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Cellular apoptosis and proliferation in testes of fathead minnow exposed to wastewater treatment plant effluent

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Title:
Cellular apoptosis and proliferation in testes of fathead minnow exposed to wastewater treatment plant effluent
Creator:
Hallagin, Andrea Elizabeth Geddes ( author )
Place of Publication:
Denver, CO
Publisher:
University of Colorado Denver
Publication Date:
Language:
English
Physical Description:
i online resoruce (225 pages) : ill. ;

Thesis/Dissertation Information

Degree:
Master's ( Master of Science)
Degree Grantor:
University of Colorado Denver
Degree Divisions:
Department of Integrative Biology, CU Denver
Degree Disciplines:
Integrative Biology

Subjects

Subjects / Keywords:
Fishes -- Effect of human beings on ( lcsh )
Minnows ( lcsh )
Water -- Purification ( lcsh )
Fishes -- Effect of human beings on ( fast )
Minnows ( fast )
Water -- Purification ( fast )
Genre:
bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )

Notes

Abstract:
Environmental toxicants and their effects on both wildlife and human populations have become more concerning as worldwide human populations grow, water sources become scarce, and anthropogenic waste becomes more prevalent. Municipal wastewater treatment plants (WWTPs) are designed to conserve water by removing nutrients and pathogens from wastewater and release effluent water back into local streams. Unfortunately, many wastewater compounds are not completely removed from the effluent and are released into the environment, sometimes causing deleterious effects even in minute quantities on biological life. Testing of anthropogenic compounds and effects on wildlife was conducted at the City of Boulder WWTP in 2005, 2006, 2008, and 2011. Biological assays were performed by maintaining adult male fathead minnow in either reference or effluent water for 28 days, and by measuring secondary sex characteristics, plasma vitellogenin, and gonad histology. It was found that significant disruption in gonad weight, sperm abundance, and vitellogenin occurred in effluent exposed fish in 2005, but little or no significant disruption occurred in any later year because of a major WWTP upgrade that improved effluent water chemistry. The aim of this study was to examine biological effects further by testing for discrete cellular changes of the same experimental fish from 2005, 2006, and 2008. This approach involved analyzing biomarkers, which identified cellular proliferation via proliferating cell nuclear antigen (PCNA) and apoptosis via terminal deoxynucleotidyl transferase biotin-dUTP nick end labeling (TUNEL). It was found that epithelial PCNA was higher in effluent-exposed fish in 2005 only, but was lower in sperm cells of effluent fish in 2006, and higher in sperm of reference fish in 2005 and 2008 (p<0.05). Epithelial and sperm TUNEL was greater in effluent exposed fish in 2005 and 2006 (p<0.05), but was not significantly different in 2008. PCNA was found to be inconsistent between years, but TUNEL was consistent with previous measurements, proved to be a more reliable biomarker, and showed that disruption is still occurring in 2008 fish even after the WWTP upgrade. Future use of TUNEL is recommended, as it responds to discrete cellular disruption early and shows disruption when other measurements fail to respond.
Thesis:
Thesis (M.I.S.)--University of Colorado Denver. Integrative biology
Bibliography:
Includes bibliographic references.
General Note:
Department of Integrative Biology
Statement of Responsibility:
by Andrea Elizabeth Geddes Hallagin.

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Source Institution:
|University of Colorado Denver
Holding Location:
|Auraria Library
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All applicable rights reserved by the source institution and holding location.
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868702905 ( OCLC )
ocn868702905

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Full Text
CELLULAR APOPTOSIS AND PROLIFERATION IN TESTES OF FATHEAD
MINNOW EXPOSED TO WASTEWATER TREATMENT PLANT EFFLUENT
by
Andrea Elizabeth Geddes Hallagin
B.A.University of Colorado, 2009
A thesis submitted to the
Faculty of the Graduate School of the
University of Colorado in partial fulfillment
of the requirements for the degree of
Master of Science
Integrative Biology
2013


