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Organophosphate Ester Flame Retardants: Are They a Regrettable Substitution for Polybrominated Diphenyl Ethers?

  • Arlene Blum
    Arlene Blum
    Green Science Policy Institute, Berkeley, California 94709, United States
    Department of Chemistry, University of California, Berkeley, Berkeley, California 94705, United States
    More by Arlene Blum
  • Mamta Behl
    Mamta Behl
    National Toxicology Program, National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, United States
    More by Mamta Behl
  • Linda S. Birnbaum
    Linda S. Birnbaum
    National Cancer Institute at National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, United States
  • Miriam L. Diamond
    Miriam L. Diamond
    Department of Earth Sciences, University of Toronto, Toronto, Ontario, Canada M5S 3B1
  • Allison Phillips
    Allison Phillips
    Risk Assessment and Natural Resource Sciences Inc., Arcadis, Raleigh, North Carolina 27607, United States
  • Veena Singla
    Veena Singla
    Program on Reproductive Health and the Environment, Department of Obstetrics, Gynecology and Reproductive Sciences, University of California, San Francisco, San Francisco, California 94143, United States
    More by Veena Singla
  • Nisha S. Sipes
    Nisha S. Sipes
    National Toxicology Program, National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, United States
  • Heather M. Stapleton
    Heather M. Stapleton
    Nicholas School of the Environment, Duke University, Durham, North Carolina 27708, United States
  • , and 
  • Marta Venier*
    Marta Venier
    O’Neill School of Public and Environmental Affairs, Indiana University, Bloomington, Indiana 47401, United States
    *E-mail: [email protected]
    More by Marta Venier
Cite this: Environ. Sci. Technol. Lett. 2019, 6, 11, 638–649
Publication Date (Web):October 21, 2019
https://doi.org/10.1021/acs.estlett.9b00582

Copyright © 2019 American Chemical Society. This publication is licensed under these Terms of Use.

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Abstract

As the use of polybrominated diphenyl ethers (PBDEs), and the entire class of organohalogen flame retardants, is declining, the use of organophosphate ester flame retardants (OPFRs) is increasing. In this paper, we ask whether OPFRs are a better choice than PBDEs. To address this question, we compared OPFRs with PBDEs for a wide range of properties. Exposure to OPFRs is ubiquitous in people and in outdoor and indoor environments, and OPFRs are now often found at higher levels compared to PBDE peak exposure levels. Furthermore, data from toxicity testing, epidemiological studies, and risk assessments all suggest that there are health concerns at current exposure levels for both halogenated and nonhalogenated OPFRs. Obtaining the scientific evidence needed for regulation of OPFRs can take many years. Given the large number of OPFRs in use, manufacturers can move toward healthier and safer products by developing innovative ways to reduce the risk of fire for electronics enclosures, upholstered furniture, building materials, and other consumer products without adding flame retardant chemicals.

Introduction

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Flame retardants are added to consumer products and building materials to reduce the risk of fire. Their use is driven by flammability standards, usually based on small-scale fire testing, which may not accurately predict real life fire behavior. (1,2) Beginning in the 1970s, polybrominated diphenyl ethers (PBDEs) were added to consumer products, including furniture, children’s products, and electronics. After extensive research showed that PBDEs were persistent, bioaccumulative, and toxic, in 2004 the European Commission and California banned the use of Penta- and OctaBDE, two commercial mixtures primarily used in North America. Also, in 2004, the U.S. Environmental Protection Agency (US EPA) negotiated a phase-out of new production of these two PBDE commercial mixtures with U.S. manufacturers. Subsequently in 2009, the US EPA negotiated the phase-out of DecaBDE production, the PBDE with the largest production volume. Penta- and OctaBDE were added to the Stockholm Convention in 2009, prompting more than 150 signatories to legislate their phase-out (see Figure 1 and Table S1 for more details about the regulatory timeline). DecaBDE was added to the Stockholm Convention in 2017 and similarly phased out of use in most countries.

Figure 1

Figure 1. Timeline of major regulatory milestones for PBDEs and OPFRs. Policy enactment dates are listed; implementation/compliance dates are often later (see Table S1 for detailed information).

