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Reduced thyroxine levels in mice perinatally exposed to polybrominated diphenyl ethers
Environmental Toxicology and Pharmacology 19 (2005) 273–281
Reduced thyroxine levels in mice perinatally exposed to polybrominated diphenyl ethers a , Agneta Oskarssona,∗ ¨ Ellen Skarmana , Per Ola Darnerudb , Helena Ohrvik a
Department of Biomedical Sciences and Veterinary Public Health, Swedish University of Agricultural Sciences, PO Box 7078, SE-75007 Uppsala, Sweden b Toxicology Division, National Food Administration, PO Box 622, SE-751 26 Uppsala, Sweden Received 12 December 2003; accepted 3 August 2004 Available online 27 September 2004
Abstract The aim of the study was to follow plasma thyroxine levels and hepatic enzyme activities in offspring after maternal gestational and lactational exposure to polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls. Mice were given 10 equimolar oral doses from gestational day (GD) 4 to postnatal day (PND) 17 of either Bromkal 70-5DE, 2,2 ,4,4 ,5-pentabrominated diphenyl ether (BDE-99) or Aroclor 1254 (total dose of 0.80 mmol/kg, b.w.). Plasma thyroxine levels were reduced in offspring in the Aroclor and Bromkal groups on PND11 but had returned to control levels by PND37. No effects on thyroxine levels were seen in the dams. Hepatic activity of EROD was increased in all treated offspring groups and so was UDP-GT in Aroclor-exposed offspring on PND11 and PND18. This study shows that PBDEs and PCBs, probably after microsomal transformation, have endocrine disrupting properties in perinatally exposed juvenile mice, most pronounced at PND11. However, BDE-99 had no effect on thyroxine levels, suggesting that other components in Bromkal are responsible for the hypothyroxinemia. © 2004 Elsevier B.V. All rights reserved. Keywords: Polybrominated diphenyl ethers; PBDEs; Thyroxine; Endocrine disruptor; Bromkal; BDE-99
1. Introduction Polybrominated diphenyl ether (PBDE) mixtures are currently used as flame retardants in various commercial products such as electronic equipment and textiles (Darnerud et al., 2001; De Wit, 2002). Due to the bio-accumulative characteristics of PBDEs, such as environmental persistence and high lipophilicity, these compounds have been detected in increasing concentrations in wildlife species (Ikonomou et al., 2002; Derocher et al., 2003) and in human tissues (Nor´en and Meironyt´e, 2000; Thomsen et al., 2002). An exponential increase of PBDEs in breast milk was found in Sweden between 1972 and 1997, after which the levels de∗ Corresponding author. Present address: Molecular Toxicology, School of Biomedical and Molecular Sciences, University of Surrey, Guildford Surrey, GU2 7XH, UK. Tel.: +44 1483 689714. E-mail address: [email protected] (A. Oskarsson).
1382-6689/$ – see front matter © 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.etap.2004.08.001
creased (Meironyt´e et al., 1999; Guvenius and Nor´en, 2001). The concentrations of PBDEs in maternal and fetal serum samples collected in the US have recently been reported to be 20–106-fold higher than those reported in Swedish serum samples (Mazdai et al., 2003; Guvenius et al., 2003). The predominant PBDE congeners found in human breast milk are tetra- and penta-BDEs and in Swedish breast milk 2,2 ,4,4 tetrabromodiphenyl ether (BDE-47) has been found to be the most common congener (Meironyt´e et al., 1999; Lind et al., 2003). Both PBDEs and PCBs bear structural resemblances to thyroid hormones and may thereby possess endocrine disrupting properties, causing interference with thyroid hormone homeostasis (McKinney and Waller, 1994; McDonald, 2002). Previous studies have shown that perinatal maternal exposure to PCB mixtures, such as Aroclor 1254, or individual PCB congeners, such as tetra- and hexa-chlorobiphenyls, reduces circulating thyroxine (T4) levels in adult and juve-
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nile rats (Morse et al., 1993, 1996; Seo et al., 1995; Crofton et al., 2000). Therefore, in the current study, Aroclor 1254 was chosen to serve as a positive control. Reduced T4 levels have also been reported in adult mice following exposure to PBDE mixtures, Bromkal 1254 or DE-71, or to the pure tetra-BDE congener, BDE-47 (Fowles et al., 1994; Hallgren et al., 2001), and in juvenile rats following perinatal maternal exposure to the PBDE mixture, DE-71 (Zhou et al., 2002). In humans, adverse effects on infant thyroid hormone homeostasis have been shown to be correlated to maternal exposure to background levels of PCBs (Nagayama et al., 1998). In a study by Koopman-Esseboom et al. (1994) it was observed that the PCB levels in breast milk collected from 105 women living in the Netherlands were negatively correlated to serum total T4 (TT4) levels in the newborn infants. Numerous attempts have been made to elucidate the mechanisms by which PCBs and PBDEs affect thyroid hormone homeostasis. Suggestions have been made that these compounds may reduce thyroid hormone levels following a cytochrome P450-mediated, metabolic conversion to hydroxylated metabolites. These metabolites exhibit properties enabling them to competitively bind to transthyretin (TTR), the major thyroid hormone transporting protein in rodents (Brouwer and van den Berg, 1986; Lans et al., 1993; Darnerud et al., 1996). Reduced T4 levels accompanied by induced cytochrome P450 activities, have been reported in PCB and PBDE exposed rats and mice (De Vito et al., 1993; Zhou et al., 2002; Craft et al., 2002). Also, other mechanisms for reduction of T4 levels have been discussed and investigated, including activation of the hepatic phase II metabolic enzyme, uridine-5-diphosphoglucuronosyltransferase (UDP-GT), leading to increased glucuronidation of T4 and increased bilary excretion of T4 (Bastomsky, 1974), direct effects (of PCB) on the thyroid gland (Collins and Capen, 1980a, 1980b; Ness et al., 1993; Brouwer et al., 1998) and competetive binding to the thyroid hormone receptor (Porterfield, 1994; Cheek et al., 1999). Other possible mechanisms may be an induction of 5 deiodinase activities, the enzyme that is responsible for the local conversion of T4 to tri-iodothyronin (T3) (Morse et al., 1993; Hood and Klaassen, 2000), and a non-functioning hypothalamic–pituitary–thyroid feedback system, possibly due to a suppression of the thyroid-stimulating hormone (TSH) (Brouwer et al., 1998; Zoeller et al., 2000). In conclusion, until the mechanisms of action of PBDEs are fully understood, it must be considered that more than one mechanism of action may operate. The objective of the present study was to compare the thyroid hormone disrupting properties of the pure BDE-99 congener to a commercial PBDE mixture and a commercial PCB mixture, in juvenile mice following maternal gestational and lactational exposure. To achieve a greater understanding of the mechanisms by which these compounds affect thyroid hormone levels, we also studied their effects on hepatic metabolic enzymes.
