AMMONIUM PERCHLORATE-INDUCED LESIONS

IN ZEBRAFISH KIDNEYS

by

TIM CAPPS, B.S.

A THESIS

IN

ENVIRONMENTAL TOXICOLOGY

Submitted to the Graduate Faculty of Texas Tech University in Partial Fulfillment of the Requirements for the Degree of

MASTER OF SCIENCE

Approved

Chairperson of the Committee

Accepted

Dean of the Graduate School

December, 2003 ACKNOWLEDGMENTS

1 would like to especially thank Dr. Reynaldo Patino for being the chairperson of my committee who made this project possible, and for giving me a positive and enjoyable educational experience. 1 would also like to thank the other members of my committee:

Dr. Todd Anderson and Professor Victoria Sutton 1 would like to thank Dr. Vickie

Blazer for her histological expertise. I would like to thank The Institute of

Environmental and Human Health at Texas Tech for giving me monetary support, and being a good organization with great people. I would like to thank the U.S. Geological

Survey through the Texas Cooperative Fish and Wildlife Research Unit at Texas Tech, and all of the people there, for the use of facilities and for providing a great work environment. I would also like to thank the U.S. Department of Defense through the

Strategic Environmental Research and Development Program, under a cooperative agreement with the US. Air Force Institute for Environmental Safety and Occupational

Health, for providing further flinding for this project.

Finally, I would like to thank all of my friends, family, and coworkers for their support and understanding. Without all of you this would not have been possible. TABLE OF CONTENTS

ACKNOWLEDGMENTS ii

ABSTRACT v

LIST OF FIGURES vii

CHAPTER

I. BACKGROUND AND OBJECTIVES 1

Introduction 1

Ammonium Perchlorate 2

Biological Effects of Perchlorate 3

Thyroidal Effects of Ammonium Perchlorate 4

Dose-Response Studies of Thyroidal Effects 4

Developmental Effects of Perchlorate 7

Reproductive Effects of Perchlorate 8

Renal Effects of Perchlorate 9

Macrophage Aggregates 11

II. LEGAL ASPECTS OF AMMONIUM PERCHLORATE 13

III. AMMONIUM PERCHLORATE -INDUCED LESIONS

IN ZEBRAFISH KIDNEYS 18

Introduction 18

Materials and Methods 20

Results 22 Discussion 24

IV. CONCLUSIONS 35

LITERATURE CITED 37 ABSTRACT

Ammonium perchlorate (AP) is a highly reactive chemical that is primarily used as a rocket propellant for military and aerospace industries. Widespread AP contamination has caused increased concern over its effect on biological systems.

Perchlorate derived from AP, and other perchlorate species, inhibits iodide uptake by the thyroid follicles, thus impairing the production of thyroid hormones and disrupting the regulatory feedback mechanisms of the thyroid system. As a result, the secretion of thyroid-stimulating hormone (TSH) by the pituitary gland increases. Effects of AP on other physiological systems, such as the renal system, are not well understood. The mammalian and teleost kidney are both known to contain TSH and thyroid hormone receptors, as well as iodide transport mechanisms similar to those found in thyroid follicles. Therefore, the kidney is also a potential target of direct (association with iodide transporters) or indirect (decreased thyroid hormones and increased TSH) actions of perchlorate. This study examined the histopathological effects of AP on the zebrafish kidney. Adult zebrafish were exposed to water-borne perchlorate at concentrations of 18- ppm for eight weeks and 677-ppm for 4 weeks. At the end of the exposure period, the fish were processed for histological analyses of the trunk region of the kidney. The samples were then analyzed for increased presence of macrophage aggregates, focal granulomas, inflammation, and generating nephrons. Quantitative analysis was conducted by determining the relative average area covered by a lesion in randomly selected histological sections. The results indicated that exposure to AP at 18-ppm for eight weeks caused a significant increase in the incidence of renal macrophage aggregates when compared with the control fish. The other lesions showed no changes in incidence at any AP dose. The study also suggests a link between thyroid disruption and increases in renal macrophage aggregates.

VI LIST OF FIGURES

1. Hemotoxilyn-Eosin stained macrophage Aggregate 27

2. PAS stained macrophage aggregate 28

3. Iron-stained macrophage aggregate 29

4 Percent kidney area covered by lesions in zebrafish exposed to ammonium perchlorate 30

5. Mean area of kidney histological sections covered by lesions, and /^-values

of statistical comparison 31

6. Hemotoxylin-Eosin stained focal granuloma with basophilic encapsulation 32

7. Hemotoxylin-Eosin stained inflammation with eosinophilic cells and core 33

8. Hemotoxylin-Eosin stained generating tubule cluster 34

Ml CHAPTER I

BACKGROUND AND OBJECTIVES

Introduction

Ammonium perchlorate (AP) is a chemical primarily used as an ingredient in soUd fuel rocket engines. In the environment, the perchlorate anion derived from AP persists for a relatively long period of time, and occurs in drinking water supplies in some parts of the United States. Perchlorate inhibits thyroid hormone production by limiting the uptake of iodide, an essential component of thyroid hormones, into the thyroid follicles. Decreased thyroid hormone production may be associated with some physiological dysfijnctions.

The purpose of this study is to determine if AP exposure affects the histopathology of zebrafish kidneys. Effects of perchlorate on the thyroidal system of zebrafish, and other organisms, are well documented. Due to some similarities between the thyroid and kidney, namely the presence of iodide transporters, thyroid hormone receptors, and thyroid stimulating hormone receptors, there exists a potential for a perchlorate-induced effect on the kidney. Whether or not there is a direct or indirect eflfect of perchlorate on the kidney will also be analyzed. To my knowledge, this is the first laboratory-controlled study where the effects of AP at concentrations found in the environment (Smith et al., 2001) are determined for the kidney of any species. Ammonium Perchlorate

Ammonium perchlorate has a variety of uses in the public and private sectors of the United States. Since the 1940s, AP has mainly been used as an oxidizing agent in military and aerospace solid fuel rocket engines and munitions (Fisher et al., 2000). Non- military or aerospace uses of AP include airbag inflators, pyrotechnics, industrial explosives, and batteries. Further, perchlorate salts can be found in products as diverse as

fertilizers, leather tarming solutions, and electrical tubes (Sharp and Walker, 2001). In the past, perchlorate was used for the treatment of Grave's disease, or hyperthyroidism, in

himians (Fisher et al., 2000).

One major problem associated with AP is the environmental persistence of perchlorate. In water, the ammonium ion and perchlorate anion of AP dissociate, and although the perchlorate anion is a strong oxidant, the activation energy required for its

reduction is high (Fisher et al., 2000). As a result, perchlorate can continue in the

envirormient for many years (Urbansky, 2002). Perchlorate is susceptible to various

natural and artificial degradation processes, including induced ion exchange and natural

microbial degradation. However, remediation measures for perchlorate-contaminated

sites have resulted in varying outcomes (Logan, 2001).

