* Canadian Wildlife Service, Burlington, Ontario, Canada, L7R 4A6; Canadian Wildlife Service, Burlington, Ontario, Canada, L7R 4A6;
National Wildlife Research Centre, Canadian Wildlife Service, Ottawa, Ontario, Canada K1A 0H3;
Patuxent Wildlife Research Center, Laurel, Maryland; ¶ Great Lakes Institute for Environmental Research, University of Windsor, Windsor, Ontario, Canada, N9B 3P4; and || Avian Science and Conservation Centre, McGill University, Ste Anne de Bellevue, Quebec, Canada, H9X 3V9
1 To whom correspondence may be addressed at Canadian Wildlife Service, 867 Lakeshore Rd., Burlington, Ontario, Canada, L7R 4A6. Fax: (905) 336-6434. E-mail: kim.fernie{at}ec.gc.ca.
Received March 2, 2005; accepted August 18, 2005
![]() |
ABSTRACT |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Key Words: PBDEs; thyroid; retinol; oxidative stress; glutathione; birds.
![]() |
INTRODUCTION |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Exposure of wild birds to PBDEs is of concern, since these and other structurally similar chemicals (e.g., polychlorinated biphenyls or PCBs) alter blood thyroid hormone homeostasis (Darnerud et al., 2001; McDonald, 2002
) and/or vitamin A stores (Hallgren et al. 2001
; Rolland, 2000
). Changes in these parameters may alter development, immuno-competence, reproductive success, and other physiological processes in birds. The mechanism by which PBDE exposure alters thyroid dynamics remains unknown, but PBDEs are structurally similar to thyroxine (T4). In addition, apparent metabolites of PBDEs such as hydroxylated PBDEs (OH-PBDEs), have been shown to have competitive binding affinity relative to T4 with human transthyretin (TTR) (Hakk and Letcher, 2003
; Meerts et al., 2000
). TTR is a plasma protein crucial for the transport and controlled distribution of thyroid hormones and vitamin A. T4 is known to have low binding affinity for TTR (Chang et al., 1999
) and a short half-life (Astier, 1980
) in birds, which may render them susceptible to PBDE-induced alterations in circulating thyroid hormone concentrations. Studies with laboratory rodents have indicated that some PBDE congeners affect thyroid hormone transport and metabolism (Zhou et al., 2001
, 2002
) as well as thyroid hormone and vitamin A levels (Hallgren et al., 2001
).
In addition to thyroid toxicity, there is mounting evidence that alterations in vitamin A homeostasis in animals are associated with exposure to persistent organohalogenated compounds, including PCBs (Chen et al., 1992; Palace et al., 1997
), polybrominated biphenyls (Chen et al., 1992
), and polychlorinated dibenzo-p-dioxins and dibenzofurans (Spear et al., 1990
). Vitamin A stores may be depleted by induction of hepatic biotransformation enzymes that produce more easily excreted vitamin A metabolites (Palace et al., 1997
). In addition, toxicant-induced mixed-function oxidase enzymes and redox recycling can also produce reactive oxygen species that damage lipid membranes, cellular proteins, and DNA in birds (Hoffman and Heinz, 1998
; Hoffman et al., 1998
). Since retinol serves as an effective biological antioxidant that protects against oxyradicals, it can become depleted when animals are exposed to organic contaminants (Palace et al., 1997
). Consequently, exposure to persistent contaminants such as PBDEs may induce oxidative stress and result in changes in antioxidant stores and enzymatic activity. Cellular defense mechanisms such as glutathione metabolism, thiols, and lipid peroxidation serve as sensitive measures of oxidative stress when animals are exposed to mercury and selenium (Hoffman et al., 1998
) and PCBs (Palace et al., 1997
).
Despite the potential risks to wildlife exposed to increasing levels of PBDEs in the environment, there has yet to be a report that addresses PBDE-induced alterations in thyroid function, vitamin A dynamics, and/or oxidative stress in any wildlife species. In vertebrates, early developmental processes, growth, metabolic regulation, thermoregulation, and reproduction are dependent, in part, upon adequate and constant stores of thyroid hormones and vitamin A (e.g., Oppenheimer et al., 1995). Oxidative stress contributes to cancers, neurodegenerative diseases including immune and brain disorders (e.g., Parkinson's disease), and aging (von Schantz et al. 1999
and references therein). We have shown previously that environmentally relevant concentrations of PBDE congeners elicited changes in growth and food consumption (Fernie et al., 2006), as well as immune function and structure (Fernie et al., 2005), in nestling American kestrels. The objectives of this study are to further determine if such BDE exposure affects the thyroid system, retinol concentrations, and oxidative stress in the same cohort of exposed kestrels.
