The metabolism of novel flame retardants and the internal exposure and toxicity of their major metabolites in fauna - a review
0Abstract
The worldwide production and usage of novel flame retardants increase their exposure to non-human fauna. Animals can accumulate and metabolize these novel flame retardants including novel halogenated flame retardants (NHFRs) and organophosphate flame retardants (OPFRs), which is of considerable significance to their internal exposure and final toxicities. In this review, recent studies on the metabolic pathways and kinetics of the two classes of novel flame retardants and the internal exposure and toxicity of their major metabolites are summarized. The results showed that the metabolic pathways of OPFRs were similar among various animals, while the metabolism kinetics (or toxicokinetics) were variable among species. O-dealkylation, hydroxylation and phase II conjunction were the most likely pathways for OPFRs. NHFRs might be metabolized through the pathways of debromination, hydroxylation, dealkylation, and phase II conjunction. We also suggested that di-alkyl phosphates (DAPs) and hydroxylated OPFRs (OH-OPFRs) were the predominant metabolites in the animal body. DAPs, 2,3,4,5-tetrabromobenzoic acid (TBBA) and 2-ethylhexyl tetrabromophthalate (TBMEHP) have relatively higher internal exposure levels in fauna, which might attribute to their high conversion rate and stability in the body. The metabolism of OPFRs and NHFRs in non-human animals may eliminate their acute toxicity but not their chronic toxicities (especially for endocrine-disrupting effects), which suggests attention should also be paid to the major metabolites. Based on the issues mentioned above, we proposed that the metabolic processes in multitrophic organisms, the transfer of major metabolites across the food web, and the co-exposure of the novel flame retardants and their metabolites in fauna are worth studying in the future.
Keywords
INTRODUCTION
In recent years, the use of traditional brominated flame retardants such as polybrominated diphenyl ethers (PBDEs), tetrabromobisphenol A (TBBPA), and hexabromocyclododecanes (HBCDs) has been restricted or prohibited[1]. As a result, novel halogenated flame retardants (NHFRs) and organophosphate flame retardants (OPFRs) have been increasingly used as substitutes in plastics, lubricants, rubber products, electronic equipment, furniture, food packaging, and other products[2-4]. Scholars have recently defined newly produced or newly detected brominated flame retardants as NHFRs[2,5-7]. The most representative NHFRs are 2-ethylhexyl tetrabromobenzoic acid (TBB), decabromodiphenyl ethane (DBDPE), and 1,2-bis (2,4,6-tribromophenoxy) ethane (BTBPE). Organophosphate flame retardants (OPFRs) have also been widely used as another kind of substitute in recent years[4]. The annual global production and usage of these novel flame retardants have also been growing rapidly in recent decades[8,9]. According to a research report from Ceresana, the global demand for flame retardants in 2018 was approximately 2.26 million tons, with brominated flame retardants (BFRs) and OPFRs accounting for 29% and 18% of the flame retardants used in the Asia Pacific region, respectively[10].
Similar to PBDEs, NHFRs have a stable brominated benzene ring structure, low solubility in water, and durability to physical, chemical, or biological degradation[9,11]. OPFRs can be divided into chlorinated (Cl-OPFRs), alkyl substituted (alkyl-OPFRs), and aryl substituted (aryl-OPFRs), according to the different ester bonds of substituents. Among them, Cl-OPFRs are more resistant to photolysis, chemical decomposition, and microbial degradation[4,12]. As a series of non-reactive additives[8,9], NHFRs and OPFRs can easily escape from the products, and distribute in various environmental matrices, such as indoor dust[13,14], atmosphere[15,16], soil[17,18], surface water[19-23], groundwater[24], and sediments[7,25], and enter into wastewater treatment plants[26-30]. With the extensive usage of new flame retardants, an increasing number of studies have gradually focused on the bioaccumulation, toxicity mechanism, and ecological risks of these pollutants.
Due to their lipophilicity, NHFRs and OPFRs can accumulate in various aquatic organisms[31-38]. Relatively higher concentrations of NHFRs and OPFRs have been detected in marine invertebrates, fish, marine mammals, and other biological samples (up to mg/g level by lipid weight), which were close to or even higher than those for traditional flame retardants (such as PBDEs and HBCDs)[11,20,39,40]. In addition, these novel flame retardants can be effectively transferred across the food chain/web and have shown potential biomagnification effects, for example, the NHFRs in food webs from the Bohai Sea, South China Sea, and Taihu Lake[41-44] and the OPFRs in food webs from the Laizhou Bay, South China Sea, and Taihu Lake[39,45,46]. Ecotoxicological studies have verified acute and chronic toxicity[47,48], reproductive toxicity[49,50], developmental toxicity[51-53], neurotoxicity[54-56], and endocrine-disrupting effects for several OPFRs[52,57,58]. The toxicological profile of NHFRs has been characterized for animals and humans[59], e.g., direct neurotoxicity, endocrine-related effects including dioxin-like effects, agonistic activity, steroidogenesis, estrogenic activity, disruption of the neuroendocrine system, reproductive developmental toxicity, hepatotoxicity, and cytogenotoxicity[9,59,60].
Toxicokinetics is of particular relevance for understanding pollutant accumulation and toxicity within an organism, which determines the relationship between external exposure and internal exposure[61]. The metabolism of pollutants in organisms leads to the formation of products with different toxicities to their parent, which results in variations in their biological toxicity. In addition, novel FRs and metabolites share similar structures and might exhibit combined toxicity to organisms[62]. Therefore, the metabolism of novel FRs and the body burden of their metabolites were both important to reflect their actual risks to fauna. Several recent reviews have summarized the production, physicochemical properties, usage, environmental occurrence, analytical methods, bioaccumulation, human exposure, and toxicities of novel FRs[2,3,6,8,9,13,18,59,63-67]. However, very few studies have reviewed the mechanisms and kinetics of the metabolism of novel FRs in various organisms. Our previous two reviews of novel FRs only partly focused on metabolic processes[11,68]. Smythe et al. reviewed the biotransformation processes of FRs, but only BFRs were considered[69]. Another transformation review only provided information specific to the plant accumulation and transformation of the novel FRs[70]. A review by Yang et al. only provided information specific to the human internal exposure and health risks of OPFRs and their metabolites[71]. Accordingly, this review aims to summarize all of the published studies on the animal-mediated metabolism of NBFR and OPFRs, to compare the compound-specific metabolism pathways of these novel FRs, and to systematically collect the internal exposure results of the major metabolites in fauna. In addition, this study proposed the current and key knowledge gaps and research needs for future research on novel FR biomonitoring.
METHODOLOGY
Systematic searches covering the period from 1966 to 2023 were conducted on Web of Science and Google Scholar using the keywords of BFRs, OPFRs, organophosphorus esters (OPEs), or dechlorane plus (DPs) and keywords of metabolism, biotransformation, metabolites, toxicity, or internal exposure. The retrieved literature was carefully checked, and peer-reviewed studies related to non-human animals were selected. A total of 69 publications were finally selected and included in the review [Table 1].
Summary of studies on metabolism of novel FRs in non-human fauna
| No. | Locations | Species studied | Compounds | Studied area | Reference |
| Field study | |||||
| 1 | Great Lakes, USA | Herring gull egg | TNBP, TBOEP, TPHP, TDCPP, and TCPP | In vitro transformation pathway, kinetics, and metabolites formation | [31] |
| 2 | Great Lakes, USA | Bald eagle eggs | TBBA and TBMEHP | Internal exposure | [96] |
| 3 | Lake Huron, Canada | Herring gull plasma | BCPP, BDCPP, BBOEP, DNBP, DEHP, and DPHP | Internal exposure | [124] |
| 4 | Taihu Lake, China | Freshwater fish liver microsome | TCEP, TCPP, TDCPP, TIBP, TPHP, TCP, and EHDPHP | In vitro transformation kinetics | [39] |
| 5 | Taihu Lake, China | Freshwater fish liver microsome | ATE, BTBPE, TBPH, PBBA, TBCT, DBDPE, and TBECH | In vitro transformation kinetics | [41] |
| 6 | Troutman Lake, Austria | Stickleback | BCEP, DNBP, and DPHP | Internal exposure | [143] |
| 7 | Rivers in Beijing, China | Topmouth gudgeon (Pseudorasbora parva), crucian carp (Carassius auratus), and loach (Misgurnus anguillicaudatus) | BBOEP, DNBP, DEHP, and DPHP | Internal exposure | [38] |
| 8 | E-waste dismantling site in Guangdong, China | Chinese water snake (Enhydris chinensis), snake egg, and commo carp | BCPP, DNBP, DPHP, BBOEP, BCIPHIPP, and EHPHP, BBOEHEP, OH-TBOEP, OH-TPHP, 5-OH-EHDPHP | Internal exposure | [42] |
| 9 | South China Sea | Marine fish liver microsome | TBECH, PBT, PBP, TBB, HBB, TBPH, DBDPE, and TBBPA-DBPE | In vitro transformation kinetics | [44] |
| 10 | Costal area of Korea | Marine fish liver microsome | BTBPE, HBB, PBEB, PBT, TBB, and TBCT | In vitro transformation kinetics | [111] |
| 11 | Arctic sea | Marine mammal liver microsomes | DBDPE | In vitro transformation kinetics, and metabolites formation | [101] |
| 12 | East Greenland | Liver microsomes of polar bears and ringed seals | TNBP, TBOEP, TPHP, TDCPP, and TEHP | In vitro transformation pathway, kinetics, and metabolites formation | [144] |
| 13 | Pearl river estuary, China | Marine food web | BBOEP, DNBP, DPHP, BBOEHEP, OH-TBOEP, and OH-TNBP | Internal exposure | [125] |
| 14 | Across the globe | Fishmeal | BCEP, BDCPP, DMP, DPHP, DNBP, and DEHP | Internal exposure | [127] |
| 15 | Tarragona, Spain | Seafood species | BCEP, DPHP, DNBP, BDCPP, BBOEP, and DEHP | Internal exposure | [145] |
| 16 | Australia | Egg | BCEP, BCPP, BDCPP, DNBP, DEHP, BBOEP, and DCP | Internal exposure | [129] |
| 17 | 30 countries | Cow milk | BCPP, DPHP, BDCPP, BBOEP, DCP, DNBP, BBOEHEP, and OH-BBOEP | Internal exposure | [126] |
| 18 | Beijing, China | Cow milk | BCPP, BDCPP, BBOEP, DNBP, DPHP, and DCP | Internal exposure | [146] |
| 19 | China | Meat meal, feather meal, and blood meal | BCEP BCPP, BDCPP, BBOEP, DNBP, DCP, DEHP, and DPHP | Internal exposure | [131] |
| 20 | Chengdu, China | Chickens, ducks, pigs, cattle, sheep, fish, and shrimp | BCEP, BCPP, BDCPP, DPHP, BBOEP, DNBP, and DEHP | Internal exposure | [130] |
| 21 | Southeast Queensland, Australia | Meat, fish, seafood, and egg | BCEP, BCPP, BDCPP, DNBP, DEHP, BBOEP, and DCP | Internal exposure | [128] |
| Laboratory study | |||||
| 1 | - | Embryonated eggs and chicks of Japanese quail | TPHP | In ovo transformation kinetics and metabolites formation | [73] |
| 2 | - | Embryonated eggs of Japanese quail | TDCPP and DPs | In ovo transformation kinetics and metabolites formation | [74] |
| 3 | - | American kestrel (Falco sparverius) egg | TBBPA-DBPE and BTPBE | In ovo transformation kinetics | [98] |
| 4 | - | Laying hens and egg | TCPP, TPHP, TNBP, TBOEP, and TEHP | In vivo transformation kinetics and metabolites formation | [76] |
| 5 | - | Chicken embryos | TCPP and TDCPP | In vivo transformation kinetics | [147] |
| 6 | - | Chicken embryos | DPs | In vivo transformation kinetics | [97] |
| 7 | - | Chicken embryo | TDCPP | In vitro transformation kinetics and metabolites formation | [75] |
| 8 | - | Bird and rat liver microsomes | BPA-BDP | In vitro transformation pathway, kinetics, and metabolites formation | [77] |
| 9 | - | Zebrafish | TPHP | In vivo transformation pathway, kinetics, and metabolites formation | [78] |
| 10 | - | Zebrafish | TPRP, TNBP, TBOEP, TCEP, TDCPP, and TCP | In vivo transformation pathway and metabolites formation | [148] |
| 11 | - | Zebrafish | EHDPHP | In vivo transformation pathway and metabolites formation | [84] |
| 12 | - | Zebrafish | TBECH and TBP | In vivo transformation pathway and kinetics | [113] |
| 13 | - | Zebrafish | PBT, HBB, BTBPE, and DBDPE | In vivo transformation kinetics | [105] |
| 14 | - | Zebrafish | DBDPE | In vivo transformation pathway and kinetics | [102] |
| 15 | - | Zebrafish | DBDPE | In vivo transformation pathway and kinetics | [103] |
| 16 | - | Chinese rare minnow | TNBP, TBOEP | In vivo transformation kinetics and metabolites formation | [82] |
| 17 | - | Chinese rare minnow | TEHP | In vivo transformation pathway, kinetics, and metabolites formation | [81] |
| 18 | - | Rainbow trout (Oncorhynchus mykiss) | BTBPE and TBPH | In vivo transformation kinetics | [149] |
| 19 | - | Rainbow trout (Oncorhynchus mykiss) | DPs | In vivo transformation kinetics | [109] |
| 20 | - | Rainbow trout (Oncorhynchus mykiss) | BTBPE | In vivo transformation kinetics | [112] |
| 21 | - | Rainbow trout (Oncorhynchus mykiss) liver microsome | TBBA | In vitro transformation pathway | [104] |
| 22 | - | Crucian carp | TNBP, TBOEP | In vitro transformation kinetics and metabolites formation | [80] |
| 23 | - | Crucian carp | CDP | In vitro transformation kinetics | [85] |
