Levels of Non-PBDE Halogenated Fire Retardants and Brominated Dioxins and their Toxicological Effects in Indoor Environments-A Review

Non-polybrominated diphenyl ether (non-PBDE) halogenated fire retardants (HFRs) such as new or novel brominated fire retardants (NBFRs) and dechlorane plus (DP) have been widely spread in the environment and recognized as emerging persistent organic pollutants (POPs) in the recent years, mainly due to the continuous increase in their global demand, especially after the worldwide restrictions on PBDE use. Polybrominated dibenzo-p-dioxins/furans (PBDD/Fs) are the unintentional byproducts of PBDE commercial formulations in the indoor environment. Although HFRs, including NBFRs, DP, and PBDD/Fs, are ubiquitous in the indoor environment due to the large-volume release from the surfaces of consumer products, only a few in vitro and in vivo studies have addressed their toxic effects. In this review article, global data on NBFRs, including decabromodiphenyl ethane (DBDPE), 1,2-bis(2,4,6-tribromophenoxy) ethane (BTBPE), bis(2-ethylhexyl)-3,4,5,6-tetrabromophthalate (BEH-TEBP), and 2-ethylhexyl-2,3,4,5-tetrabromobenzoate (EH-TBB), DP, including syn-DP and anti-DP, and PBDD/Fs in indoor aerosol and dust are summarized from recent literature. Based on the gathered data, indoor dust is a major sink for indoor contamination and is of great concern due to the fact that dust ingestion is one of the primary routes for human exposure to these chemicals. Lastly, the toxic effects of NBFRs, DP, and PBDD/Fs identified in in vitro and in vivo studies are summarized and discussed based on the current published reports. However, there is still a lack of sufficient toxicity data for assessing their risks. Future works are encouraged to focus on indoor PM2.5-bound HFR levels to further evaluate their toxic effects on human health.


INTRODUCTION
Halogenated fire retardants (HFRs), including polybrominated diphenyl ethers (PBDEs), dechlorane plus (DP), and new or novel brominated fire retardants (NBFRs), as well as polybrominated dibenzo-p-dioxins/furans (PBDD/Fs) are globally emerging persistent organic pollutants (POPs) in the indoor environment especially in the case of indoor fine particulates (PM 2.5 ) and indoor dust.PBDEs, DP, and NBFRs are emitted from various electronic products, building materials, furniture, mattresses, carpet pads, and textiles (Alaee et al., 2003;Covaci et al., 2011;Sverko et al., 2011;Chao et al., 2014a;Redfern et al., 2017).The commercial formulation of octa-BDEs and penta-BDEs have been listed in the United Nations Stockholm Convention on Persistent Organic Pollutants Annex A in 2009 (Stockholm Convention, 2016) and were banned in the European Union (EU) and has been phased out in the US market on a voluntary basis since 2004 (Costa and Giordano, 2007).Decabromodiphenyl ether  and DP may possibly be persistent and bio-accumulative in the biota and environment through global long-range transport of air pollutants (LRTAP).These two chemicals may be candidates for Annex A and D, respectively, and are under either review or evaluation by the Stockholm Convention (Stockholm Convention, 2016).The EU and some states in the US are recommending restriction or phase out of deca-BDE use, but deca-BDE is still manufactured and widely used in most countries (Chao et al., 2014a).Owing to current legislative restrictions on PBDEs, the demand for flame retardant alternatives to PBDEs has recently increased.PBDD/Fs are unintentionally released from various textiles, plastics, and building materials combustion processes, and consumer products containing BFRs, and are also present as impurities in technical mixtures of BFRs, particularly in technical mixtures of penta-BDEs, octa-BDEs, and deca-BDEs (Hanari et al., 2006).Certain brominated POPs, including unintentional PBDE-by-products such as PBDD/Fs and flame-retardant alternatives to PBDEs, including NBFRs, are emerging.Thus, monitoring their levels and fate in the indoor environment is of great importance.PBDD/Fs have structural similarity with PBDEs.The process of pyrolysis of certain commercial PBDEs results in the formation and release of PBDD/Fs (Hanari et al., 2006;Ren et al., 2011).Despite the prohibition of the use of PBDEs, alternative BFRs, such as NBFRs, are continuously produced and applied to decrease flammability, delay ignition, and to meet the fire safety requirements for combustible materials.Global demands of NBFRs have increased due to their being replacements for banned PBDE commercial formulations such as decabromodiphenyl ethane (DBDPE) as a replacement for deca-BDE, 1,2-bis(2,4,6-tribromophenoxy) ethane (BTBPE) as a replacement for octa-BDE, and bis(2ethylhexyl)-3,4,5,6-tetrabromophthalate (BEH-TEBP) and 2-ethylhexyl-2,3,4,5-tetrabromobenzoate (EH-TBB) as a replacement for penta-BDE (Ezechiáš et al., 2014;Hassan and Shoeib, 2015;Brits et al., 2016;Yu et al., 2016).In current global studies on HFRs related to environmental contamination, most scientists are interested in PBDE contamination, and very little attention is being paid to DP and NBFRs.There have been a significant number of PBDE studies considering PBDE contamination of foodstuff, in human specimens, soil, indoor dust, and indoor and outdoor air, including gas and total suspended particulate matter (TSP) (Wu et al., 2015;Chou et al., 2016;Gou et al., 2016a).However, only a few studies have focused on PBDEs as respirable fine particulates (PM 2.5 ) (Li et al., 2015b;Chao et al., 2016).Unlike PBDEs, the levels and fate of DP, NBFRs, and PBDD/Fs in indoor air and dust are scarce.Contamination of PBDD/Fs and NBFRs in the indoor environment is emerging based on possibly higher levels of these chemicals indoors than outdoors and the fact that people tend to stay indoors more than 90% of the time.
Exposure to HFRs and their unintentionally released chemicals, PBDD/Fs, occurs in the indoor environment and enters the human body through several routes, including ingestion, inhalation, and dermal contact.Dietary intake and dust ingestion are probably the two main pathways for human exposure to PBDEs (Fernandes et al., 2008;Johnson-Restrepo and Kannan, 2009;Chen et al., 2012).Much attention has been paid to indoor dust, which is a major sink for accumulation of PBDEs and an important non-dietary pathway for human exposure to PBDEs in the indoor environment (Lorber, 2008;Kang et al., 2011;Chao et al., 2014b;Newton et al., 2015).Until now, no reports have shown a non-dietary exposure route of non-PBDE HFRs for humans in the indoor environment.PBDEs and PBDD/Fs possibly have the same environmental fate as DP and NBFRs in the indoor environment based on their similar chemical and physical properties.Currently, most environmental scientists recognize that DP and NBFRs have the same exposure routes and bioaccumulation in the human body as PBDEs.