![]() |
MATERIAL AND METHODS |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
The dosing of the kestrel eggs mimicked the PBDE concentrations and congener profile measured in herring gull eggs collected from numerous sites throughout the Great Lakes (range 1981400 ppb ww) (Norstrom et al., 2002). The most prevalent PBDE congeners in these eggs were, in descending order, BDE-47, -99, -100, and -153. A concentration of 1500 ng/g
PBDEs dissolved in safflower oil, or safflower oil only, was injected into the air cell of fertile eggs at approximately 19 days of incubation, nearly half-way through the 28-day incubation period. The actual dosage used was 2.1 µg/µl of oil with the measured BDE proportions as follows: 56.4% of BDE-47; 27.2% of BDE-99; 24.8% of BDE-100; and 0.6% of BDE-153. Unlabelled standards for each BDE congener (in nonane) at a chemical purity of more than 98% were purchased from Cambridge Isotope Laboratories (Andover, MA). Post-hatch exposure occurred by oral gavage prior to the first daily meal, with the dosage determined by nestling age (hatch through to day 29), and the
PBDE concentrations approximating those measured in Great Lakes lake trout (100 ng/g wet weight) (H. Stapleton, personal communication). The
PBDE concentrations measured in the oil administered to the nestlings was 35.0 ± 0.8 ng/µl with a mean daily
PBDE dosage of 15.6 ± 0.3 ng/g bird/day.
Husbandry of kestrel nestlings.
Kestrels were fed ad libitum frozen-thawed day-old cockerels. Newly hatched kestrel chicks were individually color-marked until they were banded at 10 days of age. Hatchlings were transferred to an open-air brooder where they were kept in groups of five (the average size of wild broods (Smallwood and Bird, 2002)), receiving the same exposure treatment. At 6 days of age, the birds were transferred to cardboard boxes of approximately the same size as nest boxes used by wild kestrels.
Immediately prior to euthanization by cervical dislocation at 36 days of age, blood samples were taken by jugular venipuncture, centrifuged, and the plasma aliquoted for plasma T4, T3, and vitamin A determination. The liver was completely removed and dissected into separate aliquots for retinol, retinyl palmitate, and oxidative stress analysis. In addition, the brain, kidney, and thyroid glands were dissected from the carcass. All tissues were immediately stored in vapor-phase liquid nitrogen.
Thyroid hormones.
Total triiodothyronine (TT3) and total thyroxine (TT4) levels in plasma samples were determined by radioimmunoassay (RIA) using Coat-A-Count canine TT3 and TT4 kits (Diagnostic Products Corp. Ca. (DPC)). The DPC 125I-T3/T4 solid phase radioimmunoassay kits use tubes coated with monoclonal T3 or T4 antibody. After sample incubation, the bound and free fractions were separated, and the radioactivity was counted during 1-min intervals using a Canberra-Packard model E-5002 gamma counter. Sample results were calculated from standard curves ranging from 0.15 to 6.45 ng/ml and 2.5 to 60.0 ng/ml for TT3 and TT4 respectively.
Thyroid histology.
Thyroids were removed fresh and fixed in 10% buffered formalin. Tissues were placed in coded plastic cassettes to facilitate blind histological evaluation and embedded in paraffin. Serial sections (5 µm) were cut then stained with eosin and haematoxylin. Quantitative histological evaluation of thyroid sections was performed using a Zeiss light microscope (400x magnification) with a camera and digital image analyzing software (AxioVision 3.0, Carl Zeiss Vision, München-Hallbermoos, Germany). There was no follicular damage in any of the thyroid glands despite the use of cervical dislocation; all follicles were round and did not show any distortion or crushing. As a quantitative index of thyroid gland activity, epithelial cell height was measured in each individual follicle (n = 50 follicles per gland) lying along the vertical axis of the midline of each thyroid section (n = 2 thyroid lobes per bird) to ensure that both peripheral and interior follicles were evaluated using a standardized methodology. In each follicle, the cell height was measured at 90°, 180°, 270°, and 360° starting from an arbitrary point. The means of all measurements were calculated for each follicle and averaged for each kestrel thyroid gland.