Unfortunately, old furniture, electronics, vehicles, and other products containing PBDEs continue to be used and reused in spite of the phase-outs and bans that prevent their use in newly manufactured products. The resultant long-term exposure to PBDEs, especially in low-income communities, is another reason to avoid alternative flame retardants that could be similarly harmful.
As the use of PBDEs and the class of organohalogen flame retardants (flame retardants containing carbon and halogen elements, most often bromine or chlorine) is declining due to regulatory action and manufacturers’ voluntary actions, the use of organophosphate ester flame retardants (OPFRs) is increasing. For example, DecaBDE production in the United States dropped from 50 million pounds (23000 tonnes) in 2012 to <25000 pounds (11 tonnes) in 2015. (3) During the same time frame, the U.S. production volume of various OPFRs has remained constant or increased. (3)
OPFRs are organic esters of phosphoric acid-containing either alkyl chains or aryl groups, and they may be halogenated or nonhalogenated (see Figure S1). In addition to their use as flame retardants, OPFRs are used as plasticizers in consumer products and construction materials. (4,5) OPFRs are also oxidation products of phosphites, which are commonly used antioxidants in plastic products. (6,7) Commercial flame retardant formulations that contain mainly OPFRs have replaced PentaBDE in residential furniture. (8,9) OPFRs have also been used as substitutes for Octa- and DecaBDE in electronics, with resorcinol bis(diphenylphosphate) (RDP or PBDPP) and triphenyl phosphate (TPHP) measured in televisions at milligram per gram levels. (10−12)Table 1 lists the OPFRs discussed in this paper.
Table 1. List of Major OPFRs Cited in the Text Along with CAS Numbers (adapted from refs (17) and (106)), Molecular Formula, Molecular Weight (g/mol), LogKOW, Log KOA, TSCA, DSL
OPFRfull nameCAS Registry No.molecular formulamolecular weight (g/mol)LogKOWaLogKOAaTSCAbDSLc
TEPtriethyl phosphate78-40-0C6H15O4P182.160.95.5yes (imported) (450–4500 tonnes)yes
TNBPtri-n-butyl phosphate126-73-8C12H27O4P266.323.87.7yes (450–4500 tonnes)yes
TCEPtris(2-chloroethyl) phosphate115-96-8C6H12Cl3O4P285.481.67.6yes (imported) (11–45 tonnes)yes
TCIPP+tris(2-chloro-1-methylethyl) phosphate13674-84-5C9H18Cl3O4P   yes (23000–45000 tonnes)yes
TPPtripentyl phosphate2528-38-3C15H33O4P308.40  nono
TDMPPtris(3,5-dimethylphenyl) phosphate9006-37-5C24H27O4P410.44813.5  
TDCIPPtris(1,3-dichloro-2-propyl) phosphate13674-87-8C9H15Cl6O4P430.903.710.6 yes
TPHPtriphenyl phosphate115-86-6C18H15O4P326.284.710.5yes (450–4500 tonnes)yes
EHDPP2-ethylhexyl diphenyl phosphate1241-94-7C20H27O4P362.406.311.3yes (imported) (450–4500 tonnes)yes
TBOEPtris(2-butoxyethyl) phosphate78-51-3C18H39O7P398.473.012.3  
DOPPdioctyl phenyl phosphonate1754-47-8C22H39O3P382.52    
TmCPtri-m-cresyl phosphate563-04-2C21H21O4P368.366.312.0  
TpCPtri-p-cresyl phosphate (tri-p-tolyl phosphate)78-32-0C21H21O4P368.376.312.0  
ToCPtri-o-cresyl phosphate (tri-o-tolyl phosphate)78-30-8C21H21O4P   yesyes
TPPPtris(2-isopropyl phenyl) phosphate64532-95-2C27H33O4P452.529.114.0  
TDBPPtris(2,3-dibromopropyl) phosphate126-72-7C9H15Br6O4P697.614.214.1yesno
TTBPPtris(4-tert-butylphenyl) phosphate78-33-1C30H39O4P494.6210.415.0yes (withheld)yes
IPPisopropylated phenyl phosphate68937-41-7 (mix of isomers)C21H18O4P–C27H30O4P368.28–452.28NRNRyes 
IDDPisodecyl diphenyl phosphate29761-21-5C22H31O4P390.457.312.0  
a

Toxic Substances Control Act Chemical Substance Inventory (TSCA Inventory). https://www.epa.gov/tsca-inventory/how-access-tsca-inventory (accessed May 2018).

b

US EPA; CDR database, 2016. https://www.epa.gov/chemical-data-reporting (accessed May 2018). (17) Withheld means that the production volume was withheld (values in parentheses are estimated production volumes for 2015).

c

DSL indicates inclusion or not on the Canadian domestic substance list.

Extensive scientific research now suggests that the entire class of organohalogen flame retardants may have hazardous properties, and some authoritative bodies are now addressing this problem with a chemical class approach (see Table S1). (13) In 2017, the U.S. Consumer Product Safety Commission (CPSC) accepted a petition to ban furniture, children’s products, electronic enclosures, and mattresses containing any member of the class of organohalogen flame retardants. In May 2019, at the request of the CPSC, the National Academies of Sciences (NAS) released a report titled “Scoping Report for Conducting a Hazard Assessment of Organohalogen Flame Retardants as a Class”. (14) This report states that “the number of chemicals in use today demands a new approach to risk assessment, and the class approach is a scientifically viable option”. The authors of the report concluded that “the best approach is to define subclasses as broadly as is feasible for the analysis”. Although the challenges to a class approach might appear daunting, the alternative—individual assessments of hundreds of chemicals—is unrealistic. The report supports the idea that the class approach can prevent the problem of “regrettable substitution” of a replacement lacking adequate toxicity information for a phased out known hazardous substance. Regrettable substitution occurs because of the difficulty of changing industrial processes and a lack of toxicological information, causing manufacturers to replace a phased-out chemical with a “drop in” substitute chemical that has a similar structure, function, and potential for harm. The NAS report also highlights the point that the cumulative exposures and risks are ignored when chemicals are assessed individually. In 2018, the European Commission (EU) proposed to prohibit the entire class of organohalogen chemicals in electronic display enclosures and stands, (15) effective April 1, 2021. The manufacturers of televisions and other electronics will need to find an alternative solution to meet flammability codes, and it is likely that industry will look to nonhalogenated OPFRs.
In this paper, we ask whether OPFRs, as a class, have a reduced potential for harm compared to PBDEs or if they are an example of regrettable substitution. To address this question, we compared OPFRs and PBDEs with regard to their environmental fate, measured levels indoors, exposure levels among the general population, and evidence of adverse health effects. Our comparison was performed by reviewing the literature, although the review was not comprehensive. We conclude by discussing current policy changes and opportunities to innovate for less hazardous materials and products with reduced potential for harm compared to that of flame retardants.

