THE ADVERSE EFFECTS OF 1,2-BIS(TRIBROMOPHENOXY)ETHANE IN MINK (MUSTELA VISON) By Stephanie Smith-Edwards A THESIS Submitted to Michigan State University in partial fulfillment of the requirements for the degree of Animal Science – Master of Science 2013     ABSTRACT THE ADVERSE EFFECTS OF 1,2-BIS(TRIBROMOPHENOXY)ETHANE IN MINK (MUSTELA VISON) By Stephanie Smith-Edwards Brominated flame retardants (BFRs) have been incorporated into a variety of consumer products for several years. Demonstration of BFRs in the environment, wildlife and humans has prompted concern for these emerging contaminants. Two of the commercial polybrominated diphenyl ether (PBDE) BFRs (octa-BDE and penta-BDE) are no longer being produced because of environmental concerns. As a result, the production and use of non-PBDE BFR alternatives, such as 1,2-bis(2,4,6-tribromophenoxy) ethane (BTBPE), have increased. It was of interest to determine the sensitivity of mink, a sentinel wildlife species, to BTBPE, which has been detected in the environment. Forty adult female mink were fed one of four diets containing 0, 0.014, 0.13 or 2.3 mg BTBPE/kg feed two months prior to breeding. Females were bred to untreated males. At whelping and at 3 and 6 weeks of age, kits were counted and weighed. At 6 weeks of age, six offspring from each treatment group, as well as the adult females, were necropsied. Samples of plasma, liver, fat, lungs, and feces were processed for chemical analysis and thyroids were processed for histological assessment. Ten offspring per group were maintained on their respective treatments through 7 months of age at which time the juvenile mink were necropsied and tissues processed as described above. The results of this study indicate that exposure to BTBPE at dietary concentrations up to 2.3 mg/kg feed had no effect on the reproductive performance of mink and the survivability and growth of their offspring.     DEDICATION To my parents and friends, my children Jayla, Julian and Stephon, and my husband Julius who has sustained me through this endeavor. iii   ACKNOWLEDGEMENTS My thesis research would not have been possible without the support and guidance of many kind people around me, too numerous to mention here. First and foremost I would like to thank my academic advisor, Dr. Steve Bursian. Dr Bursian has offered me invaluable assistance, patience, and guidance through this academic journey. I would like to give a special thanks to my Dr. Scott Fitzgerald, for his valuable expertise in histology, granting me an opportunity to do an independent histology study, and reviewing my thesis paper. I appreciate the patience and support of committee members Dr. Karen Chou, and Dr. John Newsted through my Master’s thesis process. I owe my deepest gratitude to the animal science staff, Mr. Angelo Napolitano, and Mrs. Jane Link for their support in data collection and analysis. Special thanks to Mrs. Link in looking closely at the final version of my thesis writing editing and offering suggestions for improvement. Also, the Canadian Wildlife Services which provided the necessary funding for this research and assistance with sample analysis. I appreciate the support of my colleagues at the Office of Environmental Health and Safety, Biosafety staff, who inspired me to finish, and listened to presentations, read abstracts, and offered thesis revision suggestions. I would also like to thank Mrs. Rachel Ellick, Valencia Moses, and Mrs. Cona Marshall, my fellow graduate students in their support and encouragement through my thesis writing. Lastly I would like to thank my family and friends, without their love and support, none of this would have been possible. iv   TABLE OF CONTENTS LIST OF TABLES ......................................................................................................................... vi LIST OF FIGURES ..................................................................................................................... viii LIST OF ABBREVIATIONS ........................................................................................................ ix CHAPTER 1 ................................................................................................................................... 1 LITERATURE REVIEW ............................................................................................................ 1 Introduction ............................................................................................................................. 1 Tetrabromobisphenol A ........................................................................................................... 4 Hexabromocyclododecane....................................................................................................... 8 Polybrominated Diphenyl Ethers........................................................................................... 11 OctaBDE................................................................................................................................ 16 DecaBDE ............................................................................................................................... 20 Polybrominated Diphenyl Ethers and Wildlife ..................................................................... 24 1,2-Bis (2,4,6-tribromophenoxy) Ethane ............................................................................... 30 CHAPTER 2 ................................................................................................................................. 36 THE PHARMOKINETICS OF 1,2 BIS (2,4,6-TRIBROMOPHENOXY) ETHANE IN MINK ....................................................................................................................................................... 36 Introduction ........................................................................................................................... 36 Materials and Methods .......................................................................................................... 39 Results ................................................................................................................................... 45 Discussion.............................................................................................................................. 48 CHAPTER 3 ................................................................................................................................. 51 THE EFFECTS OF 1,2 BIS (2,4,6-TRIBROMOPHENOXY) ETHANE ON REPRODUCTION OF ADULT FEMALE MINK AND SURVIVAL AND GROWTH OF THEIR OFFSPRING................................................................................................................. 51 Introduction ........................................................................................................................... 51 Materials and Methods .......................................................................................................... 56 Results ................................................................................................................................... 63 Discussion.............................................................................................................................. 78 Conclusion ............................................................................................................................. 85 CHAPTER 4 ................................................................................................................................. 87 CONCLUSION ......................................................................................................................... 87 REFERENCES ............................................................................................................................. 90 v   LIST OF TABLES Table 2.1: Composition of standard mink diet............................................................... 41 Table 2.2: Mean relative weight of liver, lung, and brain of mink exposed to 2.3 mg 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE)/kg feed ................. 46 Table 2.3: Mean concentration µg/kg wet weight) of 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) in various tissues of mink fed diets containing 2.3 mg BTBPE/kg feed for 21 days and then placed on clean feed for 21 days ............................................................................... 47 Table 3.1: Dietary concentrations (mg/kg wet weight) of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) ...................................................................... 64 Table 3.2: Estimated daily feed intake of adult female mink fed 0 or 2.3 mg 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE)/kg feed ............................ 64 Table 3.3: Effect of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on female mink pre-breeding weight ................................................................................ 66 Table 3.4: Effect of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on mink reproduction ........................................................................................ 67 Table 3.5: Effect of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on dam and kit body weight (g) ........................................................................ 68 Table 3.6: Effect of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on body weight (g) of juvenile males ............................................................... 69 Table 3.7: Effect of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on body weight (g) of juvenile females ............................................................ 70 Table 3.8: Effects of dietary 1,2bis (2, 4, 6-tribromophenoxy) ethane (BTBPE) on adult female mink relative (% of body weight) organ weight...................................................................................................................... 71 Table 3.9: Effects of dietary 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on female mink kit relative (% of body weight) organ weight at six weeks of age ................................................................................................ 72 Table 3.10: Effects of dietary 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on male mink kit relative (% of body weight) organ weight at six weeks of age ............................................................................................................ 73 vi   Table 3.11: Effects of dietary 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on juvenile female mink relative (% of body weight) organ weight at 27 weeks of age ................................................................................................ 74 Table 3.12: Effects of dietary 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on juvenile male mink relative (% of body weight) organ weight at 27 weeks of age ................................................................................................. 76 Table 3.13: Effects of dietary 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) on plasma triiodothyronine (T3) and thyroxine (T4) concentrations in adult female mink ................................................................................. 79 Table 3.14 Hepatic concentrations (µg/kg wet weight) of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) in juvenile mink ............................................ 80 Table 3.15 Adipose concentrations (µg/kg wet weight) of 1,2bis (2,4,6-tribromophenoxy) ethane (BTBPE) in juvenile mink ............................................ 81 vii   LIST OF FIGURES Figure 1.1. Structure of polybrominated biphenyls (PBBs) ............................................. 3 Figure 1.2. Structure of tris (2,3-dibromopropyl) phosphate (tris-BP) ............................. 4 Figure 1.3. Structure of tetrabromobisphenol A (TBBPA)............................................... 5 Figure 1.4. Structure of hexabromocyclododecane (HBCDD) ......................................... 9 Figure 1.5. Structures of α- (left), β- (middle) and γ-HBCDD (right) .............................. 9 Figure 1.6. Generic structure of polybrominated diphenyl ethers (PBDEs) ..................... 12 Figure 1.7. Structure of octabrominated diphenyl ether (OctaBDE) ................................ 17 Figure 1.8. Structure of decabrominated diphenyl ether (DecaBDE) ............................... 21 Figure 1.9. Structure of 1,2-bis (2,4,6-tribromophenoxy) ethane (BTBPE) ..................... 31 Figure 3.1. Photomicrographs of mink thyroids ............................................................... 77 viii   LIST OF ABBREVIATIONS ABS Acrylonitrile-butadiene-styrene Ach Acetylcholine ADFI Average daily feed intake BDE Brominated diphenyl ether BFRs Brominated flame retardants BMDL Bench mark dose limit BTBPE 1,2-bis(2,4,6-tribromophenoxy)ethane ChE Cholinesterase CNPase Cyclic nucleotide 3’-phosphodiesterase DBDPE Decabromodiphenyl ethane DCM:HEX Dichloromethane:n-hexane ECNI Electron capture negative ionization EPA Environmental Protection Agency EROD Ethoxyresorufin-O-deethylase F1 First filial generation FF-680 FireMaster 680 GC-MS Gas chromatography-mass spectrometry HBB Hexabromobenzene HBCDD Hexabromocyclododecane HIPS High-impact polystyrene IS Internal standard Kow Concentration in octanol phase/concentration in aqueous phase LD Lethal Dose ix   LD50 Lethal dose to 50% of population Lsmeans Least square means mAChR Muscarinic acetylcholine receptor MLOD Method limit of detection MLOQ Method limit of quantification MS Mass spectrometer MSD Mass spectrometer detector MSU Michigan State University m/z Mass to charge ratio n Number of subjects nAChR Nicotinic acetylcholine receptor NOAEL No observed adverse effect level OECD Organization for Economic Cooperation and Development PBBs Polybrominated biphenyls PBEB Pentabromoethylbenzene PBDE Polybrominated diphenyl ether PCBs Polychlorinated biphenyls PHA Phytohemagglutinin POPs Persistent organic pollutants POTWs Publicly owned wastewater treatment works PROD Pentaoxyresorufin-O-deethylase SAS Statistical Analysis Systems SE Standard error SIM Selected ion monitoring T3 Triiodothyronine x   T4 Thyroxine TBBPA Tetrabromobisphenol A TRI Toxic Reduction Inventory tris-BP Tris (2,3-dibromopropyl)phosphate TSH Thyroid stimulating hormone UGT Uridine diphosphate-glucuronosyltransferases UDPGT Uridine diphosphate-glucuronosyltransferases UK United Kingdom US United States UV Ultraviolet xi   CHAPTER 1 LITERATURE REVIEW Introduction More than 3,000 people have been killed, more than 20,000 people have been injured and an estimated $11 billion in property damage has resulted from fires in the United States (US) alone in 2001 (Birnbaum and Staskal, 2004). The US has higher standards for protection against fire than Europe and as a result, more lives have been saved, fewer fire-related injuries have occurred and less economic loss have resulted by using flame retardants (Birnbaum and Staskal, 2004). Flame retardants are used to prevent or retard the initial phase of a developing fire (Sjödin et al., 2001). The incorporation of flame retardants into materials provides for longer escape time during fires by allowing less release of heat, smoke, and toxic gases, thus saving lives (Silva et al., 2004) Brominated flame retardants (BFRs) are the largest group of flame retardants used worldwide. They have been used widely since 1975 and their use continues to increase. An estimated 410,000 metric tons of BFRs are used in the global market each year (Covaci et al., 2011). The major commercial BFRs used are tetrabromobisphenol A (TBBPA), hexabromocyclododecane (HBCDD) and polybrominated diphenyl ethers (PBDEs; Birnbaum and Staskal, 2004). Brominated flame retardants are incorporated into electronics, clothes and furniture. The electronic industry uses these compounds in printed circuit boards, connectors, plastic covers, and cables. They are also used in consumer products such as plastic covers of televisions, microwaves and other kitchen appliances as well as in carpets, paints, wood products, textiles, and paper (Birnbaum and Staskal, 2004). 1   Due to their chemical structure, brominated flame retardants are lipophilic and are resistant to physical and biochemical degradation, making them a potential environmental hazard. Brominated flame retardants enter the environment through industrial waste, leaching from the product itself, or from landfills, and as a result are found in the environment far from where they are produced or used. Many BFRs are persistent and bioaccumulative and can biomagnify in the food web. As a result of those properties, increased concentrations have been found in terrestrial, freshwater, and marine ecosystems, affecting the environment, wildlife and humans. Therefore they are classified as persistent organic pollutants (POPs) (Alaee and Wenning, 2002; de Wit, 2002, and de Wit et al., 2006). The use of some BFRs has been discontinued or restricted due to the incidence of environmental contamination and/or because of adverse health effects. Specific BFRs that have been withdrawn from the market or have been restricted in use include polybrominated biphenyls (PBBs) and tris (2,3-dibromopropyl) phosphate (tris-BP) (Birnbaum and Staskal, 2004). One specific incidence of environmental contamination involved FireMaster FF-1, which was a commercial PBB product manufactured by Michigan Chemical Company and was used as a flame retardant in industrial and consumer products (DiCarlo et al., 1978; Fries, 1985). FireMaster FF-1 was a mixture of PBB congeners formed by substituting bromine for hydrogen on the biphenyl molecule that consists of two benzene rings. Theoretically, there are 209 possible congeners considering the five sites available on each ring for binding of bromine (Figure 1.1). The hexabromobiphenyl congeners were the predominant congeners in the commercial mixture (http://www.inchem.org/documents/ehc/ehc/ehc162.htm). In 1973, several hundred pounds of FireMasterFF-1 were accidentally mixed into animal feed that ultimately resulted in the disposal of almost 30,000 contaminated dairy cows and over two million 2   Fi igure 1.1. Structure of polybromina S p ated bipheny (PBBs) yls nated chicken (de Wit, 2002; http://n ns 2 ntp.niehs.nih h.gov/ntp/ro oc/twelfth/pr rofiles/contamin Polybrom minatedBiph henyls.pdf). More than 90% of the re 9 esidents in M Michigan had detectable d quantities of PBBs st tored in their body fat as a result of c r s consumption of contami n inated dairy , m son 978; Miceli e al., 1985). The use of PBBs was et . f products, eggs, and meat (Anders et al., 19 discontin nued in the US in 1976 as a result of the Michiga contamina U a f an ation inciden (Di Carlo et nt al., 1978) ). Tris (2,3-dibr T romopropyl) phosphate (Figure 1.2), was manufa ( , factured by th former he Michigan Chemical Company, al the major BFR used in sleepwea between 19 and 197 to n C lso ar 973 77 comply with federal regulations designed to reduce burn injuries and death. As a result, it w w d r d was estimated that over 50 million ch d hildren were exposed to t chemica (Blum et a 1978). T this al al., Tris (2,3-dibr romopropyl)p phosphate, which was sh w hown to be m mutagenic an nephrotox (Dybing et nd xic al., 1980; Söderlund et al., 1980) was withdr ), rawn from th US marke in 1977 (P he et Prival et al., um es, 1977; Blu and Ame 1977). 