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The Effects of Dioxin Exposure on the Hepatic Development of The Effects of Dioxin Exposure on the Hepatic Development of

Larval Shortnose Sturgeon (Acipenser brevirostrum) Larval Shortnose Sturgeon (Acipenser brevirostrum)

Kelsey Moon University of Maine

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THE EFFECTS OF DIOXIN EXPOSURE ON THE HEPATIC DEVELOPMENT

OF LARVAL SHORTNOSE STURGEON (ACIPENSER BREVIROSTRUM)

By

Kelsey G. Moon

A Thesis Submitted in Partial Fulfillment of the Requirements for a Degree with Honors

(Marine Science)

The Honors College

University of Maine

May 2017

Advisory Committee: Heather J. Hamlin, Assistant Professor of Marine Science, Advisor Rebecca J. Van Beneden, Professor and Director of Marine Science

Rhian G. Waller, Associate Professor of Marine Science Gayle B. Zydlewski, Associate Professor of Marine Science David S. Gross, Adjunct Associate Professor of English, Honors College

ABSTRACT

Exposure to environmental pollutants is considered to be detrimental to the

larval development of various fish species. The liver is thought to be a target organ

for many chemicals because of its high lipid content which bioaccumulates lipophilic

contaminants. Common impacts of dioxins found in fish livers include lipidosis and

glycogen depletion. However, very few studies have been conducted to show these

developmental issues in the early stages of long-lived species such as sturgeon. Given

their bottom dwelling nature, sturgeon are more vulnerable to environmental

pollutants such as dioxin which have long half-lives in aquatic sediments. This study

sought to determine the impacts of the persistent organic pollutant, dioxin, on the

hepatic development of the 25 larval shortnose sturgeon (Ascipenser brevirostrum).

The average area and numbers of vacuoles located in livers of larval sturgeon exposed

to different dioxin concentrations including 50 ppb, 0.5 ppb, and 0.005 ppb, as well as

a treatment (water) and a vehicle (acetone) control, were examined. There were no

significant differences between the concentration of dioxin exposure and the average

area and numbers of vacuoles in the livers of larval sturgeon examined in this study.

In the future, larger sample sizes should be used to determine the effects of dioxin on

liver development and other organs should be examined to determine overall

developmental impacts.

iii

ACKNOWLEDGEMENTS

I would like to gratefully acknowledge Dr. Heather Hamlin for providing me

access to her laboratory and the tools necessary to conduct the research in this study. I

would also like to acknowledge LeeAnne Thayer, who helped me learn the necessary

techniques in histology to conduct this study. Finally, I would like to thank my

parents for helping me proofread this thesis and for providing me with opportunity to

attend and study at the University of Maine.

iv

TABLE OF CONTENTS

I. Introduction 1

Persistent Organic Pollutants and Dioxins 1

Dioxins and Humans 2

TCDDs and Fish 2

Impacts of TCDDs on Fish Livers 3

Environmental Pollutants and Sturgeon 5

The shortnose sturgeon (A. brevirostrum) and TCDD 6

II. Materials and Methods 8

Larval Exposure to Dioxin 8

Dehydration and Embedding 8

Sectioning, Staining, and Cover-slipping 9

Photographing Histology Samples 9

Statistical Analysis 9

III. Results 10

IV. Discussion 13

References 16

Appendix 18

Protocols Used 18

Author’s Biography 21

v

LIST OF FIGURES AND TABLES

Figure 1 – Histological Photographs of Sturgeon Livers 11

Figure 2 – Average Vacuole Numbers for Each Sturgeon Liver at Each

Dioxin Concentration

12

Figure 3 – Average Vacuole Areas for Each Sturgeon liver at Each

Dioxin Concentration

12

1

INTRODUCTION

Section 1 - Persistent Organic Pollutants and Dioxins

Persistent organic pollutants, or POPs, have been plaguing the environment

for almost a century. Among this large group of pollutants are the polychlorinated

dibenzo-p-dioxins (PCDDs) and the tetrachlorodibenzo-p-dioxins (TCDDs). These

chemicals are unintentionally produced by humans through the combustion or

industrial synthesis of other chemicals, meaning that they are not primary industrial

products. Specifically, the dioxin family is formed and released into the environment

from the combustion of deliberately manufactured compounds like

pentachlorophenols (PCPs) and polychlorinated biphenyls (PCBs), as well as the

resulting chemical, industrial, and municipal wastes from industries such as metal

refining (Jones and Voogt, 1999; Safe, 1986).