A heightened public awareness of AP has stemmed fi-om the increased presence of perchlorate in ground and surface waters throughout the country. The use of AP in military munitions and aerospace operations accounts for the majority of environmental contamination attributed to AP. Nearly all surface and ground waters contaminated by perchlorate are near or downstream from facilities that manufacture or use AP (Sharp and

Walker, 2001). In the United States, facilities associated with AP occur in at least 39 states, 19 of which have confirmed perchlorate contamination in some ground or surface water sources (Sharp and Walker, 2001). Perchlorate contamination is a major public concern in California. As of 2001, perchlorate contamination in California occurred in

197 drinking water sources that serviced over seven million people. Some of the perchlorate concentrations found in California were above recommended safe levels for drinking water. Parts of the Colorado River, which flows through and supplies drinking and irrigation water to people in CaUfomia, Nevada, and Arizona contains levels of perchlorate as well (Sharp and Walker, 2001). Further, perchlorate contamination is not always limited to water. Recent studies of a military munitions plant in central Texas foxmd the presence of perchlorate in ground and surface water, soil, vegetation, aquatic insects, fish, and mammals (Smith et al, 2001).

Biological Effects of Perchlorate

In recent years, the heightened concern over possible public health effects of perchlorate contamination led to studies concerning the eflfects of perchlorate on biological activities such as thyroid fimction, development, and reproduction. However, studies concerning the effects of perchlorate on other biological flinctions are currently insufficient. Considering the increase in perchlorate contamination, and consequent availability to organisms, additional research of all systems conceivably affected by perchlorate seems justified. Thvroidal Effects of Ammonium Perchlorate

The longest known effect of perchlorate on biological systems is the alteration of thyroid gland fiinction (Sellivanova et al., 1968). The primary action of the perchlorate anion on thyroid follicles (the anatomical units of the thyroid gland that produce thyroid hormones) is the competitive inhibition of sodium-iodide symporters, which regulate the amount of iodide taken into the thyroid (Wolf, 1998). Since iodide is an essential component of thyroid hormones (T3 and T4), the inhibition of iodide uptake results in decreased thyroid hormone production. Decreased thyroid hormone blood levels disrupt homeostasis and triggers an increased production and release of thyroid-stimulating hormones (TSH) by the pituitary gland, which in turn causes the thyroid to produce more thyroid hormones. Because thyroid follicles are unable to increase thyroid hormone production, pituitary TSH production continues at relatively high levels eventually resulting in the hyperactive stimulation of thyroid follicles and excessive circulating TSH levels (Soldinetal, 2001).

Dose-Response Studies of Thyroidal Effects

Considerable amounts of research have focused on determining the levels of perchlorate exposure that induce thyroidal effects. An early study of AP toxicity determined the effects on dogs of AP repeatedly introduced into their stomachs in a summary dose of 4500 mg/kg over a three-week period (SeUvanova et al, 1969). This treatment caused an acute and prolonged inhibition of thyroid hormone production as well as several other severe physiological malfiinctions (Selivanova et al., 1969).

However, more recent studies focus on smaller AP concentrations and their effect on thyroid health.

Changes in thyroid hormone and TSH levels is one effect of AP. Studies conducted on rats showed that AP exposure levels in the 3-30 mg/kg/day range (given for

30 days) resuhed in increased TSH levels, and decreased T4 and T3 levels (York et al.,

2001a). A similar thyroid hormone effect occurred in pregnant rabbits dosed with AP concentrations greater than 30 mg/kg/day (administered on gestation days 6 through 28)

(York et al., 2001b). In frogs, exposures of 14-140 ppb AP result in alterations of thyroid hormone levels. However, removal from AP reversed the effects on thyroid hormone levels after 28 days (Goleman et al., 2002a).

In addition to alterations in thyroid hormone and TSH levels, studies reveal AP- induced changes in thyroid follicle histopathology. Most thyroid follicular changes manifest themselves in the occurrence of thyroid follicle hypertrophy, hyperplasia, colloid depletion, and angiogenesis. In pregnant rabbits, AP caused increased thyroid follicular hypertrophy at dose levels greater than 10 mg/kg/day (administered on gestation days 6 through 28) (York et al, 2001b). Doses of 3-30 mg/kg/day for 14 or 90 days induced increased thyroid weight, follicular cell height (hypertrophy), hyperplasia, and colloid depletion in rats (Siglin et al, 2000). Further, thyroid adenomas occurred in rats dosed with 30 mg/kg/day AP for 90 days (Wolf, 2000). In frogs dosed with 14-140 ppb AP, frogs exposed to the higher concentrations of AP exhibited thyroid follicular hypertrophy (Goleman et al, 2002a). Several studies suggest a time relevant effect of AP on thyroid histopathology.

One histological study suggests a time-dependent effect of perchlorate on the thyroid

structure of rats. Rats continuously dosed with perchlorate for 1 to 12 months showed a continuous increase in thyroid weight and increased thyroid hypertrophy, hyperplasia, colloid depletion, and increased vascularization starting after two months of treatment

(Fernandez Rodriguez, 1991). Further, in a study conducted with zebrafish, envirormientally relevant concentrations of AP (18 ppm) (Smith et al, 2001), administered for eight weeks significantly increased thyroid follicular cell hypertrophy

and hyperplasia, angiogenesis, and colloid depletion. However, larger doses (677 ppm) administered for four weeks resulted in no significant increase in thyroid folUcular hystopathological parameters (Patiiio et al, 2003).

Human AP toxicity data is less extensive than animal toxicity data. The effects of perchlorate on human thyroid function have been established through perchlorate's past use as a treatment for hyperthyroidism (Wolf, 1998). However, information on thyroidal effects of perchlorate at smaller doses possibly found in the environment is limited to epidemiological studies, and a small number of laboratory studies.

Studies of thyroidal effects of envirormiental exposures to AP reveal conflicting effects of perchlorate. An epidemiological study of newborns in Arizona concluded that those residing in areas with contaminated drinking water from the Colorado River had higher TSH levels than newborns from areas without perchlorate-contaminated water

(Brechner et al, 2000). Other epidemiological studies yielded contradictory or inconsistent information. For example, the thyroid function of people from Las Vegas, Nevada exposed to 4 to 24 |ig/L of perchlorate in their drinking water was not significantly different from the thyroid function of unexposed people (Li Zet al, 2000; Li

F.X. et al, 2000; Li et al, 2001). Further, children and newborns from a Chilean city with perchlorate-contamination, at levels up to 120 ng/L, had lower TSH levels than children and newborns from some uncontaminated areas (Crump et al, 2000).

