(Nbfrs) in Aquatic Environments —A Critical Review

Water Research 198 (2021) 117168

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Review

Occurrence, bioaccumulation, fate, and risk assessment of novel brominated ﬂame retardants (NBFRs) in aquatic environments —A critical review

∗ Rui Hou a, Lang Lin a, Hengxiang Li a, Shan Liu a, Xiangrong Xu a, , Yiping Xu b, Xiaowei Jin c,

Yong Yuan d, Zijian Wang e a Key Laboratory of Tropical Marine Bio-resources and Ecology, Guangdong Provincial Key Laboratory of Applied Marine Biology, South China Sea Institute of Oceanology, Chinese Academy of Sciences, Guangzhou 510301, China b Key Laboratory of Drinking Water Science and Technology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing, 10 0 085, China c China National Environmental Monitoring Center, Beijing 10 0 012, China d Guangzhou Key Laboratory Environmental Catalysis and Pollution Control, Guangdong Key Laboratory of Environmental Catalysis and Health Risk Control, School of Environmental Science and Engineering, Institute of Environmental Health and Pollution Control, Guangdong University of Technology, Guangzhou 510 0 06, China e State Key Laboratory of Environmental Aquatic Chemistry, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 10 0 085, China

a r t i c l e i n f o a b s t r a c t

Article history: Novel brominated flame retardants (NBFRs), which have been developed as replacements for legacy flame Received 18 December 2020 retardants such as polybrominated diphenyl ethers (PBDEs), are a class of alternative flame retardants

Revised 16 April 2021 with emerging and widespread applications. The ubiquitous occurrence of NBFRs in the aquatic environ- Accepted 17 April 2021 ments and the potential adverse effects on aquatic organisms have initiated intense global concerns. The Available online 21 April 2021 present article, therefore, identifies and analyzes the current state of knowledge on the occurrence, bioac- Keywords: cumulation, fates, and environmental and health risks of NBFRs in aquatic environments. The key findings Novel brominated flame retardants (NBFRs) from this review are that (1) the distribution of NBFRs are source-dependent in the global aquatic envi- Aquatic environments ronments, and several NBFRs have been reported at higher concentrations than that of the legacy flame Bioaccumulation retardants; (2) high bioaccumulative properties have been found for all of the discussed NBFRs due to

Persistence their strong hydrophobic characteristics and weak metabolic rates; (3) the limited information available Risk assessment suggests that NBFRs are resistant to biotic and abiotic degradation processes and that sorption to sludge and sediments are the main fate of NBFRs in the aquatic environments; (4) the results of ecological risk assessments have indicated the potential risks of NBFRs and have suggested that source areas are the most vulnerable environmental compartments. Knowledge gaps and perspectives for future research regarding the monitoring, toxicokinetics, transformation processes, and development of ecological risk assessments of NBFRs in aquatic environments are proposed. ©2021 Elsevier Ltd. All rights reserved.

1. Introduction mocyclododecanes (HBCDs), have been banned or strictly phased out due to their persistent, bioaccumulative, and toxic (PBT) ef- Brominated flame retardants (BFRs) have been extensively fects on environmental and human health ( Covaci et al. 2006 , used as additives worldwide in products such as foams, resins, Covaci et al. 2011 , Yu et al. 2016 ). In particular, octa-, penta- rubbers, adhesives, plastics, textiles, electronics, and construc- , and deca-brominated diphenyl ethers (octa-, penta-, or deca- tion materials to comply with fire safety standards and regu- BDEs) and HBCDs have been listed as persistent organic pol- lations ( Alaee et al. 2003 , Sjödin et al. 2003 ). In recent years, lutants (POPs) under the United Nations Stockholm Convention several traditional BFRs, such as some polybrominated diphenyl ( UNEP 2013 , 2009 , 2014 ); both deca-BDE and HBCD were also ethers (PBDEs), polybrominated biphenyls (PBBs) and hexabro- added to the list of priority substances in China in January 2018 ( MEE 2017 ). As a result, novel (or new) brominated flame retardants (NBFRs) have been utilized as alternatives, as have ∗ Correspondence

E-mail address: [email protected] (X. Xu). organophosphate ﬂame retardants (OPFRs), dechlorane plus (DP),

https://doi.org/10.1016/j.watres.2021.117168 0043-1354/© 2021 Elsevier Ltd. All rights reserved. R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168 and other flame retardants ( Covaci et al. 2011 , Ezechiáš et al. 2014 ). et al. 2016 ). Therefore, the bioaccumulation, fate, and ecological A mixture of NBFRs with 2-ethylhexyl-2,3,4,5-tetrabromobenzoate and health risks of NBFRs in aquatic environments are issues of (TBB) and bis(2-ethylhexyl)-3,4,5,6-tetrabromo-phthalate (TBPH) major concern. has been used as the main component to replace penta-BDEs Over the past few years, several articles have summarized de- ( Tao et al. 2016 ), while 1,2-bis(2,3,4,5,6-pentabromophenyl)ethane tailed information on the production, physico-chemical properties, (DBDPE) and 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE) are the usage, environmental distribution, analytical methods, human ex- main substitutes used to replace deca- and octa-BDEs, respectively posure, and toxicities of some NBFRs ( Covaci et al. 2011 , Ezechiáš ( Covaci et al. 2011 , Ezechiáš et al. 2014 , McGrath et al. 2018 ). The et al. 2014 , McGrath et al. 2017a , Papachlimitzou et al. 2012 , global production of NBFRs is reported to range from 100 to 180 kt Xiong et al. 2019 , Zuiderveen et al. 2020 ). However, there has annually ( Zuiderveen et al. 2020 ). been no systematic review of the current state of knowledge re- NBFRs can be divided into three classes: monoaromatic, which garding the presence, bioaccumulation, and ecological and health mainly includes polybromobenzene analogues; multi-aromatic, risks of NBFRs in aquatic environments. Recently, abundant data which includes BTBPE, DBDPE, and several tetrabromobisphenol A on NBFRs (including those formerly called alternative, emerging, (TBBPA) derivatives; and cycloaliphatic, which includes polybro- or non-PBDE BFRs) in water and in certain aquatic compart- mocyclohexane analogues and 1,2-dibromo-4-(1,2-dibromoethyl) ments (biota and sediments) have been investigated throughout cyclohexane (TBECH) isomers ( Table 1 ) ( Chen et al. 2013 , the world, which has provided an opportunity to assemble and in- Covaci et al. 2011 , Ezechiáš et al. 2014 , Li et al. 2017 ). How- tegrate a review on the environmental occurrence, fate, bioaccu- ever, these chemicals share a similar structure with halo- mulation and ecological effects of NBFRs. The main objectives of genic substitution in cyclic hydrocarbons, which can cause their this review are: 1) to obtain a clear understanding of the envi- physico-chemical properties to be generally analogous to those ronmental occurrence, temporal and spatial distributions, and fate of PBDEs and HBCDs. Most NBFRs have hydrophobic proper- of NBFRs in aquatic environments; 2) to identify NBFRs exhibiting ties, semivolatile characteristics (vapour pressures of 6.35 ×10 −15 - high bioaccumulation, persistence, and potential risks in aquatic 1.46 ×10 −3 mm Hg at 25 °C), relatively high octanol-water par- environments; and 3) to identify knowledge gaps and propose fu- > < tition coefficients (Log KOW 4), and low water solubility ( ture research needs. 1 mg/L at 25 °C). NBFRs are additively incorporated into poly- meric materials ( Harju et al. 2008 ), which leads to ready leach- 2. Source and environmental occurrence ing into various environmental matrices through industrial production ( Gouteux et al. 2008 , Li et al. 2016b ), use ( Hong et al. 2018 , In recent years, many monitoring studies have focused Olukunle and Okonkwo 2015 ), and disposal ( McGrath et al. 2017a , on the distribution of NBFRs in aquatic environment matri- Xiong et al. 2019 ). NBFRs have been frequently reported in various ces and a diverse group of aquatic animals. NBFR production environmental matrices globally, such as dust ( Abbasi et al. 2016 , sites ( Wei et al. 2012 ), manufacturing areas ( He et al. 2012 ), Bu et al. 2019 , Qi et al. 2014 , Stapleton et al. 2008 ), air and electronic waste (e-waste) recycling and disposal sites ( Hoh et al. 2005 , Liu et al. 2020a , Ma et al. 2013 ), sediments ( Morin et al. 2017 , Wu et al. 2010 ) have been identified as the ( Guerra et al. 2010 , Mo et al. 2019 , Sim et al. 2009 ), soils main sources of NBFRs in aquatic environments. WWTPs constitute ( Jans 2016 , McGrath et al. 2017b ), biota (La Guardia et al. 2012 , both an important entry pathway into the aquatic environments Shi et al. 2009 , Wu et al. 2010 ), and humans ( Chen et al. 2019a , for NBFRs and a large repository of NBFRs released from indus- Chen et al. 2019b , Kawashiro et al. 2008 , Sales et al. 2017 ). trial and living activities (La Guardia et al. 2012 , Morin et al. 2017 , Some NBFRs have also been detected in the Arctic environment, Wang et al. 2020a ). Overall, these global investigations also iden- which indicates their potential persistence ( De Wit et al. 2010 , tified the transfer of NBFR concentrations from potential emission Gewurtz et al. 2020 , Harju et al. 2008 ). Numerous concerns have source areas to receiving environments, including rivers, lakes, es- been expressed about their ubiquitous environmental occurrences tuaries, and marine environments. However, the NBFR profiles in ( Gramatica et al. 2016 , Xiong et al. 2019 ). aquatic environments were spatially specific, which may be mainly The emergence of NBFRs in the aquatic environments and their attributed to the influence of various waste emission sources. adverse effects could pose threats to aquatic ecosystems and hu- Therefore, we considered sampling regions and divided the col- man health. NBFRs enter the aquatic environments through both lected data into three groups for discussion: (i) emission sources point and nonpoint sources and sink in sediments after vari- (i.e., the BFR and product manufacturing facilities, disposal sites, ous transport, partitioning, and degradation processes. Though and WWTPs); (ii) receiving environments; and (iii) remote areas effective removal of NBFRs was consistently found during the (i.e., Arctic, Antarctic, and oceans). The occurrences of individual treatment processes in wastewater treatment plants (WWTPs), NBFRs from these three categories in aquatic environments are most of the compounds were preferentially disposed into sludge shown in Fig. 1 and Figs. S1-S4 (with details in Tables S1-S3). ( Zeng et al. 2014 ). These behaviours are consistent with those of legacy BFRs, depending on their similar physic-chemical prop- 2.1. DBDPE erties and sources ( Covaci et al. 2011 , Iqbal et al. 2017 ). Due to lipophilicity, NBFRs can accumulate at high levels in vari- 2.1.1. Pollution sources ous aquatic organisms and possibly spread throughout the food China and the USA are the major global producers of DB- chain, thereby posing potential threats to aquatic ecosystems DPE ( Covaci et al. 2011 , Shen et al. 2019 , Wei et al. 2012 ). DB- and humans ( De Wit 2002 , Zuiderveen et al. 2020 ). In ad- DPE production is currently the second largest among those of dition, NBFRs have been determined to have patterns of eco- brominated flame retardants produced in China, with a production logical and health toxicities similar to those of legacy BFRs level of 230 thousand tons (kt) between 2006 and 2016, and the ( Stieger et al. 2014 ). NBFRs have been documented as toxic to amount of DBDPE produced in China is expected to gradually rise aquatic organisms ( Ezechiáš et al. 2014 , Stieger et al. 2014 ), and to 88 kt in 2026 ( Shen et al. 2019 ). DBDPE was reported at very available evidence suggests that several NBFRs cause oxidative high levels (3.0-870 ng/g dw) in lake sediments in Arkansas, USA, stress, potential neurotoxicity, and endocrine disruption in aquatic which are located close to the major BFR manufacturing facility in animals ( Bearr et al. 2010 , Feng et al. 2013 , Ma et al. 2018 , the USA ( Wei et al. 2012 ). DBDPE concentrations ranging from 4.6 Pradhan et al. 2013 , Tomy et al. 2007 , Usenko et al. 2016 , to 34,0 0 0 ng/g dw were also detected in soil surrounding the BFR Wang et al. 2019 ) and humans ( Brown et al. 2004 , Klop ciˇ cˇ manufacturing plants in Shouguang City, China ( Li et al. 2016b ).