2. Materials and methods 2.1. Animals Primiparous NMRI mice, mated at 8 weeks of age and with confirmed vaginal plugs, were obtained from Taconic M & B Animal Resources, Denmark. The dams were housed in groups of five in standard plastic cages with sterilized pine shavings as bedding. The animals were given tap water and a standard pelleted diet (R36, Lactamin AB, Sweden) ad libitum and maintained on a 12 h light:12 h dark photoperiod. The animals were housed at a temperature of 22–23 ◦ C and with a 50–60% humidity. The study was approved by the Swedish Ethical Committee for Laboratory Animals. 2.2. Chemicals The pure congener BDE-99 (2,2 ,4,4 ,5-pentabrominated diphenyl ether) was synthesised (>99% purity) and generously supplied by G¨oran Marsh at the Department of Environmental Chemistry, Stockholm University, Stockholm. The technical mixture Bromkal 70-5DE (Bromkal, Kalk Chemische Fabrik, Germany) is a PBDE mixture with the main constituents being 2,2 ,4,4 ,5-penta-BDE (37%) and 2,2 ,4,4 tetra-BDE (35%) (Sj¨odin et al., 1998). The technical PCB mixture, Aroclor 1254 (Aroclor, Monsanto Chemical Company, USA) contains many different chlorinated biphenyls of which about 5% are CB-105. Solutions were prepared by mixing the compounds with corn oil (Di Luca & Di Luca, Sweden) and sonicating at 40 ◦ C for 30 min. 2.3. Experimental design The dams were randomly allocated to three groups including 13 animals per group and a larger control group consisting of 22 animals. Each dam was administered 80 mol/kg b.w. of either Aroclor, Bromkal, BDE-99 or pure corn oil every third day from GD4 through PND17, on a total of 10 occasions. The chemicals were administered orally and given in 3 ml oil solution/kg body weight. We allowed a 3-day administration withdrawal period before and after the estimated day of parturition, so as not to distress the dams. Equimolar doses were used and the total dose administered was 0.80 mmol/kg for all three chemicals, corresponding to 452 mg/kg BDE-99 or Bromkal and 262 mg/kg Aroclor. The dams were weighed prior to each administration, enabling correct dose adjustments. One week prior to the estimated day of parturition the dams were placed in individual cages and were given household paper as housing material. On GD17 four dams from each treatment group were euthanized using CO2 and liver and blood samples were collected. The blood was immediately centrifuged at 6000 rpm for 10 min and plasma was collected. All samples were stored at −70 ◦ C until analysis. Beginning on GD18, the dams were checked twice a day for signs of parturition and the day of parturition was denoted PND0. On
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PND3, the size of the litters was adjusted to 10 pups. The offspring were weighed by litter on GD17, PND3, PND6, PND13, and PND20 and individually on PND37. On PND11, PND18 and PND37, three to four pups from each litter were euthanized using CO2 and liver and plasma samples were collected. Regretfully, samples from offspring in the Bromkal group sampled on PND37 failed to be analysed. All samples were stored at −70 ◦ C until analysis. Due to the limited amount of tissue available, samples collected from offspring on PND11 and PND18 were pooled by litter. The dams were euthanized on PND20 and liver and blood samples were collected and stored at −70 ◦ C.
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is based on the competitive binding of plasma T4 and the 125 Ilabelled T4 to a limited number of binding sites on sheep antiT4 antibody. The proportion of the 125 I-labelled T4, bound to the antibody, is inversely related to the concentration of T4 present in the plasma. By measuring the proportion of 125 I-labelled T4 bound in the presence of reference plasma containing various known amounts of T4, the concentration of T4 in the unknown samples can be determined. The accuracy was checked by comparing with a standard solution provided by the Amerlex kit and by comparison with rodent reference samples from our laboratory. 2.7. Statistical analysis
2.4. Hepatic microsomal cytochrome P450 activity assay Liver samples, weighing approximately 0.6 g, were homogenized in cooled Potter–Elvehjem homogenizers containing 1 ml cooled 0.25 M sucrose–5 mM Tris–1 mM EDTA solution. The homogenates were centrifuged at 10,000 × g for 15 min at 4 ◦ C and microsomal pellets were obtained after centrifugation of the resulting supernatants at 105,000 × g for 1 h at 4 ◦ C. Once the second supernatant had been decanted, the pellets were rinsed twice in 0.1 M phosphate buffer pH 7.8, homogenized in 1.5 ml cooled 0.1 M phosphate buffer and stored in Eppendorf tubes at −70 ◦ C until assayed. Prior to performing the protein concentration assay and EROD analysis, 100 l of the microsomal suspensions were diluted with 900 l of 0.1 M phosphate buffer pH 7.8. The microsomal ethoxyresorufin-O-deethylase (EROD) activity (a marker for CYP1A1 activity) was determined fluorometrically using a method described by Brunstr¨om and Halldin (1998). The EROD assay was carried out in black multiwell plates and the incubation period was 10 min. The EROD values were calculated as pmol resorufin/(mg protein min). Microsomal protein concentrations were assayed using the method described by Smith et al. (1985), modified for microtiter plates. 2.5. Hepatic microsomal UDP-GT activity assay The UDP-GT activity measurements have been described in detail in Hallgren et al. (2001), using a modified method from Beetstra et al. (1991). Briefly, T4-UDP-GT activity was determined in microsomal incubations using 125 I-labelled T4 as substrate and uridine diphosphoglucuronic acid (UDPGA) as cofactor. After chromatographic separation, the formed glucuronyl product radioactivity was measured in selected fractions. 2.6. Thyroid hormone assay Plasma TT4 and FT4 levels were measured using radioimmunoassay (RIA) standard kits. To quantify the levels Amerlex-M and Amerlex-MAB kits (Ortho-Clinical Diagnostics, Amersham, UK) were used, respectively. The method
The distribution of the data was tested using a standard normality test, which showed that the data did not follow a normal distribution. Using Bartlett’s test, it was concluded that the data was heteroscedastic. Taking these observations into account, the Mann–Whitney U-test was chosen to compare treatment groups. A total of six pair-wise comparisons were performed at each sampling occasion. To protect against mass significance, the requirements for significant results were increased by a factor of six, moving the significance limit to p < 0.008. All results are presented as median values ± the interquartile range. All statistical analyses were performed using the software package Statview 5.0 (SAS Institute Inc., Cary. NC).
3. Results 3.1. Maternal and offspring body weights and liver/body weight ratios There were no significant treatment-related effects on dam or offspring body weights (Fig. 1A and B). Significantly increased liver/body weight ratios were seen on GD20 in dams treated with Aroclor and BDE-99 compared to controls (Fig. 2A). The ratios were also increased in the offspring from Aroclor treated dams relative to controls on PND11 and PND18 (Fig. 2B). On PND37, the liver/body weight ratios were significantly reduced in the Bromkal group relative to controls. 3.2. Reproductive parameters The pregnancy rate in the BDE-99 group was slightly lower relative to the other treatment groups. Of the 13 vaginal plugs confirmed animals in BDE-99 group, only 10 dams became pregnant compared to 12 and 13 in the Aroclor and Bromkal groups, respectively (Table 1). One litter in the Bromkal group was stillborn and two pups from two different litters in the BDE-99 group died during the day of parturition. There were no other apparent treatment-related effects on any reproductive parameters. Data from non-pregnant dams were excluded from all analyses.
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Fig. 1. Body weights of dams (A) and offspring (B) following perinatal exposure to Aroclor, Bromkal or BDE-99. Results are expressed as median ± interquartile range, n = 9–20 dams and litters.
3.3. Plasma thyroxine levels No treatment-related effects on plasma total thyroxine (TT4) or plasma free thyroxine (FT4) levels were seen in the pregnant or postweaning dams sampled on GD17 and PND20, respectively (Fig. 3). Following perinatal maternal exposure to Aroclor or Bromkal, significantly reduced plasma TT4 and FT4 levels were detected in offspring relative to controls (Fig. 4). In samples collected on PND11 both the Aroclor and the Bromkal group showed significantly reduced plasma TT4 and FT4 levTable 1 Reproductive parameters following perinatal exposure to Aroclor, Bromkal or BDE-99 Parameters
Aroclor
Bromkal
BDE-99
Control
Number of animals Pregnancy rate Gestation length (days)
13 12/13 19
13 13/13 19
13 10/13 19
22 20/22 19
Litter size Mean number Range; number
10.5 8–13
11 6–13
11 9–13
12 10–14
Litter size was determined at birth. Data from non-pregnant dams were excluded from all analyses.
