Reproductive and developmental toxicity following exposure to organophosphate ester flame retardants and plasticizers, triphenyl phosphate and isopropylated phenyl phosphate, in Sprague Dawley rats

Abstract

Two organophosphate esters used as flame retardants and plasticizers, triphenyl phosphate (TPHP) and isopropylated phenyl phosphate (IPP), have been detected in environmental samples around the world. Human exposure primarily occurs via oral ingestion with reported higher concentrations in children. Currently, there are no data to evaluate potential risk from exposure to either TPHP or IPP during fetal development. These short-term perinatal studies in rats provide preliminary toxicity data for TPHP and IPP, including information on transfer to fetus/offspring and across the pup blood-brain barrier. In separate experiments, TPHP or IPP were administered via dosed feed at concentrations 0, 1000, 3000, 10 000, 15 000, or 30 000 ppm to time-mated Hsd:Sprague Dawley SD rats from gestation day (GD) 6 through postnatal day (PND) 28; offspring were provided dosed feed at the same concentration as their dam (PND 28–PND 56). TPHP- and IPP-related toxicity resulted in removal of both 30 000 ppm groups on GD 12 and 15 000 ppm IPP group after parturition. Body weight and organ weights were impacted with exposure in remaining dams. Reproductive performance was perturbed at ≥10 000 ppm TPHP and all IPP exposure groups. In offspring, both TPHP- and IPP-related toxicity was noted in pups at ≥10 000 ppm as well as reduction in bodyweights, delays in pubertal endpoints, and/or reduced cholinesterase enzyme activity starting at 1000 ppm TPHP or IPP. Preliminary internal dose assessment indicated gestational and lactational transfer following exposure to TPHP or IPP. These findings demonstrate that offspring development is sensitive to 1000 ppm TPHP or IPP exposure.

Organophosphate esters (OPE) compounds have increasingly been used as flame retardants (OPFR) and plasticizers in consumer products and building materials over the past 2 decades (Wei et al., 2015; Fu et al., 2022). As noted in a Patisaul (2021) review, the increased use and worldwide exposure of OPEs is concerning especially since OPE exposure begins early in life and the long-term effects of exposure remain unknown (Patisaul et al., 2021). The U.S. Consumer and Products Safety Commission (CPSC) nominated 2 OPEs, triphenyl phosphate (TPHP) and isopropylated phenyl phosphates (IPPs), following concerns of increased hand to mouth exposure in children when used as OPFRs (Butt et al., 2016; Hoffman et al., 2017; Van den Eede et al., 2015). As use as flame retardants was the primary concern for exposure, TPHP and IPP will hereafter be referred to as OPFRs.

OPFRs are either additive or reactive ingredients, they are not chemically bound, and readily migrate out of treated products over time (Jayatilaka et al., 2017; Wensing et al., 2005). Widespread human exposure has been demonstrated by the detection of TPHP, IPP, and/or metabolites in urine, breast milk, and blood samples from populations around the world at concentrations of 1–100 ng/ml or 1–100 ng/g lipids (Araki et al., 2018; Butt et al., 2014; Hoffman et al., 2014, 2015b, 2017; Ospina et al., 2018; Van den Eede et al., 2015). Several studies have reported that children (0.36–798 ng/ml) have higher OPFR metabolite concentrations compared to adults (0.39–104 ng/ml), but very few have investigated factors that influence children’s OPFRs exposure (Butt et al., 2014, 2016; Hoffman et al., 2015a).

Although the use of OPEs as OPFRs is on the rise, the toxicological hazard has not been well-characterized. Specifically, the CPSC is concerned about the potential toxicity related to neurodevelopmental and general developmental effects following hand-to-mouth ingestion in toddlers from the OPFRs. Additionally, there is a body of data suggesting developmental toxicity and neurodevelopmental effects in OPFR exposed populations (reviewed in Doherty et al. [2019], Liu et al. [2020], van der Veen and de Boer [2012]) as well as developmental and neurodevelopmental toxicity in in vitro and alternate animal models (Bajard et al., 2019; Behl et al., 2015; Glazer et al., 2018; Hogberg et al., 2021; Klose et al., 2020). However, no extensive in vivo studies have been conducted yet following developmental exposure. Additionally, there is currently no Average Daily Intake (ADI) value that has been established for these compounds.

Hence, the CPSC nominated these compounds as a part of a larger flame-retardant program to the National Toxicology Program for the evaluation of developmental and neurodevelopmental toxicity. These 2 chemicals were specifically selected for in vivo study to serve as class representatives and anchors for read-across of the entire class. Other ongoing in vitro and alternate animal assessments have been conducted to evaluate their developmental and neurodevelopmental potential in this class of chemicals (Behl et al., 2015; Blum et al., 2019; Hogberg et al., 2021; Klose et al., 2022; OECD, 2022; Patisaul et al., 2021). TPHP and IPP were selected as representatives of the class based on production volume (Crawford et al., 2012; Wei et al., 2015), TPHP being a common backbone to all of the compounds within the class, and the ability to compare toxicity associated in the absence (TPHP) versus presence of a side-chain (IPP) (Crawford et al., 2012) (Table 1). Herein, we present data from short-term repeat dose studies spanning gestation, lactation, and early offspring development that included reproductive parameters, F0 and F1 body weight, limited organ weights, and cholinesterase and accompanying toxicokinetics to evaluate transfer of TPHP and IPP from dam to fetus/pup and across the pup blood-brain barrier. Although these were preliminary studies, cholinesterase was included because it is a well-known marker associated with acute neurotoxicity following OPE pesticide exposure and previous studies have shown serum cholinesterase expression to be the most sensitive serum and gene expression marker following acute exposure to these compounds (NTP, 2018). However, its role following developmental exposure to this class still remains unknown. Feed was used as the route of exposure to simulate human exposure including mouthing behavior in toddlers, the primary concern for this class of chemicals (Butt et al., 2014). Due to a lack of evidence in the literature at the time this study was conducted, dose selection was intended to sufficiently challenge exposed animals and capture dose-response information for maternal and pup toxicity including a high enough top-dose level to be conclusive for hazard classification.

Table 1.

Structure of compounds and percent composition in IPP mixture^a

Chemical name	Structure	Percent composition
TPHP		21.5
Mono-IPP		36.9
Bis-IPP		21.9
Tris-IPP		8.5

^aThe IPP mixture is composed of TPHP and phenyl phosphate with mono-, bis-, and tris-isopropyl side chains.