3   The five majo BFRs curr T or rently used in commerce are tetrabro i e omobisphenol A (TBBP PA), hexabrom mocyclodode ecane (HBCDD), and thr commerc mixture of polybro ree cial es ominated diphenyl ethers (PBD DEs). The three commercial PBDE m mixtures are pentabrominated diphen nyl ether (pentaBDE or “penta”), oct “ tabrominated diphenyl e d ether (octaBD or “octa” and DE ”) decabrom minated diph henyl ether (d decaBDE or “deca”). r Figure 1.2. Structure of tris (2,3 e 3-dibromopropyl)phosp phate (tris-B BP) Tetrabro omobisphen A nol The most wid used BF is TBBPA (Figure 1.3 with an es T dely FR A 3) stimated use worldwide of e 150,000 metric tons (Covaci et al., 2011). Te m ( etrabromobi isphenol A is used as a r s reactive flam me retardant in printed circuit boards and as an additive BFR in several polymers. A a reactive t c s, R As BFR, TB BBPA is inco orporated int the produc and releas into the en to ct se nvironment i minimal, is 4   Figure 1.3. Structu of tetrabromobisph ure henol A (TB BBPA) whereas when TBBP is used as an additive BFR, it can be released into the env PA s e n d vironment m more readily (B Birnbaum an Staskal, 2004). nd 2 Tetrabromobi T isphenol A and its metab a bolites have been detecte in the atm ed mosphere, so oil and sedim ment, but usu ually not in water. Tetra w abromobisph henol A was detected in t atmosphere the 3 at a concentration of 1.8 µg TBB BPA/m at a production s in Japan (Birnbaum and Staskal, site n 2004). Soil and sedim sample collected in Japan con S ment es ntained TBB BPA at conce entrations ranging from 0.5 to 140 µg/kg (d weight) and 2 to 150 µg/kg (dry weight), resp f 1 dry a pectively (Watanab et al., 198 1983b). The presen of TBBP is limited in biota bec be 83a, nce PA d cause of its relatively short half-l in water, air, and sed y life , diment. Env vironmental b biodegradati studies o ion of TBBPA indicate that the chemica partially breaks down under both aerobic and anaerobic i t al b n condition with an approximate half-life of two months (Birnbaum a Staskal, 2004). ns, a t and 5   Szymanska et al. (2001) investigated the metabolism of TBBPA in rats. After a single dose of 14 C-TBBPA, 51 to 95% of the dose was rapidly excreted in the feces largely unchanged. After intraperitoneal (IP) administration of 14 C-TBBPA (250 or 1,000 mg/kg body weight), Szymanska et al. (2001) reported that the feces contained 90% TBBPA and 10% of the metabolite tribromobisphenol A. Fecal analysis suggested rapid elimination of TBBPA in the bile and possible debromination by gastrointestinal flora. The adipose tissue had the highest concentration of the compound within the first 60 minutes. At 72 hours post–injection, adipose tissue and muscle still retained a percentage of the TBBPA (3 to 6% and 11 to 14%, respectively), suggesting that TBBPA or a metabolite has the potential to bioaccumulate with repeated exposure (Szymanska et al., 2001). Toxicity studies conducted with rodents have shown TBBPA to be relatively non-toxic. The oral LD50 (dose lethal to 50% of the population) is greater than 5 g/kg body weight in rats and greater than 4 g/kg body weight in mice (http://www.inchem.org/documents/ehc/ehc/ehc172.htm). In vivo studies have not shown TBBPA to be an inhalation, dermal or ocular toxicant or teratogen (http://www.inchem.org/documents/ehc/ehc/ehc172.htm). A 28-day repeat-dose study assessing the effects of TBBPA in rats using the Organization for Economic Cooperation and Development (OECD) guidelines was conducted by Van der Ven et al. (2008). Dietary concentrations were 0, 30, 100 or 300 mg TBBPA/kg body weight and each treatment group had 10 animals per sex. The animals were necropsied at 12 weeks of age. The effects reported were decreased plasma total thyroxine (TT4) concentration at a benchmark dose (BMDL10) of 48 mg/kg body weight/day and increased plasma total triiodothyronine (TT3) concentration at a BMDL10 of 124 mg/kg body weight/day in males. 6   A one-generation reproduction study assessing the effects of TBBPA in rats was conducted using the OECD guidelines (Van der Ven et al., 2008). Rats were assigned to one of eight dose groups (0, 3, 10, 30, 100, 300, 1000, 3000 mg/kg body weight/day). The design allowed a dose-response analysis and calculation of BMDL. Parental exposure started 70 days prior to mating in males and 14 days prior to mating in females. The F1 animals remained on their respective doses until necropsy, which was up to 17 weeks of age. The most sensitive effects were an increase in testis weight and an increase in pituitary weight in adult F1 males. The BMDL5 and BMDL10 based on these effects were 0.5 and 0.6 mg/kg body weight/day, respectively. The testis weight in F1 males at postnatal day 21 was also affected resulting in a BMDL5 of 0.5 mg/kg body weight/day. Plasma T4 concentration was decreased in F1 animals at a BMDL10 of 31 and 16 mg/kg body weight/day for males and females, respectively. Plasma T3 concentrations were increased in F1 females at a BMDL10 of 2.3 mg/kg body weight/day (Van der Ven et al., 2008). Saegusa et al. (2009) exposed pregnant female rats to 0, 100, 1000, or 10,000 mg TBBPA/kg feed from gestational day 10 to post-natal day 20. The dams did not have any significant toxic effects, although there was evidence of thyroid follicular cell hypertrophy in the 1000 mg TBBPA/kg feed treatment group and non-dose related increased relative thyroid gland weight in all treatment groups. The only effect in the offspring was a reduction in serum T3 concentrations in the 100 and 1000 mg TBBPA/kg feed groups that was not related to dose. Serum T4 and thyroid stimulating hormone (TSH) concentrations were not changed. Also, there was no evidence of impaired brain development such as hypothyroidism-related neuronal 7   migration or oligodendroglial development. The results suggest that TBBPA can induce developmental hypothyroidism. Hexabromocyclododecane Hexabromocyclododecane (Figure 1.4) is a nonaromatic, brominated cyclic alkane primarily used in thermoplastic polymers with final applications in styrene resins (Birnbaum and Staskal, 2004). It is an additive flame retardant, with total production around 22,000 metric tons per year (de Wit et al., 2010), making it a minor contributor to the BFR economy. Hexabromocyclododecane is highly lipophilic and has low water solubility (0.0034 mg/L) (Birnbaum and Staskal, 2004). Hexabromocyclododecane has been shown to be persistent in the environment, with a half-life of three days in air and two to 25 days in water (Lyman et al., 1990). Studies have shown that HBCDD has a strong propensity to bioconcentrate, with a bioconcentration factor of approximately 18,100 in fathead minnows and fish-to-sediment ratios as great as 15 (Veith and Defoe 1979; Sellström et al., 1998). The commercial HBCDD product contains three diastereomers, α- (10 to13%), β- (1 to 12%) and γ- (75 to 89%) HBCDD (Figure 1.5). Hexabromocyclododecane is usually found in sediments as γ-HBCDD (> 90%). However, small amounts of the α- and β-diastereomers have been found in some regions with high concentrations of HBCDD (http://www.epa.gov/opptintr/tsca8e/pubs/8ehq/2002/-oct02/8ehq_-1002_15166b.pdf). The lower level organisms contain mostly the γ-diastereomer, while the top predators contain mostly the α-diastereomer (http://www.epa.gov/opptintr/tsca8e/pubs/8ehq/-2002/oct02/8ehq_1002_15166b.pdf). 8   Fig gure 1.4. Str ructure of he exabromocyc clododecane (HBCDD) e Figure 1.5. Structu of α-, (l ures left), β-, (mid ddle) and γ-H HBCDD (right) g 9   Laboratory studies in rats have shown HBCDD to be capable of causing adverse effects. In a HBCDD study examining endocrine effects (Van der Ven et al., 2006), rats were given oral doses of 0, 0.3, 1, 3, 10, 30, 100, and 200 mg HBCDD/kg body weight/day for 28 days. Daily dosing resulted in dose-dependent increase in liver concentrations of HBCDD, indicating absorption of the HBCDD and increased bioavailability. The marked HBCDD dose-related effects were mostly seen in the thyroid hormone axis, including decreased concentration of total T4 (BMDL10 of 55.5 mg/kg body weight/day), increased immunostaining for TSH in the pituitary, increased weight of the pituitary (BMDL10 of 29 mg/kg body weight/day) and thyroid glands (BMDL10 of 1.6 mg/kg body weight/day), and induction of T4-UGT in the liver (BMDL10 of 4.1 mg/kg body weight/day) resulting in increased liver weight (BMDL10 of 22.9 mg/kg body weight/day). T4-UGT is involved in the metabolism of T4, thus its induction decreases plasma total T4, which in turn leads to an increase in TSH and an increase in the weight of the thyroid gland. These HBCDD effects were only seen in females. Additional effects seen in the females were increased cholesterol (BMDL10 of 7.4 mg/kg body weight/day) and increased tibial bone mineral density (BMDL10 greater than 49 mg/kg body weight/day). Males had decreased splenocyte counts (BMDL20 of 0.3–6.3 mg/kg body weight/day). In another rat study, the developmental toxicity of HBCDD was evaluated. Hexabromocyclododecane effects were assessed in rat offspring after maternal exposure from mid-gestation through lactation at feed concentrations of 100, 1000, or 10,000 mg/kg (Saegusa et al., 2009). There was an increase in relative thyroid gland weight and incidence of thyroid follicular cell hypertrophy in dams in the 10,000 mg HBCDD/kg feed treatment group at 10   weaning. The offspring in the 10,000 mg HBCDD/kg feed treatment group, at this time point, had a decrease in serum T3 concentration and an increase in serum TSH concentration. An increase in thyroid gland weight and decrease in serum T3 concentration persisted until the adult stage in offspring exposed to 1,000 mg HBCDD/kg feed. Hexabromocyclododecane reduced density of CNPase-positive oligodendrocytes at 10,000 mg HBCDD/kg feed, which suggested impaired oligodendroglial development, probably as a result of developmental hypothyroidism. A no observed adverse effect level (NOAEL) of 100 mg HBCDD/kg feed was selected based on the developmental brain effects and changes in thyroid parameters. Polybrominated Diphenyl Ethers Polybrominated diphenyl ethers are another class of BFRs that are recognized as global environmental contaminants. Roughly 70,000 metric tons of PBDEs are produced annually worldwide, with half being produced in the US and Canada (Renner, 2000; Hites, 2006). The structure of a generic PBDE molecule is illustrated in Figure 1.6. Polybrominated diphenyl ethers consist of two phenyl rings with an ether linkage between the two rings. Diphenyl ether molecules contain 10 hydrogen atoms, any of which can be exchanged with bromine, resulting in 209 PBDE congeners (Alaee et al., 2003). There are three commercial mixtures of PBDEs that have been marketed as flame retardants: “penta” or pentaBDE, “octa” or octaBDE and “deca” or decaBDE (Costa and Giordano, 2007). The commercial pentaBDE product is a mixture of 50 to 60% pentaBDEs, 24 to 38% tetraBDEs, 4 to 8 % hexaBDEs and 0 to 1% triBDEs and is sold under numerous trade names including Bromkal G1, DE 60FTM, Bromkal 70, and DE-71 (http://www.inchem.org/documents/ehc/ehc/ehc162.htm). 11   Figure 1.6. Generic str ructure of po olybrominate diphenyl ethers (PBD ed DEs) at PentaBDE is used as an additive flam retardant a concentra a me ations rangin from 5 to 30% ng in many different pol lymers, resin and other substrates. The primary use of pent ns y taBDE is in polyureth hane foams, where up to 30% of the weight of th foam can be accounte for by this o he ed s flame ret tardant (Hale et al. 2002) Penta-bro e ). ominated dip phenyl ether also has min uses in nor phenolic resins, polyesters, and epoxy resins (Birnbaum and Staskal, 2004). App e , proximately y etric tons of pentaBDE are used annu a ually with 95 being us in the US 5% sed S 7,500 me (http://ww ww.ec.gc.ca a/lcpe-cepa/0 09F567A7-B B1EE-1FEE-73DB-8AE6 6C1EB7658 8/sar_pbdeeng.pdf). . Penta-bromin nated diphenyl ether may be released into the env y d vironment fr rom various sources. Wastewater streams are an environm r e mental sourc of release as a result o handling t ce e of the ο erial. During the curing phase of foa productio temperat g am on, tures reach 1 C for sev 160 veral raw mate hours, po otentially releasing volat tilized PBDE congeners E (http://ww ww.atsdr.cdc.gov/toxpro ofiles/tp68-c c8.pdf). The estimated o e overall releas of pentaB se BDE to wastew water is appr roximately 0.11% with approximatel half of tha quantity e 0 a ly at entering the 12   atmosphere (http://www.atsdr.cdc.gov/toxprofiles/tp68-c8.pdf). Polyurethane foam scraps, which can contain up to 30% penta, are used in products such as car seats or carpet underlay and may eventually be landfilled releasing PBDEs into the environment (http://www.atsdr.cdc.gov/toxprofiles/tp68-c8.pdf). Penta-brominated diphenyl ether can be very persistent in the environment, binding to the organic fraction of particulate matter and the lipid fraction of biota. Components of the mixture will be expected to partition primarily to sediment (approximately 59%), followed by soils (approximately 40%), water (1.2%), and air (0.2%). The estimated half–lives of pentaBDE, with atmospheric degradation, are 600 days in aerobic sediment, 150 days in soil and 150 days in water (http://www.ec.-gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB8AE6C1EB7658/sar_pbde-eng.pdf). The bioaccumulation of pentaBDE increases as the trophic level increases, indicating biomagnification in the food web (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB8AE6C1EB7658/sar_pbde-eng.pdf). Increasingly elevated concentrations have been detected in fish and the piscivorous avian in the food chain. Concentrations of pentaBDE at the mg/kg lipid level have been documented in marine mammals such as whales, dolphins and seals. In blubber samples collected from San Francisco Bay harbor seals, concentrations of pentaBDE rose from 88 µg/kg lipid in 1989 to 8325 µg/kg lipid in 1998 (http://www.ec.gc.ca/lcpe-cepa/09F567A7B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbde-eng.pdf). Commercial mixtures of pentaBDE have been shown to induce both phase I (ethoxyresorufin-O-deethylase [EROD] and pentoxyresorufin-O-deethylase [PROD]), and phase II (uridinediphosphate-glucuronosyltransferase [UDPGT]) metabolic enzymes. T4 glucuronidation by phase II UDPGT in the liver has been suggested as one of the mechanisms 13   contributing to circulating T4 depletion by PBDEs and other polyhalogenated aromatic hydrocarbons (PHAHs). There are a number of studies that assessed the effects of pentBDE on thyroid hormone concentration. There are many studies that assessed the toxicity of various commercial pentaBDE mixtures in rodents. Mice were administered a commercial pentaBDE mixture, DE-71, as either a single dose (0.8, 4, 20, 100, or 500 mg DE-71/kg body weight) or daily doses over 14 days (18, 36 or 72 mg DE-71/kg body weight/day for total doses of 250, 500, or 1000 mg DE-71/kg body weight) (Fowles et al., 1994). In the acute trial, total serum T4 concentrations were significantly lower compared to controls at all doses except 100 mg/kg body weight. In the 14-day study, there was a significant induction of total hepatic microsomal cytochrome P450 and hepatic EROD and PROD activities at doses greater than 250 mg/kg body weight. There was also a dose-dependent decrease in the concentrations of total and free serum T4 concentrations and an increase in serum corticosterone concentrations. An immunosuppressive effect was demonstrated by a decreased sheep erythrocyte plaque forming cell response and a decrease in thymus weight at 1000 mg DE-71/kg body weight. In a study by Zhou et al. (2001), weanling rats were dosed with 0.3, 1.0, 3.0, 10, 30, 100, or 300 mg DE-71/kg body weight/day for four days. Hepatic EROD and PROD activities were significantly greater compared to control activities and there was a dose-related increase in uridine diphosphate glucuronyl transferase (UDPGT) activity. A dose-dependent decrease in total serum T4 concentrations was observed. Bromkal 70, a commercial pentaBDE mixture containing 64% pentaBDEs and 36% tetraBDEs, was administered to rats using one of three dosing regimens: a single dose of 300 14   mg/kg body weight, 100 mg/kg body weight/day for 4 days or 50 mg/kg body weight/day for 28 days. Relative liver weight was greater in treated animals than in controls at all doses. Additionally, there was significant induction of hepatic EROD activity in all treated animals, compared to controls at all doses (Von Meyerinck et al., 1990). The study concluded that Bromkal 70, and pentabrominated diphenyl ethers, act as mixed type inducers of liver enzymes (Von Meyerinck et al., 1990). A 90-day feeding trial was conducted in which the commercial pentaBDE mixture DE-71 was administered to rats at doses of 2, 10 or 100 mg DE-71/kg body weight/day for 90 days. Absolute liver weight were increased by 11% in animals dosed with 10 mg DE-71/kg body weight/day and by 50 and 70% in females and males, respectively, at 100 mg DE-71/kg body weight/day. Histopathological changes in the liver included hepatocytomegaly and hepatocyte degeneration and necrosis. Slight hyperplasia of the thyroid occurred in animals dosed with 100 mg DE-71/kg body weight/day and thyroid gland weight was 30% greater compared to controls. Serum T4 concentrations were decreased by greater than 20% at doses of 10 and 100 mg DE71/kg body weight/day (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB8AE6C1EB7658/sar_pbde-eng.pdf). Individual congers in commercial pentaBDE products may be neurotoxic to mammalian systems if exposure is during a critical time in development. Neonatal mice administered a single oral dose of 2,2',4,4'- tetrabromodiphenyl ether (BDE-47) at 0.7 or 10.5 mg/kg body weight or 2,2',4,4',5-pentabromodiphenyl ether (BDE-99)/kg body weight at 0.8 or 12.0 mg/kg body weight on day 10 of development had decreases in habituation, locomotion, rearing and total activity that became more pronounced with age. Adult memory and learning were also affected (Eriksson et al., 1998). 