Qualitative assessment of thyroid follicles was also performed. The presence of colloid-containing follicles was expressed as a percentage of all follicles evaluated. Colloid vacuolation was recorded as "absent," "mild" when few vacuoles were present, "moderate" when vacuoles were present around the entire periphery of the lumen, or "marked" when the colloid was filled with vacuoles. Similarly, follicular epithelial cell hyperplasia was rated as "absent," "mild" (few foci), or "marked" when the hyperplastic epithelium was multifocal and dissected the thyroid interstitium.
Plasma and hepatic vitamin A.
To quantify plasma retinol, retinyl acetate was added to an aliquot of plasma (50100 µl) as an internal standard. The retinol-protein complex was then dissociated with acetonitrile, and the retinol extracted with successive volumes of hexane. The separation of the organic and aqueous phases was achieved by centrifugation. The organic phases were combined and evaporated to dryness with nitrogen. The residues were redissolved in methanol, filtered, and analyzed by HPLC. The separation was achieved in less than 6 min with a 15-cm-long, 5-µm ODS Zorbax column using solvent programming. Retinol was detected with a UV detector set at a maximum absorption wavelength of 326 nm.
A 100-mg sample of liver was dehydrated with anhydrous sodium sulphate and transferred into an amber vial. Vitamin A compounds (retinol and retinyl palmitate) were extracted with 5 ml dichloromethane:methanol (1:9) after the addition of the retinyl acetate internal standard. After centrifugation, an aliquot of the supernatant was filtered, and 50 µl were analyzed by nonaqueous reverse-phase HPLC. The separation was achieved in less than 10 min with a 15-cm-long, 5-µm ODS Zorbax column using solvent and flow programming. The vitamin A compounds were detected with a UV detector set at 326 nm.
Biomarkers of oxidative stress, glutathione metabolism, and lipid peroxidation.
A 1-g portion of each liver was homogenized (1:10 w/v) in ice-cold 1.15% KCL0.01 M Na, potassium phosphate buffer (pH 7.4). The homogenate was centrifuged at 10,000 x g for 20 min at 4°C, and the supernatant was used for enzymatic assays related to glutathione metabolism and antioxidant activity. Following the methods described in Hoffman et al. (1998), oxidized glutathione (GSSG), reduced glutathione (GSH), and total sulfhydryls (total SH), and protein-bound sulfhydryls (PBSH) (calculated as the difference between total SH and GSH), were measured, since they are related to glutathione metabolism and antioxidant activity. In addition, thiobarbituric acid-reactive substances (TBARS) were measured as estimates of hepatic lipid peroxidation.
PBDE analysis of kestrel homogenates.
Immediately following euthanization and dissection, each carcass was plucked, and the feet removed prior to processing and homogenization. The remaining carcass was then homogenized into a uniform sample using a stainless steel waring blender. A subsample of 4 g of homogenate was used for PBDE analysis. The homogenate was ground by mortar and pestle with 24 g Na2SO4 and transferred to a glass extraction column (60 x 2.5 cm i.d.) containing 10 g Na2SO4 and 100 ml dichloromethane (DCM):hexane (50:50% v/v). The column was subsequently spiked with 2.1 ng of BDE-30 for use as a recovery standard. BDE-30 was not detectable in unspiked samples. After 1 h, the column was eluted with an additional 250 ml DCM:hexane (50:50% v/v), and the combined extracts roto-evaporated to a volume of 10 ml. Ten percent by weight of the extracts were removed for lipid determination.
The remaining extracts were cleaned by gel permeation chromatography (GPC) by adding extracts to a 50 cm x 22 cm i.d. glass column wet packed with 50 g Biobeads S-X3 (Bio-Rad Laboratories, CA, USA) in DCM:hexane (50:50% v/v). The sample was eluted from the GPC column with 300 ml DCM:hexane (50:50% v/v), the first 125 ml (containing lipids and biogenic material) was discarded, and the second fraction (containing PBDEs) was concentrated to a volume of 2 ml under reduced pressure. Further cleaning of the sample was performed by adding the concentrated extract to a 35 cm x 1 cm i.d. glass column containing 8 g of 1.2% deactivated Florisil wet packed in hexane. The sample was eluted from the column with 38 ml hexane followed by 34 ml DCM:hexane (15:85% v/v). The combined extracts were supplemented with 2 ml isooctane and roto-evaporated to a final volume of 1 ml.
Chemical analysis was performed using an Agilent (USA) 6890 GC equipped with a 63Ni µ-ECD detector and Agilent 7673 automated injector. The GC was fitted with a fused silica HP-5 column (5% phenyl methyl siloxane, 30 m (length) x 250 µm (internal diameter), 0.25 µm film thickness). The carrier gas was ultra-high purity helium, and the ECD makeup gas was 5% methane95% argon. All injections were 1 µl in volume and made in the splitless mode. The temperature ramping program was as follows: initial 80°C (held for 2 min) to 290°C at a rate of 10°C/min (held at 290°C at 15 min).