Fig. 2. Liver/body weight ratio in dams (A) and offspring (B) following perinatal exposure to Aroclor 1254, Bromkal DE 70-5 or BDE-99. Results are expressed as median ± interquartile range, n = 9–20 dams and litters. Bars not sharing a letter within an age group are significantly different, p < 0.008.
els. At this sampling occasion the Bromkal groups also had significantly lower plasma TT4 and plasma FT4 levels relative to the BDE-99 group. On PND18 a significant reduction of the plasma TT4 and FT4 levels were only seen in the Aroclor group. The treatment-related plasma thyroxine reduction seen in the Aroclor group was of greater magnitude on PND11 than on PND 18 (plasma TT4 being 44% and 72%, respectively, of controls). Plasma TT4 and FT4 levels followed a similar pattern of treatment-related reduction in offspring on both PND11 and PND18. The reduction of both plasma TT4 and plasma FT4 levels were significantly larger in the Aroclor group relative to the Bromkal group on PND11. Offspring plasma TT4 and FT4 levels had returned to control levels on PND37. No plasma TT4 or FT4 alterations were detected in offspring following maternal exposure to BDE-99. 3.4. EROD activity In pregnant dams, sampled on GD17, an induced hepatic microsomal EROD activity was seen in both the Aroclor and the Bromkal group relative to controls, but not in the BDE-99 group (Fig. 5A). The induction was significantly higher in the
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Fig. 3. Plasma TT4 (A) and FT4 (B) levels in dams following exposure to Aroclor, Bromkal or BDE-99 during gestation and lactation. Results are expressed as median ± interquartile range, n = 4–11. Bars not sharing a letter within an age group are significantly different, p < 0.008.
Fig. 5. Hepatic EROD activity in dams (A) and offspring (B) following perinatal exposure to Aroclor, Bromkal or BDE-99, n = 4–16. Results are expressed as median ± interquartile range. Bars not sharing a letter within an age group are significantly different, p < 0.008.
Fig. 4. Plasma TT4 (A) and FT4 (B) levels in offspring following perinatal exposure to Aroclor, Bromkal or BDE-99. Results are expressed as median ± interquartile range, n = 6–16. Bars not sharing a letter within an age group are significantly different, p < 0.008.
Aroclor group relative to the Bromkal group. No significant differences in EROD activity were seen in the Bromkal group relative to the BDE-99 group. In postweaning dams sampled on PND20, the EROD activity was significantly increased in the Aroclor group relative to controls. The hepatic EROD induction was of a greater magnitude in the pregnant dams compared to the induction in the postweaning dams. The Aroclor group showed a 3.2-fold increase in EROD activity on GD17 and a 1.4-fold increase on PND20. Offspring sampled on PND11 and PND18 showed increased hepatic microsomal EROD activity in all treatment groups relative to controls (Fig. 5B). The hepatic EROD activity was increased in the Aroclor group, relative to controls, with a factor of 15 on PND11 versus 4.8 on PND18. In all treatment groups the hepatic EROD activity had resumed control levels by PND37. On both PND11 and PND18 the EROD activity was significantly higher in the Aroclor group relative to the Bromkal group. However, no significant differences in EROD activity were seen in the Bromkal group relative to the BDE-99 group on neither PND11 nor PND18.
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Fig. 6. Hepatic UDP-GT activity in offspring following perinatal exposure to Aroclor, Bromkal or BDE-99. Only samples from PND11 and PND18 were analyzed. Results are expressed as median ± interquartile range. Bars not sharing a letter within an age group are significantly different, p < 0.008; (b) denotes borderline significance, p < 0.0084.
3.5. Hepatic UDP-GT activity The UDP-GT activity was studied only in offspring at PND11 and PND18. A significant increase in enzyme activity was observed in the Aroclor group at both time-points (Fig. 6). The median activity in the Aroclor group was about twice that of the control group. At PND18, the increase in the Bromkal group was of borderline significance (p = 0.0084).
4. Discussion The current study enables a comparison of the effects following exposure to the commercial PBDE mixture, Bromkal 70-5DE, to effects following exposure to the PCB mixture, Aroclor 1254, and the pure pentabrominated (BDE-99) congener. Perinatal maternal exposure to Bromkal caused significantly reduced plasma T4 levels in the offspring, indicating that PBDEs possess endocrine disrupting properties. Exposure to the pure pentabrominated congener BDE-99, however, did not significantly reduce plasma T4 levels in offspring relative to controls, at any sampling occasion. This may be an indication that BDE-99 does not affect or is a less potent reducer of plasma T4 levels than the other main component in the Bromkal mixture, BDE-47. This is consistent with the finding, reported by Hallgren et al. (2001) that BDE-47 was a slightly more potent reducer of plasma TT4 and FT4 levels in mice than Bromkal 70-5DE, when administered at equimolar doses. The hypothesis that lower brominated congeners possess greater plasma T4 reducing abilities is also supported by the finding that BDE-mixtures, containing lower brominated BDE congeners, were shown to be more potent plasma T4 reducers in weanling rats, than mixtures including higher brominated congeners. This was seen following exposure to DE-71 (58% penta-BDE, 25% tetra-BDE), DE-79 (31% octa-BDE and 45% hepta-BDE) or DE-83R (98% deca-BDE) (Zhou et al., 2001).
Both BDE-99 and Bromkal exposure caused significant increases in the hepatic EROD activity in perinatally exposed offspring. In earlier studies an association between microsomal CYP induction and the reduction of circulating thyroid hormone level have been suggested (Hallgren et al., 2001; Hallgren and Darnerud, 2002). In a study by Meerts et al. (2000) three structurally different hydroxylated PBDE congeners were tested for their in vitro potency to displace T4 from TTR. The three hydroxylated compounds, resembling different thyroid hormones, showed a large difference in this respect. The results indicated that microsomal metabolism of PBDE will create hydroxylated products that are able to bind to TTR and affect serum T4 levels, but that the structure of the metabolite will be decisive for the binding affinity and subsequent T4 decrease. In the case of BDE-47 and BDE-99 an in vitro metabolism model demonstrated that phenobarbitalinduced microsomes (reflecting CYP2B activity) were most effective in producing extracts competing with T4 for TTR binding (Meerts et al., 2000). Consequently, CYP2B seems to produce BDE-47 and BDE-99 metabolites with TTR binding properties, which could result in decreased serum T4 levels (however not seen in our in vivo study after BDE-99 exposure). Aroclor exposure resulted in an induction of EROD and UDP-GT concurrent with a decrease in plasma T4. There was no induction of UDP-GT in offspring after perinatal Bromkal exposure at PND11, when there was a reduction in plasma T4. A limited induction of UDP-GT activities has previously been reported in rats following both PBDE and PCB exposure (Hallgren et al., 2001; Zhou et al., 2002) and in mice following exposure to PCBs (Craft et al., 2002) but not following exposure to PBDEs (Hallgren et al., 2001). This indicates that UDP-GT to some extent may be involved in thyroid hormone reduction in rats and mice, following exposure to PCBs, but only in rats following PBDE exposure. However, the limited enzyme induction suggests that other mechanisms than elevated UDP-GT activities play the major role in reducing T4 levels. The reduced plasma T4 levels and induced EROD activities seen in the offspring were significantly higher in the Aroclor group compared to the Bromkal group. This indicates that the PCB mixture is a more potent thyroid hormone disruptor and inducer of hepatic metabolic enzymes than the PBDE mixture. These results are consistent with previous findings in rats and mice (von Meyerinck et al., 1990; Hallgren et al., 2001) and may indicate that PCBs and PBDEs have different mechanisms of action or that PCBs merely possess properties rendering them greater thyroid hormone disrupting abilities. Exposure to Aroclor and Bromkal caused reduced plasma T4 levels in offspring. However, no significant plasma T4 reductions could be detected in the exposed dams. This indicates that offspring are more sensitive to the thyroid hormone disrupting properties of PBDEs and PCBs than pregnant and postweaning dams. These results are consistent with findings in rats reported by Zhou et al. (2002), in which a maternal
