* Curriculum in Toxicology, University of North Carolina, Chapel Hill, North Carolina;
Neurotoxicology Division and
Experimental Toxicology Division, National Health and Environmental Effects Research Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina
Received July 24, 2001; accepted December 5, 2001
![]() |
ABSTRACT |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Key Words: brominated diphenyl ether; development; thyroid hormone; hepatic enzyme activity; endocrine disrupter; rat.
![]() |
INTRODUCTION |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Recent evidence from animal models suggests that exposure to some PBDEs results in disruption of thyroid hormone homeostasis (for reviews seeDarnerud et al., 2001;Hooper and McDonald, 2000
;McDonald, in press
). Studies in both rats and mice showed that exposure to 2,2`,4,4`-tetraBDE as low as 18 mg/kg/day for 14 days decreased circulating thyroxine (T4) concentrations (Darnerud and Sinjari, 1996
;Hallgren and Darnerud, 1998
). Short-term exposures to some commercial PBDE mixtures such as DE-71 and Bromkal 70 (consisting mainly of tetra- and penta-BDE) induced hypothyroxinemia in both rats and mice (Darnerud and Sinjari, 1996
;Fowles et al., 1994
;IPCS, 1994
;Zhou et al., 2001
). Furthermore, both commercial mixtures and BDE-47 have been demonstrated to induce both phase I (ethoxyresorufin-O-deethylase [EROD] and pentoxyresorufin-O-deethylase [PROD]), and phase II metabolic enzyme activity (uridinediphosphate-glucuronosyltransferase [UDPGT]) (Carlson, 1980a
,b
;Fowles et al., 1994
;Hallgren and Darnerud, 1998
;von Meyerinck et al., 1990
;Zhou et al., 2001
). T4 glucuronidation by phase II UDPGT enzymes in liver has been suggested as one of the mechanisms contributing to circulating T4 depletion by PBDEs and other polyhalogenated aromatic hydrocarbons (PHAHs) (Brouwer et al., 1998
;Hallgren and Darnerud, 1998
;Zhou et al., 2001
).
Normal thyroid hormone homeostasis is essential for development of many organs including the brain (Chan and Kilby, 2000;Dussault and Ruel, 1987
). Chemicals that disrupt thyroid hormone systems during pregnancy may have profound adverse impact on the normal development process of the brain (Brouwer et al., 1998
;Porterfield, 2000
;Porterfield and Hendrich, 1993
;Zoeller and Crofton, 2000
). In humans, developmental hypothyroidism leads to a characteristic deafness and mental retardation (Boyages and Halpern, 1993
). Hypothyroxinemia during fetal and/or postnatal periods, even when serum triiodothyronine (T3) concentration is normal, can lead to permanent functional abnormalities in children (e.g.,Haddow et al., 1999
;Larsen, 1982
;Morreale de Escobar et al., 2000
). Brouwer et al. (1998) proposed that thyroid hormone disruption induced by PHAHs might be at least partially responsible for neurochemical and behavioral changes observed in laboratory animal studies. Adverse neurobehavioral effects were found in neonatal mice exposed developmentally to single PBDE congeners (Eriksson et al., 1998
,1999
;Viberg et al., 2000
).
Considering that the rapid increase in PBDEs in human breast milk (Meironyte et al., 1999) suggests an increasing risk of developmental exposure, and previous data from animal exposures indicate an adverse effect of PBDEs on thyroid hormone homeostasis (Darnerud and Sinjari, 1996
;Fowles et al., 1994
;IPCS, 1994
;Zhou et al., 2001
), the purpose of the present study was to characterize the disruptive effects on thyroid hormones of developmental exposure to a commercial PBDE mixture (DE-71) in both dams and offspring. Induction of hepatic enzyme activity, EROD, PROD, and UDPGTs, were also examined to help characterize possible biochemical mechanisms for thyroid hormone disruption. EROD and PROD activity were monitored as biomarkers of aryl hydrocarbon (Ah)-type and phenobarbital (Pb)-type induction of UDPGT isoforms.
![]() |
METHODS AND MATERIALS |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
On GD 21, dams were checked for the number of pups delivered at 0800, 1000, 1200, 1500, and 1700 h, and pups were aged as PND 0 on the date of birth. All nonpregnant rats were euthanized. On PND 4, 7, 14, and 21, offspring were counted, sexed, and group-weighed by sex. Average pup weight by sex was calculated by dividing the group weight by the number of pups. In addition, body weights were recorded on PNDs 36 and 90, prior to tissue collection. Litters were culled to either 8 or 10 pups per litter with the number of pups kept similar, to the degree possible within 1 or 2 pups, throughout the preweaning period. Pups were checked daily for eye opening (pups with at least one eye open) from PNDs 11 through 18. Pups were weaned on PND 21, and housed by gender in groups of 2 or 3 per cage.
Chemicals and treatment.
DE-71 (penta-BDE, lot 7550OK20A) was generously supplied by the Great Lakes Chemical Corporation (West Lafayette, IN). DE-71 is a mixture that consists primarily of tetra and penta congeners (seeSjodin, 2000). The stock DE-71 solution (300 mg/ml) was prepared by mixing the compound with corn oil and sonicating it for 30 min at 40°C. The desired dosing solutions (1, 10, or 30 mg/ml) were obtained by serial dilution with corn oil.