This thesis for the Master of Science degree by
Andrea Elizabeth Geddes Hallagin
has been approved for the
Department of Integrative Biology
by
Alan Vajda, Chair
Michael Greene
Amanda Charlesworth
April16,2013
li


Hallagin, Andrea, Elizabeth Geddes (M.S., Integrative Biology)
Cellular Apoptosis and Proliferation in Testes of Fathead Minnow Exposed to
Wastewater Treatment Plant Effluent
Thesis directed by Assistant Professor Alan Vajda
ABSTRACT
Environmental toxicants and their effects on both wildlife and human populations
have become more concerning as worldwide human populations grow, water sources
become scarce, and anthropogenic waste becomes more prevalent. Municipal wastewater
treatment plants (WWTPs) are designed to conserve water by removing nutrients and
pathogens from wastewater and release effluent water back into local streams.
Unfortunately, many wastewater compounds are not completely removed from the
effluent and are released into the environment, sometimes causing deleterious effects
even in minute quantities on biological life. Testing of anthropogenic compounds and
effects on wildlife was conducted at the City of Boulder WWTP in 2005, 2006, 2008, and
2011.Biological assays were performed by maintaining adult male fathead minnow in
either reference or effluent water for 28 days, and by measuring secondary sex
characteristics, plasma vitellogenin, and gonad histology. It was found that significant
disruption in gonad weightsperm abundance, and vitellogenin occurred in effluent
exposed fish in 2005but little or no significant disruption occurred in any later year
because of a major WWTP upgrade that improved effluent water chemistry. The aim of
this study was to examine biological effects further by testing for discrete cellular
changes of the same experimental fish from 2005, 2006, and 2008. This approach
involved analyzing biomarkers, which identified cellular proliferation via proliferating
cell nuclear antigen (PCNA) and apoptosis via terminal deoxynucleotidyl transferase


biotin-dUTP nick end labeling (TUNEL). It was found that epithelial PCNA was higher
in effluent-exposed fish in 2005 only, but was lower in sperm cells of effluent fish in
2006, and higher in sperm of reference fish in 2005 and 2008 (p<0.05). Epithelial and
sperm TUNEL was greater in effluent exposed fish in 2005 and 2006 (p<0.05), but was
not significantly different in 2008. PCNA was found to be inconsistent between years
but TUNEL was consistent with previous measurements, proved to be a more reliable
biomarkerand showed that disruption is still occurring in 2008 fish even after the
WWTP upgrade. Future use of TUNEL is recommended, as it responds to discrete
cellular disruption early and shows disruption when other measurements fail to respond.
The form and content of this abstract are approved. I recommend its publication.
Approved: Alan Vajda
IV


ACKNOWLEDGMENTS
I would like to thank members of the Vajda lab including Zia Faizi and Ethan
Cabral for all of their help. I also want to thank all of my friends and family for all of
their support and for putting up with me during this process.


TABLE OF CONTENTS
CHAPTER
I. INTRODUCTION.............................................................1
Worldwide Water Concerns.............................................1
Endocrine Disruption from Water Contaminants...................2
Endocrine Disrupting Compounds.......................................3
Natural and Synthetic Steroid..................................3
Industrial Waste Compounds.....................................4
Pharmaceutical Compounds.......................................5
Boulder Creek........................................................7
Stream Background..............................................7
Effluent Source and Contaminants...............................8
Fish Exposure in Boulder Creek.................................9
Biological Findings from Fish Exposures.......................12
Testing for Discrete Changes........................................13
Endpoints of Disruption.......................................16
Spermatogenesis...............................................16
Proliferation and Spermatogenesis.............................17
PCNA..........................................................17
Apoptosis and Spermatogenesis.................................18
TUNEL.........................................................18
PCNA and TUNEL................................................19
Expected Outcomes...................................................20
vi