The danger behind any POP is its persistence in the environment. These toxic

pollutants have long half-lives in soils, sediments, air, and biota (Jones and Voogt,

1999). Due to their propensity to enter the gas phase, POPs can easily evaporate into

the atmosphere, and due to their long half-life and recalcitrant nature, do not readily

break down in air (Jones and Voogt, 1999). This allows these pollutants to travel long

distances before eventually being re-deposited (Jones and Voogt, 1999). POPs are

both hydrophobic and lipophilic. This means that they partition into solids, such as

organic matter that settles into sediments, and into the lipids of organisms (Jones and

Voogt, 1999). This issue, along with the long half-life, causes chemicals such as

dioxins to bioaccumulate and magnify in food chains.

2

Section 2 – Dioxins and Humans

Humans can be exposed to dioxins accidentally through their occupation or

through the environment. Industrial workers who are engaged in manufacturing

PCBs, chlorinated phenols, and other derived products involved with dioxins may

become exposed to these contaminants (Safe, 1986). Several accidents involving the

release of dioxin into the environment have occurred, resulting in high levels of

exposure to humans. An example of this is the Seveso accident in Italy which

occurred in a small chemical manufacturing plant (Safe, 1986). Environmentally,

trace levels (parts per trillion) of PCDDs have been detected in higher trophic levels

including fish, wildlife, and human tissues (Safe, 1986). However, PCDDs only make

up a small fraction of the total bioavailable lipophilic environmental pollutants (Safe,

1986). The major products of trichlorophenol, TCDDs, are another family of

extremely toxic chemicals that have been the primary focus of studies conducted on

the biological and toxicological impacts of the dioxin family (Safe, 1986). Impacts

may include specific organ dysfunction, such as thyroid, liver, or heart, or general

morphological dysfunction, such as changes in body size (Safe, 1986).

Section 3 – TCDDs and Fish

Many studies suggest that fish are among the most sensitive vertebrates to

TCDD exposure, particularly in their early life stages, although this thought has yet to

be proven (Zodrow et al., 2004). Common signs of exposure in larval fish include

pericardial edema, yolk sac edema, craniofacial malformations, reduced blood flow,

anemia, underdeveloped swim bladders, decreased growth, and mortality (Zodrow et

al., 2004). Histopathologic examination is a way to look more in depth into these

specific symptoms of TCDD toxicity. Studies have found TCDD-exposed embryos to

3

display epithelial lesions and arrested gill development, as well as ballooning

degeneration and necrosis of renal tubules, hepatocytes, pancreas, and major brain

regions (Zodrow et al., 2004). The symptoms of TCDD exposure ultimately are

detrimental to the normal development of key organs in fish.

Section 3.1 – Impacts of TCDDs on Fish Livers.

As stated previously, pollutants like TCDDs are lipophilic, and, therefore,

readily partition into lipids. Given the liver’s high lipid content, it has been considered

a target organ for environmental toxicants such as TCDD which can alter liver

structure and metabolism (Brusle and Anadon, 1996). The liver also plays vital

physiological roles, specifically anabolism, the storage of energy into proteins, lipids,

and carbohydrates, and catabolism, the release of energy through glycogenolysis,

detoxication, etc., in fish as well as other vertebrates (Brusle and Anadon, 1996). In

addition to these functions, the liver plays an important role in vitellogenesis, by

producing vitellogenin, the yolk precursor that is sequestered by developing oocytes

(Brusle and Anadon, 1996). The characteristics and functions of the liver may be

negatively impacted by TCDD making it an important organ to study in regards to

environmental pollution. Overall, given their high lipid content, TCDD-exposed adult

fish livers could be lacking the functions crucial for health and survival. Furthermore,

fish larvae may be exposed to TCDD via vitellogenesis from the mother.