Dams were assigned to treatment groups in a semirandom, weight-balanced fashion before being treated. Body weights of dams were recorded and dosing volumes adjusted on a daily basis. They were orally dosed, via gavage, with DE-71 (0, 1, 10, or 30 mg/kg/day) from GD 6 through PND 21, except for PND 0 (day of birth) when dams were left undisturbed. The dams (GD 20 and PND 22) and offspring (GD 20 and PND 4, 14, 36, and 90) were decapitated for collection of trunk blood. Liver samples were removed immediately and frozen in liquid nitrogen. Serum was obtained after clotting whole blood on ice for approximately 1.5 h, followed by centrifugation at 2500 rpm at 4°C for 20 min. Due to the limited amount of samples collected for GD 20, PND 4, and PND 14 pups, serum or liver samples within a litter were pooled. For pups at ages PND 36 and PND 90, 1 male and 1 female pup per litter were randomly sampled for body weight measurement and collection of serum and liver samples. All serum and liver samples for each age point were obtained from a minimum of 8 litters, and were stored at 80°C until analysis for thyroid hormone (T4 and T3) concentrations and hepatic enzyme (EROD, PROD, and UDPGT) activity.
Thyroid hormone assay.
Serum total concentrations of T4 and T3 were measured as previously described (Goldey and Crofton, 1998;Goldey et al., 1995
). Serum total T4 and T3 were measured in duplicate by using standard radioimmunoassay kits (Diagnostic Products Corp., Los Angeles, CA). Intra-assay and interassay coefficients of variance for the assays were below 10%. Since T4 concentrations for GD20 fetuses were below 10 ng/ml, a standard curve ranging from 2.5 to 120 ng/dl, and double volume of serum samples (50 µl) were used. The sensitivity for our T4 assay was 2.99 ng/ml, which resulted in 90.49% binding. Therefore, any result below this limit of quantitation (LOQ), i.e., above 90.49% specific binding, was recorded as 2.99 ng/ml. T3 was not assayed in serum from GD20 fetuses, due to the limited sample volumes available.
Hepatic enzyme activity assay.
Liver microsomal fractions were prepared as described previously (DeVito et al., 1993). Microsomal protein concentrations were determined using the Bio-Rad protein assay kit (Bio-Rad, Richmond, CA) with bovine serum albumin as the standard. Hepatic microsomal EROD activity (a marker for CYP1A1 activity) and PROD activity (a marker of CYP2B activity) were assayed using the method of DeVito et al. (1993). All substrate concentrations were 1.5 nM. Both EROD and PROD values were calculated as pmol resorufin per mg protein per min, or per 30 min for GD 20 fetuses (all data corrected to per min rate). PROD and UDPGT activity were not measured for samples obtained from GD 20 fetuses, due to limited available sample.
Hepatic microsomal T4-UDPGT activity was assayed as described in Zhou et al. (2001). Briefly, 100 µl microsomes (2 mg protein per ml Tris/HCl buffer) were incubated at 37°C with purified, radiolabeled T4, 6-n-propyl-2-thiouracil (PTU), and UDPGA (or no UDPGA for blank) over a 30-min period. The reaction was stopped by addition of ice-cold methanol followed by centrifugation and mixing the supernatant with HCl. The formed glucuronyl T4 (T4-G), separated by chromatography on lipophilic sephadex LH-20 columns, was counted on the gamma-counter. The UDPGT activity was calculated as pmol T4-G per mg protein per min.
Data analysis.
All statistical analyses were performed on SAS© 6.12 (SAS Institute, Inc., Cary, NC). The litter was the statistical unit for all analyses. Analysis of variance (ANOVA) was used to analyze for effects of treatment and interactions. If there was more than one independent variable, significant interactions were followed by step-down ANOVA tests for each independent variable (e.g., treatment and age). When more than one reading for each litter was obtained (i.e., body weights for male and female samples from the same litter) then a nested design was used, that is to say litter was nested under treatment. Repeated-measures ANOVAs were applied to data on dam body weights, offspring body weights (for PND 4, 7,14, and 21) and eye opening. Gestation and lactation body weight data for dams were analyzed with separate repeated-measures (day) ANOVAs because of the lack of lactation data for animals killed on GD 20. Body weights of dams were inadvertently not recorded on PND 1, and thus no data were used from this age in data analyses. Postweaning offspring body weights (i.e., from PNDs 36 and 90) were analyzed with a two-way ANOVA, with time and dose as independent variables, and litter nested under treatment. For eye opening data, only data from days 14 to 18 were analyzed due to an absence of eye opening in all groups prior to day 14. For significant effects of treatment, Duncan's Multiple Range test was used for mean contrast comparisons. The fetal T4 data were analyzed with the Kruskal-Wallis test followed by a Dunn Multiple Comparison test (due to the lack of homogeneity of variance, see also Results section). A significance level of 0.05 was used for all statistical tests.
Benchmark dose (BMD) estimates were determined for alterations in thyroid hormones and hepatic enzyme activity using the U.S. EPA Benchmark Dose Software (BMDS, V 1.3). For each endpoint a BMD was estimated using the data from the age or time point that demonstrated the greatest potency and efficacy (see all figures and Table 1). The EROD and PROD data, as well as the T4 and UDPGT data for neonates data were fit with the Hill model, as this function best describes the biological response. The T4 data and UDPGT data from the dams were fit using the power model and a second-order polynomial, respectively, due to a lack of significant fit for the Hill model. The benchmark effect levels were set at 20% decreases for the thyroid hormone data and 50% increases for the liver enzyme data (Zhou et al., 2001
). The BMDLs (lower-bound confidence limit) were calculated as the 95% lower confidence interval for the BMD.