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exposure to 180 mg/kg b.w. DE-71 caused significantly reduced serum T4 levels in offspring but a maternal exposure to 350 mg/kg b.w. DE-71 was not sufficient to cause reduced serum T4 levels in the exposed dams. Also, the effect of PBDEs and PCBs on EROD activity was more pronounced in the offspring compared to the dams. The absence of maternal T4 reduction in our study may be due to less sensitive thyroid hormone regulating mechanisms in dams and reduced T4 levels may, therefore, have been detected if higher doses had been administered. The hypothyroxinemia and induced hepatic metabolic enzyme activity in the offspring, following exposure to Bromkal 70-5DE, is consistent with the results reported in a study performed by Zhou et al. (2002), in which reduced serum T4 levels and induced EROD and UDP-GT activities were seen in juvenile rats following daily gestational and lactational maternal exposure to the PBDE mixture, DE-71 (58% penta-BDE, 25% tetra-BDE). Reduced T4 levels and induced EROD activities have also been reported in adult mice, following exposure to DE-71 (Fowles et al., 1994) and in adult non-pregnant rats and mice following exposure to Bromkal 70-5DE and BDE-47 (Hallgren et al., 2001). Although, in our study, no maternal T4 reduction was seen in pregnant or lactating mice (following exposure to 131 versus 262 mg/kg b.w. Aroclor 1254), reduced T4 levels have been detected in pregnant and lactating rats exposed to 90 and 210 mg/kg Aroclor 1254 (Crofton et al., 2000). This may indicate that rats are more sensitive to PCB-induced T4 reduction than mice. A greater magnitude of T4 reduction in non-pregnant rats compared to non-pregnant mice was also seen by Hallgren et al. (2001) following exposure to Bromkal 70-5DE. Previous studies in rats indicate that PCB induced hypothyroidism reaches a maximum during the second and third postnatal week, not only when maternal exposure occurs both pre- and post-natally (Goldey et al., 1995; Goldey and Crofton, 1998) but also following only prenatal exposure (Morse et al., 1993). This coincides with the peak in exposure, which is via milk. In a cross-fostering study by Crofton et al. (2000) it was found that the lactational exposure, in contrast to the in utero exposure, was the major route of PCB exposure responsible for the reduction in offspring thyroid hormone levels. In the present study, we found that the reduction of offspring plasma T4 levels was of a greater magnitude relative to controls on PND11 than on PND18, although the offspring sampled on PND11 had been exposed to lower total doses than the offspring sampled on PND18. We also found that the T4 levels had returned to control levels by PND37. This indicates that the peak sensitivity to thyroid hormone disruption occurs in mice during the second postnatal week and that the sensitivity of this hormone system thereafter declines. It is of interest to note that the peak sensitivity to PBDE-induced behavioural alterations also has been shown to occur in mice during the second postnatal week (Eriksson et al., 1999, 2002).
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Since the first few postnatal weeks in rodents is the time period of peak sensitivity of auditory (Uziel et al., 1985a, 1985b; Hebert et al., 1985) and neuron development (Eriksson et al., 2001), hypothyroxinemia during this time period may have numerous adverse effects on the development of these structures (Crofton et al., 2000). Also, the effects on the auditory system (Goldey et al., 1995) and on behaviour (Eriksson et al., 2001; Branchi et al., 2002), have been shown to persist into adult life, although the observed hypothyroxinemia in itself is transient (Morse et al., 1996; Goldey and Crofton, 1998). In a study performed by Eriksson et al. (2001) the behavioural changes seen in mice following exposure to BDE-99 were actually seen to increase with age, being more pronounced on PND120 than on PND60. The transient thyroid hormone disruption during a critical phase of neonatal development may play a role in the developmental deficiencies not identifiable until later on in life. Contradictory to the behavioural effects observed after BDE-99 exposure, in the present study BDE99 had no effect on T4 homeostasis, which seems to speak against the proposed T4-mediated mechanism for effects on behaviour. This discrepancy may however be explained by differences in exposure regimens of the respective studies. For example, in the present study the fetus/pups were indirectly exposed via the dam/mother, whereas in the study by Eriksson et al. (2001) the pups were given the substance directly, in a single oral dose. In conclusion, the results of this study suggest that perinatal PBDE exposure disrupts thyroid hormone homeostasis in juvenile mice and that the offspring seems to be more susceptible to thyroid hormone reduction compared to their dams. BDE-99 seems to be a less potent reducer, or may not be a reducer at all, of plasma thyroid hormone levels than BDE47, the other major component of Bromkal 70-5DE. On an equimolar basis, PCBs possess greater thyroid hormone disrupting properties than PBDEs. Since both PCB and PBDE exposure caused induced hepatic cytochrome P450 activities and we know that certain hydroxylated PBDE metabolites have TTR binding potencies, it is possible that this may be one of the major mechanisms involved in reducing circulating thyroid hormone levels in exposed mice. However, PCB but not Bromkal induced UDP-GT, indicating that there may be different mechanisms of action for the two chemical mixtures or at least a difference in potency. Knowing that a functioning thyroid homeostasis is essential for normal juvenile development, the results of this study are of concern when evaluating potentially hazardous properties of the studied compounds.
Acknowledgements The authors would like to thank Agneta Bostr¨om and Elvy Netzel for excellent technical assistance. This work has been financed by grants from the Foundation for Strategic Environmental Research within the program, A New Strategy for Risk Assessment and Management of Chemicals.
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E. Skarman et al. / Environmental Toxicology and Pharmacology 19 (2005) 273–281 competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicol. Sci. 56, 95. ˚ 1999. Analysis of polybrominated Meironyt´e, D., Nor´en, K., Bergman, A., diphenyl ethers in Swedish human milk. A time-related trend study, 1972–1997. J. Toxicol. Environ. Health 58, 101. Morse, D.C., Groen, D., Veerman, M., van Amerongen, C.J., Ko¨eter, H.B.W.M., Van Prooije, A.E.S., Visser, T.J., Koeman, J.H., Brouwer, A., 1993. Interference of polychlorinated biphenyls in hepatic and brain thyroid hormone metabolism in fetal and neonatal rats. Toxicol. Appl. Pharmacol. 122, 27. Morse, D.C., Klasson Wehler, E., Wesseling, W., Koeman, J.H., Brouwer, A., 1996. Alterations in rat thyroid hormone status following preand postnatal exposure to polychlorinated biphenyls (Aroclor 1254). Toxicol. Appl. Pharmacol. 136, 269. Nagayama, J., Okamura, K., Iida, T., Hirakawa, H., Matsueda, T., Tsuji, H., Hasegawa, M., Sato, K., Na, H., Yanagawa, T., Igarashi, H., Fukushige, J., Watanabe, T., 1998. Postnatal exposure to chlorinated dioxins and related chemicals on thyroid hormone status in Japanese breast-fed infants. Chemosphere 37, 1789. Ness, D.K., Schantz, S.L., Moshtaghian, J., Hansen, L.G., 1993. Effects of perinatal exposure to specific PCB congeners on the thyroid hormone concentrations and thyroid histology in the rat. Toxicol. Lett. 68, 311. Nor´en, K., Meironyt´e, D., 2000. Certain organochlorine and organobromine contaminants in Swedish human milk in perspective of past 20–30 years. Chemosphere 40, 1111. Porterfield, S.P., 1994. Vulnerability of the developing brain to thyroid abnormalities: environmental insults to the thyroid system. Environ. Health Perspect. 102 (suppl. 2), 125. Seo, B.W., Li, M.H., Hansen, L.G., Moore, R.W., Peterson, R.E., Schantz, S., 1995. Effects of gestational and lactational exposure to coplanar ploychlorinated biphenyl (PCB) congeners or 2,3,7,8,tetrachlorodibenzo-p-dioxin (TCDD) on thyroid hormone concentrations in weanling rats. Toxicol. Lett. 78, 253.