2 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

K 11.83 6.098 5.699 7.404 6.066 5.268 6.491 4.488 3.376 7.409 Log 4.547

13.64 9.15 8.75 11.95 6.99 7.33 7.48 6.34 4.18 5.24 11.52 Log K

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3.42 0.35 0.23 0.10 9.71 2.23 9.44 1.92 0.08 70 0.92 Solubility (mg/L) at25 Program

Estimation 549.92 486.62 551.49 500.65 971.22 687.64 943.61 706.14 530.67 330.80 427.82 Molecular mass

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84852-53-9 37853-59-1 21850-44-2 183658-27- 7 26040-51-7 87-83-2 87-82-1 85-22-3 35109-60-5 118-79-6 3322-93-8 CAS Protection

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1,2-bis(2,3,4,5,6- pentabromophenyl) ethane 1,2-bis(2,4,6 ethane 2,3,4,5,6-pentabromotoluene 2-ethylhexyl-2,3,4,5- tetrabromobenzoate tetrabromobisphenol bis(dibromopropyl bis(2-ethylhexyl)-3,4,5,6- tetrabromo-phthalate hexabromobenzene 2,3,4,5,6- pentabromoethylbenzene 2,3-dibromopropyl-2,4,6- tribromophenyl ether 2,4,6-tribromophenol 1,2-dibromo-4-(1,2- dibromoethyl) cyclohexane Chemical Table Chemical The

3 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

Fig. 1. Total NBFR concentrations (ng/g dw) in sediments (or sludge) from WWTPs, source neighboring areas, freshwater environments, and coastal marine environments around the world (A); NBFR concentrations (box plots) and compositions (pie plots) in WWTP sludge and sediments from various regions (B); and relationship between

Log octanol-water partition coefficient (Log K OW ) and the observed Log solid-water distribution coefficient (Log K SW ) of NBFRs in aquatic environments and WWTPs (C). Concentrations below the method quantification limit are given as one-half of the method detection, and the NBFRs composition (%) is defined as the mean concentration for individual NBFR divided by the total NBFR concentration.

Many investigations in Europe and China have also related the oc- Cristale et al. 2013 ), and Lake Maggiore (industrial area) in Italy currence and distribution of DBDPE in the aquatic environments ( Poma et al. 2014 ). The highest levels of DBDPE in environmental to BFR product manufacturing and disposal point sources. Elevated media near the source area were in sediments from the e-waste re- levels of DBDPE ( > 100 ng/g dw) have been found in sediments cycling and disposal sites of Guangdong, China ( Wu et al. 2010 ). In from the Pearl River (electronics manufacturing centre) in China addition, the production process (for DBDPE and products in which ( He et al. 2012 , Shi et al. 2009 ), the Ebro and Llobregat Rivers (ur- it is contained) was estimated to produce a much larger (90%) pro- ban and industrial discharge areas) in Spain ( Barón et al. 2014 , portion of the total emissions of DBDPE than the use and disposal

4 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168 processes during 2006-2016 in China ( Shen et al. 2019 ), indicating 2.2.2. Receiving environments the signiﬁcance of industrial sources responsible for its distribution Most investigations of BTBPE in the receiving aquatic envi- in aquatic environments. ronments of Europe and China have focused on coastal and es- The availability of information on the occurrence of DBDPE in tuarine areas, and BTBPE concentrations < 0.04 ng/g dw have WWTP sludge is greater than that for wastewater (Fig. S1). Ele- been reported in these regions (Fig. S2). From the available data, vated levels of DBDPE (mean: 1113 ng/g dry weight (dw)) were it is clear that the sediment levels of BTBPE in North Amer- found in sludge samples from Guangdong (China) ( Shi et al. 2009 ), ica exceed those in Europe and China. Signiﬁcantly, BTBPE was in Barcelona (Spain; mean: 394 ng/g dw) ( Cristale and Lacorte the most accumulated NBFR in surface water ( Law et al. 2010 , 2015 ), in Harbin (China; mean: 256 ng/g dw) ( Li et al. 2018 ), and in Venier et al. 2014 ) and sediments ( Qiu et al. 2007 , Yang et al. 2012 ) Beijing (China; mean: 184 ng/g dw) ( Wang et al. 2020a ), whereas in the Great Lakes, with concentrations even higher than those the DBDPE concentrations in sludge from North America, Oceania, of PBDEs ( Qiu et al. 2007 ). The predominance of BTBPE was also and South Africa were slightly lower. found in sediments from Ulsan Bay in Korea and the urban watershed of Singapore (up to 5.22 ng/L). In addition, more observations of the background levels of BTBPE in the ocean have revealed that

2.1.2. Receiving environments BTBPE is a global contaminant. Generally, the DBDPE concentrations exhibited a decreasing trend from the emission area to the receiving environments (Fig.

S1). In accordance with the DBDPE levels found in emission 2.2.3. Biota source areas and WWTPs, DBDPE is the most abundant NBFR A series of surveys has been conducted to evaluate the occur- found in sediment samples in Chinese rivers (i.e., the Fuhe River rence and accumulation of BTBPE in a diverse group of aquatic

(Hu et al. 2010), Yellow River (Wang et al. 2017a), Baiyang- animals, including invertebrates, ﬁshes, cetaceans, and birds from dian (Hu et al. 2010) and Jinnan Rivers (Hou et al. 2019)) and freshwater and seawater ecosystems (Fig. S2). BTBPE ( > 181 ng/g

European rivers (i.e., the Llobregat (Guerra et al. 2010), Ebro lw) was highly accumulated in aquatic animals from source areas,

(Barón et al. 2014), and Po Rivers (Luigi et al. 2015)), and it ac- such as mussels from the industrial sewage outfall area in the Yad- counted for greater than 70% of the total NBFR concentration. In kin River (La Guardia et al. 2012 ), ﬁsh from an e-waste recycling

China and Europe, DBDPE predominated in samples from the North site in China ( Wu et al. 2010 ), and ﬁsh from an industrial area

Sea (Ricklund et al. 2010), East China Sea (Zhu et al. 2013), and in South Africa ( Chokwe et al. 2020 ). The concentrations of BTBPE

South China Sea (Mo et al. 2019). in aquatic animals from receiving environments were much lower than those in aquatic animals from source areas (one-way ANOVA, < 2.1.3. Biota P 0.05). In general, the DBDPE levels in biota from the emission areas were signiﬁcantly higher than those from the receiving environments (Fig. S1; one-way ANOVA, P < 0.05). In ﬁsh samples from in- 2.3. TBB and TBPH dustrial areas of China and South Africa, the mean DBDPE concentration exceeded 300 ng/g lw ( Chokwe et al. 2020 , Wu et al. 2010 ). 2.3.1. Pollution sources

High levels of DBDPE were also found in aquatic birds from the TBB and TBPH, two closely associated NBFRs, are the two main

Pearl River, China (nd-800 ng/g lw) ( Luo et al. 2009 , Mo et al. 2012 , components of FireMaster 550 (35% TBB and 15% TBPH), BZ-54

Shi et al. 2009 ). DBDPE was detected in 53% of mussel samples, (70% TBB and 30% TBPH), and DP-45 (TBPH only), which are pro- with mean levels of less than 2.52 ng/g lw in mussels from Eu- duced by Chemtura Corporation (Arkansas, USA) as substitutes for ropean ﬁsh farms across 7 countries ( Aznar-Alemany et al. 2018 ). penta-BDE (Stapleton et al. 2012). Therefore, TBB and TBPH usu-

DBDPE has also been detected in seawater and freshwater ﬁsh in ally occur simultaneously in sediments from source areas (Fig.

