Disinfection byproducts in drinking water and regulatory compliance: A critical review

Abstract

Disinfection by-products (DBPs) are regulated in drinking water in a number of countries. This critical review focuses on the issues associated with DBP regulatory compliance, including methods for DBP analysis, occurrence levels, the regulation comparison among various countries, DBP compliance strategies, and emerging DBPs. The regulation comparison between China and the United States (US) indicated that the DBP regulations in China are more stringent based on the number of regulated compounds and maximum levels. The comparison assessment using the Information Collection Rule (ICR) database indicated that the compliance rate of 500 large US water plants under the China regulations is much lower than that under the US regulations (e.g. 62.2% versus 89.6% for total trihalomethanes). Precursor removal and alternative disinfectants are common practices for DBP regulatory compliance. DBP removal after formation, including air stripping for trihalomethane removal and biodegradation for haloacetic acid removal, have gained more acceptance in DBP control. Formation of emerging DBPs, including iodinated DBPs and nitrogenous DBPs, is one of unintended consequences of precursor removal and alternative disinfection. At much lower levels than carbonaceous DBPs, however, emerging DBPs have posed higher health risks.

Keywords

Disinfection byproducts (DBPs) / drinking water standards / regulatory compliance / alternative disinfection / information collection rule (ICR) / emerging DBPs

1 Introduction

Regarded as one of the most significant advancements in public health over the past decade, disinfection plays a vital role in the processes of water and wastewater treatment (Ksibi, 2006; Richardson et al., 2007; He et al., 2023; Ocal et al., 2023; Zhang et al., 2023). Historically, chlorine and ozone have been the primary chemical disinfectants used in water treatment (Cai et al., 2023; Hogard et al., 2023; Morrison et al., 2023; Jiang et al., 2024). There has been an increasing interest in the application of peroxymonosulfate (PMS) and peroxydisulfate (PDS) in recent years (Zuo et al., 2024).

Chlorine is globally recognized as the most commonly used oxidant for primary drinking water disinfection, displaying high reactivity toward phenolic compounds and achieving remarkable success in eradicating harmful microorganisms and acute waterborne diseases (Liu et al., 2013; Criquet et al., 2015; Zhu and Zhang, 2016). As a potent oxidizing agent, ozone demonstrates considerable efficacy in the removal of electron-rich functional compounds, including olefins, phenols and amines, pathogens, and certain micro-pollutants (Snyder et al., 2006; Yang et al., 2019; Wu et al., 2021b). It finds relatively extensive application for rapid deodorization and disinfection of wastewater while enhancing the biodegradability of organic matter (Mao et al., 2016; Wu et al., 2019; Wu et al., 2021a). PMS/PDS usually require an input of energy to activate them, and can effectively degrade organic pollutants in water (Wang et al., 2020a; Du et al., 2023). The SO₄·⁻ produced during activation can selectively oxidize biomolecules and macromolecules with electron-rich groups, making them more suitable for inactivating microorganisms (Zhou et al., 2023). In addition, SO₄·⁻ can also effectively oxidize endocrine disruptors, herbicides, phenols, perfluorinated compounds, and chlorine compounds (Fang and Shang, 2012). Table S1 in the Supplementary material provides a concise comparison of these three types of disinfectants.

However, the reaction of disinfectants and organic or inorganic substances in water can generate harmful disinfection byproducts (DBPs), which may pose potential risk to human health and ecological security (Richardson et al., 2007; Heeb et al., 2014; Criquet et al., 2015). As an important water matrix component, bromide may react with disinfectants to form bromate and organic bromine (Heeb et al., 2014). Bromate is categorized as a class II-B carcinogen and has garnered significant attention. The World Health Organization (WHO), the European Union (EU), and the USA have all established a maximum allowable limit for bromate in drinking water at 10 μg/L (Legube et al., 2004; Wu et al., 2019).

Organic bromines, which mainly refers to brominated disinfection byproducts (Br-DBPs) are also of concern due to their high toxicity and complex nature. While new Br-DBPs continue to be identified, a substantial number remain unknown (Richardson et al., 2002; Plewa et al., 2004; Zhang and Yang, 2018; Wu et al., 2020a). There is still a lack of compilation and comparative analysis of the toxicity of identified Br-DBPs, which impedes the identification of key toxic components. In addition, relying solely on a limited number of known Br-DBPs does not adequately predict the toxicity risk posed by the overall Br-DBPs. Therefore, utilizing total organic bromine (TOBr) to characterize the overall Br-DBPs and assess the combined toxicity of water samples offers a more comprehensive understanding of Br-DBPs generation and associated toxicity risks (Zhu and Zhang, 2016). Currently, the conventional method involves obtaining the organic extract through solid phase extraction and then diluting it to an appropriate concentration for toxicity testing (Yuan et al., 2023). However, previous studies have primarily focused on the formation of known Br-DBPs during specific disinfection processes and the subsequent prediction of their toxicity, while lacking a comprehensive understanding and evaluation of both the formation of TOBr and the combined toxicity of water samples derived from various disinfection processes.

In various disinfection processes, the formation mechanism of bromate and Br-DBPs exhibits certain similarities. The key intermediates in their formation are HOBr/OBr⁻, which also offer insights for the subsequent control of bromate and Br-DBPs (Von Gunten, 2003; Fang and Shang, 2012; Li et al., 2015). The control process for Br-DBPs primarily involves the addition of ammonia nitrogen (NH₃–N), hydrogen peroxide (H₂O₂), and ultraviolet (UV) irradiation (Boyer and Singer, 2005; Wu et al., 2020a; Wang et al., 2021a). The primary focus of these control process is on regulating specific Br-DBPs (such as bromate, trihalomethane, haloacetic acid), with a lack of comprehensive consideration for all Br-DBPs. Furthermore, while regulating the production of the target DBPs, there may be an increase in the concentration of other DBPs. For instance, during the ozonation process, NH₃–N can impede bromate reduction, leading to a higher concentration of brominated nitrogen-containing byproducts (Lin et al., 2014). Therefore, it is not reasonable to assess the applicability of the process based solely on a single DBP. Biological toxicity provides a more comprehensive assessment of the overall toxic effects of DBPs in water. Furthermore, various toxicity endpoints can reflect different associated risks (Gonçalves et al., 2020). Currently, the impact of Br-DBPs on the overall biotoxicity of water during disinfection has yet to be systematically elucidated.

In this review the global distribution of bromide is summarized, the formation mechanisms of bromate and Br-DBPs during different disinfection processes are concluded, and the toxicity of common Br-DBPs is compared under multiple toxicity endpoints. In addition, different control processes for Br-DBP are evaluated in relation to the comprehensive biological toxicity of water samples. Moreover, the prospects for the control of Br-DBPs are presented.

2 Distribution of bromide concentration

The formation of bromate and Br-DBPs is attributed to the presence of bromide in water/wastewater. To enhance our understanding of the global distribution of bromide concentrations, we have compiled and summarized the available data on reported bromide levels. As shown in Fig.1, data on bromide concentrations in groundwater, surface water, and wastewater from diverse regions have been collected over the past four decades. It is important to highlight that we have not included data on the concentration of bromide in seawater in this paper. This is due to the significantly higher concentration of bromide in seawater compared to fresh water. Research has indicated that the concentration of bromide in fresh water typically ranges from 10 to 1000 μg/L, whereas the concentration of bromide in seawater is approximately 67 mg/L (Alomirah et al., 2020). Since the focus of disinfection processes is generally on fresh water, it would be irrelevant to discuss marine bromide concentrations here.

Fig.1 Concentration distribution of bromide: (a) Bromide concentration (μg/L) in groundwater on various continents; (b) Bromide concentration (μg/L) in surface water on various continents; (c) Bromide concentration (μg/L) in wastewater on various continents; (d) Comparison of bromide concentration (μg/L) in different water sources in the same city of China. Data collected from (Luong et al., 1982; Cavanagh et al., 1992; Magazinovic et al., 2004; Selcuk et al., 2005; Yang et al., 2005; Huang et al., 2008; Wert et al., 2009; Liu et al., 2011b; Xie et al., 2012; Yang et al., 2012; Bo et al., 2013; Hua and Reckhow, 2013; Wang et al., 2014; Winid, 2015; Qi et al., 2016; Soltermann et al., 2016; Chuang and Mitch, 2017; Hong et al., 2017; Li et al., 2018; Fang et al., 2019; Jiang et al., 2019; Zhong et al., 2019; Gao et al., 2020; Wu et al., 2020a; Hua et al., 2021; Verwold et al., 2021; Yang et al., 2021; Falås et al., 2022; Du et al., 2023; He et al., 2023; Huang et al., 2023; Koo et al., 2023; Wu et al., 2023).

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2.1 Variations in bromide concentration due to geographical location

The concentration of bromide in water can be influenced by geographical location and various other factors. It was observed that coastal areas exhibit higher concentration of bromide in groundwater, surface water, and wastewater compared to inland regions. This phenomenon is commonly attributed to the intrusion of seawater into coastal groundwater and surface water, resulting in elevated levels of bromide (Von Gunten and Hoigne, 1994; Werner et al., 2013; Liu and Zhang, 2014; Kolb et al., 2017; Nguyen et al., 2021). The increase in sea level attributed to global warming, in conjunction with groundwater extraction, changes in land use, and other contributing factors, has led to the intensified and widespread observation of seawater intrusion (Werner et al., 2013; Ged and Boyer, 2014; Bourjila et al., 2023; Shao and Xu, 2023). And even minimal levels of seawater intrusion can have a profound adverse impact on groundwater resources by significantly elevating halide concentrations and causing pollution (Werner et al., 2013; Ged and Boyer, 2014). The impact of coastal geography on wastewater may be attributed to the reliance of coastal cities on seawater resources. In certain coastal regions, seawater is utilized for toilet flushing as a measure to mitigate the scarcity of freshwater (Yang and Zhang, 2013). Municipal wastewater treatment plants receive wastewater with elevated concentration of bromide. Consequently, the concentration of bromide detected in wastewater in coastal areas is significantly higher.