Epidemiological Effects of NBFRs, DP, and PBDD/Fs
NBFRs, DP, and PBDD/Fs are ubiquitously contaminating and existing in the environment and continuously bioaccumulating in biota and human bodies.Although levels of NBFRs and DP in the environmental and biological samples can be found in large amount of reports, very few studies have been made regarding their toxicity to human health particularly in the case of epidemiological studies (Ezechiáš et al., 2014).For PBDD/Fs, similar with NBFRs and DP, most of the studies providing data for their toxicity focused on in vivo testing and are commonly compared to their chlorinated counterparts.

Acute Toxicity
Scanlan et al. (2015) designed a study of acute toxicity and determined the lethal concentration 50 (LC 50 ) values of several BFRs, including FM-550, in Daphnia magna (daphnia) exposed to BFRs.FM-550's nominal LC 50 value was determined to be 0.0486 mg L -1 , which makes it the second most toxic chemical among the tested BFRs (Scanlan et al., 2015).Studies of acute toxicity in Wistar rats showed that a single dose of 5,000 mg kg -1 FM-550 did not produce lethality, outwardly observable effects, or gross changes at necropsy after 15 days (USEPA, 2015).Experimental data also showed that it was possible for BEH-TEBP to be absorbed after oral exposure with a commercial mixture FM-550 (Patisaul et al., 2013).BEH-TEBP was found in the tissues of pregnant Wistar rats after exposure to the commercial mixture; however, BEH-TEBP was not found in the offspring even though exposure had occurred from gestation to lactation (Patisaul et al., 2013).

Developmental Toxicity
In utero or childhood exposure to FM-550 might cause developmental toxicity.In a study by Bailey and Levin (2015), adolescent Danio rerio (zebrafish) were exposed to 0, 1.0, or 3.0 mg L -1 of FM-550 for 24 h, and larvae at 4 hours post fertilization (hpf) -5 days post fertilization (dpf) were exposed to 0.01, 0.1, or 1.0 mg L -1 of FM-550.The results indicate that the 1.0 mg L -1 dose of FM-550 caused significant malformations in larval zebrafish compared to adolescent zebrafish (Bailey and Levin, 2015).In an in vivo study by Patisaul et al. (2013), elevated body weight at a high-dose exposure (1,000 µg day -1 ) was the most significant developmental effect of FM-550 in Wistar rats.This effect became evident prior to adolescence and continued into adulthood for both sexes.An increase in body weight, especially in the males, was interpreted as a symptom of morbid obesity (Patisaul et al., 2013).

Neurotoxicity
EH-TBB might serve as an endocrine disruptor that disrupts thyroid hormone secretions or causes neurobehavioral and neurological toxicity in the prenatal, postnatal, or childhood periods.Pregnant Wistar rats were given 0 (control), 0.1 (low-dose), or 1 (high-dose) mg kg -1 bw -1 day -1 FM-550 in their diet through gestation day 8 until postnatal day 21 (Patisaul et al., 2013).It was observed that thyroxine (T4) was significantly elevated in the exposed pregnant rats (65% higher than controls both in the highdose and low-dose exposure); however, it was unclear why it increased with exposure, indicating that EH-TBB might be responsible for inhibiting thyroid hormone conjugating system activities responsible for clearing T4 from the body.On the other hand, T4 levels in offspring were significantly decreased (Patisaul et al., 2013).A study by Bailey and Levin (2015) revealed that the most sensitive indicator of FM-550's neurobehavioral toxicity in zebrafish is shoaling behavior, which was greatly affected by all doses of FM-550 following developmental exposure.Another study also showed that perinatal exposure to FM-550 at levels below no-observed-adverse-effect level (NOAEL) of 50 mg kg -1 day -1 has effects on the neurodevelopment of Wistar rats in a sex-specific manner (Baldwin et al., 2017).Behavioral outcomes of Baldwin's study showed heightened anxiety in male rats while hyperactivity was observed in female rats (Baldwin et al., 2017).The contradictory result was found in Patisaul's report, which revealed heightened anxiety in female rats and not in males (Patisaul et al., 2013).Pregnant Fischer rats that were gavaged with doses of 200 mg kg -1 (low dose) and 500 mg kg -1 (high dose) mono(2-ethyhexyl) tetrabromophthalate (TBMEHP), which is a metabolite of BEH-TEBP, showed that serum T3 was significantly reduced in a dose-dependent manner, but no effects were observed on serum T4 (Springer et al., 2012).This decrease was attributed to TBMEHP acting as deiodinase inhibitor, preventing conversion of T4 to T3 (Springer et al., 2012).This result is in accordance with a study by Patisaul et al. (2013), which showed BEH-TEBP to have inhibitory activities towards T4 being converted to T3 (Patisaul et al., 2013).