Europe and North America, reflecting its universal occurrence in S3), such as industrial areas (La Guardia et al. 2012) and land- the environment. fill sites (Olukunle and Okonkwo 2015). Relatively high concentrations of TBB (accounting for 34.5% of total NBFRs) were discovered in surface water near a coastal area of the Bohai Sea, China, and 2.2. BTBPE TBB was determined to be emitted from cargo ships rather than mariculture-associated activities ( Wang et al. 2017b ). However, it 2.2.1. Pollution sources was unknown whether TBB and TBPH were made in other coun- BTBPE is mainly produced as FR-680 by Great Lakes Chemical tries ( Lam et al. 2009 ). (Chemtura Corp., Arkansas) in the USA, but there is no report on It should be noted that TBPH has been reported to be the BTBPE concentrations in point sources neighbouring BFR produc- most abundant NBFR in WWTP sludge samples from Europe (Den- tion plants. However, many studies have outlined the high accumu- mark, Finland, Sweden, and Spain), with mean concentrations lation of BTBPE in sediments from urban areas and industrial areas exceeding 20 ng/g dw, implying that TBB and TBPH might be in Singapore ( Wang and Kelly 2017 ), China ( Wang et al. 2017b ), two of the most commonly used two alternatives for traditional and South Africa ( Olukunle and Okonkwo 2015 ). In the USA, textile BFRs in Europe ( Cristale and Lacorte 2015 , Schlabach et al. 2011 ). manufacturers have been identified as very important contributors TBPH has also been found to be the second most abundant NBFRs of BTBPE to local river sediments (La Guardia et al. 2012 ). Sedi- in WWTP sludge from Harbin and Beijing, China ( Li et al. 2018 , ment samples collected from the e-waste recycling and disposal Wang et al. 2020a ). In two WWTPs from Washington state (WA), sites of Guangdong, China exhibited the highest BTBPE concentra- USA, TBB and TBPH have been detected in laundry wastewater tion (4554 ± 608 ng/g ww), indicating the significance of e-waste and the influent samples, which indicated that laundry wastewa- recycling in BTBPE emissions ( Wu et al. 2010 ). ter may be the potential source of these two FRs in aquatic en- Generally, BTBPE concentrations were reported in the range of vironment ( Schreder and La Guardia 2014 ). Because North Amer- 0.1 to several ng/g dw in sludge samples from Europe and China ica, especially the Great Lakes regions, is the manufacturing base ( Nyholm et al. 2013 , Schlabach et al. 2011 , Shi et al. 2009 ), but its for TBB- and TBPH-containing FR products, the contribution of concentration in sludge from Harbin, China exceeded 200 ng/g dw NBFR emissions from WWTPs to aquatic environments cannot be ( Li et al. 2018 ). ignored.

5 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

2.3.2. Receiving environments ( Cristale et al. 2013 ). However, very few studies have investigated Both TBB and TBPH were detected in environmental media the occurrence of TBBPA-DBPE in aquatic environments and biota, from the Great Lakes. Their levels were correlated, suggesting that and this constitutes an important information gap (Table S2 and they have a common source ( Venier et al. 2014 ). However, TBB Table S3). and TBPH were not consistently detected in other studies, such as those of coastal sediments from the Great Lakes, Europe, and 2.6. Other NBFRs China ( Chen et al. 2013 , Liu et al. 2014 , Sühring et al. 2016 , Wu et al. 2010 , Zhu et al. 2013 ). This may be due to the differences 2.6.1. Pollution sources in the chemical composition of the technical FRs used in these re- Brominated benzenes constitute another group of NBFRs gions or the different fates and environmental behaviours of TBB used to replace PBDEs in a range of polymers, and include and TBPH. hexabromobenzene (HBB), 2,3,4,5,6-pentabromotoluene (penta-BT), In accordance with the predominant contamination pattern of 2,3,4,5,6-pentabromoethylbenzene (PBEB), and 2,3-dibromopropyl- TBPH in European WWTPs, TBPH was also found to be the ma- 2,4,6-tribromophenyl ether (DPTE) ( Covaci et al. 2011 , Ezechiáš jor NBFR component in the sediments of the Elbe River and the et al. 2014 ). The production of these NBFRs was reported in the North Sea ( Sühring et al. 2016 ). In South Africa, the sum of TBB and 1980s, and they might have re-emerged in the market in recent TBPH concentrations accounted for more than 64% of the detected years ( Jans 2016 ). PBEB is produced by Albemarle Corp. in the NBFRs in inland sediments collected in the eThekwini metropoli- USA, while HBB and penta-BT are primarily used in Japan and tan municipality (La Guardia et al. 2013 ). TBPH has been recently China ( Covaci et al. 2011 ). DPTE was first produced by Chemis- observed in seawater from remote regions of the oceans and the che Fabrik Kalk in Germany under the trademark Bromkal 73-5 PE Arctic ( Möller et al. 2011a , Möller et al. 2011b ). ( Vetter et al. 2010 ), but very little current information is available on its current use and production. 2.3.3. Biota The environmental distribution of these NBFRs was in strict Extremely high concentrations of TBB and TBPH (up to accordance with their usage patterns among different countries 2220 ng/g lw) have been detected in mussels and gastropods (Fig. S4). DPTE was detected in surface water from the Elbe and from the emission source area of the Yadkin River (USA) (La Weser Rivers of Germany with concentrations up to 69.8 pg/L Guardia et al. 2012 ). TBPH has also recently been detected as the ( Möller et al. 2012 ), and PBEB was found to be the predominant major NBFR congener in the muscle of freshwater fish from South NBFR in surface water from an industrial area in the Great Lakes Africa, as well as DBDPE and BTBPE ( Chokwe et al. 2020 ). TBB and region ( Venier et al. 2014 ). Elevated levels of HBB have been found TBPH have also been detected at concentrations less than 36.8 ng/g in surface water (up to 52.0 pg/L) and sediments (up to 8672 ng/g lw in invertebrates and fish from the receiving environments and ww) from e-waste recycling sites in China, and penta-BT was also remote areas ( Abbasi et al. 2017 , Hou et al. 2019 , Houde et al. 2014 , detected ( Wu et al. 2010 ). La Guardia et al. 2012 , Labunska et al. 2015 , Li et al. 2019 , Similarly, TBB, HBB, PBEB, and DPTE have also been frequently Santín et al. 2013 , Schlabach et al. 2011 , Vorkamp et al. 2015 ) detected in WWTP worldwide (Fig. S4). DPTE and PBEB were the predominant compounds for these NBFRs detected in WWTPs in 2.4. TBECH Europe and North America. In an investigation of WWTPs across 24 provinces in China, the results revealed that HBB was the pre- TBECH is produced by the Albermarle Corporation (Charlotte, dominant compound among the targeted NBFRs in sludge, and USA) under the commercial name Saytex BCL-462, and the only the simultaneous presence of DBEPE and BTBPE were also found available report on annual production volume was 230 tons in ( Zeng et al. 2014 ). the USA in 2002 ( Covaci et al. 2011 ). TBECH was consistently detected in WWTP wastewater or sludge in Europe, with α- 2.6.2. Receiving environments and β- TBECH as the predominant TBECH diastereomers (Table S1) Penta-BT, PBEB, and DPTE are also consistently detected ( Arp et al. 2011 , Nyholm et al. 2013 ). In China, previous studies in receiving environments worldwide but with lower concen- have reported relatively high concentrations of TBECH in sludge, tration ranges (Fig. S4) ( Chen et al. 2013 , Liu et al. 2014 , with reports of 126 ng/g dw from Harbin ( Li et al. 2018 ) and Sühring et al. 2016 , Wu et al. 2010 , Zhu et al. 2013 ). The highest 23.5 ng/g dw from Hong Kong ( Ruan et al. 2019 ). levels of these four NBFRs in surface water were found in the Jin- In recent studies of environmental media, TBECH showed mean jiang River of China, in which the levels of penta-BT (nd-9.21 ng/L) concentrations of 1.27 ng/g dw and 3.74 ng/g dw in river sediments and HBB (nd-2.14 ng/L) were the most prominent ( Hou et al. 2019 ). from an urban watershed in Singapore ( Wang and Kelly 2017 ) and Some studies also focused on the background levels of these from the Yangzi River Estuary of China ( Yang et al. 2012 ), respec- NBFRs in seawater samples from remote areas in the Arctic, Pa- tively. TBECH was found to be one of the predominant NBFRs in cific, and Atlantic Oceans ( Möller et al. 2011a , Möller et al. 2011b , mussels from Taihu Lake, with a mean concentration of 5.07 ±4.57 Xie et al. 2011 ). ng/g lw ( Zheng et al. 2018 ). 2.6.3. Biota 2.5. TBBPA-DBPE In a study on freshwater fish from e-waste recycling sites in China, the peak concentrations of penta-BT, HBB, and PBEB in Tetrabromobisphenol A bis(dibromopropyl ether) (TBBPA- fish were approximately 3.24, 2451, and 37 ng/g lw, respectively DBPE) is a globally produced NBFR ( Covaci et al. 2011 , ( Wu et al. 2010 ). DPTE was found to be the predominant NBFR in Liu et al. 2016 ). TBBPA-DBPE had the highest level (80 ng/L) elvers from German rivers ( Sühring et al. 2013 ), and HBB, PBEB, in wastewater at a metal recycling and car dismantling site and DPTE were found at comparably high concentrations (up to in Norway ( Nyholm et al. 2013 ), which indicated that waste 7.6, 29, and 460 ng/g lw, respectively) in the muscles of dabs recycling activities are an important potential source of TBBPA- and lees from the North Sea ( Sühring et al. 2016 ). Several studies DBPE. Herring gull eggs from the Great Lakes of North America have investigated the NBFRs in fish and aquatic birds inhabiting were found to be contaminated with TBBPA-DBPE (nd to 42.8 the Great Lakes and the St. Lawrence River ( Gauthier et al. 2007 , ng/g ww) ( Gauthier et al. 2019 ), while TBBPA-DBPE was not Gauthier et al. 2019 , Gentes et al. 2012 , Venier et al. 2010 ), where detected in sediments collected from the Besòs River of Spain relatively high concentrations of PBEB were detected.