II METHODS..............................................................22
On-Site Laboratory and Experimental Design........................22
Double Labeling Procedure.........................................22
Slide Analysis....................................................25
Statistical Analysis..............................................28
III RESULTS..............................................................30
Testicular Morphometries of 2005, 2006, and 2008 Tissue...........30
PCNA and TUNEL Immunoreactivity...................................31
IV. DISCUSSION...........................................................35
Cellular Proliferation............................................35
TUNEL.............................................................36
Use as Biomarkers.................................................38
Use of Biomarkers in Fathead Minnow Testes..................38
Use of Biomarkers in Vertebrates............................39
Use of Biomarkers in Humans.................................40
REFERENCES..............................................................42
vii


List of Tables
Table
111.1 Previous Measurements, Water Chemistry, and Biomarkers
34


LIST OF FIGURES
Figure
1.1 Boulder Creek EEq.............................................................9
1.2 Fathead Minnow {Pimephales promelas).........................................10
1.3 Plasma Vitellogenin in Reference and Effluent Exposed Fish...................11
1.4 Nuptial Tubercles in Effluent and Reference Exposed Fish.....................11
1.5 Gonadosomatic Index of Reference and Effluent Exposed Fish..................12
1.6 Disruption of the Hypothalamus-Pituitary-Gonadal Axis.......................15
11.1 Example of Sectioned Slide Used for Analysis...............................24
11.2 Blind Analysis of Slides....................................................25
11.3 Tubule Morphometries........................................................27
11.4 TUNEL and PCNA Positive Cell Count.........................................28
111.1 Averages of Sperm, Lumen and Epithelium Areas..............................30
111.2 Epithelial and Luminal TUNEL Reactive Cells in 2005........................31
111.3 Epithelial and Luminal PCNA Reactive Cells in 2005.........................31
111.4 Epithelial and Luminal TUNEL Reactive Cells in 2006........................32
111.5 Epithelial and Luminal PCNA Reactive Cells in 2006.........................32
111.6 Epithelial and Luminal TUNEL Reactive Cells in 2008........................33
111.7 Epithelial and Luminal PCNA Reactive Cells in 2008.........................33
IV.1 Use of TUNEL in Combination with Hematoxylin and Eosin Staining............39
ix


LIST OF ABBREVIATIONS
BPA bisphenol-A
EDC endocrine disrupting compound
EDTA ethylenediaminetetraacetic acid
EE2 ethinyl estradiol
ER estrogen receptor
FSH follicle stimulating hormone
GnRH gonadotropin releasing hormone
HPG hypothalamic-pituitary-gonadal
LH luteinizing hormone
NPE nonylphenol ethycarboxylates
SSRI selective serotonin reuptake inhibitor
WWTP wastewater treatment plant


CHAPTER I
INTRODUCTION
Worldwide Water Concerns
We live in a world of finite resources, yet increasing worldwide consumption of
these resources is decreasing our supply at an alarming rate. One of the most rapidly
diminishing essential resources is fresh water, which is a result from increased
consumption over the past century from dramatic increases in human populations,
expanding urbanization, and increased agricultural use (Hinrichsen and Tacio, 2012). Not
only is there a worldwide shortage of water, but increased pollution and the presence of
increasingly varied environmental toxicants pose additional threats as anthropogenic
waste compounds are contaminating what little fresh water is available (Hinrichsen and
Tacio, 2012).
One effort to conserve water and minimize waste is the implementation and
continual improvement of municipal wastewater treatment plants (WWTPs), which are
principally designed to effectively remove nutrients and pathogens from urban
wastewater and release the effluent into local streams and rivers. This process occurs by
first filtering wastewater from debris such as trash, branches, and rocks, and then
removing large particulate matter with grit chambers and settling tanks (Lee et al., 2006).
Smaller compounds are then absorbed or biochemically converted and removed using
trickling filters and/or activated sludge. After these processes, advanced treatment of the
water occurs where pathogens are destroyed, disinfection processes occur, and harmful
1