| 24 | - | Common carp | TCEP, TNBP, TBOEP, TCIPP, TDCPP, TPHP, and EHDPHP | In vivo transformation pathway, kinetics, and metabolites formation | [83] |
| 25 | - | Common carp | DPs | In vivo transformation kinetics | [108] |
| 26 | - | Fathead minnows (Pimephales promelas) | BTBPE, TBBPA-DBPE, TBPH, and TBB | In vivo transformation kinetics | [106] |
| 27 | - | Killifish (Fundulus heteroclitus) | TBPH | In vivo transformation kinetics | [114] |
| 28 | - | Redtail catfish and oscar fish | DPs | In vivo transformation kinetics | [110] |
| 29 | - | White rat | TEHP | In vivo transformation pathway and kinetics | [150] |
| 30 | - | Rat | TPHP | In vivo transformation pathway and metabolites formation | [134] |
| 31 | - | Rat | TCEP, TCPP, TDCPP, TCP, TPHP, and TNBP | In vivo transformation pathway and kinetics | [87] |
| 36 | - | Rat | BTBPE | In vivo transformation pathway and kinetics | [116] |
| 37 | - | Rat | DBDPE | In vivo transformation pathway and kinetics | [115] |
| 38 | - | Rat | TBB and TBPH | In vivo transformation pathway and kinetics | [120] |
| 39 | - | Rat | TBB and TBPH | In vivo transformation pathway, kinetics, and metabolites formation | [117] |
| 40 | - | Rat | TBPH | In vivo transformation kinetics and metabolites formation | [119] |
| 41 | - | Rat liver microsome | TPHP and TDCPP | In vitro transformation kinetics and enzyme mechanisms | [88] |
| 42 | - | Rat liver and intestinal subcellular fractions | TBB, TBPH | In vitro transformation pathway, kinetics, and metabolites formation | [118] |
| 43 | - | Earthworm (Eisenia fetida) | TNBP | In vivo transformation pathway and kinetics | [93] |
| 44 | - | Earthworm (Eisenia fetida) | TPHP | In vivo transformation pathway and kinetics | [94] |
| 45 | - | Earthworm (Eisenia fetida) | TBOEP | In vivo transformation pathway and kinetics | [95] |
| 46 | - | Earthworm (Eisenia fetida) | PBT, HBB, BTBPE, and DBDPE | In vivo transformation kinetics | [123] |
| 47 | - | Mudsnails (Bellamya aeruginosa) | PBT, HBB, and DBDPE | In vivo transformation pathway and kinetics | [121] |
| 48 | - | Marine mussel | TCP, TNBP, TBOEP, TPHP, TCPP, and EHDPHP | In vivo transformation kinetics | [92] |
| 49 | - | Clam (Corbicula fluminea) | PBT, HBB, BTBPE, and DBDPE | In vivo transformation kinetics | [122] |
| 50 | - | Daphnia magna | TPHP | In vivo transformation pathway and kinetics | [89] |
| 51 | - | Daphnia magna | TBOEP, TCEP, TDCPP, and TPHP | In vivo transformation kinetics | [91] |
| 52 | - | Invertebrates (Daphnia magna) and fish (Oryzias latipes) | TPHP | In vivo transformation pathway and kinetics | [90] |
In this review, three kinds of typical OPFRs were included, such as Cl-OPFRs [tris(2-chloroethyl) phosphate (TCEP), tris(2-chloroiso-propyl) phosphate (TCPP), and tris(2-chlorol-chloromethy) phosphate (TDCPP)], four alkyl-OPFRs [tributyl phosphate (TNBP), tris(2-butoxyethyl) phosphate (TBOEP), tri(2-ethylhexyl) phosphate (TEHP), and tripropyl phosphate (TPRP)], and five aryl-OPFRs [tripheny phosphate (TPHP), tricresyl phosphate (or so-called tris(methylphenyl) phosphate) (TCP or so-called TMPP), cresyl diphenyl phosphate (CDP), 2-ethylhexyl diphenyl phosphate (EHDPHP), and bisphenol A bis (diphenylphosphate) (BPA-BDP)]. NHFRs in this review are divided into the monoaromatic NHFRs [TBB, tis(2-ethylhexyl)-2,3,4,5-tetrabromophtalate (TBPH), pentabromotoluene (PBT), pentabromophenol (PBP), hexabromobenzene (HBB), pentabromoethylbenzene (PBEB), 2,3,4,5-tetrabromo-6-chlorotoluene (TBCT), tribromophenol (TBP), 2,4,6-tribromophenyl allyl ether (ATE), and pentabromobenzyl acrylate (PBBA)], polyaromatic NHFRs [DBDPE, BTBPE, tetrabromobisphenol A-bis(2,3-dibromopropylether) (TBBPA-DBPE)], naphthenic NHFRs [tetrabromoethylcyclohexane (TBECH)], and DPs.
In addition, in silico analysis was used for preliminary bioaccumulation and toxicity assessments of the major metabolites of novel FRs. The Log KOW and BAF values were predicted for all the major metabolites using USEPA EPI suit v4.1. The EPA T.E.S.T., distributed by the EPA, was applied to estimate acute and chronic toxicities. For acute toxicity data, fathead minnow (96 h), Daphnia magna (48 h), and T. pyriformis (48 h) were considered based on the LC50. Developmental toxicity and mutagenicity were selected as the endpoints for chronic toxicity.
METABOLIC TRANSFORMATION PROCESS OF NOVEL FLAME RETARDANTS
OPFRs
Birds
Studies on the metabolic transformation of OPFRs in avian species are mainly conducted by in vitro (i.e., liver microsome experiment) and in ovo methods (i.e., egg exposure experiment) [Table 2]. An in vitro study using liver microsomes of herring gulls from the Great Lakes found a general metabolic pathway for OPFRs in forming their respective di-alkyl phosphates (DAPs)[31]. The O-dealkylation pathway was confirmed for TPHP in vitro in chicken embryonic hepatocytes[72] and in ovo in embryonated eggs and chicks of Japanese quail[73], where this pathway was suggested to depend on cytochrome P450 (CYP) enzymes. Briels et al. also showed the formation of bis(1,3-dichloropropyl) phosphate (BDCPP) in the embryo of Japanese quail during in ovo exposure with TDCPP[74]. An efficient transformation from TDCPP to BDCPP was found in chicken embryonic hepatocytes with a molar conversion ratio of 1:1, indicating the significance of O-dealkylation in the metabolism of Cl-OPFRs[75]. BDCPP could not be metabolized further in chicken embryonic hepatocytes after 36 h of exposure[75].
Available information on metabolism pathways and toxicokinetics of OPFRs in non-human fauna
| Compounds | Species/assays | Methods | Metabolism pathway | Major metabolites | Available toxicokinetic constants | References |
| TCEP | Laying hens | In vivo | O-dealkylation | BCEP | 22.6 d (t1/2) | [76] |
| Zebrafish | In vivo | O-dealkylation and hydroxylation | BCEP and OH-BCEP | - | [79] | |
| Common fish | In vivo | - | - | 9.2-18.3 h (t1/2) | [83] | |
| Liver microsomes of yellow catfish, catfish and crucian carp | in vitro | - | - | 0.50-1.10 mL/min/mg protein | [39] | |
| Daphnia magna | In vivo | - | - | 4.13-6.03 h (t1/2 for waterborne exposure) | [91] | |
| TCPP | Laying hens | In vivo | O-dealkylation | 30.1 d (t1/2) | [76] | |
| Herring gull liver microsome | In vitro | O-dealkylation | BCPP | 27 ± 1 | [31] | |
| Catfish liver microsome | In vitro | - | - | 1.33 mL/min/mg protein | [39] | |
| Common fish | In vivo | O-dealkylation and hydroxylation | BCIPP and BCIPHIPP | 10.5-14.5 h (t1/2) | [83] | |
| TDCPP | Herring gull liver microsome | In vitro | O-dealkylation | BDCPP | 8 ± 1 mL/min/mg protein | [31] |
| Embryonated eggs of Japanese Quail | In vitro | O-dealkylation | BDCPP | - | [74] | |
| Chicken embryo | In vitro | O-dealkylation | BDCPP | - | [75] | |
| Zebrafish | In vivo | O-dealkylation | BDCPP | - | [79] | |
| Common fish | In vivo | Hydroxylation | BDCPP and OH-BDCPP | 9.4-19.8 h (t1/2) | [83] | |
| Liver microsomes of yellow catfish, catfish and crucian carp | In vitro | - | - | 0.944-0.778 mL/min/mg protein | [39] | |
| Rat | In vivo | Glutathione conjugation | GSH-TDCPP | - | [151] | |
| Rat | In vivo | O-dealkylation | BDCPP | - | [152] | |
| Rat liver microsome | In vitro | O-dealkylation | 1,3-dichloro-2-propanol, 3-chloro-1,2-propanediol, and BDCPP | - | [151] | |
| Rat liver microsome | In vitro | - | - | 1.8083 h (t1/2) | [88] | |
| Daphnia magna | In vivo | - | - | 4.36-6.60 h (t1/2 for waterborne exposure) | [91] | |
| TNBP | Laying hens | In vivo | O-dealkylation | 82.5 d (t1/2) | [76] | |
| Herring gull liver microsome | In vitro | O-dealkylation | DNBP | 73 ± 4 mL/min/mg protein | [31] | |
| Marine mammal liver microsome | In vitro | O-dealkylation | DNBP | - | [73] | |
| Zebrafish | In vivo | O-dealkylation, hydroxylation, and GLU conjugation | DNBP, OH-TNBP, and GLU-TNBP | - | [79] | |
| Rare minnow | In vivo | O-dealkylation and hydroxylation | DNBP and OH-TNBP | 0.6-2.0 d (t1/2) | [82] | |
| Crucian carp liver microsomes | In vitro | O-dealkylation and hydroxylation | DNBP and OH-TNBP | 3.1 mL/min/mg protein | [80] | |
| Common fish | In vivo | O-dealkylation and hydroxylation | DNBP | 8.8-15.9 h (t1/2) | [83] | |
| Liver microsomes of yellow catfish, catfish and crucian carp | In vitro | - | - | 0.74-1.17 mL/min/mg protein | [39] | |
| Mice | In vivo | O-dealkylation | DNBP | - | [87] | |
| Marine mussel | In vivo | - | - | 1.93 d (t1/2) | [92] | |
| Earthworm | In vivo | O-dealkylation, hydroxylation, ethylene glycol conjugation, sulfation, and phosphate conjugation | DNBP, OH-TNBP, PA-TNBP, DNBHEP, SUL-TPHP, and GLU-TPHP | - | [93] | |
| TBOEP | Laying hens | In vivo | O-dealkylation | BBOEP | 11.3 d (t1/2) | [76] |
| Herring gull liver microsome | In vitro | O-dealkylation | BBOEP | 53 ± 8 mL/min/mg protein | [31] | |
| Marine mammal liver microsome | In vitro | O-dealkylation | BBOEP | - | [73] | |
| Zebrafish | In vivo | O-dealkylation, hydroxylation, and GLU conjugation | BBOEP, BOEHEP, BBOEHEP, GLU-TBOEP, and GLU-BBOEHEP | - | [79] | |
| Rare minnow | In vivo | O-dealkylation and hydroxylation | BBOEHEP, BBOEP, and OH-TBOEP | 0.7-2.3 d (t1/2) | [82] | |
| Crucian carp liver microsomes | In vitro | O-dealkylation and hydroxylation | BBOEHEP, BBOEP, and OH-TBOEP | 3.9 mL/min/mg protein | [80] | |
| Common fish | In vivo | O-dealkylation and hydroxylation | BBOEHEP, BBOEP, and OH-TBOEP | 10.5-17 h (t1/2) | [83] | |
| Daphnia magna | In vivo | - | - | 4.28-5.33 h (t1/2 for waterborne exposure) | [91] | |
| Earthworm | In vivo | O-dealkylation and hydroxylation | BBOEP, BOEHEP, BBOEHEP, OH-TBOEP, etc. | - | [95] | |
| TEHP | Laying hens | In vivo | O-dealkylation | DEHP | 43.3 d (t1/2) | [76] |
| Marine mammal liver microsome | In vitro | O-dealkylation | DEHP | - | [144] | |
| Rare minnow | In vivo | O-dealkylation, hydroxylation, and GLU conjugation | DEHP, OH-TEHP, and GLU-TEHP | 1-2.57 d (t1/2) | [81] | |
| TPRP | Zebrafish | In vivo | O-dealkylation, hydroxylation, and GLU conjugation | DPRP, OH-DPRP, and GLU-TPRP | - | [79] |
| TPHP | Laying hens | In vivo | O-dealkylation | DPHP | - | [76] |
| Herring gull liver microsome | In vitro | O-dealkylation | DPHP | 22 ± 2 mL/min/mg protein | [31] | |
| Marine mammal liver microsome | In vitro | O-dealkylation | DPHP | - | [73] | |
| Embryonated eggs and chicks of Japanese Quail | In vivo | O-dealkylation | DPHP, OH-TPHP, 2OH-TPHP, and OH-DPHP, | 1.1-1.8 d (t1/2) | [74] | |
| Zebrafish | In vivo | O-dealkylation, hydroxylation, di-hydroxylation, and GLU conjugation | MPHP and GLU-TPHP | 20.5 h (t1/2) | [78] | |
| Common fish | In vivo | O-dealkylation and hydroxylation | DPHP and OH-TPHP | 9.7-18.6 h (t1/2) | [83] | |
| Black carp (O. latipes) | In vivo | O-dealkylation, hydroxylation, methylation, GLU conjugation, CYS conjugation, and sulfation | DPHP, OH-TPHP, SH-TPHP, SUL-TPHP, GLU-TPHP, and MET-TPHP | - | [90] | |
| Liver microsomes of yellow catfish, catfish and crucian carp | In vitro | - | - | 1.33-1.50 mL/min/mg protein | [39] | |
| Rat liver microsome | In vitro | - | - | 0.1531 h (t1/2) | [88] | |
| Mice | In vivo | O-dealkylation | DPHP | - | [87] | |
| Marine mussel | In vivo | - | - | 1.47 d (t1/2) | [92] | |
| Daphnia magna | In vivo | O-dealkylation, hydroxylation, GSH conjugation, CYS conjugation, and sulfation | DPHP, OH-TPHP, GSH-TPHP, CYS-TPHP, and SUL-TPHP | - | [89] | |
| Daphnia magna | In vivo | - | - | 6.66-7.88 h (t1/2 for waterborne exposure) | [91] | |
| Earthworm | In vivo | O-dealkylation, hydroxylation, CYS conjugation, mercaptolactic acid conjugation, mercaptoethanol conjugation, and GLU conjugation | DPHP, OH-TPHP, CYS-TPHP, MCL-TPHP, MCH-TPHP, and GLU-TPHP | - | [94] | |
| TCP | Zebrafish | In vivo | O-dealkylation, hydroxylation, and GLU conjugation | DCP, OH-DCP, and GLU-TCP | - | [79] |
| Mice | In vivo | O-dealkylation | DCP | - | [87] | |
| Marine mussel | In vivo | - | - | 3.15 d (t1/2) | [92] | |
| Liver microsomes of yellow catfish, catfish and crucian carp | In vitro | - | - | 2.11-2.71 mL/min/mg protein | [39] | |
| CDP | Crucian carp liver microsomes | In vitro | - | - | 1,2700 ± 2,120 mL/min/mg protein | [85] |
| EHDPHP | Common fish | In vivo | O-dealkylation and hydroxylation | EHPHP and OH-EHDPHP | 8.8-17.6 h (t1/2) | [83] |
| Zebrafish | In vivo | O-dealkylation, hydroxylation, GLU conjugation, and sulfation | EHPHP, OH-EHDPHP, DPHP, OH-DPHP, GLU-DPHP, MPHP, and SUL-EHDPHP | - | [84] | |
| Liver microsomes of yellow catfish, catfish and crucian carp | In vitro | - | - | 2.44-3.86 mL/min/mg protein | [39] | |
| Marine mussel | In vivo | - | - | 5.78 d (t1/2) | [92] | |
| BPA-BDP | Rat liver microsomes | In vitro | O-dealkylation and hydroxylation | DPHP, BPA, phenol, BPA-(diphenyl phosphate), BPA-(diphenyl phosphate)-(monophenyl phosphate), BPA-BDP + O metabolite, etc. | - | [77] |
| Bird liver microsomes | In vitro | Too slow | - | - | [77] |
In a 14 d exposure and 28 d depuration experiment of laying hens, the half-lives (t1/2) of five OPFRs were in the range of 11.3-106 d in the egg, with DAPs detected as main metabolites[76]. Other kinetic results showed that the non-halogenated OPFRs (i.e., TNBP, TBOEP, and TPHP) were more quickly metabolized by the liver microsomes, whereas the halogenated OPFRs were transformed to their metabolites (DAPs) more efficiently to non-halogenated OPFRs[31]. In another study, no significant metabolism of BPA-BDP was found in the herring gull liver microsomes[77].