6 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

2.7. General discussion ( Wei et al. 2012 ), since 1989 in Lake Maggiore of Northern Italy ( Poma et al. 2014 ), and since 1980 in the Great Lakes These global observations indicated that high levels of NBFRs ( Qiu et al. 2007 , Yang et al. 2012 ). As a consequence, the sediment generally occurred in source areas, such as the e-waste recycling concentrations of DBDPE approximately doubled every 3-5 years area in South China, the Pearl River in China, the Great Lakes in Lake Michigan, every 7 years in Lake Ontario ( Yang et al. 2012 ), region, the FR-producing area in Arkansas in the USA, the Ebro every 59 years in Daya Bay of South China ( Liu et al. 2014 ), and and Llobregat Rivers in Spain, and industrial and landfill areas in every 2.8-4.3 years in Arkansas in the USA ( Wei et al. 2012 ), South Africa. There is a general gradient, with higher concentra- whereas the doubling time for BTBPE was estimated to be 21 tions in WWTP sludge and sediments from source emission ar- years in Daya Bay of South China ( Liu et al. 2014 ), 5 years in Lake eas while lower concentrations in the receiving areas ( Fig. 1 -A). Ontario, and 7 years in the State of Michigan ( Yang et al. 2012 ). This can be explained by the dilution effect occurred in the re- The clear shift from the application of PBDEs to the application ceiving environments ( Liu et al. 2020a , Möller et al. 2012 ). More- of NBFRs warrants more concern regarding the predominance of over, estuarine sediments exhibited an obvious decrease in NBFR NBFRs in aquatic environments. levels with increasing distance between the riverine and the marine environments of the Yangzi River estuary ( Zhu et al. 2013 ), 3. Bioaccumulation and biomagnification Potential the Pearl River estuary ( Chen et al. 2013 ), and the Elbe River estuary ( Sühring et al. 2016 ). Terrestrial input was found as a signif- 3.1. Bioaccumulation icant source of NBFRs in marine environments ( Chen et al. 2013 , Sühring et al. 2016 , Zhu et al. 2013 ). NBFRs have also been widely Bioaccumulation factors ( BAFs ) are expressions of the net bioac- detected in marine environments globally, including those of the cumulation of the chemicals by an organism as a result of up- Atlantic, Arctic, Pacific, and Southern Oceans ( Möller et al. 2011a , take from all environmental sources and processes such as water Möller et al. 2011b , Xie et al. 2011 ), but at relatively low levels. and food, whereas bioconcentration factors ( BCFs ) express the up- This suggested that the concentrations and distribution of NBFRs take of chemicals originating from water. High field-derived BAF s were source-dependent in freshwater aquatic environments and of NBFRs have been reported ( Fig. 2 and Table S4). The BAFs coastal areas. of BTBPE, penta-BT, PBEB, and HBB were respectively determined The NBFR profiles in aquatic environments varied among dif- to be 2.09 ×10 3 -1.20 ×10 6 , 110-5.89 ×10 4 , 2.04 ×10 3 -3.47 ×10 5 , and ferent areas, indicating dissimilar NBFR usage and emission pat- 525-1.23 ×10 4 , respectively, for different aquatic species in an elec- terns across countries ( Fig. 1 -B and Fig. S5). DBDPE, a substitute tronic waste recycling site from the Pearl River ( Wu et al. 2011 ), for deca-BDE, is the predominant NBFR worldwide, mainly due to while the lipid - normalized BAFs for DPDPE were found to be its large production and usage volume. The results also show that 5.01 ×10 6 -1.26 ×10 7 in fish species of the Pearl River in another BTBPE, TBB and TBPH are more abundant than the local pollutants. study ( He et al. 2012 ). TBB and TBPH also presented high BAFs Although these NBFRs are known to be produced in the USA, they (2.49 ×10 3 -3.18 ×10 3 and 1.85 ×10 3 -1.05 ×10 5 , respectively on lipid are likely imported into other countries in manufactured products, basis) in mussels from the Yadkin River (La Guardia et al. 2012 ). leading to emissions into aquatic environments worldwide. High Only one study is available that considered the whole body biocon- penta-BT and HBB concentrations in China possibly could be ex- centration factors ( BCFs whole body) for NBFRs based on lab expo- plained that they are more frequently used in China than in other sure experimentation ( Oliver and Niimi 1985 ), where the BCFs of countries. DPTE and PBEB were the major NBFRs in aquatic en- penta-BT, HBB, and PEBE were 270, 1.10 ×10 3 , and 330 for rainbow vironment samples from Europe and North America, respectively, trout. Notwithstanding the variations of these bioaccumulation fac- which is consistent with their production and usage trends (as dis- tors among studies, all the studied NBFRs can be considered to be cussed in 2.6.1). highly bioaccumulative ( BAF or BCF > 500) or even very bioaccu- It is worth noting that NBFR concentrations recently exhibited mulative (vB, BAF or BCF > 5,0 0 0) in some cases, based on clas- increases in aquatic environments worldwide, indicating the effect sification criterion of the European Union Registration, Evaluation of the phase-out PBDEs and their replacement by NBFRs. In an and Authorization of Chemicals (REACH) regulations ( EU 2006 ) 8-year investigation of WWTP influents in Harbin, China (2009- ( Fig. 2 ). 2016), the sum of the concentrations of NBFRs in the influents The biota to sediment accumulation factor ( BSAF ) is especially increased from 2009 to 2014, whereas the deca-BDE concentra- useful for the predicting or describing NBFRs bioaccumulation in tion decreased significantly from 2010 to 2012 ( Li et al. 2018 ), aquatic animals, and a BSAF value exceeding 1 implies that NBFR consistent with the replacement of PBDEs with NBFRs in China bioaccumulation has occurred from sediments ( Burkhard 2009 ). As since 2009. The levels of DBDPE (a replacement for deca-BDE) determined by field studies, the BSAFs < 1 generally occurred for have already approached or exceeded those of deca-BDE in sedi- NBFRs in mussel and fish species, possibly due to the lower lipid ment samples taken from the Pearl River Delta ( Chen et al. 2013 , contents present in mussel and fish than the organic matter in sed- Zhang et al. 2009 ), Yangtze River Delta ( Zhu et al. 2013 ), coastal iments. DBDPE, BTBPE, TBB, and TBPH exhibited lower BSAFs ( < 1) area near Hainan Island ( Mo et al. 2019 ), lakes and marine environ- in mussels from Europe ( Aznar-Alemany et al. 2018 ) and the USA ments in Sweden ( Ricklund et al. 2010 ), and the Ebro and Llobregat (La Guardia et al. 2012 ). BSAFs less than 1 were reported for DB- river basins ( Barón et al. 2014 ). It is worth noting that the NBFR DPE, TBBPA-DBPE and HBB in eels from the German Bight and the concentrations in marine environments in several studies were Elbe River ( He et al. 2012 , Sühring et al. 2016 ). TBB and TBECH higher than those of PBDEs ( Liu et al. 2020a , Möller et al. 2011a , showed the larger BSAFs (3.82 and 5.10, respectively) in crucian Möller et al. 2011b , Xie et al. 2011 ), even in high Arctic areas from Jinnan River ( Hou et al. 2019 ). Since sediments are not the ( Möller et al. 2011b ). These results suggested that the emerging only source of NBFRs for aquatic animals, the BSAFs from these contamination from NBFRs is replacing that from the deca-BDE field studies may be overestimated ( Hou et al. 2019 ). mixture in these regions. Investigations of NBFRs in sediment cores can provide 3.2. Biomagnification additional information for constructing temporal trends in NBFR contamination. A rapid increase in the NBFR concentra- In aquatic animals from the receiving environments, the tions of sediments has occurred since 1990 in Daya Bay of NBFRs concentration showed an increasing trend along with the South China ( Liu et al. 2014 ), since 20 0 0 in Arkansas, USA trophic level, showing the possible biomagnification along the food

7 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

Fig. 2. Summary of the investigated and estimated bioaccumulation properties (A: estimated BAFs or BCFs ; B: investigated BAFs or BCFs ; C: investigated BSAFs ; and D: investigated TMFs ) of NBFRs. The criteria values for bioaccumulation (“B”), very bioaccumulation (“vB”), bioaccumulation (“B”) from sediments, and biomagnification are noted. Detailed data are compiled in Table S4. chain (Fig. S6). The high trophic transfer potentials of NBFRs has trast, the large range of BMFs for BDBPE (0.2-9.2) and BTBPE (0.2- already been supported by high concentrations reported in aquatic 2.5) from the Lake Winnipeg food web indicated differences in animals with upper trophic levels compared to those in animals trophic transport potential among predator/prey feeding relation- with lower trophic levels from aquatic food chains in South China ships ( Zheng et al. 2018 ). ( Wu et al. 2011 , Zhang et al. 2011 ). A trophic magnification factors ( TMFs ) > 1 is indicative Available studies on the estimated biomagnification factors of biomagnification in aquatic food webs and has been re- ( BMFs ) for predators are summarized in Table S4. A BMF of > 1 rep- ported for DBDPE and BTBPE in the Lake Winnipeg food web resents the state of the pollutants that may experience biomagni- ( Law et al. 2010 ), HBB in the freshwater food web from the Pearl fication, while a BMF < 1 represents the possible nutrient dilution. River ( Wu et al. 2010 , Zhang et al. 2010 ), and BTBPE and TBPH in The laboratory-derived BMFs of BTBPE in rainbow trout derived the aquatic food web from Taihu Lake ( Zheng et al. 2018 ). Trophic from its diet were reported to be 2.30 ± 0.90 by Tomy et al. (2007) , dilution ( TMFs < 1), otherwise, have been observed for penta-BT suggesting that biomagnification of BTBPE occurred in the studied and BTBPE in the food web of the Pearl River ( Zhang et al. 2010 ) fish. The presence of NBFRs in aquatic birds demonstrated the sig- and for DBDPE and TBECH in the aquatic food web from Taihu nificance of diets as an accumulation route. BMFs of DBDPE and Lake ( Zheng et al. 2018 ). These controversial results indicate more BTBPE calculated from common kingfishers and their prey fish needed researches to elucidate the biomagnifications and trophic were 0.10-0.77 and 1.90-3.60, respectively ( Mo et al. 2012 ). By con- transfer of NBFRs in various food webs.