Quantification of individual congeners was performed using an external standard approach containing BDE-47, -100, -99, -153, -138 and -183. The standards were synthesized (Marsh et al., 2003) and generously supplied by Dr. Aake Bergman (Department of Environmental Chemistry, Stockholm University, Stockholm, Sweden). BDE-154 was obtained from Wellington Scientific (ON, Canada). Among all samples analyzed, the mean recoveries of the BDE-30 spike were 93 ± 9% (mean ± SD). The mECD linear response range was determined using seven serial dilutions of the PBDE standard mixture, where the concentrations ranged from 10 to 1000 pg/µl. Linear regression of the calibration curves for individual BDEs gave coefficients of determination (R2 > 0.99). Based on the BDE-30 recovery standard, the method limit of quantification (MLOQs) for µECD based on a signal/noise ratio of 10 was 0.01 ng/g (ww). Representative sample and blank fractions were also analyzed by GC-electron-capture negative ionization-MS in the selected ion monitoring mode (79Br and 81Br ions) to confirm BDE congener identities.
For every batch of five samples, a method blank (containing 2 g sodium sulfate) and an in-house reference tissue homogenate (Detroit River homogenized fish sample) were extracted to monitor laboratory performance. Small quantities of (mainly) BDE-47 or interfering peaks were encountered in blanks, but were generally between two to three orders of magnitude lower than BDEs measured in treatment birds. In a few cases the µECD response of BDE-47 in the blanks approached the response in the PBDE fractions from the control animals. Therefore, all samples were blank corrected by subtracting, e.g., BDE-47 response in blank relative to its corresponding five sample batch. A value of nondetection was established for cases where the blank BDE area was greater than 30% of the sample peak area for the equivalent congener.
The four BDE standards were analyzed for 11 brominated dioxin and furan congeners by Axys Analytical Services (Sidney, British Columbia, Canada). Samples were spiked with 13C/12C-labelled analogues of brominated dioxins and furans and analyzed by high resolution GC/MS. Concentrations were determined using internal standards quantified against the 13C/12C-labelled standards added to the samples.
Statistical analysis.
Statistical analysis was performed using SASg. Data failed to meet the conditions of normality and homogeneity of variance, so were analyzed using the nonparametric Wilcoxon two-sample one-tail test (Zar, 1996). Differences between the two groups were also assessed for biological importance using effect size (ES) analysis (Cohen, 1988
; Hayes, 1987
). Since the standard deviations of each group were not equal, we calculated the ES by determining the absolute value of the difference between the means of the two groups and dividing that result by the result produced when taking the square-root of one half of the sum of the squared standard deviations of each group. The ES increases as the difference between population means increases, and decreases as the variability in response increases (Hayes, 1987
). Spearman's correlation analysis was used to identify specific, predetermined associations. In order to increase statistical power and reduce the chances of making a Type II error, statistical significance was considered to be p < 0.10 (Zar, 1996
). Means ± standard errors are presented.
![]() |
RESULTS |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Polybrominated Diphenyl Ether Concentrations in Kestrel Chicks
Although the treatment birds were exposed to four PBDE congeners, an additional two congeners, BDE-138 and BDE-183, were also detected in the PBDE-dosed birds. The corporal PBDE concentrations of the PBDE-dosed birds were 86.09 ± 8.40 ng/g ww (796.18 ± 119.52 ng/g lipid weight) while those of the controls birds were 0.72 ± 0.15 ng/g ww (5.83 ± 1.15 lipid weight). These findings are discussed in detail in Fernie et al. (2006).
Thyroid Hormone and Vitamin A Concentrations
There were no statistical differences in the plasma T4 concentrations of the PBDE-exposed and control kestrels, although the lower mean concentrations of the PBDE-exposed birds may be biologically important (ES = 0.34) (Table 1). Plasma T4 concentrations were negatively correlated with BDE-47 (n = 11; r = 0.71; p = 0.02) (Fig. 1A), BDE-99 (n = 13; r = 0.65; p = 0.02) (Fig. 1B), and BDE-100 (n = 16; r = 0.46; p = 0.07), but not the PBDE concentrations. However, there were no differences in T3 concentrations between groups (Table 1), and T3 levels were not associated with concentrations of PBDE congeners. Furthermore, the PBDE exposure had no effect on the morphology or pathology of the thyroid gland.