Fish and marine mammals
The metabolism of OPFRs in fish was found to be more complex than that in birds. Wang et al. first elucidated the metabolic pathways of TPHP, TPRP, TNBP, TBOEP, TCEP, TDCPP, and TCP in zebrafish[78,79], including O-dealkylation, hydroxylation, di-hydroxylation, dichlorination (for Cl-OPFRs) and glucuronic acid (GLU) conjugation after hydroxylation. DAPs were detected as the major metabolite of OPFRs, which were mainly distributed in the fish liver and intestine[78,79]. Our previous in vivo and in vitro studies have also identified that the hydroxylation, other oxidation pathways, and GLU conjugation, as well as the O-dealkylation process from TBOEP, TNBP, and TEHP to their respective DAPs, are significant for the metabolism of OPFRs in fish (Chinese rare minnow)[80-82]. In addition, Tang et al. quantified the formation of DAPs and hydroxylated OPFRs (OH-OPFRs) metabolites in common fish exposure experiments of TCPP and EHDPHP[83]. The in vitro biotransformation pathways [including O-dealkylation, hydroxylation, sulfation (SUL), and GLU conjugation] of EHDPHP were also identified in the liver and intestine homogenates of zebrafish[84]. Furthermore, the gut microbiota of zebrafish was analyzed to possess CYP450 catalysis-related enzymes, which might also be involved in the EHDPHP transformation[84].
Considering the liver to be the most important tissue for the metabolism of flame retardants in fish, liver microsomes isolated from various fish species have been used as a promising approach to evaluate the metabolism kinetics of OPFRs. According to two previous in vitro studies, the hepatic metabolism rates of OPFRs in fish were structure dependent, where aryl-OPFRs or OPFRs with larger Log KOW have faster metabolism rates than others under the same conditions[39,85]. In a study of hepatic in vitro metabolism of OPFRs in East Greenland polar bears and ringed seals, the mass balance results indicated a very efficient conversion from TDCPP and TPHP to their respective DAPs[77], which was similar to the findings in fish[78,80,86]. Both NADPH-dependent enzymes (e.g., CYP450 enzymes) and NADPH-independent enzymes are involved in the transformation of OPFRs into DAPs in the marine mammal liver[77].
Rodents
Our previous review provided a basic discussion on the metabolic processes of OPFRs in rodents (including rats and mice), where dealkylation, hydroxylation, glutathione (GSH) conjugation, and GLU conjunction were proposed as the main metabolic pathways[68].
The latest studies can provide novel insights into the metabolism of OPFRs in rodents. A study of BPA-BDP metabolism in rat liver microsomes suggested that the metabolism rate of BPA-BDP was much slower than TPHP and O-dealkylation and oxidation were the main biotransformation pathways for BPA-BDP[77]. In PM2.5-bound OPFR exposure at environmentally realistic concentrations, chlorinated OPFRs (TDCPP, TCEP, and TCPP) accumulated more in mice than other OPFRs (TCP, TPHP, and TNBP)[87]. The DAPs [dicresyl phosphate (DCP), diphenyl phosphate (DPHP), and di-n-butyl phosphate (DNBP)] were detected as urinary metabolites for their corresponding parents in the mice[87]. TPHP was found to be more easily metabolized than TDCPP by rat liver microsomes, which can explain the accumulation potential for chlorinated OPFRs in rodents[88]. NADPH-independent enzymes play an important role in the metabolism of OPFRs in rodents. CYP2E1, CYP2D6, CYP1A2, and CYP2C19 were identified as the specific enzymes for the metabolism of TDCPP, whereas CYP2E1 was the primary CYP450 isoform for the in vitro metabolism of TPHP[88].
Invertebrates
Although many studies are available concerning the metabolism of OPFRs in vertebrates, studies in invertebrates are insufficient. Daphnia magna is a primary consumer in aquatic ecosystems and prey for higher-level consumers, which was used to identify the biotransformation mechanism of TPHP in aquatic invertebrates[89,90]. Both phase I reactions (hydrolysis, hydroxylation, reduction, and (de)hydration) and phase II reactions [GSH conjugation, cysteine (CYS) formation, and sulfate conjugation] were identified for the metabolism of TPHP in D. magna[89]. More than 70% of the TPHP in water was accumulated by the
As a hotspot terrestrial specie, earthworms have recently been used to assess OPFR metabolism. The metabolism of TPHP and TNBP has previously been studied in vivo in earthworms (E. fetida)[93,94]. Major phase I metabolites for TPHP are DPHP, para- and meta-hydroxyphenyl diphenyl phosphate (OH-TPHP), and (OH)2-TPHP[94], while DNBP and dibutyl hydroxybutyl phosphate (OH-TNBP) were the major phase I metabolites for TNBP[93]. Reported phase II metabolites included the thiol conjugates and glucoside conjugates of TPHP and TNBP[93,94]. TBOEP can accumulate in E. fetida and activate the CYP and glutathione pathways to promote the metabolism of TBOEP[95]. Bis(2-butoxyethyl) phosphate (BBOEP), 2-butoxyethyl hydroxyethyl phosphate (BOEHEP), bis(2-butoxyethyl) hydroxyethyl phosphate (BBOEHEP), and bis(2-butoxyethyl) 3-hydroxyl-2-butoxyethyl phosphate (3-OH-TBOEP) were identified as the main metabolites of TBOEP in earthworms[95].
NHFRs
Birds
To the best of our knowledge, very few studies on the metabolism and biotransformation of NBFRs in birds have been reported [Table 3]. 2,3,4,5-Tetrabromobenzoic acid (TBBA) and 2-ethylhexyl tetrabromophthalate (TBMEHP) were respectively detected as the metabolites of TBB and TBPH in eagle eggs from the Great Lakes Region, indicating that O-dealkylation occurred for the metabolism processes of the two NHFRs[96]. In an exposure experiment with Japanese quail eggs, neither single- nor mixture-exposed DPs showed metabolism during incubation[74]. However, relatively rapid depurations for DP isomers (t1/2 of 2.46-5.59 d for anti-DP and 2.76-5.87 d for syn-DP) were found in chicken embryos, indicating the species-specific metabolism of DPs[97]. Although several other studies have been conducted in ovo exposure to NHFRs (including BTBPE and TBBPA-DBPE)[98-100], no evidence for their metabolic pathways was reported.
Available information on metabolism pathways and toxicokinetics of NHFRs in non-human fauna
| Compounds | Species/assays | Methods | Metabolism pathway | Major metabolites | Available toxicokinetic constants | References |
| DBDPE | Zebrafish larvae | In vivo | Debromination | Dibrominated metabolites without confirmed structures | - | [102] |
| Zebrafish larvae | In vivo | Debromination | nona-BDPE, octa-BDPE, hepta-BDPE, hexa-BDPE, and penta-BDPE | - | [103] | |
| Zebrafish | In vivo | Debromination | nona-BDPE, octa-BDPE, hepta-BDPE, hexa-BDPE, and penta-BDPE | 1.50-8.33 d (t1/2) | [105] | |
| Marine mammal liver microsome | In vitro | - | No metabolites detected | - | [101] | |
| Marine fish liver microsome | In vitro | - | - | 0.044-0.050 mL/h/mg protein | [44] | |
| Freshwater fish liver microsome | In vitro | - | - | 0.073-0.162 mL/h/mg protein | [41] | |
| Marine mammal liver microsomes | In vitro | Debromination | Phenolic metabolites | ≈ 0.185 mL/h/mg protein | [101] | |
| Rat | In vivo | Debromination | MeSO2-nona-BDPE and EtSO2-nona-BDPE | - | [115] | |
| Clam | In vivo | Debromination | nona-BDPE, hexa-BDPE, and penta-BDPE | 0.9-11.6 d (t1/2) | [122] | |
| Mudsnails | In vivo | Debromination | nona-BDPE, octa-BDPE, hepta-BDPE, hexa-BDPE, and penta-BDPE | 3.0-3.8 d (t1/2) | [121] | |
| BTBPE | Rainbow trout juvenile | In vivo | - | No metabolites detected | 54.1 ± 8.5 d (t1/2) | [112] |
| Zebrafish | In vivo | Debromination and O-dealkylation | TBP and vinyl tribromobenzene ether | 1.00-7.25 d (t1/2) | [105] | |
| Fathead minnow | In vivo | O-dealkylation and hydroxylation | DBP | - | [106] | |
| Rainbow trout Liver microsome | In vitro | O-dealkylation and hydroxylation | TBP and TBPE | - | [104] | |
| Marine fish S9 fraction | In vitro | - | - | 0.13-0.20 mL/h/mg protein | [111] | |
| Rat | In vivo | Hydroxylation, debromination, and O-dealkylation | OH-BTBPE, (OH)2-BTBPE, TBP, and TBPE | - | [116] | |
| Clam | In vivo | O-dealkylation, hydroxylation, and methylation | OH-BTBPE, MeOH(OH)-BTBPE, TBP, and TBPE | 2.07-5.87 d (t1/2) | [122] | |
| TBB | Bald eagle eggs | In ovo | O-dealkylation | TBBA | - | [96] |
| Fathead minnow liver S9 fraction | In vitro | O-dealkylation and methylation | TBBA, Di-BB, and TBMB | 2.40 ± 0.15 pmol/h/mg protein | [107] | |
| Common carp liver S9 fraction | In vitro | O-dealkylation and methylation | TBBA, Di-BB, and TBMB | 2.34 ± 0.12 pmol/h/mg protein | [107] | |
| Rainbow trout liver microsome | In vitro | O-dealkylation | TBBA | - | [104] | |
| Marine fish liver microsome | In vitro | - | - | 0.053-0.065 mL/h/mg protein | [44] | |
| Marine fish S9 fraction | In vitro | - | - | 0.18-0.50 mL/h/mg protein | [111] | |
| Rat | In vivo | O-dealkylation | TBBA | - | [120] | |
| Rat | In vivo | O-dealkylation | TBBA and TBPA | - | [117] | |
| Rat liver microsome | In vitro | O-dealkylation | TBBA | 6.25 ± 0.58 nmol/min/mg protein | [118] | |
| Rat liver cytosol | In vitro | O-dealkylation | TBBA | 0.203 ± 0.004 nmol/min/mg protein | [118] | |
| Rat intestinal microsome | In vitro | O-dealkylation | TBBA | 0.422 ± 0.093 nmol/min/mg protein | [118] | |
| TBPH | Bald eagle eggs | In ovo | O-dealkylation | TBMEHP | - | [96] |
| Killifish (Fundulus heteroclitus) | In vivo | 22 d (t1/2) | [114] | |||
| Fathead minnow liver S9 fraction | In vitro | O-dealkylation and methylation | TBBA, Di-BB, and TBMB | 0.629 ± 0.066 pmol/h/mg protein | [107] | |
| Common carp liver S9 fraction | In vitro | O-dealkylation and methylation | TBBA, Di-BB, and TBMB | 0.620 ± 0.103 pmol/h/mg protein | [107] | |
| Rat | In vivo | O-dealkylation | TBBA | - | [120] | |
| Rat | In vivo | O-dealkylation | TBBA and TBPA | - | [117] | |
| Rat liver microsome | In vitro | - | No metabolites found | - | [118] | |
| Marine fish liver microsome | In vitro | - | - | 0.016-0.017 mL/h/mg protein | [44] | |
| TBBPA-DBPE | Fathead minnow | In vivo | O-dealkylation | TBBPA | - | [106] |
| Marine fish liver microsome | In vitro | - | - | 0.047-0.048 mL/h/mg protein | [44] | |
| PBT | Zebrafish | In vivo | Debromination | Tetra-BT, Tri-BT, and Di-BT | 1.14-10.37 d (t1/2) | [105] |
| Marine fish liver microsome | In vitro | - | - | 0.043-0.049 mL/h/mg protein | [44] | |
| Marine fish S9 fraction | In vitro | - | - | 0.05-0.28 mL/h/mg protein | [111] | |
| Clam | In vivo | Debromination | Tetra-BT | 3.22-6.48 d (t1/2) | [122] | |
| Mudsnails | In vivo | Tetra-BT, Tri-BT, and Di-BT | 4.7-5.9 d (t1/2) | [121] | ||
| PBP | Marine fish liver microsome | In vitro | - | - | 0.053-0.055 mL/h/mg protein | [44] |