Nevertheless, certain data points exhibit deviations from the principles previously outlined. These discrepancies may be influenced by specific geographical conditions, local industries, and various environmental factors. For instance, the Barossa region in South Australia is not situated along the coast, yet its surface water contains a high concentration of bromide (Magazinovic et al., 2004). Due to its proximity to the coastline, the ocean still exerts an influence on local surface water. In North Carolina, a coastal region, bromide concentration can be as low as 30 µg/L. This is attributed to the use of a lake within the city for testing purposes, which is less affected by seawater intrusion (Cavanagh et al., 1992). And it is evident that the concentration of bromide in urban wastewater in Calgary, Canada reaches a staggering level of 1700 µg/L, potentially influenced by local industries and other contextual factors (Verwold et al., 2021).

Additionally, it is noteworthy that the variation in evapotranspiration rates resulting from climatic conditions will also impact the concentration of bromide in water. The concentration of bromide in the groundwater and surface water of the Tarim Basin and Datong Basin in China is notably high, exceeding even the levels found in coastal areas. This disparity can be attributed to elevated evapotranspiration rates. The annual evapotranspiration rate for the Datong Basin stands at approximately 2000 mm, while that of the Tarim Basin soars to a staggering 2750 mm; however, China's average annual evapotranspiration rate merely amounts to 359.1 mm (Xie et al., 2012; Werner et al., 2013; Cheng et al., 2021; Nguyen et al., 2021; Huo et al., 2022).

2.2 Differences in bromide concentration in water across different continents

As shown in Fig.1(a)–Fig.1(c), data on bromide concentrations in groundwater, surface water, and wastewater are collected for each continent. It is worth noting that the majority of the reported data pertained to North America, Europe, Asia, and Australia and there is a scarcity of studies on bromide concentrations in Africa and South America, necessitating further research. The concentration of bromide in groundwater exhibits minimal variation across different continents. However, the concentration of bromide in Australian surface water exceeds that found in surface waters collected from other continents. Additionally, European wastewater appears to contain more bromide. Currently, there is limited research on the variations in bromide concentrations in water across different continents. Nevertheless, based on the aforementioned data, it is hypothesized that discrepancies in bromide concentration among continents do exist and may be attributed to geographical location, climatic conditions and other factors that warrant further investigation.

2.3 Disparities in bromide concentration within various water

To further investigate potential difference in bromide concentrations among groundwater, surface water and wastewater, this study compares their levels within the same city (Fig.1(d)). In the data we collected, the concentration of bromide in surface water is observed to be higher than that in groundwater. Conversely, a study conducted in the USA has concluded that the concentration of bromide in groundwater exceeds that in surface water (Weisman et al., 2023). In fact, the concentration of bromide in surface and groundwater is influenced by a variety of factors. It has been observed that the concentration of bromide in surface and groundwater is dependent on their source, whether natural or anthropogenic (Gregov et al., 2022). An experiment conducted in Canada suggests that a significant portion of bromide is present on the soil surface rather than deep down (Parsons et al., 2004). Therefore, the existing research findings regarding the disparity in bromide concentrations between surface water and groundwater remain inconclusive. In the data collected in this review, there are also instances where the concentration of bromide in groundwater exceeds that found in surface water. In addition, groundwater and surface water can be mutually replenished (Wu et al., 2020b). Therefore, absolute conclusions are difficult to form, and need to be further explored. Furthermore, the evidence collected in this review and other studies do not reveal any significant difference in the concentration of bromide between wastewater and other water sources, which may be due to the greater influence of human factors in wastewater.

The increase in bromide concentration due to seawater intrusion is well established, and the impact of climatic conditions on bromide has been thoroughly investigated. In regions with high levels of bromide content, greater attention should be given to the generation and control of DBPs containing bromine during disinfection processes. Furthermore, further research is needed to explore the variation in bromide concentration across different continents as well as the differences of bromide concentration in different water sources.

3 Mechanisms of bromate and Br-DBPs formation during disinfection processes

Chlorine-based disinfection, ozonation, and persulfate-based disinfection are common disinfection processes. Understanding the mechanism of Br-DBPs generation during these processes is a prerequisite for achieving effective control over Br-DBPs.

3.1 Chlorine-based disinfection processes

The chlorine-based disinfection process primarily involves various methods of disinfection using chlorine, chloramine, and UV/chlorine as the disinfectants. Fig.2(a) illustrates the mechanism of bromate and Br-DBPs formation in chlorine-based disinfection processes. The reaction equations depicted in Table S2 are as follows. Chlorine or chloramine disinfection is completely a chemical disinfection method, in which bromide is transformed under the action of disinfectants (Zhu and Zhang, 2016). UV/ chlorine is a combination of physical and chemical methods in which free radicals like hydroxyl radicals (OH·) and chlorine radicals (Cl·) are generated through UV and play an important role in the formation of Br-DBPs (Fang et al., 2017). Therefore, this review categorizes the pathways of Br-DBPs formation during chlorine-based disinfection into two pathways: those involving radical participation and those without radical participation.

Fig.2 Formation mechanism of brominated DBPs in (a) Chlorine-based disinfection process; (b) Ozonation; (c) Persulfate-based disinfection process (Legube et al., 2004; Fang and Shang, 2012; Langsa et al., 2017; Yang et al., 2019; Wu et al., 2020a).

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In the pathway without radical, HOCl and NH₂Cl exhibit the capability to oxidize bromide into HOBr/BrO⁻, which serves as the crucial intermediate in the generation of Br-DBPs (Table S2 R1–R3) (Von Gunten, 2003; Fang and Shang, 2012; Li et al., 2015). The reaction between HOBr/BrO⁻ and organic matter in water leads to the formation of the Br-DBPs (Table S2 R13). But the oxidation of bromide to bromate cannot be achieved solely by the action of HOCl or NH₂Cl (Yang et al., 2019). Therefore, the generation of bromate seems to impose greater demands regarding chemical disinfectants when compared to Br-DBP.

It is important to note that the approach to chloramine disinfection discussed here is simplified. Chloramines are formed through a substitution reaction between free chlorine and ammonia, resulting in three forms: monochloramine (NH₂Cl), dichloramine (NHCl₂), and trichloramine (NCl₃). Among these, NH₂Cl serves as the primary disinfectant in drinking water treatment (Feng et al., 2024). Therefore, this study focuses on the generation of Br-DBPs during chloramine disinfection using NH₂Cl as the research object. When bromide come into contact with NH₂Cl, monobromamine (NH₂Br), dibromamine (NHBr₂), and HOBr/OBr⁻ are formed. Theoretically, all three substances mentioned above can react with dissolved organic matter to produce Br-DBPs. However, it is important to note that NH₂Br and NHBr₂ have a slow reaction rate with dissolved organic matter. Therefore, only the pathway of HOBr/OBr⁻ reacting with DOM to generate Br-DBPs is depicted in the figure (Zhu and Zhang, 2016).

In the pathway with radical (UV/chlorine), HOCl, OH· and Cl· are present at the same time. In the presence of OH·, the bromide can undergo conversion to the bromine radical (Br·), which subsequently transforms into BrO· over time (R8–R10 in Table S2 ). Besides, HOBr/BrO⁻ can react with OH· and Cl· to generate BrO· (R5–R7 in Table S2 ). The produced BrO· rapidly self-converts into BrO₂⁻ and BrO⁻ (R11 in Table S2 ), the former being able to form bromate in the presence of OH· ( R12 in Table S2). The resulting BrO⁻ can undergo reactions with organic matter to generate Br-DBPs, or it can further undergo reactions with HOCl to produce BrO₂⁻ which can eventually be converted to bromate in the presence of OH· (R4, R12, R13 in Table S2 ). The data indicates that the reaction of bromate formation from BrO· is fast, with a rate constant reaching 10⁹ L/(mol·s). In contrast, the generation of Br-DBPs from HOBr exhibits a lack of uniform reaction rate due to the complexity of DOM. But studies have suggested that its value falls within the range of 500–5000 L/(mol·s) (Westerhoff et al., 2004), which is significantly lower than the formation rate of bromate. However, given the potential toxicity of Br-DBPs, it is crucial to give equal attention to controlling both Br-DBPs and bromate in processes where a combination of ultraviolet light and chlorine is widely used.

The combination of UV and chlorine is widely utilized and exhibits great versatility (Moore et al., 2023; Wang et al., 2023). Therefore, both bromate and Br-DBPs need to be given full attention in chlorine-based disinfection process.

3.2 Ozonation

The ozonation disinfection process is depicted in Fig.2(b), illustrating the conversion of bromide into bromate and Br-DBPs. The reaction equations depicted in Table S3 are as follows. In addition to direct oxidation by ozone (O₃), O₃ can also undergo self-decomposition, leading to the generation of OH· (Wu et al., 2021b).

During the process of ozonation, the formation of bromate is typically categorized into three pathways: “direct”, “direct-indirect”, and “indirect-direct” (Pearce et al., 2022). In the direct pathway, bromide can be directly oxidized by O₃ to form HOBr/OBr⁻, which further converts to bromate (R1–R3 in Table S3 ). In the direct-indirect pathway, the HOBr/OBr⁻ produced by O₃ oxidation reacts with OH· to generate BrO·, which then undergoes disproportionation to form BrO₂⁻ and ultimately gets oxidized to bromate (R3, R9–R11 in Table S3 ) (Von Gunten, 2003; Qi et al., 2016). In the indirect-direct pathway, bromide is directly oxidized by OH· to produce BrO·, followed by disproportionation reaction, oxidation reaction and eventual formation of bromate (R4–R6, R8, R11 in Table S3 ). Furthermore, during the bromate formation process, HOBr/OBr⁻ generated through any of these pathways will react with organic matter in water, leading to the production of Br-DBPs (R12 in Table S3 ).