Human laboratory studies give further data on the thyroid toxicity of AP. A study with humans dosed wdth AP at 10 mg/day for 14 days found a decrease in radioactive iodide uptake levels in thyroid follicles without a change in circulating thyroid hormone or TSH levels (Lawrence et al, 2000). Further, concentrations of perchlorate in the range of 0.02 to 0.5 mg/kg/day administered in drinking water for 14 days did not affect iodide uptake in humans, thereby indicating a no-observable-effect level of perchlorate in drinking water of 250 |xg/L (ppb) (Greer et al, 2000). While these studies administered low, environmentally relevant doses, the length of exposure is arguably not indicative of envirormiental exposure through drinking water.

Developmental Effects of Perchlorate

The effects of AP on the thyroid have led to concerns over possible developmental effects. Thyroid hormones are vital for the proper development of physiological fiinctions such as neurological and bone development. Thyroid hormone irregularities during development are a major cause of neurological impairment in humans (Sharp & Walker, 2001). Studies reveal varying AP effects on development. In rabbits, the fetuses of pregnant females continuously dosed via drinking water with 0-102.3 mg/kg/day during gestation days 6 through 28 showed no increased incidence of skeletal alterations, neurological abnormaUties, reduced body weight, or increased mortality when compared with historical incidences of the abnormalities in the test species. These observations led to a finding of a no-observable-adverse developmental effect level of 100.0 mg/kg/day

(York et al., 2001b). However, in a study with zebrafish exposed to goitrogens

(including sodivmi perchlorate). the inhibition of thyroid hormone production was found to have substantial developmental effects. The inhibition of transition from larva to juvenile fish, inhibition of the formation of scales, and stimted fin grow1:h were all found to occur in the perchlorate treated groups (Brown, 1997). In a study of field mice dosed with 0, 1 nM, 1 |J,M, or I mM of AP from the time of breeding pair cohabitation to the time of pup euthanasia, a period of 20 to 70 days, pups from the dosed dams were found to have lower body and organ weights (Thuett et al, 2002). Further, tadpoles exposed to

AP at concentrations of 0-140 ppb, showed inhibited metamorphosis, and a lower male sex ratio in groups treated with higher concentrations. However, there were no incidences of structural abnormalities, and the inhibition of metamorphosis was corrected after removal of AP (Goleman et al, 2002a).

Reproductive Effects of Perchlorate

The potential of AP to interfere with reproductive fiinctions, through alterations of thyroid activity, led to experimental determinations of AP's effect on reproductive performance. In a two-generation study, pregnant rats were exposed to AP at 0-30.0 mg/kg/day for 30 days in drinking water (York et al, 2001a). Reproductive, developmental and physiological parameters such as body weights, thyroid and major organ weights, fertility, sperm counts, estrogen levels, litter sizes, and fetus and pup mortality were recorded. The only adverse effects observed were on thyroid hormone levels and thyroid weights of both parents and offspring at the higher doses, but there were no observable reproductive effects on either generation of rats at any of the doses tested (York et al., 2001a). In a study with zebrafish, groups of eight AP-exposed females were each paired with four AP-exposed males to determine the effect of AP on reproductive performance at exposures of 0-ppm, 18- ppm, and 677-ppm for eight weeks

(Patiiio et al, 2003). At 677-ppm, spawned egg volume was decreased immediately, but the effect was likely due to extrathyroidal effects of AP (either ammonium or perchlorate). Although the group dosed withl 8-ppm AP for eight weeks exhibited drastic thyroid follicle histopathological effects, no effect on spawned egg volume or egg fertilization rate occurred, thereby showing a no observable reproductive performance effect in zebrafish at environmentally recorded AP concentrations (Smith et al, 2001;

Patiiio et al, 2003).

Renal Effects of Perchlorate

Effects of AP on the renal system are relatively unknown. Early studies found that

AP is toxic to the kidney, but these studies often used extremely high doses of AP that usually ended with injury to several other organs (SeUivanova et al, 1968). Some epidemiological studies used kidney function as one of the parameters to determine the risks of AP exposure in the workplace. One such study found that employees exposed to up to 38 mg/kg/day of AP through inhalation and dermal absorption showed no signs of kidney injury (Gibbs et al, 1998; Lamm et al, 1999).

Perchlorate has the potential to affect the renal system in several ways. Thyroid hormones affect and regulate several important renal functions such as kidney growth and development, plasma flow, filtration rates, and electrolyte metabohsm (Katz et al, 1975;

Capasso et al, 1999; Prasad et al, 1999;). Further, TSH receptors as well as thyroid hormone receptors occur in mammaUan and teleost kidneys (MacLatchy and Eales, 1992;

Dutton et al, 1997; Sellitti et al, 2000; Vischer and Bogerd, 2003). Although the role of

TSH receptors in kidneys is unknown, it is conceivable that excess amounts of TSH and decreased levels of thyroid hormones, as during exposure to AP, could affect TSH and thyroid hormone sensitive renal fiinctions. Further, sodium-iodide symporters occur in mammalian kidneys (Spitzweg et al, 2001). Human renal iodide symporters are sensitive to perchlorate, thereby suggesting a similar iodide transport mechanism in the kidney and thyroid, and similar susceptibility to perchlorate-induced inhibition (Spitzweg et al,

2001). Overall, these fmdings indicate that perchlorate may affect kidney fimction directly, through interactions with sodium-iodide symporters or indirectly, through alterations in TSH and thyroid hormone levels.

Considering the current attention on the biological effects of AP and the existence of similarities between the thyroid and kidney, such as iodide transport, thyroid hormone receptors, and TSH receptors (Spitzweg et al, 2001: Dutton et al, 1997; Sellitti et al.

10 2000), there is much that remains unknown about the affects of thyroid toxicants, like

AP, on renal function. The possibility that AP causes kidney dysfunction warrants

fiirther research on this topic.

Macrophage Aggregates

A commonly used histopathological indicator of toxic exposure in fish is the

increased number and size of macrophage aggregates (MAs) (Wo Ike, 1992).

Macrophage aggregates most often occur in spleen, kidney, and liver tissues of both

healthy and diseased or distressed fish (Jolly, 1923; Yoffey, 1929). Factors such as fish

age, size, diet, heaUh, and exposure to environmental contaminants can affect the number

and size of MAs (Brown and George, 1985; Blazer et al, 1987). Macrophage aggregates

play key roles in the capture of cell debris and foreign materials for reuse, destruction, or

detoxification (Ellis et al, 1976). Further, macrophage aggregates play important

immunological roles by serving as a primitive form of germinal centers, found in

mammals, which introduce antigens to lymphocytes which produce antibodies to fight

infection (Wolke, 1992). Macrophage aggregates also serve in cellular and tissue responses to injury. Inflammation is a common response to tissue injury, which releases

chemical signals summoning defensive measures within an organism. Macrophages are one of the responding defensive agents that act to prevent fiirther damage to tissues by destroying invading organisms, isolating foreign materials, or removing cellular debris

(Wolke, 1992).

Often, the constituents of MAs can be determined by the pigments they contain.