Reproductive Toxicity
We found several in vitro studies showing that EH-TBB affects reproduction (Saunders et al., 2013;Mankidy et al., 2014).The findings of Mankidy's study showed that EH-TBB did not affect sex-steroid production; however, it affected the synthesis of aldosterone and cortisol by upregulating CY21A2, which is the enzyme responsible for directing substrates away from sex hormone synthesis and towards aldosterone and cortisol synthesis in primary porcine testicular cells (Mankidy et al., 2014).Another study using rat hepatoma cell line H4IIE and human adrenocarcinoma cell line H295R demonstrated EH-TBB's antagonism with both estrogen receptors (ER) and androgen receptors (AR) and elevated steroidogenesis of estrogen (Saunders et al., 2013).In this in vitro study, the anti-androgenic response of EH-TBB was weak, which might have been due to limitations in the dosing concentration.The maximum anti-estrogenic effect (62%) was observed at 0.5 mg L -1 EH-TBB.BEH-TEBP was shown to produce a maximal anti-androgenic effect of 74% at 300 mg L -1 (Saunders et al., 2013).Also, maximum exposure of 30 mg L -1 BEH-TEBP resulted in a moderate 1.96-fold increase in testosterone concentration compared to controls.BEH-TEBP exposure also resulted in the greatest increase of 17-β-estradiol concentrations, eliciting a maximal response of a 5.29-fold change as compared to the controls (Saunders et al., 2013).

Genotoxicity
In vivo exposure to EH-TBB might be associated with DNA or chromosomal damage or differential gene expression.A study by Bearr et al. (2010) showed EH-TBB to induce repairable DNA damage in the liver tissues of Pimephales promelas (fathead minnows) via dietary exposure.During the exposure period, significant increases in DNA strand breaks from the liver cells were observed.However, such increases were not found after the recovery period (Bearr et al., 2010).Scanlan et al. (2015) reported that daphnids (Daphnia magna) that were exposed to five different concentrations of FM-550 (1/2 LC 50 (0.243 mg L -1 ), 1/10 LC 50 (0.0486 mg L -1 ), where LC 50 was equal to 0.0486 mg L -1 , and three additional dilutions: 0.243 µg L -1 , 0.0486 µg L -1 , and 0.0486 ng L -1 were used) caused differential mRNA levels at all concentrations.Exposure at 0.0486 µg L -1 , however, resulted in the largest number of differentially expressed genes.In all five concentrations, three genes (a trichohyalin-like protein, peroxidase, and an unknown protein) were differentially expressed, which may be useful biomarkers of exposure to FM-550 (Scanlan et al., 2015).
Dietary exposure to BEH-TEBP caused a significant increase in percent tail DNA in liver tissues of fathead minnows on days 28 and 56, which was 3.4 and 6.3 times greater than in the controls, respectively (Bearr et al., 2010).Several in vivo studies in Wistar rats exposed to BEH-TEBP through dermal and intraperitoneal injection failed to identify genotoxic effects in the form of micronucleated erythrocytes in the bone marrow (USEPA, 2015).Consistent results were also found in various in vitro studies using isolated human lymphocytes, where no elicited chromosomal aberrations were found (USEPA, 2015).Egloff et al. (2011) indicated that chicken embryonic hepatocytes (CEH) administered with BEH-TEBP did not affect the mRNA expression of any of the genes of interest in CEH and did not induce any toxic effects.
BEH-TEBP and its metabolite, TBMEHP, was shown to cause other toxic effects in in vitro studies.In vitro experiments using murine fatty acid oxidation (FAO) and NIH 3T3 L1 cells showed that TBMEHP activates both PPARα-and PPARγ-mediated gene transcription and is capable of stimulating PPARγ-mediated adipocyte differentiation (Springer et al., 2012).Mankidy et al. (2014) showed that expression of CYP11A1 was 3.5 and 6.6 times greater in primary porcine testicular cells exposed to 0.15 and 15.0 mg L -1 BEH-TEBP, respectively, relative to the controls.Cells exposed to 0.15 mg L -1 BEH-TEBP showed no change in the expression of CYP19A1, but expression of mRNA for CYP19A1 was 3.3 times greater when the cells were exposed to 15 mg L -1 BEH-TEBP (Mankidy et al., 2014).

Toxicity Information in the in vitro and in vivo Studies for BTBPE Acute Toxicity
In a study by Nomeir et al. (1993) of Fischer 344 rats given a diet of 0.05-5% 14 C-labeled BTBPE for one day, a very limited amount of radioactivity of less than 1% of the ingested dose was eliminated in the urine, but of 80-100% of the dose ingested, a high percentage of fecal excretion was observed.In most of the tissues, there were undetectable levels of radiolabeled compounds.It was concluded that BTBPE gastrointestinal absorption is poor in rats.However, it was found that the adipose tissue, kidney, skin, and the thymus contained the highest concentration in rats given a diet of 500 mg kg -1 bw -1 day -1 14 C-labeled BTBPE for a duration of 10 days.In the majority of the tissues, less than 0.01% of the dose was found (Nomeir et al., 1993).In another study by Hakk et al. (2004), male Sprague-Dawley (SD) rats were given a single dose of 2 mg kg -1 bw -1 14 Clabeled BTBPE by gavage.100% of the dose was recovered in the feces.The same research group also demonstrated that elimination of radioactivity by bile is less than 1%, which suggests that fecal elimination was primarily from unabsorbed BTBPE.Due to this low level of absorption, the tissue levels of BTBPE were low.72 h after being given a single dose, more than 0.1% of the dose was found only in the gastrointestinal tract and carcass.No compound-related effects were observed in rats after being fed up to 10% BTBPE in the diet, at an estimated concentration of 35 mg kg -1 bw -1 day -1 , for 14 days (Hakk et al., 2004).Nomeir et al. (1993) also reported that exposure through inhalation of 5 or 20 mg L -1 BTBPE in the atmosphere for 21 days showed no gross pathological changes in Fischer 344 rats; however, unspecified histopathological lesions were observed in the lungs.