3.3 Persulfate-based disinfection process

The conversion of the bromide into Br-DBPs and bromate during persulfate-based disinfection process is depicted in Fig.2(c). The reaction equations depicted in Table S4 are as follows. Persulfate mainly includes PMS and PDS (Zhang et al., 2014). When in a non-activated state, they exhibit lower reactivity and are thus frequently activated for utilization (Ding et al., 2021). The persulfate-based disinfection process can be primarily categorized into two approaches: one relying solely on disinfectants and the other involving the activation of free radicals. It should be noted that PMS and PDS exhibit limited reactivity with halides in their non-activated state (Ding et al., 2021), so the bromide can only undergoes oxidation to form HOBr/BrO⁻ (R1, R2 in Table S4 ), without further oxidation to form bromate. During this period, the reaction between HOBr/BrO⁻ and DOM leads to the formation of Br-DBPs (R13 in Table S4 ).

In the activation process of persulfate, SO₄·⁻ can be produced on the PMS and PDS as the substrate by heat, UV, alkali, transition metal, etc (Fang and Shang, 2012). While OH· can only be generated when PMS is employed as the substrate. The disparity in electron transfer pathways between PMS and PDS is the primary determinant of this phenomenon (Lee et al., 2020). The mechanism by which bromide undergoes oxidation by OH· during persulfate process exhibits similarities to the oxidative process involving bromine and OH· in other oxidation reactions (R4–R6, R11 in Table S4 ). The SO₄·⁻ can also catalyze the oxidation of the bromide, leading to the gradual formation of BrO·, HOBr/BrO⁻ and bromate (R3, R12 in Table S4 ).

4 Factors affecting the formation of bromate and Br-DBPs

The generation of bromate and Br-DBPs during the disinfection process is influenced by a multitude of factors. Recent studies have identified several key factors, including the dosage and type of disinfectants used, as well as the concentration of bromide present (Yang et al., 2019; Wang et al., 2020b). To comprehensively investigate the impact of bromide and disinfectant dosage on the generation of Br-DBPs, this review consolidates findings related to the generation of bromate, trihalomethanes (THMs), haloacetic acids (HAAs) and TOBr in chlorine-based disinfection, ozonation, and persulfate-based disinfection.

4.1 Influence of bromide and ozone dose on bromate production in ozonation

The generation of bromate is primarily considered in the context of ozonation. The concentration of bromate produced under different Br⁻ and O₃ dosage is logarithmic to obtain Fig.3(a). It is evident that an increase in bromide concentration, along with a higher dosage of O₃, can significantly enhance the formation of bromate. Furthermore, it is also observed that at low bromide concentration (O₃ dosage), an increase in O₃ dosage (bromide concentration) does not result in an increase in bromate production.

Fig.3 Effects of concentration of Br− and dosage of disinfectants on the generation of brominated DBPs: (a) Bromate; (b) Trihalomethane (THMs); (c) Haloacetic acid (HAAs); (d) TOBr (Haag and Hoigne, 1983; Song et al., 1997; Chang et al., 2001; Zhang et al., 2005; Hua and Reckhow, 2006; Hua and Reckhow, 2007; Wert et al., 2007; Kristiana et al., 2009; Liu et al., 2011a; Liu et al., 2011b; Tian et al., 2013; Bond et al., 2014; Lin et al., 2014; Langsa et al., 2017; Jiang et al., 2018; Li et al., 2018; Liu et al., 2018; Mao et al., 2018; Fang et al., 2019; Gu et al., 2020; Wu et al., 2021a; Yu et al., 2021; Huang et al., 2022; Liu et al., 2022; Pearce et al., 2022; Wu et al., 2022; Zhao et al., 2022; Du et al., 2023).

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It has been shown that there is a strong correlation between the concentration of bromate and the initial concentration of Br⁻ (r² = 0.992) (Moslemi et al., 2011). When the concentration of bromide is low, the formation of bromate is influenced by the concentration of bromide (Wen et al., 2018). Consequently, an increase in oxidant concentration over time, under conditions of lower bromide concentration, does not effectively promote the generation of bromate. The ozonation process has the ability to decompose organic matter, and research indicates that the reaction rate between O₃ and organic compounds is significantly faster than that of O₃ with Br⁻. This leads to a predominant interaction between O₃ and organic matter (Lin et al., 2014; Zhang et al., 2023). Consequently, when the O₃ concentration is low, only a negligible amount of O₃ will react with Br⁻ to form bromate and Br⁻ is always in excess. Therefore, at low O₃ concentration, variations in Br⁻ concentration have minimal impact on bromate formation.

4.2 Influence of bromide and disinfectants dose on brominated THMs and HAAs

The impact of bromide and disinfectant dose on the generation of brominated byproducts during ozonation and chlorine-based disinfection was investigated (Fig.3(b) and Fig.3(c)), with brominated THMs and HAAs employed as representative components of known Br-DBPs.

However, based on the limited data collected, it remains unclear how the concentration of Br⁻ or disinfectants affects the production of THMs and HAAs. The formation of these compounds is influenced by a multitude of complex factors. In particular, within the chlorine-based disinfection process, there exists a competitive interaction between chlorine and bromine (Liu et al., 2018). This observation further confirms that the formation of Br-DBPs is influenced by multiple factors.

4.3 Influence of bromide and disinfectants dose on TOBr formation

Total organic halogens (TOX), encompassing total organic chlorine, total organic bromine, and total organic iodine compounds, can serve as a reliable indicator to evaluate the overall toxicity of water samples (Richardson et al., 2007; Kristiana et al., 2009). Currently available research has only uncovered a limited number of known halogenated organic DBPs. And their contribution to TOX is relatively low. It has been demonstrated that recognized DBPs contribute merely 8.3% to the TOX content (Richardson et al., 2002; Plewa et al., 2004). Given the significant carcinogenic potential of TOBr, they warrant substantial attention. The production data of TOBr under various disinfection methods, different concentrations of Br⁻, and varying amounts of disinfectants are summarized in Fig.3(d).

Increasing the concentration of Br⁻ or the dosage of disinfectants can lead to the generation of TOBr by augmenting the HOBr/OBr⁻ content in water (Song et al., 1997; Lin et al., 2014). For the ozonation, the production of TOBr significantly increased with the elevation of Br⁻ concentration or O₃ dosage. However, the impact of O₃ concentration on TOBr is not directly proportional under identical Br⁻ concentration. When the Br⁻ concentration is approximately 1 mg/L, the O₃ dose increases from 41.67 to 104 µmol/L, while the TOBr production shows an increase from 210 to 260 nmol/L, which is not statistically significant. However, when the O₃ dosage further escalates from 104 to 208 µmol/L, there is a substantial rise in TOBr production from 260 to 650 nmol/L, representing an approximate two and a half-fold increase. It is speculated that when the O₃ concentration is not high enough, the increase of O₃ is mainly used for the generation of bromate or the reaction with organic matter (Soltermann et al., 2017).

For the chlorination, the production of TOBr differs significantly between bromide concentration below and above 0.1 mg/L. However, the effect of bromide on TOBr production is found to be minimal when comparing TOBr generated at concentrations below or above this threshold. Additionally, the increase in disinfectant dosage has little impact on promoting TOBr production especially at higher concentrations of bromide. When the concentration of bromide reaches values such as 58 mg/L, and chlorine disinfectant amounts to 28.17, 70.42, and 140.85 µmol/L, respective yields of TOBr amount to 969, 1244, and 1913 nmol/L incrementally, and these increases are not substantial. The increase in TOBr production exhibits a significantly smaller magnitude compared to the corresponding increase in chlorine dosage. The lack of significant changes in TOBr production may be attributed to the relatively constant initial concentration of bromide. (Liu et al., 2018) research shown, an increase in chlorine facilitated the formation of total organic chlorine at a consistent initial bromide concentration, while the level of TOBr remained relatively stable.

Both HOCl and HOBr can react with organic matter, and there exists a competitive relationship between them (Yuan et al., 2023). Chlorine acts not only as an oxidant that promotes the formation of Br-DBPs but also as a competitor that suppresses their formation. The chlorine/bromide ratio determines the extent of Br-DBPs generation. Only when the chlorine concentration is relatively low and the bromide concentration is relatively high does HOBr become an active species, resulting in substantial production of Br-DBPs (Magazinovic et al., 2004). This may explain the previously observed separate changes in the dosage of Br⁻ or chlorine that are not proportional to changes in the production of TOBr.

It is worth noting that chlorine appears to promote TOBr production more than O₃. When the Br⁻ concentration ranges from 0.1 to 2 mg/L and the chlorine disinfectant dosage is approximately 77 μmol/L, the yield of TOBr ranges from 1611 to 3478 nmol/L. In contrast, with an O₃ content of about 208 µmol/L, the TOBr yield varies between 28 and 2828 nmol/L. Notably, while the O₃ dosage is roughly 2.7 times that of chlorine, the production of TOBr under ozonation is significantly lower compared to that under chlorination. This may be attributed to the higher oxidizing potential of O₃ compared to HOCl, leading to its more rapid consumption by organic matter and consequently reduced production of HOBr (Xue et al., 2024). Furthermore, the potent oxidative properties of O₃ enables direct oxidation of HOBr to bromate, resulting in decreased Br-DBPs formation during the ozonation (Pearce et al., 2022). Additionally, in the chlorination, substitution reactions convert chlorine disinfection byproducts into Br-DBPs, significantly increasing the quantity of Br-DBPs in this process.

4.4 Other factors affecting bromate and Br-DBPs formation

The formation of Br-DBPs and bromate is also influenced by other factors such as organic matter, pH, and temperature (Shah et al., 2015; Chen et al., 2018; Li et al., 2024b). The organic matter is the primary precursors that reacts with disinfectants to form DBPs. The water contains a substantial amount of organic matter, and its composition is complex (Jing et al., 2023). Studies have indicated that hydrophobic components in organic compounds are more prone to the formation of DBPs compared to hydrophilic components. Furthermore, bromides have a tendency to incorporate low molecular weight organics (Kitis et al., 2002; Farré et al., 2013). Therefore, variations in organic matter composition have a certain impact on DBPs generation.