| HBB | Zebrafish | In vivo | Debromination | PBB, Tetra-BB, Tri-BB, and Di-BB | 0.85-10.34 d (t1/2) | [105] |
| Marine fish liver microsome | In vitro | - | - | 0.017-0.025 mL/h/mg protein | [44] | |
| Marine fish S9 fraction | In vitro | - | - | 0.048-0.13 mL/h/mg protein | [111] | |
| Clam | In vivo | Debromination | PBB, Tetra-BB, Tri-BB, and Di-BB | 1.82-.6.54 d (t1/2) | [122] | |
| Mudsnails | In vivo | Debromination | PBB, Tetra-BB, Tri-BB, and Di-BB | 2.5-3.5 d (t1/2) | [121] | |
| PBEB | Marine fish S9 fraction | In vitro | - | - | 0.052-0.40 mL/h/mg protein | [111] |
| TBECH | Zebrafish | In vivo | - | - | 1.3 d (t1/2) | [113] |
| Marine fish liver microsome | In vitro | - | - | 0.061-0.067 mL/h/mg protein | [44] | |
| Freshwater fish liver microsome | In vitro | - | - | 0.006-0.027 mL/h/mg protein | [41] | |
| TBP | Zebrafish | In vivo | - | - | 0.9-1.3 d (t1/2) | [113] |
| Rat | In vivo | Glucuronic acid conjugation, sulfation | GLU-TBP and SUL-TBP | 2-5 h (t1/2) | [119] | |
| TBCT | Marine fish S9 fraction | In vitro | - | - | 0.052-0.40 mL/h/mg protein | [111] |
| Freshwater fish liver microsome | In vitro | - | - | 0.015-0.114 mL/h/mg protein) | [41] | |
| PBBA | Freshwater fish liver microsome | In vitro | - | - | 0.122 mL/h/mg protein | [41] |
| DPs | Embryonated eggs of Japanese Quail | In vivo | Too slow | - | - | [74] |
| Chicken embryos | In vivo | - | - | 2.46-5.59 d (t1/2 for anti-DP) 2.76-5.87 d (t1/2 for syn-DP) | [97] | |
| Common carp | In vivo | - | - | 16.3-50.2 d (t1/2 for anti-DP) 17.8-45.6 d (t1/2 for syn-DP) | [108] | |
| Rainbow trout (Oncorhynchus mykiss) | In vivo | - | - | 53.3 ± 13.1 d (t1/2 for anti-DP) 30.4 ± 5.7 d (t1/2 for syn-DP) | [109] | |
| Redtail catfish | In vivo | - | No metabolites found | 19.1-39.7 d (t1/2 for anti-DP) | [110] | |
| Oscar fish | In vivo | - | No metabolites found | 22.3-34.5 d (t1/2 for syn-DP) | [110] |
Fish and marine mammals
DBDPE could be rapidly metabolized (39.6-66.6 pmol in 90 min) to phenolic metabolites by marine mammal liver microsomes from arctic areas (polar bear, beluga whale, and ringed seal)[101]. DBDPE debromination (7 unknown compounds) was also confirmed in zebrafish after water-borne exposure[102]. They tentatively assigned them to nona-BDPE, nona-brominated products, octa-BDPE, hepta-BDPE, and other-brominated products[103]. BTBPE can be transformed into TBP and tribromophenoxyethanol (TBPE) during in vitro incubation using rainbow trout liver microsomes[104]. The formation of TBP was also confirmed in metabolism of BTBPE in zebrafish[105], whereas dibromophenol (DBP) was identified as a metabolite of BTBPE in fathead minnow[106]. HBB went through multiple debromination to metabolites of penta-bromobenzene (PBB), 1,2,4,5-tetra bromobenzene (Tetra-BB), 1,2,4-tribromobenzene (Tri-BB), and dibromobenzene (Di-BB) in zebrafish, and PBT could be gradually transformed to tetrabromotoluene (Tetra-BT), tribromotoluene (Tri-BT), and dibromotoluene (Di-BT)[105]. Ganci et al. identified TBBA as the major metabolite of TBB by trout liver microsomes[104]. Except for TBBA, Di-BB, and 2,3,4,5-tetrabromomethylbenzoate (TBMB), formed via dealkylation and methylation, were detected as metabolites for the mixture of TBB and TBPH in fathead minnow and common carp liver S9 fraction[107]. Fathead minnow (P. promelas) exposed to TBBPA-DBPE was found to produce TBBPA via hydrolysis (O-dealkylation)[106]. DPs have been inferred to be metabolized in the liver of freshwater fish[108-110], but no metabolite could be detected in the fish body.
In vitro incubation using liver microsomes was conducted in several studies to assess the biotransformation clearance rates of NHFRs in fish. Lee et al. first found chemical-to-chemical variations in the metabolism rate of 6 NHFRs (BTBPE, HBB, PBEB, PBT, TBB, and TBCT) in marine fish (Epinephelus septemfasciatus, Konosirus punctatus, Lateolabrax japonicus, Mugil cephalus, and Sebastes schlegelii) from Koera[111]. Generally, the fully brominated NHFRs were metabolized slower than the less brominated NHFRs in fish. TBB exhibited the fastest metabolism rate in fish liver S9 fractions, whereas HBB and TBCT were the two slowest depleted NHFRs[111]. Our previous study using marine fish from the South China Sea liver microsome also reported the lowest in vitro clearance rate constants for HBB compared with TBB and PBT[44]. The clearance rates of NHFRs in marine fish from our study were 1.16 (TBB) - 7.68 (PBT) times lower than the values obtained in the marine fish from Korea, which might be attributed to the difference in enzyme activities between liver S9 and microsomes. In a study using freshwater fish liver microsomes (crucian carp, catfish, and yellow-head catfish), ATE, BTBPE, and TBPH showed no significant metabolism, and the clearance rate of DBDPE was much higher than that in marine fish from our previous study[41]. These results imply the occurrence of species-specific metabolism of NHFRs in aquatic animals.
The t1/2 of BTBPE was estimated to be approximately 54.1 ± 8.5 d in juvenile rainbow trout (Oncorhynchus mykiss)[112], and the estimated t1/2 for TBECH and TBP were < 2 d in zebrafish[113]. Qiao et al. also found that the liver, intestine, and gill were the top three tissues for the accumulation of PBT, HBB, BTBPE, and DBDPE in zebrafish with t1/2 lower than 7 d[105]. In a dietary exposure of TBPH to Atlantic killifish (Fundulus heteroclitus), only a very small proportion of the TBPH in diet (< 0.5%) was bioaccumulated in fish by 28 d and the depuration t1/2 was estimated to be
Rodents
Recent in vitro and in vivo studies in humans and rodents have confirmed the basic metabolic pathways of typical NHFRs. DBDPE is slowly metabolized in rats to MeSO2-nona-BDPE and EtSO2-nona-BDPE[115]. A study based on in vivo exposure of rats found that BTBPE could be metabolized into monohydroxylated and polyhydroxylated BTBPE and the debromination products (i.e., TBP and TBPE)[116]. In addition, 2,3,4,5-tetrabromo phthalic acid (TBPA) is another urine metabolite in rats that results from the metabolism of the TBB and TBPH mixture[117]. In previous studies using rat liver microsomes, TBBA was identified as an in vitro metabolite for TBB, whereas no metabolites were found for TBPH[118]. TBP can be phase II metabolized to GLU-TBP and SUL-TBP by both pregnant and nursing rats[119].
The metabolism of TBB was significantly slower in rat intestinal microsomes and liver cytosol than in rat liver microsomes[118]. In DBDPE-exposed rats, adipose tissue accumulated the majority of DBDPE rather than liver and kidney tissues at 90 d of exposure[115]. For BTBPE, a high proportion of 14C (> 94%) was excreted in the feces at 72 h rather than accumulated in rat tissue[116]. In addition, the lactational transfer of TBB and TBPH was found to be approximately 200- to 300-fold higher than that of placental transfer in dosed Wistar rats, and their common metabolite TBBA was detected in the urine of pups[120]. The TBP-administrated rat could rapidly accumulate in kidney and plasma at 30 min, and the exposed TBP pregnant and nursing rats resulted in the distribution of TBP and its metabolites in their offspring[119].
Invertebrates
In the sediment-water-mudsnail system, nona-BDPE, octa-BDPE, hepta-BDPE, hexa-BDPE, and penta-BDPE were found to be debromination metabolites of DBDPE by snails[121]. The debromination process for DBDPE also occurred in clams, where nona-BDPE and penta-BDPE were detected as major metabolites[122]. Debromination was also investigated as the main metabolic pathway for PBT and HBB in both snails and clams[121,122]. Tetra-BT, Tri-BT, and Di-BT were found to be the major metabolites of PBT, and PBB, Tetra-BB, Tri-BB, and Di-BB were the major metabolites of HBB in the two invertebrate species[121,122]. The hydrolysis and hydroxylation products of BTBPE also had been confirmed in clams[122].
The highest distribution of NHFRs in viscera was found for both snail and clam, and the t1/2 values for PBT, HBB, and DBDPE for snail and clam were 2.5-5.9 d and 0.911-11.6 d, respectively[121,122]. In a study of NHFRs in oil-earthworm systems, HBB and PBT were mainly distributed in the intestine and epidermis
Summary of transformation processes of novel flame retardants
In general, based on the above literature, the metabolic pathways of NHFRs and OPFRs in the fauna can be clarified. The main metabolic pathways of OPFRs include dealkylation (ester hydrolysis) and hydroxylation, and phase II conjunction. DAPs and OH-OPFRs are the most important metabolites in the body. Debromination, hydroxylation, dealkylation, and phase II conjunction occupied the major metabolic pathways of NHFRs in fauna. The most important metabolic pathway for NHFRs with ether bonds is O-dealkylation (hydrolysis), such as BTBPE, TBB, TBPH, and TBPPA-DBPE. Other NHFRs share general metabolic pathways of mono- and multiple hydroxylation and debromination, and phase II metabolism can occur subsequently once hydroxyl is formed for the intermediates. Toxicokinetic results suggest that NHFRs are more resistant to metabolism than OPFRs, especially for DBDPE, DPs, and the monoaromatic NHFRs. For OPFRs, the metabolism of non-chlorinated OPFRs is faster than Cl-OPFRs. Species-specific metabolism of novel flame retardants can be concluded according to the collected studies, where their metabolism rate in birds and rodents is generally faster than in fish and invertebrates.
INTERNAL EXPOSURE OF THE MAJOR NOVEL FR METABOLITES
Several studies are available concerning the internal exposure of novel FR metabolites in fauna. DAPs, formed from in vivo dealkylation, can act as biomarkers for assessing the internal exposure of OPFRs. The DAPs of bis(2-chloroethyl) phosphate (BCEP), bis(1-chloro-2-propyl) phosphate (BCPP), 1-hydroxy-2-propyl bis(1-chloro-2-propyl) phosphate (BCIPHIPP), BDCPP, BBOEP, DNBP, di(2-ethylhexyl) phosphate (DEHP), DPHP, DCP [or so-called bis(methylphenyl) phosphate (BMPP)], and 2-ethylhexyl phenyl phosphate (EHPHP) and OH-OPFRs of BBOEHEP, OH-TBOEP, OH-TNBP, OH-TPHP, and hydroxylated EHDPHP (OH-EHDPHP) were recently detected in fauna biomonitoring studies [Table 4]. The metabolite/parent ratio (MPR) was recently used in internal exposure studies to compare the relative persistence of OPFRs and metabolites [Figure 1], where an MPR ratio higher than one indicates that the metabolites, rather than the parent contaminants, should receive greater concern regarding their accumulation potentials.
Figure 1. The metabolite/parent ratios (MPRs) of OPFRs in fauna across the internal exposure studies (A) Cl-OPFRs, (B) Alkyl-OPFRs, and (C) Aryl-OPFRs). Detailed data are compiled in Table 4.