Melanin containing MAs are common in many fish tissues, including the kidney, and

11 may aid in the neutralization of harmfiil oxidative species, the recycling and removal of cellular debris, and the sequestering, destruction, and removal of contaminants (Wolke,

1992). The presence of hemosiderin, a protein-bound iron pigment, indicates the breakdown and storage of iron-containing cellular products (Weeks et al, 1992; Blazer et al, 1997). PAS-positive MAs indicate the presence of polysaccharide deposits within the

MA.

12 CHAPTER 11

LEGAL ASPECTS OF AMMONIUM PERCHLORATE

Due to the increased environmental occurrence of perchlorate, steps have been

initiated to regulate the amount of perchlorate in water systems. While perchlorate is

currently listed under the Drinking Water Contaminant Candidate List (CCL) (Motzer,

2000), the United States Environmental Protection Agency (USEPA) is compiling data to

determine the threat that perchlorate poses to the public by compiUng extensive

toxicological data as well as data pertaining to perchlorate's distribution and availability

(USEPA, 2003). In 2006, the USEPA will decide the threat perchlorate poses to the

pubUc and if it should be listed under the National Primary Drinking Water Regulation

(NPDWR). If perchlorate becomes listed imder the NPDWR, a rrmximum contaminant

level (MCL) for perchlorate will be determined and used as a standard for regulating

perchlorate contamination in drinking water supplies (Motzer, 2000).

Part of the data dealing with the persistence and availability of perchlorate the

USEPA is compiling comes from the listing of perchlorate as a List One unregulated

chemical under the Unregulated Contaminant Monitoring Regulation (UCMR). The

UCMR requires the monitoring of chemicals not listed under the NPDWR, but which may pose a threat to public health, in public water supplies using set analytical methods.

As a List One chemical, the presence of perchlorate is monitored in pubUc drinking water sources that service over 10, 000 people, as well as in smaller drinking water sources

13 selected to obtain a representative sample of small water systems within a state (Federal

Register, 2000).

In 1999, the USEPA established a reference dose (RfD) for perchlorate based on doses of perchlorate found to cause thyroidal effects and disruption of neurological development. The RfD for perchlorate is between .0001 and .0005 mg/kg/day, and a hypothetical conversion of 4-18 ppb, which assumes 70 kg body weight and 2 liters of water consumption per day. This RfD is used as an advisory level for perchlorate levels in water, and as a guideline for RCRA and CERCLA remediation measures (USEPA,

2003).

The pending federal regulation of perchlorate caused several states to take steps to control perchlorate contamination. California enacted an interim health standard of 18 ppb for perchlorate, which requires the removal of a water source from the public or public notification of perchlorate levels found above the interim health standard (CDHS,

2003). Further, California enacted legislation calling for a risk assessment of perchlorate in order to establish a primary drinking water standard that would take effect on January

I, 2004. California's primary drinking water standard will have the same effect of a national drinking water standard in regards to regulating the cleanup and discharge of perchlorate, and the removal of contaminated water sources from pubUc use (California

Health & Safety Code, 2003). Other southwestern states have estabUshed interim action levels based on the USEPA's RfD that require remediation action in areas with perchlorate contamination exceeding these levels (Logan, 2001). Arizona implemented a

14 health based guidance level of 14-ppb. Nevada implemented an action level at 18-ppb, while the Texas action level is the highest at 22-ppb (Motzer, 2001).

Though federal drinking water regulation is awaiting review of data on the threat perchlorate poses, several bills have been introduced in the United States Congress which would accelerate regulation. The Preventing Perchlorate Pollution Act (The Act),

introduced to the United States House of Representatives on May 15, 2003, would

require the USEPA to list perchlorate as a contaminant under the NPDWR, and establish an MCL for perchlorate by July 1, 2004. The Act also seeks to increase community

information on perchlorate contamination by requiring parties involved with a perchlorate

discharge into any water source to notify the appropriate state authorities of the

discharge. Further, annual pubUcation in the Federal Register of discharge information

and locations where perchlorate is detected in the groundwater would advance public

awareness of perchlorate contamination. The Act also requires owners or operators of perchlorate storage facilities to report the storage conditions and amounts of perchlorate

stored at their facihty since 1950. To promote perchlorate remediation. The Act establishes a fund, augmented by fines resulting from violations of The Act, which used for perchlorate remediation efforts (H.R.2123, 2003).

Although perchlorate is not a chemical currently regulated by the NPDWR, perchlorate falls under several other regulations. Among these are the Comprehensive

Environmental Response, Compensation and Liability Act (CERCLA), and the Resource

Conservation and Recovery Act (RCRA). These two acts serve to control the fate of

15 hazardous wastes and chemicals, and makes parties responsible for contamination liable for remediation measures.

For a chemical to be regulated by RCRA, the chemical must either be listed as a hazardous waste or have the characteristics of a hazardous waste. Though perchlorate is not Usted as a hazardous waste, perchlorate has the requisite characteristics, namely its reactive and explosive nature, to make it a hazardous waste under RCRA (Castaic Lake

Water Agency v. Whittaker Corp., 2003; 40 C.F.R. §§ 261.21-261). By coming under

RCRA regulation, the manufacture, transport, use, and disposal of perchlorate is closely monitored and regulated (40 C.F.R. §§ 261.21-261).

CERCLA imposes liability on past and present owners of facilities involved in releases of hazardous wastes. Owners, past and present, responsible for the contamination of a site become responsible for the clean up and remediation of the contaminated site (40 C.F.R. §§ 261.21-261). The liability of owners of toxic sites may extend to areas outside of their facility. In California, an owner of a facihty that used perchlorate was held liable for the cleanup and remediation of a public water source well outside of the facility's boundaries after perchlorate seeped into the ground water

(Castaic Lake Water Agency v. Whittaker Corp., 2003).

The threat of civil liability for damages caused by perchlorate contamination may aid in stemming future perchlorate contamination. California has been the hotbed of perchlorate related civil Utigation. Most cases filed are class action suits due to the large number of plaintiffs seeking similar relief Class action suits allow citizens, who individually would not have the resources to pursue a claim, to pool their monetary

16 resources with others with like afflictions in order to seek legal relief The certification of plaintiffs as a class is one of many obstacles in class action perchlorate litigation. In

California, plaintiffs from a city sued a corporation that released several harmful chemicals, including perchlorate, into the water supply. The plaintiffs sought for the corporation to fiind a health-monitoring program as well as pay punitive damages due to the health effects caused by the released chemicals. The Superior Court of California ultimately denied a class action certification of the plaintiffs, who potentially numbered between 50,000 and 100,000 people, due to a lack of a demonstration of predominant issues. While all of the plaintiffs experienced exposure to harmfiil chemicals, the level of exposure was different, either due to differing lengths of exposure or amount of exposure due to differing quantities of water consumption between the plaintiffs. The plaintiffs' differing doses made every plaintiffs issues and potential remedies unique to themselves, thus preventing their certification as a class (Lockhead Martin Corp. v. Superior Court,

2003). This limitation on class action certification severely hinders chemical contamination Utigation. When a class certification requires plaintiffs to be exposed to relatively the same cimounts of a chemical in order to obtain class certification, the suit, along with many of the plaintiffs' chances of bringing a successful claim separately is denied.