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Sj¨odin, A., Jakobsson, E., Kirkegaard, A., Marsh, G., Sellstr¨om, U., 1998. Gas chromatographic identification and quantification of polybrominated diphenyl ethers in a commercial product, Bromkal 70-5DE. J. Chromatogr. 822, 83. Smith, P.K., Krohn, R.I., Hermanson, G.T., Mallia, A.K., Gartner, F.H., Provenzano, M.D., Fujimoto, E.K., Goeke, N.M., Olson, B.J., Klenk, D.C., 1985. Measurement of protein using bicinchoninic acid. Anal. Biochem. 150, 76. Thomsen, C., Lundanes, E., Becher, G., 2002. Brominated flame retardants in archived serum samples from Norway; a study on temporal trends and the role of age. Environ. Sci. Technol. 36, 1414. Uziel, A., Legrand, C., Rabie, A., 1985a. Corrective effects of thyroxine on cochlear abnormalities induced by congenital hypothyroidism in the rat. I. Morphological study. Develop. Brain Res. 19, 111. Uziel, A., Marot, M., Rabie, A., 1985b. Corrective effects of thyroxine on cochlear abnormalities induced by congenital hypothyroidism in the rat. II. Electrophysiological study. Brain Res. 351, 123. von Meyerinck, L., Hufnagel, B., Schmoldt, A., 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. Toxicology 61, 259. Zhou, T., Ross, D.G., DeVito, M.J., Crofton, K.M., 2001. Effects of shortterm in vivo exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats. Toxicol. Sci. 61, 76. Zhou, T., Taylor, M.M., DeVito, M.J., Crofton, K.M., 2002. Developmental exposure to brominated diphenyl ethers results in thyroid hormone disruption. Toxicol. Sci. 66, 105. Zoeller, R.T., Dowling, A.L.S., Vas, A.A., 2000. Developmental exposure to polychlorinated biphenyls exerts thyroid hormone-like effects on the expression of RC3/neurogranin and myelin basic protein messenger ribonucleic acids in the developing rat brain. Endocrinology 141, 181.
Reduced thyroxine levels in mice perinatally exposed to polybrominated diphenyl ethers a , Agneta Oskarssona,∗ ¨ Ellen Skarmana , Per Ola Darnerudb , Helena Ohrvik a
Department of Biomedical Sciences and Veterinary Public Health, Swedish University of Agricultural Sciences, PO Box 7078, SE-75007 Uppsala, Sweden b Toxicology Division, National Food Administration, PO Box 622, SE-751 26 Uppsala, Sweden Received 12 December 2003; accepted 3 August 2004 Available online 27 September 2004
Abstract The aim of the study was to follow plasma thyroxine levels and hepatic enzyme activities in offspring after maternal gestational and lactational exposure to polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls. Mice were given 10 equimolar oral doses from gestational day (GD) 4 to postnatal day (PND) 17 of either Bromkal 70-5DE, 2,2 ,4,4 ,5-pentabrominated diphenyl ether (BDE-99) or Aroclor 1254 (total dose of 0.80 mmol/kg, b.w.). Plasma thyroxine levels were reduced in offspring in the Aroclor and Bromkal groups on PND11 but had returned to control levels by PND37. No effects on thyroxine levels were seen in the dams. Hepatic activity of EROD was increased in all treated offspring groups and so was UDP-GT in Aroclor-exposed offspring on PND11 and PND18. This study shows that PBDEs and PCBs, probably after microsomal transformation, have endocrine disrupting properties in perinatally exposed juvenile mice, most pronounced at PND11. However, BDE-99 had no effect on thyroxine levels, suggesting that other components in Bromkal are responsible for the hypothyroxinemia. © 2004 Elsevier B.V. All rights reserved. Keywords: Polybrominated diphenyl ethers; PBDEs; Thyroxine; Endocrine disruptor; Bromkal; BDE-99
1. Introduction Polybrominated diphenyl ether (PBDE) mixtures are currently used as flame retardants in various commercial products such as electronic equipment and textiles (Darnerud et al., 2001; De Wit, 2002). Due to the bio-accumulative characteristics of PBDEs, such as environmental persistence and high lipophilicity, these compounds have been detected in increasing concentrations in wildlife species (Ikonomou et al., 2002; Derocher et al., 2003) and in human tissues (Nor´en and Meironyt´e, 2000; Thomsen et al., 2002). An exponential increase of PBDEs in breast milk was found in Sweden between 1972 and 1997, after which the levels de∗ Corresponding author. Present address: Molecular Toxicology, School of Biomedical and Molecular Sciences, University of Surrey, Guildford Surrey, GU2 7XH, UK. Tel.: +44 1483 689714. E-mail address: [email protected] (A. Oskarsson).
1382-6689/$ – see front matter © 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.etap.2004.08.001
creased (Meironyt´e et al., 1999; Guvenius and Nor´en, 2001). The concentrations of PBDEs in maternal and fetal serum samples collected in the US have recently been reported to be 20–106-fold higher than those reported in Swedish serum samples (Mazdai et al., 2003; Guvenius et al., 2003). The predominant PBDE congeners found in human breast milk are tetra- and penta-BDEs and in Swedish breast milk 2,2 ,4,4 tetrabromodiphenyl ether (BDE-47) has been found to be the most common congener (Meironyt´e et al., 1999; Lind et al., 2003). Both PBDEs and PCBs bear structural resemblances to thyroid hormones and may thereby possess endocrine disrupting properties, causing interference with thyroid hormone homeostasis (McKinney and Waller, 1994; McDonald, 2002). Previous studies have shown that perinatal maternal exposure to PCB mixtures, such as Aroclor 1254, or individual PCB congeners, such as tetra- and hexa-chlorobiphenyls, reduces circulating thyroxine (T4) levels in adult and juve-
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nile rats (Morse et al., 1993, 1996; Seo et al., 1995; Crofton et al., 2000). Therefore, in the current study, Aroclor 1254 was chosen to serve as a positive control. Reduced T4 levels have also been reported in adult mice following exposure to PBDE mixtures, Bromkal 1254 or DE-71, or to the pure tetra-BDE congener, BDE-47 (Fowles et al., 1994; Hallgren et al., 2001), and in juvenile rats following perinatal maternal exposure to the PBDE mixture, DE-71 (Zhou et al., 2002). In humans, adverse effects on infant thyroid hormone homeostasis have been shown to be correlated to maternal exposure to background levels of PCBs (Nagayama et al., 1998). In a study by Koopman-Esseboom et al. (1994) it was observed that the PCB levels in breast milk collected from 105 women living in the Netherlands were negatively correlated to serum total T4 (TT4) levels in the newborn infants. Numerous attempts have been made to elucidate the mechanisms by which PCBs and PBDEs affect thyroid hormone homeostasis. Suggestions have been made that these compounds may reduce thyroid hormone levels following a cytochrome P450-mediated, metabolic conversion to hydroxylated metabolites. These metabolites exhibit properties enabling them to competitively bind to transthyretin (TTR), the major thyroid hormone transporting protein in rodents (Brouwer and van den Berg, 1986; Lans et al., 1993; Darnerud et al., 1996). Reduced T4 levels accompanied by induced cytochrome P450 activities, have been reported in PCB and PBDE exposed rats and mice (De Vito et al., 1993; Zhou et al., 2002; Craft et al., 2002). Also, other mechanisms for reduction of T4 levels have been discussed and investigated, including activation of the hepatic phase II metabolic enzyme, uridine-5-diphosphoglucuronosyltransferase (UDP-GT), leading to increased glucuronidation of T4 and increased bilary excretion of T4 (Bastomsky, 1974), direct effects (of PCB) on the thyroid gland (Collins and Capen, 1980a, 1980b; Ness et al., 1993; Brouwer et al., 1998) and competetive binding to the thyroid hormone receptor (Porterfield, 1994; Cheek et al., 1999). Other possible mechanisms may be an induction of 5 deiodinase activities, the enzyme that is responsible for the local conversion of T4 to tri-iodothyronin (T3) (Morse et al., 1993; Hood and Klaassen, 2000), and a non-functioning hypothalamic–pituitary–thyroid feedback system, possibly due to a suppression of the thyroid-stimulating hormone (TSH) (Brouwer et al., 1998; Zoeller et al., 2000). In conclusion, until the mechanisms of action of PBDEs are fully understood, it must be considered that more than one mechanism of action may operate. The objective of the present study was to compare the thyroid hormone disrupting properties of the pure BDE-99 congener to a commercial PBDE mixture and a commercial PCB mixture, in juvenile mice following maternal gestational and lactational exposure. To achieve a greater understanding of the mechanisms by which these compounds affect thyroid hormone levels, we also studied their effects on hepatic metabolic enzymes.