Materials and methods

 

 

Chemical and reagent procurement

Chemical analysis was conducted at MRIGlobal (Kansas City, Missouri). The test articles, TPHP (CAS No. 115-86-6; lot A0343614) and IPP (CAS No. 68937-41-7; lot 20111210) were procured from Acros Organics (Fair Lawn, New Jersey) and from AmplaChem (Carmel, Indiana), respectively. The identity of TPHP and IPP was confirmed by infrared spectroscopy. The isomeric composition of IPP was evaluated by liquid chromatography-mass spectrometry (LC-MS) and determined to be 21.5% TPHP, 36.9% mono-isopropylated-IPP (mono-IPP), 21.9% bis-isopropylated IPP (bis-IPP), and 8.5% tris-isopropylated IPP (tris-IPP) (Table 1). Remaining components included propofol or unidentified impurities. The purity for TPHP and IPP (as sum of isomers) was determined to be approximately 99% using high-performance liquid chromatography (HPLC) with ultraviolet (UV) or gas chromatography (GC) with flame ionization detection (FID), respectively.

The internal standard for quantitation of TPHP (¹³C₁₈-TPHP lot YS-2016-009B1) was purchased from Isosciences (King of Prussia, Pennsylvania). Individual ortho-substituted IPP compounds were obtained from Wellington Laboratories, Inc (Guelph, Ontario, Canada); 2-Isopropylphenyl Diphenyl Phosphate; Lot No. 2IPPDPP0716, 4-Isopropylphenyl Diphenyl Phosphate; Lot No. 4IPPDPP0716, Bis(2-isopropylphenyl) Phenyl Phosphate; Lot No. B2IPPPP0716 and Bis(4-isopropylphenyl) Phenyl Phosphate; Lot No. B4IPPPP0716. Sprague Dawley (Hsd:Sprague Dawley SD) female rat plasma and amniotic fluid, gestation day (GD) 18 fetuses, and postnatal day (PND) 4 male pups to be used in calibration standards and quality control (QC) sample preparation were obtained from BioIVT (Westbury, New York). Other reagents and standards were obtained from commercial sources.

Dose formulation and exposure

Dosed-feed formulations (0, 1000, 3000, 10 000, 15 000, 30 000 ppm) of TPHP and IPP were prepared in NIH-07 Open Formula (certified, irradiated) meal feed (Zeigler Brothers Inc, Gardners, Pennsylvania). All dosed feed required one premix except for the 30 000 ppm formulation, which required preparation of 2 premixes to allow thorough mixing and elimination of clumping. Test article was blended with the feed using a twin-shell blender for approximately 15 min. Pre- and post- (collected from animal room on last day of study) feed formulations for TPHP and IPP were analyzed using a validated HPLC with UV or GC with FID, respectively, and all formulations were within 10% of target concentration. Formulations of 30 000 ppm TPHP and IPP were homogenous with relative standard deviation (RE) of 2%. Feed formulations were stored at −20°C and were used within 42 days of preparation during study conduct. As previously established, TPHP and IPP in feed are stable up to 42 days when stored at −20°C.

Animals

Studies were conducted according to Good Laboratory Practices at Battelle Memorial Institute (West Jefferson, Ohio) under the auspices of the National Toxicology Program. Eleven to 12-week-old time-mated female SD (Hsd:Sprague Dawley SD) rats (F0) were obtained from Harlan Laboratories (now Envigo) (Indianapolis, Indiana) and housed in solid-bottom, polycarbonate cages lined with irradiated hardwood-chip bedding (Sani-Chips cage litter, P.J. Murphy Forest Products, Montville, New Jersey). F0 females were singly housed except when with their respective litters during lactation. Upon receipt, irradiated NIH-07 certified rodent diet (Zeigler Brothers Inc) was available ad libitum. Beginning on GD 6, irradiated NIH-07 meal feed, either untreated or treated, was available ad libitum. Tap water (Village of West Jefferson, Ohio) was available ad libitum via an automatic water delivery system.

Animal rooms were maintained at 69–75°F, humidity 35%–65%, with 10 filtered room air changes per hour, and a 12-h light/dark cycle per day.

Study design

The TPHP and IPP studies were conducted separately. The study design is the same for both unless noted otherwise (Figure 1).

Figure 1.

Study design summary. Studies examining perinatal exposure to TPHP or IPP followed the same study design outline though conducted separately. Abbreviations: IPP, isopropylated phenyl phosphate; TPHP, triphenyl phosphate.

The day of confirmed mating via presence of a vaginal plug or sperm was considered GD 0. Time-mated females (n = 15 or 22 in groups selected for biological sampling) were randomly assigned to control and experimental groups on GD 3. F0 females were provided control or dosed feed (1000, 3000, 10 000, 15 000, or 30 000 ppm) on GD 6 through PND 28. Adult females were observed twice daily for mortality and morbidity, clinical observations conducted once daily and body weights were recorded once daily during gestation, beginning on GD 3 and every 3–4 days during lactation (PND 1–28). Care was taken so that dams in the process of parturition were not disturbed. Feed consumption was recorded for overlapping 3–4 day periods GD 3 through PND 28. Feed intake was calculated as either TPHP or IPP in feed (ppm) multiplied by feed consumed (kg) and divided by animal bodyweight (kg), on a per-cage basis.

Parturition checks were performed twice-daily beginning GD 20 until pups were delivered, or through GD 25. The date of delivery was considered the start of lactation and designated PND 0. The number of live and dead pups per litter was recorded daily and individual pups were sexed and weighed every 3–4 days from PND 1 to 28. Clinical observations for pups were recorded on PND 1, 4, 7, 17, and 24.

All pups not selected for PND 28 biological sampling were weaned on PND 28. Offspring were group housed by sex and exposure group with up to 5 animals per cage. F1 pups were administered the same dosed-feed concentrations as their respective dam from PND 28 through study termination. Offspring were observed twice daily for mortality and morbidity. Clinical observations and body weights were recorded on PND 35 and 42 (females) and PND 42, 49, and 53 (males). Female pups were examined daily beginning on PND 23 until the achievement of vaginal opening (VO) or PND 45. Examination of males for balanopreputial separation (BPS) began on PND 35 and continued until acquisition or PND 53. Body weight was recorded on the day of acquisition for each animal.