Temperature exerts a significant influence on the rate constant, product yield, and substance distribution in chemical reactions (Li et al., 2019). Hua and Reckhow (2008) observed in their study that during chlorination, the concentration of TOX increased by 15% and 29%, respectively, as the temperature rose from 5 °C to 15 °C and 30 °C. When examining the impact of temperature on individual DBPs, it was found that elevating the temperature from 5 to 30 °C resulted in a maximum increase of THMs production by up to 100%, accompanied by a corresponding rise of 73% for dihaloacetic acids (DHAAs), while trichloroacetic acids (THAAs) only experienced a modest increment of merely 16%. Thus, the order in which temperature affects these three specific DBPs is as follows: THMs > DHAAs > THAAs. Furthermore, other research has indicated that the production of THMs is more influenced by temperature compared to the production of HAAs (Kavanaugh et al., 1980; Mcclellan, 2000). It can be inferred that the influence of temperature on DBPs formation is highly intricate.

The impact of pH value on the formation of bromate and Br-DBPs has been substantiated by numerous studies (Krasner et al., 1993; Cai et al., 2022). For most disinfection processes, such as ozonation, chlorine-based disinfection, and persulfate-based disinfection, pH plays a significant role in the generation of Br-DBPs. The pH level can impact the existence forms of disinfectants, the charge and reactivity of organic precursors, as well as the formation mechanism of individual DBPs (Liu et al., 2011b; Li et al., 2020; Siddique et al., 2023). The experimental findings by Wen et al. demonstrate a significant increase in bromate generation during ozonation or O₃/PMS disinfection when the pH is raised from 4 to 10 (Wen et al., 2018). During the process of chlorination, the formation of bromate and Br-DBPs is also influenced by changes in pH. This is primarily due to the fact that the conversion between HOBr and BrO⁻ is also pH-dependent (Westerhoff et al., 2004). And the specific peak point varies according to the actual circumstances (Liu and Croué, 2016; Qadafi et al., 2023).

5 Toxicity of Br-DBPs

Since the identification of DBPs, over 800 DBPs have been recognized, with more than 100 of these undergoing systematic and quantitative comparative toxicological analysis (Plewa and Wagner, 2015; Wu et al., 2022). Numerous studies have demonstrated that these DBPs exhibit cytotoxic, genotoxic, carcinogenic, teratogenic effects (Richardson et al., 2007). It is concerning that the majority of Br-DBPs exhibit relatively high levels of toxicity.

Due to the fact that many substances in DBPs are considered carcinogenic, mutagenic, and reproductive poisons, the potential threat posed by DBPs to human health has been under investigation since the last century (Mian et al., 2018). It has been confirmed that DBPs have an impact on human liver function and nervous system (Huang et al., 2021). Different types of DBPs have varying effects on health (Chaves et al., 2019). THMs and HAAs, which are the main DBPs, have been formally linked to bladder, rectal, and colon cancer (Kumari and Gupta, 2022). Halogenated aldehydes may lead to chromosomal mutations and DNA damage (Chaves et al., 2019). Some studies have also mentioned a correlation between DBPs and cardiovascular defects as well as abortion (Wright et al., 2017). As for Br-DBPs - the focus of this paper - although bromine itself does not directly impact human health, its associated DBPs have officially been recognized as having higher toxicity than common chlorine-containing DBPs, posing a greater threat to human health (Sharma et al., 2023).

The assessment of toxicity is essential for investigating the potential toxicity and adverse effects of DBPs. Therefore, it is imperative to consolidate the toxicity profiles of commonly encountered Br-DBPs. The cytotoxicity, genotoxicity, and developmental toxicity of specific Br-DBPs are summarized in this paper. The initial two categories pertain to in vitro toxicity, with cells serving as the primary experimental subjects. Relevant data on Chinese hamster ovary (CHO) cells and human hepatoma cells (HepG2) are collected here. The latter category pertains to in vivo toxicity, with organisms as the experimental subjects. Mainly, data on zebrafish, seaweed, and polychaete plants are collected here.

5.1 Cytotoxicity of Br-DBPs to mammalian cells

CHO cells, have been widely used in toxicology (Wagner and Plewa, 2017; Zhang et al., 2018; Qin et al., 2019; Lau et al., 2020). HepG2, a human hepatoma cell line with well-differentiated cell, exhibits functional characteristics similar to primary hepatocytes and is also a common subject for toxicity testing (Wu et al., 2021a). The cytotoxicity of common Br-DBPs to CHO and HepG2 was collected and shown in Fig.4(a).

Fig.4 Assessment of the toxicity of common Br-DBPs: (a) Cytotoxicity (LC50(μmol/L)) of Br-DBPs to CHO and HepG2 cells; (b) Genotoxicity (50% TDNA or midpoint of Tail moment (μmol/L)) to CHO cells; (c) Developmental toxicity (LC50(μmol/L)) to zebrafish embryos, marine algae and polychaete Platynereis dumerilii (Yang and Zhang, 2013; Liu and Zhang, 2014; Wagner and Plewa, 2017; Wu et al., 2021b).

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Common Br-DBPs can be cyclic or aliphatic. It is obvious that the cytotoxicity of aliphatic DBPs to HepG2 is much higher than that of cyclic DBPs. Wu et al. (2022) have indicated an approximate 8-fold increase in in vitro HepG2 cytotoxicity index for aliphatic DBPs when compared to cyclic DBPs. This disparity in toxicity mechanisms among different types of DBPs may elucidate these findings. Previous research has demonstrated that certain aliphatic DBPs such as monohaloacetic acid, iodoacetic acid, and bromoacetic acid act as potent inhibitors of glyceraldehyde-3-phosphate dehydrogenase (GAPDH), leading to a significant reduction in cellular ATP levels (Dad et al., 2018).

Br-DBPs are also categorized based on the presence of nitrogen (N). It has been observed that the toxicity of Br-DBPs containing N is significantly greater than that of those without N. Currently, nitrogen-containing disinfection byproducts (N-DBPs) have been studied relatively comprehensively, focusing on their high toxicity (Zhuang et al., 2023). N-DBPs found in drinking water primarily consist of nitrosamines, cyanogen halides (CNX), haloacetonitrile (HANs), haloacetamide (HAcAms), and halonitromethane (HNMs) (Shen, 2021). The N in N-DBPs may originate from nitrogen-containing precursor organics or from disinfectants (Ding and Chu, 2017). Numerous studies have shown that N-DBPs, especially brominated N-DBPs is more toxic than N-free disinfection byproducts (C-DBPs) (Bond et al., 2011; Qian et al., 2021; Zhuang et al., 2023). Currently, it is widely believed that N-DBPs have strong mutagenic properties, can induce sister chromatid exchange, DNA strand breaks, and then show high toxicity. It is important to note that nitrogen-rich colloids, hydrophilic neutral and hydrophilic alkaline components often play a dominant role in the formation of N-DBPs in nitrogen-containing organic precursors, which is significantly different from the most prevalent DBPs-trihalomethane. Therefore, conventional control strategies exhibit limited effectiveness on N-DBPs (Bond et al., 2011). At present, acid rain continues to introduce N into aquatic systems, and the issue of water eutrophication remains unresolved due to various factors (Hu et al., 2023; Xie et al., 2024). Furthermore, water utilities increasingly shift from chlorination to alternative disinfectants like chloramines, N-DBPs as a highly toxic DBPs, its production and control should receive continuous attention (Bond et al., 2011).

In addition, in aliphatic C-DBPs, it is observed that halogenated acetaldehydes are also more cytotoxic than the others. The toxicity of several halogenated acetaldehydes collected is similar to or even higher than that of N-DBPs. The halogenated acetaldehydes have been extensively discussed in previous studies due to its high toxicity. Its toxicity level is several orders of magnitude higher than that of THMs and HAAs (Chuang et al., 2015; Mao et al., 2016). According to the traditional mechanism, halogenated acetaldehyde is formed through the gradual alpha-hydrogen substitution of acetaldehyde. In other words, its formation is primarily associated with aldehydes (Li et al., 2024a).

5.2 Genotoxicity of Br-DBPs to mammalian cells

Considering the ability of numerous substances to induce DNA damage and potentially carcinogenic outcomes, studying genotoxicity has become widespread (Bianchi et al., 2017). Consequently, Fig.4(b) summarizes the genotoxicity data of common Br-DBPs to CHO.

The presence of N is observed to enhance toxicity, and the same trend applies to genotoxicity (Ding and Chu, 2017). The genotoxic 50% TDNA or midpoint of Tail moment values for most N-DBPs range from 100 to 1000 μmol/L, while the genotoxic 50% TDNA or midpoint of Tail moment values for most C-DBPs range from 1000 to 10000 μmol/L. Furthermore, halogenated acetaldehydes exhibit higher genotoxicity compared to other aliphatic DBPs such as trihalomethanes and haloacids (Xue et al., 2023). This observation aligns with the consistent pattern observed for both cytotoxicity and genotoxicity.

5.3 Developmental toxicity of Br-DBPs to aquatic organism

Taking zebrafish embryo as the test object, the toxicity of Br-DBPs in vivo can be studied. In addition, developmental toxicity data for Br-DBPs to marine algae and polychaete platynereis dumerilii are also collected (Fig.4(c)). The zebrafish embryo is widely recognized as the standard model for toxicological experimentation (Wang et al., 2017), and its data selection is universal. The reason for selecting marine organisms as the research subject here is that water containing DBPs is likely to be discharged into the sea, thus necessitating an assessment of the harm to marine organisms (Hsu and Singer, 2010).

The presence of N also increased developmental toxicity and the developmental toxicity of cyclic Br-DBPs and aliphatic Br-DBPs is still different. The developmental toxicity of cyclic Br-DBPs was higher than that of aliphatic Br-DBPs. However, this rule is contrary to the difference in cytotoxicity between cyclic and aliphatic Br-DBPs mentioned above. This indicates a significant disparity between in vivo toxicity and in vitro toxicity. Wu et al. (2022) conducted a comprehensive comparison between in vivo and in vitro toxicity of DBPs using toxicological study data and observed no correlation between in vivo and in vitro toxicity based on Pierre’s correlation coefficient.

In vivo toxicity studies have demonstrated that halogenated phenylquinones can induce the generation of reactive oxygen species (ROS) in zebrafish embryos and impede cellular antioxidant response, resulting in mortality, deformities, oxidative DNA damage, and apoptosis. The precise mechanism underlying the hypertoxicity induction of cyclic DBPs remains unclear. Nevertheless, it is plausible to speculate that the heightened toxicity of cyclic DBPs in vivo may be attributed to their interference with normal physiologic functions of cells and infliction of DNA damage (Wang et al., 2018).