15   OctaBDE Octabrominated diphenyl ether (OctaBDE) (Figure 1.7) is another commercial PBDE product that is composed of 45% heptaBDEs, 33% octaBDEs, 12% hexaBDEs, 10% nonaBDEs, 0.7% decaBDE and 0.05% pentaBDE (http://chm.pops.int/Convention/POPsReviewCommittee/Reports/tabid/2301/Default.aspx). Trade names for OBDE include Bromkal 79-8DE, CD-79, DE-79, EB-8, FR-1208, FR-143, Tardex 80, Saytex 111 and Adine 404. Approximately 3,790 metric tons of octaBDE have been used annually with the majority being used in the US (1,500 metric tons) and Asia (1,500 metric tons) (Birnbaum and Staskal, 2004). Approximately 70% of the octaBDE manufactured globally is added to acrylonitrilebutadiene-styrene (ABS) polymers that are then used to produce computers and business cabinets (http://www.inchem.org/documents/ehc/ehc/ehc162.htm). The polymers are also used in pipes and fittings, automotive products, business machines and appliances (http://www.ec.gc.ca/lcpecepa/09F567A7-B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbde-eng.pdf). Octabrominated diphenyl ether is also included in high-impact polystyrene (HIPS), polybutylene, terephthalate, polyamide polymers, nylon and low-density polyethylene, polycarbonate, phenol-formaldehyde resins and unsaturated polyesters, and adhesive coatings (http://www.ec.gc.ca/lcpecepa/09F567A7-B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbde-eng.pdf). There are many potential routes by which octaBDE may be released into the environment. Manufacturing of plastic is one of many potential routes for release (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbdeeng.pdf). During the handling of octaBDE, there is loss of the raw material, estimated at 0.21% for powders of particle size greater than 40 µm (http://www.ec.gc.ca/lcpe-cepa/09F567A7B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbde-eng.pdf). The losses are initially to the 16   atmosphe but the dust rapidly settles, creat ere, d ting primaril solid wast which ma be recycled or ly te, ay disposed of, or it may y e r w.ec.gc.ca/lc cpe-cepa/09F F567A7-B1E EE-1FEE-73 3DBenter the wastewater (http://www 8AE6C1E EB7658/sar_ _pbde-eng.p pdf). It is est timated that approximate 0.05% octaBDE is ely released into the atmosphere and 0.05% into wastewater during the p d processing of thermoplas stic polymers (http://www s w.ec.gc.ca/lc cpe-cepa/09F F567A7-B1E EE-1FEE-73 3DB8AE6C1E EB7658/sar_ _pbde-eng.p pdf). Figure 1.7. Structure of Octabrom e minated diph henyl ether ( (OctaBDE) Congeners comprising th commerc octaBDE mixture ha been dete c he cial E ave ected in air f emote sites in the US ran nging from a approximately 0.2 to 0.9 9 samples from urban, rural, and re 3 pg/m (S Strandberg et al., 2001). In the west, central, and eastern bas of Lake Ontario, the t d sins ere are measurable total PBDE (mon through heptaBDE co no ongeners) con ncentrations of 0.0039, s 17   0.0065 and 0.0053 ng/L, respectively (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE73DB-8AE6C1EB7658/sar_pbde-eng.pdf). Concentrations of octaBDE in United Kingdom (UK) sediments ranged from less than 0.44 to 3030 µg/kg dry weight (Allchin et al., 1999; http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbdeeng.pdf). Octabrominated diphenyl ether was detected in sediment samples in Japan in 1987 and 1988 at concentrations of 8 to 21 μg/kg and 15 to 22 µg/kg, respectively (http://chm.pops.int/Convention/POPsReview-Committee/Reports/tabid/2301/Default.aspx). Concentrations of octaBDE components (hexaBDE and heptaBDE congeners) detected in lake trout collected from lakes Superior, Huron, and Ontario ranged from 11 to 53 µg/kg lipid (Alaee et al., 1999). Sampling of wild bird eggs in western and northern Canada between 1983 and 2000 determined that concentrations of total hexa- and heptaBDE congeners ranged from 0.148 to 52.9 µg/kg wet weight in great blue heron (Ardea herodias) eggs, 0.03 to 0.68 µg/kg wet weight in northern fulmer (Fulmarus glacialis) eggs and 0.009 to 0.499 µg/kg wet weight in thick billed murre (Uria lomvia) eggs. The presence of these congeners in arctic bird eggs on Canada’s west coast and in the Canadian arctic suggests long-range transport of the hexa- and heptaBDE (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB8AE6C1EB7658/sar_pbde-eng.pdf). There are a number of rodent studies that have assessed the toxicity of the commercial octaBDE mixture. The data available for rodents indicate a very low level of acute toxicity. An acute oral LD50 of 28,000 mg/kg body weight for octaBDE in rats was reported (http://www.inchem.org/-documents/ehc/ehc/ehc162.htm). In a study by Zhou et al. (2001), weanling rats were administered DE-79 at doses of 0.3, 1.0, 3.0, 10, 30, 60 or 100 mg DE-79/kg body weight/day for a four-day period. Hepatic EROD and PROD activities were significantly 18   increased in the highest dose group, with 20-fold induction for EROD activity and 26-fold induction for PROD activity. Additionally, there was a dose-dependent decrease in total serum T4 concentrations, with a maximum decrease of 70% in the animals dosed with 100 mg DE79/kg body weight/day. In another study conducted by Great Lakes Chemical Corporation (http://www.inchem.org/documents/ehc/ehc/ehc162.htm), rats were administered DE-79 for 28 days or 13 weeks. The doses used in the 28-day study were 0, 100, or 1000 mg DE-79/kg feed. At the end of the 28-day study, absolute and relative mean liver weight were significantly greater in the 100 and 1000 mg/kg feed treatments. The doses used in the 13-week study were 0, 100, 1,000, or 10,000 mg/kg feed. At the termination of the 13-week feeding study, statistically significant increases in absolute and relative liver weight were evident at all feed concentrations. Liver cells were enlarged, with finely granular appearing cytoplasm and increased vacuolation. Hepatocyte necrosis at the 1,000 and 10,000 mg DE-79/kg feed concentrations was also detected. A rat fetal toxicity study was performed using octaBDE preparations of DE-79 and Saytex 111. Maternal doses of 0, 2.5, 10, 15, 25 or 50 mg/kg body weight/day were administered by gavage from days 6 to 15 of gestation. Adult females dosed with 50 mg/kg body weight/day had reduced body weight gain. An increase in mortality and fetal reabsorptions, delayed skeletal ossification and reduced fetal body weight occurred at doses of 10, 15 and 25 mg/kg body weight/day (http://www.inchem.org/documents/ehc/ehc/ehc162.htm). Viberg et al. (2003a) assessed the neurotoxic effects of BDE-153, a congener found in commercial octaBDE mixtures, in mice. Neonates were administered a single dose of 0.45, 0.9, or 9.0 mg BDE-153/kg body weight on postnatal day 10, a time associated with rapid brain development in mice. At 2, 4, and 6 months post-dosing, treated mice had significantly altered 19   motor behavior. The incidence of abnormal behavioral responses was dose-related and the condition worsened with age. Adult mice also had aberrant behavior with impaired spatial learning ability and memory function. DecaBDE The commercial decaBDE (Figure 1.8) mixture is composed of 97-98% decaBDE and small quantities of other PBDEs (mainly nonaBDEs at 0.3-3.0%) (http://www.inchem.org/documents/ehc/ehc/ehc162.htm). Some trade names for DBDE are 102 (E), Adine 505, AFR 1021, Bromkal 81 and Saytex 102. DecaBDE is the second most used BFR globally with approximately 56,400 metric tons being used annually (de Wit et al., 2010). The majority of its use is in the Americas (24,500 metric tons) and Asia (23,000 metric tons). DecaBDE is used in a variety of polymer applications at a rate of approximately 10 to 30% of the total product weight (Kierkegaard et al., 2009). DecaBDE is also used in terephthalates to make moldings, connectors and electrical equipment (http://www.oekorecherche.de/english/berichte/volltext/Flame%-20Retardants.pdf). There is potential for decaBDE to be released into the environment during the manufacturing process, throughout the service life of the materials containing decaBDE and during disposal (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB8AE6C1EB7658/sar_pbde-eng.pdf). The US EPA’s Toxic Reduction Inventory (TRI) indicated that most US decaBDE manufacturer emissions were directed to land and publicly owned wastewater treatment works (POTWs). The 2001 TRI estimated the emissions from the textile industry were by far the largest decaBDE discharge to US surface waters and POTWs. Publicly owned wastewater treatment works consequently affect sewage sludge burdens and effluent receiving streams. About 98% of the 20   estimated air emissio were attr d ons ributed to fug gitive dust re eleases from facilities be m elonging to t the two majo US decaB or BDE manufac ctures (Hale et al., 2006) ). Once decaBD is release into the en O DE ed nvironment, it can be fou in sedim and sew und ment wage sludge. Sediments tak near a manufacturin site in Sw S ken m ng weden had de ecaBDE conc centrations o of 68-390 µg/kg dry we µ eight and dec caBDE conc centrations in river sedim n ments collect from an ted urban are in Japan were 33-390 µg/kg dry weight (Sells ea w w ström et al., 1998). Deca abrominated diphenyl ether has be found at higher conc een centrations in samples fr n rom riverine, estuarine an , nd ediments, in comparison to other bro n ominated dip phenyl ether such as he rs, exaBDE, marine se pentaBDE and tetraB BDE. The de ecaBDE con ncentrations ranged from less than 25 to 11,600 m 5 µg/kg dry weight, wh y hereas conce entrations of the latter co f ompounds w in the ra were ange of less t than the limit of detection of 70 µg/kg dry weight (Darnerud e al., 2001). n g t et . Figure 1.8. Structu of decab e ure brominated d diphenyl ethe (DecaBDE er E) 21   Decabromodiphenyl ether has also been identified in air and water samples. DecaBDE was identified in 10 samples of air from a manufacturing site at atmospheric concentrations that 3 were below the analytical detection limit of 25 µg/m , while aquatic concentrations of decaBDE were reported to be in the range of 0.2 to 2.5 µg/L (http://www.inchem.org/documents/ehc/ehc/ehc162.htm#PartNumber:1). The atmospheric half-lives of tetraBDE to decaBDE ranged from 7.14 to 317.53 days (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB8AE6C1EB7658/sar_pbde-eng.pdf). DecaBDE may be debrominated in the natural environment to lower brominated PBDEs by sunlight and UV light based on laboratory studies (Watanabe and Sakai, 2003). It is also possible that decaBDE can undergo anaerobic biodegradation and dehalogenation (http://www.atsdr.cdc.gov/ToxProfiles/tp68-c8.pdf; http://www.ec.gc.ca/lcpe-cepa/09F567A7B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbde-eng.pdf). DecaBDE can absorb light up to 325 nm, which indicates its susceptibility to photodegradation at environmental wavelengths (http://www.atsdr.cdc.gov/ToxProfiles/tp68-c8.pdf). Although it was first believed that bioaccumulation of decaBDE in aquatic organisms was minimal due in part to the large molecular size of the decaBDE constituents (Opperhuizen et al. 1985; http://www.inchem.org/documents/ehc/ehc/ehc162.htm), recent research has shown that decaBDE can bioaccumulate in aquatic organisms. Juvenile rainbow trout (Oncorhynchus mykiss) were dosed with DOW FR-300 (77.4% decaBDE, 21.8% nonaBDEs and 0.8% octaBDEs). Several hexa-, octa- and nonaBDEs were detected in the liver and muscle of treated fish with concentrations increasing with duration of exposure. The presence of various PBDE congeners suggested metabolic debromination of decaBDE (Kierkegaard et al., 1999). It has been reported that peregrine falcon eggs had high concentrations of decaBDE, indicating that the 22   commercial mixture has the potential to bioaccumulate in terrestrial organisms (http://bfr2010.com/abst/2001/BFR2001del5.pdf#page=33). In general, toxicity studies have not shown decaBDE to have deleterious effects. The decaBDE studies performed with aquatic organisms have shown the compound to be acutely nontoxic at concentrations up to and exceeding its limit of water solubility (< 0.1 μg/L) (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbdeeng.pdf). There are numerous rodent studies that have assessed the toxicity of decaBDE. Commercial decaBDE has been shown to have low oral toxicity. Male and female rats were fed a commercial decaBDE mixture (> 97% decaBDE with nonaBDE present at 0.3 to 3%) at concentrations of 100 and 1000 mg/kg feed for a period of 28 days. Corresponding approximate doses were 10 and 100 mg/kg body weight/day. The only observable change that could be attributed to treatment was increased bromine content in liver and fat samples from both treatment groups (http://www.inchem.org/documents/ehc/ehc/ehc162.htm). A 30-day study in male rats that were fed Dow FR-300-BA (77.4% decaBDE, 21.8% nonaBDE, 0.8% octaBDE) at concentrations of 0.01, 0.1 and 1.0%, which corresponded to approximate daily doses of 8, 80 and 800 mg/kg body weight. The significant effects reported were liver enlargement and thyroid hyperplasia in the 0.1 and 1.0% groups, and renal and hepatic lesions in the high dose group (Norris et al., 1975; Darnerud et al., 2001). In a decaBDE carcinogenicity study, rats and mice were fed diets containing commercial decaBDE (94 to 99% purity) for 103weeks. Tumor incidence was observed at 1,200 and 2,500 mg/kg body weight/day in rats. A dose related increase in liver adenomas was observed in rats and mice fed decaBDE. Hepatocellular adenomas and carcinomas were increased in mice at doses of 3,500 and 7,000 mg/kg body weight/day (Darnerud et al., 2001) 23   Pregnant rats were administered Dow-300-BA at doses of 10, 100, or 1000 mg/kg body weight on days 6 to 15 of gestation. Fetuses in the high dose group had evidence of subcutaneous edema and a significantly increased incidence of delayed skull ossification (Norris et al. 1975; Darnerud et al., 2001). Viberg et al. (2003b) evaluated the potential neurotoxicity of BDE-209 in mice. A single oral dose of 2.22 or 20.1 mg BDE-209/kg body weight was administered to mice on postnatal days 3 or 19. An additional study used a single oral dose of 1.34, 13.4 or 20.1 mg BDE-209/kg body weight administered to 10-day-old mice (Viberg et al., 2003b). Mice exposed to the 20.1 mg BDE-209 dose on postnatal day 3 had disturbances in spontaneous behavior (motor behavior, locomotion, rearing, and total activity) as adults. The disruption of spontaneous behavior appeared to worsen with age. When comparing the habituation capacity (decrease in response to a stimulus after repeated exposure to the stimulus) in 2-, 4-, and 6-month-old mice, there was a decrease in locomotion, rearing and total activity in response to diminishing novelty of the test chamber. The mice exposed on postnatal day 3 showed the aberrant behavior pattern at 6 months of age only. Mice exposed on postnatal days 10 and 19 did not have disrupted habituation behavior. The results from this study indicate BDE-209 can be neurotoxic in adult mice exposed neonatally. Polybrominated Diphenyl Ethers and Wildlife Polybrominated diphenyl ethers are known global environmental contaminants, posing health risks to humans and wildlife (Segev et al., 2009). Lower brominated congeners (i.e. mono- to hexaBDEs) are generally more persistent (Gerecke et al., 2006; Law et al., 2006), bioaccumulative (Burreau et al., 2004, 2006) and toxic (Darnerud, 2003) than higher brominated congeners. DecaBDE is the PBDE used to the greatest extent in North America. However, the 24   lower brominated congeners are the most prominent PBDEs found in wildlife and humans. Higher brominated congeners can be debrominated to lower brominated congeners in mammals (Morck et al., 2003). The most abundant PBDE congeners found in biota and demonstrated to elicit the most adverse health effects are those that are components of the commercial pentaBDE mixture DE-71 (http://www.atsdr.cdc.gov/toxprofiles/index.asp). The concentrations of PBDEs in freshwater fish in North America, the greatest user of PBDEs, are high and increasing exponentially in certain species and locations (Hale et al., 2001; Johnson and Olson, 2001; Rayner et al., 2003; Chernyak et al., 2005). The dietary exposure of piscivorous wildlife species such as mink to PBDEs may pose a health risk to these species. While there are numerous studies that have assessed the effects of PBDEs in rodents, relatively few studies have focused on wildlife species. Wildlife species can serve as sentinels of environmental hazards. Laboratory animals reared under controlled conditions are used in determining dose-response relationships and the molecular mechanisms of toxicity. The laboratory studies typically do not give enough information about integrated biologic responses to chronic exposures to real-world mixtures of environmental chemicals. Wildlife species can provide information on the types, characteristics, amounts and bioavailability of pollutants in an environment. Interactive effects of environmental chemicals and the role of the other environmental factors in the final toxicological response are provided by wildlife (Fox et al., 2001). Mammalian wildlife species have physiological systems that are similar to humans and rodents in terms of uptake, distribution, metabolism and distribution of toxicants. Humans and many wildlife species inhabit the same ecosystems and are exposed to the same food sources and pollutants (Basu et al., 2007). Because of these 25   characteristics, the use of wildlife species to examine environmental pollutants can be advantageous. The mink is an excellent wildlife model to address issues related to environmental pollutants. Mink satisfy the requirements of a sentinel species, characteristics of which include: a wide spread distribution, high trophic status, ability to bioaccumulate pollutants, capable of being maintained and studied in captivity, captured in sufficient numbers, restricted home range, well known biology and sensitive to pollutants (Basu et al., 2007). The mink is one of the most widespread carnivores in North America and is generally found in forested regions, especially in areas containing wetlands (Arnold and Fritzell, 1990). Mink can be used in epidemiological studies because of their occurrence across wide geographical areas, which ensures their presence in polluted and non-polluted regions. Mink are abundant and frequently trapped, and they provide data on spatial and temporal trends in environmental pollutants (Basu et al., 2007). The life span of mink in the wild is approximately three years. However, in captivity mink can live up to eight years, permitting pollutants time to bioaccumulate to appreciable concentrations. Wild mink are opportunistic predators that consume a range of prey items available in their local habitat, including small mammals, frogs, snakes and birds (Basu et al., 2007). Fifty percent of a