Biological sampling

Biological sampling on GD 18 and PND 4 served as an initial assessment to investigate gestational and lactational transfer. Biological samples collected on GD 18 include dam plasma, amniotic fluid, and fetus. On PND 4, whole pups (1 per sex) were collected for sampling. On GD 18, 3 randomly selected pregnant dams from the 0, 1000, and 10 000 ppm groups were heavily sedated by CO₂ inhalation, euthanized via decapitation, and trunk blood collected into tubes containing tripostassium ethylene diaminetetraacetic acid (K₃EDTA). Blood samples were processed to plasma and stored at −70°C. Amniotic fluid was collected at necropsy and fetuses from these litters euthanized, pooled by litter, and euthanized by decapitation prior to flash freezing (head and body) in liquid nitrogen. Samples were stored frozen in −70°C. On PND 4, 3 litters from the 0, 1000, and 10 000 ppm groups were randomly selected. One pup per sex from each dam was euthanized via decapitation and individually flash frozen in liquid nitrogen (head and bodies) and stored at 70°C. Due to a low number of littering dams (including controls), samples from dams were not collected on PND 4.

Ten dams per group were randomly selected from the 0, 1000, 3000, 10 000, and 15 000 (TPHP only) ppm exposure groups for bio sampling on PND 28. Dams were anesthetized with CO₂/O₂ and blood collected via intra-cardiac puncture. Blood samples were collected in tubes containing K₃EDTA which was processed for assessment of acetylcholinesterase (AChE) and butyryl cholinesterase (BChE) activity. The brain, liver, and thymus of each animal were excised and weighed. All surviving dams not selected for biological sampling on PND 28 were euthanized via CO₂ inhalation and disposed appropriately without further evaluation.

Selected PND 28 pups (8–11 per sex) from the 0, 1000, 3000, 10 000, and 15 000 ppm groups were used for cholinesterase activity assessments. When possible, pups were selected from different litters. Due to toxicity, no samples were collected for cholinesterase activity measurements in the 15 000 ppm IPP group; the remaining viable pups in the 10 000 ppm IPP group (10 males and 11 females from 4 and 3 litters, respectively) were used for blood cholinesterase activity. Offspring were anesthetized with CO₂/O₂ and blood collected via intracardiac puncture. Samples were processed for assessment of AChE and BChE activity. Following blood collection, the pups were euthanized and the whole brain (0, 1000, and 3000 ppm) excised, weighed, and rinsed with isotonic saline and flash frozen. Tissues were stored at approximately −80°C until assessment for brain cholinesterase activity.

A separate set of PND 28 pups (8–10 per sex, 1/sex/litter) were selected from the 0, 1000, and 10 000 (TPHP only) ppm groups for assessment of TPHP and IPP plasma and brain AChE and BChE concentrations. Samples were processed to plasma, flash frozen and stored at −80°C until analysis. Following blood collection, the pups were euthanized, and the brain, liver, and thymus of each animal excised and weighed. The brain was rinsed with isotonic saline, flash frozen, and stored at −80°C until used for analysis of TPHP and IPP concentrations.

Acetylcholinesterase and butyryl cholinesterase activity assay

Blood samples from PND 28 dams and male/female offspring were analyzed for AChE and BChE activity as described previously (McGarry et al., 2013). Briefly, 80 μl of each whole blood sample was diluted with 720 μl of Millipore water and mixed with 800 μl of HemogloBind (H0145-500; BioTech Support Group, LLC). The samples were placed on a rocker for 15–20 min at room temperature (RT), centrifuged at 9000 × g for 2 min at 4°C and the supernatant was removed, split into a primary and a retention sample, and stored at ≤−70°C. Whole brains collected from PND 28 offspring were flash frozen in liquid nitrogen, and stored at ≤−70°C until processing. At the time of processing, samples were mechanically homogenized with a bead beater homogenizer at approximately 200 mg/ml in homogenization buffer (1% TritonX-100 in PBS, with protease inhibitor tablets added [Thermo Scientific Pierce: 88266 or equivalent]). Samples were subsequently rocked for at least 1 h at 2°C–8°C to ensure solubilization of the membranes. The homogenized samples were then centrifuged at 10 000 × g for 2 min at 4°C. The supernatant was removed and retained at ≤−70°C for cholinesterase (ChE) activity analysis.

On the day of analysis, the samples were thawed at room temperature. The sample master box was prepared by diluting each sample 2-fold (blood) or 10-fold (brain) in assay buffer (1× PBS). Activity was then assessed using a spectrophotometric assay conducted in a manner similar to Ellman et al. (1961) and McGarry et al. (2013). Samples were diluted an additional 2-fold into the test plate by adding 100 μl of sample to a total volume of 200 μl in each well of a 96-well plate. To reduce the background caused by free thiols present in tissue, brain samples were first incubated at RT with the indicator 5,5-dithio-bis-(2-nitrobenzoic acid) (DTNB; 0.5 mM final) for 10–15 min. A 1:1 mixture of 20× substrate: indicator (substrate was acetylthiocholine iodide [ATC] or S-butyrylthiocholine [BTC] iodide at a final concentration of 1 mM or 3 mM respectively; indicator was 5,5′-dithiobis at a final concentration of 0.5 mM) was prepared and added to the samples in the plate. The plate was sealed, and a spectrophotometric kinetic read was performed by reading the plate every 47 s for a period of 10 min at a wavelength of 412 nm. The pathlength for each well was standardized to 1 cm using the Pathlength Correction feature in the Gen5 software. Additionally, to determine the amount of protein present in each sample, a bicinchoninic acid (BCA) assay was performed using the Pierce BCA Protein Assay Kit (catalog number 23227, lot number PK211391) from Thermo Scientific. ChE data were subsequently reported as active enzyme units per gram of total protein.

TPHP and IPP analyte quantitation

The quantitation of TPHP and IPP isomers was performed using the liquid chromatography with tandem mass spectrometry (LC-MS/MS) method. On the day of analysis, plasma and amniotic fluid samples were stabilized by the addition of 1100 mM sodium fluoride to achieve a final concentration of 4.8 ng/ml and thawed on ice. To a 100-µl aliquot of each stabilized sample, 5 µl of 2.75 M aqueous formic acid and 10 µl of internal standard (approximately 1 µg/ml ¹³C₁₈-TPHP in ethanol) were added and vortex-mixed after each addition. For TPHP samples, 285 µl of acetonitrile were added, vortex-mixed, and placed on ice for 15 min, then centrifuged for 5 min at 16 000 × g. The supernatant was filtered through individual Phree cartridges to remove phospholipids and the eluent was transferred to autosampler vials for analysis. For IPP, the entire acidified sample with internal standard was transferred to a 15-ml centrifuge tube containing a Phree cartridge. Then 285 µl of acetonitrile was added, vortex-mixed, and placed on ice for 15 min. After standing on ice, samples were centrifuged for 5 min at 16 000 × g and the resulting filtrate was analyzed by LC-MS/MS.