The impact of TCDD toxicity on fish livers has been reported in various

studies. A study done on adult zebrafish found that the most significant differences

between the unexposed control and a TCDD exposed sample of fish were observed in

the liver (Zodrow et al., 2004). In this study, zebrafish (Danio rerio) were injected

with3.5 ng/g PC (control) and 70 ng/g TCDD (treatment). Hypertrophy of

hepatocytes, a reduction in the number of hepatocyte nuclei, glycogen depletion, and

4

lipidosis of the liver were observed in adult zebrafish that were exposed to 70 ng/g

TCDD (Zodrow et al., 2004).

A similar study done on yellow perch (Perca flavescens) yielded similar

impacts on the liver. Yellow perch were exposed to various concentrations of TCDD,

ranging from 1 µg/kg to 125 µg/kg, some being exposed via injection of toxins

dissolved in corn oil (study one) and others by direct injection intraperitoneally (study

two) (Spitsbergen et al., 1988). Using histopathology on the epithelial tissues of the

perch livers, researchers found that hepatocyte lipidosis, an excess amount of fats

inside liver cells, occurred in study one where perch had received a dose of 1 µg/kg or

more of TCDD (Spitsbergen et al., 1988). Additionally, all study two perch receiving

a dose of 25 or 125 µg/kg and over 50% of perch receiving a dose of 1 or 5 µg/kg had

hepatocyte lipidosis occurring in sections throughout the liver (Spitsbergen et al.,

1988). Glycogen depletion was also evident in study two perch treated at 25 or 125

µg/kg TCDD (Spitsbergen et al., 1988).

The impacts of TCDD exposure on the larval and early developmental stages

of fish have also been explored. This is an important area to study because TCDDs

and other pollutants can bioaccumulate in the food chain and may, therefore, also be

present in the bodies of fish offspring. A study exploring the impacts of TCDD

exposure on adult zebrafish and their offspring before and after reproduction found

that, at the post-spawning stage, some of the males displayed pronounced lesions with

reduced glycogen stores in the hepatocyte cytoplasm (Heiden et al., 2009). This study

found that exposure to concentrations of waterborne TCDDs may be too low to cause

larval mortality, but it still induced multiple, subtle, toxic responses that can result in

reduced health and survival in the future of an exposed organism (Heiden et al.,

2009). These toxic responses include increased incidence of blue sac syndrome,

5

decreased incidence of swim bladder inflation, and increased incidence of altered

heart function (Heiden et al., 2009).

Section 3.2 – Environmental Pollutants and Sturgeon.

There are 26 species of sturgeon, all of which are in decline. One of the main

reasons for this decline is exposure to chemical toxicants (Chambers et al., 2012). The

ecology of sturgeon make them particularly vulnerable to persistent organic pollutants

like TCDD. As stated previously, pollutants like this have long half-lives in soils,

sediments, air, and biota. The long half-life in sediments is of particular concern to

sturgeon because this fish is a bottom dweller, feeding on benthic prey items that live

in these toxic sediments (Feist et al., 2005). Due to their long life span of up to 100

years, TCDD has an extended amount of time to accumulate, even before

reproduction which occurs at 16-35 years, possibly increasing the chances of larvae

being exposed to TCDD through their mothers (Feist, et al. 2005). Furthermore, there

have been recent studies to suggest that the early life stages of sturgeon are more

sensitive to pollutants than the early life stages of other fish (Kynard et al, 2016). A

study investigating the contaminant sensitivity of endangered aquatic species exposed

both Atlantic and shortnose sturgeon to five different chemicals including carbaryl,

copper, 4-nonylphenol, pentachlorophenol, and permethrin (Dwyer et al., 2005). It

was observed that both species of sturgeon were two of the most sensitive species

used in this study based on the compared ranks of LC50s (mg/L), or the lethal

concentration required to kill 50% of the population (Dwyer et al., 2005). Atlantic and

shortnose sturgeon had the lowest LC50 values when combining all five chemicals,

suggesting they could be the most sensitive of the other species used (Dwyer et al.,

2005). Thus, while more studies are needed, current research suggests that both adult

6

and larval sturgeon (specifically shortnose sturgeon) bioaccumulate dioxins at high

levels that could be physically damaging (Kynard et al, 2016).