2. Materials and methods 2.1. Animals Primiparous NMRI mice, mated at 8 weeks of age and with confirmed vaginal plugs, were obtained from Taconic M & B Animal Resources, Denmark. The dams were housed in groups of five in standard plastic cages with sterilized pine shavings as bedding. The animals were given tap water and a standard pelleted diet (R36, Lactamin AB, Sweden) ad libitum and maintained on a 12 h light:12 h dark photoperiod. The animals were housed at a temperature of 22–23 ◦ C and with a 50–60% humidity. The study was approved by the Swedish Ethical Committee for Laboratory Animals. 2.2. Chemicals The pure congener BDE-99 (2,2 ,4,4 ,5-pentabrominated diphenyl ether) was synthesised (>99% purity) and generously supplied by G¨oran Marsh at the Department of Environmental Chemistry, Stockholm University, Stockholm. The technical mixture Bromkal 70-5DE (Bromkal, Kalk Chemische Fabrik, Germany) is a PBDE mixture with the main constituents being 2,2 ,4,4 ,5-penta-BDE (37%) and 2,2 ,4,4 tetra-BDE (35%) (Sj¨odin et al., 1998). The technical PCB mixture, Aroclor 1254 (Aroclor, Monsanto Chemical Company, USA) contains many different chlorinated biphenyls of which about 5% are CB-105. Solutions were prepared by mixing the compounds with corn oil (Di Luca & Di Luca, Sweden) and sonicating at 40 ◦ C for 30 min. 2.3. Experimental design The dams were randomly allocated to three groups including 13 animals per group and a larger control group consisting of 22 animals. Each dam was administered 80 mol/kg b.w. of either Aroclor, Bromkal, BDE-99 or pure corn oil every third day from GD4 through PND17, on a total of 10 occasions. The chemicals were administered orally and given in 3 ml oil solution/kg body weight. We allowed a 3-day administration withdrawal period before and after the estimated day of parturition, so as not to distress the dams. Equimolar doses were used and the total dose administered was 0.80 mmol/kg for all three chemicals, corresponding to 452 mg/kg BDE-99 or Bromkal and 262 mg/kg Aroclor. The dams were weighed prior to each administration, enabling correct dose adjustments. One week prior to the estimated day of parturition the dams were placed in individual cages and were given household paper as housing material. On GD17 four dams from each treatment group were euthanized using CO2 and liver and blood samples were collected. The blood was immediately centrifuged at 6000 rpm for 10 min and plasma was collected. All samples were stored at −70 ◦ C until analysis. Beginning on GD18, the dams were checked twice a day for signs of parturition and the day of parturition was denoted PND0. On
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PND3, the size of the litters was adjusted to 10 pups. The offspring were weighed by litter on GD17, PND3, PND6, PND13, and PND20 and individually on PND37. On PND11, PND18 and PND37, three to four pups from each litter were euthanized using CO2 and liver and plasma samples were collected. Regretfully, samples from offspring in the Bromkal group sampled on PND37 failed to be analysed. All samples were stored at −70 ◦ C until analysis. Due to the limited amount of tissue available, samples collected from offspring on PND11 and PND18 were pooled by litter. The dams were euthanized on PND20 and liver and blood samples were collected and stored at −70 ◦ C.
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is based on the competitive binding of plasma T4 and the 125 Ilabelled T4 to a limited number of binding sites on sheep antiT4 antibody. The proportion of the 125 I-labelled T4, bound to the antibody, is inversely related to the concentration of T4 present in the plasma. By measuring the proportion of 125 I-labelled T4 bound in the presence of reference plasma containing various known amounts of T4, the concentration of T4 in the unknown samples can be determined. The accuracy was checked by comparing with a standard solution provided by the Amerlex kit and by comparison with rodent reference samples from our laboratory. 2.7. Statistical analysis
2.4. Hepatic microsomal cytochrome P450 activity assay Liver samples, weighing approximately 0.6 g, were homogenized in cooled Potter–Elvehjem homogenizers containing 1 ml cooled 0.25 M sucrose–5 mM Tris–1 mM EDTA solution. The homogenates were centrifuged at 10,000 × g for 15 min at 4 ◦ C and microsomal pellets were obtained after centrifugation of the resulting supernatants at 105,000 × g for 1 h at 4 ◦ C. Once the second supernatant had been decanted, the pellets were rinsed twice in 0.1 M phosphate buffer pH 7.8, homogenized in 1.5 ml cooled 0.1 M phosphate buffer and stored in Eppendorf tubes at −70 ◦ C until assayed. Prior to performing the protein concentration assay and EROD analysis, 100 l of the microsomal suspensions were diluted with 900 l of 0.1 M phosphate buffer pH 7.8. The microsomal ethoxyresorufin-O-deethylase (EROD) activity (a marker for CYP1A1 activity) was determined fluorometrically using a method described by Brunstr¨om and Halldin (1998). The EROD assay was carried out in black multiwell plates and the incubation period was 10 min. The EROD values were calculated as pmol resorufin/(mg protein min). Microsomal protein concentrations were assayed using the method described by Smith et al. (1985), modified for microtiter plates. 2.5. Hepatic microsomal UDP-GT activity assay The UDP-GT activity measurements have been described in detail in Hallgren et al. (2001), using a modified method from Beetstra et al. (1991). Briefly, T4-UDP-GT activity was determined in microsomal incubations using 125 I-labelled T4 as substrate and uridine diphosphoglucuronic acid (UDPGA) as cofactor. After chromatographic separation, the formed glucuronyl product radioactivity was measured in selected fractions. 2.6. Thyroid hormone assay Plasma TT4 and FT4 levels were measured using radioimmunoassay (RIA) standard kits. To quantify the levels Amerlex-M and Amerlex-MAB kits (Ortho-Clinical Diagnostics, Amersham, UK) were used, respectively. The method
The distribution of the data was tested using a standard normality test, which showed that the data did not follow a normal distribution. Using Bartlett’s test, it was concluded that the data was heteroscedastic. Taking these observations into account, the Mann–Whitney U-test was chosen to compare treatment groups. A total of six pair-wise comparisons were performed at each sampling occasion. To protect against mass significance, the requirements for significant results were increased by a factor of six, moving the significance limit to p < 0.008. All results are presented as median values ± the interquartile range. All statistical analyses were performed using the software package Statview 5.0 (SAS Institute Inc., Cary. NC).