Frozen fetus (GD 18), whole pup (PND4), and brain samples (PND 28) were homogenized in the presence of sodium fluoride, subsampled for analysis, and acidified as described for control tissue homogenates. Aliquots (100 mg) of each homogenized sample were transferred to individual microcentrifuge tubes, extracted with 280 µl of acetonitrile, placed on ice 15 min, centrifuged for 5 min at 16 000 × g (TPHP), and then passed through a 1-ml Phree phospholipid removal cartridge (by positive pressure for TPHP; by centrifugation at 1800 rpm for 5 min for IPP). The filtrate was transferred into vials for LC-MS/MS analysis. An additional centrifugation step (14 000 rpm for 5 min) was incorporated into the IPP preparation procedure for pup homogenate extracts prior to phospholipid removal.

Statistical methods

Before statistical analysis, F0 body weight and food consumption extreme values were identified by the outlier test of Dixon and Massey (Dixon and Massey, 1957). For F1 pup weights, all observations across dose groups were fit to a linear mixed effects model with a random litter effect, and the residuals were tested by dose group for outliers using Tukey’s outer fences method (Tukey, 1977). All outliers were examined by in-house experts, and implausible values were eliminated from the analysis. Statistical analyses were performed using SAS software version 9.4 (SAS Institute, Cary, North Carolina). F0 body weights were analyzed using Jonckheere’s trend test (Jonckheere, 1954) and Williams’ or Dunnett’s (pairwise) test depending on detection of a significant trend at 0.01 level (Dunnett, 1955; Williams, 1971, 1972). Gestational length, feed consumption, litter size, and survival endpoints were analyzed using the nonparametric multiple comparison methods of Shirley(Shirley, 1977) (as modified by Williams [Williams, 1971] or Dunn [Dunn, 1964]). Jonckheere’s trend test was used to assess the significance of dose-related trends (Jonckheere, 1954). For reproductive performance endpoints, statistical analysis was performed by Cochran-Armitage (trend) and Fisher’s Exact (pairwise) 2-sided tests (Gart et al., 1979). F1 body weights were first adjusted for litter size and then analyzed using mixed effects linear models with a random litter effect, and a Dunnett-Hsu adjustment (Hsu, 1992) was used to adjust for multiple comparisons. Dosed groups different from control at p ≤ .05 were considered statistically significant.

Results

Herein, we summarize the key findings of the study. All data from the toxicological evaluation and chemical transfer measurement described here are available in the National Toxicology Program’s Chemical Effects in Biological Systems (CEBS) database (https://doi.org/10.22427/NTP-DATA-002-02974-0007-0000-1; last accessed December 23, 2022).

The chemical transfer measurements are reported from TPHP and IPP measured concentrations of biological samples collected on GD 18 (dam plasma and amniotic fluid and fetus), PND 4 (whole pup), and PND 28 (F1 plasma and brain).

 

Survival, clinical observations, and body weight

These preliminary studies tested a large range of exposure (1000 ppm to 30 000 ppm) to ensure maximum tolerated dose was appropriately captured. All 30 000 ppm TPHP and IPP dams were removed from study (GD 12) due to overt toxicity as evidenced by clinical signs and a lack of body weight gain (CEBS I03, I05). The 15 000 ppm IPP group was removed after parturition due to litter loss and clinical observations. In addition, survival issues were also noted at 15 000 ppm TPHP and 10 000 ppm IPP with 3 and 2 dams lost during lactation, respectively. Clinical observations related to toxicity during gestation and/or lactation for both compounds included hunched posture, thin appearance, and/or ruffled coat at 10 000 ppm TPHP and 15 000 ppm IPP. Additional clinical observations were also noted at lower exposure levels for IPP (≥3000 ppm) and included hyperactivity and twitches and/or tremors (CEBS I05).

Consistent exposure effects on body weight were observed beginning on GD 7 and continued through lactation with dams exposed to ≥10 000 ppm TPHP weighing less (6%–25%) than controls (Figure 2A). At necropsy, relative liver weights were higher in dams exposed to ≥3000 ppm TPHP (25%–48%). Additionally, a significant decrease in the relative weight of the thymus was noted in the 10 000 and 15 000 ppm TPHP dams (52%–65%) and an increase in relative brain size in the 15 000 ppm TPHP dams (approximately 5%) (CEBS PA06). Organ weight changes were not associated with any gross macroscopic observations.

Figure 2.

Mean body weight of (A) TPHP and (B) IPP exposed F0 dams during gestation and lactation. Graphs depict mean ± SEM. Abbreviations: IPP, isopropylated phenyl phosphate; TPHP, triphenyl phosphate.

Beginning GD 8 and continuing through lactation, dams consistently weighed less than controls in the 3000 (5%–11%) and 10 000 (6%–18%) ppm IPP exposure groups (Figure 2B). Relative liver (3000 and 10 000 ppm) and brain weights (10 000 ppm) of IPP dams were higher compared to controls (15%–24% and 7%, respectively) (CEBS PA06), but no effect on thymus weight was noted (≤10 000 ppm). These liver weight changes were not associated with any gross macroscopic observations.

Feed consumption and chemical intake

For the majority of gestation and early lactation, feed consumption (g/kg/day) for the ≥10 000 ppm TPHP dams was 17%–50% higher than dams in the control group (Figure 3A). Increased feed consumption (up to 68%) was only observed during gestation in dams exposed to 10 000 ppm IPP (Figure 3B), it remains unclear if feed consumption is related in part to wastage. The calculations of chemical intake for TPHP and IPP (mg/kg/day) during gestation and lactation can be found online (CEBS I08). Across the 1000–15 000 ppm TPHP exposure groups, dams were exposed to approximately 2–2.5 times more TPHP during lactation than gestation. Dams in the 1000 and 3000 ppm IPP groups were exposed to approximately 2.5 times the amount of IPP during lactation than gestation. Exposure in the 10 000 ppm IPP dams was consistent across gestation and lactation. In general, intake of both TPHP and IPP increased proportionally to the feed concentration in lower exposure groups (1000 and 3000 ppm). The increase in food consumption at ≥10 000 ppm TPHP and IPP exposure groups resulted in the chemical intake to be greater than the proportional feed concentration.