typical mink diet consists of fish, because mink usually forage close to aquatic habitats. The fish in the mink’s diet represents the primary route by which persistent chemicals are accumulated (Chan et al., 2003; Wiener et al., 2003). The mink’s high trophic status and limited home range provides relatively steady-state information on the types, bioavailability, and concentration of pollutants in specific regions. The pollutants measured in mink tissues reflect levels in the local prey base (Cumbie, 1975; Kucera, 1983; Foley et al., 1988). Mink are highly susceptible to many pollutants (Calabrese et al., 1992; Aulerich and Bursian, 1996). Fur ranchers observed that 26   mink consuming fish collected from the Great Lakes during the 1960s failed to reproduce (Aulerich et al., 1971). Subsequent feeding studies demonstrated that polychlorinated biphenyls (PCBs) present in the tissues of the fish collected from the region were the cause of the reproductive problems. Mink can be raised in captivity, allowing derivation of quantitative exposure-response relationships, which makes them a particularly useful model in toxicology. Toxicity data can be compared between wild and captive animals, strengthening environmental risk characterization. Most studies performed in captivity address reproductive and lethal effects of environmental pollutants. Most PBDE toxicity studies are conducted in rodents, but the rodent may not be an adequate model to analyze higher trophic wildlife. Because mink are sensitive to other halogenated environmental contaminants such as PCBs, they were chosen by Environment Canada as a sentinel species to assess the effects of the commercial pentaBDE mixture DE-71. Initially, Martin et al. (2007) conducted a sub-chronic study that assessed the immunotoxicity of DE-71 in ranch mink. Forty 20-week-old male mink were randomly assigned to four treatment groups, with 10 animals per treatment group. Mink were provided feed containing 0, 1.0, 10, or 100 mg DE-71/kg feed for nine weeks. Within the first week, mink on the 100 mg DE-71/kg feed diet rejected the food and were then switched to a diet containing 5 mg DE-71/kg feed for the duration of the trial. Hepatic concentrations of ∑PBDEs (lipid corrected) increased with increasing dietary DE-71 exposure. Predominant congeners detected in the livers of the mink included BDEs 47, 153, 99, 85 and 100. The percent contribution of BDE 47 to ∑PBDEs ranged from 39 to 44 %. Mink exposed to DE-71 did not differ significantly in the phytohemagglutinin (PHA) skin response, relative to control mink, and the PHA-induced response was not associated with individual liver PBDE congener concentrations. Primary and 27   secondary antibody responses of mink in the 5.0 and 10 mg DE-71/kg feed groups exceeded that of control mink, suggesting an alteration, perhaps an up-regulation, in one or more aspects of the antibody response. There was also an increase in the spleen somatic index and greater germinal center development in spleens of the DE-71 treated mink. Hepatic microsomal EROD activity was positively associated with ∑PBDE concentrations and EROD induction correlated with the liver somatic index, reflecting the significantly enlarged livers in treated mink. Mean hematocrit decreased significantly in mink exposed to 10 mg DE-71/kg feed over the course of the trial. A significant decline in mean hematocrit with increasing ∑PBDEs in liver indicated a contaminantinduced effect. The percentage of neutrophils increased and the percentage of lymphocytes decreased significantly in the two highest treatment groups. Results from this study provided evidence for the vulnerability of mink to the immune effects of bioaccumulative PBDE congeners. A chronic trial was then conducted to assess the accumulation, disposition and metabolism of DE-71 and its effects on reproductive performance of female mink and survivability, growth and neurodevelopment of their offspring (Bull et al., 2007; Zhang et al., 2008, 2009). Mink were fed diets containing 0, 0.1, 0.5, or 2.5 mg DE-71/kg feed beginning seven weeks prior to breeding through weaning of kits at six weeks post-parturition. Portions of the offspring were continued on their respective diets until 33 weeks of age. These dietary concentrations bracketed environmentally relevant concentrations. Zhang et al. (2008) evaluated the accumulation, disposition and metabolism of DE-71 in the adults and offspring from the above study. Similar lipid-normalized concentrations of PBDEs were detected in most tissues of adult mink with the exception of the brain. Six-weekold kits had a greater proportion of PBDEs in the brain compared to adults, presumably because 28   of incomplete development of the blood-brain barrier. Lesser brominated congeners were transferred from the mother to the kit and the bulk of the body burden in kits at weaning resulted from lactational rather than transplacental transfer. Lipid-normalized, whole body biomagnification factors ranged from 0.5 to 5.2 for the major congeners and were greatest for BDEs 47 and 153. Hydroxylated PBDEs accounted for 28 to 32% of the excreted fraction, indicating that metabolism was an important elimination pathway. Zhang et al. (2009) presented data from the chronic study related to effects of DE-71 on reproductive performance of adult female mink, survivability and growth of their offspring, histological and biochemical effects in the liver and thyroid, and concentrations of circulating thyroid hormones in adult, kit and juvenile mink. The dietary concentration of 2.5 mg DE-71/kg feed resulted in complete reproductive failure, while reproduction was unaffected at the lesser dietary concentrations. Juvenile mink at 33 weeks of age had disrupted thyroid hormone homeostasis as evidenced by a significant decrease of T3 in males and females exposed to 0.5 mg DE-71/kg feed, despite a compensatory increase in total T4 in females, but not males. Additionally, thyroid follicular epithelium cell height was increased in the 0.5 mg DE-71/kg feed males and females. Ethoxyresorufin-O-deethylase activity was significantly increased in all offspring at 33 weeks of age, but this increase was attribute to polybrominated dioxin, polybrominated furan and/or polybrominated biphenyl impurities in DE-71. Finally, Bull et al. (2007) assessed the effect of DE-71 on cholinergic parameters in adult female mink and their offspring exposed to 0, 0.1, 0.5 or 2.5 mg DE-71/kg feed. Cholinergic parameters, including muscarinic acetylcholine receptor (mAChR) and nicotinic acetylcholine receptor (nAChR) binding, cholinesterase (ChE) activity and acetylcholine (ACh) concentration, were assessed in the cerebral cortex and ChE activity was measured in the plasma. Despite a 29   dose-related increase in brain PBDE concentrations, DE-71 had no significant effect on cortical cholinergic parameters. Concern about the environmental impact and deleterious health effects of some commercial PDBE mixtures has led to discontinuation or restriction of their use. The commercial pentaBDE mixture, whose use was concentrated in North America, and octaBDE were both phased out by the European Union in 2004 and by Canada in 2006. Twelve states in the US have banned at least two commercial PBDEs, and four states (Washington, Maine, Vermont, and Oregon) have banned all three of the commercial PBDE mixtures. In the US, the producer of pentaBDE- and octaBDE voluntarily ceased production in 2004 (Costa and Giordano, 2007). DecaBDE is still the most widely used and cost effective PBDE in the polymer industry, but alternative BFRs have been introduced as replacements for the PBDEs (Ward et al, 2008). Alternatives to PBDEs include hexabromocyclodedecane (HBCDD), decabromodiphenyl ethane (DBDPE), hexabromobenzene (HBB), pentabromoethylbenzene (PBEB) and 1, 2-bis (2,4,6tribromophenoxy) ethane (BTBPE). 1,2-Bis (2,4,6-tribromophenoxy) Ethane 1,2-Bis (2,4,6-tribromophenoxy) ethane (Figure 1.9) has been introduced as a replacement for the commercial octaBDE mixture (Renner, 2004). It is manufactured in the US by Great Lakes Chemical Corporation and is marketed commercially as FireMaster 680 or FF680. The total annual world production was approximately16,710 metric tons in 2001 (Verreault et al., 2007). 1,2-Bis (2,4,6-tribromophenoxy) ethane is used in applications such as the production of plastic materials that require high manufacturing temperatures and light stability such as acrylonitrile-butadinene-polystyrene and high impact polystyrene (Hakk et al., 2004). The chemical structure of BTBPE is similar to that of PBBs. 1,2-Bis (2,4,6-tribromophenoxy) 30   ethane is very hydrop phobic (Kow of 3.14) and, like many BFRs, is ex y xpected to be persistent and w cumulate in the environm (Hakk et al., 2004). t ment e . to bioacc 1,2-Bis (2,4,6 6-tribromoph henoxy) etha was first detected in the environm in the l ane t ment late 1970s (Q et al., 200 Release of BTBPE into the envi Qiu 07). e i ironment ma be throug the ay gh manufact turing process, as well as through va a arious waste streams suc as househ ch hold dust (Karlsson et al., 2007 and dust collected from electronic recycling c n 7) c m c centers (Pette ersson-Julan nderet al., 2004) 1,2-Bis (2,4,6-tribrom ). mophenoxy) ethane, in recent years, h also been detected in e has n n Figure 1.9. Structure of 1,2-bis (2, o ,4,6-tribromo ophenoxy) e ethane (BTB BPE) ambient air and sedim sample from vario sites acro the US (H et al., 2 a ment es ous oss Hoh 2005a; Qiu e al., et 2007). Hoh et al. (20 H 005b) found concentratio of BTBP in the atm ons PE mosphere tha were great at test near its manufacturin source in El Dorado, Arkansas. T observa m ng A This ation suggest that ts atmosphe transpor and deposi eric rt ition is a significant sour of BTBP in the Gre Lakes. rce PE eat 31   Sediment cores collected from Lake Ontario began to have increasing concentrations of BTBPE in the early 1980s. The maximum concentration detected was 6.7 µg/kg dry weight (Xinghua et al., 2007). This concentration was similar to the concentration detected in Lake Michigan sediment cores (7.2 µg/kg dry weight) (Xinghua et al., 2007). However, these concentrations are low compared to concentrations detected in sediment collected near a manufacturing site in Arkansas (470 µg/kg dry weight) (http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=40001GYV.txt). In Southern China, BTBPE concentrations found in dust collected from an electronic waste area were in the range of 14.6 to 232 µg/kg dry weight with a mean value of 107 µg/kg dry weight (Shi et al., 2009). The Shi et al. (2009) study demonstrates that concentrations of BTBPE are increasing in the environment with its increased usage. Because BTBPE is being used to replace the octaBDE, the concentrations will likely increase in the future (Xinghua et al., 2007). A study that examined the exposure of rainbow trout to BTBPE, demonstrated the great potential for uptake of BTBPE in the aquatic food web. However, it was also demonstrated that BTBPE was rapidly degraded and depurated in the rainbow trout (Tomy et al., 2007). Quantifiable concentrations of BTBPE were reported in herring gull eggs collected from seven colonies on the Laurentian Great Lakes from the mid 1990’s to 2006. The BTBPE concentrations found in Lake Superior, Lake Michigan, Channel-Shelter Island (Lake Huron), Chantry Island (Lake Huron), the Detroit River, the Niagara River and Lake Ontario were 0.11, less than 0.06, 0.08, 0.10, 0.20, 0.11, and 0.09 µg/kg wet weight, respectively (Gauthier et al., 2007, 2008). The concentrations of BTBPE found in gull eggs may reflect exposure of adult birds through soil/sediments and atmosphere, or possible dietary exposure in the form of fish such as alewife (Alosa pseudoharengus) and rainbow smelt (Osmerus mordax) present in the Great Lakes 32   aquatic food web (Gauthier and Letcher, 2009). Biotransformation of BTBPE may be greater in herring gulls relative to fish because gulls may possess a greater metabolic capability (Gauthier and Letcher, 2009). In gull egg pools, a number of lower brominated and unidentified compounds were detected, which could have been BTBPE debromination products (Gauthier et al., 2007). Marine mammal monitoring studies have detected BTBPE in blubber from Canadian Arctic beluga whales (Delphinapterus leucas) collected from several sites from 2002 to 2005 (de Wit et al., 2010). The sample concentrations ranged from 0.1 to 2.5 µg/kg lipid weight. 1,2-Bis (2,4,6-tribromophenoxy) ethane was also found in 10% of ringed seal (Phoca hispida) blubber samples collected in five locations in the Canadian Arctic in 2006. The concentrations ranged from less than 0.01 to 0.29 µg/kg lipid weight (Covaci et al., 2011). During preparation of dosing solutions for pharmacokinetic studies, it was observed that BTBPE was virtually insoluble in all common vehicles used for oral dose preparations (Hakk et al., 2004). It was concluded from the solubility studies that mammalian absorption of BTBPE via ingestion would be minimal. Rats were fed 100 to 1000 mg FF-680/kg feed for 28 days. There was some accumulation of the compound in fat, liver, and muscle during the treatment period, but the compound disappeared from the tissues after cessation of dosing (Nomeir et al. 1993). In another study, rats were administered a single oral dose of 14 C-labeled FF-680. Eighty percent of the dose was excreted in the feces and 5% in the urine within 96 hours of administration. At 10 days following dosing, the greatest concentration was in the fat (0.38 mg/kg), whereas the maximum concentration in other tissues was 0.05 mg/kg. A maximum concentration of 0.58 mg/L occurred in the blood at 24 hours post-dosing, decreasing to 0.15 mg/L by 96 hours post-dosing. It was concluded from these studies that FF-680 was absorbed poorly through the rat gastrointestinal tract and it accumulated in the fat (Nomeir et al. 1993). 33   Nomeir et al. (1993) investigated the extent of absorption and disposition of 14 C-labeled FF-680 following dietary administration. On day one, rats were provided dietary concentrations of 0.05, 0.5, and 5% FF-680 and on days two through eleven they were provided 0.05% FF-680. The rational for the concentrations chosen were based on an LD50 for FF-680 that exceeded 10 g/kg of body weight. FF-680 was poorly absorbed through the gastrointestinal tract after the first 24 hours based on the lack of the radioactivity in the major tissues, excretion of less than 1% of the dose in the urine, and excretion of more than 99% of the dose in the feces. After 11 days, the excretion profile was similar to the day-one profile. The parent compound accounted for 94% of the radioactivity excreted in the feces. The gastrointestinal tract had the greatest level of radioactivity, followed by adipose tissue. The adipose tissue had detectable amounts of FF-680 in the one and ten day feed trial. The concentration in adipose tissue was 0.38 ± 10 nmole/g tissue for the 0.05 dose and 1.76 ± 3.45 10 nmole/g tissue for the 5.0 dose of the one-day trial. In the ten-day feed trial, the concentration determined in adipose tissue was 3.19 ± 0.58 nmole/g tissue. This study indicated that FF-680 accumulated in adipose tissue, but the potential for systemic exposure was minimal. Hakk et al. (2004) conducted a study that examined the metabolism, tissue disposition, and excretion of BTBPE in male Sprague-Dawley rats. The radiolabeled BTBPE (2.0 mg/kg body weight) was administered by stomach tube to seven conventional and six bile ductcannulated rats. It initially appeared that the intestinal absorption of BTBPE was low, with cumulative fecal excretion of BTBPE approximating 100% and 94% for conventional and bile duct-cannulated rats, respectively. Approximately 82% of the extractable fecal 14 C was the parent BTBPE compound. Additionally, cumulative biliary excretion of BTBPE was only 34   0.22% of the dose, indicating that hepatic metabolism of BTBPE was minimal. The study by Hakk et al. (2004) demonstrated that 39% of the fecal 14 14 C collected 24 hours after dosing with C-BTBPE was extractable into solvents ranging from anisole to water. This value increased to 44% and 83% at 24 to 48 hours and 48 to 72 hours post-dosing, respectively. The biotransformation of BTBPE, as determined by the metabolites detected in the feces, included a series of oxidations, debrominations and ether cleavages with less than 4% of the 2.0 mg/kg body weight dose undergoing these reactions (Hakk et al., 2004). It was also concluded that biliary and urinary excretion of BTBPE was minimal following oral exposure, resulting in minimal tissue accumulation. Acute toxicity studies indicated an oral LD50 for FF-680 exceeding 10 g/kg body weight for rats and dogs. The dermal LD50 of FF-680 exceeded 2 g/kg body weight for rats and 10 g/kg body weight for rabbits (Nomeir et al., 1993). No compound-related effects were seen in rats fed diets containing up to 10% FF-680 for 14 days. A 28-day dermal toxicity study indicated that rabbits dosed daily with up to 5 g/kg body weight had no clinical signs indicative of toxicity. Rats that inhaled FF-680 at 5 or 20 mg/liter for 21 days had no gross pathological changes. However, unspecified histopathological lesions were observed in the lungs. FF-680 was nonmutagenic in the Ames Salmonella/microsome test system (Nomeir et al., 1993). Environmental persistence, metabolism/elimination and toxicity data suggest that BTBPE is not a major concern. However, because mink are sensitive to DE-71, a kinetic study and reproductive study were conducted in mink to test the hypothesis that BTBPE would not be absorbed to a great extent in mink and as a result, BTBPE would not have adverse effects on reproduction of adults and growth and survivability of offspring. 