|
![]() |
RESULTS |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Body and Organ Weights
No evidence of treatment-related effects were found for maternal (Fig. 1A) or offspring body weights (Fig. 1B
). For body weights of dams, repeated-measures analysis showed neither a treatment-by-age interaction during gestation (F(42,451) = 0.83,p < 0.7549) or lactation (F(57,287) = 0.98,p < 0.5186). Nor was there any main effect of treatment (gestation: F(3,165) = 0.22,p < 0.8853; lactation: F(3,114) = 0.27,p < 0.8438). Consistent with the overall changes in body weight due to pregnancy, there were main effects of age during both gestation (F(14,152) = 402.22,p < 0.0001) and lactation (F(19,96) = 37.16,p < 0.0001;Fig. 1
). For preweaning offspring body weights, there were no significant interactions of treatment with any other variables (all F values < 1.44,p values > 0.2445), nor was there a main effect of treatment (F(3,41) = 0.94,p < 0.4296). Consistent with postnatal growth, there was a main effect of age (F(3,39) = 618.37,p < 0.0001). There were no significant interactions of treatment with any other variables (all F values < 0.42,p values > 0.6963), nor was there a main effect of treatment (F(3,70) = 0.25,p < 0.8620). There was a significant interaction of gender and age (F(1,70) = 313.63,p < 0.0001) that was consistent with age-related increases in body weight (Fig. 1
), and more so in males than in females (body weight data not shown by gender).
|
|
|
Thyroid Hormones
Perinatal maternal exposure to DE-71 caused a decrease in serum total T4 in dams, fetuses, and offspring (Fig. 4). The effects in the dam were present during both gestation and lactation. On GD 20 there was a significant decrease only in the high dose (48% relative to controls). On PND 22 there again was only a significant decrease in the high-dose group (44% relative to controls). These inferences were supported by a significant treatment-by-age interaction (F(3,83) = 4.30,p < 0.0071) that resulted from a slightly larger dose effect on PND 22 and a high overall serum concentration of T4 on PND 22 (Fig. 4A
). For dams, there were main effects of treatment at GD 20 (F(3,47) = 4.23,p < 0.0099) and PND 22 (F(3,36) = 27.37,p < 0.0001). Results of mean contrast comparisons at each age sampled are illustrated inFigure 4A
.
|
There were no treatment-related effects of developmental DE-71 exposure on serum total T3 concentrations in either the dams or the offspring (data not shown). There were no significant main effects of treatment, nor any interactions of treatment and age for either dams or pups (allp > 0.5). There was a significant effect of age in dams (F(1,83) = 7.35,p < 0.0081) that reflects a slightly lower (10%) concentration of T3 in the PND 22 dams (88.75 ± 5.3 ng/dl) compared to GD 20 dams (98.72 ± 3.7 ng/dl). There was also a significant effect of age in offspring (F(3,159) = 214.57,p < 0.0001) that reflects an age-related increase in serum total T3 (in controls: 23.89 ± 3.33 ng/dl at PND 4; 84.27 ± 6.38 ng/dl at PND 14; 119.08 ± 5.06 ng/dl at PND 36; and 109.79 ± 5.22 ng/dl at PND 90).
Hepatic Enzyme Activity
Maternal exposure to DE-71 resulted in significant increases in hepatic EROD, PROD, and UDPGT activity in both dams and offspring (Figs. 5, 6, and 7).
|
|
|
Hepatic EROD activity in fetuses and offspring was increased as a result of maternal exposure to DE-71 (Fig. 5B). Fetal EROD activity was significantly increased 2.5-fold in the high-dose group on GD 20. Offspring EROD activity was significantly increased, 39-fold and 95-fold, on PND 4 and 20-fold and 57-fold on PND 14 in the 10 and 30 mg/kg/day groups, respectively. There was a much smaller, yet significant, increase of 0.5-fold in the high-dose group on PND 36. There were no treatment-related changes in EROD activity on PND 90. These effects were confirmed by a significant treatment-by-age interaction (F(12,219) = 100.99,p < 0.0001) and significant main effects of treatment on GD 20 (F(3,26) = 7.32,p < 0.0010), PND 4 (F(3,49) = 219.61,p < 0.0001), PND 14 (F(3,48) = 137.40,p < 0.0001), and PND 36 (F(3,40) = 3.96,p < 0.0146). There was also a main effect of age (F(4,219) = 227.15,p < 0.0001) that resulted from the large treatment-related increases at the younger ages, as well as increases in basal activity in control samples as a function of age (Fig. 5B
). Results of mean contrast comparisons at each age sampled are illustrated inFigure 5B
.
Hepatic PROD activity was increased slightly more in dams on PND 22 compared to GD 20 (Fig. 6A). There were dose-dependent increases in PROD activity of 9-fold and 19-fold on GD 20, and 9-fold and 24-fold on PND 22 in the 10 and 30 mg/kg/day groups, respectively. These conclusions were confirmed statistically with a significant treatment-by-age interaction (F(3,83) = 5.87,p < 0.0011), and significant main effects of treatment on GD 20 (F(3,45) = 24.52,p < 0.0001) and PND 22 (F(3,36) = 56.49,p < 0.0001). There was no effect of the 1 mg/kg/day dose on PROD activity.
Hepatic PROD activity in offspring was increased as a result of maternal exposure to DE-71 (Fig. 6B). Offspring PROD activity was significantly increased by 21- and 26-fold on PND 4 and 19- and 21-fold on PND14, in the 10 and 30 mg/kg/day groups, respectively. There was a significant increase of 10-fold in the high-dose group on PND36. There were no treatment-related changes in PROD activity on PND 90. These effects were confirmed by a significant treatment-by-age interaction (F(9,193) = 40.64,p < 0.0001) and significant main effects of treatment on PND 4 (F(3,49) = 125.90,p < 0.0001), PND 14 (F(3,48) = 81.59,p < 0.0001), and PND 36 (F(3,40) = 11.89,p < 0.0001). There was also a main effect of age (F(3,193) = 158.79,p < 0.0001) that resulted from the large treatment-related increases at the younger ages. In contrast to EROD activity, PROD activity did not vary significantly as a function of age in control samples (Fig. 6B
). Results of mean contrast comparisons at each age sampled are illustrated inFigure 6B
.