Internal concentration of NBFR metabolites in fauna (ng/mL or ng/g ww)
| Sample types | Study area | BCEP | BCPP | BCIPHIPP | BDCPP | BBOEP | BBOEHEP | OH-TBOEP | DNBP | OH-TNBP | DEHP | DPHP | OH-TPHP | DCP | EHPHP | OH-EHDPHP | ΣmOPFRs | Reference |
| Cow milk | Asia | - | 0.044 ± 0.079 | - | 0.037 ± 0.056 | 0.02 ± 0.027 | 0.017 ± 0.029 | 0.002 ± 0.003 | 0.024 ± 0.026 | - | - | 0.005 ± 0.016 | - | 0.156 ± 0.139 | - | - | 0.02 ± 0.025 | [126] |
| Europe | - | 0.036 ± 0.046 | - | 0.078 ± 0.118 | 0.011 ± 0.014 | 0.023 ± 0.04 | 0.002 ± 0.006 | 0.044 ± 0.079 | - | - | 0.002 ± 0.004 | - | 0.821 ± 0.181 | - | - | 0.135 ± 0.716 | [126] | |
| North America | - | 0.039 ± 0.031 | - | 0.084 ± 0.109 | 0.005 ± 0.007 | 0.018 ± 0.01 | 0.002 ± 0.002 | 0.036 ± 0.043 | - | - | 0.001 ± 0.001 | - | 0.215 ± 0.128 | - | - | 0.043 ± 0.044 | [126] | |
| South America | - | 0.036 ± 0.019 | - | 0.081 ± 0.146 | 0.004 ± 0.004 | 0.032 ± 0.058 | 0.001 ± 0.001 | 0.048 ± 0.057 | - | - | 0.001 ± 0.002 | - | 0.261 ± 0.125 | - | - | 0.049 ± 0.058 | [126] | |
| Oceania | - | 0.024 ± 0.019 | - | 0.083 ± 0.071 | 0.021 ± 0.017 | 0.011 ± 0.013 | 0.002 ± 0.002 | 0.018 ± 0.018 | - | - | 0.002 ± 0.002 | - | 0.099 ± 0.165 | - | - | 0.017 ± 0.018 | [126] | |
| Cow milk | Beijing, China | - | 0.998 ± 0.45 | - | 0.053 ± 0.12 | 0.274 ± 0.29 | - | - | 0.279 ± 0.15 | - | - | 0.917 ± 0.57 | - | 0.1 ± 0.03 | - | - | 2.62 ± 0.98 | [146] |
| Fishmeal (in dry weight) | United States | 5.36 ± 3.25 | 11.01 ± 3.67 | - | 1.5 ± 2.87 | 0.53 ± 0.63 | - | - | 0.32 ± 0.16 | - | 3.49 ± 5.87 | 3.6 ± 3.43 | - | 29.0 ± 9.1 | - | - | 41.9 ± 13.0 | [127] |
| China | 4.08 ± 2.33 | 0.85 ± 0.75 | - | 2.21 ± 4.18 | 0.11 ± 0.43 | - | - | 0.65 ± 1.69 | - | 7.31 ± 6.8 | 2.05 ± 2.6 | - | 36.6 ± 19.6 | - | - | 52.8 ± 23.0 | [127] | |
| Europe | ND | 1.39 ± 0.62 | - | 1.99 ± 2.73 | 0.2 ± 0.38 | - | - | 0.08 ± 0.21 | - | 1.26 ± 1.68 | 1.86 ± 3.21 | - | 20.6 ± 6.57 | - | - | 28.9 ± 5.68 | [127] | |
| South America | 6.23 ± 3.45 | 1.18 ± 3.68 | - | 1.09 ± 2.37 | 0.09 ± 0.62 | - | - | 0.26 ± 0.68 | - | 2.69 ± 23.3 | 1.24 ± 2.18 | - | 31.8 ± 13.2 | - | - | 42.1 ± 33.9 | [127] | |
| Southeast Asia | ND | 1.7 ± 4.35 | - | 0.98 ± 0.4 | 0.09 ± 0.04 | - | - | 0.21 ± 0.41 | - | 3.46 ± 2.56 | 1.01 ± 1.17 | - | 35.7 ± 11.8 | - | - | 43.6 ± 10.4 | [127] | |
| Meat meal (in dry weight) | China | - | 16.5 | 7.25 | - | 0.27 | - | - | 0.81 | - | 7.98 | 14.9 | - | 2.20 | - | - | 49.9 | [131] |
| Feather meal (in dry weight) | - | 12.5 | 1.83 | - | 0.04 | - | - | 0.21 | - | 2.24 | 4.15 | - | 2.36 | - | - | 23.3 | [131] | |
| Blood meal (in dry weight) | - | 5.57 | 4.29 | - | 0.63 | - | - | 0.77 | - | 8.87 | 8.01 | - | 24.0 | - | - | 52.1 | [131] | |
| Beef | Southeast Queensland, Australia | ND | ND | ND | ND | ND | ND | ND | - | - | 0.043 ± 0.026 | 0.098 ± 0.039 | - | ND | - | - | 0.152 ± 0.033 | [128] |
| Lamb | ND | ND | ND | ND | ND | ND | 0.11 | - | - | ND | 0.207 ± 0.112 | - | ND | - | - | 0.205 ± 0.205 | [128] | |
| Pork | ND | ND | ND | ND | ND | ND | ND | - | - | 0.12 | 0.14 | - | ND | - | - | 0.114 ± 0.163 | [128] | |
| Chicken | ND | ND | ND | ND | ND | ND | ND | - | - | 0.038 | 0.245 ± 0.078 | - | ND | - | - | 0.204 ± 0.182 | [128] | |
| Prawn | ND | ND | 0.42 | ND | 0.23 | ND | 0.124 ± 0.034 | - | - | 0.297 ± 0.147 | 1.17 ± 1.42 | - | 5.00 ± 3.14 | - | - | 7.06 ± 4.79 | [128] | |
| Oyster | ND | ND | ND | ND | 0.52 ± 0.156 | ND | 0.1 | - | - | 0.335 ± 0.149 | 4.53 ± 1.89 | - | 0.407 ± 0.161 | - | - | 6.45 ± 1.88 | [128] | |
| Salmon | ND | ND | ND | ND | ND | ND | 0.083 | - | - | ND | 0.44 | - | ND | - | - | 0.202 ± 0.309 | [128] | |
| Egg | ND | ND | ND | ND | 1.13 ± 0.603 | ND | 0.082 | - | - | 1.63 ± 1.53 | 3.86 ± 1.29 | - | 0.313 ± 0.118 | - | - | 8.28 ± 4.64 | [128] | |
| Egg albumin | Australia | ND | ND | 0.33 | ND | 0.26 | ND | ND | 0.15 | - | 2.3 | 5.3 | - | 0.079 | - | - | 9.7 | [129] |
| Egg yolk | ND | ND | ND | 0.32 | 0.063 | ND | ND | 0.05 | - | 0.72 | 1.2 | - | ND | - | - | 3 | [129] | |
| Herring gull plasma (ng/g ww) | Lake Huron, Canada | - | ND | - | 2.13 ± 1.13 | 5.32 ± 11.8 | - | - | 0.410 | - | 0.120 ± 0.079 | ND | - | - | - | - | 7.98 ± 11.4 | [124] |
| Bald eagle eggs | Great Lakes, USA | 5.4 ± 1.7 | 1.8 ± 0.25 | - | 2.5 ± 0.21 | 1.3 ± 0.3 | - | - | 2.4 ± 0.49 | - | - | 1 ± 0.23 | - | - | - | - | 27 ± 3 | [96] |
| Water snake | E-waste dismantling site in Guangdong, China | - | 0.17 ± 0.13 | 0.029 ± 0.013 | - | 0.076 ± 0.12 | ND | ND | 0.47 ± 0.30 | - | 0.39 ± 0.37 | 0.061 ± 0.057 | - | 0.11 ± 0.033 | ND | 1.3 ± 0.49 | [42] | |
| Snake egg | - | 0.073 ± 0.13 | 0.037 ± 0.013 | - | 0.29 ± 0.37 | 0.022 ± 0.038 | 0.031 ± 0.033 | 0.39 ± 0.29 | - | 0.50 ± 0.15 | 0.28 ± 0.22 | - | 0.32 ± 0.10 | 0.046 ± 0.08 | 2.0 ± 0.41 | [42] | ||
| Common carp | - | 0.54 ± 0.11 | 0.19 ± 0.16 | - | 0.41 ± 0.24 | 0.019 ± 0.010 | 0.019 ± 0.0090 | 0.51 ± 0.32 | - | 0.61 ± 0.37 | 1.3 ± 1.9 | - | 0.24 ± 0.24 | 0.059 ± 0.045 | 2.8 ± 0.41 | [42] | ||
| Topmouth gudgeon (in lipid weight) | Rivers in Beijing, China | - | - | - | - | 33.4 ± 32.2 | - | - | 23.3 ± 15.3 | 26 ± 11.1 | 10.4 ± 6.3 | - | - | - | - | 93.1 ± 46.2 | [38] | |
| Crucian carp (in lipid weight) | - | - | - | - | 25.1 ± 17.5 | - | - | 34 ± 18.7 | 30.9 ± 16.6 | 12.2 ± 9.1 | - | - | - | - | 102 ± 43.6 | [38] | ||
| Loach (in lipid weight) | - | - | - | - | 32.9 ± 28.7 | - | - | 113 ± 92.4 | 58.6 ± 52.3 | 16.3 ± 12.9 | - | - | - | - | 220 ± 150 | [38] | ||
| Marine snail (in lipid weight) | Pearl river estuary, China | - | - | - | - | 55.6 ± 97.6 | 0.49 ± 0.55 | 0.01 ± 0.01 | 68.9 ± 36.7 | 1.01 ± 0.93 | - | 11.5 ± 8.00 | - | - | - | - | 137 ± 134 | [125] |
| Marine shrimp (in lipid weight) | - | - | - | - | 18.2 ± 11.5 | 0.55 ± 0.31 | 0.29 ± 0.47 | 98.6 ± 63.7 | 1.55 ± 0.82 | - | 8.47 ± 7.64 | - | - | - | - | 140 ± 62.5 | [125] | |
| Marine crabs (in lipid weight) | - | - | - | - | 23.9 ± 13.4 | 1.05 ± 0.43 | 0.19 ± 0.08 | 314 ± 360 | 1.70 ± 1.61 | - | 13.3 ± 9.21 | - | - | - | - | 384 ± 341 | [125] | |
| Marine fish (in lipid weight) | - | - | - | - | 9.99 ± 10.0 | 0.40 ± 0.44 | 0.18 ± 0.31 | 66.9 ± 47.7 | 2.78 ± 3.27 | - | 11.4 ± 12.1 | - | - | - | - | 89.1 ± 59.4 | [125] | |
| 8 marine fish species | Tarragona, Spain | ND | - | - | ND | ND | - | - | 47.6 ± 18.2 | - | ND | 62.6 ± 18.4 | - | - | - | - | 110 ± 34.9 | [145] |
| Stickleback | Troutman Lake, Alaska, USA | 0.081 ± 0.009 | ND | - | ND | - | - | - | 0.436 ± 0.066 | - | - | 0.410 ± 0.143 | - | - | - | - | 0.927 ± 0.218 | [143] |
In a recent report by Su et al., BBOEP and BDCPP were detected at concentrations higher than 2 ng/g ww in herring gull plasma from the Great Lakes[124]. Our previous study investigated the accumulation of four DAPs (i.e., BBOEP, DNBP, DEHP, and DPHP) in wild freshwater fish from Beijing, China, and found that ΣDAPs concentrations were approximately 0.10-1.12 times (MPR) those of their parent compounds in fish[38]. The four DAPs in crucian carp and loach were mainly distributed in the fish liver (135 and
In a global survey of OPFR metabolites in cow milk, samples from European countries presented higher OPFR metabolite concentrations in all countries (ΣmOPFRs = 0.135 ng/mL), while the metabolite levels in Asian countries were much lower (mean level < 0.021 ng/mL)[126]. TDCPP/BDCPP and TCPP/BCPP pairs presented significantly positive correlations, which indicated that they shared similar sources in milk[126]. BBOEP and BBOEHEP showed much higher concentrations than the hydroxyl metabolites (i.e., OH-TBOEP) in milk, which might be attributed to the high conversion rate from OPFRs to their corresponding DAPs[126]. However, the concentration of ΣmOPFRs in fishmeal showed a geographic order of China
Relatively little information exists regarding the internal exposure of NBFR metabolites in fauna. TBBA [from not detectable (ND) to 330 ng/g ww] and TBMEHP (ND-330 ng/g ww) were detected in the bald eagle (Haliaeetus leucocephalus) eggs from the Great Lakes region[96]. For the two metabolites, their corresponding parent compounds (i.e., TBB and TBPH) were not detected in the eggs, suggesting greater concern should be paid to the two metabolites rather than their parents[96].
In general, DAPs have relatively higher internal exposure concentrations in fauna than OH-OPFRs, which is related to their high conversion rate and stability in the body [Figure 1]. The higher MPRs than 1 were frequently reported for the DAP/alkyl-OPFR pairs, which may be related to the easy-to-metabolism characteristics of the alkyl-OPFRs [Figure 1]. However, the sources of novel FR metabolites in the body are complex. In addition to being formed from metabolic processes in the body, some can also be formed from biotic and abiotic degradation processes in the environment before accumulation by the fauna. Some of the metabolites can also be applied as industrial products. For example, DEHP, DPHP, DCP, DMPP, and DNBP can be used as FRs or plasticizers[65]. Thus, the internal exposure of metabolites in the body is not always relative to the external exposure to FRs.
TOXICITY OF THE MAJOR NOVEL FR METABOLITES
OPFR metabolites
The predicted Log KOW values (using EPI suite v4.1) for the major OPFR metabolites were lower than those of their parent compounds [Table 5], which indicated their comparably limited potential for bioaccumulation. The estimated results from the EPA T.E.S.T. program indicate that DAPs and OH-OPFRs exhibit lower acute toxicities to aquatic animals. However, the estimated developmental toxicity for OPFRs is not eliminated after metabolism. BCEP, DNBP, OH-TNBP, and DCP show significantly positive developmental toxicity, while their parent compounds do not.