Environmental Behavior

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The idea that OPFRs are less harmful than PBDEs was largely based on the presumption that OPFRs are less environmentally persistent and hence have a lower potential for widespread environmental distribution and exposure. However, mounting evidence calls this presumption into question.
PBDEs have been classified as persistent organic pollutants (POPs) under the Stockholm Convention due to their persistence, ability to undergo long-range transport, bioaccumulation potential, and toxicity to both humans and wildlife. (16) PBDEs have been documented in air, water, and terrestrial and aquatic biota at a global scale (see Tables S1–S4). They are nonpolar compounds with generally low volatility and vapor pressures (see Figure S2). PBDEs can degrade, albeit slowly, in environmental matrices through photodegradation and microbial debromination, which can create more bioavailable and toxic congeners.
OPFRs are expected to be less persistent in the environment than PBDEs. (17) Data on their persistence are scant, and the physical and chemical properties of OPFRs are difficult to measure or estimate, which makes prediction of their environmental behavior more uncertain than that of PBDEs. Although their higher vapor pressures lead to the expectation of higher air concentrations compared to those of PBDEs (see Figure S2), they are also expected to have shorter half-lives in air and thus reduced atmospheric long-range transport potential. (18) Compared with PBDEs, OPFRs and especially chlorinated OPFRs are more soluble and can persist in water, which gives them the ability to undergo long-range transport via waterborne routes. (19,20) Thus, rather than being POPs, as is the case with PBDEs, chlorinated OPFRs appear to be persistent mobile organic compounds (PMOCs), which is equally concerning. (18,21)
Although OPFRs were not expected to accumulate in the environment on the basis of their physical and chemical properties, multiple measurements show that the concentrations of many OPFRs have reached values that are orders of magnitude higher in air and water in numerous environments ranging from urban areas to remote Arctic and Antarctic locations, compared to those of PBDEs when they were at peak use (see Figure 2 and Tables S2 and S3). OPFRs are readily scavenged from air by precipitation and then transported to surface waters because of their higher solubility and weak tendency for sequestration in soil and other carbon-rich matrices compared to PBDEs. This efficient scavenging of OPFRs by rain, coupled with high emission rates, results in some urban surface water concentrations being similar to those from treated final wastewater treatment plant effluent (see Table S1). (18) In Great Lakes water, for example, concentrations of OPFRs were in the range of 10–100 ng/L while ∑BDE (defined as the sum of total measured BDE congeners, which can vary from study to study) values were significantly lower, with a range of 0.05–0.25 ng/L (see Table S2 for specific data). Despite their high aqueous solubility, OPFRs also accumulate in sediment by virtue of high emissions and their ability to be transported to aquatic systems. For example, ∑14OPFR values were ∼0.5–50 ng/g of dry weight (g dw) in the Great Lakes sampled in 2010–2013, (22) comparable to concentrations of 0.5–6.7 and <4 to >240 ng/g dw for ∑9BDE and BDE-209, respectively, measured at their peak usage. (23)

Figure 2

Figure 2. Bar chart representing selected median concentrations of total OPFRs and total PBDEs in (A) outdoor air (picograms per cubic meter), (B) water (nanograms per liter), and (C) indoor dust (nanograms per gram). Each bar represents data from a different study (see Tables S2–S5 for details about the studies included and for a more comprehensive list of locations).

Finally, numerous measurements confirm not only the presence but also the relative abundance of OPFRs in remote locations, which cannot be explained by local releases. For example, total OPFRs have reached a median concentration of 237 pg/m3 in Canadian Arctic air (19,24−26) while ∑BDE air concentrations, at the time of peak use, were orders of magnitude lower. OPFRs have also accumulated in Arctic sediments at concentrations 10–100 times greater than those of PBDEs, which, as noted above, is unexpected because OPFRs are less likely to deposit due to of their high water solubility. (27)
Overall, environmental measurements clearly show that OPFRs demonstrate long-range transport and accumulation in the environment that rival those of PBDEs, despite expectations to the contrary. In addition, whereas concerns with PBDEs were due to them being POPs, OPFRs fit into a different class of concern, that of PMOCs.