nutrients are removed (Lee et al.2006). After these WWTP processesthe effluent water
is then released into a local stream where it provides necessary stream flow for the
survival of many ecosystems (Vajda and Norris, 2011).
Although most compounds are removed at the WWTP, many others such as
steroids, pharmaceuticals, and other inorganic and organic compounds are not completely
removedand are therefore released into the environment with effluent water. Most of
these compounds are relatively innocuous, but unfortunately, some have been found to
have deleterious effects on biological life, even in minute quantities (Vajda and Norris,
2011). In addition, effluent compounds are found in complex mixtures and sometimes
can act synergistically when consumed, compounding and multiplying the effects of each
other, resulting in physiological systems being affected at multiple sites and via multiple
mechanisms (Ankley et al., 2009).
Endocrine Disruption from Water Contaminants
Although effluent compounds can affect multiple physiological processes in many
vertebrate speciesa concerning affect these compounds can have is disruption of
reproductive pathways, because not only is the fitness and health of the individual
affectedbut reproductive disruption can decrease fecundity of individuals and can
ultimately affect populations of the species (Ankley et al., 2008). In addition,
reproductive and endocrine pathways are mostly conserved amongst vertebrate species,
so the negative impacts effluent compounds have on one species can also occur in a wide
range of vertebrate populations (Nagahama, 1994). Therefore, because reproductive
pathways are conserved in vertebrate speciesand because disruption of reproductive
2


pathways by effluent compounds can have populations level effects, it is imperative that
the impact of effluent compounds on reproduction be studied to understand not only
individual health affects, but individual fitness, population growth, and ecosystem
impacts as well.
Endocrine Disrupting Compounds
Natural and Synthetic Steroids
The most notable and concerning endocrine and reproductive disrupting
compounds (EDCs) found in effluent water include natural and synthetic steroids, organic
and inorganic industrial waste, and pharmaceutical compounds. Steroidal compounds
such as the natural estrogens estrone (El)17p_estradiol (E2)and estriol (E3)and the
synthetic estrogen 17a-ethynyl estradiol (EE2)are found in wastewater downstream
from populated areas where widespread use and improper disposal of birth control occurs
(Barber et al., 2012). Although these compounds are present in significantly smaller
quantities than industrial byproducts or pharmaceuticals, they are potent endocrine
disrupting compounds because of their strong affinity for both a and P estrogen receptors
(ERs) in vertebrate species (Thorpe et al., 2003). More specifically, EE2 has substantial
estrogen disrupting potential because of its identical structure to endogenous steroid
hormones in vertebrates. When this exogenous steroid is present at just ng/L
concentrations and able to bind to ERs, biological effects such as behavioral changes,
gonadal intersex, increased plasma vitellogenin levels, and decreased sperm abundance
occur in male fish (Vajda and Norris, 2011). However, the binding of exogenous steroids
to a and P ERs throughout the body are not the only sites of reproductive disruption.
3


Their ability to modulate the hypothalamic-pituitary-gonadal (HPG) axis by negative
feedback on the hypothalamus reduces follicle stimulating hormone (FSH) and
luteinizing hormone (LH) release from the pituitary, which ultimately decreases
gametogenesis and sex hormone production (Tmdeau et al., 1997). The strong affinity
these exogenous steroids have for ERs induces widespread disruption of endocrine and
reproductive processes, ultimately causing reduced fecundity and decreased populations
in aquatic vertebrates (Ankley et al.2008).
Industrial Waste Compounds
In addition to estrogenic steroidal compounds found in effluent watersindustrial
waste compounds have been present in large quantities for decades, and also affect
reproductive and endocrine systems. The most prevalent compounds include the organic
metal complexing agent EDTAthe organic nonionic surfactants nonylphenol
ethycarboxylates (NPEs), and the organic plastic byproduct bisphenol-A (BPA). EDTA
in concentrations similar to that found in effluent water is cytotoxic and has been found
to cause reproductive and developmental effects in zebrafish (Lanigan and Yamarik,
2002). Likewisein a wide range of vertebrate speciesNPEs are known to disrupt
reproductive and developmental pathways, because of their small affinity for both a and
P ERs. For example, in a study by Sumpter and Jobling, male juvenile fish exposed to
NPEs produced vitellogenin, a hepatic egg yolk precursor protein that is normally
produced by mature female fish in response to estrogen (2005). Because the immature
fish produced vitellogenin, NPEs were found to have, albeit weak, estrogenic effects in
vivo, and although the affinities of NPEs for estrogen receptors are weak compared to
endogenous estrogens, the high concentrations of these compounds in effluent water
4