The estimated ecotoxicities and bioaccumulation values for the major novel FR metabolites
| FRs | Metabolites | Fathead minnow LC50 (mg/L 96 h)a | Daphnia magna LC50 (mg/L 48 h)a | T. pyriformis IGC50 | Developmental Toxicitya | Mutagenicityaa | Estimated | Estimated BCFb |
| TCEP | 14.53 | 0.040 | 228.51 | - | + | 1.63 | 3.465 | |
| BCEP | 17.91 | 0.095 | NA | + | + | 0.83 | 1.457 | |
| TCPP | 5.80 | 0.018 | 150.20 | + | + | 2.89 | 36.66 | |
| BCPP | 12.77 | 0.180 | NA | + | + | 1.19 | 2.251 | |
| BCIPHIPP | 9.22 | 0.049 | 511.98 | + | + | 1.17 | 1.557 | |
| TDCPP | 0.22 | 0.016 | 154.86 | + | - | 3.65 | 126.3 | |
| BDCPP | 2.09 | NA | NA | + | NA | 1.70 | 5.511 | |
| TBOEP | 28.57 | 0.040 | 93.78 | + | - | 3.65 | 54.19 | |
| BBOEP | 14.88 | 0.27 | NA | + | - | 1.74 | 5.094 | |
| BBOEHEP | 61.71 | 0.061 | 310.34 | + | - | 0.82 | 1.079 | |
| 3-OH-TBOEP | 36.66 | 0.15 | 465.24 | + | - | 1.53 | 1.737 | |
| TEHP | 0.56 | 0.021 | NA | - | - | 9.49 | 1.4 | |
| DEHP | 0.42 | NA | NA | - | NA | 5.60 | 823.7 | |
| TNBP | 18.60 | 0.030 | 124.47 | - | - | 3.82 | 69.65 | |
| DNBP | 5.20 | 0.66 | NA | + | - | 2.29 | 16.37 | |
| 3-OH-TNBP | 11.15 | 0.20 | 383.34 | + | - | 2.28 | 7.905 | |
| TPHP | 1.12 | 0.10 | 12.12 | + | - | 4.70 | 73.18 | |
| OH-TPHP | 0.12 | 0.16 | 19.69 | - | - | 4.22 | 46.75 | |
| DPHP | 6.95 | NA | NA | + | - | 2.88 | 40.14 | |
| EHDPHP | 0.21 | 0.062 | 2.73 | + | - | 5.73 | 273.1 | |
| EHPHP | 1.08 | NA | NA | + | - | 4 | 195.5 | |
| OH-EHDPHP | 0.34 | 0.036 | 3.16 | + | - | 5.82 | 149 | |
| TCP | 0.19 | 0.54 | 2.48 | - | - | 6.34 | 2.98 × 104 | |
| DCP | 4.90 | NA | NA | + | NA | 3.50 | 241.5 | |
| TBB | 0.12 | 0.096 | 0.063 | NA | + | 8.75 | 2072 | |
| TBBA | 1.02 | 10.08 | 18.02 | NA | - | 5.09 | 835.2 | |
| TBPH | 0.007 | 0.089 | 0.017 | NA | - | 11.95 | 2.401 | |
| TBMEHP | 0.032 | 1.16 | 0.54 | NA | - | 7.53 | 169.1 | |
| TBBPA-DBPE | 0.004 | 0.003 | 3.83 × 104 | NA | - | 11.52 | 1.215 × 104 | |
| TBBPA | 0.069 | 0.033 | 0.11 | NA | - | 2.856 | 717.5 |
According to the literature, some transformation products might be more toxic than parent compounds, especially for endocrine-disrupting endpoints. TNBP shows both androgen receptor and glucocorticoid receptor antagonistic activity, whereas its metabolite DNBP cannot exhibit any nuclear receptor activity[132]. 5-OH-EHDPHP can elicit approximately 3.1 times the androgen receptor antagonistic activity of EHDPHP in Japanese medaka (Oryzias latipes)[133]. The metabolites BBOEHEP and 3-OH-TBOEP can act as pregnancy X receptor agonists at similar levels to their parent TBOEP[132]. DPHP can significantly dysregulate the avian genes associated with lipid/cholesterol metabolism, which is more than two times that of TPHP[72]. Low-dose chronic exposure to DPHP can interrupt the fatty acid metabolism in the rat liver and exert adverse consequences on overall physiology[134]. Similar adverse results were also observed in male zebrafish[135]. OH-TPHP elicited the upregulation of estrogenic genes and thyroid genes to induce growth inhibition in zebrafish embryos[136]. Both BCPP and BDCPP upregulated the genes encoding for estrogenic synthesis enzymes in H295R cells, which indicated that these metabolites may produce comparable or even higher endocrine-disrupting effects than OPFRs[137].
NBFR metabolites
All the estimated NBFR metabolites had lower Log KOW values and aquatic toxicities (including LC50 to Fathead minnow and Daphnia magna and IGC50 to T. pyriformis) than those of their parent compounds using the in silico methods [Table 5]. However, certain metabolites of NHFRs also exhibit other adverse effects on organisms, according to previous studies. The metabolites TBBA and TBMEPH were shown to have comparable thyroid hormone, androgen, glucocorticoid, and pregnancy X receptor agonist activities[138,139] and induced stronger cytotoxicity than their parent compounds (TBB and TBHP)[140]. TBBA and TBMEHP exhibited binding potency to human PPARγ, but TBB and TBPH did not[141]. TBP, one of which was reported as a BTBPE metabolite, is an industrial additive with stronger neurotoxicity and can inhibit the expression of human steroidogenic enzymes, leading to a certain degree of endocrine-disrupting effect[60]. Bromophenol, another BTBPE metabolite, was found to have strong cytotoxic and genotoxic effects on aquatic organisms[142].
CONCLUSION AND PERSPECTIVES
To date, great efforts have been made to study the metabolism of novel FRs in fauna, such as metabolic pathways and kinetics, metabolite formation, internal exposure of metabolites, and their toxicities. OPFRs share similar metabolic pathways in various animals, where O-dealkylation, hydroxylation, and phase II conjunction are the most likely pathways. DAPs and OH-OPFRs are the predominant metabolites in the body. O-dealkylation (hydrolysis) is the key pathway controlling the metabolism of NHFRs with ether bonds, while other NHFRs might metabolize through debromination, hydroxylation, dealkylation, and phase II conjunction. However, compared with OPFRs, there is still a lack of metabolism information on most of the NHFRs including their full metabolism pathways, the conversion efficiency of specific metabolites, and the stability of the intermediates in the body[6,11,69]. The metabolism kinetics (or toxicokinetics) of novel FRs are CYP enzyme-related and variable among species. Research has progressed to often evaluating the metabolism of novel FRs in a single species, but comparative studies of biotransformation between species remain insufficient. When invertebrates, which are at the lower levels of the food chain, are exposed to FRs, the parental compounds and their metabolites can affect the organisms at the upper levels[125]. Therefore, future research is necessary on the metabolic processes in multitrophic organisms and the transfer of major metabolites across the food web.
DAPs, as important OPFR metabolites, have been investigated as biomarkers for OPFR exposure in fauna. The occurring higher internal exposure of DAPs than the respective OPFRs also highlights their potential risk for animals and their importance in understanding the metabolism processes of OPFRs. Nevertheless, few studies have focused on the internal exposure of NBFR metabolites, and we recommended employing these biomarkers for biomonitoring fauna. A few studies have indicated that the residues of the major FR metabolites in the body may have adverse effects on fauna. These results underscore the importance of studying the occurrence and ecological risks of metabolites in organisms. In addition, internal exposure data of metabolites can provide valuable information for human exposure and risk assessments of novel FRs. Hence, more attention should concentrate on the co-exposure of FRs and their metabolites, especially for those FRs with easy-metabolic characteristics and stable metabolites in the body.
DECLARATIONS
Authors’ contributionsConceptualization and methodology, data analysis, writing-review & editing: Hou R
Reviewing and editing: Sun C, Zhang S, Huang Q, Liu S, Lin L, Li H
Project administration, resources and supervision: Hou R, Xu X
Availability of data and materialsAll the data were included in this paper. No additional data are available.
Financial support and sponsorshipThis work was supported jointly by the National Natural Science Foundation of China (No. 41907339), Foundation of MNR Key Laboratory of Eco-Environmental Science and Technology, China (MEEST-2021-03), the National Key Research and Development Program of China (2022YFC3105600), and the Natural Science Foundation of Guangdong Province (2022A1515011498).
Conflicts of interestAll authors declared that there are no conflicts of interest.
Ethical approval and consent to participateNot applicable.
Consent for PublicationNot applicable.
Copyright© The Author(s) 2023.
REFERENCES
1. Yu G, Bu Q, Cao Z, et al. Brominated flame retardants (BFRs): a review on environmental contamination in China. Chemosphere 2016;150:479-90.
2. Covaci A, Harrad S, Abdallah MA, et al. Novel brominated flame retardants: a review of their analysis, environmental fate and behaviour. Environ Int 2011;37:532-56.
3. Chen D, Kannan K, Tan H, et al. Bisphenol analogues other than BPA: environmental occurrence, human exposure, and toxicity-a review. Environ Sci Technol 2016;50:5438-53.
4. der Veen I, de Boer J. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis. Chemosphere 2012;88:1119-53.
5. Li Q, Yang K, Li K, et al. New halogenated flame retardants in the atmosphere of nine urban areas in China: Pollution characteristics, source analysis and variation trends. Environ Pollut 2017;224:679-88.
6. Ezechiáš M, Covino S, Cajthaml T. Ecotoxicity and biodegradability of new brominated flame retardants: a review. Ecotoxicol Environ Saf 2014;110:153-67.
7. Chen SJ, Feng AH, He MJ, Chen MY, Luo XJ, Mai BX. Current levels and composition profiles of PBDEs and alternative flame retardants in surface sediments from the Pearl River Delta, southern China: comparison with historical data. Sci Total Environ 2013;444:205-11.
8. Sunday OE, Bin H, Guanghua M, et al. Review of the environmental occurrence, analytical techniques, degradation and toxicity of TBBPA and its derivatives. Environ Res 2022;206:112594.
9. Xiong P, Yan X, Zhu Q, et al. A Review of environmental occurrence, fate, and toxicity of novel brominated flame retardants. Environ Sci Technol 2019;53:13551-69.
10. Ceresana. Marktstudie kunststoff-additive, 2019. Available from: https://www.ceresana.com/de/marktstudien/chemikalien/kunststoff-additive/ceresana-marktstudie-kunststoff-additive.html [Last accessed on 26 Apr 2023].
11. Hou R, Lin L, Li H, et al. Occurrence, bioaccumulation, fate, and risk assessment of novel brominated flame retardants (NBFRs) in aquatic environments - a critical review. Water Res 2021;198:117168.
12. Reemtsma T, Quintana JB, Rodil R, Garcı´a-lópez M, Rodrı´guez I. Organophosphorus flame retardants and plasticizers in water and air I. Occurrence and fate. TRAC-Trend Anal Chem 2008;27:727-37.
13. Chen Y, Liu Q, Ma J, Yang S, Wu Y, An Y. A review on organophosphate flame retardants in indoor dust from China: Implications for human exposure. Chemosphere 2020;260:127633.
14. Zuiderveen EAR, Slootweg JC, de Boer J. Novel brominated flame retardants - a review of their occurrence in indoor air, dust, consumer goods and food. Chemosphere 2020;255:126816.
15. Möller A, Xie Z, Cai M, et al. Polybrominated diphenyl ethers
16. Kung H, Hsieh Y, Huang B, Cheruiyot NK, Chang-chien G. An overview: organophosphate flame retardants in the atmosphere. Aerosol Air Qual Res 2022;22:220148.
17. Li WL, Ma WL, Zhang ZF, et al. Occurrence and Source Effect of Novel Brominated Flame Retardants (NBFRs) in soils from five asian countries and their relationship with PBDEs. Environ Sci Technol 2017;51:11126-35.
18. McGrath TJ, Ball AS, Clarke BO. Critical review of soil contamination by polybrominated diphenyl ethers (PBDEs) and novel brominated flame retardants (NBFRs); concentrations, sources and congener profiles. Environ Pollut 2017;230:741-57.
19. Law K, Halldorson T, Danell R, et al. Bioaccumulation and trophic transfer of some brominated flame retardants in a Lake Winnipeg (Canada) food web. Environ Toxicol Chem 2006;25:2177-86.
20. He MJ, Luo XJ, Chen MY, Sun YX, Chen SJ, Mai BX. Bioaccumulation of polybrominated diphenyl ethers and decabromodiphenyl ethane in fish from a river system in a highly industrialized area, South China. Sci Total Environ 2012;419:109-15.
21. Zhang L, Lu L, Zhu W, et al. Organophosphorus flame retardants (OPFRs) in the seawater and sediments of the Qinzhou Bay, Northern Beibu Gulf: occurrence, distribution, and ecological risks. Mar Pollut Bull 2021;168:112368.
22. Chen M, Gan Z, Qu B, Chen S, Dai Y, Bao X. Temporal and seasonal variation and ecological risk evaluation of flame retardants in seawater and sediments from Bohai Bay near Tianjin, China during 2014 to 2017. Mar Pollut Bull 2019;146:874-83.
23. Gao X, Lin Y, Li J, Xu Y, Qian Z, Lin W. Spatial pattern analysis reveals multiple sources of organophosphorus flame retardants in coastal waters. J Hazard Mater 2021;417:125882.
24. Hou L, Jiang J, Gan Z, et al. Spatial distribution of organophosphorus and brominated flame retardants in surface water, sediment, groundwater, and wild fish in Chengdu, China. Arch Environ Contam Toxicol 2019;77:279-90.
25. Shi T, Chen SJ, Luo XJ, et al. Occurrence of brominated flame retardants other than polybrominated diphenyl ethers in environmental and biota samples from southern China. Chemosphere 2009;74:910-6.
26. Rodil R, Quintana JB, López-Mahía P, Muniategui-Lorenzo S, Prada-Rodríguez D. Multi-residue analytical method for the determination of emerging pollutants in water by solid-phase extraction and liquid chromatography-tandem mass spectrometry. J Chromatogr A 2009;1216:2958-69.
27. Martínez-Carballo E, González-Barreiro C, Sitka A, Scharf S, Gans O. Determination of selected organophosphate esters in the aquatic environment of Austria. Sci Total Environ 2007;388:290-9.
28. Woudneh MB, Benskin JP, Wang G, Grace R, Hamilton MC, Cosgrove JR. Quantitative determination of 13 organophosphorous flame retardants and plasticizers in a wastewater treatment system by high performance liquid chromatography tandem mass spectrometry. J Chromatogr A 2015;1400:149-55.