|
|
|
|
![]() |
DISCUSSION |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Exposure to persistent organohalogens other than PBDEs is associated with changes in thyroid hormone and retinol concentrations in free-ranging vertebrates (Rolland, 2000). Comparisons with PBDE-related exposure effects on other vertebrates are difficult due to the limited number of published studies. However, the lower plasma T4 concentrations, particularly with the increasing concentrations of BDE-47, -99, and -100, and the maintenance of T3 levels in the PBDE-exposed kestrels, is consistent with studies involving the exposure of laboratory rodents to PBDEs. Exposure of mice and rats to the commercial pentabromodiphenyl ether mixture, DE-71, DE-79, the PBDE technical mixture Bromkal 705, or BDE-47 decreased plasma total and free T4 concentrations but not T3 concentrations (Fowles et al., 1994
; Hallgren et al., 2001
; Zhou et al., 2001
). Gestational rats were exposed to DE-71 at varying doses up to 30 mg/kg/day until postnatal day 21; dams, their fetuses, and offspring all showed dose-dependent reductions in plasma T4 concentrations (Zhou et al., 2002
). While these changes in mammalian T4 concentrations may have been influenced by the presence of low levels of dioxin-like compounds within the technical PBDE mixtures, in this study, the kestrels were exposed to mixtures of individual BDE congener standards of high chemical purity (>98%) and with evidence of background exposure to the octa-BDE congener, BDE-183. Furthermore, in these previous mammalian studies the doses used (0.8030 µg/g) exceeded the dose used in the present kestrel study (15.6 ng/g) by three orders of magnitude. In comparison to PBDE-related thyroid system effects in mammalian studies, our results illustrate that the kestrel, and perhaps the avian thyroid system in general, is sensitive to modulation as a result of PBDE exposure in vivo.
Several mechanisms may be involved with the changes in the thyroid system and retinol concentrations of the PBDE-exposed kestrels in this study, which warrant further investigation. The reduction in plasma T4 concentrations but not plasma T3 concentrations may be the result of increased conversion of T4 to T3 through hepatic mono-deiodinase enzymes, induction of sulfotransferase enzymes, and/or increased glucuronidation of T4 (Barter and Klaassen, 1994; Morse et al., 1993
) as reported in rodents exposed to PBDEs (Zhou et al., 2001
, 2002
). In addition, competitive displacement of circulating T4 and retinol from TTR by PBDE congeners or OH-PBDE metabolites may also have contributed to their reductions in the kestrels. TTR is highly conserved in all vertebrates (Chang et al., 1999
). Thyroid hormone-like OH-PBDE congeners have high T4-human TTR competitive binding potency in vitro, possibly causing a conformational change in the TTR protein (Meerts et al., 2000
), which may also inhibit the formation of the serum retinol transport protein complex (Brouwer et al., 1985
). Subsequently, the excess unbound T4 and retinol would be more susceptible to hepatic catabolism, resulting in a greater clearance rate and a decrease in circulating concentrations. The high avidity for TTR by the lower brominated BDE metabolites causes concern for thyroid homeostasis in wild raptors, as lower brominated PBDEs are detected in these species (Lindberg et al., 2004
).
In the present study, the lack of change in thyroid follicular epithelial cell size indicates that glandular activity was unaffected by PBDE exposure, which initially may seem contrary to the thyroid hormone results reported above. Decreases in circulating T4 concentrations are assumed to increase secretions of thyroid stimulating hormone from the pituitary gland, with subsequent compensatory hypertrophy and/or hyperplasia of the thyroid gland (Capen et al., 2002). However, thyroid stimulating hormone concentrations did not increase as a result of PBDE exposure (Darnerud and Sinjari, 1996
; Hallgren et al., 2001
; Zhou et al., 2001
). The reductions in plasma T4 concentrations may be due to hepatic microsomal enzymes and/or the receptor at the tissue level being occupied by the PBDEs/HO-PBDEs.
Our results also indicate a change in retinol tissue concentrations as a function of PBDE exposure. The PBDE-exposed kestrels had lower plasma retinol and hepatic retinol and retinyl palmitate concentrations, consistent with other studies reporting reductions in vitamin A analogue concentrations in birds exposed to various organohalogen compounds including dioxins and furans (Spear et al., 1990). Reductions of plasma and hepatic vitamin A levels have also been observed in laboratory rodents following exposure to Bromkal 705DE but not BDE-47 (Hallgren et al., 2001
). The suppression of the various vitamin A analogues is more likely a function of PBDE exposure, since higher concentrations of the same three BDE congeners, BDE-47, -100, and -99, associated with suppressed T4 concentrations, were also associated with suppressed plasma and hepatic retinol concentrations.