Indoor Behavior and Human Exposure

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Concentrations of PBDEs and OPFRs in indoor air, in house dust, and on hand wipes provide critical information about the exposure potential because inhalation, hand-to-mouth contact, and dermal absorption are all important routes of human exposure to flame retardants. Below, we provide evidence of relatively high levels of exposure to OPFRs compared to PBDEs, originating from elevated levels in indoor air, house dust, and food.
The widespread use of PBDEs since the late 1970s resulted in near-ubiquitous human exposure, with ∑BDE in breast milk and serum levels (both lipid-adjusted) peaking in the early to mid-2000s. (28,29) However, the increasing use of OPFRs following the phase-out of PBDEs has also led to greater human exposure, and urinary OPFR biomarker levels have been steadily climbing since the early 2000s. (30) Interestingly, OPFRs have been found in dust at substantial levels for at least two decades. For instance, ∑OPFR levels are similar to ∑PBDE levels in house dust standard reference material (SRM 2585; ∼5000 ng/g), which was prepared from hundreds of vacuum cleaner bags collected in several U.S. states in the mid-1990s. (9,31−33) This comparison suggests that exposure to OPFRs was similar to exposure to PBDEs in the 1990s but now appears to have increased since the phase-out of PBDEs.
The growing prevalence of OPFRs and decreasing amounts of decaBDE in indoor house dust correlate with their changing production levels in the United States. Common OPFRs such as tricresyl phosphate (TCP), tri-n-butyl phosphate (TNBP), and 2-ethylhexyl diphenyl phosphate (EDHPP) each have had production volumes of ≤10 million pounds (4500 tonnes) yearly since 2012, (34) compared to 10–50 million pounds (4500–23000 tonnes) of DecaBDE in 2012, dropping to <25000 pounds (11 tonnes) in 2015. (35)
Although the percentage by weight application of PBDEs and OPFRs to polyurethane foam is roughly identical (∼3–7%), OPFRs are also heavily used in electronics. As a result, OPFRs have been detected at much higher levels in indoor air (Table S4). (8,36) While this is partly accounted for by OPFRs being used as both flame retardants and plasticizers, it is also likely reflective of the higher vapor pressure of OPFRs compared to that of PBDEs, leading to increased off-gassing of OPFRs from treated products into indoor air (Figure S2). Studies from North America and Europe conducted in the early to mid-2000s, when PBDE use and exposure were at their height, reported average ∑BDE indoor air concentrations in the range of 100–600 pg/m3. (37−39) By contrast, recent studies report levels of OPFRs in the nanogram per cubic meter range, at least an order of magnitude higher than for PBDEs (Table S4). Much like those of PBDEs, OPFR air concentrations have been found to fluctuate seasonally. They also vary depending on the microenvironment, with cars and offices often having higher concentrations compared to living spaces, reflecting high usage and emissions. (39−41)
Because OPFRs have vapor pressures that are higher than those of PBDEs, it might be assumed that OPFR dust concentrations would be lower than PBDE dust concentrations. However, despite regional differences in flame retardant dust concentrations, recently reported ∑OPFR geometric mean and median dust levels are either equivalent to (if considering the measured sum of PBDE congeners in commercial mixtures) or higher than ∑PBDE geometric mean and median dust levels from the early to mid-2000s (Figure 2). Reported ∑OPFR geometric mean and median dust concentrations range from low micrograms per gram to low milligrams per gram (Table S5). (3,16,33,41−46) Several recent studies have estimated the daily exposure to OPFRs and PBDEs (e.g., nanograms per kilogram per day) via levels measured in indoor dust, and the results demonstrate that exposure to OPFRs is higher. (47,48)
Although data are limited for PBDEs, OPFRs have been found on hand wipes at levels higher than those of PBDEs, suggesting that the magnitude of exposure via hand-to-mouth and dermal transfer pathways is potentially greater for OPFRs than for PBDEs (Table S3). As for PBDEs, OPFR exposure may also result from dietary or diet-associated intake. (49,50) The use of EHDPP, TPHP, and TNBP in food packaging material is approved by the U.S. Food and Drug Administration, and their migration from plasticized film wrappers into food has been documented. (51) The presence of OPFRs has also been reported in butter, bread bags, fish, and drinking water. (52−54) Likewise, a recent study detected tris(2-chloroisopropyl) phosphate (TCIPP) and TNBP in >70% of 87 food samples and five tap water samples collected in Australia. (55) On the basis of the levels found in food, the estimated daily dietary intake was 4.1 ng/kg of body weight (kg bw) for tris(2-chloroethyl) phosphate (TCEP), 25 ng/kg bw for TCIPP, and 6.7 ng/kg bw for TNBP. In Sweden, Poma et al. detected EHDPP in composite food samples from multiple food categories and estimated a daily intake of ∑OPFRs [TCEP, TPHP, EHDPP, tris(1,3-dichloro-2-propyl) phosphate or TDCIPP, and TCIPP] in the range of 6–49 ng/kg bw. (56) The total dietary intake of ∑7OPFRs in fish from the Philippines was 22 ng (kg bw)−1 day–1. (57) In a U.S. market basket study from 1988, the dietary intake for TPHP was calculated as 0.3–4.4 ng (kg bw)−1 day–1. (58) For comparison, the estimate of the daily intake of ∑13BDE congeners for U.S. adults was 0.9 ng/kg bw and that for ∑18BDE congeners for Canadian adults was 0.7 ng/kg bw during the 2000s. (59,60)
Because of differences in metabolism between the two chemical classes, the comparison of PBDE and OPFR internal dose levels is challenging. PBDEs are bioaccumulative and have long half-lives (weeks to years) in the human body, while OPFRs are rapidly metabolized with relatively short half-lives (hours to days). (61,62) PBDEs have been measured in human serum as biomarkers of exposure, while diester metabolites have been measured in human urine as indicators of OPFR exposure. Serum PBDE levels have been documented in the picomolar to low nanomolar range (on a wet weight basis), with BDE-47 generally detected most frequently and at the highest concentration of all BDE congeners (although it should be noted that BDE-209 and higher-molecular weight congeners were often not included in past serum analyses). (63−65) Urinary OPFR metabolite levels have been reported in the low to mid nanomolar range, with diphenyl phosphate (DPHP, metabolite of multiple OPFRs), 1-hydroxy-2-propyl bis(1-chloro-2-propyl) phosphate (BCIPHIP, metabolite of TCIPP), and bis(1,3-dichloro-2-propyl) phosphate (BDCIPP, metabolite of TDCIPP) frequently detected at higher levels compared to other urinary metabolites, (66,67) although regional differences exist. (68) While it is difficult to compare serum PBDE levels to urinary OPFR metabolite levels, detection frequencies among biospecimens are similar. BDE-47 was detected in 97% of serum samples (n = 2062) tested in the 2003–2004 National Health and Nutrition Examination Survey (NHANES), and recent urine analyses have detected BDCIPP, BCIPHIPP, and DPHP in >95% of tested samples. (40,55,63,68−71) From these data, it is evident that OPFR exposure is pervasive in the human population and that vulnerable subpopulations like infants and children may be more highly exposed. (66,71−74)
Continuing OPFR biomonitoring efforts will help define trends regarding human exposure to OPFRs. To summarize, OPFRs and their metabolites can be detected at relatively high levels and frequencies in indoor air, dust, food items, and human biospecimens. Given that OPFR exposures may still be increasing, it is clear that the current potential for indoor exposure to OPFRs is substantially greater than it was for PBDEs in the early to mid-2000s, when PBDE use was at its height.

