The metabolism of novel flame retardants and the internal exposure and toxicity of their major metabolites in fauna - a review
Abstract
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.
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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
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