8 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

3.3. Metabolism and elimination Considering the liver to be the most important tissue for me- tabolized pollutants in the body, liver microsomes isolated from An important aspect of the bioaccumulation potential for NBFRs various fish species have been used as an in vitro approach to is their resistance to metabolism. Several studies have confirmed evaluate the biotransformation kinetics of NBFRs in several stud- the enzymatic biotransformation of NBFRs into either more or ies ( Table 2 ). The in vitro biotransformation rates of NBFRs were less toxic metabolites (Table S5 and Fig. S7). Nona-bromodiphenyl also investigated in seawater fish liver microsomes (i.e., DBDPE, ethane (nona-BDPE), octa-BDPE, hepta-BDPE, hexa-BDPE, and BTBPE, TBB, TBPH, penta-BT, HBB, PBEB, and α- and β-TBECH; penta-BDPE were found to be debromination metabolites of DB- Zheng et al. (2018) ), marine mammalian liver microsomes (i.e., DB- DPE in snails ( Wang et al. 2020b ). A fathead minnow adminis- DPE; McKinney et al. (2011) ), fathead minnow liver microsomes tered BTBPE and TBBPA-DBPE, produced dibromophenol (DBP) and (i.e., TBB and TBPH; Bearr et al. (2012) ), and common carp liver tetrabromobisphenol A (TBBPA) via O-dealkylation, respectively microsomes (i.e., TBB and TBPH; Bearr et al. (2012) ). Accordingly,

(De Jourdan et al. 2014). Ganci et al. (2017) found the formation of the t1/2 values of NBFRs were estimated to range from 15.8 to 1873 tribromophenol (TBP) and tribromophenoxyethanol during in vitro days in fish, which showed that NBFRs are sensitive to metabolism biotransformation of BTBPE in rainbow trout liver microsomes. by aquatic animals. In fathead minnow and common carp liver microsomes, 2,3,4,5- tetrabromobenzoic acid (TBBA), 2-ethylhexyl dibromobenzoate 3.4. Factors influencing NBFR bioaccumulation (DBB), and 2,3,4,5-tetrabromomethylbenzoate (TBMB), formed via dealkylation and methylation, were detected as metabolites for the 3.4.1. Physico-chemical properties mixture of TBB and TBPH ( Bearr et al. 2012 ). Ganci et al. (2017) fur- The physico-chemical properties of NBFRs may influence their ther confirmed TBBA and mono (2-ethylhexyl) tetrabromophtha- bioaccumulation potential. Based on all the available collected late (TBMEHP) as the specific metabolites of TBB and TBPH, re- data, correlations between the Log BAF and Log KOW values and be- spectively in fish liver. The debromination were also investi- tween Log BAF and molecular weight (MW) were both found to be gated to be the main metabolism pathway for penta-BT and nonsignificant, with weak positive relationships for NBFRs in fish HBB in snail ( Wang et al. 2020b ). Tetrabromotoluene (tetra- ( P > 0.05, Fig. S8). However, when a limit was prescribed for the BT), dibromotoluene (di-BT), and 2,4,6-tribromotoluene (2,4,6-tri- investigated species, a significant positive linear relationship ex-

BT) were found as the major metabolites of penta-BT, and the isted between Log KOW and Log BAFs and relationships between pentabromobenzene (PBB), tetrabromobenzene (1,2,4,5-tetra-BB), molecular weight and Log BAFs (though not significant) were found 1,2,4-tribromobenzene (1,2,4-tri-BB), and dibromobenzene (di-BB) for NBFRs in mussels (La Guardia et al. 2012 , Wu et al. 2011 ) were the metabolites for HBB. Collectively, for the NBFRs with and fishes ( Wu et al. 2011 ). Generally, multi-aromatic NBFRs have ether or ester bond, such as BTBPE, TBB, TBPH, and TBPPA-DBPE, higher theoretical BCFs / BAFs than other NBFRs, and cycloaliphatic hydrolysis is the major metabolic pathway, whereas other NBFRs NBFRs exhibit the lowest accumulation potential. Thus the Log KOW share general metabolic pathways of mono- and multiple hydroxy- values of the NBFRs may act as a significant predictor of their lation and debromination. bioaccumulation potential in a certain species. In addition, certain metabolites of NBFRs also exhibit toxic A negative parabolic relationship was observed between Log effects on organisms. An in vivo study evaluating the biocon- BSAF values and their corresponding Log KOW values for molluscs centration and endocrine impact of DBDPE using zebrafish lar- (La Guardia et al. 2012 ). This observation suggested that NBFRs vae reported that the potential metabolites comprised 88.2% of with higher hydrophobicity tend to have higher accumulation po- the total bromine-containing compounds present in the larvae tential but poorer bioavailability from sediments and that the most at 14 dpf, suggesting the significance of metabolites accumula- important absorption pathway of NBFRs for filter-feeding benthic tion of DBDPE in fish tissues ( Wang et al. 2019 ). The metabo- organisms (mussels) originates from sediment pore water rather lites TBBA and TBMEPH were shown to have comparable thyroid than directly from sediments. hormone, androgen, glucocorticoid, and pregnancy X receptors ag- onist activities (Gramec Skledar et al. 2016 , Klop ciˇ cˇ et al. 2016 ) 3.4.2. Metabolism capacity and induced stronger cytotoxicity than their parent compounds Although physico-chemical properties can be useful for predict- ( Chen et al. 2020 ). Bromophenols, one of which was reported ing NBFR bioaccumulation potentials, they can lead to an over- as a BTBPE metabolite (i.e., 2,4,6-tribromophenol), was found to estimation if metabolism/biotransformation occurs. The BCFs / BAFs have a strong cytotoxic and genotoxic effect on aquatic organ- from laboratory-derived and field investigations were relatively isms ( Yang and Zhang 2013 ). The hydroxylation and debromination lower than the theoretical values estimated by the ACD/Labs Per- pathways were considered to enhance toxicological effects for PB- cepta and USEPA EPI Suite ( Fig. 2 ). In a study of NBFR metabolism DEs ( Dingemans et al. 2008 , Lv et al. 2020 , Meerts et al. 2001 ); in fish liver microsomes, the whole-body biotransformation rates thus, the metabolites of NBFRs derived from the same pathways of NBFRs were used to evaluate BCFs/BAFs ( Lee et al. 2019 ). Sig- may also pose increased risks to freshwater ecosystems. nificant correlations have been observed between BCFs/BAFs and Toxicokinetic or metabolism studies highlighted the slow the model-derived or laboratory-measured metabolic rates of or- metabolism or elimination of these NBFRs from the aquatic ganic chemicals, and the derived whole-body BCFs for BTBPE, TBB, animals ( Table 2 ). In zebrafish exposed to NBFRs (a mixture HBB, penta-BT, and PBEB ranged from 16 to 7100 and are close to of TBECH and eight other BFRs) by diet administration for 42 the values calculated from field studies ( Lee et al. 2019 ). There- days ( Nyholm et al. 2009 ), over 60 % of the TBECH in food was fore, metabolism may play an important role in the bioaccumu- < absorbed, and the estimated half-life (t1/2 ) of TBECH was 2 lation of NBFRs in aquatic animals, but its contribution warrants day. Tomy et al. (2007) reported a mean assimilation efficiency and further evaluation. a t1/2 value of BTBPE in rainbow trout juveniles at environmen- It has also been suggested that the metabolic rate is an im- tally relevant dose were 27 ± 3 % and 54.1 ± 8.5 days, respectively. portant factor influencing the biomagnification of organic com- In a study of NBFRs (i.e., penta-BT, HBB, and DBDPE) in sediment- pounds, in which the decrease in metabolic rate with increasing water-mudsnail systems, the distribution of NBFRs in snail viscera trophic levels contributes to the biomagnification potential in the was approximately 3 times that of pleopods, and the t1/2 values aquatic food web ( Fisk et al. 2001 , Zheng et al. 2016 ). NBFRs with for penta-BT, HBB, and DBDPE were in the range of 2.5-5.9 days higher trophic magnification factors ( TMFs ) were found to have ( Wang et al. 2020b ). slow metabolic rates in fish livers and moderate hydrophobicity

9 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

Table 2 Summary of the investigated toxicokinetic constants of NBFRs in aquatic animals.

Depuration rate Disposition Uptake rate con- Assimilation constant (/day or a Compounds Species Exposureapproaches tissues stant(nmol/day) eﬃciency mL/(h ·mg protein)) t 1/2 (day) References DBDPE snail sediments exposure viscera - - 0.181-0.272 2.5-3.8 ( Wang et al. 2020b ) marine ﬁsh liver in vitro exposure - - - 0.073-0.162 (in vitro) 15.8-17.3 ( Zheng et al. 2018 ) microsome 0.040-0.044 (whole

body) b marine mammal in vitro exposure - - - ≈ 0.185 ≈ 3.75 ( McKinney et al. 2011 ) liver microsome BTBPE rainbow trout diet exposure muscle 0.0069 ±0.0012 27 ±3% 0.0128 ± 0.002 54.1 ± 8.5

juvenile ( Tomy et al. 2007 )

marine ﬁsh liver in vitro exposure - - - 0.13-0.18 (in vitro) 92.4-1873 ( Lee et al. 2019 ) microsome 0.0004-0.0075 (whole body) marine ﬁsh liver in vitro exposure - - - 0.005-0.007 (in vitro) 43.3-57.8 ( Zheng et al. 2018 ) microsome 0.012-0.016 (whole

body) b TBB marine ﬁsh liver in vitro exposure - - - 0.18-0.50 (in vitro) 28.9-1035 ( Lee et al. 2019 ) microsome 0.0007-0.024 (whole body) fathead minnow in vitro exposure - - - 0.0024 ± 0.0006(in 100 ( Bearr et al. 2012 ) liver microsome vitro) 0.0069 ± 0.0019

(whole body) b common carp liver in vitro exposure - - - 0.0016 ± 0.0006 (in 144 ( Bearr et al. 2012 ) microsome vitro) 0.0048 ± 0.0019

(whole body) b TBPH marine ﬁsh liver in vitro exposure - - - 0.001-0.009 (in vitro) 36.5-231 ( Zheng et al. 2018 ) microsome 0.003-0.019 (whole

body) b fathead minnow in vitro exposure - - - 0.0006 ± 0.0002 (in 365 ( Bearr et al. 2012 ) liver microsome vitro) 0.0019 ± 0.0007

(whole body) b common carp liver in vitro exposure - - - 0.0003 ± 0.0002 (in 693 ( Bearr et al. 2012 ) microsome vitro) 0.0010 ± 0.0007

(whole body) b penta-BT snail sediments exposure viscera - - 0.118-0.122 5.7-5.9 ( Wang et al. 2020b ) marine fish liver in vitro exposure - - - 0.05-0.28 (in vitro) 14.7-1284 ( Lee et al. 2019 ) microsome 0.0005-0.047 (whole body) HBB snail sediments exposure viscera - - 0.197-0.272 2.5-3.5 ( Wang et al. 2020b ) marine fish liver in vitro exposure - - - 0.048-0.13 (in vitro) 16.9-533 ( Lee et al. 2019 ) microsome 0.0013-0.041 (whole body) PBEB marine fish liver in vitro exposure - - - 0.052-0.40 (in vitro) 15.1-1733 ( Lee et al. 2019 ) microsome 0.0004-0.046 (whole body) TBP zebrafish diet exposure - - - 0.54 1.3 ( Nyholm et al. 2009 ) TBECH zebrafish diet exposure - - > 60 % 0.52-0.76 0.9-1.3 ( Nyholm et al. 2009 ) mixture α-TBECH marine fish liver in vitro exposure - - - 0.011-0.027 (in vitro) 25.7-63.0 ( Zheng et al. 2018 ) microsome 0.021-0.031 (whole

body) b β-TBECH marine ﬁsh liver in vitro exposure - - - 0.006-0.023 (in vitro) 23.1-49.5 ( Zheng et al. 2018 ) microsome 0.014-0.030 (whole

body) b

a the depuration rate constants for whole body and in vitro liver microsomes were expressed as /day and mL/(h ·mg protein), respectively;

b the whole body clearance rate contants were estimated in this study using an in vitro to in vivo extrapolation (IVIVE) model, where the used hepatosomatic index was 0.02, the value of cardiac output was 0.259, and microsome index was 37.8 mg/g liver (Nichols et al. 2006).