Toxicity and Health Effects

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Another important aspect of this comparison between PBDEs and OPFRs is understanding how the toxicity profiles of these classes of compounds compare with each other. As a class, 97 OPFR flame retardants were shown to have potential toxicity based on the Quick Chemical Assessment Tool (QCAT). (75) The US EPA categorized some of the alternative flame retardants as high-priority compounds that critically need more toxicological studies or for which regulatory measures could be envisaged. (75)
Although the use of OPFRs is on the rise, their toxicological hazard has not yet been well-characterized, and indeed, a more complete understanding of the toxicity of PBDEs has emerged only after their regulation. This section summarizes recent toxicity data and describes a linkage between toxicity data noted in in vitro and in vivo studies with human exposure, ultimately identifying data gaps that need to be addressed. In addition to OPFRs and PBDEs, we also compared the toxicity profile of tetrabromobisphenol A (TBBPA) because it is the most highly produced brominated flame retardant in the world. (76) Its production has remained fairly stable over the years, likely a reflection of its primary use as a reactive flame retardant (as opposed to PBDE additive use). The EU noted “no health effects of concern” for TBBPA for adults or infants based on adequate margins of safety. (77) However, evidence in the literature suggests effects on reproductive and nervous system development, including brain and thyroid function. (76,78)
OPFRs, PBDEs, and TBBPA appear to have comparable developmental and neurodevelopmental toxicity potential in a variety of in vitro assays that represent processes critical to neurodevelopment such as neuronal proliferation, neurite outgrowth, synaptogenesis, and network formation. (75,79) They have also been shown to affect reproduction, development, and motor activity in a multitude of alternative animal models such as zebrafish, Caenorhabditis elegans, and Planaria. (79−87) Furthermore, similar to that of PBDEs, exposure to OPFR and TBBPA appears to elicit behavioral alterations that persist into adulthood, long after cessation of developmental exposure. (88,89)
To relate toxicity data from in vitro and alternative animal models, as well as traditional in vivo animal studies to human exposure, we used a high-throughput toxicokinetic model (HTTK) (90,91) to convert the data, thereby allowing comparisons across the various exposure scenarios (Figure 3). The main purpose of this analysis is twofold. The first is to compare in vitro and in vivo PODs (points of departure) to inform readers about how novel rapid testing strategies relate to traditional rodent studies in their toxicological outcomes, so that they may be used as complementary approaches. The second is to compare the minimum risk levels (MRLs, i.e., PODs values including safety factors, which are equivalent to RfDs or reference doses) to measured human exposures.

Figure 3

Figure 3. Flame retardant plasma concentrations measured or estimated from ingestion using data from house dust, breast milk, and/or handwipe samples (colored bars and circles) are compared to the most potent in vitro concentration per chemical (black dots) and in vivo point of departure (POD; triangles). The filled triangles represent rat plasma concentrations based on in vivo POD values (when available), and the empty triangles represent the minimum risk levels (MRLs). The colored bars represent the range of concentrations, and the circles represent the mean, median, or maximum median (see the Supporting Information for further details about these calculations).