increases their overall estrogenicity and potential for reproductive disruption (Sumpter
and Jobling, 2005). BPA, another industrial byproduct with estrogenic effects, has a small
affinity for both a and P ERs throughout tissue in fish and mammalian species (Hatef et
al.2012). In a study by Hatef et al.male goldfish exposed to environmentally relevant
concentrations of BPA (0.2 and 20 |ig/L) had increased levels of estrogen receptor
mRNA, brain and testis specific aromatase, and vitellogenin (2012). In addition, sperm
abundance and motility were reducedfurther proving its ability to disrupt reproduction
and reduce fitness of an animal (Hatef et al., 2012). From observations such as these, it is
clear that adverse effects of industrial waste products and their ability to affect a wide
range of speciesimplicates them as agents of biological disruption. Not only is their
ability to disrupt cellular processes at many sites and through multiple mechanisms
concerning, but their presence in such large quantities worldwide renders them as major
contributors to the negative effects anthropogenic compounds are having on aquatic and
mammalian species (Paterson and Metacalfe, 2008).
Pharmaceutical Compounds
In addition to steroids and industrial waste products, other compounds that are
found in WWTP effluent worldwide and adversely affect aquatic and mammalian species
are pharmaceuticals (Corcoran et al., 2010). Although many pharmaceuticals act on
evolutionarily conserved pathways disrupting fitness and life histories in diverse
organismsneuroactive drugs such as selective serotonin reuptake inhibitors (SSRIs) and
monoamine oxidase inhibitors (MAOIs) are most concerning. This is because their
mechanism of action is only known at the cellular level, but how they affect neural
pathways overall is still unknown (Corcoran et al., 2010). In addition, they are engineered
5


to have potent and long lasting effects on neural tissue, and because anatomy is
conserved in neuroendocrine systems such as the serotonergic system across vertebrates,
they have afrimties for binding at evolutionarily conserved target sites in a wide range of
non-target species (Corcoran et al., 2010). The effects of psychopharmaceuticals vary
between species as these drugs act at multiple sitesand can disrupt physiological
processes through multiple mechanisms. For example, fluoxetine, a widely prescribed
SSRIaffects the hypothalamus of fishand was shown to greatly disrupt feeding and
metabolism in male goldfish (Mennigen et al.2010). In addition, disruption of
aggressive and sexual behavior, and also modulation of the HPG axis has been reported
from fish exposure to psychopharmaceuticals (Kreke and Dietrich, 2008). In a study by
Kreke and Dietrich, fish exposed to SSRIs had increased levels of serotonin, which
stimulated release of gonadotropins from the hypothalamus, and therefore stimulated
release of FSH and LH from the pituitary (2008). Meanwhile fish exposed to
pharmaceuticals that increased presence of dopamine had a decrease in the release of
gonadotropins, and therefore a depression in all aspects of the HPG axis (Kreke and
Dietrich2008). SSRIs and other neuroactive drugs not only affect the central nervous
system, they also directly affect organs such as the gonads. In a study by Cerda et al.,
female killifish {Fundulus heteroclitus) exposed to 0.05 |iM of 5-HT hydrochloride, an
SSRI, demonstrated a significant reduction in fecundity from increased serotonin levels
blocking meiotic maturation of the oocytes (1998). From studies involving
pharmaceutical exposure and observations of biological disruptionit is clear that
neuroactive compounds pose a complex problem because they are present in effluent
waters worldwide, and are chronically acting on native species with known and
6