( Zheng et al. 2018 ). The metabolic capacity of organisms at high tissue distribution of NBFRs in ﬁsh. NBFRs were more likely to trophic levels plays an important role in the biomagniﬁcation po- be distributed in muscle and viscera rather than in gill tissues, tential of NBFRs ( Zheng et al. 2018 ). suggesting that the accumulation of NBFRs is essentially associated with lipids ( Hou et al. 2019 ). Several studies have investigated the plasma, liver, and eggs of aquatic bird species (herring gulls, 3.4.3. Biometric parameters bald eagles, and ring-billed gulls) inhabiting the Great Lakes and In addition, NBFRs have a tissue-dependent distribution in the St. Lawrence River ( Gauthier et al. 2007 , Gauthier et al. 2019 , biota, just as the bioaccumulation of other lipophilic compounds Gentes et al. 2012 , Venier et al. 2010 ). Relatively high concentra- is related to hydrophobicity. Only one study has investigated the

10 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168 tions of BEBPE, HBB, and PBEB (mean values of 2.47, 4.96, and 3.72 mosphere ( Liu et al. 2020a , Möller et al. 2011a , Möller et al. 2011b , ng/g lw, respectively) were detected in the eggs. Compared with Xie et al. 2011 ). Therefore, several studies have suggested air- muscle, substantially higher DBDPE levels were found in the liver seawater exchange and dry particle-bound deposition are the im- and kidney of aquatic birds from the Pearl River ( Shi et al. 2009 ). portant contributors to the transport of NBFRs in remote oceans, This result is likely explained by the higher lipid contents in the especially for highly hydrophobic coumpounds ( Möller et al. 2011a , liver relative to those in muscle for the aquatic birds sampled from Möller et al. 2011b ). this region ( Shi et al. 2009 ).

4. Environmental behaviors and fates 4.2. Biotic and abiotic degradation

4.1. Phase distribution characteristics Studies on the fate of NBFRs in aquatic environments describe their abiotic and biotic degradation processes. Only one half-life

The distribution of NBFRs between aqueous and particulate estimate is available for the hydrolysis of TBBPA-DBPE with a t1/2 phases and between water and sediments are important factors of 3.23 days at 50 °C ( ECHA 2020 ), whereas at ambient temperature that affect their environmental behaviours and fates in aquatic en- conditions, the selected NBFRs were determined to be very recal- vironments. The estimated Log organic carbon adsorption coeffi- citrant to hydrolysis via the HydroWin model using the USEPA EPI cients from the EPI Suite (Log KOC ; from 3.376 to 11.83) indicate Suite (Table 3). It can be suggested that hydrolysis is not an impor- the relatively high hydrophobicities of NBFRs ( Table 1 ). In a study tant fate process for NBFRs in aquatic environments. of the Singapore urban watershed ( Wang and Kelly 2017 ), the field- The very limited number of studies on photodegradation has derived Log KOC values of NBFRs were also found to be generally suggested that several NBFRs have the potential to undergo consistent with the estimated Log KOC values. Hydrophobic com- degradation in the aqueous phase (Davis and Stapleton 2009, > pounds with Log KOC 4 are preferentially found in the organic Kajiwara et al. 2008, Zhang et al. 2016). Under natural sunlight matter in sediments, and therefore, these compounds have a high conditions, the estimated aqueous t1/2 of BTBPE and DPTE de- affinity for suspended and bottom sediments within aquatic envi- creased considerably and range between 1.5 and 17.1 days and ronments (see the estimated high emission t1/2 in sediments from 2.9-33.9 days, respectively (Zhang et al. 2016). Debromination and Table 3 ). In this review, we have also compiled the observed solid- ether bond cleavage (if possible) were determined to be the main water distribution coefficients (Log KSW , L/kg dw) of NBFRs from phototransformation pathways for the selected NBFRs (TBB, TBPH, all the available field data ( Table 3 and Table S6). The relationships BTBPE, and DPTE; Table S7), and it was concluded that similar between these Log KSW values and Log KOW values were not signif- degradation pathways can also occur for other NBFRs when con- < icant (P 0.001; Fig. 1-C), which may be attributed to the varied sidering their structural similarity. However, the estimated t1/2 val- organic carbon (OC) content found among the solid samples from ues of NBFRs in water ( Table 3 ) were only relevant in the upper various studies. Despite that, the high Log KSW values indicate that layer and clean water (without natural organic matter as an accel- suspended particles are important carriers of NBFRs in aquatic en- erator or inhibitor of photodegradation). The persistence of NBFRs vironments. against photodegradation depends on several conditions, such as Sediments are major sinks for NBFRs that have been trans- sunlight irradiation, the presence of oxidants and natural organic ported through the atmosphere, entered water bodies and were fi- matter, and the persistence of NBFRs in natural water bodies needs nally deposited in aquatic environments ( De Wit 2002 ); as a result, to be further studied. sediments have been widely viewed as useful environmental me- When compared with photochemical transformation, field and dia for the investigation of their pollution by NBFRs. In addition, laboratory studies have indicated that biodegradation by aerobic it can be suggested that a competition exists between the adsorp- or anaerobic microorganisms can be the more influential process tion of NBFRs on sediments and their bioaccumulation in aquatic driving the fate of halogenated pollutants in aquatic environments animals ( Wang and Kelly 2017 ). The affinity for sediments may ef- ( Chen et al. 2015 ). Although there has been no direct study con- fectively reduce the freely dissolved water concentrations of these cerning the microbial degradation of NBFRs in surface waters and NBFRs, thus reducing their bioavailability and bioaccumulation po- sediments, aerobic and anaerobic biodegradation of HBB, DPTE, tentials as revealed by the relatively low BSAFs from recent studies and TBECH has been observed in soil. Generally, the degrada- ( Fig. 2 ). tion of HBB was faster under aerobic conditions than in anaerobic The majority of NBFRs have relatively low volatilities, with es- soil, and TBECH was degraded quickly in both aerobic and anaer- timated vapor pressures between 6.35 ×10 −15 and 1.46 ×10 −3 mm obic soils ( Nyholm et al. 2010 ). DPTE and TBECH also reported

Hg at 25°C (Table 1). Thus, volatilization from water surfaces is poor biodegradabilities (t1/2 of 98 and 63-68 days, respectively) not expected to be an important fate process for these compounds. in soils from other anaerobic screening tests ( Wong et al. 2012 ). The fugacity ratios for penta-BT, HBB, and DPTE between water TBPH can be degraded by activated sludge under aerobic condi- and air (fW /fA ) were derived from field studies (Möller et al. 2011a, tions according to OECD Guideline 302 C, with a reported degra- Möller et al. 2011b ), with ratios ranging from < 0.01 to 1.52 dation of 7% within 28 days (ECHA 2020), whereas no biodegra- ( Table 3 ). These results suggest a net deposition of the three NBFRs dation was found for DBDPE or TBBPA-DBPE under the test con- from the atmosphere to seawater in the Arctic, Atlantic and remote ditions of OECD Guideline 302 C (ECHA 2020). The total degra- oceans ( Möller et al. 2011a , Möller et al. 2011b , Xie et al. 2011 ). The dation periods for other NBFRs were estimated to range from air-seawater exchange fluxes from recent studies have also demon- days to months by USEPA EPI suite ( Table 3 ). However, little strated high-magnitude deposition of NBFRs from the atmosphere is known about the possible biotransformation processes and no into the ocean ( Table 3 ), which indicates that seawater is an im- further analysis of the functional microbial organisms has been portant sink for NBFRs arriving after atmospheric transportation. done. Microbially mediated reductive dehalogenation is one of the However, no significant correlation ( P > 0.05) was found between most important routes of environmental transformation for per- the Log transformed vapor pressure ( Table 1 ) and the Log trans- sistent halogenated compounds such as PCBs, PBBs, and PBDEs formed NBFR concentrations in the atmosphere or surface waters ( Chen et al. 2015 ). Therefore, we suggest that the microbial dehalo- from the Arctic, Atlantic, and other oceans (Fig. S9). This may be genation pathways of NBFRs are strikingly similar to those of PB- due to variations of total suspended particulates among studies, as DEs and may be the most likely removal pathway of contaminants NBFRs proved to be mainly presented in particulate phase in at- from sediments. Future studies need to explore the degradation

11 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

)

) ) )

)

S5 S5

)

page

2010 2010 2010

2011a 2016 next 2012 Table Table 2011b

2012) 2012)

al. al. al. Stapleton

suite suite suite suite suite suite suite al.

in in al.

on al. al.

al. al.

et et et

et et et et

et et EPI EPI EPI EPI EPI EPI EPI 2020) and

)

detail detail

Zhang Davis Möller Nyholm Wong Nyholm Nyholm

continued

Möller

2009 (ECHA (Wang (ECHA 2020) USEPA USEPA USEPA USEPA ( (Wang ( USEPA ( ( ( ( USEPA ( USEPA see see References ( (

3 3

10 10

× ×

900 1.80 8.10 weeks 17-32 19-23 63-68 52.8 days- 32-42 86.4 2.200 4.71 2.39-4.43 TBECH

3 3 4

10 10 10

× × ×

0.001-0.017

4.32 8.64 3.89 weeks 63.6 98 0.02-1.52 2.9-33.9 2.650 2.20-3.0 3.62-3.85 ∞ < DPTE

3 3 4

10 10 10

× × ×

months 233 4.32 8.64 3.89 weeks- 9.301 3.14-4.34 1.20-3.76 ∞ PBEB

4 3 3 4

10 10 10 10

× × × ×

0.001-0.007

100-148 months 18-27 2.24 4.32 8.64 3.89 18-26 934 2.34-5.45 2.27-3.39 < ∞ HBB

3 3 3 4

10 10 10 10 × × × ×

0.001-0.35

months 4.32 8.64 3.89 2.56-3.84 57.8 1.62-3.85 weeks- 1.39 < ∞ penta-BT

3 3 4

10 10 10

× × ×

3.89 4.32 8.64 1.018 months 24.4 ∞ ∞ TBBPA-DBPE compartments.