First, plasma concentrations were simulated in the model using data from house dust, breast milk, or hand wipes samples (Figure 3 and Table S7). Important input parameters for the model were the estimated oral exposures based on child feeding (for breast milk), ingestion due to hand-to-mouth actions (for dust), and chemical-specific parameters, such as the fraction of the chemical bound to plasma, intrinsic metabolic clearance, pKa, LogP (lipophilicity), and molecular weight. A detailed description of these values and calculations can be found in Tables S7–S9. The internal exposure values calculated with the HTTK model were then compared to POD values from developmentally or neurodevelopmentally associated in vitro assays and/or in vivo alternative animal (i.e., zebrafish) assays to understand how in vitro PODs relate to human exposure. A POD is defined as a dose–response point that marks the starting point for low-dose extrapolation; that is, the POD is the exposure level at which an effect is seen.
The in vitro POD to human plasma concentration comparison found that the in vitro POD for TPHP, BDE-47, TDCIPP, and TBBPA lies within the range of estimated plasma concentrations from human exposure, as indicated by an overlap of the colored bars with the black filled circles (Figure 3). For other compounds such as trimethyl phenyl phosphate (TMPP), isodecyl diphenyl phosphate (IDDP), EHDPP, isopropylated phenyl phosphates (IPP), tert-butylated phenyl diphenyl phosphate (BPDP), and TCEP, the biological activity in vitro occurred at concentrations higher than those estimated from human exposure data. However, it should be noted that, even though the in vitro POD appears to be higher for these latter compounds, there are limited exposure data for these OPFRs. This is important because, while we do not know the potential health effects of these OPFRs, they have patterns of in vitro activity at comparable concentrations similar to that of the phased-out flame retardant (e.g., BDE-47) and TBBPA.
Where available (TMPP, TDCIPP, and TCEP), we then compared the HTTK-modeled plasma concentrations to the POD values obtained from in vivo rat studies that are currently used to set the minimum risk levels (MRLs, empty triangles). According to ATSDR, an MRL is an estimate of the daily human exposure to a hazardous substance that is likely to be without appreciable risk of adverse health effects over a specified duration of exposure. (92) This comparison showed that for some compounds such as TDCIPP, the in vivo rodent POD lies within the range of human exposure, while for others such as TCEP, toxicity in animal studies is noted at a higher concentration compared with human exposure. Differences in sensitivities between the findings in rodents to that in human-derived cell-based models or toxicokinetic, toxicodynamic and/or exposure characteristics should be taken into account. For some of the less well-studied OPFRs such as IPP, BPDP, EHDPP, and IDDP, even though it may appear that there is a wide margin of exposure between current serum concentrations in humans and predicted toxicological effects, it should be noted that there are relatively sparse exposure data available. Furthermore, due to recent CPSC and EU regulations (Table S1), the use of the nonhalogenated OPFRs is projected to be on the rise.
While this general strategy provides insights into those OPFRs for which concern could be greatest, it has several caveats. For example, it does not consider sensitive populations or even genetic variability within a population, which could significantly change the interpretation. (93) This approach is also limited by several model constraints; e.g., the in vivo POD used to set MRLs for some compounds such as BDE-47 and TBBPA could not be plotted because in silico model parameter estimates were unavailable. Additionally, as only developmentally or neurodevelopmentally associated in vitro assays were used to evaluate the in vitro potency of these compounds, other assays surveying a broader biological space could indicate additional biological disruption.
Nonetheless, these findings indicate that the in vitro activity for some of the OPFRs (i.e., TDCIPP and TPHP) is comparable to that of the phased-out BDEs (e.g., BDE-47) and lies within the range of human exposure (TPHP). Importantly, animal data are sparse for many of these OPFRs, but for compounds where data do exist, the in vitro activity appears to be at levels comparable to the in vivo PODs for some compounds (e.g., TDCIPP). Hence, it is imperative to consider novel strategies for integrating data to provide rapid and timely relevant information for human health protection, especially for sensitive populations, to complement time and cost intensive animal studies.

Epidemiological Evidence

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Epidemiological evidence points to concern about PBDEs and, more recently, OPFRs. PBDEs are well-established neurodevelopmental toxicants (see Table 2). In a meta-analysis, Lam and co-workers concluded that there was sufficient evidence supporting an association between developmental PBDE exposure and reduced IQ. (94) A similar conclusion was also reached by the extensive review of the literature conducted by the National Academies of Sciences, Engineering and Medicine, although their work was restricted to BDE-47. (93)
Table 2. Overview of Data for Specific OPFRs, Including CAS Registry Numbers, Major Uses, U.S. Production Volumes for 2015, a Summary of Biomonitoring Data, Major Regulatory Actions, and Risk Assessment/Epidemiological Dataa
chemical/CAS Registry No.usesU.S. production volume (2015)biomonitoringregulatory actionsassociations in epidemiological studies and risk assessment
TPHP/115-86-6industrial; commercial; consumer: foam seating and bedding, plastic and rubber products, and nail polish1–10 million lbincreasing exposure; (30) ubiquitous exposure in NHANES: 92%, higher exposures in women and higher exposures in 6–11-year-old children (71)CA SCP,b reproductive and neurological toxicity; WA CSPAcadverse reproductive outcomes; (107) prenatal exposures associated with decreased IQ and working memory in children; (108) in combination with other OPFRs, increase in behavioral problems (96)
 FDA Indirect Additives used in Food Contact Substances    
TDCIPP/13674-87-8industrial; commercial; consumer: foam seating and bedding10–50 million lbincreasing exposure; (30) ubiquitous exposure in NHANES: ≥92% exposure in women (71)CA Prop 65, carcinogen; CA SCP, carcinogen; WA CSPAin combination with other OPFRs, prenatal exposures associated with decreased IQ and working memory in children; (108) cancer risks of concern for infants; (97) cancer risks of concern for children (99)
IPP/68937-41-7 (mix of isomers)industrial; commercial; consumer: foam seating and bedding, and plastic and rubber products1–10 million lb CA SCP, reproductive and neurological toxicity; WA CSPA 
EHDPP/1241-94-7industrial; commercial; consumer: foam seating and bedding, and plastic and rubber products1–10 million lb WA CSPA; MDHd chemical of high concern 
 FDA Indirect Additives used in Food Contact Substances    
IDDP/29761-21-5industrial; commercial; consumer: foam seating and bedding, and plastic and rubber productswithheld CA SCP, reproductive and neurological toxicity; MDH chemical of high concern 
TMPP (TCP)/1330-78-5industrial; commercial; consumer: plastic and rubber products, lubricants and greases, and other unspecified uses1–10 million lb CA SCP, reproductive and neurological toxicity; WA CSPA 
TCEP/115-96-8industrial, unspecified25000–100000 lbubiquitous exposure in NHANES: ≥89% exposures in 6–11-year-old children (71)CA Prop 65, carcinogen; EU toxic to reproduction; CA SCP carcinogen; WA CSPAin combination with other OPFRs, increase in behavior problems; (96) risks of concern for cancer and reproductive effects for infants (97)
RDP (PBDPP)/57583-54-7industrial; commercial; consumer: plastic and rubber products and other unspecified uses1–10 million lb CA SCP 
TCIPP/13674-84-5industrial; commercial; consumer: foam insulation, building/construction materials, foam seating and bedding products, and electronic products50–100 million lbwidespread exposure in NHANES: ≥61% exposures in 6–11-year-old children and higher exposures in women (71)CA SCP, carcinogen; WA CSPAin combination with other OPFRs, increase in behavior problems; (96) paternal levels associated with decreased fertilization; (95) risks of concern for cancer and reproductive effects for infants (97)
TNBP/126-73-8industrial; commercial; consumer: adhesives/sealants and inks/toners1–10 million lbubiquitous exposure in NHANES: ≥81% exposures in women (71)EU CMRe; MDH chemical of high concern; WA CSPA 
 FDA Indirect Additives used in Food Contact Substances    
a