3. Results 3.1. Maternal and offspring body weights and liver/body weight ratios There were no significant treatment-related effects on dam or offspring body weights (Fig. 1A and B). Significantly increased liver/body weight ratios were seen on GD20 in dams treated with Aroclor and BDE-99 compared to controls (Fig. 2A). The ratios were also increased in the offspring from Aroclor treated dams relative to controls on PND11 and PND18 (Fig. 2B). On PND37, the liver/body weight ratios were significantly reduced in the Bromkal group relative to controls. 3.2. Reproductive parameters The pregnancy rate in the BDE-99 group was slightly lower relative to the other treatment groups. Of the 13 vaginal plugs confirmed animals in BDE-99 group, only 10 dams became pregnant compared to 12 and 13 in the Aroclor and Bromkal groups, respectively (Table 1). One litter in the Bromkal group was stillborn and two pups from two different litters in the BDE-99 group died during the day of parturition. There were no other apparent treatment-related effects on any reproductive parameters. Data from non-pregnant dams were excluded from all analyses.
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Fig. 1. Body weights of dams (A) and offspring (B) following perinatal exposure to Aroclor, Bromkal or BDE-99. Results are expressed as median ± interquartile range, n = 9–20 dams and litters.
3.3. Plasma thyroxine levels No treatment-related effects on plasma total thyroxine (TT4) or plasma free thyroxine (FT4) levels were seen in the pregnant or postweaning dams sampled on GD17 and PND20, respectively (Fig. 3). Following perinatal maternal exposure to Aroclor or Bromkal, significantly reduced plasma TT4 and FT4 levels were detected in offspring relative to controls (Fig. 4). In samples collected on PND11 both the Aroclor and the Bromkal group showed significantly reduced plasma TT4 and FT4 levTable 1 Reproductive parameters following perinatal exposure to Aroclor, Bromkal or BDE-99 Parameters
Aroclor
Bromkal
BDE-99
Control
Number of animals Pregnancy rate Gestation length (days)
13 12/13 19
13 13/13 19
13 10/13 19
22 20/22 19
Litter size Mean number Range; number
10.5 8–13
11 6–13
11 9–13
12 10–14
Litter size was determined at birth. Data from non-pregnant dams were excluded from all analyses.
Fig. 2. Liver/body weight ratio in dams (A) and offspring (B) following perinatal exposure to Aroclor 1254, Bromkal DE 70-5 or BDE-99. Results are expressed as median ± interquartile range, n = 9–20 dams and litters. Bars not sharing a letter within an age group are significantly different, p < 0.008.
els. At this sampling occasion the Bromkal groups also had significantly lower plasma TT4 and plasma FT4 levels relative to the BDE-99 group. On PND18 a significant reduction of the plasma TT4 and FT4 levels were only seen in the Aroclor group. The treatment-related plasma thyroxine reduction seen in the Aroclor group was of greater magnitude on PND11 than on PND 18 (plasma TT4 being 44% and 72%, respectively, of controls). Plasma TT4 and FT4 levels followed a similar pattern of treatment-related reduction in offspring on both PND11 and PND18. The reduction of both plasma TT4 and plasma FT4 levels were significantly larger in the Aroclor group relative to the Bromkal group on PND11. Offspring plasma TT4 and FT4 levels had returned to control levels on PND37. No plasma TT4 or FT4 alterations were detected in offspring following maternal exposure to BDE-99. 3.4. EROD activity In pregnant dams, sampled on GD17, an induced hepatic microsomal EROD activity was seen in both the Aroclor and the Bromkal group relative to controls, but not in the BDE-99 group (Fig. 5A). The induction was significantly higher in the
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Fig. 3. Plasma TT4 (A) and FT4 (B) levels in dams following exposure to Aroclor, Bromkal or BDE-99 during gestation and lactation. Results are expressed as median ± interquartile range, n = 4–11. Bars not sharing a letter within an age group are significantly different, p < 0.008.
Fig. 5. Hepatic EROD activity in dams (A) and offspring (B) following perinatal exposure to Aroclor, Bromkal or BDE-99, n = 4–16. Results are expressed as median ± interquartile range. Bars not sharing a letter within an age group are significantly different, p < 0.008.
Fig. 4. Plasma TT4 (A) and FT4 (B) levels in offspring following perinatal exposure to Aroclor, Bromkal or BDE-99. Results are expressed as median ± interquartile range, n = 6–16. Bars not sharing a letter within an age group are significantly different, p < 0.008.
Aroclor group relative to the Bromkal group. No significant differences in EROD activity were seen in the Bromkal group relative to the BDE-99 group. In postweaning dams sampled on PND20, the EROD activity was significantly increased in the Aroclor group relative to controls. The hepatic EROD induction was of a greater magnitude in the pregnant dams compared to the induction in the postweaning dams. The Aroclor group showed a 3.2-fold increase in EROD activity on GD17 and a 1.4-fold increase on PND20. Offspring sampled on PND11 and PND18 showed increased hepatic microsomal EROD activity in all treatment groups relative to controls (Fig. 5B). The hepatic EROD activity was increased in the Aroclor group, relative to controls, with a factor of 15 on PND11 versus 4.8 on PND18. In all treatment groups the hepatic EROD activity had resumed control levels by PND37. On both PND11 and PND18 the EROD activity was significantly higher in the Aroclor group relative to the Bromkal group. However, no significant differences in EROD activity were seen in the Bromkal group relative to the BDE-99 group on neither PND11 nor PND18.
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Fig. 6. Hepatic UDP-GT activity in offspring following perinatal exposure to Aroclor, Bromkal or BDE-99. Only samples from PND11 and PND18 were analyzed. Results are expressed as median ± interquartile range. Bars not sharing a letter within an age group are significantly different, p < 0.008; (b) denotes borderline significance, p < 0.0084.
3.5. Hepatic UDP-GT activity The UDP-GT activity was studied only in offspring at PND11 and PND18. A significant increase in enzyme activity was observed in the Aroclor group at both time-points (Fig. 6). The median activity in the Aroclor group was about twice that of the control group. At PND18, the increase in the Bromkal group was of borderline significance (p = 0.0084).
4. Discussion The current study enables a comparison of the effects following exposure to the commercial PBDE mixture, Bromkal 70-5DE, to effects following exposure to the PCB mixture, Aroclor 1254, and the pure pentabrominated (BDE-99) congener. Perinatal maternal exposure to Bromkal caused significantly reduced plasma T4 levels in the offspring, indicating that PBDEs possess endocrine disrupting properties. Exposure to the pure pentabrominated congener BDE-99, however, did not significantly reduce plasma T4 levels in offspring relative to controls, at any sampling occasion. This may be an indication that BDE-99 does not affect or is a less potent reducer of plasma T4 levels than the other main component in the Bromkal mixture, BDE-47. This is consistent with the finding, reported by Hallgren et al. (2001) that BDE-47 was a slightly more potent reducer of plasma TT4 and FT4 levels in mice than Bromkal 70-5DE, when administered at equimolar doses. The hypothesis that lower brominated congeners possess greater plasma T4 reducing abilities is also supported by the finding that BDE-mixtures, containing lower brominated BDE congeners, were shown to be more potent plasma T4 reducers in weanling rats, than mixtures including higher brominated congeners. This was seen following exposure to DE-71 (58% penta-BDE, 25% tetra-BDE), DE-79 (31% octa-BDE and 45% hepta-BDE) or DE-83R (98% deca-BDE) (Zhou et al., 2001).