The effects of perinatal maternal exposure to DE-71 on hepatic UDPGT activity, measured as T4 glucuronidation, is illustrated inFigure 7. Exposure to DE-71 caused increases in the rate of glucuronidation of T4 in both dams and offspring. In dams, there was a similar increase of about 1.6-fold in UDPGT activity, only in the high-dose groups, on both GD 20 and PND 22 (Fig. 7A
). There were no effects on UDPGT activity detected in the two lower doses. These conclusions were confirmed statistically by a nonsignificant treatment-by-age interaction (F(3,55) = 1.04,p < 0.3839), and a significant main effect of treatment (F(3,55) = 13.52,p < 0.0001). There was no main effect of age (F(1,55) = 0.05,p < 0.8190) reflecting a similar basal activity level on both days.
Hepatic UDPGT activity in offspring was increased as a result of maternal exposure to DE-71 (Fig. 7B). Offspring UDPGT activity was significantly increased 1.9-fold and 4.7-fold on PNDs 4 and 14, respectively, in the 30 mg/kg/day group. There was no significant effect of any lower doses on PNDs 4 or 14. There were no effects of exposure on PND 36 or PND 90. These effects were confirmed by a significant treatment-by-age interaction (F(9,146) = 5.89,p < 0.0001) and significant main effects of treatment on PND 4 (F(3,44) = 10.65,p < 0.0001) and PND14 (F(3,50) = 14.17,p < 0.0001). There was also a main effect of age (F(3,193) = 158.79,p < 0.0001) that resulted from the large treatment-related increases at the younger ages. There was a small decrease in basal UDPGT activity in control offspring on PND 14 (0.39 ± 03 pmol T4-G/mg protein/min) compared to PND 4 (0.75 ± 0.06), PND 36 (0.59 ± 0.07) and PND 90(0.89 ± 0.12). Results of mean contrast comparisons at each age sampled are illustrated inFigure 7B
.
NOELs and model estimates for BMDs and BMDLs are shown inTable 2. The relationship for NOELs and BMDs vary by endpoint. These variations are likely due to the effects of data variability and dose spacing on the NOEL estimates. Based on visual inspection of the data, BMD estimates appeared to be better approximations of potency. BMD estimates for all endpoints were lower for neonates compared to dams. BMDs for EROD and PROD were up to an order of magnitude lower in neonates when compared to dams, whereas there was only a 24-fold difference for T4 and UDPGT (Table 2
).
|
![]() |
DISCUSSION |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
The current study demonstrated a perinatal hypothyroxinemia following DE-71 developmental exposure. DE-71 caused dose- and time-dependent reductions in serum total T4 concentrations in fetal and postnatal rats (GD 20, PND 4, and PND 14), with a maximal reduction of 66% occurring at the highest doses on PND 14. There was a complete recovery in hypothyroxinemia in rats on PND 36, 15 days after cessation of lactation exposure. It is important to note that the large number of samples below the LOQ suggest that the magnitude of reduction in T4 during the fetal period is uncertain and may be underestimated. The effect of DE-71 on T4 was less pronounced in dams than in offspring. This is reflected in lower NOEL (10-fold) and BMD (2-fold) values in the offspring compared to dams and indicates that the offspring are more sensitive to the effects of DE-71 than pregnant dams. Serum total T3 was not affected by DE-71 in either dams or offspring at any time point sampled.
The effects of DE-71 reported here, decreases in T4 with no significant changes in T3, are consistent with previous reports on the effects of PBDE exposures, and extends those findings to include effects in the dam, fetus, and developing offspring. Norris (1975) first reported the thyrotoxic effects of PBDE compounds in a 30-day exposure in adult rats to octa- and deca-BDE mixtures, which resulted in thyroid hyperplasia. Fowles et al. (1994) first showed that T4 was decreased in mice exposed for 14 days to 18, 36, or 72 mg/kg/day of DE-71. Darnerud and Sinjari (1996) demonstrated decreased total plasma T4 in both rats and mice exposed for 14 days to 18 or 36 mg/kg/day of Bromkal 70. These same authors also exposed mice to 18 mg/kg/day of BDE-47 and found a 31% decrease in total plasma T4. Hallgren and Darnerud (1998) found decreases in both total and free plasma T4 with no increase in TSH following a 14-day exposure of female rats to 18 mg/kg/day BDE-47. In a 90-day study, T4 concentrations were decreased, but T3 concentrations were not altered in rats administered doses as high as 100 mg/kg/day DE-71 in the diet (IPCS, 1994). This is consistent with a previous 4-day exposure study in weanling rats (Zhou et al., 2001
), where there was no significant reduction in T3 at doses up to 30 mg/kg/day. Rosiak et al. (1997) reported that maternal exposure to individual chlorinated diphenyl ether congeners (2,2`,4,4`,5,5`-hexachlorodiphenyl ether; 2,2`,4,5,6`-pentachlorodiphenyl ether) depressed T4 in dams during gestation and in preweaning offspring. These chlorinated diphenyl ether congeners did not affect serum T3 or TSH concentrations in maternal or juvenile rats. The mechanism(s) responsible for the lack of effects of DE-71 on serum T3 are currently unknown. Possible mechanisms include increased tissue-specific conversion of T4 to T3 due to increased deiodinase activity (Raasmaja et al., 1996
). Alternatively, there may be an increased hepatic catabolism and clearance of T4, and not T3. Recent work by Hood and Klaassen (2000) has demonstrated that Aroclor 1254 (A1254), a polychlorinated biphenyl mixture, decreases serum T4 and induces glucuronidation of T4, but does not alter serum T3 concentrations or T3 glucuronidation (see also discussion below). In general, the data presented herein clearly demonstrate that PBDEs adversely impact circulating concentrations of T4.