17 CHAPTER III

AMMONIUM PERCHLORATE-INDUCED LESIONS IN ZEBRAFISH KIDNEYS.

Introduction

Ammonium perchlorate (AP) is a strong oxidant used primarily as a rocket propellant in military and aerospace industries and operations (Fisher et al, 2000). After the dissociation of AP's ammonium and perchlorate ions in water, the resulting perchlorate anion persists in the environment for a relatively long time (Fisher et al,

2000; Urbansky, 2002). In the United States, of the 39 states with facilities where AP is manufactured or used, indicating possible perchlorate contamination, 19 states have confirmed perchlorate contamination in ground or surface water (Sharp and Walker,

2001).

Perchlorate affects the fimction of the thyroid gland by inhibiting sodium-iodide symporters that regulate the amount of iodide taken into thyroid follicles (Wolf, 1998).

Reduced thyroid foUicle uptake of iodide retards the production and release of th5Toid hormones thereby reducing the amount of circulating thyroid hormones. In response, the pituitary gland acts to trigger the release of thyroid hormones by releasing thyroid- stimulating hormones (TSH). However, the impaired production of thyroid hormones causes the continual hyperactive stimulation of thyroid follicles by TSH, and an excess of circulating TSH (Katz, 1975; Wolf, 1998; Soldin et al, 2001). Laboratory studies have confirmed these effects of perchlorate on thyroid function by finding that environmentally relevant perchlorate exposures decrease circulating thyroid hormones

18 levels and increase circulating TSH levels in mammals and amphibians (Brechner et al,

2000; York et al, 2001b; Goleman et al, 2002a). Further, studies reveal that perchlorate exposure leads to increased thyroid follicular cell hypertrophy, hyperplasia, angiogenesis, and colloid depletion in mammals and zebrafish (Fernandez Rodriguez et al, 1991; Siglin et al, 2000; York et al, 2001b; Patiiio et al, 2003).

Altered thyroid functions caused by perchlorate may consequently affect thyroid- dependent reproductive and developmental processes such as bone and neurological development in humans (Sharp and Walker, 2001). Environmentally relevant exposures to perchlorate cause some developmental abnormaUties in frogs (Goleman et al,

2002a,b), mice (Thuett et al, 2002), and zebrafish (Brown, 1997). However, no clear effects of environmentally relevant exposures to perchlorate on reproduction function have been observed in zebrafish (Patino et al, 2003) or rodents (York et al, 2001 a,b).

Whereas the effects of perchlorate on the thyroid, development, and reproduction are relatively well studied, effects on the renal system are somewhat unknowTi. Thyroid hormones influence renal fiinctions such as kidney growth, development, plasma flow, filtration rates, and electrolyte metaboUsm (SelUvanova et al, 1968; Capasso et al, 1999;

Prasad et al, 1999). Further, the discovery of mammaUan and teleost renal thyroid hormone and TSH receptors (MacLatchy and Eales, 1992; Dutton et al, 1997; SeUitti et al, 2000; Vischer and Bogerd, 2003), as weU as mammaUan perchlorate-sensitive sodium-iodide transporters (Spitzweg et al, 2001), opens the possibiUty of perchlorate- induced effects of renal function by indirectly affecting thyroid hormones and TSH sensitive renal processes, or by directly inhibiting renal iodide transport.

19 Materials and Methods

Fish and Exposure Regimens

Samples for this study were generated during a previous study of AP effects on zebrafish reproduction and thyroid function (Patifio et al, 2003). Adult zebrafish were exposed to AP using a static-renewal procedure at concentrations of 6-ppb (control), 18- ppm, and 677-ppm. Four to five female tank repUcates of 8 fish, and two male tank repUcates of twelve fish were used per concentration for a total of 6-7 repUcates per treatment. Because of the differential effects of AP at different concentrations on reproductive performance (the focus of the original study; Patifio et al, 2003), the 677- ppm groups offish were sampled at four weeks of exposure whereas the control and 18- ppm groups were sampled at eight weeks. For the present study, at least three fish per treatment repUcate (aquarium) were prepared for histological analysis of the kidney.

Histological Preparations

After euthanasia by immersion in 1 g/L MS-222, the fish were fixed whole in

Bouin's solution for 48 h at 4°C, rinsed in water for several hours, and placed in 70% ethanol The trunk region of the fish was embedded in paraffin and serial sections of the kidney (5-6 |im) were obtained with a microtome. SUdes were stained with Harris's hematoxyUn stain and counterstained with eosin (Sigma Diagnostics, St. Louis, MO).

Select sUdes were also stained with Mason's Trichrome Stain (Sigma Diagnostics, St.

20 Louis, MO), periodic acid Shiff (PAS) reagent (Sigma Diagnostics, St. Louis, MO), and

Wright's Iron Stain (Sigma Diagnostics, St. Louis, MO).

Histological Observations

A preliminary microscopical analysis at lOOx total magnification for general abnormaUties was performed on all of the fish samples. If these observations resulted in the finding of lesions or chemges of potential interest, further examination was conducted.

Following these preliminary observations, all subsequent histological observations were recorded using the same tissue section for each fish sample. The sections were chosen according to their histological integrity and quaUty when viewing the first row of sections on the sUde from left to right. To measure the incidence of each histological character of

interest, a 1 x l-mm grid containing 121 crosshairs was positioned over the left lobe of the kidney (oriented so that the spine was at the top of the view) covering as much kidney

tissue area as possible at lOOx total magnification. Crosshairs of the grid that fell on each

type of lesion were counted. Crosshairs that feU outside the kidney or on large blood

vessels were regarded as misses and subtracted from the total crosshair count.

Estimations of the area of kidney sections covered by a particular lesion were determined

by dividing the number of crosshairs faUing on the lesion by the total crosshair count.

These measurements were performed three separate times on each fish sample using the

same tissue section.

21 Statistical Analvsis

For each lesion, the average of the three measurements was designated as the fish value. The fish values within each aquarium were then averaged to determine the tank value. The tank values, grouped by treatment, were then subjected to statistical analysis

(Patiiio et al, 2003).

Analyses were conducted using the STATISTICA for Windows 1998 software package (StatSoft®; Tulsa, OK, USA). Because of the difference in treatment durations

(Patifio et al, 2003), the 18-ppm and 677-ppm groups could not be compared to each other. Thus, the following plarmed comparisons were made: control versus 18 ppm, and control versus 677 ppm. The assumption made for the latter comparison is that the kidney condition in the control group (sampled at 8 weeks) is unlikely to have changed significantly from the time the 677-ppm group was sampled at 4 weeks. The data was analyzed using Student's ^tests to determine treatment effects, and a 2-way ANOVA test to determine effects caused by treatment and sex. Homogeneity of variances was assessed using the Cochran C statistic and the Bartlett Chi-square test, and appropriate data transformations were used if needed to correct non-homogeneities. Statistical differences were considered significant at overaU a of 0.05.