Retinol is physiologically essential. It is stored as a fatty acid ester, retinyl palmitate, in hepatic stellate cells and is released in free form only on demand and, then, only bound to TTR in the blood. Exposure to the PBDEs likely interfered with the retinoid dynamics in the kestrels, resulting in the inability to maintain or replace declining retinol stores. The displacement of retinol from TTR by PBDEs, in concert with increased uridine diphosphate-glucuronosyl transferase (UDPGT) activity seen in laboratory animals exposed to PBDEs (Zhou et al., 2001, 2002
), would greatly increase the clearance of conjugated retinol, rendering it unavailable for reabsorption by the kidney. With increasing clearance of conjugated retinol, the hepatic vitamin A stores would be mobilized to maintain homeostatic balance. This mobilization may have involved: increased induction of hepatic phase I and phase II enzymes resulting in greater metabolism and clearance of retinol (Hallgren et al., 2001
); the loss or transformation of stellate storage cells in the liver; and changes in hepatic vitamin A enzyme activity involving retinol esterification (Chen et al., 1992
). Induction of phase I and phase II liver detoxification enzymes, proposed as mechanisms of retinoid reduction, occurred in rodents exposed to technical PBDE products or brominated mixtures (Fowles et al., 1994
; Hallgren et al., 2001
; von Meyerinck et al., 1990).
Retinol also is an antioxidant that reduces oxidative stress (Käkelä et al., 2003; Palace et al., 1997
). PBDE-related oxidative stress seemed apparent in the kestrels with the increased ratio of GSSG to GSH, particularly in the females which also showed marginal signs of lipid peroxidation. Two of the BDE congeners, BDE-99 and especially BDE-183, may have been primarily involved, as they were the only congeners associated with the hepatic biochemical endpoints. Increasing concentrations of BDE-99 (males only) and -183 were associated with increasing GSH and thiol concentrations, respectively, which would increase protection against oxidative stress and lipid peroxidation, hence possibly explaining the negative association between BDE-99 concentrations and hepatic TBARS. Consistent with this hypothesis of increased protection against oxidative stress is the increased synthesis of oxidized GSSG in the PBDE-exposed birds. Potentially, the PBDE exposure was beginning to exceed the capabilities of the antioxidants involving retinol, since the retinol:retinyl palmitate concentrations declined as lipid peroxidation increased, and both retinol and hepatic retinyl palmitate declined in the exposed birds. Unfortunately, the compensation against the PBDE-induced oxidative stress was not successful in the kestrels, particularly females, despite their carotenoid-rich diet; this raises concerns for wild birds exposed to PBDEs. Diet is the primary source of antioxidants, and wild birds generally do not consume a diet as rich in antioxidants and carotenoids as that of our captive kestrels in which oxidative stress was present.
In conclusion, this study demonstrates that exposure of kestrels to environmentally relevant PBDEs alters thyroid and retinol concentrations and results in hepatic oxidative stress, marginal lipid peroxidation, and ensuing changes in glutathione metabolism. Circulating T4 and retinol concentrations, and hepatic stores of retinol, were negatively associated with individual PBDE congeners. The mechanisms of action remain unclear. Since HO-PBDE metabolites have been shown to competitively bind with human TTR and/or alter hepatic enzymatic activity, HO-PBDEs may be mediating changes in the plasma T4 and retinol concentrations in the present kestrels, thereby requiring future assessments. Oxidative stress and marginal lipid peroxidation were induced by PBDE exposure, the latter in only the females, with BDE-183 and -99 associated with changes in the thiols, reduced glutathione, and TBARs. The effects observed in captive kestrels in this study raise concerns for free-ranging birds and other wildlife exposed to PBDEs.