Figure 3.

Mean food consumption of (A) TPHP and (B) IPP exposed F0 dams during gestation and lactation. Graphs depict mean ± SEM. Abbreviations: IPP, isopropylated phenyl phosphate; TPHP, triphenyl phosphate.

Reproductive performance and litter parameters

The 30 000 ppm TPHP and IPP dams were removed from the study due to toxicity on GD 12 thus no reproductive performance parameters are reported. Exposure levels ≥10 000 ppm TPHP negatively affected reproductive performance. The number of live pups was lower at 15 000 ppm TPHP (34% less than controls) and offspring survival from PND 1–4 and from 5 to 28 was 89% and 62% of controls, respectively (Table 2). Although it did not reach significance, pup survival was also decreased at 10 000 ppm TPHP during the same periods. Combined male and female pup weights were significantly lower than controls for all timepoints measured for the 10 000 ppm (12%–41%) and 15 000 ppm (19%–68%) TPHP exposure groups (Table 2). IPP exposure affected reproductive performance in all exposure groups (≥1000 ppm). Decreases in the number of live pups were noted at ≥3000 ppm with the largest impact in the 15 000 ppm IPP (76% less than controls). Pup survival was also affected at all exposure levels of IPP for PND 0 (7%–56% less than controls) and the 15 000 ppm IPP exposure group was removed by PND4 due to total litter loss. Pup survival for 3000 ppm and 10 000 ppm IPP were 27%–71% as compared to controls on PND 1–4 and 16%–25% as compared to controls on PND 5–28, respectively. Combined male and female pup weights were significantly lower than controls from PND 7–21 for the 1000 ppm exposure group (10%–15%) and for all timepoints measured for the 3000 ppm (16%–29%) and 10 000 ppm IPP (34%–65%) exposure groups (Table 2).

Table 2.

Mean TPHP and IPP reproductive performance and litter parameters

	PPM
	0	1000	3000	10 000	15 000
TPHP
No. of time-mated females	22	22	15	22	15
No. of pregnanta	19	19	12	18	11
Female littered	15	16	12	15	11
Gestation length	22.2 ± 0.1	22.3 ± 0.1	22.3 ± 0.1	22.1 ± 0.1	22.4 ± 0.2
Litter parameters (PND0)
No. of viable litters	15	16	12	15	11
Fecundity ratioc	94%	100%	100%	100%	100%
Total no. of pups	12.2 ± 0.8	11.9 ± 0.9	13.4 ± 0.5	12.9 ± 0.6	11.8 ± 0.5
No live pups	11.8 ± 0.8	11.7 ± 0.9	13.2 ± 0.5	12.7 ± 0.7	11.7 ± 0.5
% Survival PND 0	96 ± 2	98 ± 1	98 ± 1	98 ± 1	89 ± 9
% Survival PND 1–4	99 ± 1***	100 ± 0	100 ± 0	90 ± 6	89 ± 3**
% Survival post cull PND 4–28	97 ± 2***	94 ± 5	97 ± 1	86 ± 8	62 ± 11**
PND 1 pup weights
Combined	7.2 ± 0.13***	7.0 ± 0.11	7.4 ± 0.13	6.4 ± 0.19**	5.9 ± 0.19**
Male	7.4 ± 0.17***	7.2 ± 0.11	7.7 ± 0.13	6.6 ± 0.2**	6.1 ± 0.18**
Female	7.1 ± 0.13***	6.9 ± 0.12	7.2 ± 0.14	6.2 ± 0.18**	5.7 ± 0.2**
IPP
No. of time-mated females	22	22	15	22	15
No. of pregnanta	15	21	12	16	12
Female littered	12	17	12	13	8
Gestation length	22.5 ± 0.2	22.2 ± 0.1	22.3 ± 0.2	22.3 ± 0.2	22.8 ± 0.3
Litter parameters (PND 0)
No. of viable litters	12	17	12	12	8
Fecundity ratioc	100%	94%	100%	100%	80%
Total no. of pups	12.1 ± 1.2***	12.3 ± 0.5	10.8 ± 0.9	10.7 ± 1.1	8.8 ± 1*
No. live pups	13 ± 0.7***	10.5 ± 0.6	9.9 ± 0.9	10.5 ± 1	7 ± 2.1*
% Survival PND 0	91 ± 8***	85 ± 3*	82 ± 8*	74 ± 12*	40 ± 16**
% Survival PND 1–4	99 ± 1***	98 ± 1	72 ± 12*	29 ± 12**	0 ± 0**
% Survival post cull PND 5–28	99 ± 1***	96 ± 3	83 ± 10*	74 ± 14*	NAb
PND 1 pup weights
Combined	7.4 ± 0.14***	6.8 ± 0.14	6.0 ± 0.34**	4.9 ± 0.37**	NAb
Male	7.7 ± 0.16***	6.8 ± 0.13*	6.1 ± 0.41**	5 ± 0.38**	NAb
Female	7.2 ± 0.16***	6.6 ± 0.15	5.8 ± 0.34**	4.8 ± 0.36**	NAb

^*Significantly different from control at p ≤ .05;

^** p ≤ .01.

^***Significant trend at p ≤ .01.

^aIncludes 3 pregnant animals per group removed on GD 18 and PND 4 for biological sampling from 0, 1000, and 10 000 ppm dose groups for biological sampling.

^bValues were not available (NA) due to early termination of the groups on PND 1–3.

^cCalculated as the number of litters divided by the number of pregnant females (excluding those taken for biological sampling).