potentially unknown adverse effects. Therefore, pharmaceuticals must be further studied
to understand how chronic exposure affects species reproduction and overall survival.
Boulder Creek
Stream Background
Locally, the impact of treated effluent water containing compounds such as
estrogens, industrial waste, and pharmaceuticals was examined using a fish model in a
long-term study at the City of BoulderCO WWTP. (Barber et al.2012, Vajda et al.
20062011).Boulder Creek receives its pristine water source from snow melt at the
continental divide, and flows through the mountain foothills relatively unmodified
besides minor nonpoint source pollution from human activity such as fishing and
swimming (City of Boulder, 2012). The native stream water flows until it reaches the 75th
St. WWTP, where 30-80% of the stream flow is then composed of effluent water
depending on the season and amount of snowmelt (Barber et al.2012). The change of
season and subsequent variations in snowmelt create substantial fluctuations in stream
flow, where low flow occurs during fall and winter months from October to March, and
high flow during spring and summer months from April to July. The unpolluted
snowmelt source, relatively undisturbed flow through the city of Boulder, single effluent
source from the WWTPand the fluctuations in water flow leading to a dynamic effluent
load, has provided an ideal site for examining effects of anthropogenic effluent
compounds on organisms.
7


Effluent Source and Contaminants
Wastewater from Boulder and Gunbarrel urban populations enters the City of
Boulder WWTP where water is filtered from debris, compounds are absorbed or
biochemically converted and removed, and pathogens are destroyed by purification and
disinfection processes before the effluent water is released into Boulder Creek (City of
Boulder, 2012). Although the WWTP has undergone massive upgrades to improve
removal efficiency and purification methods, many compounds have still not been
completely removed and have been released into the creek. From over a decade of water
chemistry analyses of the Boulder Creek, effluent water has included hundreds of
compounds, including natural and synthetic steroids, industrial wastes, and
pharmaceuticals (Barber et al.2012).
It is clear from this Boulder Creek study that the anthropogenic compounds have
varied over time, reflecting a change in demographics, chemical usage, and WWTP
treatment processes (USGS2012). The most notable changes took place between 2006
and 2007when a sharp decline of NPEs occurred due to the establishment of
governmental policies limiting the commercial use of these compounds (City of Boulder,
2012). In addition, a fifty million dollar upgrade of the WWTP in 2007 occurred that
implemented the activated sludge process, which significantly improved the removal of
steroidal estrogen compoundsincluding the potent EE2 (Figure 1.1) (USGS2012).
Although these changes improved effluent water chemistry immensely, other compounds
that have been found in effluent waters for decades are still being found todayincluding
compounds that are known to affect vertebrate species such as pharmaceuticals, EDTA,
and BPA (Barber et al., 2012).
8


60 -i
Effluent Estrogenicity (EEq)
2005
2006 2008
Year
2011
Figure LI Boulder Creek EEq.
Estrogenicity in effluent waters from 2005, 2006, 2008, and 2011 from estrogen
compounds such as El,E2, E3, and EE2, shows that significant reductions occurred with
a major WWTP upgrade in 2007.
Fish Exposure in Boulder Creek
The biological effects of the City of Boulder WWTP effluent were evaluated in
2005, 2006, 2008, and 2011 with controlled flow-through fish exposure experiments by
pumping water just upstream (reference) and just downstream (effluent) from the Boulder
Creek 75th WWTP into a mobile lab housing several tanks of fathead minnow
{Pimephales promelas) purchased from Aquatic Biosystems in Fort Collins, Colorado.
Fathead minnow were chosen as an animal model for several reasons, most notably
because they have short generation times, they are large enough for adequate tissue
samples, their geographical distributions are broad, they swim throughout the entire water
column, encountering all contaminants regardless of contaminant partitioning behavior,
their steroid-dependent secondary sex characteristics are distinct and easy to measure,
and because there are over forty years of basic biology and established parameters in their
neuroendocrine and reproductive processes (Ankley et al.2001;Vajda et al.2011).
o o o o o
5 4 3 2 1
(Tooc)bww