Reproductive and Developmental Toxicity
Egloff et al. (2011) studied the reproductive and developmental toxicity of BTBPE in chickens.It was found that BTBPE had no hatching effects and did not cause a delay in embryonic development.

Genotoxicity
Egloff et al. (2011) have also observed differences in β-actin amplification.BTBPE was also found to induce CYP1A4/5 mRNA levels to a maximum of 115-and 18fold at doses greater than or equal to 0.03 µM and 0.1 µM, respectively.Moreover, iodothyronine deiodinase 3 (DIO3) expression was down-regulated in a concentrationdependent manner to a maximum of 2.5-fold following exposure to BTBPE (Egloff et al., 2011).In bacteria from an Ames test and in the yeast, S. cerevisiae, BTBPE was found to be not mutagenic according to the World Health Organization/International Programme on Chemical Safety (WHO/IPCS) evaluation ( 2005) for genotoxicity and carcinogenicity.No information was available for BTBPE on human health endpoints.

Toxic Effects in the in vitro and in vivo Studies for DBDPE Acute Toxicity
In a study by Hardy et al. (2012), DBDPE toxicity was tested on five different aquatic species, including Chironmus riparius (sediment midge), Lumbriculus variegates (sediment oligochaete), Oncorhynchus mykiss (rainbow trout), daphnia, and Pseudokirchneriella subcapitata (algae) at 313, 625, 1,250, 2,500, and 5,000 mg kg -1 dry sediment concentrations of DBDPE.It was observed in midges that 313, 625, and 2,500 mg kg -1 of exposure caused lethargy and loss of equilibrium.Two dead midges were also observed from 2,500 and 5,000 mg kg -1 exposure.Larval midges were also found dead at all concentrations except at the 1,250 mg kg -1 exposure.However, these responses were not concentrationdependent and were therefore considered insignificant.In oligochaetes, no mortalities or sublethal effects were observed at any of the concentrations.No mortality and overt signs of toxicity were observed in trout at all concentrations after 96 hours.It was assumed that one of the replicates might have been contaminated.No toxic effects have been observed in either daphnids or algae (Hardy et al., 2012).In repeated dose studies by Hardy et al. (2002), Hardy et al. (2010), and Hardy et al. (2011), the lowest NOAELs were the highest doses tested: 1,000 mg kg -1 day -1 (SD rats, 90-day) and 1,250 mg kg -1 day -1 (Female Crl:CD ® BR VAF/Plus ® rats and New Zealand White rabbits exposed prenatally).

Genotoxicity
In an in vitro study by Egloff et al. (2011), DBDPE was administered to CEH, and it was found that DBDPE significantly up-regulated the expression of CYP1A4/5 at 0.1 and 0.2 µM to a maximum of 29-and 53-fold, respectively.Also, iodothyronine deiodinase 1 (DIO1) mRNA levels significantly increased in CEH treated with 0.1 µM DBDPE (Egloff et al., 2011).In another study, no clinical signs of toxicity were observed in male SD rats exposed to DBDPE (Wang et al., 2010).It was observed that exposure to 100 mg kg -1 day -1 DBDPE did not cause any significant changes in body, liver, or kidney weight.However, a decrease in creatinine (Cr) levels and aspartate aminotransferase (AST) and alkaline phosphatase (ALP) activities, and elevated total bile acid (TBA) levels indicated that DBDPE exposure induces hepatotoxicity in rats, indicating possible oxidative stress due to the accumulation of DBDPE or its metabolites.An increase in T3 levels was also observed, and it was noted that DBDPE may alter thyroid hormone homeostasis.DBDPE was also shown to significantly increase the expression of CYP3A2 by 1.24fold in the liver tissue of rats (Wang et al., 2010).

Toxic Effects of DP in the in vitro and vivo Studies
Acute Toxicity A study by Dou et al. (2015) investigated the acute toxicity and mutagenicity of DP using luminous bacteria, Vicia faba and Tetrahymena thermophile, as test organisms.The DP concentrations used in the study were 0.591, 2.95, 14.8, 73.8, and 369 µg L -1 for a bioassay measuring light emission, which is representative of the bacteria's cellular metabolism.The luminosities fluctuated at around 100% after exposure to different DP concentrations, which suggests that there is no acute toxicity to luminous bacteria under these concentrations.

Neurotoxicity
An in vivo study by Chen et al. (2017) investigated the developmental neurobehavioral toxicity of DP using embryolarval stages of zebrafish.The embryos were waterborneexposed to DP at 15, 30, and 60 mg L -1 beginning from 6 hpf.Larval teratology, motor activity, motoneuron axonal growth, and muscle morphology were examined at different developmental stages.Reactive oxygen species (ROS) levels and lipid peroxidation levels (LPO) product malonaldehyde (MDA), as well as mRNA transcript expression levels of axonal growth-related and apoptosisrelated genes were also analyzed to elucidate the potential mechanisms of DP-induced developmental neurobehavioral toxicity.The results showed that DP exposure significantly altered embryonic spontaneous movement, reduced touchinduced movement and free-swimming speed, and decreased the swimming speed of larvae in response to dark stimulation.These changes occurred at DP doses that resulted in no significant teratogenesis in zebrafish.Exposure to DP also significantly inhibited axonal growth of primary motoneurons and induced apoptotic cell death and lesions in the muscle fibers of zebrafish.ROS and MDA formation as well as the mRNA transcript levels of apoptosis-related genes bax and caspase-3 were significantly increased at 30 and 60 mg of L -1 DP exposure.