The data in this paper do not show a clear relationship between serum T4 depletion and induction of the T4-UDPGT activity. Significant reduction in T4 was observed in doses as low as 10 mg/kg/day for dams at GD 20 and PND 22, and for offspring at PNDs 4 and 14. However, only the 30 mg/kg/day treatment groups showed significant induction of T4-UDPGT activity. One explanation for the inconsistency between T4 concentrations and UDPGT activity is that the T4-UDPGT assay may be a better biomarker for the isoforms of UDPGT that are induced by Ah-receptor agonists compared to phenobarbital-like agonists (Craft et al., 2001). A second possible explanation for this lack of correlation is that PBDE congeners and metabolites, as well as structurally similar PCBs, are known to displace T4 from transthyrethin (TTR), the major protein that transports thyroid hormones in rats and mice (Chauhan et al., 2000;Cheek et al., 1999
;Meerts et al., 2000
). While the exact role this mechanism plays in the regulation of serum concentrations of T4 is unknown, displacement of serum T4 could lead to increased glucuronidation and a consequent lowered serum concentration of T4. PBDEs may also have direct effects on the thyroid gland (Brouwer et al., 1998
). However, previous studies (Darnerud and Sinjari, 1996
;Zhou et al., 2001
) found no evidence of increased thyroid stimulating hormone (TSH). Previous work with other PHAHs, such as A1254 and some chlorinated diphenyl ethers, have also failed to find any upregulation of TSH during development (Goldey et al. 1995
;Morse et al., 1996
;Rosiak et al., 1997
). This indicates a lack of activation of the hypothalamic-pituitary-thyroid feedback process normally found with direct acting thyrotoxicants (Capen, 1997
;DeVito et al., 1999
). Interestingly, long-term exposure studies to the deca-BDE have found small increases in the rate of thyroid hyperplasia and neoplasia (Norris et al., 1975
;NTP, 1986
). However, no long-term cancer bioassays have been conducted on the tetra-, penta-, or octa-BDEs. A combination of the above mentioned mechanisms might be ultimately responsible for the difference between measured increases in UDPGT activity and decreases in serum T4 concentrations.
Developmental exposure to DE-71 resulted in increased hepatic metabolic activity in dams, fetuses, and offspring. DE-71 exposure resulted in increased EROD, PROD, and UDPGT activity in dams during both gestation and lactation, with the amount of induction fairly similar at both time points. There was a slightly larger increase in EROD induction on GD 20 (3.7-fold) compared to PND 22 (2.9-fold) and a slightly higher PROD induction on PND 22 (24-fold) compared to GD 20 (19-fold). There was no statistically significant effect of age for UDPGT induction (1.6-fold). It is unlikely that the small difference in EROD activity prenatally versus postnatally is biologically significant. These data suggest that the level of induction of hepatic metabolizing enzymes, as measured by EROD, PROD, and UDPGT activity, is relatively similar at the end of pregnancy and the end of lactation. This is the first report of increased hepatic Phase-1 and Phase-2 activity by PBDEs in pregnant animals. Previous reports have found increased liver weights in pregnant rabbits exposed to Saytex 111, a commercial mixture consisting mostly of hepta- and octa-BDEs (Breslin et al., 1989). Norris et al. (1975) reported no effect of a deca-BDE mixture on liver weights in dams from a rat teratology study.
Developmental exposure to DE-71 resulted in increased hepatic microsomal enzyme activity at both fetal and early postnatal time periods. EROD and PROD activity were increased on PND 4, PND 14 and PND 36, with activity returning to control levels by PND 90. UDPGT activity was increased on PND 4 and PND 14, and recovered to control levels by PND 36. Consistent with increased hepatic metabolic activity were increased liver weights in the offspring. Important to note was that EROD was also increased on GD 20. This supports a conclusion of significant fetal exposure to DE-71 and/or metabolites. Liver weights in fetuses were not statistically different. These effects are consistent with a number of previous reports on the effects of PBDE exposure in weanling or adult rats and mice. Carlson (1980a; 1980b) showed that both 14- and 90-day exposures to penta- and octa-BDE mixtures increased hepatic benzo-[a]-pyrene and p-nitroanisole metabolism. von Meyerinck et al. (1990) found increased EROD and benzphetamine N-demethylation activity in hepatic tissue from mice treated for 14 days with Bromkal 70. Increased EROD and PROD activity were found in mice exposed to DE-71 for 14 days, but only increased PROD was found following acute exposure (Fowles et al., 1994). More recently, Hallgren and Darnerud (1998) reported increased EROD (3-fold) and PROD (14-fold) activity in rats after 14 days of exposure to BDE-47.