35   CHAPTER 2 THE PHARMOKINETICS OF 1,2 BIS (2,4,6-TRIBROMOPHENOXY) ETHANE IN MINK Introduction Brominated flame retardants (BFRs) are the largest group of flame retardants used worldwide because of their low cost and high performance efficiency (Birnbaum and Staskal, 2004). As of 2008, an estimated 410,000 metric tons of BFRs were used annually in the global market, which was an increase from 311,000 metric tons in 2005 (Covaci et al., 2011). Brominated flame retardants are incorporated into electronics, clothes, and furniture to prevent or to slow combustion, thus allowing for longer escape time and saving lives. The extensive use of BFRs for a variety of purposes has resulted in increased concentrations of BFRs in the environment, in wildlife, and in humans. The increase in worldwide environmental contamination by BFRs and their presence in biota, even in remote locations, have prompted concern about the potential health effects of these chemicals (deWit, 2002; Law et al., 2006; Burreau et al., 2006; Voorspoels et al., 2007). The polybrominated diphenyl ethers (PBDEs) constitute a major class of BFRs used in industry. An estimated 70,000 metric tons of PBDEs are produced annually worldwide with half being used in the US and Canada (Renner, 2000; Hites, 2006). Polybrominated diphenyl ethers are an important class of BFRs from an environmental and health standpoint because they have been found in the environment and they persist in abiotic compartments, wildlife, and humans (Pijnenburg et al., 1995; de Boer et al., 1998, 2000; de Wit, 2002; Ikonomou et al., 2002). Two of the commercial PBDEs, octaBDE (octa) and pentaBDE (penta), are no longer produced in North America, the European Union, and China because of their presence in the 36   environment and concern about potential health effects in humans and wildlife (Birnbaum and Staskal, 2004; Renner, 2004; Zhou, 2006). In the US, the manufacturer of pentaBDE- and octaBDE voluntarily ceased production of these two PBDE mixtures in 2004 (Hites, 2006) and production throughout North America was phased out by 2005, effectively eliminating their use on this continent (http://www.ec.gc.ca/lcpe-cepa/09F567A7-B1EE-1FEE-73DB-8AE6C1EB7658/sar_pbde-eng.pdf). Penta-brominated diphenyl ether, which was used primarily in North America, and octaBDE were banned by the European Union in 2004, and by several states in the US (California, Maine and Hawaii in 2006 and Washington in 2008). A number of alternative BFRs have been developed to replace PBDEs. One alternative is 1,2-bis (2,4,6-tribromophenoxy) ethane (BTBPE; Figure 1.9), which is marketed under the trade name of FireMaster 680 as a replacement for the commercial octa mixture. Worldwide production and usage of BTBPE was estimated to be 16,710 metric tons in 2001 (Covaci et al., 2011). 1,2-Bis (2,4,6-tribromophenoxy) ethane has been found in the environment, being detected in ambient air and sediment samples from various sites across the US (Hoh et al., 2005a; Qiu et al., 2007). Additionally, BTBPE has been found in the eggs of northern fulmars from the Faroe Islands and in glaucous gulls (Larus hyperboreus) from Svalbard and herring gulls (Larus argentatus) from across the Great Lakes (Karlsson et al., 2006; Verreault et al., 2007; Gauthier and Letcher, 2009). The concentrations found in the herring gull eggs in the Great Lakes region ranged from less than 0.06 to 0.20 µg/kg wet weight (Gauthier et al., 2007, 2008). Marine mammal monitoring studies have detected BTBPE in blubber from Canadian Arctic beluga whales collected from several sites from 2002 to 2005 (de Wit et al., 2010). The sample concentrations ranged from 0.1 to 2.5 µg/kg lipid weight (de Wit et al., 2010). 1,2-Bis 37   (2,4,6-tribromophenoxy) ethane was also found in 10% of ringed seal blubber samples collected from five locations in the Canadian Arctic in 2006. The concentrations ranged from less than 0.01 to 0.29 µg/kg lipid weight (Covaci et al., 2011). With increased usage of BTBPE, there is a potential for environmental concentrations of this BFR to increase (Xinghua et al., 2007). Because BTBPE has been detected in the environment, there are concerns about its potential health effects in wildlife species. Although toxicity studies of BTBPE have been conducted using rodent models (Nomeir et al., 1993; Hakk et al., 2004), the rodent may not be an adequate model for higher trophic wildlife. A recent study that assessed the kinetics of DE-71, which is a commercial pentaBDE mixture, in mink demonstrated that there were major differences in the biotransformation capabilities of mink and rodents (Zhang et al., 2008). Additionally, Zhang et al. (2008, 2009) reported that mink were more sensitive to DE-71 than rodents based on reproductive endpoints and thyroid homeostasis. Because of its sensitivity to a related BFR, it was of interest to determine the effects of BTBPE in mink, which is regarded as a sentinel wildlife species for potential environmental contaminants (Basu et al., 2007). The first study conducted was a pharmacokinetic study that assessed the distribution, metabolism and excretion of BTBPE. Kinetic studies in rodents suggest that BTBPE is absorbed to a limited extent and that the majority of the compounded is excreted unchanged in the feces (Nomeir et al., 1993; Hakk et al., 2004). Thus the hypothesis of this study is that 1,2-bis(tribromophenoxy)ethane will appear in feces unchanged and will not be absorbed, or metabolized to a great extent. Materials and Methods Chemical. 1,2-Bis (2,4,6-tribromophenoxy) ethane (purity greater than 97%, as indicated by the manufacturer) was purchased from Wellington Laboratories (Guelph, Ontario, Canada). 38   Diet preparation. Because of the insolubility of BTBPE, the powder was added directly to other dry ingredients of the standard Michigan State University Experimental Fur Farm diet (Table 2.1). An appropriate quantity of BTBPE was added to a small container containing vitamin premix, trace mineral mix and biotin, which was then tumbled for 30 minutes. Diets were mixed in two batches (453 kg of each diet for batch 1 and 138 to 231 kg for batch 2). The dry ingredients were then added to a 550 kg capacity paddle mixer containing water, oil and phosphoric acid and mixed for one minute. The remaining ingredients of the mink diet were added to the mixer and the feed continued to mix for an additional 20 minutes. Samples of feed were collected as the feed was being dispensed into labeled plastic containers for subsequent freezing. Feed samples were used for proximate analysis and for verification of BTBPE concentration in the feed (National Wildlife Research Centre, Ottawa, Ontario, Canada). Study design. Forty first-year, virgin female, natural dark mink, housed at the MSU Experimental Fur Farm, were randomly assigned to one of two treatment groups (20 animals per group). The two groups consisted of a control group that was fed the standard experimental fur farm diet and a treatment group that was fed the standard diet containing BTBPE at a targeted concentration of 2.0 mg BTBPE/kg feed (actual concentration of 2.3 mg BTBPE/kg feed). Animals were started on their respective treatment diets on January 15, 2009 and exposure continued for 21 days. Feed and water were available ad libitum. Dietary intake of BTBPE was estimated by determining feed consumption for two consecutive days each week. Fecal and urine samples were taken on days 0, 7, 14 and 21. Five animals from both the control and treatment group were euthanized and necropsied at day 7, 14 and 21. Prior to euthanasia, animals were anesthetized with an intramuscular injection of ketamine HCl (30 mg/kg body weight) and blood samples were taken by cardiac puncture. Blood was collected in a Vacutainer tube containing 39   sodium heparin. The blood was gently mixed for two minutes and centrifuged for five minutes at room temperature. The isolated plasma was then stored at -80°C until shipment to the National Wildlife Research Centre for analysis of T3 and T4. After blood collection, animals were euthanized with CO2, weighed and necropsied. The brain, liver and lung were removed and weighed. Samples of brain, liver, lung and abdominal fat were frozen at -80°C for subsequent contaminant analysis. The remaining 10 animals were fed the control diet for an additional 21 days. Urine and fecal samples were taken on day 42. Animals were then anesthetized, sampled for blood, weighed, euthanized and necropsied on day 42 as described above. Chemical analyses. Diet samples were analyzed for BTBPE using a gas chromatography-mass spectrometry (GC-MS) (electron capture negative ionization [ECNI])based method (MET-ORGRES-NEW BFR/PBDE; Revision #2, April, 2010) with minor modifications. 1,2-Bis (2,4,6-tribromophenoxy) ethane and 13C12-BTBPE standard solutions (purity greater than 97%, as indicated by the manufacturer) were purchased from Wellington Laboratories (Guelph, Ontario, Canada). Solvents used for extraction and clean up were OmniSolve quality (EMD Chemicals, Gibbstown, NJ, USA) and suitable for gas chromatography and residue analysis. Diatomaceous earth was heated at 600ºC overnight in a muffle furnace prior to use. Silica was activated at 120°C overnight in an oven and then treated with sulfuric acid (equal weight). The mixture was homogenized for more than 24 hours prior to use. 40   Table 2.1 Composition of standard mink diet Ingredients % of diet 1 Mink cereal 16.65 2 Chicken 21.55 Spray-dried liver 3 6.86 Spray-dried egg3 6.86 4 Spray-dried blood cells 1.47 5 Soybean oil 2.45 Water 35.26 Fish meal 5 6.86 6 Vitamin premix 0.42 7 Trace mineral premix Phosphoric acid, 85% 1.42 8 0.98 9 d-biotin 0.03 10 Larvacide (mL/kg feed) 1.32 mL/kg diet 0, 0.014, 0.13 or 2.3 mg/kg diet 11 1,2-Bis (tribromophenoxy) ethane (BTBPE) 1 2 GnF-20, National Fur Foods, Division of Milk Specialties Co., New Holstein, WI. Fresh whole ground chicken-hens, Michigan State University Poultry Research and Teaching Center, East Lansing, MI. 41   Table 2.1 (con’t) 3 4 5 VanElderen, Inc., Martin, MI. AP301G, APC, Ankeny, IA. North American Nutrition, Lewisburg, OH. 6 Vitamin A, 916,652 IU/kg; vitamin D3, 91,674 IU/kg; vitamin E, 11,000 IU/kg; vitamin K activity, 2200 mg/kg; menadione, 733 mg/kg; vitamin B12, 5.5 mg/kg; riboflavin, 733 mg/kg; dpantothenic acid, 2935 mg/kg; niacin, 4400 mg/kg; thiamine, 7 Calcium, 13.40%; copper, 2000 mg/kg; iodine, 30 mg/kg; iron, 2.0 %; manganese, 2000 mg/kg; selenium, 60 mg/kg; zinc, 2.0 %; Akey, Louisburg, OH. 8 9 Astaris, St. Louis, MO. Biotin 100 (222.2 mg/kg), ADM, Des Moines, IA. 10 Lavadex, active ingredient: cyromazine (N-cyclopropyl-1,2,5-triazine-2,4,5-triamine, 2%), Novatis Animal Health, Greensboro, NC. 11 Kinetic trial: 0 or 2.3 mg/kg diet; reproduction trial: 0, 0.014, 0.13 or 2.3 mg/kg diet. Great Lakes FF-680, Great Lakes Chemical, West Lafayette, IN. 42   The approximate size of the samples selected for analysis was 2.0 g, excluding fat and feces, which were 0.2 and 0.5 g, respectively. These samples were weighed and ground with diatomaceous earth and then spiked with internal standard (IS – see below). The samples were then extracted with 50:50 dichloromethane: n-hexane (DCM: HEX) solution. A 10 % portion of the extract was used for gravimetric lipid determination. For the remaining 90% of the extract, sulfuric acid-impregnated silica (50%) was used to clean up the sample of biogenic material. Acid silica (5.0 g) cartridges were conditioned with successive 4 × 5 mL of 25% DCM:HEX. After the sample was loaded on the cartridge, the analytes were eluted with 4 × 5 mL of 25% DCM:HEX (v/v). The eluent was then concentrated and solvent exchanged to isooctane in preparation for analysis (GC-MS [ECNI]). The IS used was mass-labeled 13C12-BTBPE, which was validated previously as representative for quantification of native BTBPE. A series of calibration standard solutions was prepared, which contained six concentrations of the target compound with a constant concentration of IS (100 µg13C12-BTBPE/kg). These standard solutions were analyzed by GC/MS (ECNI) and the results indicated that there was a highly and linearly correlated 2 concentration versus ECNI response (r > 0.996) over a concentration range of approximately 1.0 µg/kg to 500 µg/kg. Samples were analyzed on an Agilent gas chromatograph 6890 equipped with a 5973 quadrupole mass spectrometer (MS) detector (Agilent Technologies, Mississauga, Ontario, Canada). The injector temperature was set to 240°C. The oven temperature was programmed as follows: 100°C for 2 minutes, 25°C/minute to 325°C, held for 2 minutes. A volume of 1 µL was injected with pulsed-splitless injection mode (injection pulse at 25.0 psi until 1.00 minute; purge 43   flow to split vent of 96.4 mL/minute to 0.80 min; gas save flow of 20 mL/minute at 2.0 minutes). The analytical column was a 15 m × 0.25 mm i.d. DB-5 HT capillary column with a film thickness of 0.10 µm. The ECNI source temperature was 250°C, the quadrupole temperature was 150°C and the transfer line temperature was 280°C. Methane was used as the reagent gas. Selected ion monitoring (SIM) was used for quantification of the target compounds based on the most selective mass fragment under the above mass spectrometry detection operating conditions. 1,2-Bis (2,4,6-tribromophenoxy) ethane was monitored with ions of 251 and 330 (m/z) for di- and tri-bromophenoxy anions, respectively. Ion selection for the internal standard was 257 and 336 (m/z) (for di- and tri-bromophenoxy anions) for 13C12-BTBPE. Quantification for these samples was achieved via Agilent ChemStation Advanced Data Analysis software (Agilent Technologies, Mississauga, Ontario, Canada), based on the internal standard quantification approach. The average recovery of the internal standard varied, depending on the sample matrix under study. Generally, recoveries of 13C12-BTBPE internal standard were 80 ± 32%. An internal standard method was used for quantification, where the concentrations of target compounds were inherently recovery-corrected by internal standard. A method blank was included in each batch of samples (n of 10 to 12) to monitor any potential contamination resulting from the entire analysis process. Blank samples did not contain any detectable (below the method limit of detection [MLOD]) BTBPE, and thus background subtraction was not necessary. A pork liver homogenate was spiked with target compounds (25 ng BTBPE) as a method control (for liver, lung, brain and diet) and olive oil (for fat) was analyzed with each batch of samples to test the recovery and reproducibility. An acceptable 44   coefficient of variation of less than 5 % was obtained. To monitor instrumental reproducibility and response stability, a set of external and internal standards was injected for every eight to 10 sample analyses, with isooctane solvent injected prior to and after analysis of standards. The method limit of quantification (MLOQ) and MLOD were defined as the minimum amount of analyte that produced a peak with a calculated signal to noise ratio of 10 and 3, respectively. Depending on the tissue or sample matrix, the BTBPE MLOQ for the method ranged from 0.16 µg/kg wet weight to 1.36 µg/kg wet weight and the MLOD ranged from 0.05 µg/kg wet weight to 0.46 µg/kg wet weight. Results The target dietary concentration of BTBPE in the present study was 2.0 mg/kg feed and the analyzed concentration was 2.3 mg/kg feed. The concentration of BTBPE in the control diet was below the MLOD. Relative liver, lung and brain weight of mink fed 2.3 mg BTBPE/kg feed for up to 21 days and then placed on clean feed for an additional 21 days were not significantly different compared to control weight (Table 2.2). 1,2-Bis (2,4,6-tribromophenoxy) ethane was detected predominately in the feces with lesser concentrations in adipose tissue of mink fed diets containing 2.3 mg BTBPE/kg feed for 21 days. Minimal concentrations of BTBPE were detected in the liver, brain, lung and urine (Table 2.3). The concentration of BTBPE in the feces was consistent over the 21-day exposure period while the concentration in adipose tissue appeared to reach steady state after the second week of exposure. When animals were placed on the control feed, the concentration of BTBPE in the feces decreased to below the MLOD by day 42, while the concentration in the adipose tissue decreased by almost 70% at day 42 compared to the concentration at day 21. 45   Table 2.2 Mean relative weight of liver, lung, and brain of mink exposed to 2.3 mg 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE)/kg feed 2 Relative organ weight (percent of body weight) 1 Body Weight (g) Liver Lung Brain Day 7 BTBPE 1215.0 (64.5) Control 3.42 (2.59 – 4.24) BTBPE 3.49 (2.79 - 4.19) Control 0.60 (0.53 – 0.67) BTBPE 0.64 (0.57 - 0.70) Control 0.60 (0.48 – 0.72) BTBPE 0.68 (0.63- 0.73) 5 1173.6 (64.5) 1251.2 (64.5) 3.55 (3.12 -3.99) 3.53 (3.11 – 3.94) 0.65 (0.52 –0.78) 0.54 (0.48 – 0.60) 0.71 (0.62 – 0.80) 0.71 (0.61 – 0.82) 21 5 1302.0 (64.5) 1321.8 (64.5) 3.59 (3.09 -4.09) 3.33 (2.98 – 3.69) 0.56 (0.49 – 0.63) 0.52 (0.45 – 0.60) 0.66 (0.56 – 0.77) 0.65 (0.52 -0.77) 42 2 Control 1324.4 (64.5) 14 1 n 5 5 1248.6 (64.5) 1249.2 (64.5) 3.34 (3.17 – 3.50) 3.19 (2.72 – 3.66) 0.69 (0.60 – 0.79) 0.61 (0.52 – 0.69) 0.69 (0.61 – 0.76) 0.62 (0.53 -0.71) Data presented as mean body weight with standard error in parentheses. Data presented as mean with 95% confidence limits in parentheses. 46   Table 2.3 Mean concentration (µg/kg, wet weight) of 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) in various tissues of mink fed 1 diets containing 2.3 mg BTBPE/kg feed for 21 days and then placed on clean feed for 21 days Day n Treatment Liver Lung Brain Fat Feces Urine 7 5 Control not analyzed not analyzed not analyzed 0.996) over a concentration range of approximately 1.0 µg/kg to 500 µg/kg. The GC operation parameters were as follows: injector temperature, 240°C; oven temperature/time, 100°C for 2 minutes, 25°C/minute to 325°C, held for 2 minutes; a volume of 1 µL was injected with pulsed-splitless injection mode (injection pulse at 25.0 psi until 1.00 minute; purge flow to split vent of 96.4 mL/minute to 0.80 minutes; gas save flow of 20 mL/minute at 2.0 minutes); analytical column: 15 m × 0.25 mm i.d. DB-5 HT capillary column with a film thickness of 0.10 µm. The MS operation parameters were as follows: electron capture negative ionization (ECNI) source temperature, 250°C; quadrupole temperature, 150°C; transfer line temperature, 280°C. Methane was used as the reagent gas. Selected ion monitoring (SIM) was used for quantification of the present target compounds based on the most selective mass fragment under the above mass spectrometer detector (MSD) operating conditions. 1, 2-Bis (2,4,6-tribromophenoxy) ethane was monitored with ions of 251 and 330 (m/z) for di- and tribromophenoxy anions, respectively. Ion selection for the internal standard 13C12-BTBPEwas 257 and 336 (m/z) for di- and tribromophenoxy anions. Quantification of these samples was achieved via Agilent ChemStation Advanced Data Analysis software, based on the internal standard quantification approach. The average recovery of the internal standard was 99 ± 11% for fat, 79 ± 21% for feed, and 68 ± 15% for liver. An internal standard method was used for quantification, where the concentrations of target compounds were inherently recovery-corrected by internal standard. 