humans. Environ. Int. 119, 133–151. Du, H., Li, J., Moe, B., McGuigan, C.F., Shen, S., Li, X.F., 2013. Cytotoxicity and oxidative damage induced by halobenzoquinones to T24 bladder cancer cells. Environ. Sci. Technol. 47 (6), 2823–2830. Du, Y., Zhang, X., Li, C., Wu, Q.Y., Huang, H., Hu, H.Y., 2017a. Transformation of DON in reclaimed water under solar light irradiation leads to decreased haloacetamide formation potential during chloramination. J. Hazard. Mater. 340, 319–325. 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, 681–690. Du, Y., Wu, Q.Y., Lv, X.T., Ye, B., Zhan, X.M., Lu, Y., Hu, H.Y., 2018a. Electron donating capacity reduction of dissolved organic matter by solar irradiation reduces the cytotoxicity formation potential during wastewater chlorination. Water Res. 145, 94–102. Du, Y., Wu, Q.Y., Lv, X.T., Wang, Q.P., Lu, Y., Hu, H.Y., 2018b. Exposure to solar light reduces cytotoxicity of sewage effluents to mammalian cells: Roles of reactive oxygen and nitrogen species. Water Res. 143, 570–578. Elmore, S., 2007. Apoptosis: a review of programmed cell death. Toxicol. Pathol. 35 (4), 495–516. Graña, X., Reddy, E.P., 1995. Cell cycle control in mammalian cells: role of cyclins, cyclin dependent kinases (CDKs), growth suppressor genes and cyclindependent kinase inhibitors (CKIs). Oncogene 11 (2), 211–220. Halliwell, B., Whiteman, M., 2004. Measuring reactive species and oxidative damage in vivo and in cell culture: how should you do it and what do the results mean?. Br. J. Pharmacol. 142 (2), 231–255. Han, Y., Ma, M., Li, N., Hou, R., Huang, C., Oda, Y., Wang, Z., 2018. Chlorination, chloramination and ozonation of carbamazepine enhance cytotoxicity and genotoxicity: Multi-endpoint evaluation and identification of its genotoxic transformation products. J. Hazard. Mater. 342, 679–688. 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. 551, 133–142. 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 (6), 3435–3444. Jeong, C.H., Postigo, C., Richardson, S.D., Simmons, J.E., Kimura, S.Y., Marinas, B.J., Barcelo, D., Liang, P., Wagner, E.D., Plewa, M.J., 2015. Occurrence and comparative toxicity of haloacetaldehyde disinfection byproducts in drinking water. Environ. Sci. Technol. 49, 13749–13759. Kang, M.A., So, E.Y., Simons, A.L., Spitz, D.R., Ouchi, T., 2013. DNA damage induces reactive oxygen species generation through the H2AX-Nox1/Rac1 pathway. Cell Death Dis. 3 (1), e249.

Y. Du et al. / Science of the Total Environment 701 (2020) 134881 Komaki, Y., Mariñas, B.J., Plewa, M.J., 2014. Toxicity of drinking water disinfection byproducts: cell cycle alterations induced by the monohaloacetonitriles. Environ. Sci. Technol. 48 (19), 11662–11669. Kryston, T.B., Georgiev, A.B., Pissis, P., Georgakilas, A.G., 2011. Role of oxidative stress and DNA damage in human carcinogenesis. Mut. Res. 711 (1–2), 193–201. Lee, H.R., Cho, J.M., Shin, D.H., Yong, C.S., Choi, H.G., Wakabayashi, N., Kwak, M.K., 2008. Adaptive response to GSH depletion and resistance to L-buthionine-(S, R)sulfoximine: involvement of Nrf2 activation. Mol. Cell. Biochem. 318 (1–2), 23– 31. Li, J., Moe, B., Vemula, S., Wang, W., Li, X.F., 2016. Emerging disinfection byproducts, halobenzoquinones: effects of isomeric structure and halogen substitution on cytotoxicity, formation of reactive oxygen species, and genotoxicity. Environmental Sci. Technol. 50 (13), 6744–6752. Li, Y., Zhang, X., Yang, M., Liu, J., Li, W., Graham, N.J., Li, X., Yang, B., 2017. Three-step effluent chlorination increases disinfection efficiency and reduces DBP formation and toxicity. Chemosphere 168, 1302–1308. Lin, H., Liu, X., Yu, J., Hua, F., Hu, Z., 2013. Antioxidant n-acetylcysteine attenuates hepatocarcinogenesis by inhibiting ros/er stress in tlr2 deficient mouse. Plos One 8 (10), e74130. Liu, Y., Zhang, Q., Hong, Y., 2017. Formation of disinfection byproducts from accumulated soluble products of oleaginous microalga after chlorination. Front. Environ. Sci. Eng. 11 (6), 1. Maffei, F., Carbone, F., Forti, G.C., Buschini, A., Poli, P., Rossi, C., Marabini, L., Radice, S., Chiesara, E., Hrelia, P., 2009. Drinking water quality: an in vitro approach for the assessment of cytotoxic and genotoxic load in water sampled along distribution system. Environ. Int. 35 (7), 1053–1061. Maraldi, T., Riccio, M., Zambonin, L., Vinceti, M., De Pol, A., Hakim, G., 2011. Low levels of selenium compounds are selectively toxic for a human neuron cell line through ROS/RNS increase and apoptotic process activation. Neurotoxicology 32 (2), 180–187. 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 DBPs more toxic?. Environ. Sci. Technol. 41, 645–651. Pals, J., Attene-Ramos, M.S., Xia, M., Wagner, E.D., Plewa, M.J., 2013. Human cell toxicogenomic analysis linking reactive oxygen species to the toxicity of monohaloacetic acid drinking water disinfection byproducts. Environ. Sci. Technol. 47 (21), 12514–12523. Pan, Y., Zhang, X., Li, Y., 2016. Identification, toxicity and control of iodinated disinfection byproducts in cooking with simulated chlor(am)inated tap water and iodized table salt. Water Res. 88, 60–68. Plewa, M.J., Muellner, M.G., Richardson, S.D., Fasano, F., Buettner, K.M., Woo, Y.T., McKague, A.B., Wagner, E.D., 2008. Occurrence, synthesis and mammalian cell cytotoxicity and genotoxicity of haloacetamides: an emerging class of nitrogenous drinking water disinfection by-products. Environ. Sci. Technol. 42, 955–961. Plewa, M.J., Wagner, E.D., 2009. Mammalian Cell Cytotoxicity and Genotoxicity of Disinfection by-Products. Water Research Foundation, Denver CO. Plewa, M.J., Wagner, E.D., Jazwierska, P., Richardson, S.D., Chen, P.H., McKague, A.B., 2004. Halonitromethane drinking water disinfection byproducts: chemical characterization and mammalian cell cytotoxicity and genotoxicity. Environ. Sci. Technol. 38, 62–68. Postigo, C., Emiliano, P., Barceló, D., Valero, F., 2018. Chemical characterization and relative toxicity assessment of disinfection byproduct mixtures in a large drinking water supply network. J. Hazard. Mater. 359, 166–173.