The increase in hepatic metabolic capacity has a number of important implications for the toxicity of PBDEs. First, PBDE mixtures have been suggested to be either solely phenobarbital-type inducers, such as DE-71 and DE-79 (Carlson, 1980a), or mixed-type inducers (i.e., phenobarbital and dioxin-like substances) of xenobiotic metabolism such as Bromkal 70 (von Meyerinck et al., 1990
) or DE-71 (Zhou et al., 2001
). The present data support the conclusion that DE-71 is a mixed-type inducer in both pregnant rats and offspring during the early postnatal period. Second, the effects of DE-71 on fetal EROD activity following maternal exposure, together with decreased fetal T4 concentrations, suggest placental transfer and fetal exposure to DE-71 congeners and/or metabolites. This is consistent with data from maternal exposure to other PHAHs such as dioxins and PCBs. Third, a comparison of maternal/fetal and maternal/offspring ratios of EROD activity suggests a much greater postnatal exposure to DE-71 components or their metabolites. EROD activity has been suggested as a biomarker of exposure to Ah-active compounds (Lagueux et al., 1999
;Sewall et al., 1995
;Whyte et al., 2000
). EROD activity was induced to a fairly similar degree in dams and fetuses on GD 20 (3.7-fold for dams, and 2.5-fold for fetuses). In contrast, induction of EROD activity was much greater in the postnatal offspring (95-fold) compared to the postnatal dam (2.9-fold). Thus, it is interesting to speculate whether there was a much greater magnitude of exposure to the postnatal offspring compared to both the fetus and the dam. Greater postnatal exposure to offspring via lactation has been demonstrated previously where compounds like TCDD and PCBs are transferred placentally to the fetus in limited quantities compared to the amount delivered via lactation (Crofton et al., 2000
;Masuda et al., 1978
;Takagi et al., 1986
;Vodicnik and Lech, 1980
). Confirmation of this hypothesis will require developmental toxicokinetic studies of PBDEs.
Lastly, there were neither significant adverse effects on dam or offspring body weight, nor effects on postnatal survival, sex ratio at birth, or litter size. These data indicate that DE-71, at the dosages examined, did not produce overt toxicity in either dams or offspring. These findings were consistent with other studies on commercial penta-BDE (IPCS, 1994). Commercial penta-BDE, as high as 100 mg/kg in the diet during gestation and lactation, had no effects on the number of pregnancies or on survival and weight of the neonates. For pregnant rats given commercial penta-BDE from GD 6 through 15, inhibition of maternal body weight gain only occurred above doses of 100 mg/kg.
In conclusion, developmental exposure to DE-71 reduced circulating T4 concentrations and induced hepatic EROD, PROD, and UDPGT activity in both dams and offspring. The T4-depleting effects of DE-71 are likely to involve multiple mechanisms of action. These data demonstrate that DE-71 is an endocrine disrupter in rats during development.
![]() |
ACKNOWLEDGMENTS |
---|
![]() |
NOTES |
---|
1 To whom correspondence should be addressed at the Neurotoxicology Division, MD-74B, National Health and Environmental Effects Laboratory, U.S. EPA, Research Triangle Park, NC 27711. Fax: (919) 541-4849. E-mail: crofton.kevin{at}epa.gov.
![]() |
REFERENCES |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Breslin, W. J., Kirk, H. D., and Zimmer, M. A. (1989). Teratogenic evaluation of a polybromodiphenyl oxide mixture in New Zealand white rabbits following oral exposure.Fundam. Appl. Toxicol.12,151157.[ISI][Medline]
Brouwer, A., Morse, D. C., Lans, M. C., Schuur, A. G., Murk, A. J., Klasson-Wehler, E., Bergman, A., and Visser, T. J. (1998). Interactions of persistent environmental organohalogens with the thyroid hormone system: Mechanisms and possible consequences for animal and human health.Toxicol. Ind.. Health14,5984.[ISI][Medline]
Capen, C. C. (1997). Mechanistic data and risk assessment of selected toxic end points of the thyroid gland.Toxicol. Pathol.25,3948.[ISI][Medline]
Carlson, G. P. (1980a). Induction of xenobiotic metabolism in rats by short-term administration of brominated diphenyl ethers.Toxicol. Lett.5,1925.[ISI][Medline]
Carlson, G. P. (1980b).Induction of xenobiotic metabolism in rats by brominated diphenyl ethers administered for 90 days.Toxicol. Lett.6,207212.[ISI][Medline]
Chan, S., and Kilby, M. D. (2000). Thyroid hormone and central nervous system development.J. Endocrinol.165,18.
Chauhan, K. R., Kodavanti, P. R., and McKinney, J. D. (2000). Assessing the role of ortho-substitution on polychlorinated biphenyl binding to transthyretin, a thyroxine transport protein.Toxicol. Appl. Pharmacol.162,1021.[ISI][Medline]
Cheek, A. O., Kow, K., Chen, J., and McLachlan, J. A. (1999). Potential mechanisms of thyroid disruption in humans: Interaction of organochlorine compounds with thyroid receptor, transthyretin, and thyroid-binding globulin.Environ. Health Perspect.107,273278.[ISI][Medline]
Craft, E. S., Ross, D. G., DeVito, M. J., and Crofton, K. M. (2000). Comparative responsiveness of Long-Evans rats versus C57BL/6J mice given TCCD-like and phenobarbital-like PCB (polychlorinated biphenyl) congeners.Toxicologist54352 (abstract).
Crofton, K. M., Kodavanti, P. R. S., Derr-Yellin, E. C., Casey, A. C., and Kehn, L. S. (2000). PCBs, thyroid hormones, and ototoxicity in rats: Cross-fostering experiments demonstrate the influence of postnatal lactation exposure.Toxicol. Sci.57,131140.
Darnerud, P. O., Eriksen, G. S., Johannesson, T., Larsen, P. B., and Viluksela, M. (2001). Polybrominated diphenyl ethers: Occurrence, dietary exposure, and toxicology.Environ. Health Perspect.109,(Suppl.1)4968.[ISI][Medline]
Darnerud, P. O., and Sinjari, T. (1996). Effects of polybrominated diphenyl ethers (PBDEs) and polybrominated biphenyls (PCBs) on thyroxine and TSH blood levels in rats and mice.Organohalogen Compounds29,316319.