![]() |
NOTES |
---|
![]() |
ACKNOWLEDGMENTS |
---|
![]() |
REFERENCES |
---|
![]() ![]() ![]() ![]() ![]() ![]() ![]() |
---|
Barter, R. A., and Klaassen, C. D. (1994). Reduction of thyroid hormone levels and alteration of thyroid function by four representative UDP-glucuronosyltransferase inducers in rats. Toxicol. Appl. Pharmacol. 128, 917.[CrossRef][ISI][Medline]
Birnbaum, L. S., Staskal, D. F., and Diliberto, J. J. (2003). Health effects of polybrominated dibenzo-p-dioxins (PBDDs) and dibenzofurans (PBDFs). Environ. Int. 29, 855860.[CrossRef][ISI][Medline]
Brouwer, A., van den Berg, K. J., and Kukler, A. (1985). Time and dose responses of the reduction in retinoid concentrations in C57BL/Rij and DBA/2 mice induced by 3,4,3',4'- tetrachlorobiphenyl. Toxicol. Appl. Pharmacol. 78, 180189.[CrossRef][ISI][Medline]
Capen, C. C., DeLellis, R. A., and Yarrington, J. T. (2002). Endocrine system. In Handbook of Toxicologic Pathology (W. M. Haschek, C. G. Rousseau, and M. A. Waalig, Eds), pp. 681783. Academic Press, San Diego.
Chang, L., Munro, S. L. A., Richardson, S. J., and Schreiber, G. (1999). Evolution of thyroid hormone binding by transthyretins in birds and mammals. Eur. J. Biochem. 259, 534542.[CrossRef][ISI][Medline]
Chen, L. C., Berberian, I., Koch, B., Mercier, M., Azais-Braesco, V., Glauert, H. P., Chow, C. K., and Robertson, L. W. (1992). Polychlorinated and polybrominated biphenyl congeners and retinoid levels in rat tissues: Structure-activity relationships. Toxicol. Appl. Pharmacol. 114, 4755.[CrossRef][ISI][Medline]
Cohen, J. (1988). Statistical Power Analysis for the Behavioural Sciences. Lawrence Erlbaum Associates, New Jersey.
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, 4968.[ISI][Medline]
Darnerud, P. O., and Sinjari, T. (1996). Effects of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) on thyroxine and TSH blood levels in rats and mice. Organohalogen Comp. 29, 316319.
Fernie, K. J., Mayne, G., Shutt, J. L., Pekarik, C., Grasman, K. A., Letcher, R. J., and Drouillard, K. (2005). Evidence of immunomodulation in nestling American kestrels (Falco sparverius) exposed to environmentally relevant PBDEs. Environ. Pollut. 138, 485493.[ISI][Medline]
Fernie, K. J., Shutt, J. L., Ritchie I. J., Letcher, R. J., Drouillard, K., and Bird, D. M. (2006). Effects of exposure to environmentally relevant polybrominated diphenyl ethers on survival and growth of American kestrel (Falco sparverius) nestlings. J. Toxicol. Environ. Health.
Fowles, J. R., Fairbrother, A., Baecher-Steppan, L., and Kerkvliet, N. I. (1994). Immunologic and endocrine effects of the flame-retardant pentabromondiphenyl ether (DE-71) in C57BL/6J mice. Toxicology 86, 4961.[CrossRef][ISI][Medline]
Fox, G. A. (1993). What have biomarkers told us about the effects of contaminants on the health of fish-eating birds in the Great Lakes? The theory and a literature review. J. Great Lakes Res. 19, 722736.[ISI]
Hakk, H., and Letcher, R. J. (2003). Metabolism in the toxicokinetics and fate of brominated flame retardantsA review. Environ. Int. 29, 801826.[CrossRef][ISI][Medline]
Hallgren, S., Sinjari, T., Håkansson, H., and Darnerud, P. O. (2001). Effects of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) on thyroid hormone and vitamin A levels in rats and mice. Arch. Toxicol. 75, 200208.[CrossRef][ISI][Medline]
Hayes, J. P. (1987). The positive approach to negative results in toxicology studies. Ecotoxicol. Environ. Saf. 14, 7377.[CrossRef][ISI][Medline]
Hoffman, D. J., and Heinz, G. H. (1998). Effects of mercury and selenium on glutathione metabolism and oxdative stress in mallard ducks. Environ. Toxicol. Chem. 17, 161166.[CrossRef][ISI]
Hoffman, D. J., Ohlendorf, H. M., Marn, C. M., and Pendleton, G. W. (1998). Association of mercury and selenium with altered glutathione metabolism and oxidative stress in diving ducks from the San Francisco Bay region, USA. Environ. Toxicol. Chem. 17, 167172.[CrossRef][ISI]
Käkelä, R., Käkelä, A., Hyvärinen, H., and Nieminen. P. (2003). Effects of aroclor 1242 and different fish-based diets on vitamins A1 (retinol) and A2 (3,4-didehydroretinol), and their fatty acyl esters in mink plasma. Environ. Res. 91, 104112.[CrossRef][ISI][Medline]
Lindberg, P., Sellstrom, U., Haggberg, L., and de Wit, C. A. (2004). Higher brominated diphenyl ethers and hexabromocyclododecane found in eggs of peregrine falcons (Falco peregrinus) breeding in Sweden. Environ. Sci. Technol. 38, 9396.[CrossRef][ISI][Medline]
Marsh, G., Stenutz, R., and Bergman, Å. (2003). Synthesis of hydroxylated and methoxylated polybrominated diphenyl ethers natural products and potential polybrominated diphenyl ether metabolites. Eur. J. Org. Chem. 14, 25662576.[CrossRef]
McDonald, T. A. (2002). A perspective on the potential health risks of PBDEs. Chemosphere 46, 745755.[CrossRef][ISI][Medline]
Meerts, I. A. T., van Zanden, J. J., Luijks, E. A. C., van Leeuwen-Bol, A., Marsh, G., Jokobsson E., Bergamna, A., and Brouwer, A. (2000). Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicol. Sci. 56, 95104.