5.4 Comprehensive toxicity changes during disinfection of water containing bromide

The impact of bromide presence or absence on cytotoxicity and genotoxicity of water samples during typical disinfection processes is explored to provide a more fully response to the toxicity risk posed by Br-DBPs. The toxicity ratio between disinfected water samples containing bromide and those without bromide is depicted in Fig.5(a) and Fig.5(b), respectively.

Fig.5 The impact of bromide concentration on the overall enhancement of toxicity in water: (a) Effect on cytotoxicity; (b) effect on genotoxicity (Toxicity increase refers to the ratio of toxicity after disinfection when bromide is present to when it is not present) (Wu et al., 2010; Chai et al., 2017; Wu et al., 2019; Wu et al., 2020b; Huang et al., 2022; Du et al., 2023).

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Overall, the toxicity of water samples containing bromide is found to be higher than that without bromide following disinfection with ozone, chlorine, chloramine, and PMS. Furthermore, within the same disinfection process, the cytotoxicity and genotoxicity of disinfected water samples increased proportionally with the concentration of bromide. These findings suggest that the formation of Br-DBPs is an important reason for the increased cytotoxicity and genotoxicity.

However, different disinfection processes exhibit varying sensitivities to bromide. First, the toxicity of water samples differs after different disinfection processes at the same bromide concentration. At a bromide concentration of about 2000 μg/L, the cytotoxicity rise to 1.60 times during chlorination, 2.70 times during UV/PMS disinfection, and 3.06 times during ozonation (compared to water samples without bromide). When the bromide concentration is about 1000 μg/L, the genotoxicity increased to 1.59 times during UV/PMS disinfection, 2.20 times during PMS disinfection, 3.49 times during chlorination, and 4.72 times during ozonation (Tables S5 and S6).

Secondly, the pattern of toxicity changes due to changes in bromide concentration after various disinfection processes is different. For example, it is evident that the continuous increase in bromide concentration has a significantly greater impact on toxicity during chloramine disinfection than during chlorination. In Fig.5(a), an increase in bromide concentration from 500 to 10000 μg/L resulted in a corresponding increase in cytotoxicity from 1.13 times to 2.12 times during chlorination, and from 1.43 to 3.64 times during chloramine disinfection, compared with water disinfection without bromide. Chloramine disinfection introduces N and forms N-DBPs which are more toxic than C-DBPs (Bond et al., 2011). The rise in bromide concentration during chloramine disinfection not only elevates the levels of Br-DBPs but also increases the overall quantity of highly toxic Br-N-DBPs, thus supporting the findings depicted in the Fig.5(a) (Cao et al., 2024).

Currently, there is a limited number of articles comparing the increased toxicity of bromide in water across various disinfection processes. However, the heightened toxicity resulting from Br-DBPs formed from bromide during ozonation has raised concerns (Wu et al., 2019; Wu et al., 2020a; Wu et al., 2021a). Careful consideration should be given to the application of chlorination, chloramination and ozonation in water with high bromide concentration.

6 Strategies for bromate and Br-DBP control

Common strategies for the control of bromate and Br-DBPs include the removal of bromide prior to disinfection processes, as well as the inhibition of bromate and Br-DBPs formation during disinfection (Yang et al., 2019).

6.1 Removal of bromide from water

Reducing the levels of bromide prior to disinfection is a critical strategy for controlling Br-DBPs and bromate formation (Liao et al., 2023). Currently, anion exchange represents a more innovative and advanced approach (Boyer and Singer, 2005). The predominant anion exchange membrane primarily employs traditional anion exchange resin or magnetic anion exchange resin.

The relationship between the ability of ion exchange resin to remove bromide and the amount of resin is summarized, and the results are presented in Fig.6. As expected, an increase in the amount of ion exchange resin leads to a higher removal rate of bromide from the raw water. The higher concentration of ion exchange resin enables a greater number of ion exchange sites, thereby enhancing the efficiency of bromide removal (Hsu and Singer, 2010). However, as the resin concentration continues to rise, the increasing trend of removal rate slows down. The same trend is observed in other studies, suggesting that the saturation of ion exchange sites may no longer be the determining factor for bromide removal as the amount of ion exchange resin increases in later stages (Johnson and Singer, 2004; Hsu and Singer, 2010).

Fig.6 Removal rate of bromide by ion exchange resins (Johnson and Singer, 2004; Boyer and Singer, 2005; Humbert et al., 2005; Hsu and Singer, 2010; Yin et al., 2020).

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Furthermore, a comparison is made between the bromide removal efficiency of traditional anion exchange resin and magnetic ion exchange resin. When it comes to bromide removal, magnetic ion exchange resins do not seem to possess an advantage over ordinary anion exchange resins. Data collected indicates that under identical amounts of resin used, the bromide rate by magnetic ion exchange resin consistently remains lower than that achieved by anion exchange resin. For instance, at a resin amount of 1 mL/L, the maximum bromide removal rate using magnetic ion exchange resin reaches only 41%, whereas with anion exchange resin it reaches 56.8%. Similarly, at a higher amount of 10 mL/L, respectively, the maximum bromide removal rate using magnetic ion exchange resin is 66.2% compared to 90.1% with anion exchange resin.

The primary distinction between magnetic ion exchange resin and traditional anion exchange resin lies in the incorporation of magnetic substances into the polymer matrix, resulting in a pore size for magnetic ion exchange resin that is 2–5 times smaller than that of conventional ion exchange resin (Singer and Bilyk, 2002). Previous study proved that magnetic ion exchange resins exhibit superior removal efficiency for dissolved organic carbon (DOC) compared to traditional anion exchange resins. However, during the removal process, DOC may displace some of the bromide previously removed by the resin, leading to their subsequent re-release (Hsu and Singer, 2010). This may be the reason that the removal rate of bromide by magnetic ion exchange resin is lower than that of traditional ion exchange resin. The majority of relevant articles primarily focus on the disparities between the two ion-exchange resins in terms of DOC removal, with less emphasis on comparing their effectiveness in bromide removal.

6.2 Inhibition of bromate and Br-DBP formation by masking HOBr

The inhibition of HOBr formation is a promising strategy for controlling the generation of bromate and Br-DBPs, as HOBr serves as an intermediate in their formation process. Currently, there exist two prevalent approaches for the mitigation of HOBr: the addition of ammonia nitrogen (NH₃–N) or hydrogen peroxide (H₂O₂).

6.2.1 Inhibiting bromate and Br-DBPs by NH₃–N/O₃ process

In the ozonation, the addition of NH₃–N leads to a reaction between ammonia (NH₃) and HOBr in water, resulting in the formation of NH₂Br, which subsequently converts to NHBr₂. This effectively reduces the concentration of HOBr during oxidation and thereby inhibits the production of bromated and Br-DBPs (Wu et al., 2020a; Wu et al., 2020b).

The changing trend of the removal rate of various DBPs (the percentage reduction in DBPs after adding NH₃–N compared to without) is depicted in Fig.7(a) as NH₃–N/O₃ (mass ratio) increases. For bromate, the inhibition rate gradually increases with increasing NH₃–N dosage. For TOBr, THMs and HAAs, NH₃–N showed overall inhibitory effect. However, the results indicate that for TOBr, THMs and HAAs, the addition of NH₃–N even enhances their formation in some cases. Mechanistically, this could be because both NH₂Br and NHBr₂ can react with O₃ to generate bromide while also directly reacting with organic matter to produce bromine-nitrogen-containing organic DBPs. The impact of NH₃–N on bromate and Br-DBPs is intricate, with numerous factors such as the quantity of NH₃–N influencing its functionality (Wu et al., 2021b).

Fig.7 Inhibition rate of bromate and Br-DBPs during ozonation by (a) NH3–N; (b) H2O2 (Kishimoto and Nakamura, 2012; Wang et al., 2013; Qi et al., 2016; Antoniou et al., 2017; Wu et al., 2020b; Wang et al., 2021a; Wu et al., 2022).

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The addition of NH₃–N inhibits the formation of bromine-containing DBPs. While its effect on bromate is evident, the mechanism influencing Br-DBPs is more complex. Thus, when utilizing NH₃–N for the control of bromate and Br-DBPs, a comprehensive consideration of these two types of DBPs is necessary.

6.2.2 Inhibiting bromate and Br-DBPs by H₂O₂/O₃ process

The efficacy of H₂O₂ addition in bromate control has been substantiated, and it has the advantages of simplicity, high efficiency and relatively low cost (Wen et al., 2017; Chen et al., 2018; Wang et al., 2021b). Compared to NH₃–N addition, the incorporation of H₂O₂ may possess the potential advantage of augmenting pollutant oxidation in source water through OH· generation (Wang et al., 2014). Fig.7(b) illustrates the removal rates of various DBPs as the H₂O₂/O₃ mass ratio increases (the percentage reduction in DBPs after adding H₂O₂ compared to without).

The inhibition rate of bromate exhibits a consistent upward trend with the progressive increase in H₂O₂/O₃. Particularly, when H₂O₂/O₃ ranges from 0.2 to 2.13 g/g, the addition of H₂O₂ significantly inhibits bromate formation. The inhibitory effect of a specific H₂O₂/O₃ ratio on bromate has been experimentally demonstrated (Wang et al., 2014). The mechanism underlying the impact of H₂O₂ addition during ozonation on bromate primarily involves the consumption of HOBr/OBr⁻ and the decomposition of O₃. Upon addition, H₂O₂ rapidly reacts with HOBr at a significantly high rate, leading to a substantial reduction in bromate formation. Furthermore, H₂O₂ accelerates both the decomposition of O₃ and the generation of free radicals, which also influences bromate formation (Yu et al., 2020).