DeVito, M., Biegel, L., Brouwer, A., Brown, S., Brucker-Davis, F., Cheek, A. O., Christensen, R., Colborn, T., Cooke, P., Crissman, J., Crofton, K., Doerge, D., Gray, E., Hauser, P., Hurley, P., Kohn, M., Lazar, J., McMaster, S., McClain, M., McConnell, E., Meier, C., Miller, R., Tietge, J. and Tyl, R. (1999). Screening methods for thyroid hormone disrupters.Environ. Health Perspect.107,407415.[ISI][Medline]
DeVito, M. J., Maier, W. E., Diliberto, J. J., and Birnbaum, L. S. (1993).Comparative ability of various PCBs, PCDFs, and TCDD to induce cytochrome P450 1A1 and 1A2 activity following 4 weeks of treatment.Fundam. Appl. Toxicol.20,125130.[ISI][Medline]
Dussault, J. H., and Ruel, J. (1987). Thyroid hormones and brain development.Annu. Rev. Physiol.49,321334.[ISI][Medline]
Eriksson, P., Jakobsson, E., and Fredriksson, A. (1998). Developmental neurotoxicity of brominated flame-retardants, polybrominated diphenyl ethers, and tetrabromo-bis-phenol A.Organohalogen Compounds35,375377.
Eriksson, P., Viberg, H., Jakobsson, E., Orn, U., and Fredriksson, A. (1999).PBDE, 2,2`,4,4`,5-pentabromodiphenyl ether causes permanent neurotoxic effects during a defined period of neonatal brain development.Organohalogen Compounds40,333336.
Fowles, J. R., Fairbrother, A., Baecher-Steppan, L., and Kerkvliet, N. I. (1994).Immunologic and endocrine effects of the flame-retardant pentabromodiphenyl ether (DE-71) in C57BL/6J mice.Toxicology86,4961.[ISI][Medline]
Goldey, E. S., and Crofton, K. M. (1998).Thyroxine replacement attenuates hypothyroxinemia, hearing loss, and motor deficits following developmental exposure to Aroclor 1254 in rats.Toxicol. Sci.45,94105.[Abstract]
Goldey, E. S., Kehn, L. S., Lau, C., Rehnberg, G. L., and Crofton, K. M. (1995). Developmental exposure to polychlorinated biphenyls (Aroclor 1254) reduces circulating thyroid hormone concentrations and causes hearing deficits in rats.Toxicol. Appl. Pharmacol.135,7788.[ISI][Medline]
Haddow, J. E., Palomaki, G. E., Allan, W. C., Williams, J. R., Knight, G. J., Gagnon, J., O'Heir, C. E., Mitchell, M. L., Hermos, R. J., Waisbren, S. E., Faix, J. D., and Klein, R. Z. (1999). Maternal thyroid deficiency during pregnancy and subsequent neuropsychological development of the child.N. Engl. J. Med.341,549555.
Hallgren, S., and Darnerud, P. O. (1998). Effects of polybrominated diphenyl ethers (PBDEs), polychlorinated biphenyls (PCBs), and chlorinated paraffins (CPs) on thyroid hormone levels and enzyme activities in rats.Organohalogen Compounds35,391394.
Hood, A., and Klaassen, C. D. (2000). Differential effects of microsomal enzyme inducers onin vitro thyroxine (T4) and triiodothyronine (T3) glucuronidation.Toxicol. Sci.55,7884.
Hooper, K., and McDonald, T. A. (2000). The PBDEs: An emerging environmental challenge and another reason for breast milk-monitoring programs.Environ. Health Perspect.108,387392.[ISI][Medline]
IPCS (1994). Environmental Health Criteria 162: Brominated diphenyl ethers. International Programme on Chemical Safety, World Health Organization, Geneva.
Kierkegaard, A., Sellstrom, U., Bignert, A., Olsson, M., Asplund, L., Jansson, B., and de Wit, C. (1999). Temporal trends of a polybrominated diphenyl ether (PBDE), a methoxylated PBDE, and hexabromocyclododecane (HBCD) in Swedish biota.Organohalogen Compounds40,367370.
Lagueux, J., Pereg, D., Ayotte, P., Dewailly, E., and Poirier, G. G. (1999).Cytochrome P450 CYP1A1 enzyme activity and DNA adducts in placenta of women environmentally exposed to organochlorines.Environ. Res.80,369382.[ISI][Medline]
Larsen, P. R. (1982). Thyroid-pituitary interaction: Feedback regulation of thyrotrophin secretion by thyroid hormones.N. Eng. J. Med.306,2332.[ISI][Medline]
Lindstrom, G., Wingfors, H., Dam, M., and van Bavel, B. (1999). Identification of 19 polybrominated diphenyl ethers (PBDEs) in long-finned pilot whales (Globicephala melas) from the Atlantic.Arch. Environ. Contam. Toxicol.36,355363.[ISI][Medline]
Masuda, Y., Kagawa, R., Tokudome, S., and Kuratsune, M. (1978). Transfer of polychlorinated biphenyls to the foetuses and offspring of mice.Food Cosmet. Toxicol.16,3337.[ISI][Medline]
McDonald, T. A. (in press). A perspective on the potential health risks of PBDEs.Chemosphere.
Meerts, I. A., van Zanden, J. J., Luijks, E. A., van Leeuwen-Bol, I., Marsh, G., Jakobsson, E., Bergman, A., and Brouwer, A. (2000). Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretinin vitro.Toxicol. Sci.56,95104.