29. Shi Y, Gao L, Li W, Wang Y, Liu J, Cai Y. Occurrence, distribution and seasonal variation of organophosphate flame retardants and plasticizers in urban surface water in Beijing, China. Environ Pollut 2016;209:1-10.
30. Xu L, Hu Q, Liu J, et al. Occurrence of organophosphate esters and their diesters degradation products in industrial wastewater treatment plants in China: Implication for the usage and potential degradation during production processing. Environ Pollut 2019;250:559-66.
31. Greaves AK, Su G, Letcher RJ. Environmentally relevant organophosphate triesters in herring gulls:
32. Kim JW, Isobe T, Chang KH, et al. Levels and distribution of organophosphorus flame retardants and plasticizers in fishes from Manila Bay, the Philippines. Environ Pollut 2011;159:3653-9.
33. Sundkvist AM, Olofsson U, Haglund P. Organophosphorus flame retardants and plasticizers in marine and fresh water biota and in human milk. J Environ Monit 2010;12:943-51.
34. Su G, Letcher RJ, Moore JN, et al. Spatial and temporal comparisons of legacy and emerging flame retardants in herring gull eggs from colonies spanning the Laurentian Great Lakes of Canada and United States. Environ Res 2015;142:720-30.
35. Sala B, Giménez J, de Stephanis R, Barceló D, Eljarrat E. First determination of high levels of organophosphorus flame retardants and plasticizers in dolphins from Southern European waters. Environ Res 2019;172:289-95.
36. McGoldrick DJ, Letcher RJ, Barresi E, et al. Organophosphate flame retardants and organosiloxanes in predatory freshwater fish from locations across Canada. Environ Pollut 2014;193:254-61.
37. Giulivo M, Capri E, Kalogianni E, et al. Occurrence of halogenated and organophosphate flame retardants in sediment and fish samples from three European river basins. Sci Total Environ 2017;586:782-91.
38. Hou R, Liu C, Gao X, Xu Y, Zha J, Wang Z. Accumulation and distribution of organophosphate flame retardants (PFRs) and their di-alkyl phosphates (DAPs) metabolites in different freshwater fish from locations around Beijing, China. Environ Pollut 2017;229:548-56.
39. Wang X, Zhong W, Xiao B, et al. Bioavailability and biomagnification of organophosphate esters in the food web of Taihu Lake, China: Impacts of chemical properties and metabolism. Environ Int 2019;125:25-32.
40. Bekele TG, Zhao H, Wang Y, Jiang J, Tan F. Measurement and prediction of bioconcentration factors of organophosphate flame retardants in common carp (Cyprinus carpio). Ecotoxicol Environ Saf 2018;166:270-6.
41. Zheng G, Wan Y, Shi S, et al. Trophodynamics of emerging brominated flame retardants in the aquatic food web of lake taihu: relationship with organism metabolism across trophic levels. Environ Sci Technol 2018;52:4632-40.
42. Liu YE, Tang B, Liu Y, et al. Occurrence, biomagnification and maternal transfer of legacy and emerging organophosphorus flame retardants and plasticizers in water snake from an e-waste site. Environ Int 2019;133:105240.
43. Liu AF, Qu GB, Yu M, Liu YW, Shi JB, Jiang GB. Tetrabromobisphenol-A/S and nine novel analogs in biological samples from the chinese bohai sea: implications for trophic transfer. Environ Sci Technol 2016;50:4203-11.
44. Hou R, Huang Q, Pan Y, et al. Novel brominated flame retardants (NBFRs) in a Tropical marine food web from the south china sea: the influence of hydrophobicity and biotransformation on structure-related trophodynamics. Environ Sci Technol 2022;56:3147-58.
45. Bekele TG, Zhao H, Wang Q, Chen J. Bioaccumulation and Trophic Transfer of Emerging Organophosphate Flame Retardants in the Marine Food Webs of Laizhou Bay, North China. Environ Sci Technol 2019;53:13417-26.
46. Ding Y, Han M, Wu Z, et al. Bioaccumulation and trophic transfer of organophosphate esters in tropical marine food web, South China Sea. Environ Int 2020;143:105919.
47. Giraudo M, Douville M, Houde M. Chronic toxicity evaluation of the flame retardant tris (2-butoxyethyl) phosphate (TBOEP) using Daphnia magna transcriptomic response. Chemosphere 2015;132:159-65.
48. Yan S, Wu H, Qin J, Zha J, Wang Z. Halogen-free organophosphorus flame retardants caused oxidative stress and multixenobiotic resistance in Asian freshwater clams (Corbicula fluminea). Environ Pollut 2017;225:559-68.
49. Xu Q, Wu D, Dang Y, Yu L, Liu C, Wang J. Reproduction impairment and endocrine disruption in adult zebrafish (Danio rerio) after waterborne exposure to TBOEP. Aquat Toxicol 2017;182:163-71.
50. Chen R, Hong X, Yan S, Zha J. Three organophosphate flame retardants (OPFRs) reduce sperm quality in Chinese rare minnows (Gobiocypris rarus). Environ Pollut 2020;263:114525.
51. Liu Y, Wu D, Xu Q, Yu L, Liu C, Wang J. Acute exposure to tris (2-butoxyethyl) phosphate (TBOEP) affects growth and development of embryo-larval zebrafish. Aquat Toxicol 2017;191:17-24.
52. Ma Z, Tang S, Su G, et al. Effects of tris (2-butoxyethyl) phosphate (TBOEP) on endocrine axes during development of early life stages of zebrafish (Danio rerio). Chemosphere 2016;144:1920-7.
53. Tran CM, Lee H, Lee B, Ra JS, Kim KT. Effects of the chorion on the developmental toxicity of organophosphate esters in zebrafish embryos. J Hazard Mater 2021;401:123389.
54. Jiang F, Liu J, Zeng X, Yu L, Liu C, Wang J. Tris (2-butoxyethyl) phosphate affects motor behavior and axonal growth in zebrafish (Danio rerio) larvae. Aquat Toxicol 2018;198:215-23.
55. Jiang X, Yang Y, Liu P, Li M. Transcriptomics and metabolomics reveal Ca2+ overload and osmotic imbalance-induced neurotoxicity in earthworms (Eisenia fetida) under tri-n-butyl phosphate exposure. Sci Total Environ 2020;748:142169.
56. Sun L, Xu W, Peng T, et al. Developmental exposure of zebrafish larvae to organophosphate flame retardants causes neurotoxicity. Neurotoxicol Teratol 2016;55:16-22.
57. Hong X, Chen R, Yuan L, Zha J. Global microRNA and isomiR expression associated with liver metabolism is induced by organophosphorus flame retardant exposure in male Chinese rare minnow (Gobiocypris rarus). Sci Total Environ 2019;649:829-38.
58. Liu X, Ji K, Choi K. Endocrine disruption potentials of organophosphate flame retardants and related mechanisms in H295R and MVLN cell lines and in zebrafish. Aquat Toxicol 2012;114-115:173-81.
59. Dong L, Wang S, Qu J, You H, Liu D. New understanding of novel brominated flame retardants (NBFRs): Neuro(endocrine) toxicity. Ecotoxicol Environ Saf 2021;208:111570.
60. Harju M, Heimstad E S, D H. Current state of knowledge and monitoring requirements: emerging “new” brominated flame retardants in flame retarded products and the environment. Available from: https://hdl.handle.net/11250/2718681 [Last accessed on 26 Apr 2023].
61. Tennekes HA, Sánchez-Bayo F. The molecular basis of simple relationships between exposure concentration and toxic effects with time. Toxicology 2013;309:39-51.
62. Li X, Zhang Q, Wang P, Fu J, Jiang G. Post dioxin period for feed: cocktail effects of emerging POPs and analogues. Environ Sci Technol 2020;54:6-8.
63. Papachlimitzou A, Barber JL, Losada S, Bersuder P, Law RJ. A review of the analysis of novel brominated flame retardants. J Chromatogr A 2012;1219:15-28.
64. Chokwe TB, Abafe OA, Mbelu SP, Okonkwo JO, Sibali LL. A review of sources, fate, levels, toxicity, exposure and transformations of organophosphorus flame-retardants and plasticizers in the environment. Emerging Contaminants 2020;6:345-66.
65. Liu Y, Gong S, Ye L, et al. Organophosphate (OP) diesters and a review of sources, chemical properties, environmental occurrence, adverse effects, and future directions. Environ Int 2021;155:106691.
66. Yan Z, Feng C, Leung KMY, et al. Insights into the geographical distribution, bioaccumulation characteristics, and ecological risks of organophosphate esters. J Hazard Mater 2023;445:130517.
67. Zhang Q, Wang Y, Zhang C, Yao Y, Wang L, Sun H. A review of organophosphate esters in soil: implications for the potential source, transfer, and transformation mechanism. Environ Res 2022;204:112122.
68. Hou R, Xu Y, Wang Z. Review of OPFRs in animals and humans: absorption, bioaccumulation, metabolism, and internal exposure research. Chemosphere 2016;153:78-90.
69. Smythe TA, Su G, Bergman Å, Letcher RJ. Metabolic transformation of environmentally-relevant brominated flame retardants in Fauna: a review. Environ Int 2022;161:107097.
70. Zhang Q, Yao Y, Wang Y, et al. Plant accumulation and transformation of brominated and organophosphate flame retardants: a review. Environ Pollut 2021;288:117742.
71. Yang Y, Chen P, Ma S, Lu S, Yu Y, An T. A critical review of human internal exposure and the health risks of organophosphate ester flame retardants and their metabolites. Crit Rev Environ Sci Technol 2022;52:1528-60.
72. Su G, Crump D, Letcher RJ, Kennedy SW. Rapid
73. Marteinson S, Guigueno MF, Fernie KJ, Head JA, Chu S, Letcher RJ. Uptake, deposition, and metabolism of triphenyl phosphate in embryonated eggs and chicks of Japanese Quail (Coturnix japonica). Environ Toxicol Chem 2020;39:565-73.
74. Briels N, Løseth ME, Ciesielski TM, et al.
75. Farhat A, Crump D, Porter E, et al. Time-dependent effects of the flame retardant tris(1,3-dichloro-2-propyl) phosphate (TDCPP) on mRNA expression,
76. Yin Y, Zhao N, Yifei L, et al. Deposition, bioaccumulation and depletion of organophosphate triesters (tri-OPEs) and their organophosphate diester metabolites (di-OPEs) from feed to laying hens’ eggs. J Hazard Mater 2022;440:129858.
77. Herczegh SM, Chu S, Letcher RJ. Biotransformation of bisphenol-A bis(diphenyl phosphate):
78. Wang G, Du Z, Chen H, Su Y, Gao S, Mao L. Tissue-specific accumulation, depuration, and transformation of triphenyl phosphate (TPHP) in adult zebrafish (Danio rerio). Environ Sci Technol 2016;50:13555-64.
79. Wang G, Shi H, Du Z, Chen H, Peng J, Gao S. Bioaccumulation mechanism of organophosphate esters in adult zebrafish (Danio rerio). Environ Pollut 2017;229:177-87.
80. Hou R, Huang C, Rao K, Xu Y, Wang Z. Characterized
81. Hou R, Xu Y, Rao K, Feng C, Wang Z. Tissue-specific bioaccumulation, metabolism and excretion of tris (2-ethylhexyl) phosphate (TEHP) in rare minnow (Gobiocyprisrarus). Environ Pollut 2020;261:114245.
82. Hou R, Yuan S, Feng C, Xu Y, Rao K, Wang Z. Toxicokinetic patterns, metabolites formation and distribution in various tissues of the Chinese rare minnow (Gobiocypris rarus) exposed to tri(2‑butoxyethyl) phosphate (TBOEP) and tri-n-butyl phosphate (TNBP). Sci Total Environ 2019;668:806-14.
83. Tang B, Poma G, Bastiaensen M, et al. Bioconcentration and biotransformation of organophosphorus flame retardants (PFRs) in common carp (Cyprinus carpio). Environ Int 2019;126:512-22.
84. Yang R, Ye Y, Chen Y, et al. First Insight into the formation of
85. Yan Z, Feng C, Jin X, et al.
86. Xu T, Wang Q, Shi Q, Fang Q, Guo Y, Zhou B. Bioconcentration, metabolism and alterations of thyroid hormones of Tris(1,3-dichloro-2-propyl) phosphate (TDCPP) in Zebrafish. Environ Toxicol Pharmacol 2015;40:581-6.
87. Chen M, Liao X, Yan SC, et al. Uptake, Accumulation, and biomarkers of pm2.5-associated organophosphate flame retardants in C57BL/6 mice after chronic exposure at real environmental concentrations. Environ Sci Technol 2020;54:9519-28.
88. Chen MH, Zhang SH, Jia SM, Wang LJ, Ma WL.
89. Choi Y, Jeon J, Choi Y, Kim SD. Characterizing biotransformation products and pathways of the flame retardant triphenyl phosphate in Daphnia magna using non-target screening. Sci Total Environ 2020;708:135106.
90. Choi Y, Jeon J, Kim SD. Identification of biotransformation products of organophosphate ester from various aquatic species by suspect and non-target screening approach. Water Res 2021;200:117201.
91. Liu W, Zhang H, Ding J, He W, Zhu L, Feng J. Waterborne and dietary bioaccumulation of organophosphate esters in zooplankton daphnia magna. Int J Environ Res Public Health 2022;19:9382.
92. Mata MC, Castro V, Quintana JB, Rodil R, Beiras R, Vidal-Liñán L. Bioaccumulation of organophosphorus flame retardants in the marine mussel Mytilus galloprovincialis. Sci Total Environ 2022;805:150384.
93. Wang L, Huang X, Laserna AKC, Li SFY. Metabolism of tri-n-butyl phosphate in earthworm Perionyx excavatus. Environ Pollut 2018;234:389-95.
94. Wang L, Huang X, Laserna AKC, Li SFY. Untargeted metabolomics reveals transformation pathways and metabolic response of the earthworm Perionyx excavatus after exposure to triphenyl phosphate. Sci Rep 2018;8:16440.