Offspring development

Male and female offspring in the 10 000 and 15 000 TPHP exposure groups were significantly smaller throughout the study (approximately 11%–70% as compared to controls). Males and females in both the 1000 and 3000 ppm TPHP exposure groups had similar body weights, within approximately 10% of controls, for the majority of the study (Figs. 4A and 4B). Evaluation of male pubertal onset revealed that all 15 000 ppm TPHP offspring failed to achieve BPS and a significant delay (+2 to +13 days) was observed in the 1000–10 000 ppm TPHP groups when compared to controls. However, onset of puberty was not impacted when adjusting for body weight at weaning, on PND 28, for the 1000 and 3000 ppm groups and only slightly mitigated at 10 000 ppm TPHP, changing from +13 day to a +10 days delay. All females in 15 000 ppm TPHP treatment groups and most females in the 10 000 ppm group did not achieve VO and an approximate 3 days delay of VO was observed in the 3000 ppm TPHP group, 2 days when adjusting for body weight at weaning (Table 3). No impact on VO was noted for the 1000 ppm TPHP exposure group. At necropsy, the relative brains of male and female offspring were 55%–126% larger than controls when exposed to ≥10 000 ppm TPHP. The relative thymus weight was significantly reduced (59%) in the 10 000 ppm TPHP males (CEBS PA06).

Figure 4.

Mean body weight (g) of male and female offspring (TPHP [A and B] and IPP [C and D], respectively). Graphs depict mean ± SEM. Abbreviations: IPP, isopropylated phenyl phosphate; TPHP, triphenyl phosphate.

Table 3.

Summary of balanopreputial separation and vaginal opening of offspring exposed to THPH or IPP

	ppm	0	1000	3000	10 000
TPHP a
Male	Day of balano-separationb	37.6 ± 0.6****	39.7 ± 0.6*	40 ± 0.3*	50.6 ± 0.5**
			+2.1	+2.4	+13
	No. examined (litters)	31 (10)	52 (10)	59 (12)	32 (9)
	BW on day of BPS (g)	134.9 ± 3.7	140.3 ± 2.7	139.5 ± 2.8	129.9 ± 5.8
Female	Day of vaginal openingb	36.6 ± 0.4****	37.2 ± 0.3	39.2 ± 0.5**	c
			+0.6	+2.6
	No. examined (litters)	55 (11)	45 (10)	62 (12)
	BW on day of VO	108.8 ± 2.7	108.2 ± 2.7	110.2 ± 2
IPP
Male	Day of balano-separationb	41.6 ± 0.6***	42 ± 0.4	43.1 ± 0.7
			+0.4	+1.5
	No. examined (litters)	32 (7)	60 (16)	17 (7)
	BW on day of BPS (g)	167.2 ± 3.8***	164.6 ± 3.9	151.6 ± 5.1
Female	Day of vaginal openingb	36.4 ± 0.5****	36.5 ± 0.4	38.8 ± 0.7*
			+0.1	+2.4
	No. examined (litters)	38 (10)	43 (14)	23 (7)
	BW on day of VO (g)	111.2 ± 2.5	108 ± 2.8	115 ± 4.4

^aNo animals in the 15 000 ppm TPHP achieved BPS or VO during time of examination.

^bMean analysis, litter mean ± SE.

^cOnly one animal in 10 000 ppm TPHP achieved vaginal opening during time of examination. No statistical analysis able to be completed.

^*Significantly different from control at p ≤ .05;

^** p ≤ .01.

^***Significant trend at p ≤ .05,

^**** p ≤ .01.

BW, body weight; BPS, balanopreputial separation; VO, vaginal opening.

Although male and female offspring across all IPP exposures continuously weighed less compared to controls, differences in body weight for the 1000 and 3000 ppm groups only reached significance at sporadic intervals throughout gestation and lactation (Figs. 4C and 4D). All animals in the 10 000 ppm were selected for bioanalysis at PND 28, hence weights are not reported past this study day and no pubertal endpoints were assessed in the 10 000 ppm exposure group. For the IPP exposed males evaluated for pubertal factors, no impacts on BPS were observed; however, a delay (approximately +2 days) in VO was observed in 3000 IPP female offspring, even with adjustment for weaning weight (Table 3 and CEBS R16). The relative weight of the brain was also significantly increased at 10 000 ppm IPP for males and females (61%–89%) (CEBS PA06). Thymus weights were only collected in 1000 ppm offspring and no differences were observed.

Cholinesterase activity in blood

The AChE and BChE activity in blood decreased in a dose-dependent manner in dams as TPHP dose increased reaching significance at ≥3000 ppm TPHP. In the 15 000 ppm, TPHP exposure group AChE and BChE were approximately 13% activity of controls. TPHP exposed offspring had increased or equivalent AChE and BChE activity levels at 1000 and 3000 ppm compared to controls. A reduction in AChE and BChE activity was observed in male (42%–63% of controls) and female (46%–75%) offspring at ≥10 000 ppm TPHP (Figs. 5A and 5B). The reduction in offspring cholinesterase activity only reached significance in the BChE of 15 000 ppm TPHP exposed females.

Figure 5.

Blood AChE and BChE activity in dams and pups exposed to TPHP (A and B) or IPP (C and D). Each point represents group mean as percent of control (control = 100%) and indicates a dose-dependent decrease in activity. TPHP dams n = 10 (except 15 000 ppm n = 8), male offspring n = 7–8/group and female offspring n = 7–8/group. IPP dams n = 9–10/group, male offspring n = 10 (except 3000 ppm n = 7) and female offspring n = 8–11/group. Abbreviations: AChE, acetylcholinesterase; BChE, butyryl cholinesterase; IPP, isopropylated phenyl phosphate; TPHP, triphenyl phosphate.

The AChE and BChE activity in blood from dams and offspring significantly decreased in a dose-dependent manner as IPP dose increased across all IPP exposure groups. In IPP exposed dams, the AChE and BChE activity ranged from 19%–8% to 29%–11% of controls, respectively, across exposure groups. A similar reduction in offspring was reported, with the AChE and BChE activity 47%–14% and 60%–26% of controls, respectively, in males and 50%–18% and 61%–25% of controls, respectively, in females. No sex differences were observed in offspring (Figs. 5C and 5D).

Cholinesterase activity in brain

Cholinesterase activity in the brain is reported for offspring only and brain samples were not collected from the dams for evaluation. In general, a reduction in brain AChE and BChE activity occurred as the TPHP concentration increased. In males, AChE activity was affected in all TPHP exposed offspring groups and BChE activity was decreased in the 3000 ppm and 10 000 ppm exposure groups, ranging 80%–40% and 75%–45% of controls, respectively. Across all exposure groups, the AChE and BChE activity in TPHP exposed females ranged 103%–56% and 112%–59% of controls, respectively. The decrease in AChE and BChE activities was significant in the ≥10 000 ppm and ≥3000 ppm TPHP exposed females, respectively. A significant sex difference in brain activity was observed in offspring exposed to TPHP at 1000 and 3000 ppm for AChE and 1000 ppm for BChE, with males having decreased levels compared to females (Figs. 6A and 6B).