9

Rice, J., Westerhoff, P., 2014. Spatial and temporal variation in de facto wastewater reuse in drinking water systems across the USA. Environ. Sci. Technol. 49 (2), 982–989. Rice, J., Wutich, A., Westerhoff, P., 2013. Assessment of de facto wastewater reuse across the US: trends between 1980 and 2008. Environ. Sci. Technol. 47 (19), 11099–11105. 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. 636 (1), 178–242. Samuni, Y., Goldstein, S., Dean, O.M., Berk, M., 2013. The chemistry and biological activities of N-acetylcysteine. BBA-Gen Subjects 1830 (8), 4117–4129. Shi, F., Nie, B., Gan, W., Zhou, X.Y., Takagi, Y., Hayakawa, H., Sekiguchi, M., Cai, J.P., 2012. Oxidative damage of DNA, RNA and their metabolites in leukocytes, plasma and urine of Macaca mulatta: 8-oxoguanosine in urine is a useful marker for aging. Free Radical Res. 46 (9), 1093–1098. Suzuki, Y.J., Carini, M., Butterfield, D.A., 2010. Protein carbonylation. Antioxid. Redox Sign. 12 (3), 323–325. Stiff, T., O’Driscoll, M., Rief, N., Iwabuchi, K., Löbrich, M., Jeggo, P.A., 2004. ATM and DNA-PK function redundantly to phosphorylate H2AX after exposure to ionizing radiation. Cancer Res. 64 (7), 2390–2396. Uhlmann, F., Bouchoux, C., Lopez-Avilés, S., 2011. A quantitative model for cyclindependent kinase control of the cell cycle: revisited. Phil. Trans. R. Soc. B 366 (1584), 3572–3583. Wang, S., Tian, D., Zheng, W., Jiang, S., Wang, X., Andersen, M.E., Zheng, Y., He, G., Qu, W., 2013. Combined exposure to 3-chloro-4-dichloromethyl-5-hydroxy-2 (5H)furanone and microsytin-LR increases genotoxicity in Chinese hamster ovary cells through oxidative stress. Environ. Sci. Technol. 47 (3), 1678–1687. Wang, Z., Shao, D., Westerhoff, P., 2017. Wastewater discharge impact on drinking water sources along the Yangtze River (China). Sci. Total Environ. 599, 1399– 1407. Wagner, E.D., Plewa, M.J., 2017. CHO cell cytotoxicity and genotoxicity analyses of disinfection by-products: an updated review. J. Environ. Sci. 58, 64–76. 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, 43–52. Yang, Y., Komaki, Y., Kimura, S.Y., Hu, H.Y., Wagner, E.D., Mariñas, B.J., Plewa, M.J., 2014. Toxic impact of bromide and iodide on drinking water disinfected with chlorine or chloramines. Environ. Sci. Technol. 48 (20), 12362–12369. Ye, Y., Liu, J., Chen, M., Sun, L., Lan, M., 2010. In vitro toxicity of silica nanoparticles in myocardial cells. Environ. Toxicol. Pharmacol. 29 (2), 131–137. Zhang, X., Huang, Q., Deng, F., Huang, H., Wan, Q., Liu, M., Wei, Y., 2017. Musselinspired fabrication of functional materials and their environmental applications: progress and prospects. Appl. Mater. Today 7, 222–238. Zhang, X., Huang, Q., Liu, M., Tian, J., Zeng, G., Li, Z., Wang, K., Zhang, Q., Wan, Q., Deng, F., Wei, Y., 2015. Preparation of amine functionalized carbon nanotubes via a bioinspired strategy and their application in Cu2+ removal. Appl. Surf. Sci. 343, 19–27. Zhang, Y., Bu, Y., Han, J., Liu, Y., Chen, B., Zhang, X., Yang, M., Sui, Y., 2018. A facile and green pretreatment method for nonionic total organic halogen (NTOX) analysis in water–Step I. Using electrodialysis to separate NTOX and halides. Water Res. 145, 631–639.