Data from US EPA’s Chemical Data Reporting, NIH PubChem, and the Chemical Hazard Data Commons.

b

CA SCP indicates CA Safer Consumer Products Program Candidate Chemical List.

c

WA CSPA indicates WA Children’s Safe Products Act Chemicals of High Concern for Children.

d

MDH indicates Toxic Free Kids Act Chemicals of High Concern or Priority Chemicals.

e

EU CMR indicates EU Carcinogen, Mutagen or Reproductive Toxicant.

Evidence from epidemiological studies is now indicating that OPFRs could also be causing adverse effects at ambient exposures. For example, Carignan et al. reported that urinary concentrations of DPHP, a metabolite of several OPFRs, as well as a stand-alone flame retardant, were significantly associated with adverse reproductive outcomes (e.g., failed fertilization and implantation) in 211 U.S. women undergoing in vitro fertilization. (30) Carignan et al. also found that paternal urinary concentrations of BDCIPP were associated with reduced rates of fertilization. (95) Castorina et al. reported evidence of the developmental toxicity of OPFRs, finding that ∑OPFR metabolites measured in pregnant women, and particularly DPHP, were significantly associated with decreased IQ and working memory of their 7-year-old children in the CHAMACOS birth cohort in California. (95) Lipscomb et al. found a significant dose-dependent relationship between the exposure of 92 U.S. children (3–5 years of age) to ∑OPFRs (TCIPP, TCEP, and TPHP) measured using silicone wristbands and teacher ratings of less responsible behavior and more externalizing behavior problems. (96)
Limited risk assessments of OPFRs have been conducted in the past 10 years. The exposure of infants to TCEP, TDCIPP, and TCIPP present in children’s products and residential upholstered furniture was associated with an increased risk of cancer (97) and reproductive effects for TCEP and TCIPP. (72,98) This finding is further supported by a study investigating TDCIPP exposure in U.S. infants. (72) Bradman et al. identified cancer risks exceeding California health risk-based guidelines for TDCIPP in early childhood education environments. (99) Canada issued a draft risk assessment proposing that TCIPP be considered toxic on the basis of human exposure to foam-containing upholstered furniture, but TDCIPP was not considered to pose a threat to human health at current exposures. (100) Authoritative government bodies have evaluated current data and listed several OPFRs as chemicals of concern or known to have specific health hazards (see Table 1 and Table S4).
Overall, data from traditional toxicity testing, new approach methods, epidemiological studies, and risk assessments all indicate health concerns for both halogenated and nonhalogenated OPFRs.