Genotoxicity
DP did not show any signs of overt cytotoxicity in CEH injected with 0.01 and 3 µM DP and did not affect the mRNA expression levels of 11 transcripts, which cover mechanisms such as xenobiotic metabolism, thyroid hormone homeostasis, lipid regulation, and growth (Crump et al., 2011).Available toxicity data from the USEPA HPV ( 2008) and the data presented by Crump et al. (2011) indicate that DP has minimal to no overt toxic effects on test organisms even with dose exposures exceeding its environmental prevalence.In another study, Acipenser sinensis (juvenile Chinese sturgeons) were injected intraperitoneally with DP at doses of 1, 10, and 100 mg kg -1 wet weight -1 (ww) (Liang et al., 2014).Liver proteasomes were collected after 14 days and were analyzed using two-dimensional (2D) electrophoresis.39 protein spots out of the 740 spots that were detected were found to be altered in abundance in more than one DP exposure concentration compared to the controls.The 39 proteins were analyzed using matrixassisted laser desorption/ionization tandem time-of-flight (MALDI-TOF-TOF), and 27 were identified using mass spectrometry (MS).Exposure to DP down-regulated signal transduction-related proteins (Ras-related protein Rab-6B and BAI1-associated protein 2-like 1b), Annexin A4 (ANXA4), and T-complex protein 1 subunit epsilon (CCT5).In contrast, increased abundance of CDHR2 and upregulation of heat shock cognate protein 70 (HSC70) were observed following DP exposure.These differentially expressed proteins may induce cell proliferation, apoptosis, and oxidative stress in juvenile Chinese sturgeons (Liang et al., 2014).Oxidative stress in other test organisms such as quail, mice, and earthworms caused by DP exposure was also investigated (Wu et al., 2012;Li et al., 2013;Zhang et al., 2014).A recent study by Gagné et al. (2017) reported that DP induces oxidative stress in Mytilus edulis (blue mussels.)The study investigated the in vivo and in vitro effects of DP exposure on histopathology, LPO levels, cyclooxygenase (COX) activity, phagocytosis capacity and efficiency, and DNA strand breakage in the blue mussel following a 29-day exposure (0.001, 0.01, 0.1, and 1.0 µg L -1 DP).No significant change in hemocyte phagocytosis rate or viability was found for either in vivo or in vitro exposure.Effects of in vivo DP exposure included a lack of histopathological lesions found in the gonads of blue mussels at any of the doses.In vitro DP exposure effects included an 82% and 67% increase in LPO levels observed at 0.01 and 1.0 µg L -1 DP doses, respectively, a decrease in COX activity, and no significant difference in DNA strand breakage frequency measured in hemolymph cells.However, a previous study was the first to report that DP induced DNA damage in the hemocytes of blue mussels at all of the concentrations tested, which were 5.6, 56, and 100 µg L -1 DP (Barón et al., 2016).Dou et al. (2015) investigated the genotoxicity of DP using V. faba and T. thermophile.Micronucleus tests were performed on the root tips of V. faba, which can predict the genotoxic potential of DP with exposures at 2.4, 12, 60, 300, 1,500 µg L -1 DP.The test showed no significant difference between the treatment and control groups, indicating no genotoxicity of DP.A comet assay was performed on T. thermophile to measure DNA damage at 2.4, 12, 60, 300, and 1,500 µg L -1 DP.DP concentrations ranging from 300 to 1,500 µg L -1 showed that high levels elicited DNA damage while lower concentrations did not.This suggests that DP may pose a potential risk at concentrations greater than or equal to 300 µg L -1 .

Disrupting Hormones
A study by Kang et al. (2016) was the first to identify DP as disruptor of thyroid hormone balance in zebrafish.There was a previous study that reported an association between thyroid hormone levels and DP concentrations in serum in mother-infant pairs near an e-waste recycling area in China (Ben et al., 2014), but its endocrine disruption potential was not investigated.A study by Kang et al. (2016) examined oral and water-borne exposure pathways of DP to zebrafish, and toxicological responses, including oxidative stress and endocrine disruption, were evaluated.DP was delivered to adult male zebrafish via gavage feeding and was carried out twice on days 0 and 2, at up to 3 µg g -1 bw -1 ww -1 .Blood, liver, testis, and brain were collected on day 6 and were evaluated for oxidative damage and endocrine disruption.Hepatic catalase activity significantly increased following DP exposure, implying its oxidative damage potential.Moreover, plasma T4 concentrations increased along with up-regulation of corticotropin releasing hormone and thyroid stimulating hormone β genes in the brain.Transcriptional responses of sex hormone-related genes in the brain were also observed following DP exposure, suggesting possible sex hormone disrupting potentials of DP.In female C57BL/6J mice, 2,3,7,8-TeCDD and 2,3,7,8-TeBDD showed nearly identical potencies for endpoints that detect immunotoxicity, such as thymus weight and thymocyte numbers and spleen weight and splenocyte numbers (Ao et al., 2009).Additionally, PeBDD caused dose-independent inhibition of IL-5 production by splenocytes (Ao et al., 2009).Another recent study presented relative potencies for immune suppression of several PBDD/Fs and their chlorinated analogs in female B6C3F1/N mice.1PeBDF, 4PeBDF, and some chlorinated analogs suppressed the immunoglobulin M (IgM) antibody response (Frawley et al., 2014).It was also reported that thyroxine transport protein (Ttr) and xenobiotic metabolizing enzyme (XME) gene expression, which are used to assess aryl hydrocarbon (AhR)-mediated responses, were suppressed and up-regulated, respectively, by PBDD/Fs.TriBDD, however, had no effect on antibody response.It was concluded in the study that the brominated analogs were more potent for immunotoxicity compared to their chlorinated analogs (Frawley et al., 2014).AhR-mediated responses to PBDD/Fs were studied using three mammalian and one fish dioxin-specific bioassays (Olsman et al., 2007).All of the tested PBDD/Fs induced AhR-mediated activity.Further, the relative potencies of individual PBDD/Fs were similar to those of their chlorinated analogs using the three mammalian cell lines under investigation.On the other hand, the relative potencies obtained using a fish cell line were comparatively lower than those in the mammalian cell lines, which could be attributed to the differences in the cell line and reporter system characteristics (Olsman et al., 2007).Furthermore, pregnant C57BL/6J mice exposed to TeCDD or TeBDD in utero and via lactation showed nearly identical responses to fear conditioning tests and had deficits in contextual and auditory retention tests, indicating that both TeCDD and TeBDD disrupt emotional and memory functions and have similar developmental toxicities (Haijima et al., 2010).