Results

There were 6, 7, and 6 aquarium repUcates for the control, 18-ppm, and 677-ppm treatments, respectively (Patiiio et al, 2003). Observations from the present study showed that fish from one male control aquarium and one male 18-ppm aquarium had signs of

22 kidney mycobacterial infection, which exacerbated the incidence of some of the lesions of interest. These two aquaria were omitted from the statistical analysis, leading to a final 5, 6, and 6 replicate count for the control, 18-ppm, and 677-ppm treatments, respectively. The omission of these two male tanks resulted in one male tank for the control and 18-ppm groups for which to conduct a sex-based analysis. Due to the low sample size of males in this study, a statistical analysis based on sex was not feasible.

Based on the observations made during the preliminary scan of histological sections, the specific kidney lesions recorded in subsequent detailed observations were macrophage aggregates (MAs), focal granulomas, inflammation, and generating tubules.

In sections stained with hematoxylin-eosin, macrophage aggregates were observed as clear or light-brown clusters of macrophages of varying sizes (Fig. 1). CeUs in some of the aggregates were positive for PAS (Fig. 2) and Wright's Iron Stain (Fig. 3) indicating the presence of polysaccharides and iron, respectively. Macrophage aggregates were most readUy observed in the 18-ppm group, and the results of the Student's /-test indicated significantly higher levels in the 18-ppm group when compared with the control group {P

= 0.015639, square-root data transformation; Fig. 4, Table 1). However, no significant differences were observed between the 677-ppm and control groups (P > 0.05; square- root data transformation; Fig. 4, Table 1). Focal granulomas were characterized by dark- staining, basophilic, encapsulated, cyst-Uke masses (Fig. 5). The occurrence of focal granulomas was not statisticaUy significant between the control and 18-ppm groups

(Student's /-test, P > 0.05, square-root data transformation; Fig. 4, Table 1), or the control and 677-ppm groups (Student's /-test, P > 0.05; square-root data transformation; Fig. 4,

23 Table 1). Inflammation was sometimes associated with focal granulomas and was characterized by aggregations of eosinophiUc cells around a darkly staining eosinophiUc core (Fig. 6). The incidence of inflammation was not significantly different between the control and 18-ppm groups (Student's /-test; P > 0.05, square-root data transformation;

Fig. 4, Table 1), or the control and 677-ppm groups (Student's /-test; P > 0.05; square- root data transformation; Fig. 4, Table 1). Newly forming kidney tubules were observed

as darkly stained (basophilic) small tubules or cell clusters (Fig. 7). The incidence of

generating kidney tubules was not significantly different between the control and 18-ppm

groups (Student's /-test; P > 0.05; Fig. 4, Table 1), or the control group and the 677-ppm

group (Student's /-test; P> 0.05; Fig. 4, Table 1).

Discussion

The present study showed that kidneys of zebrafish exposed to 18-ppm AP for

eight weeks have an increased incidence of macrophage aggregates (MAs). Increases in the number and size of MAs are considered indicators of exposure to environmental contaminants (Wolke, 1992), and may serve as an immunotoxicologic biomarker (Weeks et al, 1992; Blazer et al, 1997). MAs are a common response to tissue injury (WoUce,

1992); therefore, the increase in kidney MAs in zebrafish exposed to AP suggests that AP affects renal condition.

The renal MAs of zebrafish in the present study contained some melanin, hemosiderin, and PAS-positive components. Melanin containing MAs are common in many tissues of fishes, including the kidney, and aid in several responses to ceUular

24 injury (Weeks et al, 1992; Blazer et al, 1997). The presence of hemosiderin, a protein- bound iron pigment, indicates the breakdown and storage of iron-containing cellular products. Therefore, an increase in the amount of hemosiderin-containing MAs could indicate alterations in the hemopoietic function of kidneys (Wolke, 1992). PAS-positive

MAs indicate the presence of polysaccharide deposits.

Based on the findings of a similar study conducted on mosquitofish, the increase in MAs found in zebrafish kidneys in the present study is likely due to perchlorate exposure and not ammonium exposure. In the mosquitofish study, adult mosquitofish were exposed to 0, 1, 10, and 100-ppm of sodium perchlorate for eight weeks. The 100- ppm group, when compared with the control, had a significant increase in MAs, like is

seen in the present study (Mukhi et al, unpubUshed data). Due to the similar effects from

exposure to different perchlorate species, it can be assumed that in the present study,

ammonium did not play a substantial role in the increased incidence of MAs.

The present study showed that perchlorate exposures at 18-ppm for eight weeks,

but not at 677-ppm for four weeks, induced an increase in kidney MAs in zebrafish. This

pattern of AP effects in the kidney closely paraUels the pattern of histopathological

effects in the thyroid reported in a previous study with the same fish (Patino et al, 2003).

The effects of AP on foUicular ceU hypertrophy, hyperplasia, and coUoid depletion in fiish

exposed to 18-ppm for eight weeks were much greater than in fish exposed to 677-ppm

for four weeks (Patiiio et al, 2003).

The similarity in the patterns of AP effects in the thyroid (Patino et al, 2003) and

the kidney (present study) suggests that the effects of AP in both tissues are linked in

25 zebrafish. The presence of thyroid hormone receptors and TSH receptor mRNA in teleost kidneys (MacLatchy and Eales, 1992; Vischer and Bogerd, 2003) suggests the possibiUty of perchlorate-dependent effects on the kidneys via the reduction of thyroid hormone levels or the subsequent increase in circulating TSH. A time-dependent inhibitory effect of perchlorate on thyroid condition, as seen in mammals (Fernandez Rodriguez et al,

1991), possibly explains the greater thyroid histopathological effects (Patiiio et al, 2003) and higher incidence of renal MAs (present study) in zebrafish exposed to 18 ppm perchlorate for eight weeks when compared to zebrafish exposed to 677 ppm for four weeks. An ahemative scenario is that perchlorate directly affected kidney condition thus

leading to the increased incidence of renal MAs. However, this scenario does not clearly

account for the similar time-concentration patterns of response of the thyroid (Patifio et

al, 2003) and kidney (present study) to AP exposure.

26 Figure 1: Hemotoxilyn-Eosin stained macrophage aggregate. Arrow pomts to melamn containing macrophage aggregate. Bar = 50 |um.

27 r r

k1 Figure 2: PAS-stained macrophage aggregates. Arrow points to red staining polysaccharide deposits. Bar = 100 ^im.