Policy Approaches to OPFRs

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As concerns about OPFR exposure and toxicity have emerged, regulators in the United States, EU, and other jurisdictions have responded with policies that gather data, inform consumers, limit flame retardants of concern, and/or change flammability standards to reduce the need for flame retardants. Such responses have been aimed especially at vulnerable populations (e.g., children) and uses in common consumer products that result in widespread exposures (see Figure 1, Table 2, and Table S1). In the United States, the federal law covering most industrial, commercial, and consumer product chemicals, the 1976 Toxic Substances Control Act (TSCA), was acknowledged to be ineffective. (101) Faced with a lack of regulation for decades, individual U.S. states have stepped in to issue their own regulations. As shown in Figure 1, initial policies for individual OPFR chemicals focused more on data gathering (for example, the Washington Children’s Safe Product Act requiring reporting of chemicals use to the state) or labeling (for example, California Proposition 65 requiring warnings).
The past decade has seen a shift toward changing flammability standards so that flame retardants are not needed in consumer products when they do not provide a significant fire safety benefit. In 2016, the TSCA was updated; although there were potential improvements in the law, it remains to be seen if public health protections from toxic chemicals will be improved. (102,103) Canada, through its Chemical Management Plan, has regulated PBDEs and TCEP. (99) OPFRs and other flame retardants in Canada are being assessed with a recommendation for declaring TCIPP potentially harmful to human health, (100) but these assessments remain constrained to a chemical-by-chemical basis rather than taking a more comprehensive approach to managing OPFRs and other flame retardants as a class.
The trends in Figure 1 and Table S1 indicate that policy makers are increasingly concerned with the use of hazardous chemicals in consumer products, especially those that result in exposures to children. An emerging trend is to require informed substitution (e.g., 2015 Minnesota House Bill 1100-MN 2015) or to avoid regrettable substitution (e.g., 2016 Washington, DC, Carcinogenic Flame Retardant Prohibition Amendment Act (104)), by creating health criteria for replacement chemicals.

Looking Forward

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Here we have shown that, as with PBDEs in the past, OPFRs are now being used in large volumes, are sufficiently persistent to be detected globally, present health hazards, and may cause harm to humans, especially children, at current exposure levels. Given the large number of OPFRs on the market, obtaining the level of evidence a government often requires to regulate each compound would prove to be expensive and lengthy. However, manufacturers and purchasers can make informed choices now to eliminate use of potentially harmful chemicals.
As PBDEs and other organohalogen flame retardants are phased out, instead of replacing them with OPFRs with a similar potential for harm, we suggest pursuing creative “out-of-the-box” strategies such as improved product design to minimize flammability and using alternative materials that are inherently flame resistant. For example, one manufacturer reduced flame retardants in TV cases by removing the power supply from inside the TV. An external power source was used (like in a laptop power cord), thus eliminating the need for flame retardants in the plastic case around the TV. (105) Furthermore, only using flame retardants when they provide a proven benefit can reduce health risks without impacting fire safety.
This paper has shown that the replacement of PBDEs with OPFRs is likely a regrettable substitution. The time has come for manufacturers, with the help of the scientific community, to stop moving from the use of one family of harmful chemicals to the next and to instead find innovative ways to reduce both fire hazard and the use of hazardous chemicals.

Supporting Information

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The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.estlett.9b00582.

  • Plasma bioequivalents using high-throughput toxicokinetic (HTTK) modeling, Tables S1–S9, Figures S1 and S2, and additional references (PDF)

Terms & Conditions

Most electronic Supporting Information files are available without a subscription to ACS Web Editions. Such files may be downloaded by article for research use (if there is a public use license linked to the relevant article, that license may permit other uses). Permission may be obtained from ACS for other uses through requests via the RightsLink permission system: http://pubs.acs.org/page/copyright/permissions.html.

Author Information

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  • Corresponding Author
  • Authors
    • Arlene Blum - Green Science Policy Institute, Berkeley, California 94709, United StatesDepartment of Chemistry, University of California, Berkeley, Berkeley, California 94705, United States
    • Mamta Behl - National Toxicology Program, National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, United States
    • Linda S. Birnbaum - National Cancer Institute at National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, United States
    • Miriam L. Diamond - Department of Earth Sciences, University of Toronto, Toronto, Ontario, Canada M5S 3B1Orcidhttp://orcid.org/0000-0001-6296-6431
    • Allison Phillips - Risk Assessment and Natural Resource Sciences Inc., Arcadis, Raleigh, North Carolina 27607, United States
    • Veena Singla - Program on Reproductive Health and the Environment, Department of Obstetrics, Gynecology and Reproductive Sciences, University of California, San Francisco, San Francisco, California 94143, United States
    • Nisha S. Sipes - National Toxicology Program, National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina 27709, United StatesOrcidhttp://orcid.org/0000-0003-4203-6426
    • Heather M. Stapleton - Nicholas School of the Environment, Duke University, Durham, North Carolina 27708, United StatesOrcidhttp://orcid.org/0000-0002-9995-6517
  • Notes
    The authors declare no competing financial interest.

Acknowledgments

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The authors thank Anna Soehl for coordinating this effort, Shaorui Wang for help with compiling data and references, and Swati Rayasam for help with Figure S1.

References

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  • Abstract

    Figure 1

    Figure 1. Timeline of major regulatory milestones for PBDEs and OPFRs. Policy enactment dates are listed; implementation/compliance dates are often later (see Table S1 for detailed information).

    Figure 2

    Figure 2. Bar chart representing selected median concentrations of total OPFRs and total PBDEs in (A) outdoor air (picograms per cubic meter), (B) water (nanograms per liter), and (C) indoor dust (nanograms per gram). Each bar represents data from a different study (see Tables S2–S5 for details about the studies included and for a more comprehensive list of locations).

    Figure 3

    Figure 3. Flame retardant plasma concentrations measured or estimated from ingestion using data from house dust, breast milk, and/or handwipe samples (colored bars and circles) are compared to the most potent in vitro concentration per chemical (black dots) and in vivo point of departure (POD; triangles). The filled triangles represent rat plasma concentrations based on in vivo POD values (when available), and the empty triangles represent the minimum risk levels (MRLs). The colored bars represent the range of concentrations, and the circles represent the mean, median, or maximum median (see the Supporting Information for further details about these calculations).

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