95. Wu X, Zhu Y, Yang M, Zhang J, Lin D. Biological responses of Eisenia fetida towards the exposure and metabolism of tris (2-butoxyethyl) phosphate. Sci Total Environ 2022;811:152285.
96. Stubbings WA, Guo J, Simon K, Romanak K, Bowerman W, Venier M. Flame retardant metabolites in addled bald eagle eggs from the Great Lakes Region. Environ Sci Technol Lett 2018;5:354-9.
97. Li ZR, Luo XJ, Luo YL, Zeng YH, Mai BX. Comparative study of dechlorane plus (DP) in adult chickens and developing embryos: Stereo-selective bioaccumulation of DP in chickens. Environ Pollut 2019;247:550-5.
98. Eng ML, Karouna-Renier NK, Henry PFP, et al.
99. Goodchild CG, Karouna-Renier NK, Braham RP, Henry PFP, Letcher RJ, Fernie KJ. Hepatic gene expression profiling of american kestrels (Falco sparverius) exposed
100. Goodchild C, Karouna-Renier NK, Henry PFP, et al. Thyroid disruption and oxidative stress in American kestrels following embryonic exposure to the alternative flame retardants, EHTBB and TBPH. Environ Int 2021;157:106826.
101. McKinney MA, Dietz R, Sonne C, et al. Comparative hepatic microsomal biotransformation of selected PBDEs, including decabromodiphenyl ether, and decabromodiphenyl ethane flame retardants in Arctic marine-feeding mammals. Environ Toxicol Chem 2011;30:1506-14.
102. Wang X, Ling S, Guan K, et al. Bioconcentration, biotransformation, and thyroid endocrine disruption of decabromodiphenyl ethane (dbdpe), a novel brominated flame retardant, in Zebrafish Larvae. Environ Sci Technol 2019;53:8437-46.
103. Wang X, Sun Y, Fu M, et al. Nano-TiO2 adsorbed decabromodiphenyl ethane and changed its bioavailability, biotransformation and biotoxicity in zebrafish embryos/larvae. Front Environ Sci 2022;10:860786.
104. Ganci AP, Abdallah MA, Nguyen KH, et al. Investigating the in vitro metabolism of NBFRs by trout liver microsomes using a high resolution accurate mass benchtop Q-Exactive Orbitrap mass spectrometer. 2017. Available from: https://assets.thermofisher.com/TFS-Assets/CMD/posters/PO-HRAM-MS-Metabolism-NBFRS-BFR2017-EN.pdf [Last accessed on 26 Apr 2023].
105. Qiao Z, Wang Y, Lu C, et al. Environmental fate of five brominated flame retardants co-exposure in a water-sediment-zebrafish microcosm system: Enrichment, removal, and metabolism mechanisms. J Clean Prod 2023;387:135916.
106. Jourdan BP, Hanson ML, Muir DC, Solomon KR. Fathead minnow (Pimephales promelas Rafinesque) exposure to three novel brominated flame retardants in outdoor mesocosms: bioaccumulation and biotransformation. Environ Toxicol Chem 2014;33:1148-55.
107. Bearr JS, Mitchelmore CL, Roberts SC, Stapleton HM. Species specific differences in the
108. Tang B, Luo XJ, Huang CC, et al. Stereoselective bioaccumulation of syn- and anti-Dechlorane plus isomers in different tissues of common carp (Cyprinus carpio). Sci Total Environ 2018;616-617:1339-46.
109. Tomy GT, Thomas CR, Zidane TM, et al. Examination of isomer specific bioaccumulation parameters and potential
110. Yang S, Gu S, Tang B, et al. Tissue-specific and stereoselective accumulation of Dechlorane Plus isomers in two predator fish in a laboratory feeding study. Ecotox Environ Safe 2023;249:114469.
111. Lee HJ, Jung JH, Kwon JH. Evaluation of the bioaccumulation potential of selected alternative brominated flame retardants in marine fish using
112. Tomy GT, Palace VP, Pleskach K, et al. Dietary exposure of juvenile rainbow trout (Oncorhynchus mykiss) to 1,2-bis(2,4,6-tribromophenoxy)ethane: bioaccumulation parameters, biochemical effects, and metabolism. Environ Sci Technol 2007;41:4913-8.
113. Nyholm JR, Norman A, Norrgren L, Haglund P, Andersson PL. Uptake and biotransformation of structurally diverse brominated flame retardants in zebrafish (Danio rerio) after dietary exposure. Environ Toxicol Chem 2009;28:1035-42.
114. Nacci D, Clark B, La Guardia MJ, et al. Bioaccumulation and effects of dietary exposure to the alternative flame retardant, bis(2-ethylhexyl) tetrabromophthalate (TBPH), in the Atlantic killifish, Fundulus heteroclitus. Environ Toxicol Chem 2018;37:2350-60.
115. Wang F, Wang J, Dai J, et al. Comparative tissue distribution, biotransformation and associated biological effects by decabromodiphenyl ethane and decabrominated diphenyl ether in male rats after a 90-day oral exposure study. Environ Sci Technol 2010;44:5655-60.
116. Hakk H, Larsen G, Bowers J. Metabolism, tissue disposition, and excretion of 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE) in male Sprague-Dawley rats. Chemosphere 2004;54:1367-74.
117. Silva MJ, Hilton D, Furr J, et al. Quantification of tetrabromo benzoic acid and tetrabromo phthalic acid in rats exposed to the flame retardant Uniplex FPR-45. Arch Toxicol 2016;90:551-7.
118. Roberts SC, Macaulay LJ, Stapleton HM.
119. Knudsen GA, Sanders JM, Birnbaum LS. Disposition of the emerging brominated flame retardant, bis(2-ethylhexyl) tetrabromophthalate, in female Sprague Dawley rats: effects of dose, route and repeated administration. Xenobiotica 2017;47:245-54.
120. Phillips AL, Chen A, Rock KD, Horman B, Patisaul HB, Stapleton HM. Editor’s highlight: transplacental and lactational transfer of firemaster® 550 components in dosed wistar rats. Toxicol Sci 2016;153:246-57.
121. Wang Y, Ling S, Lu C, et al. Exploring the environmental fate of novel brominated flame retardants in a sediment-water-mudsnail system: Enrichment, removal, metabolism and structural damage. Environ Pollut 2020;265:114924.
122. Zhou S, Fu M, Luo K, et al. Fate and toxicity of legacy and novel brominated flame retardants in a sediment-water-clam system: bioaccumulation, elimination, biotransformation and structural damage. Sci Total Environ 2022;840:156634.
123. Qiao Z, Lu C, Han Y, et al. Enrichment and removal of five brominated flame retardants in the presence of co-exposure in a soil-earthworm system. Environ Pollut 2022;310:119877.
124. Su G, Greaves AK, Gauthier L, Letcher RJ. Liquid chromatography-electrospray-tandem mass spectrometry method for determination of organophosphate diesters in biotic samples including Great Lakes herring gull plasma. J Chromatogr A 2014;1374:85-92.
125. Huang Q, Hou R, Lin L, et al. Bioaccumulation and trophic transfer of organophosphate flame retardants and their metabolites in the estuarine food web of the Pearl River, China. Environ Sci Technol 2023;57:3549-61.
126. Yao S, Shi Z, Cao P, et al. A global survey of organophosphate esters and their metabolites in milk: Occurrence and dietary intake via milk consumption. J Hazard Mater 2023;442:130080.
127. Li X, Zhao N, Fu J, et al. Organophosphate diesters (Di-OPEs) play a critical role in understanding global organophosphate esters (OPEs) in fishmeal. Environ Sci Technol 2020;54:12130-41.
128. He C, Wang X, Tang S, et al. Concentrations of organophosphate esters and their specific metabolites in food in southeast Queensland, Australia: Is dietary exposure an important pathway of organophosphate esters and their metabolites? Environ Sci Technol 2018;52:12765-73.
129. Li Z, He C, Thai P, et al. Organophosphate esters and their specific metabolites in chicken eggs from across Australia: Occurrence, profile, and distribution between yolk and albumin fractions. Environ Pollut 2020;262:114260.
130. Li Y, Li J, Fu X. Analysis and occurrence of organophosphorus flame retardants and their metabolites in animal derived food. Researchsquare 2022:preprint.
131. Zhao N, Fu J, Liu Y, Wang P, Su X, Li X. Be aware of organophosphate diesters as direct sources in addition to organophosphate ester metabolites in food supplies. J Agric Food Chem 2021;69:1283-90.
132. Kojima H, Takeuchi S, Van den Eede N, Covaci A. Effects of primary metabolites of organophosphate flame retardants on transcriptional activity via human nuclear receptors. Toxicol Lett 2016;245:31-9.
133. Li Y, Kang Q, Chen R, et al. 2-Ethylhexyl diphenyl phosphate and its hydroxylated metabolites are anti-androgenic and cause adverse reproductive outcomes in male Japanese medaka (Oryzias latipes). Environ Sci Technol 2020;54:8919-25.
134. Selmi-Ruby S, Marín-Sáez J, Fildier A, et al.
135. Chen Q, Lian X, An J, et al. Life cycle exposure to environmentally relevant concentrations of diphenyl phosphate (DPhP) inhibits growth and energy metabolism of zebrafish in a sex-specific manner. Environ Sci Technol 2021;55:13122-31.
136. Lee JS, Kawai YK, Morita Y, Covaci A, Kubota A. Estrogenic and growth inhibitory responses to organophosphorus flame retardant metabolites in zebrafish embryos. Comp Biochem Physiol C Toxicol Pharmacol 2022;256:109321.
137. Zhang Q, Yu C, Fu L, Gu S, Wang C. New Insights in the endocrine disrupting effects of three primary metabolites of organophosphate flame retardants. Environ Sci Technol 2020;54:4465-74.
138. Gramec Skledar D, Tomašič T, Carino A, Distrutti E, Fiorucci S, Peterlin Mašič L. New brominated flame retardants and their metabolites as activators of the pregnane X receptor. Toxicol Lett 2016;259:116-23.
139. Klopčič I, Skledar DG, Mašič LP, Dolenc MS. Comparison of
140. Chen Y, Guo M, Liu R, Ma LQ, Cui X. Effects of novel brominated flame retardants and metabolites on cytotoxicity in human umbilical vein endothelial cells. Chemosphere 2020;253:126653.
141. Fang M, Webster TF, Ferguson PL, Stapleton HM. Characterizing the peroxisome proliferator-activated receptor (PPARγ) ligand binding potential of several major flame retardants, their metabolites, and chemical mixtures in house dust. Environ Health Perspect 2015;123:166-72.
142. Yang M, Zhang X. Comparative developmental toxicity of new aromatic halogenated DBPs in a chlorinated saline sewage effluent to the marine polychaete Platynereis dumerilii. Environ Sci Technol 2013;47:10868-76.
143. Zheng G, Miller P, von Hippel FA, Buck CL, Carpenter DO, Salamova A. Legacy and emerging semi-volatile organic compounds in sentinel fish from an arctic formerly used defense site in Alaska. Environ Pollut 2020;259:113872.
144. Strobel A, Willmore WG, Sonne C, Dietz R, Letcher RJ. Organophosphate esters in East Greenland polar bears and ringed seals: adipose tissue concentrations and
145. Hidalgo-Serrano M, Borrull F, Pocurull E, Marcé RM. Determination of organophosphate ester metabolites in seafood species by QuEChERS-SPE followed by LC-HRMS. Molecules 2022;27:8635.
146. Chen X, Zhang N, Li L, et al. A simple method for simultaneous determination of organophosphate esters and their diester metabolites in dairy products and human milk by using solid-phase extraction coupled to liquid chromatography-tandem mass spectrometry. Anal Bioanal Chem 2022;414:4255-65.
147. Farhat A, Crump D, Chiu S, et al.
148. Wang G, Chen H, Du Z, Li J, Wang Z, Gao S.
149. Giraudo M, Douville M, Letcher RJ, Houde M. Effects of food-borne exposure of juvenile rainbow trout (Oncorhynchus mykiss) to emerging brominated flame retardants 1,2-bis(2,4,6-tribromophenoxy)ethane and 2-ethylhexyl-2,3,4,5-tetrabromobenzoate. Aquat Toxicol 2017;186:40-9.
150. MacFarland HN, Punte CL Jr. Toxicological studies on tri-(2-ethylhexyl)-phosphate. Arch Environ Health 1966;13:13-20.
151. Nomeir AA, Kato S, Matthews HB. The metabolism and disposition of tris(1,3-dichloro-2-propyl) phosphate (Fyrol FR-2) in the rat. Toxicol Appl Pharmacol 1981;57:401-13.
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Hou R, Sun C, Zhang S, Huang Q, Liu S, Lin L, Li H, Xu X. The metabolism of novel flame retardants and the internal exposure and toxicity of their major metabolites in fauna - a review. J Environ Expo Assess 2023;2:10. http://dx.doi.org/10.20517/jeea.2023.08
AMA Style
Hou R, Sun C, Zhang S, Huang Q, Liu S, Lin L, Li H, Xu X. The metabolism of novel flame retardants and the internal exposure and toxicity of their major metabolites in fauna - a review. Journal of Environmental Exposure Assessment. 2023; 2(2): 10. http://dx.doi.org/10.20517/jeea.2023.08
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Hou, Rui, Chuansheng Sun, Siqi Zhang, Qianyi Huang, Shan Liu, Lang Lin, Hengxiang Li, Xiangrong Xu. 2023. "The metabolism of novel flame retardants and the internal exposure and toxicity of their major metabolites in fauna - a review" Journal of Environmental Exposure Assessment. 2, no.2: 10. http://dx.doi.org/10.20517/jeea.2023.08
ACS Style
Hou, R.; Sun C.; Zhang S.; Huang Q.; Liu S.; Lin L.; Li H.; Xu X. The metabolism of novel flame retardants and the internal exposure and toxicity of their major metabolites in fauna - a review. J. Environ. Expo. Assess. 2023, 2, 10. http://dx.doi.org/10.20517/jeea.2023.08
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