60   A method blank was included in each batch of samples (n =8) to monitor any potential contamination resulting for the entire analysis process. Blank samples did not contain any detectable BTBPE, thus background subtraction was not necessary. A pork liver homogenate, which was spiked with target compounds (25 ng BTBPE) as a method control (for liver and feed) and olive oil (for fat) were analyzed with each batch of samples (n = 8 to 10) to test the recovery and reproducibility. An acceptable coefficient of variation of less than 5 % was obtained. To monitor instrumental reproducibility and response stability, a set of external and internal standards was injected for every 8 to 10 sample analyses, with isooctane solvent injected prior to and after. The method limit of quantification and detection (MLOD) were defined as the minimum amount of analyte that produced a peak with a signal to noise ratio of 10 and 3, respectively. The MLOQ for the method ranged from 0.12 µg/kg wet weight to 0.72 µg/kg wet weight for BTBPE and the MLOD ranged from 0.04 µg/kg wet weight to 0.24 µg/kg wet weight depending on the matrix (feed, liver, adipose tissue). Thyroid Histology. Both thyroid lobes were embedded into single paraffin block using a Tissue Tek VIP 5 tissue processor (Sakura, Torrance, CA, USA). Sections (5 µm) were affixed to glass microscope slides and stained with hematoxylin and eosin. The thyroid was examined for proliferative follicular lesions using a Nikon Eclipse 50i compound microscope (Mississauga, Ontario, Canada). Photomicrographs were taken with an Olympus Qcolor 3 digital camera (Olympus, Richmond Hill, Ontario, Canada) and assessed using ImagePro Express software (version 5.1; Media Cybernetics, Bethesda, MD, USA). A representative thyroid section was used to quantify epithelial cell activity, and cell heights were measured at four cardinal points on all colloid-containing follicles along the axis. 61   Thyroid hormone analysis. The T3 and T4 enzyme-immunoassay protocols were a modification of previously published protocols (Graham et al., 2001). The optimum concentrations and dilutions of antibodies and biotinylated T3 and T4 were determined by checkerboard titration. In brief, Nunc microtiter plates (Thermo Fisher Scientific, Rochester, NY, USA) were coated with T3 or T4 antibody (Sigma-Aldrich, Oakville, Ontario, Canada) dissolved in coating buffer (0.015 mol/L Na2CO3, 0.035 mol/L NaHCO3; pH 9.6) and incubated overnight at room temperature. Coated plates were washed with 0.04% Tween 20 (Sigma-Aldrich, Oakville, Ontario, Canada), 100 µl of assay buffer was added to each well and plates were incubated at room temperature for 1 hr. Then, 50 µl of diluted plasma sample and standards were dispensed into appropriate wells, followed immediately by 100 µl of biotinylated T3 or T4. Plates were incubated overnight at room temperature. After incubation, plates were washed and 200 µl of streptavidin-peroxidase conjugate (1 µL in 22 mL assay buffer; Roche Molecular Biochemicals, Indianapolis, IN, USA) were added to each well. After incubation (45 minutes, room temperature), plates were washed and 200 µl of substrate solution (0.5 µl of 0.016 mol/L tetramethylbenzidine in dimethyl sulfoxide and 100 µl 0.175 mol/L H2O2 diluted in 24 µl 0.01 mol/L C2H3O2Na; pH 5.0) was added to each well. After incubation (45 minutes, room temperature), the enzyme reaction was stopped with 50 µl of stop solution (3 mol/L H2SO4). The optical density was measured at 450 nm. Triiodothyronine and T4 were used as standards and serial dilutions of plasma gave displacement curves parallel to that of the standard curve. 62   Statistical analysis. Reproductive data were analyzed as a completely randomized design using the mixed procedure of SAS (version 9.2; Statistical Analysis Systems, Cary, NC, USA). The body weights for adult females and, juvenile females and males were analyzed as repeated measures. Differences between treatments for adult female and kit body weight and number of live kits at birth, three and six weeks of age were ascertained using least square means with Tukey adjustment. Kit survivability was expressed as a percent and therefore transformed for analysis using arcsine transformation. The means reported reflect the back-transformed data. Since standard error (SE) is not readily back-transformed, indication of variance is expressed as 95% confidence intervals around the mean. Organ weight at necropsy were expressed as percent of body weight and therefore transformed as indicated above. Similar analysis was used for kit growth except the model also contained the effect of sex and treatment by sex interaction. For body weight, fat and liver BTBPE concentrations, and T3 and T4 concentrations, least square means utilizing Tukey adjustment were used to ascertain differences between treatments. Differences between treatments were considered to be significant at p < 0.05. Results Dietary BTBPE concentrations. The targeted dietary concentrations of BTBPE were 0.0, 0.02, 0.2, and 2.0 mg/kg feed, wet weight. The analyzed dietary concentrations were 0.014, 0.13 and 2.3 mg/kg feed, wet weight. The MLOQ and MLOD were 0.12 and 0.04 µg/kg wet weight respectively. Table 3.1 presents the analyzed dietary concentrations of BTBPE. Effects of BTBPE on adult female feed consumption. The daily estimated feed intake was not significantly different between the control and 2.3 mg BTBPE/kg feed treatment groups (the only groups assessed) from January 23 to February 10, 2009 (Table 3.2). The data represents prebreeding feed consumption of adult females. 63   Table 3.1 Dietary concentrations of 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) Targeted dietary BTBPE concentration (mg/kg feed, wet weight) (mg/kg feed, wet weight) 0.0 < MLOD 0.02 0.014 0.2 0.13 2.0 1 Analyzed dietary concentration 2.3 1 Concentration below the method limit of detection (MLOD). Table 3.2 Average daily feed intake (ADFI) of adult female mink fed 0.0 or 2.3 mg 1, 2 bis (2, 4, 6-tribromophenoxy) ethane (BTBPE)/kg feed from January 23 to February 10, 2009 Dietary BTBPE concentration (mg/kg feed, wet weight) ADFI(g/d) 0.0 139 2.3 138 SE 1.6 p-value 0.6549 64   Reproduction, offspring survivability, body weight, and organ weight. There were no significant differences in the number of females whelping, the total number of kits whelped or the number of live kits whelped per litter. Kit survivability through six weeks of age was not adversely affected by exposure to in utero and lactational exposure to BTBPE (Table 3.3). There were no statistically significant treatment-related differences in adult pre-breeding body weights (Table 3.4), maternal and kit body weights at whelping and at three and six weeks of age (Table 3.5), and juvenile female body weights through 27 weeks of age (Table 3.7). Body weights of 27-week-old juvenile males exposed to BTBPE were significantly less compared to controls (Tables 3.6). At weaning, relative liver weight of adult control animals was significantly greater compared to relative liver weight of adult females in the 2.3 mg BTBPE/kg feed group (p = 0.0319). There were no other significant differences in relative organ weight of adult females (Table 3.8). Relative liver weight of six-week-old female kits was significantly greater in the 0.014 mg BTBPE/kg feed group compared to relative liver weight of animals in the 2.3 mg BTBPE/kg feed group (p = 0.0198). There were no other significant changes in relative organ weight of sixweek-old females (Table 3.9), nor were there significant differences in relative organ weight of six-week-old male kits (Table 3.10). Relative spleen weight of juvenile females was less in the 0.014 mg BTBPE/kg feed group in comparison to relative spleen weight of juvenile females in the 0.13 mg BTBPE/kg feed group (p = 0.0526). There were no other significant effects for relative weight of other organs for juvenile females (Table 3.11). In juvenile males, relative weight of the adrenal glands in the 0.13 mg BTBPE/kg feed group was significantly less compared to the control group 65   Table 3.3 Effect of 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on mink reproduction Dietary BTBPE concentration (mg/kg feed, wet weight) 0 % % # Females Survivability Survivability whelping/# % Total kits Live kits Survivability at 3 weeks at 6 weeks females 1 1 2 2 whelped whelped of age of age bred at birth p-value 1 2 10/10 87.4 86.1 (0.76) (83.7 -104) (68.0 – 107) (67.1 – 105) 7.0 5.4 80.7 88.3 88.3 (0.76) (62.0 - 99.3) (74.0 -103) (74.0 – 103) 7.7 6.8 88.9 93.7 92.7 (0.68) (74.7 -103) (85.9 – 101) (83.1 – 102) 8.9 7.6 86.0 81.8 81.8 (0.74) 2.3 10/10 93.7 (0.70) 0.13 7.4 (0.78) 8/10 7.9 (0.78) 0.014 8/10 (0.72) (73.0 - 98.9) (68.2 - 95.3) (68.2 – 95.3) 0.3774 0.1814 0.5610 0.3696 0.4215 Data presented as mean with the standard error in parenthesis. Data presented as mean with the 95% confidence interval in parenthesis. 66   Table 3.4 Effect of 1,2 bis (2, 4, 6-tribromophenoxy) ethane (BTBPE) on adult female mink 1 pre-breeding weight Weight (g) Dietary BTBPE concentration (mg/kg feed, wet weight) n Initial 30 d on trial 60 d on trial 0.0 10 1271 1378 1306 (52.4) (51.6) (55.9) 1285 1398 1303 (52.4) (51.6) (55.9) 1278 1377 1295 (55.2) (54.4) (58.9) 1276 1364 1295 (52.4) (51.6) (55.9) 0.014 0.13 2.3 p-value treatment p-value time p-value treatment*time 1 10 9 10 0.9966 < 0.0001 0.6557 Data are least square means (lsmeans) with the standard error in parenthesis. 67   Table 3.5 Effect of 1,2 bis (2, 4, 6-tribromophenoxy) ethane (BTBPE) on adult female and kit 1 body weights (g) at whelping and at three and six weeks post-whelping Targeted dietary BTBPE concentration (mg/kg feed, wet weight) 0.0 1 1255 1079 966 9.5 118 311 (46) (44) (0.6) (8) (23) 1300 1136 1013 9.5 118 308 (46) (44) (0.6) (8) (23) 1294 1135 985 9.5 126 340 (41) (39) (0.5) (7) (21) 1290 1111 1018 9.7 122 311 (50) p-value 10 (48) 2.3 Kits at whelping (53) 0.13 Adult females at 6 weeks postwhelping (53) 0.014 n Adult females at whelping Adult females at 3 weeks postwhelping (43) (41) (0.6) (8) (22) 0.9294 0.7854 0.8138 0.9886 0.8245 0.6702 10 9 10 Data presented as least square means with the standard error in parenthesis. 68   Kits at 3 Kits at 6 weeks of weeks of age age 1 Table 3.6 Effect of 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on body weight (g) of juvenile males Age Dietary BTBPE concentration (mg/kg feed, wet weight) n 10 weeks 14 weeks 19 weeks 23 weeks 27 weeks 0.0 10 1137 1670 2098 2283 2369 (65.8) (60.4) (60.4) (68.9) (56.7) 974 1490 1841 1988 1961 (65.8) (60.4) (60.4) (68.9) (56.7) 1103 1655 2029 2153 2143 (65.8) (60.4) (60.4) (68.9) (56.7) 966 1544 1945 2154 2087 (65.8) (60.4) (60.4) (68.9) (56.7) 0.014 0.13 2.3 10 10 10 p-value treatment < 0.0166 p-value time < 0.0001 p-value treatment*time 1 0.2281 Data are presented as least square means with the standard error in parenthesis. 69   a b b b 1 Table 3.7 Effect of 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on body weights (g) of juvenile females Age Dietary BTBPE concentration (mg/kg feed, wet weight) n 10 weeks 14 weeks 19 weeks 23 weeks 27 weeks 0.0 10 797 1145 1358 1426 1447 (45.2) (75.8) (116.7) (128.3) (145.5) 713 1057 1265 1384 1405 (45.2) (75.8) (116.7) (128.3) (145.5) 722 1141 1389 1456 1459 (45.2) (75.8) (116.7) (128.3) (145.5) 743 1036 1189 1250 1217 (45.2) (75.8) (116.7) (128.3) (145.5) 0.014 0.13 2.3 p-value treatment p-value time p-value treatment*time 1 10 10 10 0.2580 < 0.0001 0.0996 Data are presented as least square means. Standard error for all lsmeans was 66.4 70   1 Table 3.8 Effects of dietary 1,2 bis (2, 4, 6-tribromophenoxy) ethane (BTBPE) on adult female mink relative organ weights (% of body weight) Targeted dietary BTBPE concentration (mg/kg feed, wet weight) n Body weight at necropsy (g) Liver a Thyroid gland Heart Spleen Adrenal glands Kidneys Brain 0.014 10 10 842 4.1 0.0080 0.82 0.40 0.013 0.87 0.98 (40) 0.0 (3.6-4.5) (0.0070-0.0090) (0.77-0.88) (0.30-0.50) (0.0090-0.017) (0.78-0.95) (0.87-1.1) 0.0070 0.78 0.33 0.011 0.80 0.96 (0.0050-0.0080) (0.69-0.85) (0.24-0.41) (0.0090-0.013) (0.71-0.89) (0.85-1.1) 0.0080 0.77 0.34 0.011 0.80 0.95 (0.0060-0.0090) (0.72-0.82) (0.29-0.39) (0.010-0.012) (0.69-0.91) (0.87-1.0) 0.0070 0.80 0.36 0.010 0.78 0.94 864 (38) 0.13 10 852 ab 3.7 (3.2-4.2) ab 3.6 (38) p-value 10 871 3.3 (38) 2.3 (3.2-4.0) (2.9-3.6) (0.0060-0.0090) (0.74-0.86) (0.29-0.44) (0.0080-0.011) (0.72-0.83) (0.84-1.0) 0.9512 0.0526 0.4064 0.4603 0.4057 0.2253 0.4338 0.9204 b 1 Data are presented as means with standard error (body weight) or 95% confidence interval (relative organ weight) beneath in parenthesis. ab Values with different superscripts within the same column are significantly different (p < 0.05). 71   1 Table 3.9 Effects of dietary 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on female kit relative organ weights (% of body weight) at 6 weeks of age Dietary BTBPE concentration (mg/kg feed, wet weight) 0.0 n 5 Body weight (g) 393 Liver ab 5.0 (34) 0.13 5 5 359 5.5 (34) 0.014 (4.1-5.9) (5.0-6.1) 468 a ab 4.7 Thyroid gland Heart Spleen Adrenal glands Kidneys 0.0090 0.65 0.47 0.014 0.87 (0.0080-0.010) (0.53-0.78) (0.39-0.54) (0.010-0.018) (0.79-0.95) 0.0090 0.60 0.53 0.011 0.93 (0.0060-0.012) (0.50-0.70) (0.41-0.64) (0.008-0.014) (0.84-1.03) 0.0080 0.64 0.55 0.012 0.87 (0.0060-0.010) (0.56-0.71) (0.42-0.68) (0.0090-0.014) (0.75-0.98) 0.0080 0.62 0.47 0.011 0.86 (34) p-value 5 413 4.5 (34) 2.3 (4.8-5.0) (4.2-4.9) (0.0050-0.011) (0.60-0.64) (0.33-0.60) (0.0080-0.013) (0.83-0.90) 0.1795 0.0198 0.8577 0.6530 0.3962 0.2870 0.3520 b 1 Data are presented as means with standard error (body weight) or 95% confidence interval (relative organ weight) beneath in parenthesis. ab Values with different superscripts within the same column are significantly different (p < 0.05). 72   1 Table 3.10 Effects of dietary 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on male kit relative organ weights (% of body weight) at 6 weeks of age Dietary BTBPE concentration (mg/kg feed, wet weight) n 0.0 Body weight (g) Liver Thyroid Heart 5 457 (59) 5.5 (5.2-5.8) 0.0090 (0.0060-0.011) 0.014 5 404 (59) 5.4 (4.4-6.4) 0.13 5 494 (59) 2.3 5 p-value Spleen Adrenal glands Kidneys 0.60 (0.54-0.66) 0.45 (0.42-0.48) 0.016 (0.012-0.021) 1.0 (0.91-1.1) 0.0080 (0.0060-0.010) 0.59 (0.49-0.70) 0.49 (0.43-0.55) 0.014 (0.012-0.016) 1.0 (0.94-1.1) 5.1 (4.4-5.9) 0.0080 (0.0060-0.011) 0.62 (0.50-0.73) 0.55 (0.34-0.76) 0.014 (0.011-0.018) 0.91 (0.78-1.0) 488 (59) 5.1 (3.5-6.7) 0.0080 (0.0040-0.011) 0.65 (0.57-0.73) 0.54 (0.39-0.69) 0.0161 (0.002-0.030) 1.0 (0.78-1.2) 0.7005 0.8236 0.7572 0.6538 0.4134 0.9084 0.3932 1 Data are presented as means with standard error (body weight) or 95% confidence interval (relative organ weight) beneath in parenthesis. 73   1 Table 3.11 Effects of dietary 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on juvenile female relative organ weights (% of body weight) at 27 weeks of age Dietary TBPE concentration (mg/kg feed, wet weight) n Body weight (g) Liver Thyroid gland Heart Spleen ab Adrenal glands Reproductive tract Kidneys Brain 0.005 (0.0030.007) 0.52 (0.460.58) 0.62 (0.470.77) 0.07 (0.01-0.12) 0.006 (0.0040.008) 0.53 (0.490.58) 0.61 (0.510.71) 0.07 (0.03-0.11) 0.006 (0.0040.007) 0.54 (0.440.64) 0.58 (0.500.66) 0.10 (0.05-0.15) 0.0 5 1447 (86) 3.2 (2.6-3.8) 0.004 (0.0030.005) 0.50 (0.420.58) 0.21 (0.170.26) 0.014 5 1405 (86) 3.4 (2.9-3.9) 0.004 (0.0030.006) 0.54 (0.470.62) 0.20 (0.180.21) 0.013 5 1459 (86) 3.7 (3.1-4.2) 0.004 (0.0030.006) 0.58 (0.420.75) 0.25 (0.220.29) 2.3 5 1217 (86) 3.2 (2.9-3.5) 0.004 (0.0030.006) 0.62 (0.530.71) 0.22 (0.180.27) 0.007 (0.0050.010) 0.56 (0.510.60) 0.74 (0.620.86) 0.07 (0.04-0.10) 0.2080 0.2778 0.9130 0.1882 0.0526 0.2096 0.7898 0.0776 0.3569 p-value 1 a b ab Data are presented as means with standard error (body weight) or 95% confidence interval (relative organ weight) beneath in parenthesis. ab Values with different superscripts within the same column are significantly different (p < 0.05). 74   (p = 0.0320). The relative brain weight of juvenile males in the 0.13 mg BTBPE/kg feed group was significantly less than relative brain weight of animals in the 0.014 (p = 0.0127) and 2.3 mg BTBPE/kg feed groups (p = 0.0105) and relative brain weight of male juveniles in the 2.3 mg BTBPE/kg feed group was significantly greater (p = 0.0022) compared to controls (Table 3.12). Thyroid histology and hormone concentrations of juveniles. The thyroid follicles were round to ovoid, and were comprised of a single layer of cuboidal to columnar epithelial cells surrounding a colloid-filled lumen. Colloid vacuolation was absent to moderate, with no apparent difference among treatment groups (Figure 3.1A). In most of the mink examined, diffuse follicular cell hyperplasia was evident in one or both lobes of the thyroid; incidence and severity did not vary among treatment groups. Hyperplastic areas exhibited microfollicular architecture as well as larger irregular-shaped follicles with papillary in-folding and relatively sparse, lightstaining colloid. These areas were poorly demarcated, non-encapsulated, and lacked cellular atypia (Figure 3.1B). Cystic hyperplasia was seen in a single individual in the 0.014 mg BTBPE/kg feed group. This lesion formed a large discrete nodule comprised of low cuboidal epithelial cells with papillary projections surrounding a distended colloid-filled lumen (Figure 3.1C). There was no evidence of neoplasia in any groups. Mean epithelial cell height was significantly greater in the mink exposed to 0.13 mg BTBPE/kg feed compared to controls (p = 0.043). Dietary exposure to BTBPE resulted in significant changes in plasma T3 and T4 concentrations in adult female mink. Mean T3 concentration was significantly greater in animals in the 2.3 mg BTBPE/kg feed group compared to animals in the other groups and mean T4 concentration was significantly less in adult females in the 0.014 mg BTBPE/kg feed group 75   1 Table 3.12 Effects of dietary 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on juvenile male relative organ weights (% of body weight) at 27 weeks of age Dietary BTBPE concentration (mg/kg feed, wet weight) n Body weight (g) a 0.0 5 2369 (57) 0.014 5 1961 (57) 0.13 5 2143 (57) 2.3 5 2087 (57) p-value 1 b b b 0.0010 Liver Thyroid gland Heart a 3.2 (2.9-3.6) 0.004 (0.0020.005) 0.60 (0.480.73) 0.17 (0.120.22) 0.004 (0.0030.006) 3.2 (2.6-3.8) 0.003 (0.0020.004) 0.55 (0.460.64) 0.16 (0.120.19) 0.005 (0.0050.006) 3.0 (2.3-3.7) 0.003 (0.0030.003) 0.65 (0.530.77) 0.16 (0.140.19) 0.006 (0.0050.006) 2.8 (2.4-3.3) 0.003 (0.0020.003) 0.58 (0.490.66) 0.17 (0.120.22) 0.5045 0.2497 0.3479 0.8700 Kidneys Brain ac Testes 0.56 (0.500.63) 0.43 (0.390.48) 0.57 (0.500.63) 0.50 (0.450.54) 0.56 (0.450.68) 0.42 (0.370.46) 0.005 (0.0050.006) 0.52 (0.480.56) 0.52 (0.470.56) 0.10 (0.040.17) 0.0359 0.6015 0.0010 0.6662 ab b ab bc a b 0.10 (0.060.14) 0.09 (0.030.14) 0.12 (0.080.16) Data are presented as means with standard error (body weight) or 95% confidence interval (relative organ weight) in parenthesis. abc Values with different superscripts within the same column are significantly different (p < 0.05). 