Figure 6.

Brain AChE and BChE activity in male and female offspring exposed to TPHP (A and B) or IPP (C and D). Each point represents group mean as percent of control (control = 100%) indicating a dose-dependent decrease in activity. TPHP dams n = 10 (except 15000 ppm n = 8), male offspring n = 7–8/group and female offspring n = 7–8/group. IPP dams n = 9–10/group, male offspring n = 10 (except 3000 ppm n = 7) and female offspring n = 8–10/group. Abbreviations: AChE, acetylcholinesterase; BChE, butyryl cholinesterase; IPP, isopropylated phenyl phosphate; TPHP, triphenyl phosphate.

Male and female offspring exposed to either 1000 or 3000 ppm IPP had significantly reduced AChE activity in the brain, with males ranging 72%–60% of controls and females ranging 63%–59% of controls. Male offspring exposed to 3000 ppm IPP also had significantly reduced BChE activity (77% of controls), as did female offspring exposed to 1000 ppm IPP (73% of controls). No significant sex differences in brain activity were observed in offspring exposed to IPP (Figs. 6C and 6D).

Analysis for TPHP and IPP concentration

The ability to transfer TPHP and IPP during gestation and lactation and the analytes to cross the blood-brain barrier were assessed in a limited design. An analytical method was developed to quantitate TPHP and isomers of IPP in plasma, amniotic fluid, fetus/pup, and brain.

Gestational and lactational transfer

Overall, maternal transfer via gestation and lactation was evident by detection of TPHP and IPP isomers in GD 18 fetuses and PND 4 offspring (Figure 7A). On GD 18, the increase in TPHP concentration in maternal plasma was greater than the proportional exposure concentration with a 25-fold increase in measured TPHP corresponding with a 10-fold increase in exposure concentration. The increase in TPHP concentration is not due to an increase in chemical intake as this was similar to the proportional exposure (13 vs 10). In amniotic fluid, 2 samples were below the lower limit of quantitation (approximately 8.0 ng/ml) in the 1000 ppm TPHP exposure group and one sample was excluded due to technical issues in the 10 000 ppm TPHP exposure group. The TPHP concentration in amniotic fluid was 13.5 ng/ml in the 1000 ppm TPHP exposure group and 58.1 ng/ml in the 10 000 ppm TPHP exposure group. Exposure concentrations at GD 18 did not increase in a dose proportional manner as there was approximately a twenty-fold increase in fetuses exposed to 1000 ppm TPHP as compared to 10 000 ppm TPHP (31 ng/g and 658 ng/g, respectively). On PND 4, a 33- to 38-fold increase was observed in both male and female offspring exposed to 10 000 ppm TPHP as compared to 1000 ppm TPHP (4415–4471 ng/g and 114–134 ng/g, respectively).

Figure 7.

Mean concentrations of TPHP (A and B) and IPP components (C and D) during gestation and lactation. Due to toxicity, no samples were collected from dams in 10 000 ppm IPP group on PND 4. TPHP and IPP GD 18 and PND 4 samples n = 3/group. This was a preliminary study and no statistical analysis was performed. Abbreviations: IPP, isopropylated phenyl phosphate; PND, postnatal day; TPHP, triphenyl phosphate.

Due to toxicity, concentrations were only reported for the 10 000 ppm IPP exposure group on GD 18. The IPP components TPHP, mono-IPP and tris-IPP were below the lower limit of quantitation (LLOQ) (approximately 8.1 ng IPP/g) for one fetus exposed to 1000 ppm IPP on GD 18. The distribution of TPHP, mono-IPP, bis-IPP, and tris-IPP isomers was similar across plasma, amniotic fluid, and fetus with either bis-IPP or tris-IPP having greatest percent composition in the majority of samples (Figure 7B). Exposure concentrations did not increase in a dose-dependent manner. On GD 18 in amniotic fluid from 1000 ppm and 10 000 ppm IPP exposed dams the component concentrations were 2.9 and 5.8 ng/ml TPHP, 6.44 and 18.3 ng/ml mono-IPP, 13.94 and 43.6 ng/ml bis-IPP, and 5.6 and 24.2 ng/ml tris-IPP, respectively. The PND 4 whole pup female concentrations of IPP components were 2–4 times higher than male concentrations in the 1000 ppm IPP group (Figure 7B). Male and female concentration on PND 4 were 82.7 and 201.8 ng/g TPHP, 163.3 and 463.7 ng/g mono-IPP, 244.5 and 555.8 ng/g bis-IPP, and 226.4 ng/g and 378.8 ng/g tris-IPP, respectively.

TPHP and IPP concentration in the brain

To examine plasma and brain concentrations, bio samples were collected from 1000 ppm and 10 000 ppm TPHP exposure groups, male and female offspring, on PND 28. In the 1000 ppm TPHP exposure group, only 2 male and 3 female plasma samples had concentrations above the LLOQ (approximately 8.0 ng/ml). The 14-fold increase in TPHP concentration in plasma was greater than the proportional exposure concentration increase. Similar to plasma, brain concentrations of TPHP were greater with a 16-fold increase in concentration from 1000 ppm to 10 000 ppm TPHP samples. Brain tissue concentrations of TPHP were approximately 2-fold higher than plasma sample concentrations (Figure 8A).

Figure 8.

Concentration of TPHP (A) and IPP (B) isomers in plasma and brain of PND 28 offspring. Due to toxicity, no samples were collected for the 10 000 ppm IPP group and only 1000 IPP isomer concentrations reported. In IPP exposed animals, plasma TPHP isomer concentration fell below detection. TPHP samples n = 5–6/group and IPP n = 8–10/group. This was a preliminary study and no statistical analysis was performed. Abbreviations: IPP, isopropylated phenyl phosphate; PND, postnatal day; TPHP, triphenyl phosphate.