28 Figure 3: Iron-stained macrophage aggregate. Arrows point to blue staining hemosiderin deposits (Arrow). Bar = 50 |am.

29 3n •D ^•6 ppb « ^^18 ppm > 1=1677 ppm 0 U c 2- (0

C 1-

i. « Q. i I^ i ii Focal New Inflammation Macrophage Granuloma Nephrons Aggregates Lesions

Figure 4: Percent kidney area covered by lesions in zebrafish exposed to ammonium perchlorate. Asterisk indicates significant difference (P < 0.05).

30 Lesion: Control 18-ppm 677-ppm Macrophage Aggregates 0.48 2.385 0.39 /'= 0.0156 P= 0.3087 Focal Granuloma 0.0 0.367 0.0 P= 0.095 Inflammation 0.0 0.185 0.0 P= 0.1878 New Nephrons 0.18 0.98 0.22 P= 0.0996 P= 0.7759

Figure 5: Mean area of kidney histological sections covered by lesions, and/?-values of statistical comparison. The 18-ppm group and 677-ppm group were compared to the control group separately.

31 Figure 6: Hemotoxylin-Eosin stained focal granuloma with basophiUc encapsulation. Bar = 40 pm.

32 Figure 6: HemotoxyUn-Eosin stained inflammation with eosinophiUc ceUs and core (arrow). Bar = 50 |um.

33 Figure 7: HemotoxyUn-Eosin stained generating tubule cluster (Circle). Note the highly basophiUc nature of the tubules. Bar = 100 pm.

34 CHAPTER IV

CONCLUSIONS

The presence of MAs in the kidney of zebrafish exposed to AP at 18 ppm may be an indicator of a toxic renal response to environmentally relevant (Smith et al, 2001), concentrations of perchlorate. The increase in renal MAs also suggests a link to a perchlorate inhibited thyroid system. These observations indicate that perchlorate has effects on organisms beyond its classical effects on the thyroid and general development, and extends to other physiological systems that may be affected by perchlorate directly, or through an altered thyroid system. Due to the many systems affected by the thyroid system, including the renal system, research into the effects of perchlorate on systems affected by the thyroid system are necessary for establishing safe standards of perchlorate consumption.

This study also raises several other issues for fiirther study. The increased incidence of MAs suggests the need for additional and more detailed studies of the effect of perchlorate on kidney function. Increased MAs are a structural kidney abnormality.

However, their increased presence may not indicate a harmfiil effect of AP on the kidney.

Since an AP induced structural effect on zebrafish kidneys occurs, it is important to determine if renal function is adversely and irreversibly affected in zebrafish, as well as mammals. Further, the present study revealed a renal effect at an exposure of 18-ppm for eight weeks. Whether or not there is a similar effect at smaller doses for longer periods of time, or whether the same effect can be found in mammals at similar times and doses.

35 or any time and dose, needs to be determined in order to establish a safe level of AP exposure for humans.

In conclusion, while there is much known about the effects of perchlorate on certain physiological processes, such as the thyroid and development, the effects of perchlorate on other systems needs fiirther study.

36 LITERATURE CITED

Blazer VS, Fournie JW, Weeks-Perkins BA. 1997. Macrophage aggregates: biomarker for immune fiinction in fishes? In: Dwyer FJ, Doane TR, Hinman ML, editors. Environmental Toxicology and risk assessment:modeling and risk assessment Volume 6. Philadelphia (PA): American Society for Testing and Materials. Nr 1317.

Brechner RJ, Parkhurst GD, Humble WO, Brown MB, Herman WH 2000. Ammonium perchlorate contamination of Colorado River drinking water is associated with abnormal thyroid fiinction in newborns in Arizona. Journal of Occupational and Environmental Medicine 42:777-82.

Brown DD. 1997. The role of thyroid hormone in zebrafish and axolotl development. Proc National academy of Science USA 94:13011-16.

Capasso G, De Tommaso G, Pica A, Anastaio P, Capasso J, Kinne R, De Santo NG. 1999. Effects of thyroid hormones on heart and kidney fiinctions. Mineral and Electrolyte Metabolism 25:56-64.

Crump C, Michaud P, Tellez R, Reyes C, Gonzalez B, Montgomery EL, Crump KS, Lobo G, Becerra C, Gibbs JP. 2000. Does perchlorate in drinking water affect thyroid fiinction in newborns or school-age children? Journal of Occupational and Environmental Medicine 42:603-12.

California Department of Health and Safety. 2003. Perchlorate in California drinking water: Status of regulafions and monitoring results, http ://www.dhs.ca.gov/ps/ddwem/chemicaL'perchl/perchlindex.htm. Last updated April 1,2003.

Dutton CM, Joba W, Spitzweg C, Heufelder AE, Bahn RS. 1997. Thyrotropin receptor expression in adrenal, kidney, and thymus. Thyroid 7:879-84.

Federal Register. 2000. 40 CFR Part 141. Unregulated contaminant monitoring regulation for public water systems; analytical methods for perchlorate and acetochlor; armouncement of laboratory approval and performance testing (PT) program for the analysis of perchlorate. Federal Register 65:11386.

37 Fisher J, Todd P, Mattie D, Godfey D, Narayanan L, Yu K, 2000. Preliminary development of a physiological model for perchlorate in the adult male rat: a framework for further studies. Drug and Chemical Toxicology 23:243-258.

Fernandez Rodriguez A, Galera Davidson H, Salguero Villadiego M, Moreno Fernandez A, Martin Lacave I, Fernandez Sanz J. 1991, Induction of thyroid proliferative changes in rats treated with antithyroid compound. Anatomic Histological Embryology 20.2%9-9^.

Gibbs JP, Ahmad R, Crump KS, Houck DP, LeveiUe TS, Findley JE, Francis M. 1998, Evaluation of a population with occupational exposure to airborne ammonium perchlorate for possible acute or chronic effects on thyroid fiinction. Journal of Occupational and Environmental Medicine 40:1072-82.

Goleman WL, Carr JA, Anderson TA. 2002. Environmentally relevant concentrations of ammonium perchlorate inhibit thyroid fimction and alter sex ratios in developing Xenopus laevis. Environmental Toxicology Chemistry 21:590-7.

Goleman WL, Urquidi LJ, Anderson TA, Smith EE, Kendall RJ, Carr JA, 2002, Environmentally relevant concentrations of ammonium perchlorate inhibit development and metamorphosis in Xenopus Laevis. Environmental Toxicology Chemistry 21:424-30.

Greer MA, Goodman G, Pleus RC, Greer SE. 2000. Does environmental perchlorate exposure alter human thyroid function? Determination of the dose-response for inhibition of radioiodine uptake. Endocrine Journal 47(suppl.): 146.

H.R. 2123. 2003. Preventing Perchlorate Pollution Act. May 15, 2003,

Jolly J. 1923. "Traite Technique d'hematologie" Paris: A. Maloine et Fils

Jones TC, Hunt RD. 1983. Veterinary pathology. 5th ed. Philadelphia, PA, Lea & Febiger. 1792 pp.