Morse, D. C., Groen D., Veerman, M., van Amerongen, C. J., Koëter, H. B. W. M., Smits van Prooije, A. E., Visser, T. J., Koeman, J. H., and Brouwer, A. (1993). Interference of polychlorinated biphenyls in hepatic and brain thyroid hormone metabolism in fetal and neonatal rats. Toxicol. Appl. Pharmacol. 122, 2733.[CrossRef][ISI][Medline]
Norstrom, R. J., Simon, M., Moisey, J., Wakeford, B., and Weseloh, D. V. C. (2002). Geographical distribution (2000) and temporal trends (19812000) if brominated diphenyl ethers in Great Lakes herring gull eggs. Environ. Sci. Technol. 36, 47834789.[CrossRef][ISI][Medline]
Olfert, R. R., Cross, B. M., and McWilliam, A. A. (1993). Guide to the Care and Use of Experimental Animals, Vol. 1, 2nd ed. Canadian Council on Animal Care, Ottawa, ON, Canada.
Oppenheimer, J. H., Schwartz, H. L., and Strait, K. A. (1995). An integrated view of thyroid hormone action in vivo. In Molecular Endocrinology: Basic Concepts and Clinical Correlations (B. D. Weintrab, Ed.), pp. 249267. Raven, New York.
Palace, V. P., Klaverkamp, J. F., Baron, C. L., and Brown, S. B. (1997). Metabolism of 3H-retinol by lake trout (Salvelinus namoycush) pre-exposed to 3,3,3'4,4',5-pentachlorobiphenyl (PCB 126). Aquat. Toxicol. 39, 321332.[CrossRef][ISI]
Rolland, R. M. (2000). A review of chemically-induced alterations in thyroid and vitamin A status from field studies of wildlife and fish. J. Wildl. Dis. 36, 615635.
Sellström, U., Bignert, A., Kierkegaard, A., Häggberg, L., de Wit, C. A., Olsson, M., and Jansson, B. (2003). Temporal trend studies on tetra- and pentabrominated diphenyl ethers and hexabromocyclodedecane in Guillemot egg from the Baltic Sea. Environ. Sci. Technol. 37, 54965501.[CrossRef][ISI][Medline]
Smallwood, J. A., and Bird, D. M. (2002). American kestrel (Falco sparverius). In The Birds of North America, No. 602 (A. Poole, and F. Gill, Eds.), 32 pp. Academy of Natural Sciences and The American Ornithologists' Union, Washington, DC.
Spear, P. A., Norstrom, R. J., and Moon, T. W. (1990). Yolk retinoids (Vitamin A) in eggs of the herring gull and correlations with polychlorinated dibenzo-p-dioxins and dibenzofurans. Environ. Toxicol. Chem. 9, 10531061.[ISI]
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. Toxicology 61, 259274.[CrossRef][ISI][Medline]
von Schantz, T., Bensch, S., Grahn, M., Hasselquist, D., and Wittzell, H. (1999). Good genes, oxidative stress and condition-dependent sexual signals. Proc. R. Soc. London. B 266, 112.[CrossRef][ISI][Medline]
Zar, J. H. (1996). Biostatistical Analysis, 3rd ed. Prentice Hall, Englewood Cliff, New Jersey.
Zhou, T., Ross, D. G., deVito, M. J., and Crofton, K. M. (2001). Effects of short-term in vivo exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats. Toxicol. Sci. 61, 7682.
Zhou, T., Taylor, M. M., DeVito, M. J., and Crofton, K. M., (2002). Developmental exposure to brominated diphenyl ethers results in thyroid hormone disruption. Toxicol. Sci. 66, 105116.
|