3 3 4

10 10 10

× × ×

(7%

environmental

degradation) 3.23 28 1.44 2.88 1.30 29.2 0.490 weeks 0.153 4.18-5.60 11.8 TBPH various

3 3 4

10 10 10

× × ×

1.44 2.88 1.30 34.1 0.979 weeks 3.01-3.52 0.113 3.55 23.5 TBB NBFRs

for

3 3 4

10 10 10

× × ×

properties

0.720 months 4.32 8.64 3.89 3.68-5.0 3.18-6.36 1.5-17.1 17.3 BTBPE ∞

environmental 3 3 4

10 10 10

× × ×

4.466 2.80-5.19 0.167 2.93-4.05 0.021-0.042 8.64 4.32 107 3.89 DBDPE ∞ ∞ ∞ estimated

and

(day) (day)

(day) with

(L/kg 1/2

water (UV)

with with t

Arctic

50 by

day)

model)

(h)

soil

(day)

Ocean 7;

between degradation sludge sludge

soil soil

sludge

acid/water Guideline property

(day) (h)

(day)

surface WWTPs air

(pH (day) (h) investigated

(h)

weight) total ratio

fugacity

in in

Biowin Paciﬁc European of

hydrolysis hydrolysis photodegradation day)

3 distribution

water methanol methanol/water humic aerobic aerobic anaerobic soil air water soil sediments and

(OECD

dry

in in

SW SW

activated activated digested

K K weight) 3 for for for in in in in in in in in in in in in A A

AopWin f f

/ /

1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2 1/2

W W

t t Distribution f t t Phase Log (L/kg Log dry (Level Arctic HydroWin Photodegradation t by (day) time t Hydrosis t t water 302 t (day) t (day) Biodegradation expected t 0.5% 0.5% 0.5% t t t Fugacity water t biotransformation f Table Summary

12 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

) )

)

) )

2012 2011b 2011b 2011a 2012

Screening Screening Screening

al. al. al. al. al. 2011 2011 2020a 2020a

et et et et et al. al. al. al.

2020)

LRTP LRTP LRTP

et et et et & & &

Möller Liu Möller Xie Liu Möller Möller Xie Möller

Pov Tool ( (ECHA Pov ( ( Tool Pov Tool ( ( ( ( ( ( References

classiﬁed

1.01 not 111 0.425 TBECH

classiﬁed

0.01-1.04 0.01-94.5

0.4-4.0 0.030-1.608 1.0 1.41 0.005-0.225 3.10 not 518 2.89 < < DPTE

classiﬁed

3.63 not 515 7.02 PBEB

10 ×

classiﬁed

0.5 0.096-4.87 1.14 1.52 not 517 78.0 HBB

10 ×

classiﬁed

0.01-55.8

not 2.08 9.10 515 47.6 < penta-BT

10 ×

potential PBT/vPvB 2.86 519 12.7 TBBPA-DBPE

PBT/vPvB 2.86 potential 173 12.7 TBPH

10 ×

classiﬁed

not 2.23 7.54 173 TBB

10 ×

PBT/vPvB 2.83 12.4 potential 519 BTBPE

water.

degradation; to

air

10 ×

PBT/vPvB

from

10 ×

recalcitrant 0.1-3.10

2.86 2.28-9.12 potential 12.7 519 DBDPE <

was

deposition

∗

for

day)) day))

net Arctic

· ·

2 2

(day)

(%)

ECHA

and km) ﬂuxes ﬂuxes ) day))

the Southern

compound

by means

travel

exchange

(pg/(m (ng/(m Arctic Arctic

the and and

potential (CTD; Ocean Ocean Sea Sea day)) day)) day))

day)) day)) day))

sea sea

value eﬃciency

for · · ·

· · ·

persistence

continued

(ng/(m 2 2 2 2 2 2

(

potential deposition deposition 3

plus

European North European Paciﬁc North Bohai Atlantic Atlantic Bohai

a stands

Classiﬁcation distances in (ng/(m air-seawater in (pg/(m LRT transport characteristic in (pg/(m Net in in (ng/(m in (ng/(m in in in Dry transfer Southern (pg/(m Ocean overall ∗ Table ∞

13 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168 and transformation mechanisms of NBFRs in various environmental Stockholm Convention (which refers to half-lives of 60 days in wa- matrices. ter, 180 days in soil, or 180 days in sediments for persistence). The NBFRs of DBDPE, BTBPE, TBBPA-DBPE, HBB, PBEB, and DPTE can 4.3. Removal eﬃciency of NBFRs at WWTPs reach the persistence criteria. However, these results derived using chemical properties as model input should be interpreted with

The Log KSW values were estimated from the available stud- caution and future research must be pursued before the sources ies of NBFR levels in WWTPs ( Table 3 and Table S6). Generally, and LRT of NBFRs in remote environments are fully understood, in-

NBFRs with high Log KOW values tended to have high Log KSW cluding the relationship between sources and atmospheric concen- values ( P < 0.001; Fig. 1 -C), which was in accordance with their trations, the air-water exchange mechanism, key deposition pro- sorption-hydrophobicity relationship. Relatively high Log KSW val- cesses, and the contribution of various LRT pathways. ues in WWTPs were found for DBDPE, BTBPE, and TBPH (2.80-5.19, 3.68-5.00, and 4.18-5.60, respectively). 4.5. Classification A collection of the removal efficiencies of NBFRs from WWTPs worldwide is presented in Fig. S10. In WWTPs in Norway, Spain In the REACH database (ECHA 2020), the NBFRs of DBDPE, and China ( Arp et al. 2011 , Kim et al. 2014 , Li et al. 2018 , BTBPE, TBPH, and TBBPA-DBPE have been identified or classified as Nyholm et al. 2013 , Wang et al. 2020a ), NBFR concentrations in “potential PBT/vBvP” (which refers to their being under consider- influents (from several to tens of ng/L) were consistently higher ation for persistent, bioaccumulative and toxic/very persistent sta- than those in the effluent samples (one-way ANOVA, P < 0.05). tuses and having bioaccumulative properties), while other NBFRs Although with their known persistence abilities, all the inves- have not been classified as “potential PBT/vBvP” by the European tigated NBFRs (i.e., DBDPE, BTBPE, TBB, TBPH, HBB, TBEB, and Chemicals Agency (ECHA) Substance Infocard database ( Table 3 ). DPTE) have shown higher overall removal efficiencies in WWTPs in recent studies ( Cristale and Lacorte 2015 , Gewurtz et al. 2020 , 5. Risk assessment on aquatic environments Kim et al. 2014 , Li et al. 2018 , Wang et al. 2020a ). Approximately 73%-89% of the 19 NBFRs were removed via sorption on the 5.1. Ecological risk assessment sludge in WWTPs in Harbin ( Li et al. 2018 ), and DBDPE was al- most completely sequestered in the sludge of WWTPs in Stock- NBFRs are ubiquitous and highly bioaccumulative and pose holm ( Ricklund et al. 2008 ). In contrast, NBFR removal through potential ecological risks to aquatic animals. In addition to biodegradation can only account for 46%, 11%, 14%, and 16% of the the acute exposure and/or high-dose harmful effects of NBFRs overall removal of DPTE, BTBPE, TBPH, and DBDPE in the WWTPs that were explored recently, increasing evidence has revealed of Beijing, respectively ( Wang et al. 2020a ). Interestingly, the over- the long-term toxicological effects of the main NBFRs on all removal ratios of NBFRs were consistent with the removal of bioindicating aquatic organisms ( Dong et al. 2021 , Ezechiáš the total suspended solid, which was unaffected by the treatment et al. 2014 , Marteinson et al. 2021 , Ortega-Olvera et al. 2020 , processes used in WWTPs ( Arp et al. 2011 , Kim et al. 2014 ). Col- Xiong et al. 2019 ). DBDPE, BTBPE, TBB, TBPH, and TBECH can in- lectively, the data suggest that the sorption of NBFRs on sludge, duce oxidative stress ( De Jourdan et al. 2012 , Feng et al. 2013 , rather than their degradation processes, was the dominant mecha- Tomy et al. 2007 , USEPA 2021 ). All the investigated NBFRs can nism determining their fate in WWTPs. cause endocrine disorders and reproductive developmental toxicity Although low levels of NBFRs were detected in the sludge, the ( Bearr et al. 2010 , Feng et al. 2013 , Ma et al. 2018 , Nakari and Huh- estimated daily amount of NBFRs released from the WWTPs was tala 2010 , Tomy et al. 2007 , Usenko et al. 2016 , Wang et al. 2019 ). relatively high. According to the elution amount of sludge per day, DBDPE, TBB and TBPH drive abnormal behaviour and neurodevel- the discharge of the total NBFRs from the sludge of WWTPs was opment ( Jin et al. 2018 , McGee et al. 2013 ), and DBDPE, BTBPE and estimated to be 483.7 kg/year in Korea (sum of DBDPE and BTBPE) TBECH cause teratogenesis in fish embryos ( Giraudo et al. 2017 , ( Lee et al. 2014 ), 8.25 kg/year in Spain (sum of DBDPE, HBB and Pradhan et al. 2013 , Wang et al. 2019 ). PBEB) ( Gorga et al. 2013 ), 6.38 kg/year in Beijing, China (sum of The results from these toxicity tests can be used to estimate DBDPE, DPTE, BTBPE, and TBPH) ( Wang et al. 2020a ), 7.79 kg/year the safe concentrations of NBFRs and preliminarily determine their in Harbin, China (sum of DBDPE, BTBPE, and PBEB) ( Li et al. 2016a ), potential ecological risks for aquatic ecosystems (Table S8). The and 239.3 kg/year across all WWTPs in China (sum of DBDPE and predicted no effect concentration (PNEC) values for surface wa- μ μ BTBPE) (Zeng et al. 2014). Therefore, the WWTP sludge can be ter (PNECwater ; g/L) and sediments (PNECsediment ; g/g dw) can treated as a large reservoir for NBFRs with long-term persistence, be obtained based on the reported test endpoints ( EC 2003 ), such and the applications of sludge to soils in agricultural land may as the no observed effect concentration (NOEC) designation and have some adverse environmental outcomes. Future monitoring the half maximal effective concentration (EC50 ) value, and divided studies and environmental risk assessments are urgently needed into their specific assessment factors (Table S8), as described in the to provide more information on the potential risks of introducing Supporting Information (SI-1). Accordingly, the PNEC values of DB- NBFRs from sewage sludge into the aquatic environments. DPE, BTBPE, TBB, TBPH, TBBPA-DBPE, penta-BT, HBB and PBEB in surface water and sediments were calculated in this study (Table 4.4. Long-range transport (LRT) S8). However, it should be noted that due to a shortage of data on chronic and site-specific toxicities, the PNECs in this study were The persistence of NBFRs in aquatic environments can result not strictly extrapolated based on the species sensitivity distribu- in their long-range transport (LRT), which has been indicated by tion (SSD) method. Moreover, PNECs of TBECH and DPTE were not many studies carried out in remote and particularly vulnerable deduced due to the lack of aquatic toxicological data. ecosystems, such as in the arctic or pristine mountain zones (as we The global ecological risks of NBFRs for aquatic ecosystems discussed in 4.1). The LRT potential of NBFRs can be characterized were assessed using the risk quotient (RQ) method according to by their characteristic travel distance (CTD) and overall persistence the European Commission’s Technical Guidance document (TGD;