Katz Al, Emmanouel DS, Lindheimer MD. 1975. Thyroid hormone and the kidney. Nephron 15:223-49.

Lamm SH, Braverman LE, Li FX, Richman K, Pino S, Howearth G, 1999. Thyroid heahh status of ammonium perchlorate workers: a cross-sectional occupational health study. Journal of Occupational and Environmental Medicine 41:248-60.

Lawrence JE, Lamm SH, Pino S, Richman K, Braverman LE. 2000. The effect of short- term low-dose perchlorate on various aspects of thyroid fiinction. Thyroid 10:659-63.

38 Li FX, Byrd DM, Meyhle GM, Sesser DE, Skeels MR, Katkowsky SR, Lamm SH. 2000. Neonatal thyroid-stimulating hormone level and perchlorate in drinking water. teratology 62:429-'},\.

Li FX, Squartsoff L, Lamm SH 2001. Prevalence of thyroid diseases in Nevada counfies with respect to perchlorate in drinking water. .Journal of Occupational and Environmental Medicine 43:63 0-4.

Li Z, Li FX, Byrd D, Deyhle GM, Sesser DE, Skeels MR, Lamm SH. 2000. Neonatal thyroxine level and perchlorate in drinking water. Journal of Occupational and Environmental Medicine 42:200-5.

Lockheed Martin Corp. v. Superior Court, 29 Cal 4th 1096, 1101-1102 (Cal, 2003).

Logan BE. 2001. Assessing the outlook for perchlorate remediation. Environvironmental Science and Technology 35: 482A-487A.

MacLatchy DL, Eales JG. 1992. Intra- and extra-ceUular sources of T3 binding to putative thyroid hormone receptors in liver, kidney, and gill nuclei of immature rainbow trout, Oncorhynchus mykiss. Journal of Experimental Zoology 262:22-9.

Motzer WE. 2001. Perchlorate: Problems, detection, and solutions. Environmental Forensics 2:301-312.

Mukhi S, Theodarakis C. 2002. UnpubUshed data, Texas Institute of Environmental and Human Health.

Patifio R, Wainscott MR, Cruz-Li EI, Balakrishnan S, McMurry C, Blazer VS, Anderson TA. 2003. Effects of ammonium perchlorate on the reproductive performance and thyroid condition of zebrafish. Environmental Toxicology and Chemistry, in press.

Prasad R, Kumar V, Kumar R, Singh KP. 1999. Thyroid hormones modulate zinc transport activity of rat intestinal and renal brush-border membrane. American Journal of Physiology 276:E774-82.

Roberts RJ. (1975). Melanin-containing cells of the teleost fish and their relation to disease. In: Ribelin WE, Migaki G. (eds.) The pathology of fishes. Madison, WI, University of Wisconsin Press, pp 399-428.

SelUtti DF, Akamizu T, Dooi SQ, Kim GH, Kariyil JT, Kopchik JJ, Koshiyama H. 2000. Renal Expression of two 'thyroid-specific' genes: thyrotropin receptor and thyroglobulin. Experimental Nephrology 8:235-43.

39 Sellivanova LN, Vorobieva EN. 1968. Toxicology of ammonium perchlorate with its repeated action on the organism. Farmakologieiia I Toksikologieiia 32:480-2.

Sharp R, Walker B. 2001._Rocket Science: Perchlorate and the toxic legacy of the cold war. Environmental Working Group Report, Washington, DC.

Siglin JC, Mattie DR, Dodd DE, Hildebrandt PK, Baker WH. 2000. A 90-day drinking water toxicity study in rats of the environmental contaminant ammonium perchlorate. Toxicological Science 5 7:61 -74.

Smith PN, Theodorakis CW, Anderson TA, Kendall RJ, 2001. Preliminary assessment of perchlorate in ecological receptors at the Longhorn Army Ammunition Plant (LHAAP), Karnack, Texas. Ecotoxicology 105:305-13.

Soldin OP, Braverman LE, Lamm SH. 2001. Perchlorate clinical pharmacology and human heath: a review. Therapeutic DrugMonitor 23:316-31.

Spitzweg C, Dutton CM, Castro MR, Bergert ER, GoeUner JR, Heufelder AE, Morris JC. 2001. Expression of the sodium iodide symporter in human kidney. Kidney International 59:1013-23.

Thuett KA, Roots EH, Mitchell LP, Gentles BA, Anderson TA, Smith EE. 2002. In utero and lactational exposure to ammonium perchlorate in drinking water: effects on developing deer mice at postnatal day 21. Journal of Toxicology and Environmental Health 65:1061 -76.

Urbansky ET. 2002. Perchlorate as an environmental contaminant. Environmental Science Pollution Research International 9:187-92.

USEPA. 2003. Memorandum: Status of EPA's interim assessment guidance for perchlorate. January 22, 2003.

Vischer HF, Bogerd J. 2003. Cloning and fiincfionalcharacterizatio n of a tesficular TSH receptor cDNA from the African catfish (Clarias gariepinus). Journal of Molecular Endocrinology 30:227-38,

Weeks BA, Anderson DP, DeFour AP, Fairbrother A, Govern AJ, Lahvis GP, Peters G 1992. Immunological biomarkers to assess environmental stress. In: Huggett RJ, Kimerle RA, Mehrle PM, Jr., Bergman HL, editors. Biomarkers, physiological, and histological markers of anthropogenic Stress. Boca Raton (FL): Lewis PubUshers. P 211-34.

40 Wolf CD, 2000. Report of the peer review of the thyroid histopathology from rodents and rabbits exposed to ammonium perchlorate in the drinking water. United States Environmental Protection Agency Memorandum. May 5 2000.

Wolf J. 1998. Perchlorate and the thyroid gland. Pharmacology Review 50:S9-\05.

Wolke RE. 1992. Piscine macrophage aggregates: a review. Annii. Rev. Fish Dis. 2:91- 108.

Wu, J., R. Unz, H. Zhang and BE. Logan, 2001. Persistence of perchlorate and the relative numbers of perchlorate- and chlorate-respiring microorganisms in natural waters, soils and wastewater. Bioremediation Journal 5:119-130.

Yoffey JM. 1929. A contribution to the study of the comparative histology and physiology of the spleen, with reference chiefly to its cellular constituents. J. Anat. 63:314-44.

York RG, Brown WR, Girard MF, DoUarhide JS. 2001. Two-generation reproduction study of ammonium perchlorate in drinking water in rats evaluates thyroid toxicity. International Journal of Toxicology 20:183-97.

York RG, Brown WR, Girard MF, DoUarhide JS. 2001. Oral (drinking water) developmental toxicity study of ammonium perchlorate in New Zealand white rabbits. Jnternational Journal of Toxicology 20:199-205.

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