DP AND NBFRs IN THE INDOOR ENVIRONMENT
DP and NBFRs are non-PBDE HFRs that exist widely in indoor environments (Table 1) and have different physical and chemical properties showing variations in persistence and fate in such environments.The commercial formulation of DP is a highly-chlorinated fire retardant additive that has been popularly used in cable coating, computer monitors, furniture, and plastic roofing materials for over 40 years (Sverko et al., 2011;Khan et al., 2016;Li et al., 2016).Although there is limited data on toxicity and no regulations for its use and production, DP is continuously of great concern to the USEPA due to the consideration that it is produced as a high volume chemical.DP was first identified in environmental samples from the Great Lakes in 2006 (Hoh et al., 2006).Since then, DP has become an emerging POP due to its ubiquitous existence in the environment, continuous use as an alternative fire retardant in the place of PBDEs, and growing demand for global use (Sverko et al., 2011).DP has similar physicochemical properties to those of PBDEs, especially in terms of its high molecular weight and log K ow (Zheng et al., 2010).DP has also been detected in human serum and hair samples (Cequier et al., 2015;Zheng et al., 2015), as well as in the indoor environment, such as in indoor dust (Cequier et al., 2015;Khan et al., 2016;Li et al., 2016).The DP isomers (syn-DP and anti-DP) in indoor dust have been positively and significantly associated with Norwegian female adult serum, indicating that their similarities originate from the same sources and have a similar environmental fate in the indoor environment (Cequier et al., 2015).Indoor dust DP has also been found to have significant positive associations with the hair of Chinese residents living in an e-waste recycling area (Zheng et al., 2010).This study also showed similarity in f anti ratios between hair and dust suggesting that dust ingestion might be one of the major pathways for human exposure to DP (Zheng et al., 2010).According to a Pakistani study by Khan et al. (2016), DP levels were heavily present indoors but not outdoors, and high correlations were observed   Peng et al., 2017Peng et al., 2017Peng et al., 2017Peng et al., 2017Peng et al., 2017Kurt-Karakus et al., 2017Kurt-Karakus et al., 2017Kurt-Karakus et al., 2017Kurt-Karakus et al., 2017Kurt-Karakus et al., 2017Al-Omran et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Khan et al., 2016Fan et al., 2016Fan et al., 2016 a Sum of all DPs, including syn-and anti-DP.
b Not analyzed.Hsu et al., Aerosol and Air Quality Research, 18: 2047-2063, 20182056 between PBDEs and DP in both indoor and outdoor dust as well as in indoor and outdoor dust in different altitudinal zones.Wang et al. (2011) indicated that DP is positively related to BFRs including PBDEs in house dust collected from e-waste recycling areas, but no correlations were found in urban areas in south China.In constrast, there were no correlations between levels of DP and PBDEs in the same indoor dust samples from Canada (Zhu et al., 2007).Li et al. (2015a) revealed indoor dust DP in urban areas to be almost 2-fold higher than in rural areas, and f syn was siginificantly correlated with f syn latitude and longitude calculated as the ratio of syn-DP to ΣDP in an indoor dust DP survey in China.As to the individual distribution of the isomers to total DP concentrations, higher contamination levels of anti-DP compared with syn-DP were found in house dusts from Hong Kong, China, and Canada (Li et al., 2015a;Zheng et al., 2015;Deng et al., 2016;Fan et al., 2016).In contrast, syn-DP exhibited higher levels than anti-DP in Egyptian house dusts (Hassan and Shoeib, 2015).However, in Pakistan, differences in DP contamination were observed in house dusts sampled from different areas (Khan et al., 2016).Additionally, house dust samples from e-waste recycling areas have also been shown to exhibit higher DP contamination compared to samples from none-waste recycling areas (Zheng et al., 2015).Furthermore, workplaces and cars in Egypt were found to exhibit higher DP levels compared to houses (Hassan and Shoeib, 2015) while kindergarten schools exhibited lower DP contamination compared to houses in Hong Kong (Deng et al., 2016).
NBFRs are also known to widely exist, and several researchers worldwide have shown interest in the detection of contamination levels in the indoor environment.Chinese studies have reported levels of NBFR in indoor dusts collected from houses, offices, e-waste recycling sites, cars, and other public areas (Cao et al., 2014;Qi et al., 2014;Zheng et al., 2015;Peng et al., 2017).Human activities and geographical distribution influences indoor NBFR pollution levels (Qi et al., 2014).Indoor dust from urban areas and public buildings showed significantly higher levels of NBFRs than rural areas and family homes indicating the different applications of NBFRs (Qi et al., 2014).A temporal trend study on NBFRs revealed that NBFR concentrations are relatively stable during any season and that human activities do not affect seasonal variations in NBFR contamination (Cao et al., 2014).In e-waste recycling areas, different pollutant patterns have been observed at different sites, which could be attributed to the different kinds of e-waste dismantled during recycling activities at each site (Zheng et al., 2015).As to human exposure, Zheng et al. (2015) and Peng et al. (2017) both reported that human exposure to NBFRs via dust ingestion is higher in toddlers compared with adults.In these studies (Cao et al., 2014;Qi et al., 2014;Zheng et al., 2015;Peng et al., 2017), DBDPE was the dominant compound, exhibiting the highest contamination level in all sampling areas, suggesting that DBDPE had been used or released in large amounts in China.Furthermore, indoor dust samples collected from e-waste recycling areas showed relatively higher NBFR contamination compared to other areas.The levels of NBFR in the indoor environment in other Asian countries have also been investigated (Ali et al., 2014;Hassan and Shoeib, 2015;Mizouchi et al., 2015;Khan et al., 2016;Kurt-Karakus et al., 2017).In Pakistan, DBDPE levels in house dust (Khan et al., 2016) were found to be higher compared to those found in universities, clothing stores, and electronics stores (Ali et al., 2014).However, these levels were relatively lower compared to DBDPE levels in house dust collected from Japan (Mizouchi et al., 2015), except for one Pakistani rural home and one home from a colder area in Pakistan, in which the values were higher.Ali et al. (2014) reported lower BTBPE levels than DBDPE in universities, clothing stores, and electronics stores.EH-TBB, BTBPE, and BEH-TEBP concentrations in Egyptian and Turkish indoor environments were studied (Hassan and Shoeib, 2015;Kurt-Karakus et al., 2017).Low NBFR contaminations were found in homes, workplaces, and cars in Egypt.The use of flame retardants is not regulated in Egypt; therefore, imported goods incorporated with NBFRs may be one source of NBFR contamination.Consequently, the levels found in Egyptian dusts were among the lowest NBFR contamination reported compared to other countries (Hassan and Shoeib, 2015).Higher EH-TBB, BTBPE, and BEH-TEBP concentrations were found in Turkish indoor dusts (Kurt-Karakus et al., 2017).EH-TBB were among the highest concentrations, followed by BTBPE, while most BEH-TEBP levels were less than the method detection limit (Kurt-Karakus et al., 2017).In indoor air samples collected from Turkish homes and offices, EH-TBB was also found to be the most dominant compound (Kurt-Karakus et al., 2017).
Contamination levels of NBFRs in the United States and other parts of Europe have also been reported.In the United States, levels of NBFRs were measured in the indoor dust from houses and fire stations (Brown et al., 2014;Schreder and La Guardia, 2014).Dust samples from California homes and in living quarters from California fire stations were investigated for EH-TBB, BTBPE, DBDPE, and BEH-TEBP contamination.Dust samples from the fire stations showed higher concentrations compared to those from houses, with EH-TBB being the most dominant compound in both sampling areas.NBFRs released during fire events might have transferred to the living quarters of fire stations through the firefighters (Brown et al., 2014).Lower levels of EH-TBB and BEH-TEBP and higher levels of BTBPE and DBDPE were found in house dust collected from Washington compared to house dust from California, with EH-TBB also found to be the most dominant NBFR (Schreder and La Guardia, 2014).In Canadian house dusts, lower levels of EH-TBB and BTBPE were obtained compared to American house dusts, with EH-TBB present in higher concentrations than BTBPE (Fan et al., 2016).In European countries, dust samples from various indoor environments were investigated for NBFR contamination (Cequier et al., 2014;Fromme et al., 2014;Al-Omran and Harrad, 2016;Cristale et al., 2016;Kuang et al., 2016).Living rooms from the United Kingdom were shown to exhibit higher levels of NBFRs compared to kitchens, with DBDPE having the highest concentrations in both areas (Kuang et al., 2016).However, in another UK study, BEH-TEBP exhibited the highest concentrations in house dusts (Al-Omran and Harrad, 2016).Similar results were obtained in a German study, where BEH-TEBP was the dominant NBFR in house dusts (Fromme et al., 2014).In Norway, NBFR levels in school classrooms were higher compared to those of residential living rooms, with DBDPE showing the highest concentrations in both environments (Cequier et al., 2014).In Spain, NBFRs in indoor dust from schools were not detectable except for BEH-TEBP.The highest level of NBFR contamination was found in dusts collected from theatres, followed by homes, and universities/research institutions (Cristale et al., 2014).