76   Spleen Adrenal glands A  B  C  f BPE ology of juve enile mink t thyroid gland (A) Cont ds. trol Figure 3.1 Effect of dietary BTB on histo (200x): Ovoid follicl are comprised of cuboidal epithel O les lium surroun nding eosino ophilic colloid. (B) 2.3 mg BTBPE/k feed (100x Diffuse follicular ce hyperplas showing microfollicu m kg x): ell sia, ular architectu and papi ure illary folding of epitheliu (upper le adjacent to normal f g um eft), t follicles. Note that this condition oc c ccurred in all treatment groups. (C) 0 l g 0.014 mg BT TBPE/kg fee (100x): C ed Cystic follicular hyperplasia with a larg discrete nodule compr r a; ge n rised of low cuboidal ep w pithelial cells s with papi illary projections surrou unding a diste ended colloi filled lumen. id 77   compared to females in the other groups (Table 3.13). Hepatic and adipose BTBPE concentrations. 1,2 Bis (2,4,6-tribromophenoxy) ethane accumulated to a greater extent in the adipose tissue compared to the liver, although concentrations in both tissues were from one to several orders of magnitude less than dietary concentrations. Hepatic concentrations of BTBPE in exposed adult females and kits were similar to control concentrations (data not presented). In 27–week-old juveniles, mean hepatic concentration of BTBPE in the females fed 2.3 mg BTBPE/kg feed was significantly greater compared to mean concentrations in the other groups (hepatic concentration of BTBPE was not determined for the 0.014 mg/kg feed group) (Table 3.14). 1,2Bis (2,4,6-tribromophenoxy) ethane was quantifiable in the adipose tissue of juveniles. Juveniles exposed to 2.3 mg BTBPE/kg feed had a significantly greater mean concentration of BTBPE in their adipose tissue than did juveniles in the other treatment groups (Table 3.15). The bioaccumulation factor (ratio of the tissue concentration of BTBPE and the concentration of BTBPE in the feed) was approximately 0.1 for the 2.3 mg BTBPE/kg feed group. Discussion 1,2bis (2,4,6-tribromophenoxy) ethane had no adverse effects on mink feed consumption (Table 3.2). Similarly, rats exposed to BTBPE at dietary concentrations of 0.05, 0.5, and 5% for one day or 10 days had feed consumption that was not different than feed consumption of control animals (Nomeir et al., 1993). In a DE-71 mink study, adult females were fed diets containing 0, 0.1, 0.5, or 2.5 mg DE-71/kg feed. Feed consumption was significantly lower in the 0.5 and 2.5 mg DE-71/kg feed treatments compared to controls (Zhang et al., 2009). In another DE-71 mink study, adult male mink were fed diets containing 0, 1, 5, or 10 mg DE-71/kg feed for nine weeks 78   Table 3.13 Effects of dietary 1,2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) on plasma triiodothyronine (T3) and thyroxine (T4) concentrations in adult female mink Dietary BTBPE concentration (mg/kg feed, wet weight) 0.0 19.8 0.014 18.9 0.13 18.3 2.3 44.8 101 SE 6.7 10 p-value 1 T3 (ng/ml) 0.0196 0.0051 1 1 T4 (ng/ml) a a 111 a b 61.1 a a 103 b a Data are presented as lsmeans. ab Values with different superscripts within the same column are significantly different (p < 0.05). 79   Table 3.14 Hepatic concentrations (µg/kg, wet weight) of 1,2 bis (2,4,61 tribromophenoxy) ethane (BTBPE) in juvenile mink Dietary BTBPE concentration (mg/kg feed, wet weight) Male Female 0.0 0.035 (0.025) 0.035 (0.025) 0.077 (0.033) 0.21 0.07 0.21 0.07 0.28 0.09 0.035 (0.025) 2.3 MLOD µg/kg wet weight 0.035 (0.025) 0.13 MLOQ µg/kg wet weight 0.230 (0.033) a a p-value treatment 0.0010 p-value sex 0.0381 p-value treatment * sex b 0.0306 1 Data are presented as lsmeans with standard error in parenthesis. ab Values with different superscripts within the same column are significantly different (p < 0.05). 80   Table 3.15 Adipose concentrations (µg/kg, wet weight) of 1, 2 bis (2,4,6-tribromophenoxy) ethane (BTBPE) in juvenile mink Dietary BTBPE concentration (mg/kg feed, wet weight) 0.0 0.012 0.014 1.5 0.13 11 2.3 212 Standard error 9.01 p-value 1 Adipose concentration < 0.0001 a a a b Data are presented as lsmeans. ab Values with different superscripts within the same column are significantly different (p< 0.05). 81   (Martin et al., 2007). The animals in the 10 mg DE-71/kg feed group had a slight decline in feed consumption while feed consumption in the other groups was similar. 1,2 bis (2,4,6-tribromophenoxy) ethane also had no significant effect on adult mink reproduction as assessed by the number of females whelping, litter size and kit survivability through six weeks of age (Table 3.3). This is in contrast to the deleterious reproductive effects induced by dietary exposure of adult female mink to comparable concentrations of DE-71, a commercial pentaBDE mixture (Zhang et al., 2009). Female mink fed a diet containing 2.5 mg DE-71/kg feed experienced complete reproductive failure, while females fed diets containing 0.5 mg/kg feed or less had average whelping rates of 70 to 90%. The presence of implantation sites in the uteri of the females fed 2.5 mg DE-71/kg feed suggested that the animals were able to conceive and that DE-71 was inducing fetotoxic effects during gestation. Body weight of adult female mink, neonatal, lactating and weanling kits, and juvenile mink generally were not adversely affected by exposure to BTBPE (Tables 3.4-3.7) with the exception of male juveniles at 27 weeks of age. At this time point, body weights of all males fed diets containing BTBPE were significantly less than control male body weights (Table 3.12). Studies conducted with BTBPE in rats did not assess body weight (Nomeir et al., 1993; Hakk et al., 2004), so comparisons cannot be made for the compound. However, studies assessing the effects of DE-71 in mink have included body weight as an endpoint. Zhang et al. (2009) reported that body weight of adult female mink fed diets containing 0, 0.1, 0.5, or 2.5 mg DE71/kg feed were not affected at time of whelping or at six weeks post-whelping, despite the fact that feed intake was significantly lower in the 0.5 and 2.5 mg DE-71/kg feed groups. Additionally, no significant differences were reported in offspring body weight through 33 weeks of age. Martin et al. (2007) fed adult male mink diets containing DE-71 at concentrations 82   of 0, 1.0, 5.0, or 10 mg/kg feed for nine weeks. Mink from groups exposed to 5.0 and 10 mg DE-71/kg feed had significant reductions in body weight. Exposure to BTBPE resulted in a few differences in relative organ weight in the present study, although the effects were not generally dose related. The only effect that could be considered treatment related was a decrease in mean relative liver weight of adult females in the 2.3 mg BTBPE/kg feed group compared to controls. Organ weight was not evaluated in the rodent BTBPE rodent study (Nomeir et al., 1993; Hakk et al., 2004). The DE-71 mink reproduction study performed by Zhang et al. (2009) did evaluate organ weight. Adult female mink were fed diets containing 0, 0.1, 0.5, or 2.5 mg DE-71/kg feed and the liver somatic index was significantly greater in the 2.5 mg DE-71/kg feed treatment group compared to other treatment groups. Additionally, there was a significant increase in liver somatic index of juvenile mink in the 0.5 mg DE-71/kg feed treatment group (there were no offspring in the 2.5 mg DE-71/kg feed group). Martin et al. (2007) also reported changes in organ weight in mink as a result of dietary exposure to DE-71. Adult male mink exposed to 10 mg DE-71/kg feed had significantly greater mean relative spleen and absolute adrenal gland weight compared to the control group. Additionally, relative liver weight of males in the 5.0 and 10 mg DE-71/kg feed groups were significantly greater compared to controls, which is opposite to the effect on liver weight reported in the present study. The effects of BTBPE on thyroid hormone concentrations have not been reported in mammals, however, the similarity in structure between BTBPE and DE-71 suggested that BTBPE could be a disrupter of the thyroid gland such has been reported for DE-71 in mink (Zhang et al., 2009). In the latter study, plasma T3 concentrations were decreased and plasma T4 concentrations were increased as a result of exposure to DE-71, suggesting a disruption in the 83   conversion of T4 to the metabolically active T3. In the present study, mean plasma T3 concentration was significantly greater in adult females in the 2.3 mg BTBPE/kg feed group compared to other treatment groups while mean plasma T4 concentration was significantly lower in the 0.014 mg BTBP/kg feed group compared to the other treatment groups. Increased plasma T3 is related to hyperthyroidism, the clinical signs of which can include increased appetite and decreased body weight. Because the females in the 2.3 mg BTBPE/kg feed group did not have significant weight loss or increased feed intake in comparison to females in the other treatment groups, it is assumed that the increase reported here was transitory and probably not treatment related. A decrease in plasma T4 may indicate hypothyroidism, a clinical sign of which is an increase in body weight. Because the decrease in T4 reported here occurred only at the lowest dose and because an increase in body weight was not reported for mink in this treatment group, it is likely that the change in T4 was transitory and not treatment related. A lack of treatmentrelated histological changes in the thyroid gland of mink exposed to BTBPE support the conclusion that BTBPE, at the concentrations fed, has no effect on the thyroid gland of mink. The potential of BTBPE to be a thyroid axis disrupter has been assessed in rainbow trout. Dietary exposure to BTBPE had no effect on thyroid gland morphology, deiodinase activity that results in conversion of T4 to T3, or plasma concentrations of T3 and T4. These results suggest that BTBPE has no effect on the thyroid gland of the rainbow trout, supporting the conclusion that BTBPE did not affect the thyroid gland of the mink. The histological assessment of the juvenile thyroid glands in the present study indicated a significant increase in thyroid epithelial cell height in the 0.014 mg BTBPE/kg feed group 84   compared to controls. An increase in thyroid epithelial cell height can suggest toxic effects on the thyroid gland. A dose-related increased in thyroid cell height was reported in mink fed diets containing DE-71 (Zhang et al., 2009). Because the changes in plasma thyroid hormone concentrations and thyroid cell morphology were not dose-related, it is difficult to make conclusive statements regarding the effects of BTBPE on thyroid homeostasis. 1,2bis (2,4,6-tribromophenoxy) ethane was detected in the liver and was predominant in adipose tissue of juvenile mink, although at concentrations less than dietary concentrations. These results are similar to those reported in a BTBPE toxicokinetic study with rodents that indicated that little of the compound is retained in the animal and that adipose tissue is the preferred site of accumulation for that portion that is not eliminated (Nomeir et al., 1993). In a 10-day study, BTBPE was fed to rats at 0.05% of the diet daily. At the end of the study, the concentrations of BTBPE were highest in adipose tissue (0.06% of dose), followed by kidney, skin, and thymus, which had a minimal percentage of the dose (< 0.01%), suggesting that the potential for systemic exposure by ingestion was minimal in rats (Nomeir et al., 1993). Similarly, Hakk et al. (2004) administered a single oral dose of 2.0 mg BTBPE/kg body weight to rats and reported minimal tissue retention of BTBPE, with the greatest concentrations occurring in the lipophilic tissues such as adipose tissue, the adrenal glands and thymus. Conclusion The results of the present study indicated that exposure to BTBPE at dietary concentrations up to 2.3 mg/kg feed had no effect on reproductive performance of mink and survivability and growth of kits and growth of juvenile females. Body weights of juvenile males exposed to BTBPE were significantly less compared to controls at 27 weeks of age. Exposure to BTBPE had no significant effect on adult female body or organ weights, with the exception of 85   reduced relative liver weight in the 2.3 mg BTBPE/kg feed group. The significant differences seen in female kits, and juvenile female organ weights were not treatment related. The significant differences seen in juvenile male organ weight were not treatment related. However the significant effect seen in juvenile male body weights could be treatment related, and should be studied further. Histological assessment of juvenile thyroid glands indicated increased thyroid epithelial height at 0.014 mg BTBPE/kg feed. 1,2 Bis (2,4,6-tribromophenoxy) ethane affected thyroid hormone homeostasis, as indicated by an increase in adult female plasma T3 concentrations. 1,2 Bis (2,4,6-tribromophenoxy) ethane was measureable in adipose tissue, but not in the liver, suggesting that BTBPE is effectively cleared from the liver, but can accumulate in the fat. With BTBPE increased usage and the ability for it to accumulate in the environment, effects of higher doses may need to be assessed. 86   CHAPTER 4 CONCLUSION 1,2 bis (tribromphenoxy)-ethane (BTBPE) is considered to be an alternative flame retardant that is currently being used as a replacement for the commercial octaBDE. Assessment of environmental concentrations suggests that BTBPE has not accumulated to the same extent as the PBDEs. However since BTBPE is being marketed as a replacement for octaBDE, there is the potential that environmental concentrations will increase. The results from the present studies were comparable to those reported for rainbow trout exposed to BTBPE. Tomy et al. (2007) determined that BTBPE was rapidly degraded and depurated in the fish. The present mink studies with BTBPE showed that the compound was rapidly eliminated at concentrations expected to occur in the environment. At higher doses of BTBPE, concentrations would be expected to increase in the adipose tissue, and potentially could cause adverse effects. Results of the present pharmacokinetic mink study were similar to those reported in previous rodent studies (Nomeir et al., 1993), which indicated that the majority of orally administered BTBPE was excreted in feces. Lipophilic tissues of the mink, such as the adipose tissue, can accumulate BTBPE in small amounts, which is similar to the effects observed in rats (Nomeir et al., 1993; Hakk et al., 2004). Because of the minimal amount of absorption of BTBPE in the mink, few if any adverse effects would be expected as a result of prolonged exposure to this chemical, at the feed concentrations evaluated in the present studies. The results of the present mink reproduction study confirmed few adverse effects at the feed concentrations examined. Because there were no changes in feed consumption, reproductive 87   parameters, body weights of adults and kits, juvenile female body weights or organ weights of adults and offspring that were considered to be treatment related, it was concluded that BTBPE demonstrated no toxicological effects. Regardless of the exposure concentration in the present studies, BTBPE was effectively cleared from the liver, and only small amounts accumulated in the adipose tissue. 1,2 bis (tribromphenoxy)-ethane affected thyroid hormone homeostasis, as indicated by an increase in adult female T3 concentrations. The change in T3 concentrations may suggest that exposure to BTBPE can result in subtle endocrine effects that should be studied further. In future studies, an increase in dietary concentrations must be assessed as the concentration of BTBPE is increasing in the environment. The present reproductive BTBPE study in mink showed a significant effect on T3 concentration. Because deiodination and conjugation are principle pathways for thyroid hormone metabolism, hepatic T4 glucuronidation should be examined using higher dietary concentrations. Additionally, the effect of higher dietary concentrations on hepatic microsomal EROD and PROD, UDPGT activity must be assessed. Disruption of thyroid homeostasis, can cause effects on body weight and brain development. A decrease in body weight of juvenile males was reported in the reproduction study. The decrease in body weight at the highest dietary concentration compared to the control should be investigated further. The decrease in body weight was a significant effect seen in the reproductive study, but an obvious reason for the effect could not be determined since feed consumption was determined for the juveniles. Additionally, thyroid hormone concentrations did not appear to be associated with body weight changes reported for the juvenile males. 88   In conclusion, both the toxicokinetic and reproductive studies of BTBPE in the mink demonstrated that the flame retardant did not induce overt toxic effects at environmentally relevant concentrations. Because environmental concentrations of BTBPE are expected to increase in time, it is advised that higher dietary concentrations be evaluated in mink. 89   REFERENCES 90   REFERENCES Alaee, M., Sergeant, D.B., Muir, D.C.G., Whittle, D.M., Solomon, K., and Luross, J. (1999) Distribution of polybrominated diphenyl ethers in the Canadian environment. Organohalogen Compounds 40:347-350. Alaee, M., and Wenning, R.J. (2002) The significance of brominated flame retardants in the environment: current understanding, issues and challenges. Chemosphere 46:579-582. Alaee, M., Arias, P., Sjödin, A., and Bergman, A., (2003) An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environment International 29:683-689. Allchin, C. R., Morris, S., and Law, R. J. (1999) Polybrominated diphenylethers (PBDEs) in sediments and biota downstream of potential sources in the UK. Environmental Pollution 105: 197-207. Anderson, H. A., Wolff, M. S., Fischbein, A., and Selikoff, I. J. (1978) Investigation of the health status of Michigan Chemical Corporation employees. Environmental Health Perspectives 23:187-191. Arnold, T.W., and Fritzell, E.K. (1990) Habitat use by male mink in relation to wetland characteristics and avian prey abundances. Canadian Journal of Zoology 68:2205-2208. Aulerich, R. J., Ringer, R. K., Seagram, H. L., and Youatt, W. G. (1971) Effect of feeding coho salmon and other Great Lakes fish on mink reproduction. Canadian Journal of Zoology 49:611-616. Aulerich, R.J., and Bursian, S.J. (1996) Toxicology in mink. In: D.B. Hunter and N. Lemieux (Ed.), Mink: Biology, Health, and Disease. University of Guelph, Guelph, Ontario, Canada. Chap. 18, p. 1-34. Basu, N., Scheuhammer, A.M., Bursian, S.J., Elliott, J., Rouvinen-Watt, K., and Chan, H.M. (2007) Mink as a sentinel species in environmental health. Environmental Research 103:130-144. Birnbaum, L.S., and Staskal, D.F. (2004) Brominated flame retardants: cause for concern? Environmental Health Perspectives 112:9-17. Blum, A., and Ames, B.N. (1977) Flame retardant additives as possible cancer hazards. Science 195:17-23. Blum, A., Ames, B.N., Gold, M.D., Jones, F.R., Hett, E.A., Dougherty, R.C., Horning, E.C., Dzidic, I., Carroll, D.I., Stillwell, R.N., and Thenot, J.P. (1978) Children absorb tris-BP flame retardant from sleepwear: urine contains the mutagenic metabolite, 2,3dibromopropanol. Science 201:1020-1023. 