Meironyte, D., Noren, K., and Bergman, A. (1999).Analysis of polybrominated diphenyl ethers in Swedish human milk: A time-related trend study, 19721997.J. Toxicol. Environ. Health58,329341.[ISI]
Morreale de Escobar, G., Obregon, M. J., and Escobar del Rey, F. (2000). Clinical perspective: Is neuropsychological development related to maternal hypothyroidism or to maternal hypothyroxinemia?J. Clin. Endocrinol. Metab.85,39753987.
Morse, D. C., Wehler, E. K., Wesseling, W., Koeman, J. H., and Brouwer, A. (1996). Alterations in rat brain thyroid hormone status following pre- and postnatal exposure to polychlorinated biphenyls (Aroclor 1254).Toxicol. Appl. Pharmacol.136,269279.[ISI][Medline]
NTP (1986). Toxicology and Carcinogenesis Studies of Decabromodiphenyl Oxide (CAS No. 1163195) in F344/N Rats and B6C3F1 Mice (feed studies). NTP TR 309. NIH Publication No 862565. National Toxicology Program, U.S. Department of Health and Human Services, National Institutes of Health.
Norris, J. M., Kociba, R. J., Schwetz, B. A., Rose, J. Q., Humiston, C. G., Jewett, G. L., Gehring, P. J., and Mailhes, J. B. (1975). Toxicology of octabromobiphenyl and decabromodiphenyl oxide.Environ. Health Perspect.11,153162.[Medline]
Porterfield, S. P. (2000). Thyroidal dysfunction and environmental chemicalspotential impact on brain development.Environ. Health Perspect.108(Suppl. 3),433438.[ISI][Medline]
Porterfield, S. P., and Hendrich, C. E. (1993). The role of thyroid hormones in prenatal and neonatal neurological developmentcurrent perspectives.Endocr. Rev.14,94106.[ISI][Medline]
Raasmaja, A., Viluksela, M., and Rozman, K. K. (1996). Decreased liver type I 5`-deiodinase and increased brown adipose tissue type II 5`-deiodinase activity in 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-treated Long Evans rats.Toxicology114,199205.[ISI][Medline]
Rosiak, K. L., Seo, B. W., Chu, I., and Francis, B. M. (1997). Effects of maternal exposure to chlorinated diphenyl ethers on thyroid hormone concentrations in maternal and juvenile rats.J. Environ. Sci. Health B32,377393.[ISI][Medline]
Sellstrom, U., Jansson, B., Kierkegaard, A., and deWit, C. (1993). Polybrominated diphenyl ethers (PBDE) in biological samples from the Swedish environment.Chemosphere26,17031718.[ISI]
Sewall, C. H., Bell, D. A., Clark, G. C., Tritscher, A. M., Tully, D. B., Vanden Heuvel, J., and Lucier, G. W. (1995). Induced gene transcription: Implications for biomarkers.Clin. Chem.41,18291834.
She, J., Petreas, M., Winkler, J., Visita, P., McKinney, M., and Kopec, D. (in press). PBDEs in the San Francisco Bay area: Measurements in harbor seal blubber and human breast adipose tissue.Chemosphere.
Sjodin, A. (2000). Occupational and dietary exposure to organohalogen substances, with special emphasis on polybrominated diphenyl ethers. Ph.D. dissertation, Stockholm University, Stockholm, Sweden.
Stern, G. A., and Ikonomou, M. G. (2000). Temporal trends of polybrominated diphenyl ethers in SE Baffin beluga: Increasing evidence of long-range atmospheric transport.Organohalogen Compounds47,8184.
Strandman, T., Koistinen, J., Kiviranta, H., Vuorinen, P. J., Tuomisto, J., Tuomisto, J., and Vartiainen, T. (1999). Levels of some polybrominated diphenyl ethers (PBDEs) in fish and human adipose tissue in Finland.Organohalogen Compounds40,355358.
Takagi, Y., Aburada, S., Hashimoto, K., and Kitaura, T. (1986). Transfer and distribution of accumulated (14C)polychlorinated biphenyls from maternal to fetal and suckling rats.Arch. Environ. Contam. Toxicol.15,709715.[ISI][Medline]
Viberg, H., Fredriksson, A., and Jacobsson, E. (2000). Developmental neurotoxic effects of 2,2`,4,4`,5-pentabromodiphenyl ether (PBDE 99) in the neonatal mouse.Toxicologist54,1360 (abstract).
Vodicnik, M. J., and Lech, J. J. (1980).The transfer of 2,4,5,2`,4`,5`-hexachlorobiphenyl to fetuses and nursing offspring: I. Disposition in pregnant and lactating mice and accumulation in young.Toxicol. Appl. Pharmacol.54,293300.[ISI][Medline]
von Meyerinck, L., Hufnagel, B., Schmoldt, A., and Benthe, H. F. (1990).Induction of rat liver microsomal cytochrome P-450 by the pentabromo diphenyl ether Bromkal 70 and half-lives of its components in the adipose tissue.Toxicology61,259274.[ISI][Medline]
Whyte, J. J., Jung, R. E., Schmitt, C. J., and Tillitt, D. E. (2000). Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chemical exposure.Crit. Rev. Toxicol.30,347570.[ISI][Medline]
Zhou, T., Ross, D. J., DeVito, M. J., and Crofton, K. M. (2001). Effects of short-termin vivo exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats.Toxicol. Sci.61,7682.
Zoeller, R. T., and Crofton, K. M. (2000). Thyroid hormone action in fetal brain development and potential for disruption by environmental chemicals.Neurotoxicology2,935945.