A number of studies support that adult and larval sturgeon are negatively

impacted by dioxins. A study on white sturgeon in both free-flowing and reservoir

parts of the Columbia River found 18 types of pesticides and 28 types of PCBs

accumulating in the livers and gonads (Feist, et al. 2005). Specifically, researchers

found high incidences of macrophage aggregates and lymphocytes, which are known

to be present during times of environmental stress, particularly chemical pollution

(Feist, et al. 2005, Agius and Roberts, 2003). This suggests that both adult and larval

sturgeon exposed to any type of contaminant, including TCDD, could be experiencing

negative impacts on liver development and liver function.

Section 3.3 – The shortnose sturgeon (A. brevirostrum) and TCDD.

The shortnose sturgeon is a diadromous species with distributions in the large

Atlantic coastal rivers and estuaries of North America (Kynard et al., 2016). It is

known to live 50 to 60 years, growing to maximum lengths of around 120 cm, and

weighing up to 24 kg (Cope et al., 2011). The shortnose sturgeon was listed as

endangered by the U.S government and remains in danger throughout its global

distribution (Cope et al., 2011). As discussed previously, this endangerment could

possibly be due to the bioaccumulation of toxins such as TCDD, but has also been

found to be due to overfishing and the construction of dams (Kynard et al., 2016).

Ultimately, it is critical to understand the impact of environmental pollutants on

shortnose sturgeon to possibly save this species from extinction.

A study exploring the toxic effects of TCDD and PCB126 on larval shortnose

sturgeon found that TCDD was tenfold more toxic than the PCB126 (Chambers et al.,

2012). TCDD was found to significantly decrease survival to hatching in shortnose

7

sturgeon, with a larval life span of less than five days at concentrations equal to or

greater than one ppb (Chambers et al., 2012). The larvae that did survive displayed a

decrease in size, yolksac length, and eye development (Chambers et al., 2012).

Ultimately, this study showed that shortnose sturgeon will exhibit toxic responses to

TCDD at relevant environmental concentrations (as little as 0.001 ppb), thus

highlighting the importance of learning more about the impacts of these toxicants and

to finding solutions to this ongoing issue.

The primary goal of the following study was to determine the effects of

different concentrations of the persistent organic pollutant TCDD on the number and

average area of vacuoles found in larval shortnose sturgeon livers. Samples were

taken from dioxin exposed sturgeon larvae using histological techniques. The

hypothesis was that several defects would be seen in the liver tissue samples including

increased liver vacuole number and area with an increase in dioxin concentration.

Ultimately, this project sought to reveal the extent to which sturgeon liver cellular

morphology is affected by TCDD exposure and to, therefore, increase our knowledge

of the risks associated with TCDD exposure in a threatened and ecologically

important fish species.

8

MATERIALS AND METHODS

Section 1 – Larval Exposure to TCDD

Shortnose sturgeon embryos were obtained by NOAA’s NEFSC Howard

Laboratory from a hatchery in Canada and kept in an environmental control chamber

in 15°C water until they were mature enough for experimentation. Collaborators at

the Howard Laboratory then exposed 56 to 72 hour old sturgeon embryos to both

PCB-126 and TCDD in a 25µL acetone vehicle, to 25µL of acetone alone, or to

normal water conditions . At hatch, sturgeon larvae were then preserved in Bouin’s

fixative, and delivered to the Hamlin Lab at the University of Maine, Orono. The

contaminant used in this study was TCDD, a persistent organic pollutant known to

bioaccumulate in these animals. A total of 25 sturgeon were used, five per TCDD

concentration. Treatments consisted of water (treatment control) and acetone (vehicle

control), as well as 0.005 ppb, 0.5 ppb, and 50 ppb TCDD.