91   Bull, K., Basu, N., Zhang, S., Martin, J. W., Bursian, S., Martin, P., and Chan, L. H. (2007) Dietary and in utero exposure to a pentabrominated diphenyl ether mixture did not affect cholinergic parameters in the cerebral cortex of ranch mink (Mustela vison). Toxicological Sciences 96:115-122. Burreau, S., Zebuhr, Y., Broman, D., and Ishaq, R. (2004) Biomagnification of polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs) studied in pike (Esox lucius), perch (Perca fluviatilis), and roach (Rutilus rutilus) from the Baltic Sea. Chemosphere 55:1043–1052. Burreau, S., Zebühr,Y., Broman, D., and Ishaq, R. (2006) Biomagnification of PBDEs and PCBs in food webs from the Baltic Sea and the northern Atlantic Ocean. Science of the Total Environment 366:659-672. Calabrese, E.J., Aulerich, R.J., and Padgett, G.A. (1992) Mink as a predictive model in toxicology. Drug Metabolism Reviews 24:559-578. Chan, H.M., Scheuhammer, A.M., Ferran, A., Loupelle, C., Holloway, J., and Weech S. (2003) Impacts of mercury on freshwater fish-eating wildlife and humans. Human and Ecological Risk Assessment 9:867-883. Chernyak, S.M., Rice, C.P., Quintal, R.T., Begnoche, L.J, Hickey, J.P., and Vinyard, B.T. (2005) Time trends (1983–1999) for organochlorines and polybrominated diphenyl ethers in rainbow smelt (Osmerus mordax) from Lakes Michigan, Huron, and Superior, USA. Environmental Toxicology and Chemistry 24:1632-1641. Costa, L.G., and Giordano, G. (2007) Developmental neurotoxicity of polybrominated diphenyl ether (PBDE) flame retardants. Neurotoxicology 28:1047-1067. Covaci, A., Harrad, S., Abdallah, M.A.E., Ali, N., Law, R.J., Herzke, D., and de Wit, C.A. (2011) Novel brominated flame retardants: A review of their analysis environmental fate and behavior. Environment International 37:532–556. Cumbie, P.M. (1975) Mercury levels in Georgia otter mink and freshwater fish. Bulletin of Environmental Contamination and Toxicology 14:193-196. Darnerud, P.O. (2003) Toxic effects of brominated flame retardants in man and in wildlife. Environmental International 29:841-853. Darnerud, P.O., Eriksen, G.S., Jóhannesson, T., Larsen, P. B., and Viluksela, M. (2001) Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environmental Health Perspective 109:49-68. de Boer, J., de Boer, K., and Boon, J.P. (2000) Polybrominated biphenyls and diphenylethers. In: Paasivirta J. (Ed.), The Handbook of Environmental Chemistry, vol. 3, part K, pp. 61–95. de Boer, J., Wester, P.G., Klamer, H.J.C., Lewis, W.E., and Boon, J.P. (1998) Do flame retardants threaten ocean life? Nature 394: 28–29. 92   de Wit, C.A. (2002) An overview of brominated flame retardants in the environment. Chemosphere 46:583-624. de Wit, C.A., Alaee, M., and Muir, D.C.G. (2006) Levels and trends of brominated flame retardants in the Arctic. Chemosphere 64:209-233. de Wit, C.A., Herzke, D., and Vorkamp, K. (2010) Brominated flame retardants in the Arctic environment - trends and new candidates. Science of the Total Environment 408, 28852918. Di Carlo, F.J.D., Seifter, J., and De Carlo, V.J. (1978) Assessment of the hazards of polybrominated biphenyls. Environmental Health Perspectives 23:351-365. Dybing, E., Soderlund, E.J., and Nelson, S.D. (1980) Irreversible macromolecular binding of the flame retardant tris-(2, 3-dibromopropyl) phosphate in vitro and in vivo. Developments in Toxicology and Environmental Science 8:265-268. Eriksson, P., Jakobsson, E., and Fredriksson, A. (1998) Developmental neurotoxicity of brominated flame retardants, polybrominated diphenyl ethers, and tetrabromo-bis-phenol A. Organohalogen Compounds 35:375-377. Fowles, J.R., Fairbrother, A., Baecher-Steppan, L., and Kerkvliet, N.I (1994) Immunologic and endocrine effects of the flame-retardant pentabromodiphenyl ether (DE-71) in C57BL/6J mice. Toxicology 86:49-61 Foley, R.E., Jackling, S.J., Sloan, R.J., and Brown, M.K. (1988) Organochlorine and mercury residues in wild mink and otter: comparison with fish. Environmental Toxicology and Chemistry 7:363-374. Fox, W.M., Connor, L., Copplestone, D., Johnson, M.S., Leah, R.T. (2001) The contamination history with organochlorines of the Mersey estuary UK, revealed by analysis of sediment cores from salt marshes. Marine Environmental Research 51:213-227. Fries, G.F. (1985) The PBB episode in Michigan - An overall appraisal. Critical Reviews in Toxicology 16:105-156. Gauthier, L.T., Hebert, C.E., Weseloh, D.V.C., and Letcher, R.J. (2007) Current-use flame retardants in the eggs of herring gulls (Larus argentatus) from the Laurentian Great Lakes. Environmental Science and Technology 41:4561-4567. Gauthier, L.T., Hebert, C.E., Weseloh, D.V.C., and Letcher, R.J. (2008) Dramatic changes in the temporal trends of polybrominated diphenyl ethers (PBDEs) in herring gull eggs from the Laurentian Great Lakes: 1982–2006. Environmental Science and Technology 42:15241530. Gauthier, L.T, and Letcher, R.J. (2009) Isomers of Dechlorane Plus flame retardant in the eggs of herring gulls (Larus argentatus) from the Laurentian Great Lakes of North America: temporal changes and spatial distribution. Chemosphere 75:115-120. 93   Gerecke, A.C., Giger, W., Hartmann, P.C., Heeb, N.V., Kohler, H.P.E., Schmid, P., Zennegg, M., Kohler, M. (2006) Anaerobic degradation of brominated flame retardants in sewage sludge. Chemosphere 64:311-317. Gill, U., Chu, I., Ryan, J.J., and Feeley, M. (2004) Polybrominated diphenyl ethers: human tissue levels and toxicology. Reviews of Environmental Contamination and Toxicology 183:5597. Graham, L.H., Schwarzenberger, F., Mostl, E., Galama, W., Savage, A. (2001) A versatile enzyme immunoassay for the determination of progestogens in feces and serum. Zoo Biology 20:227-236 Hakk, H., Larsen, G., and Bowers, J. (2004) Metabolism, tissue disposition, and excretion of 1,2bis (2,4,6-tribromophenoxy)ethane (BTBPE) in male Sprague-Dawley rats. Chemosphere 54:1367-1374. Hale, R.C., La Guardia, M.J., Harvey, E.P, Matteson-Mainor, T., Duff, W.H., and Gaylor, M.O. (2001) Polybrominated diphenyl ether flame retardants in Virginia freshwater fishes. Environmental Science and Technology 35:4585-4591. Hale, R.C., La Guardia, M.J., Harvey, E., and Mainor, T.M. (2002) Potential role of fire retardant-treated polyurethane foam as a source of brominated diphenyl ethers to the US environment. Chemosphere 46:729-735. Hale, R.C., La Guardia, M.J., Harvey, E., Gaylor, M.O., Matteson-Mainor, T. (2006) Brominated flame retardant concentrations and trends in abiotic media. Chemosphere 64:181-186. Hites, R.A. (2006) Brominated flame retardants in the Great Lakes. Handbook of Environmental Chemistry 5:355-390. Hoh, E., Zhu, L., and Hites, R.A (2005a) Novel flame retardants, 1, 2-bis (2, 4, 6tribromophenoxy) ethane and 2, 3, 4, 5, 6-pentabromoethylbenzene, in United States' environmental samples. Environmental Science and Technology 39:2472-2477. Hoh, E., Zhu, L., and Hites, R.A. (2005b) Brominated flame retardants in the atmosphere of the East-Central United States. Environmental Science and Technology 39:7794-7802. Ikonomou M.G., Rayne, S., and Addison R.F. (2002) Exponential increases of the brominated flame retardants, polybrominated diphenyl ethers, in the Canadian Arctic from 1981 to 2000. Environmental Science and Technology 36: 1886–1892. Johnson, A., and Olson, N. (2001) Analysis and occurrence of polybrominated diphenyl ethers in Washington State freshwater fish. Archives Environmental Contamination and Toxicology 41:339-344. Karlsson, M., Ericson, I., Bavel, B.V., Jensen, J.K., and Dam, M. (2006) Levels of brominated flame retardants in Northern Fulmar (Fulmarus glacialis) eggs from the Faroe Islands. Science of the Total Environment 367:840-846. 94   Karlsson, M., Julander, A., Van Bavel, B., and Hardell, L. (2007) Levels of brominated flame retardants in blood in relation to levels in household air and dust. Environment International 33:62-69. Kierkegaard, A., Balk, L., Tjärnlund, U., de Wit, C.A., and Janssonet, B. (1999) Dietary uptake and biological effects on decabromodiphenyl ether in rainbow trout (Oncorhynchus mykiss). Environmental Science and Technology 33:1612-1617. Kierkegaard, A., Sellström, U., McLachlan, M., (2009) Environmental analysis of higher brominated diphenyl ethers and decabromodiphenyl ethane. Journal of Chromatography A 1216:364-375. Kucera, E. (1983) Mink and otter as indicators of mercury in Manitoba waters. Canadian Journal of Zoology 61:2250-2256. Law, R.J., Allchin, C.R., DeBoer, J., Covaci, A., Herzke, D., Lepom, P., Morris, S., Tronczynski, J., and de Wit, C.A. (2006) Levels and trends of brominated flame retardants in the European environment. Chemosphere 64:187-208. Lyman, W.J., Reehi, W.F. and Rosenblatt, D.H. (Editors) (1990) Handbook of Chemical Property Estimation Methods: Environmental Behavior of Organic Compounds. American Chemical Society, Washington, D.C. Martin, P. A., Mayne, G. J., Bursian, S.J., Tomy, G., Palace, V., Pekarik, C., and Smits, J. (2007) Immunotoxicity of the commercial polybrominateddiphenylether mixture DE-71 in ranched mink (Mustela vison). Environmental Toxicology and Chemistry 26:988-997. Miceli, J.N., Nolan, D.C., Marks, B., and Hariharan, M. (1985) Persistence of polybrominated biphenyls (PBB) in human post-mortem tissue. Environmental Health Perspectives 60:399-403. Morck, A., Hakk, H., Orn, U., and Wehler, E.K. (2003) Decabromodiphenyl ether in the rat: absorption, distribution, metabolism, and excretion. Drug Metabolism and Disposition 31:900-907. Nomeir, A.A., Markham, P.M., Ghanayem, B.I., Chadwick, M. (1993) Disposition of the flame retardant 1,2-bis (2,4,6-tribromophenoxy)ethane in rats following administration in the diet. Drug Metabolism and Disposition 21:209-214. Norris, J. M., Kociba, R. J., Schwetz, B. A., Rose, J. Q., Humiston, C. G., Jewett, G. L., Gehring, P. J., and Mailhes, J. B. (1975) Toxicology of octabromobiphenyl and decabromobiphenyloxide. Environmental Health Perspectives 11:153-161. Opperhuizen, A., Velde, E.W., Gobas, FAPC., Liem, D.A.K., and Steen, J.M.D. (1985) Relationship between bioconcentration in fish and steric factors of hydrophobic chemicals. Chemosphere 14:1871-1896. 95   Pettersson-Julander, A., van Bavel, B., Engwall, M., and Westberg, H. (2004) Personal air sampling and analysis of polybrominated diphenyl ethers and other bromine containing compounds at an electronic recycling facility in Sweden. Journal of Environmental Monitoring 6:874-880. Pijnenburg, A.M., Everts, J.W., de Boer, J., and Boon, J.P. (1995) Polybrominated biphenyl and diphenylether flame retardants: analysis, toxicity, and environmental occurrence. Reviews of Environmental Contamination and Toxicology 141:1-26. Prival, M., McCoy, E., Gutter, B, and Rosendranz, H. (1977) Tris(2,3-dibromopropyl) phosphate: mutagenicity of a widely used flame retardant. Science 195:76-78. Qiu X., Marvin, C.H., and Hites, R.A. (2007) Dechlorane plus and other flame retardants in a sediment core from Lake Ontario. Environmental Science and Technology 41:6014-6019. Rayner, N.A., Parker, D.E., Horton, E.B., Folland, C.K., Alexander, L.V., Rowell, D.P., Kent, E.C., and Kalan, A. (2003) Globally complete analyses of sea surface temperature, sea ice and night marine air temperature in the late nineteenth century. Journal of Geophysical Research 108:4407-4436. Renner, R. (2000) Increasing levels of flame retardants found in North American environment. Environmental Science and Technology 34:452-453. Renner, R. (2004) In US, flame retardants will be voluntarily phased out. Environmental Science and Technology 38:14A. Saegusa, Y., Fujimoto, H., Woo, G.H., Inoue, K., Takahashi, M., Mitsumori, K., Hirose, M., Nishikawa A, and Shibutani M. (2009) Developmental toxicity of brominated flame retardants, tetrabromobisphenol A and 1,2,5,6,9,10-hexabromocyclododecane, in rat offspring after maternal exposure from mid-gestation through lactation. Reproductive Toxicology 28:456-467. Segev, O., Kushmaro, A., and Brenner, A. (2009) Environmental impact of flame retardants (persistence and biodegradability). International Journal of Environmental Research and Public Health 6:478-491. Sellström, U., Kierkegaard, A., de Wit, C., and Jansson, B. (1998) Polybrominated diphenyl ethers and hexabromocycododecane in sediment and fish from a Swedish river. Environmental Toxicology and Chemistry 17:1065-1072. Shi, T., Chen, S., Luo, X., Zhang, X., Tang, C., Luo, Y., Ma, Y., Wu. J.P., Peng, X.Z., and Mai, B.X. (2009) Occurrence of brominated flame retardants other than polybrominated diphenyl ethers in environmental and biota samples from southern China. Chemosphere 74:910-916. Silva, K.D., Fernandes, A., and Rose, M. (2004) Brominated organic micropollutants--igniting the flame retardant issue. Critical Reviews in Environmental Science and Technology 34:141-207. 96   Sjödin, H., Carlsson, K., Thuresson, S., Sjölin, A., Bergman, and Ostman, C. (2001) Flame retardants in indoor air at an electronics recycling plant and at other work environments. Environmental Science and Technology 35:448-454. Söderlund, E., Dybing, E., and Nelson, S.D. (1980) Nephrotoxicity and hepatotoxicity of tris(2,3-dibromopropyl) phosphate in the rat. Toxicology and Applied Pharmacology 56:171-181. Strandberg, B., Dodder, N.G., Basu, I., and Hites, R.A. (2001) Concentrations and spatial variations of polybrominated diphenyl ethers and other organohalogen compounds in Great Lakes air. Environmental Science and Technology 35:1078-1083. Szymanska, J.A., Sapota, A., and Frydrych, B. (2001) The disposition and metabolism of tetrabromobisphenol-A after a single i.p. dose in the rat. Chemosphere 45:693-700. Tomy, G.T., Palace, V.P., Pleskach, K., Ismail, N., Oswald, T., Danell, R., Wautier, K., and Evans, B. (2007) Dietary exposure of juvenile rainbow trout (Oncorhynchus mykiss) to 1,2-bis(2,4,6-tribromo-phenoxy)ethane: bioaccumulation parameters, biochemical effects, and metabolism. Environmental Science and Technology 41:4913-4918. Van der Ven, L.T.M., Verhoef, A., van de Kuil, T., Slob, W., Leonards, P.E.G., Visser, T.J., Hamers, T., Herlin, M., Hakansson, H., Olausson, H., Piersma, A.H., and Vos, J.G. (2006) A 28-day oral dose toxicity study enhanced to detect endocrine effects of hexabromocyclododecane in Wistar rats. Toxicological Sciences 94:281-292. Van der Ven, L.T., Van de Kuil, T., Verhoef, A., Verwer, C.M., Lilienthal, H., Leonards, P.E.G, Schauer, U.M.D., Canton, R.F., Litens, S., De Jong, F.H., Visser, T.J., DeKant, W., Stern, N., Hakansson, H., Slob, W., Van den Berg, M., Vos, J.G., and Piersma, A.H. (2008) Endocrine effects of tetrabromobisphenol-A (TBBPA) in Wistar rats as tested in a one-generation reproduction study and a subacute toxicity study. Toxicology 245:76-89. Veith, G.D., and Defoe, D.L. (1979) Measuring and estimating the bioconcentration factor of chemicals in fish. Journal of the Fisheries Research Board of Canada 36:1040-1048. Verreault, J., Gebbink, W.A., Gauthier, L.T., Gabrielsen, G.W., and Letcher, R.J. (2007) Brominated flame retardants in glaucous gulls from the Norwegian Arctic: more than just an issue of polybrominated diphenyl ethers. Environmental Science and Technology 41:4925-4931. Viberg, H., Fredriksson, A., and Eriksson, P. (2003a) Neonatal exposure to polybrominated diphenyl ether (PBDE 153) disrupts spontaneous behaviour, impairs learning and memory, and decreases hippocampal cholinergic receptors in adult mice. Toxicology and Applied Pharmacology 192:95-106. Viberg, H., Fredriksson, A., Jakobsson, E., Orn, U., and Eriksson, P. (2003b) Neurobehavioural derangements in adult mice receiving decabrominated diphenyl ether (PBDE 209) during a defined period of neonatal brain development. Toxicological Sciences 76:112-120. 97   Von Meyerinck, L., Hufnagel, B., Schmoldt, A., and Benthe, H.F. (1990) Induction of rat liver microsomal cytochrome P-450 by the pentabromodiphenyl ether Bromkal 70 and halflives of its components in the adipose tissue. Toxicology 61:259-274. Voorspoels, S., Covaci, A., Neels, H., and Schepens, P. (2007) Dietary PBDE intake: a market basket study in Belgium. Environment International 33:93-97. Ward, J., Mohapatra, S.P., and Mitchell, A. (2008) An overview of policies for managing polybrominated diphenyl ethers (PBDEs) in the Great Lakes basin. Environment International 34:1148-1156. Watanabe, I., Kashimoto, T., and Tatsukawa, R. (1983a) The flame retardant tetrabromobisphenol A and its metabolite found in river and marine sediments in Japan. Chemosphere 12:1533–1839. Watanabe, I., Kashimoto, T., and Tatsukawa, R. (1983b) Identification of the flame retardant tetrabromobisphenol A in the river sediment and the mussel collected in Osaka. Bulletin of Environmental Contamination and Toxicology 31:48–52. Watanabe, I., and Sakaik, S. (2003) Environmental release and behavior of brominated flame retardants. Environment International 29:665–682. Wiener, J.G., Krabbenhoft, D.P., Heinz, G.H., and Scheuhammer, A.M. (2003) Ecotoxicology of mercury. In: Hoffman, D.J., Rattner, B.A., Burton, G.A., and Cairns, J. (Eds.), Handbook of Ecotoxicology, CRC Press, Boca Raton, FL. p. 409-463. Wise, M.H., Linn, I.J, and Kennedy, C.R. (1981) A comparison of the feeding biology of mink (Mustela vison) and otter (Lutra lutra). Journal of Zoology (London) 195:181-213. Xinghua, Q., Marvin, C.H., and Hites, R.A. (2007) Dechlorane Plus and other flame retardants in a sediment core from Lake Ontario. Environmental Science and Technology 41:60146019. Zhang, S., Bursian, S., Martin, P. A., Chan, H. M., and Martin, J. W. (2008) Dietary accumulation, disposition, and metabolism of technical pentabrominated diphenyl ether (DE-71) in pregnant mink (Mustela vison) and their offspring. Environmental Toxicology and Chemistry 27:1184–1193. Zhang, S., Bursian, S.J., Martin, P.A., Chan, H.M., Tomy, G., Palace, V.P., Mayne, G.J., Martin, J.W. (2009) Reproductive and developmental toxicity of a pentabrominated diphenyl ether mixture, DE-71, to ranch mink (Mustela vison) and hazard assessment for wild mink in the Great Lakes region. Toxicological Sciences 110:107-116. Zhou, T., Ross, D.G., DeVito, M.J., and Crofton, K.M. (2001) Effects of short-term exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats. Toxicological Sciences 61:76-78. 98   Zhou, Z.M (2006) Implement of administrative measure on the control of pollution caused by electronic information products and the exemption of deca-BDE mixture. Flame Retardant Meter Technology 4:15-16. 99