It should be noted that there are instances in the Fig.7(b) where the inhibition rate of bromate is less than 0, indicating that the addition of H₂O₂ promotes its formation. Specifically, when H₂O₂/O₃ equals 0.16, it can increase bromate formation by up to 89.1%. Wu et al. (2021b) also concluded that low concentrations of H₂O₂ may promote bromate generation through the production of HO₂⁻ in water, which reacts with O₃ to generate OH· and ultimately enhances bromate formation. However, not all low ratio of H₂O₂/O₃ promote bromate formation, which discrepancy may be attributed to variations in background water composition, such as organic matter. Different raw water sources can lead to varying levels of H₂O₂/O₃ exposure and OH· formation (Yu et al., 2020), thereby resulting in a lack of complete consistency between the control effect of H₂O₂ on bromate and its overall behavior.

The addition of H₂O₂ appears to enhance the formation of THMs and HAAs. Consistent findings from other studies also support this conclusion, indicating that the introduction of H₂O₂ can promote THMs generation (Wu et al., 2021b). This phenomenon can be attributed to the promotion of O₃ decomposition by H₂O₂, leading to a significant increase in OH· concentration within water. However, OH· exhibits non-selective oxidation toward organic matter, thereby diminishing the removal efficiency of THMs and HAAs precursors while facilitating the formation of THMs and HAAs (Wang et al., 2013).

Although the addition of H₂O₂ facilitates the formation of THMs and HAAs, it significantly impedes the formation of TOBr. This is because H₂O₂ can convert HOBr, a key intermediate required for TOBr formation, into bromide (Wu et al., 2020a; Wu et al., 2020b). Generally speaking, it is undeniable that the addition of H₂O₂ exerts an inhibitory effect on bromate; however, its impact on Br-DBPs is complex.

6.2.3 Comparison of NH₃–N/O₃ and H₂O₂/O₃

NH₃–N and H₂O₂ are commonly used as bromate control measures. As mentioned earlier, these two methods have varying effects on the production of different DBPs. Since both NH₃–N/O₃ and H₂O₂/O₃ inhibit the formation of bromate, the toxicity of water under the two disinfection methods was integrated with the bromate inhibition rate data (Fig. S1) to conduct a comprehensive evaluation of the two disinfection methods.

The Fig. S1 clearly demonstrates that, in terms of both cytotoxicity and genotoxicity, a substantial number of data points indicate that while the addition of NH₃–N may suppress bromate production, it appears to enhance toxicity. It has been verified that the inclusion of NH₃–N inhibits the production of DBPs but concurrently promotes an escalation in water toxicity (Wu et al., 2020a; Wu et al., 2020b). The addition of NH₃–N resulted in a 1.89-fold increase in cytotoxicity and a 2.86-fold increase in genotoxicity. According to the mechanism of action following the addition of NH₃–N, the addition of NH₃ elevates the formation of highly toxic Br-N-DBPs formation, such as bromoacetamide, dibromoacetamide and dibromonitromethane (Lu et al., 2021).

For H₂O₂, most of the points in the Fig. S1 showed positive bromate inhibition rate and toxicity inhibition rate. In other words, under normal circumstances, H₂O₂ plays a certain role in both DBPs inhibition and toxicity inhibition. However, it is noteworthy that all data points in the Fig. S1 lie below the 1:1 line, indicating a lower inhibition rate of toxicity compared to bromate inhibition rate. Considering this observation, the influence of H₂O₂ addition on the variation trend of Br-DBPs can not be disregarded. Under specific conditions, H₂O₂ can even facilitate the formation of certain Br-DBPs, potentially accounting for the lower toxicity inhibition rate compared to bromate inhibition rate.

Considering the disparity in toxicity levels between NH₃–N and H₂O₂, it can be inferred that H₂O₂ is a more favorable alternative than NH₃–N due to its potential for increased toxicity. Furthermore, the data also emphasize the significant contribution of Br-DBPs to overall toxicity, necessitating further attention.

6.3 Effect of UV irradiation on bromate and TOBr formation

The combination of ultraviolet and other chemical disinfection technologies presents a novel approach for controlling Br-DBPs (Ao et al., 2020). The integration of ultraviolet irradiation with other oxidation-based disinfection technologies can effectively mitigate the formation of Br-DBPs through diverse mechanisms.

In Fig.8(a), the amount of TOBr significantly reduces after UV/O₃ process compared to ozonation at the same O₃ dosage. However, the addition of UV radiation increases the bromate production in water (Zhao et al., 2013). When O₃ is added at a concentration of 0.17 mmol/L, TOBr is generated within the range of 0.055 to 0.132 μmol/L upon exposure to ultraviolet irradiation, whereas in the absence of ultraviolet irradiation, TOBr levels significantly increase to 0.323 μmol/L or higher. However, with O₃ dosage of 0.17 mmol/L, bromate production ranges between 0.253 and 0.283 μmol/L without ultraviolet irradiation and between 0.504 and 0.558 μmol/L with ultraviolet irradiation applied concurrently. Ultraviolet radiation doubles the production of bromate.

Fig.8 Effect of UV on the formation of TOBr and BrO3− in the process of disinfection: (a) Ozonation; (b) persulfate-based disinfection (Fang and Shang, 2012; Zhao et al., 2013; Mao et al., 2018; Wu et al., 2019; Yang et al., 2019; Guan et al., 2020; Luo et al., 2020; Wu et al., 2020b; Wu et al., 2021a; Chen et al., 2023; Du et al., 2023).

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At present, relevant research has been carried out on this phenomenon, and a reasonable explanation has been obtained. UV irradiation can induce the photolysis of O₃ and HOBr, resulting in a significant decrease in HOBr production (Kim et al., 2007; Zhao et al., 2013). Furthermore, the exposure to UV radiation will enhance the generation of free radicals, thereby facilitating the aforementioned bromate formation process through free radical mechanisms.

In the context of persulfate-based disinfection process, when combined with UV treatment, its behavior aligns with that observed when UV is combined with O₃. Specifically, the addition of UV inhibits the formation of TOBr, but promotes the production of bromate (Fig.8(b)) (Du et al., 2023). This phenomenon can be attributed to the generation of free radicals in the oxidation system following UV irradiation. According to the mechanism of bromate formation during persulfate-based disinfection process described above, PMS/PDS lacks the capacity to oxidize bromide to bromate in the absence of activation (Fang et al., 2017). The introduction of UV generates free radicals, facilitating the formation of bromate. This may be attributed to UV irradiation opening up pathways for the conversion of bromide to bromate, thereby reducing the conversion of bromide to Br-DBPs.

While UV can inhibit the formation of TOBr, it may also promote the generation of bromate. Therefore, when conducting a comprehensive evaluation of this technology, it is crucial to consider the potential impact of ultraviolet application on toxicity.

7 Conclusions

Increasing attention is being devoted to the risks of Br-DBPs to the ecosystem and human health, leading to a growing number of Br-DBPs being regulated and an increasing focus on exploring alternative disinfection processes as well as strategies for controlling Br-DBPs. Based on the findings summarized in this paper, the following conclusions and perspectives are drawn to achieve the reduction and control of Br-DBPs to ensure water safety:

1) The global distribution of bromide exhibits a relatively consistent pattern across continents, with concentrations predominantly hovering around 100 μg/L. While the continent of Australia exhibits a slightly higher concentration of surface water compared to others, with the lowest collected concentration being 238 μg/L. Furthermore, within the same geographical area, the concentration of surface water appears to exceed that of groundwater. Besides, coastal and arid regions exhibit relatively elevated concentrations of bromide, necessitating critical attention toward the production of bromate and Br-DBPs during disinfection.

2) HOBr plays a pivotal role as an intermediate in the generation of bromate and Br-DBPs during disinfection processes, and bromide and disinfectant dose are positively correlated with the production of bromate and Br-DBPs. The use of anion exchange membranes to remove bromide or minimize disinfectant dosage while safeguarding the disinfection efficacy is necessary in some cases.

3) Currently, the control of Br-DBPs is limited to bromate, THMs and HAAs, but this paper concludes that aldehydes Br-DBPs and nitrogen-containing Br-DBPs have higher toxicity. While aromatic Br-DBPs demonstrate varying results in terms of cytotoxicity and developmental toxicity. Therefore, there is a need to improve the understanding of the toxicity of various types of Br-DBPs under multiple toxicity endpoints.

4) Current control processes for Br-DBPs face the challenge of not being able to simultaneously regulate both the formation of Br-DBPs and the toxicity. H₂O₂/O₃ has limited effect in controlling Br-DBPs. NH₃–N/O₃ is associated with an increased risk of toxicity, up to a 2.86-fold increase in genotoxicity. UV/O₃ and UV/PS promotes bromate production. Given the complexity of Br-DBPs, systems thinking should be established in developing water treatment processes that focus not only on reducing individual types of Br-DBPs, but also on ensuring comprehensive control over biological toxicity.