Section 2 – Dehydration and Embedding

Prior to dehydration, sturgeon larvae were stored in 70% ETOH. The larvae

were then placed into increasing concentrations of ethanol: 85% ETOH (1-hr), 95%

ETOH (1-hr), 100% ETOH (2-hrs), and 100% Citrisolv (2-hrs) to dehydrate samples

prior to wax embedding. Samples were then moved into paraffin wax and placed into

a vacuum oven at 57-62°C at 12 psi. Every hour samples were moved to fresh wax,

and the vacuum oven was set to 13, 14, and 15 psi respectively. Samples were then

placed into a mold and were allowed to cool for 24-hours.

9

Section 3 – Sectioning, Staining, and Cover-slipping

The embedded samples were then sectioned at 6 µm with a rotary microtome.

Samples were mounted on slides using a water bath and were dried on a plate warmer

for 24-hours. Slides were stained using the Mason’s Trichrome Stain protocol

(Appendix), cover-slipped, and allowed to dry overnight.

Section 4 – Photographing Histology Samples

Photographs of slide images were taken using a Nikon eclipse E200

microscope with an attached MVI microscope camera. Each image included a digital

scale bar. Photographs were then analyzed for average vacuole number and area for

sturgeon exposed to each dioxin concentration and to each control using ImageJ

software. Vacuole area was then digitally traced by hand, and vacuole number and

area was calculated using ImageJ. Five photographs and five animals for each

concentration of dioxin were used to collect and analyze data with the exception of

sturgeon at a concentration of 50 ppb, which only have four animals and photographs.

Due to scale bar errors using a microscope from another lab, data accumulated from

one animal exposed to 50 ppb had to be excluded. The accumulated data from each

sample were averaged for each concentration of dioxin, including the two controls.

Section 5 –Statistical Analysis

The average vacuole area and number for each larval sample was then

analyzed statistically using an ANOVA. This one-way analysis of variance was used

to determine whether there were any statistically significant differences between the

means of the five different treatments used in this study at the two different endpoints.

Standard error values were also produced for each dioxin treatment after all larval

samples were averaged.

10

RESULTS

Histology techniques were used to examine the liver vacuoles of dioxin

exposed larval sturgeon. Vacuoles were found in each sturgeon liver, both in exposed

and control sturgeon (Fig 1). Using a one-way ANOVA, it was determined that there

was no significant impact on average liver vacuole area (p = 0.582) and number (p =

0.410) correlated with dioxin concentration or with control treatments (Fig 2 and 3).

However, this study does show the beginning of a trend in vacuole number and dioxin

concentration. As dioxin concentration increases, we can see a possible decreasing

trend in liver vacuole numbers (Fig 2).

11

a) b)

c) d)

e)

Fig 1. Representative liver histology photos, at 40X magnification, from larval sturgeon exposed to 50 ppb (a), 0.5 ppb (b), 0.005 ppb (c), water (d), and acetone (e). Scale bar represents 100 µm. Vacuoles are shown by the black arrows.

12

0  

5  

10  

15  

20  

25  

Water   Acetone   0.005   0.5   50  

Aver

age

Vacu

ole

Num

ber

TCDD Concentration (ppb)

0  

20  

40  

60  

80  

100  

120  

140  

160  

180  

Water   Acetone   0.005   0.5   50  

Aver

age

Vacu

ole A

rea

(µm

2 )

TCDD Concentration (ppb)

Fig 2. Average vacuole numbers for each concentration of TCDD. One-way ANOVA p-value = 0.410. Error bars SE. Fig 3. Average liver vacuole areas for each concentration of TCDD in µm2. One-way ANOVA p-value = 0.582. Error bars SE.

13

DISCUSSION

In this study, exposure to TCDD throughout embryonic development (56 – 72

hours to post-hatch) did not affect hepatic vacuole area or number in larval shortnose

sturgeon. However, the mean number of liver vacuoles show trends approaching

significance, specifically exhibiting a decrease in average vacuole number as 50 ppb

dioxin is approached (Fig 2). Using a larger sample size not only could show

significance in vacuole numbers, but could also influence vacuole area as the

concentration of dioxin increases. This study only used 25 larval sturgeon in total,

while a similar study had a much larger sample size, using up to 30 larval sturgeon

per concentration in multiples of 10 (n ~ 300 animals) (Chambers et al., 2012). Using

a larger sample size like this would mean that the results would be more

representative of a population, thus limiting the influence of outliers and creating a

better chance for statistically significant data.