PBDD/FS IN THE INDOOR ENVIRONMENT
The global concentrations of PBDD/Fs in the indoor environment are listed in Table 2.In indoor environments, PBDD/Fs have mainly been studied in dust.Most of the recent studies are from Asian countries including Japan, Vietnam, and Taiwan but there are also some from western countries such as Sweden and the USA.House dusts collected from a residential house in an e-waste recycling site in Vietnam showed higher average PBDD/F concentrations compared to general urban Vietnamese environment house dusts (Tue et al., 2010).Workshop floor dust collected from an e-waste recycling site in Taizhou, China, however, showed higher average PBDD/F levels compared to that of Vietnam (Ma et al., 2009).However, the maximum PBDD/F concentration found by Ma et al. (2009) were still below the maximum PBDD/F levels found in an American residential house dust (Tue et al., 2013).In Japan, PBDD/F is among one of the major dioxin-like compounds in indoor dust.Suzuki et al. (2010) reported that levels of PBDD/Fs in Japanese office dusts were similar to those of house dusts.Moreover, Japanese PBDD/F house dust levels were in a range similar to those found in Vietnam.In Taiwan, indoor dust levels of PBDD/Fs in school classrooms from rural areas were higher compared to urban school classrooms (Gou et al., 2016b).In addition, Taiwanese school classrooms' PBDD/F levels were in the same order of magnitude as those found in Japanese and Vietnamese house dusts, but lower compared to Swedish and American house dusts (Gou et al., 2016b).These differences in the indoor dust levels of PBDD/Fs emphasize large variations in the overall global data.
These variations in PBDD/F levels are mainly attributed to the presence of potential emission sources indoors.Electrical and electronic equipment, textiles, furniture, and construction materials incorporated with flame retardants are materials that could be potential sources of PBDD/Fs.Among these materials, electrical and electronic equipment is regarded to be an important source of PBDD/Fs, which is supported by the very high concentrations of PBDD/Fs recorded in dust from an electrical waste and electronic equipment recycling plant in Sweden (Remberger et al., 2014).Extremely high PBDD/F concentration of 3,800 ng g -1 dw -1 was also found in dust from a new car in Sweden compared to house dusts, which may indicate that new cars could be reservoirs for PBDD/Fs (Remberger et al., 2014).Not only are PBDD/Fs possibly correlated with different indoor emission sources, but several studies have reported significant correlations between PBDD/Fs and PBDEs (Ma et al., 2009;Gou et al., 2016b).ΣPBDD/Fs were found to be significantly correlated with ΣPBDEs in workshop floor dust from a Chinese e-waste recycling facility (r = 0.769, p < 0.01) (Ma et al., 2009) and in indoor dust from Taiwanese school classrooms (r = 0.862, p < 0.001) (Gou et al., 2016b).NBFRs and DP may also be connected with PBDD/Fs in the indoor dust, but no studies have looked into this possible correlation yet.
Although there is an ample amount of scientific evidence confirming the presence of NBFRs, DPs, and PBDD/Fs in indoor environments, there is still a lack of evidence linking their indoor levels with their toxicity.However, results from studies using animal models suggest that there is a strong emerging evidence showing that exposure to high levels of these chemicals may possibly cause adverse health-related outcomes.