Due to toxicity, the concentrations of IPP components (TPHP, mono-, bis-, and tris-IPP) were reported for only the 1000 ppm exposed offspring on PND28. In IPP exposed animals, TPHP concentrations were below the LLOQ (approximately 8.0 ng/ml) for all plasma samples in both male and female offspring as well as a few mono-IPP concentrations. No sex differences were observed in the mean IPP component concentrations of the remaining plasma samples, with bis- and tris-IPP concentrations being approximately 2-fold higher than mono-IPP. All components of IPP were detected in the brain of male and female offspring. IPP component concentrations in the brain of males were 7.26 ng TPHP/g, 21.19 ng mono-IPP/g, 23.21 ng bis-IPP/g, and 8.24 ng tris-IPP, while IPP component concentrations in the brain of females were 7.75 ng TPHP/g, 18.94 ng mono-IPP, 19.01 ng bis-IPP/g, and 7.08 ng tris-IPP (Figure 8B).

Discussion

Our study evaluated the effects of 2 representative OPFRs, TPHP and IPP in separate studies, following gestational and postnatal exposure in Sprague Dawley (Hsd:Sprague Dawley SD) rats via dosed feed. Significant maternal (F0) toxicity was noted for both chemicals and included maternal loss (or removal of an exposure group), decreases in body weight, changes in organ weights, and adverse clinical observations, noted primarily at ≥10 000 ppm for both TPHP and IPP. Developmental toxicity was observed in the F1 generation and included a decrease in the number of pups, live pups, and pup survival throughout lactation at 15 000 TPHP and ≥3000 ppm IPP. Furthermore, the F1 generation had dose-dependent decreases following developmental exposure to TPHP and IPP on body weight, organ weight, cholinesterase activity, and pubertal indices, although effects on body weight and blood cholinesterase activity were greater in IPP exposed animals. Developmental exposure to TPHP delayed puberty in both males (≥1000 ppm) and females (≥3000 ppm), whereas developmental exposure to IPP delayed puberty only in females exposed to 3000 ppm IPP. These delays were not due to the size of the animals as adjusting for body weight at weaning to account for growth retardation showed minimal to no impact on the data. Overall, the lower IPP exposure groups had more severe effects than low exposure to TPHP with exception of the onset of male puberty.

This study demonstrated TPHP and IPP were transferred to offspring during gestation and lactation and TPHP and IPP components can cross the blood-brain barrier as demonstrated on PND28. The mechanisms driving the accumulation of isomers remain unclear but in vitro studies of OPFRs are helping to further elucidate these pathways. Similar to the current work, high bioconcentrations of TPHP and reduced cholinesterase activity were observed in zebrafish larvae exposed to TPHP (Shi et al., 2018). Reproductive and developmental effects have previously been noted with these compounds in alternate animal models including daphnia, zebra fish, and Caenorhabditis elegans (Behl et al., 2015; 2016; Yuan et al., 2018).

Ubiquitous exposure to high levels of OPFR occurs in the general population (Araki et al., 2018; Hoffman et al., 2017; Ospina et al., 2018; Van den Eede et al., 2015). More importantly, developmental exposure has been suggested because OPFR metabolites have been detected in urine samples of pregnant women as well as in breast milk (Kim et al., 2014; Kuiper et al., 2020; Percy et al., 2020). Indications of continuous exposure in children have been reported with children having higher reported levels of OPFR relative to other age groups (Butt et al., 2014; 2016). Little is known about exposure risks during these critical developmental periods in this vulnerable population. The results of one study monitoring OPFR exposure in children and developmental landmarks suggest specific OPFRs may adversely impact cognitive development, including fine motor skills and early language abilities (Doherty et al., 2019). Despite the ubiquitous exposure of the general population, there is no acceptable daily intake level (ADI) set for most of the OPFRs.

The use of OPFRs is projected to be on the rise following the acceptance of a petition by the CPSC to ban furniture, children’s products, electronic enclosures, and mattresses containing organohalogen flame retardants (CPSC, 2017). In a recent review, TPHP and IPP were categorized as a “high toxicological concern” requiring additional research (Bajard et al., 2019). In response to concerns of OPFR exposure and potential toxicity, many regulatory bodies around the world (including United States and EU) have begun to gather data, inform consumers, limit flame retardants of concern, and/or change flammability standards to reduce the need for flame retardants (reviewed in Blum et al. [2019]). Following an assessment of flame retardants, including TPHP, the CPSC stated the report “yields one [exposure assessment] of the 2 factors needed to assess the potential for human health risk from these flame retardants…toxicity estimates are the second piece of the risk equation” (Patterson et al., 2016).

This current work was conducted to evaluate the short-term perinatal toxicity of TPHP and IPP as an initial step of a larger ongoing strategy. Due to effects observed in all exposure groups, a no observed adverse effect level (NOAEL) could not be determined in the current studies. Ongoing work at the NIEHS Division of Translational Toxicology (DTT) in support of both NIEHS and NTP includes using this data to conduct an in-depth assessment of developmental neurotoxicity and reproductive and developmental toxicity starting from GD 6 to PND 100 to better characterize the findings noted herein and expand to include additional endpoints. Ultimately, the goal of this exercise is to provide a case example of how a class of compounds may be evaluated by using a combination of in vitro screens and representative compounds as in vivo anchors to inform read-across thereby increasing efficiency, being prudent about time and cost of conducting individual in-depth in vivo studies on the entire class, and importantly being able to generate and provide data to agencies such as the CPSC in a timely manner for further evaluation of risk assessment.

Acknowledgments

The authors would like to thank Esra Mutlu and Rachel Dee for their review of the manuscript. Additionally, we would also like to thank Shawn Harris, Guanhua Xie, and Sandra McBride for additional help with statistical evaluation. The authors would like to acknowledge additional study design team members Barry McIntyre, Scott Auerbach, Mike Devito, Paul Foster, Dori Germolec, Greg Travlos, and Matthew Stout.

Funding

The Intramural Research Program of the National Institute of Health, National Institute of Environmental Health Sciences (ZIA ES103316-04); and performed for the Division of Translational Toxicology (formerly known as the Division of the National Toxicology Program), National Institutes of Health, U.S. Department of Health and Human Services under contracts HHSN273201600020C (Taconic Biosciences, provided animals), HHSN273201400020C, HHSN273201400001C and HHSN273201100001C (MRI Global; chemistry activities+), HHSN316201200054W (Arctic Slope Regional Corporation, data management), HHSN273201300004C (Instem, furnished data collection system), HHSN273201600011C (Social Scientific Systems; statistics), N01-ES-55536 and HHSN273201400015C (Battelle, Columbus, OH; animal study laboratory).

Declaration of conflicting interests

The authors declared no potential conflicts of interest with respect to the research, authorship, and/or publication of this article.