It is possible that the impacts of dioxin may not manifest until later life

stages, or affect endpoints other than those investigated in this study. Given the long-

lived nature of sturgeon, dioxins may bioaccumulate in the body over time, which

could cause effects not observed in early life stages. Studies conducted to investigate

hepatic function or morphology of fishes are more commonly conducted at juvenile or

adult life stages. For example, a study conducted on yellow perch in the juvenile life

stages found hepatocyte lipidosis and glycogen depletion in yellow perch livers

(Spitsbergen et al., 1998). Similarly, using adult zebrafish found that higher

concentrations of dioxin caused hepatocyte hypertrophy, glycogen depletion, and

lipidosis of the liver (Zodrow et., al 2004). However, another study investigating 96

hour old larval Atlantic and shortnose sturgeon exposed to different toxicants found

these two species of sturgeon to be the most sensitive when compared to other aquatic

14

species (Dyer et al., 2005). Ultimately, our study examined only two hepatic

morphological changes, and therefore more studies investigating a greater type and

variety of endpoints should be conducted to more clearly understand the effects of

dioxin on shortnose sturgeon.

A study conducted on TCDD-exposed zebrafish embryos found their hearts to

be elongated and string-like, with thinner and elongated atria and smaller, more

compact ventricles (Antkiewicz et al., 2005). This study also found that both the

cross-sectional area of the heart chambers and the overall heart size was reduced

(Antkiewicz et al., 2005). Another study conducted on shortnose sturgeon embryos

exposed to TCDD found that most treated fish died two weeks after hatching

(Singleman and Holtzman, 2016). Those that did survive exhibited ballooning defects

in the hearts of fish exposed to lower concentrations and reduced looping and

ballooning in fish exposed to higher concentrations (Singleman and Holtzman, 2016).

Both normal and reverse chamber orientation were also observed in TCDD treated

sturgeon (Singleman and Holtzman, 2016). Thus, future studies may consider looking

at the impacts of TCDD and other environmental pollutants on other vital organs like

the heart.

The route of contaminant exposure can affect the type of severity of the

response. In this study, larval sturgeon were exposed to dioxin via waterborne

exposure. Another study investigating adult zebrafish delivered TCDD via injection,

assuming that this method ensures a consistent delivered dose (Zodrow et al., 2004).

Livers in TCDD -injected zebrafish exhibited hepatocyte hypertrophy, glycogen

depletion, and lipidosis (Zodrow et al., 2004). A study done on yellow perch injected

one group with a dioxin that had been dissolved in corn oil, and another group

intraperitoneally (Spitsbergen et al., 1988). Fish injected with dioxin dissolved in corn

15

oil exhibited hepatocyte lipidosis, while fish injected intraperitoneally exhibited both

hepatocyte lipidosis and glycogen depletion (Spitsbergen et al., 1988). However, a

study investigating the toxicological impacts of PCB-126 and TCDD on larval

shortnose sturgeon used waterborne exposure, finding decreases in size, yolk sac

length, eye development, and overall survivability to hatch (Chambers et al., 2012).

Despite the significant results obtained by the two injection studies, it is strictly a

laboratory procedure, thus lacking the real environmental applications that waterborne

exposure studies like our study and the Chambers et al. study have.

In conclusion, there were no significant effects on the number and area of liver

vacuoles in newly hatched shortnose sturgeon larvae exposed to different

concentrations of dioxin during embryonic development. Although this study did not

observe a significant effect on the parameters measured, other studies investigating

parameters on larval as well as later stage shortnose sturgeon are warranted. The

bioaccumulation of dioxin and other environmental pollutants persists in the aquatic

environment. Ultimately, future studies should examine different organs and life

stages using larger sample sizes to see when and where sturgeon are most vulnerable

to dioxin exposure in order to preserve a declining and protected species.

16

REFERENCES

Agius, C. and Roberts, R.J. 2003. Melano-macrophage centers and their role in fish

pathology. J. Fish Dis. 26: 499-509.