(POV ) (Table 3), which can be estimated by the “OECD POV and LRTP EC (2003)) and previous studies (Bu et al. 2019, Chen et al. 2018, Screening Tool” ( Wegmann et al. 2009 ). Matthies et al. (2009) pro- Liu et al. 2020b ). The results of the risk assessment of NBFRs posed that a POV boundary of 230 days or a LRT potential boundary for surface water and sediments across the world are shown in of 5200 km can be applied as persistence criterion based on the Fig. 3 . There were only a few site-speciﬁc instances where the

14 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

Fig. 3. Global distribution of maximum risk quotients (RQs) for NBFRs in surface water and sediments (A); and box plot diagrams for estimated RQs in both surface water (B; empty box) and sediments (C; shadow box) from various locations with at least low risks (RQ > 0.01). Detailed RQs from all cases are compiled in Fig. S11 and S12. environmental concentration of NBFRs exceeded their PNECs. The 5.2. Human health risk assessment maximum measured DBDPE concentrations (MEC) in the Bohai Sea adjacent to the mariculture zone had exceeded its predicted The toxic effects of NBFRs on humans have not been thoroughly PNEC water , indicating the high risks posed by DBDPE. BTBPE and investigated ( Dong et al. 2021 , Harju et al. 2008 ). NBFRs are re- HBB posed significant high risks (mean RQ > 1) to aquatic organ- ported to elicit adverse effects in rodents from in vivo and in isms in sediments from e-waste sites of Guangdong, China (Fig. vitro studies, and the most concerned health toxic endpoints for S12). In addition, BTBPE also posed medium risks (maximum RQ > NBFRs are the effects on histopathology change, reproduction and 0.01) to the relevant sensitive aquatic organisms in the sediments the associated hormonal disruption ( Dong et al. 2021 , Ezechiáš from industrial area of South Africa, Pearl River estuary and Daya et al. 2014 , Marteinson et al. 2021 , Xiong et al. 2019 ). Since NBFRs Bay from China. The high RQs of TBPH were exhibited in the sed- have not been classified as carcinogenic, their oral carcinogenic risk iments from Durben Bay of South Africa. These results indicated (CR) cannot be assessed in this study. Among the studied 10 NBFR that the greatest potential ecological risks posed by NBFRs exhib- congeners, only DBDPE and HBB have undergone a complete eval- ited regional variability to some extent. The investigated 7 NBFRs uation and determination under the US EPA’s Integrated Risk In- mostly posed insignificant risks when considering their average formation System (IRIS) program for evidence of human noncar- concentrations in the surface water and sediments in the receiv- cinogenic risk (Table S9). For other NBFRs, RfDs from references ing environments. were adapted in this study, which were calculated by applying the

15 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

EPA’s maximum recommended composite uncertainty factors for (2) Knowledge of environmental behavior and the fate of NBFRs the lowest reported NOAEL (Table S9). should be improved. Studies addressing the solid/water partition- Previous studies have shown that the consumption of fish is ing or air-water exchange, mechanism of long-range transport, and a significant contributor to the human intake of organic pollu- extent of LRT are needed to understand the environmental behav- tants from aquatic environments, especially for BFRs with high ior of NBFRs. Experiments are also needed to determine the half- accumulation patterns in fish ( Lyche et al. 2015 , Ma et al. 2020 , life of NBFRs in natural water bodies and sediments to elucidate Nøstbakken et al. 2018 ). To comprehensively evaluate the risk of their mechanisms of biodegradation in aquatic environments and exposure to NBFRs through fish consumption, the potential non- to identify their abiotic and biotic degradation pathways. Multime- carcinogenic risk was estimated using the hazard quotient (HQ), dia environmental transport and exposure models established from which refers to the ratio of the estimated daily intake value (EDI) monitoring data can provide an overall picture of the exposure of to the RfD (SI-2; Table S10). Considering a worst-case scenario of aquatic environments and humans to NBFRs. exposure, the maximum reported NBFR concentrations in fish sam- (3) Research is needed to better elucidate the toxicokinetics of ples in the previously published literature were used as MEC. As a NBFRs in aquatic organisms, which can provide an interpretation result, the ƩHQs of NBFRs through fish consumption were in the of the external to internal exposure for consideration in toxico- range of no risk ( ≤ 0.041) for the general population across the logical studies and risk assessments, especially for NBFRs classi- different study regions (Table S11). A relatively serious risk was fied as “VB” in this study. This should include a variety of fields observed in Guangdong Province of China, which is an important and include studies on uptake and elimination kinetics, tissue dis- industrial base for manufacturing and recycling of FR-containing tribution patterns, metabolic pathways, and species-specific and products. In the vast area of the Great Lakes region, Germany, and structure-specific metabolic susceptibility in aquatic organisms. In the coastal area of South Africa, the total intake values of NBFRs via addition, there is still a lack of concise conclusions on the trophic fish exceeded 7.45 ×10 −4 , 1.52 ×10 −3 and 0.022 μg/kg body weight transfer processes and trophodynamics of NBFRs in the aquatic (bw)/day, respectively. However, the ƩHQs were far less than 1 in food web, which should be given more focus in the future. all the study regions, indicating that no immediate noncarcino- (4) It should be noted that transformation products or metabo- genic risk for NBFRs occurred though fish consumption. It should lites can behave differently in terms of environmental behavior be noted that the calculated ƩHQs might be underestimated be- and toxicity compared to their parent compounds. Novel analyti- cause not all the NBFR concentrations in fish were available as can- cal and modeling methods should be established to simultaneously didates to generate the ƩHQ values. The actual risk may be higher determine the NBFRs present and identify their major degrada- when considering the presence of all NBFR components. In ad- tion/metabolism products. Further investigation of the distribution dition, considering the gradual replacement of NBFRs with tradi- of these products and toxicological studies on their adverse effects tional FRs, their environmental and health risks will likely increase are required and should be performed. in the future. (5) At present, with the limited information available on the occurrence and toxicity of NBFRs, it is still impossible to perform a 6. Conclusions comprehensive risk assessment for all the NBFRs. With the gradual increase in the availability in monitoring and investigation data, This paper provides an overall picture of the occurrence, a higher-tier quantitative probabilistic risk assessment using the fate, bioaccumulation, and potential ecological and health-related joint probability curve (JPC) method would be helpful to rigorously risks of NBFRs in aquatic environments. The emission source of characterize their risks and screen the priority NBFR analogs for NBFRs greatly affects their distribution in aquatic environmental high aquatic ecological risk at the national scale. Future compre- compartments and aquatic animals. It is worth noting that the con- hensive health risks of NBFR mixtures, which collect the routes of centrations of some NBFRs in coastal sediments from source areas water drinking, ingestion of various aquatic animals and dermal in- can reach or exceed the PNEC, and there are obviously increas- take, should be assessed to more accurately reflect the systematic ing trends for NBFR levels in sediments worldwide. The physico- risks of NBFRs to aquatic environments. chemical properties of NBFRs have a great influence on their environmental behaviors, fates and bioaccumulation potentials. The Declaration of Competing Interest

NBFR congeners with high Log KOW values tend to be distributed in particles rather than in the aqueous phase. When compared with The authors declare that there are no conﬂicts of interest re- degradation (abiotic and biotic), the partitioning of NBFRs caused garding the publication of this paper. by their hydrophobicities more closely determined their fates in aquatic environments. There is evidence that NBFRs can be highly bioaccumulated or even transferred through the aquatic food web, Acknowledgment but the bioaccumulation potential of NBFRs varies among species.

However, the existing data are not suﬃcient to draw thorough con- This work was supported jointly by the National Natural Science clusions regarding PBT potentials, precise risk assessments, and ﬁ- Foundation of China (Nos. 41907339; 41876129), Natural Science nal proposals on prohibitions or restrictions. Foundation of Guangdong Province (2018A030313136), National

Future studies are needed in a variety of ﬁelds, with a focus on Science & Technology, Fundamental Resources Investigation Pro- contaminant monitoring, environmental behaviours, toxicokinetics, gram of China (2018FY100104), Innovation Academy of South China biodegradation processes, and ecological risk assessments: Sea Ecology and Environmental Engineering, Chinese Academy

(1) In past studies, NBFRs have rarely been investigated as a of Sciences (ISEE2019ZR03, ISEE2018PY03, ISEE2018ZD02), Key whole class of pollutants, which had limited the simultaneous Special Project for Introduced Talents Team of Southern Ma- recognition of target NBFR congers. Therefore, it is diﬃcult to ana- rine Science and Engineering, Guangdong Laboratory (Guangzhou) lyze the spatial and temporal differences in NBFR distribution pat- (GML2019ZD0404). terns in the global environment. It is then necessary to carry out further monitoring studies of NBFRs to better understand the in- Supplementary materials ﬂuence of their sources, environmental conditions, and physico- chemical properties on the occurrence of these compounds in Supplementary material associated with this article can be aquatic environments. found, in the online version, at doi: 10.1016/j.watres.2021.117168 .

16 R. Hou, L. Lin, H. Li et al. Water Research 198 (2021) 117168

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