References

[1] Sadiq R, Rodriguez M J. Disinfection by-products (DBPs) in drinking water and predictive models for their occurrence: a review. Science of the Total Environment, 2004, 321(1–3): 21–46

[2] Xie Y F. Disinfection Byproducts in Drinking Water—Formation, Analysis, and Control. Washington, DC: Lewis Publishers, 2004

[3] Richardson S D, Plewa M J, Wagner E D, Schoeny R, Demarini D M. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection by-products in drinking water: a review and roadmap for research. Mutation Research, 2007, 636(1–3): 178–242

[4] Karanfil T, Krasner S W, Westerhoff P, Xie Y F. Recent advances in Disinfection by-product: Formation, occurrence, health effects, and regulations. In: Karanfil T, Krasner S W, Westerhoff P, Xie Y F, editors, Disinfection By-Products in Drinking Water – Occurrence, Formation, Health Effects, and Control. Washington, DC: American Chemical Society, 2008, 2–19

[5] Krasner S W. The formation and control of emerging disinfection by-products of health concern. Philosophical Transactions of the Royal Society A—Mathematical Physical and Engineering Sciences, 2009, 367(1904): 4077–4095

[6] Bond T, Templeton M R, Graham N. Precursors of nitrogenous disinfection by-products in drinking water—a critical review and analysis. Journal of Hazardous Materials, 2012, 235-236: 1–16

[7] Shah A D, Mitch W A. Halonitroalkanes, halonitriles, haloamides, and N-nitrosamines: a critical review of nitrogenous disinfection byproduct formation pathways. Environmental Science & Technology, 2012, 46(1): 119–131

[8] GB5749–2006. P.R. China Standards for Drinking Water Quality. Beijing: Department of Health, P.R. China, 2006 (in Chinese)

[9] Xie Y F. Disinfection by-product analysis in drinking water. American Laboratory, 2000, 32(22): 50–54

[10] GB/T 5750.10–2006. P.R. China Standards Examination Methods for Drinking Water – Disinfection By-products Parameters. Beijing: Department of Health, P.R. China, 2006 (in Chinese)

[11] Wang X M, Zhang X L, Yang H W, Zhang J S, Liu L J, Liu B, Xie Y F. Comparison of methods for chloral hydrate analysis in drinking water. Water & Wastewater Engineering, 2013, 39 (supplementary): 125–128 (in Chinese)

[12] Xie Y F, Hwang C J. Cyanogen chloride and cyanogen bromide analysis in drinking water. In: Meyers R A, ed. Encyclopedia of Analytical Chemistry. Hoboken: John Wiley & Sons, 2006

[13] Xie Y F, Reckhow D A. A rapid and simple analytical method for cyanogen chloride and cyanogen-bromide in drinking water. Water Research, 1993, 27(3): 507–511

[14] Richardson S D. Disinfection by-products and other emerging contaminants in drinking water. Trace-Trends in Analytical Chemistry, 2003, 22(10): 666–684

[15] Weinberg H S. Modern approaches to the analysis of disinfection by-products in drinking water. Philosophical Transactions of the Royal Society A-Mathematical Physical and Engineering Sciences, 2009, 367(1904): 4097–4118

[16] Loos R, Barcel� D. Determination of haloacetic acids in aqueous environments by solid-phase extraction followed by ion-pair liquid chromatography-electrospray ionization mass spectrometric detection. Journal of Chromatography. A, 2001, 938(1–2): 45–55

[17] Meng L, Wu S, Ma F, Jia A, Hu J. Trace determination of nine haloacetic acids in drinking water by liquid chromatography-electrospray tandem mass spectrometry. Journal of Chromatography. A, 2010, 1217(29): 4873–4876

[18] Richardson S D. The role of GC-MS and LC-MS in the discovery of drinking water disinfection by-products. Journal of Environmental Monitoring, 2002, 4(1): 1–9

[19] Richardson S D. Environmental mass spectrometry: emerging contaminants and current issues. Analytical Chemistry, 2008, 80(12): 4373–4402

[20] Zhang X, Minear R A, Guo Y, Hwang C J, Barrett S E, Ikeda K, Shimizu Y, Matsui S. An electrospray ionization-tandem mass spectrometry method for identifying chlorinated drinking water disinfection byproducts. Water Research, 2004, 38(18): 3920–3930

[21] Zhang X, Talley J W, Boggess B, Ding G, Birdsell D. Fast selective detection of polar brominated disinfection byproducts in drinking water using precursor ion scans. Environmental Science & Technology, 2008, 42(17): 6598–6603

[22] Zhang X, Minear R A, Barrett S E. Characterization of high molecular weight disinfection byproducts from chlorination of humic substances with/without coagulation pretreatment using UF-SEC-ESI-MS/MS. Environmental Science & Technology, 2005, 39(4): 963–972

[23] Xie Y, Zhou H J. Biologically active carbon for HAA removal: part II, column study. Journal—American Water Works Association, 2002, 94(5): 126–134

[24] Urbansky E T. Techniques and methods for the determination of haloacetic acids in potable water. Journal of Environmental Monitoring, 2000, 2(4): 285–291

[25] Kubáň P, Makarõtševa N, Kiplagat I K, Kaljurand M. Determination of five priority haloacetic acids by capillary electrophoresis with contactless conductivity detection and solid phase extraction preconcentration. Journal of Separation Science, 2012, 35(5–6): 666–673

[26] Zwierner C, Richardson S D. Analysis of disinfection by-products in drinking water by LC–MS and related MS techniques. Trace-Trends in Analytical Chemistry, 2005, 24(7): 613–621

[27] Zwiener C, Frimmel F H. LC-MS analysis in the aquatic environment and in water treatment technology—A critical review. Part II: Applications for emerging contaminants and related pollutants, microorganisms and humic acids. Analytical and Bioanalytical Chemistry, 2004, 378(4): 862–874

[28] Liu Y, Mou S. Determination of bromate and chlorinated haloacetic acids in bottled drinking water with chromatographic methods. Chemosphere, 2004, 55(9): 1253–1258

[29] Krasner S W, McGuire M J, Jacangelo J G, Patania N L, Reagan K M, Aieta E M. The occurrence of disinfection by-products in United-States drinking-water. Journal—American Water Works Association, 1989, 81(8): 41–53

[30] Krasner S W, Weinberg H S, Richardson S D, Pastor S J, Chinn R, Sclimenti M J, Onstad G D, Thruston A D Jr, Thruston A D. Occurrence of a new generation of disinfection byproducts. Environmental Science & Technology, 2006, 40(23): 7175–7185

[31] Pressman J G, Richardson S D, Speth T F, Miltner R J, Narotsky M G, Hunter E S 3rd, Rice G E, Teuschler L K, McDonald A, Parvez S, Krasner S W, Weinberg H S, McKague A B, Parrett C J, Bodin N, Chinn R, Lee C F, Simmons J E. Concentration, chlorination, and chemical analysis of drinking water for disinfection byproduct mixtures health effects research: U.S.EPA’s Four Lab Study. Environmental Science & Technology, 2010, 44(19): 7184–7192

[32] McGuire M J, McLain J L, Obolensky A. Information collection rule data analysis. Denver: American Water Works Association Research Foundation, 2002

[33] Obolensky A, Singer P C. Halogen substitution patterns among disinfection byproducts in the information collection rule database. Environmental Science & Technology, 2005, 39(8): 2719–2730

[34] Obolensky A, Singer P C, Shukairy H M. Information collection rule data evaluation and analysis to support impacts on disinfection by-product formation. Journal of Environmental Engineering, 2007, 39(8): 53–63

[35] Mitch W A, Krasner S W, Westerhoff P, Dotson A. Occurrence and formation of nitrogenous disinfection by-products. Denver: Water Research Foundation, 2009

[36] McGuire M J, Meadow R G. AWWARF trihalomethane survey. Journal—American Water Works Association, 1988, 80(1): 61–68

[37] Williams D T, LeBel G L, Benoit F M. Disinfection by-products in Canadian drinking water. Chemosphere, 1997, 34(2): 299–316

[38] Simpson K L, Hayes K P. Drinking water disinfection by-products: an Australian perspective. Water Research, 1998, 32(5): 1522–1528

[39] Nissinen T K, Miettinen I T, Martikainen P J, Vartiainen T. Disinfection by-products in Finnish drinking waters. Chemosphere, 2002, 48(1): 9–20

[40] Lee K J, Kim B H, Hong J E, Pyo H S, Park S J, Lee D W. A study on the distribution of chlorination by-products (CBPs) in treated water in Korea. Water Research, 2001, 35(12): 2861–2872

[41] Yoon J, Choi Y, Cho S, Lee D. Low trihalomethane formation in Korean drinking water. Science of the Total Environment, 2003, 302(1–3): 157–166

[42] Ding H, Meng L, Zhang H, Yu J, An W, Hu J, Yang M. Occurrence, profiling and prioritization of halogenated disinfection by-products in drinking water of China. Environmental Science—Processes & Impacts, 2013, 15(7): 1424–1429

[43] US Environmental Protection Agency. National primary drinking water regulations: disinfectants and disinfection byproducts. Federal Register, 1998, 63: 69390–69476

[44] US Environmental Protection Agency. National primary drinking water regulations: stage 2 disinfectants and disinfection byproducts rule. Federal Register, 2006, 71: 388–493

[45] CDW. Guidelines for Canadian Drinking Water Quality Summary Table. Committee on Drinking Water of the of the Federal-Provincial-Territorial Committee on Health and the Environment, Canada, 2012

[46] EU. Council directive 98/83/EC of 3 November 1998 on the quality of water intended for human consumption, 1998

[47] Golfinopoulos S K, Nikolaou A D. Survey of disinfection by-products in drinking water in Athens, Greece. Desalination, 2005, 176(1–3): 13–24

[48] WHO. Guidelines for Drinking-water Quality. 4th edition. World Health Organization, 2011

[49] MHLW. Ministry of Health, Labour and Welfare, Japan

[50] NHMRC, NRMMC. Australian Drinking Water Guidelines 6. Commonwealth of Australia, 2011

[51] Bond T, Huang J, Templeton M R, Graham N. Occurrence and control of nitrogenous disinfection by-products in drinking water—a review. Water Research, 2011, 45(15): 4341–4354

[52] Nawrocki J, Andrzejewski P. Nitrosamines and water. Journal of Hazardous Materials, 2011, 189(1–2): 1–18

[53] Gopal K, Tripathy S S, Bersillon J L, Dubey S P. Chlorination byproducts, their toxicodynamics and removal from drinking water. Journal of Hazardous Materials, 2007, 140(1–2): 1–6

[54] Watson K, Farr� M J, Knight N. Strategies for the removal of halides from drinking water sources, and their applicability in disinfection by-product minimisation: a critical review. Journal of Environmental Management, 2012, 110: 276–298

[55] Mitch W A, Sharp J O, Trussell R P, Valentine R L, Alvarez-Cohen L, Sedlak D L. N-nitrosodimethylamine (NDMA) as a drinking water contaminant: a review. Environmental Engineering Science, 2003, 20(5): 389–404

[56] Schreiber I M, Mitch W A. Influence of the order of reagent addition on NDMA formation during chloramination. Environmental Science & Technology, 2005, 39(10): 3811–3818

[57] Hua G H, Reckhow D A. DBP formation during chlorination and chloramination: Effect of reaction time, pH, dosage, and temperature. Journal—American Water Works Association, 2008, 100(8): 82–95

[58] Tung H H, Unz R F, Xie Y F. Haloacetic acid removal by granular activated carbon adsorption. Journal—American Water Works Association, 2006, 98(6): 107–112