FUTURE WORKS FOR HFRS AND PBDD/FS IN THE INDOOR ENVIRONMENT
After over two decades of use, most studies have focused only on PBDE levels in the indoor environment as compared to DP, NBFRs, and PBDD/Fs, which are now considered to be the modern day substitutes for PBDEcontaining flame retardants.Spatial and longitudinal investigations of these emerging HFRs should be conducted to further assess the risks of these compounds to human health.Although current studies show that PBDEs levels in indoor dust are not related to indoor air PBDE concentrations (possibly because of small sample size; Fromme et al., 2009;Toms et al., 2009), it has been hypothesized that a pattern of non-dietary PBDE exposure probably exists and that the possible route of PBDE exposure is release from the surface of electronics and transfer to human bodies through accumulation of airborne PBDEs on indoor dust, as revealed in our previous study (Ni et al., 2011;Shy et al., 2015).Ni et al. (2011) indicated that human exposure to PM 2.5 -bound PBDEs might have a minor contribution to total non-dietary PBDE exposure.In terms of global scientific reports, few studies have considered PM 2.5 -bound PBDE levels in the indoor environment (Beser et al., 2014;Li et al., 2015a;Xu et al., 2015;Chao et al., 2016;Deng et al., 2016).To our knowledge, there are no known reports that have investigated the levels of PM 2.5 -bound DP, NBFRs, or PBDD/Fs in the indoor environment.Respirable fine particulates are reported to have correlations to various cardiopulmonary health effects such as cardiovascular and respiratory diseases (Dominici et al., 2006;de Oliveira et al., 2012;Zhang et al., 2013;Xing et al., 2016).Numerous studies have turned their focus on PM 2.5 levels due to the possibility of these small particles being deposited deep into the lungs, thus causing respiratory ailments and diseases, as well as other negative health problems (Zhang et al., 2013;Gautam et al., 2016;Hwang et al., 2016;Lee et al., 2016;Lu et al., 2016;Xing et al., 2016;Ma et al., 2017).People with a known history of cardiopulmonary problems face   et al., 2009Suzuki et al., 2010Suzuki et al., 2010Tue et al., 2010Tue et al., 2010Tue et al., 2013Remberger et al., 2014Remberger et al., 2014Remberger et al., 2014Gou et al., 2016Gou et al., 2016Gou et al., 2016 Gou et al.,  higher risk of mortality due to the increasing PM 2.5 concentrations in the atmosphere while normal people also face a higher risk of developing respiratory problems (Dominici et al., 2006;de Oliveira et al., 2012;Xing et al., 2016;Li et al., 2017).PM 2.5 is the essential issue for human health in both the indoor and outdoor environment.Although most environmental scientists recognize that high levels of PM 2.5 -bound HFRs including PBDEs are well known in the indoor environment, few reports have considered their toxic effects for human health via inhalation.Moreover, with PM 2.5 -bound PBDEs being a very important concern, then PM 2.5 -bound NBFRs, DP, and PBDD/Fs should be of important concern also.Future works will be continuously encouraged to investigate indoor levels of PM 2.5 -bound NBFRs, DP, and PBDD/Fs to further evaluate their body burden percentage via inhalation pathway and to assess the health risks or the health effects of direct inhalation.

CONCLUSIONS
There is currently a significant increase in global demand for DP and NBFRs as substitutes for the banned conventional brominated flame retardants.Global data on the indoor environment, particularly on indoor dust, have reported the presence of NBFRs, DP, and PBDD/Fs.Currently, there is limited global toxicological data for these chemicals.The toxicological data for NBFRs, DP, and PBDD/Fs, which includes acute toxicity, developmental toxicity, immunotoxicity, neurotoxicity, reproductive toxicity, genotoxicity, or disruption of hormones, are reported in various in vitro and in vivo studies.Most previous studies focused only on the non-dietary intake of HFRs, such as PBDEs, DP, and NBFRs, and there is a lack of study on simultaneous levels of indoor dust and fine particulate matter.Therefore, in order to assess their risks and the potential for adverse health effects through the main exposure routes, including dust ingestion and inhalation of respirable fine particles, additional studies are required to further observe their levels in the indoor environment, particularly in dust and fine particulate samples.

Table 1 .
Global data for DP and NBFR contamination in indoor environments.

Table 1 .
Global data for DP and NBFR contamination in indoor environments.(cont'd

Table 1 .
Global data for DP and NBFR contamination in indoor environments.(cont'd).

Table 2 .
Global concentrations of PBDD/Fs in indoor environments.