Antkiewicz, D.S., Burns, C.G., Carney, S.A., Peterson, R.E., and Heidemen, W. Heart malformation is an early response to TCDD in embryonic zebrafish. Toxicol. Sci. 84: 368-377.

Brusle, J. and Anadon, G.G. 1996. The structure and function of fish liver. Fish morphology. 6: 77-93.

Chambers, R.C., Davis, D.D., Habeck, E.A., Roy, N.K., and Wirgin, I. 2012. Toxic effects of PCB126 and TCDD on shortnose sturgeon and Atlantic sturgeon. Environ. Toxicol. Chem. 31(10): 2324-2337.

Cope, W.G., Holliman, F.M., Kwak, T.J., Oakley, N.C., Lazaro, P.R., Shea, D., Augspurger, T., Law, J.M., Henne, J.P., and Ware, K.M. 2011. Assessing water quality suitability for shortnose sturgeon in the Roanoke River, North Carolina, USA with an in situ bioassay approach. J. Appl. Icthyol. 27: 1-12.

Dwyer, F. J.; Hardistry, D. K.; Henke, C. E.; Ingersol, C. G.; Whites, D. W.; Augspurger, T.; Canfield, T. J.; Mount, D. R.; Mayer, F. L., 2005: Assessing contaminant sensitivity of endan- gered species: Part III. Effluent toxicity tests. Arch. Environ. Contam. Toxicol. 48, 174–183.

Feist, G.W., Webb, M.A.H., Gundersen, D.T., Foster, E.P., Schreck, C.P., Maule. A.G., and Fitzpatrick, M.S. 2005. Evidence of detrimental effects of environmental contaminants on growth and reproductive physiology of white sturgeon in impounded areas of the Columbia River. Environ. Health Persp. 113(12): 1675 – 1682.

Jones, K.C. and de Voogt, P. 1999. Persistent organic pollutants (POPs): state of the science. Environ. Pollut. 100: 209-221.

King Heiden, T.C., Spitsbergen, J., Heideman, W., and Peterson, R.E. 2009. Persistent adverse effects on health and reproduction caused by exposure of zebrafish to 2,3,7,8-tetrachloroibenzo-p-dioxin during early development and gonad differentiation. Toxicol. Sci. 109(1): 75-87.

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Kynard, B., Bolden, S., Kieffer, M., Collins, M., Brundage, H., Hilton, E.J., Litvak,

M., Kinnison, M.T., King, T., and Peterson, D. 2016. Life history and status of Shortnose Sturgeon (Acipenser brevirostrum LeSueur, 1818) J. Appl. Icthyol. 32: 208-248.

Safe, S.H. 1986. Comparative toxicology and mechanism of action of polychlorinated dibenzo-p-dioxins and dibenzofurans. Ann. Rec. Pharmacol. Toxicol. 26: 371-399.

Singleman, C. and Holtzman, N.G. 2016. Impact of aquatic toxins on heart development in Atlantic and shortnose sturgeon larvae. Section V: 1-27 pp. In S.H. Fernald, D.J. Yozzo and H. Andreyko (eds.), Final Reports of the Tibor T. Polgar Fellowship Program, 2014. Hudson River Foundation.

Spitsbergen, J.M., Kleeman, J.M., Peterson, R.E. 1988. 2,3,7,8-Tetracholordibenzo-p-dioxin toxicity in yellow perch (Perca flavescens). J. Toxicol. and Env. Health. 23: 359-383.

Zodrow, J.M., Stegeman, J.J., and Tanguay, R.L. 2004. Histological analysis of acute toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in zebrafish. Aquat. Toxicol. 66: 25-38.

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APPENDIX: PROTOCOLS

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AUTHOR’S BIOGRAPHY

Kelsey G. Moon was born in Simsbury, Connecticut on October 10, 1994. She

was raised in Simsbury, CT and graduated from Simsbury High School in 2013.

Majoring in marine science, Kelsey is concentrating in marine biology. She played for

the University of Maine Women’s Club Rugby team. She has received the Dean’s

scholarship.

Upon graduation, Kelsey plans to take a year off from her studies to obtain

experience in marine science, and then plans to attend graduate school to pursue her

Master’s degree.