Both BDE-99 and Bromkal exposure caused significant increases in the hepatic EROD activity in perinatally exposed offspring. In earlier studies an association between microsomal CYP induction and the reduction of circulating thyroid hormone level have been suggested (Hallgren et al., 2001; Hallgren and Darnerud, 2002). In a study by Meerts et al. (2000) three structurally different hydroxylated PBDE congeners were tested for their in vitro potency to displace T4 from TTR. The three hydroxylated compounds, resembling different thyroid hormones, showed a large difference in this respect. The results indicated that microsomal metabolism of PBDE will create hydroxylated products that are able to bind to TTR and affect serum T4 levels, but that the structure of the metabolite will be decisive for the binding affinity and subsequent T4 decrease. In the case of BDE-47 and BDE-99 an in vitro metabolism model demonstrated that phenobarbitalinduced microsomes (reflecting CYP2B activity) were most effective in producing extracts competing with T4 for TTR binding (Meerts et al., 2000). Consequently, CYP2B seems to produce BDE-47 and BDE-99 metabolites with TTR binding properties, which could result in decreased serum T4 levels (however not seen in our in vivo study after BDE-99 exposure). Aroclor exposure resulted in an induction of EROD and UDP-GT concurrent with a decrease in plasma T4. There was no induction of UDP-GT in offspring after perinatal Bromkal exposure at PND11, when there was a reduction in plasma T4. A limited induction of UDP-GT activities has previously been reported in rats following both PBDE and PCB exposure (Hallgren et al., 2001; Zhou et al., 2002) and in mice following exposure to PCBs (Craft et al., 2002) but not following exposure to PBDEs (Hallgren et al., 2001). This indicates that UDP-GT to some extent may be involved in thyroid hormone reduction in rats and mice, following exposure to PCBs, but only in rats following PBDE exposure. However, the limited enzyme induction suggests that other mechanisms than elevated UDP-GT activities play the major role in reducing T4 levels. The reduced plasma T4 levels and induced EROD activities seen in the offspring were significantly higher in the Aroclor group compared to the Bromkal group. This indicates that the PCB mixture is a more potent thyroid hormone disruptor and inducer of hepatic metabolic enzymes than the PBDE mixture. These results are consistent with previous findings in rats and mice (von Meyerinck et al., 1990; Hallgren et al., 2001) and may indicate that PCBs and PBDEs have different mechanisms of action or that PCBs merely possess properties rendering them greater thyroid hormone disrupting abilities. Exposure to Aroclor and Bromkal caused reduced plasma T4 levels in offspring. However, no significant plasma T4 reductions could be detected in the exposed dams. This indicates that offspring are more sensitive to the thyroid hormone disrupting properties of PBDEs and PCBs than pregnant and postweaning dams. These results are consistent with findings in rats reported by Zhou et al. (2002), in which a maternal
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exposure to 180 mg/kg b.w. DE-71 caused significantly reduced serum T4 levels in offspring but a maternal exposure to 350 mg/kg b.w. DE-71 was not sufficient to cause reduced serum T4 levels in the exposed dams. Also, the effect of PBDEs and PCBs on EROD activity was more pronounced in the offspring compared to the dams. The absence of maternal T4 reduction in our study may be due to less sensitive thyroid hormone regulating mechanisms in dams and reduced T4 levels may, therefore, have been detected if higher doses had been administered. The hypothyroxinemia and induced hepatic metabolic enzyme activity in the offspring, following exposure to Bromkal 70-5DE, is consistent with the results reported in a study performed by Zhou et al. (2002), in which reduced serum T4 levels and induced EROD and UDP-GT activities were seen in juvenile rats following daily gestational and lactational maternal exposure to the PBDE mixture, DE-71 (58% penta-BDE, 25% tetra-BDE). Reduced T4 levels and induced EROD activities have also been reported in adult mice, following exposure to DE-71 (Fowles et al., 1994) and in adult non-pregnant rats and mice following exposure to Bromkal 70-5DE and BDE-47 (Hallgren et al., 2001). Although, in our study, no maternal T4 reduction was seen in pregnant or lactating mice (following exposure to 131 versus 262 mg/kg b.w. Aroclor 1254), reduced T4 levels have been detected in pregnant and lactating rats exposed to 90 and 210 mg/kg Aroclor 1254 (Crofton et al., 2000). This may indicate that rats are more sensitive to PCB-induced T4 reduction than mice. A greater magnitude of T4 reduction in non-pregnant rats compared to non-pregnant mice was also seen by Hallgren et al. (2001) following exposure to Bromkal 70-5DE. Previous studies in rats indicate that PCB induced hypothyroidism reaches a maximum during the second and third postnatal week, not only when maternal exposure occurs both pre- and post-natally (Goldey et al., 1995; Goldey and Crofton, 1998) but also following only prenatal exposure (Morse et al., 1993). This coincides with the peak in exposure, which is via milk. In a cross-fostering study by Crofton et al. (2000) it was found that the lactational exposure, in contrast to the in utero exposure, was the major route of PCB exposure responsible for the reduction in offspring thyroid hormone levels. In the present study, we found that the reduction of offspring plasma T4 levels was of a greater magnitude relative to controls on PND11 than on PND18, although the offspring sampled on PND11 had been exposed to lower total doses than the offspring sampled on PND18. We also found that the T4 levels had returned to control levels by PND37. This indicates that the peak sensitivity to thyroid hormone disruption occurs in mice during the second postnatal week and that the sensitivity of this hormone system thereafter declines. It is of interest to note that the peak sensitivity to PBDE-induced behavioural alterations also has been shown to occur in mice during the second postnatal week (Eriksson et al., 1999, 2002).
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Since the first few postnatal weeks in rodents is the time period of peak sensitivity of auditory (Uziel et al., 1985a, 1985b; Hebert et al., 1985) and neuron development (Eriksson et al., 2001), hypothyroxinemia during this time period may have numerous adverse effects on the development of these structures (Crofton et al., 2000). Also, the effects on the auditory system (Goldey et al., 1995) and on behaviour (Eriksson et al., 2001; Branchi et al., 2002), have been shown to persist into adult life, although the observed hypothyroxinemia in itself is transient (Morse et al., 1996; Goldey and Crofton, 1998). In a study performed by Eriksson et al. (2001) the behavioural changes seen in mice following exposure to BDE-99 were actually seen to increase with age, being more pronounced on PND120 than on PND60. The transient thyroid hormone disruption during a critical phase of neonatal development may play a role in the developmental deficiencies not identifiable until later on in life. Contradictory to the behavioural effects observed after BDE-99 exposure, in the present study BDE99 had no effect on T4 homeostasis, which seems to speak against the proposed T4-mediated mechanism for effects on behaviour. This discrepancy may however be explained by differences in exposure regimens of the respective studies. For example, in the present study the fetus/pups were indirectly exposed via the dam/mother, whereas in the study by Eriksson et al. (2001) the pups were given the substance directly, in a single oral dose. In conclusion, the results of this study suggest that perinatal PBDE exposure disrupts thyroid hormone homeostasis in juvenile mice and that the offspring seems to be more susceptible to thyroid hormone reduction compared to their dams. BDE-99 seems to be a less potent reducer, or may not be a reducer at all, of plasma thyroid hormone levels than BDE47, the other major component of Bromkal 70-5DE. On an equimolar basis, PCBs possess greater thyroid hormone disrupting properties than PBDEs. Since both PCB and PBDE exposure caused induced hepatic cytochrome P450 activities and we know that certain hydroxylated PBDE metabolites have TTR binding potencies, it is possible that this may be one of the major mechanisms involved in reducing circulating thyroid hormone levels in exposed mice. However, PCB but not Bromkal induced UDP-GT, indicating that there may be different mechanisms of action for the two chemical mixtures or at least a difference in potency. Knowing that a functioning thyroid homeostasis is essential for normal juvenile development, the results of this study are of concern when evaluating potentially hazardous properties of the studied compounds.
Acknowledgements The authors would like to thank Agneta Bostr¨om and Elvy Netzel for excellent technical assistance. This work has been financed by grants from the Foundation for Strategic Environmental Research within the program, A New Strategy for Risk Assessment and Management of Chemicals.
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