[59] Xie Y F, Zhou H J, Romano J P. Development of a capillary electrophoresis method for haloacetic acids. Abstracts of Papers of the American Chemical Society, 1999, 217: 742–742

[60] Wu H W, Xie Y F. Effects of EBCT and water temperature on HAA removal using BAC. Journal—American Water Works Association, 2005, 97(11): 94–101

[61] Tung H H, Xie Y F. Association between haloacetic acid degradation and heterotrophic bacteria in water distribution systems. Water Research, 2009, 43(4): 971–978

[62] Rodriguez M J, S�rodes J B, Levallois P. Behavior of trihalomethanes and haloacetic acids in a drinking water distribution system. Water Research, 2004, 38(20): 4367–4382

[63] Boyden B H, Banh D T, Huckabay H K, Fernandes J B. Using inclined cascade aeration to strip chlorinated VOCs from drinking-water. Journal—American Water Works Association, 1992, 84(5): 62–69

[64] Velazquez C, Estevez L A. Stripping of trihalomethanes from drinking-water in a bubble-column aerator. American Institute of Chemical Engineers Journal, 1992, 38(2): 211–218

[65] Sherant S R. Trihalomethane Control by Aeration. Dissertation for the Master Degree. Pennsylvania: The Pennsylvania State University, 2008

[66] Sherant S R, Hardin Y D, Xie Y F. A simple technology for THM control in consecutive systems. In: Proceedings of the American Water Works Association Water Quality Technology Conference, 2007

[67] Brooke E, Collins M R. Post treatment aeration to reduce THMs. Journal—American Water Works Association, 2011, 103(10): 84–96

[68] Hozalski R M, Zhang L, Arnold W A. Reduction of haloacetic acids by Fe0: implications for treatment and fate. Environmental Science & Technology, 2001, 35(11): 2258–2263

[69] Zhang L, Arnold W A, Hozalski R M. Kinetics of haloacetic acid reactions with Fe(0). Environmental Science & Technology, 2004, 38(24): 6881–6889

[70] Pearson C R, Hozalski R M, Arnold W A. Degradation of chloropicrin in the presence of zero-valent iron. Environmental Toxicology and Chemistry, 2005, 24(12): 3037–3042

[71] Chen C, Wang X, Chang Y, Liu H. Dechlorination of disinfection by-product monochloroacetic acid in drinking water by nanoscale palladized iron bimetallic particle. Journal of Environmental Sciences (China), 2008, 20(8): 945–951

[72] Li T, Chen Y, Wan P, Fan M, Yang X J. Chemical degradation of drinking water disinfection byproducts by millimeter-sized particles of iron-silicon and magnesium-aluminum alloys. Journal of the American Chemical Society, 2010, 132(8): 2500–2501

[73] Wang X Y, Ning P, Liu H L, Ma J. Dechlorination of chloroacetic acids by Pd/Fe nanoparticles: Effect of dying method on metallic activity and the parameter optimization. Applied Catalysis B: Environmental, 2010, 94(1–2): 55–63

[74] Tang S, Wang X M, Yang H W, Xie Y F. Haloacetic acid removal by sequential zero-valent iron reduction and biologically active carbon degradation. Chemosphere, 2013, 90(4): 1563–1567

[75] Woo Y T, Lai D, McLain J L, Manibusan M K, Dellarco V. Use of mechanism-based structure-activity relationships analysis in carcinogenic potential ranking for drinking water disinfection by-products. Environmental Health Perspectives, 2002, 110(Suppl 1): 75–87

[76] Cancho B, Ventura F, Galceran M, Diaz A, Ricart S. Determination, synthesis and survey of iodinated trihalomethanes in water treatment processes. Water Research, 2000, 34(13): 3380–3390

[77] Chen P H, Richardson S D, Krasner S W, Majetich G, Glish G L. Hydrogen abstraction and decomposition of bromopicrin and other trihalogenated disinfection byproducts by GC/MS. Environmental Science & Technology, 2002, 36(15): 3362–3371

[78] Wang C K, Zhang X J, Wang J, Chen C. Detecting N-nitrosamines in water treatment plants and distribution systems in China using ultra-performance liquid chromatography-tadenm mass spectrometry. Frontiers of Environmental Science & Engineering, 2012, 6(6): 770–777

[79] Zhao Y Y, Boyd J, Hrudey S E, Li X F. Characterization of new nitrosamines in drinking water using liquid chromatography tandem mass spectrometry. Environmental Science & Technology, 2006, 40(24): 7636–7641

[80] Egorov A I, Tereschenko A A, Altshul L M, Vartiainen T, Samsonov D, LaBrecque B, M�ki-Paakkanen J, Drizhd N L, Ford T E. Exposures to drinking water chlorination by-products in a Russian city. International Journal of Hygiene and Environmental Health, 2003, 206(6): 539–551

[81] Boyd J M, Hrudey S E, Richardson S D, Li X F. Solid-phase extraction and high-performance liquid chromatography mass spectrometry analysis of nitrosamines in treated drinking water and wastewater. Trace-Trends in Analytical Chemistry, 2011, 30(9): 1410–1421

[82] Krasner S W, Mitch W A, McCurry D L, Hanigan D, Westerhoff P. Formation, precursors, control, and occurrence of nitrosamines in drinking water: a review. Water Research, 2013, 47(13): 4433–4450

[83] Zhao H, Liu H, Hu C, Qu J. Effect of aluminum speciation and structure characterization on preferential removal of disinfection byproduct precursors by aluminum hydroxide coagulation. Environmental Science & Technology, 2009, 43(13): 5067–5072

[84] Uyak V, Yavuz S, Toroz I, Ozaydin A, Genceli E A. Disinfection by-products precursors removal by enhanced coagulation and PAC adsorption. Desalination, 2007, 216(1–3): 334–344

[85] Hua G, Reckhow D A. Characterization of disinfection byproduct precursors based on hydrophobicity and molecular size. Environmental Science & Technology, 2007, 41(9): 3309–3315

[86] Liang L, Singer P C. Factors influencing the formation and relative distribution of haloacetic acids and trihalomethanes in drinking water. Environmental Science & Technology, 2003, 37(13): 2920–2928

[87] Ates N, Yilmaz L, Kitis M, Yetis U. Removal of disinfection by-product precursors by UF and NF membranes in low-SUVA waters. Journal of Membrane Science, 2009, 328(1–2): 104–112

[88] Chellam S. Effects of nanofiltration on trihalomethane and haloacetic acid precursor removal and speciation in waters containing low concentrations of bromide ion. Environmental Science & Technology, 2000, 34(9): 1813–1820

[89] Chellam S, Sharma R R, Shetty G R, Wei Y. Nanofiltration of pretreated Lake Houston water- Disinfection by-product speciation, relationships, and control. Separation and Purification Technology, 2008, 64(2): 160–169

[90] Harrison C J, Le Gouellec Y A, Cheng R C, Childress A E. Bench-scale testing of nanofiltration for seawater desalination. Journal of Environmental Engineering, 2007, 133(11): 1004–1014

[91] Wang C K, Zhang X J, Chen C, Wang J. Factors controlling N-nitrosodimethylamine (NDMA) formation from dissolved organic matter. Frontiers of Environmental Science & Engineering, 2013, 7(2): 151–157

[92] Park S H, Wei S, Mizaikoff B, Taylor A E, Favero C, Huang C H, Huang C H. Degradation of amine-based water treatment polymers during chloramination as N-nitrosodimethylamine (NDMA) precursors. Environmental Science & Technology, 2009, 43(5): 1360–1366

[93] Mitch W A, Sedlak D L. Formation of N-nitrosodimethylamine (NDMA) from dimethylamine during chlorination. Environmental Science & Technology, 2002, 36(4): 588–595

[94] Kohut K D, Andrews S A. Polyelectrolyte age and N-nitrosodimethylamine formation in drinking water treatment. Water Quality Research Journal of Canada, 2003, 38(4): 719–735

[95] Bichsel Y, von Gunten U. Formation of iodo-trihalomethanes during disinfection and oxidation of iodide containing waters. Environmental Science & Technology, 2000, 34(13): 2784–2791

[96] Richardson S D, Fasano F, Ellington J J, Crumley F G, Buettner K M, Evans J J, Blount B C, Silva L K, Waite T J, Luther G W, Mckague A B, Miltner R J, Wagner E D, Plewa M J. Occurrence and mammalian cell toxicity of iodinated disinfection byproducts in drinking water. Environmental Science & Technology, 2008, 42(22): 8330–8338

[97] Chen Z, Valentine R L. The influence of the pre-oxidation of natural organic matter on the formation of N-nitrosodimethylamine (NDMA). Environmental Science & Technology, 2008, 42(14): 5062–5067

[98] Shah A D, Krasner S W, Lee C F T, von Gunten U, Mitch W A. Trade-offs in disinfection byproduct formation associated with precursor preoxidation for control of N-nitrosodimethylamine formation. Environmental Science & Technology, 2012, 46(9): 4809–4818

[99] Padhye L, Luzinova Y, Cho M, Mizaikoff B, Kim J H, Huang C H. PolyDADMAC and dimethylamine as precursors of N-nitrosodimethylamine during ozonation: reaction kinetics and mechanisms. Environmental Science & Technology, 2011, 45(10): 4353–4359

[100] Hoigne J, Bader H. The formation of trichloronitromethane (chloropicrin) and chloroform in a combined ozonation chlorination treatment of drinking-water. Water Research, 1988, 22(3): 313–319

[101] Bond T, Goslan E H, Parsons S A, Jefferson B. Treatment of disinfection by-product precursors. Environmental Technology, 2011, 32(1–2): 1–25

[102] Reckhow D A, Linden K G, Kim J, Shemer H, Makdissy G. Effect of UV treatment on DBP formation. Journal—American Water Works Association, 2010, 102(6): 100–113

[103] Shah A D, Dotson A D, Linden K G, Mitch W A. Impact of UV disinfection combined with chlorination/chloramination on the formation of halonitromethanes and haloacetonitriles in drinking water. Environmental Science & Technology, 2011, 45(8): 3657–3664

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