+>-


Figure 1.2 Fathead minnow (Pimephales promelas).
The fathead minnow was used ror the effluent exposure experiments. Secondary
sex characteristics were measured including the dorsal fatpaa and nuptial tubercles on the
snout.onads were removed ror later histology and immunohistocnemical analysis.
In the mobile lab, reproductively stimulated male fathead minnow were randomly
selected and exposed for 28 days in effluent dilutions of 100% effluent, 50% effluent, or
0% effluent, with controlled temperature, photoperiod, diet, aeration, and flow. After this
28-day period, fish were sacrificed, sexed, measured, and weighed. The number of
nuptial tubercles on the snout of the fish and the prominence of the dorsal fat pad were
then measured, and blood plasma was collected from the caudal vein for plasma
vitellogenin measurement (Vajda et al.2004). Organs including the gonads were then
removed, weighed, and preserved in neutral buffered formalin (10% NBF) (Vajda et al.,
2004). Gonads were later embedded, sectioned, and stained with hematoxylin and eosin
for analysis of morphology.
10


hJ
g)
Year
REF
50%Eff
EFF
Figure 1.3 Plasma Vitellogenin in Reference and Effluent Exposed Fish.
Plasma vitellogenin was measured after fish were exposed for 28 days in either
Boulder Creeks effluent or reference waters. This egg yolk precursor protein is a
biomarker for estrogen exposure.
Nuptial Tubercles (day 28)
20 -i
2005 2006 2008 2011
Year
Figure 1.4 Nuptial Tubercles in Effluent and Reference Exposed Fish.
Secondary sex characteristics including nuptial tubercles were measured on fish
exposed to either the reference or effluent waters in 2005, 2006, 2008, and 2011.Changes
in secondary sex characteristics are indicative of endocrine disruption.
11


1.4 i
Gonadosomatic Index (day 28)
1.2 -
REF
EFF
50/50
2005
2006
2008
2011
Year
Figure 1.5 Gonadosomatic Index of Reference and Effluent Exposed Fish.
The ratio of the gonad weight to the fish weight was compared as a
gonadosomatic index in fish exposed to either reference or effluent waters in 2005, 2006
2008, and 2011.
Biological Findings from Fish Exposures
In 2005, although fish exposed to effluent water had no increased mortality,
biological observations included diminished primary and secondary sexual characteristics
evident by less abundant and prominent nuptial tubercles (figure 1.4), less prominent
dorsal fat pads, reduced gonadosomatic index (GSI) (figure 1.5)decreased sperm
abundance, and the complete absence of sperm in several male fathead minnow (Vajda et
al., 2011). In addition, plasma vitellogenin was greatly elevated in male fish exposed to
any amount of effluent in this year (figure 1.3) (Vajda et al.2006). From the biological
findings of the effluent exposed fish, it was concluded that complex mixtures in effluent
water downstream from the WWTP was significantly disrupting reproductive processes.
In 2006, fish from the 28 day exposure maintained in 100% effluent showed
significant disruption of secondary sexual characteristics with less abundant and
prominent nuptial tubercles (figure 1.4) and less prominent dorsal fat pads. Effluent
12


exposed fish also had significantly increased plasma vitellogenin (figure L3). However,
there was no significant effect at the gonad, which was evident by the GSI and sperm
abundance being similar to reference exposed fish (figure 1.5) (Vajda et al.2006). In
2006governmental policies limited the commercial use of the extremely potent
compounds NPEs, and a major WWTP upgrade occurred during this time, which
significantly reduced estrogens. Therefore, it was expected that biological observations of
the effluent exposed fish would be similar to reference exposed fish (Vajda et al., 2011).
However, from the effluent exposed fish measurements and observations, it was found
that although effluent water in 2006 did not contain compounds estrogenic enough to
induce changes in gross morphology of the gonad, it was estrogenic enough to induce
plasma vitellogenin and affect secondary sex characteristics in male fish (figures LI and
1.3).
From observations in 2008 and 2011effluent exposed fish had statistically
similar observations to the reference fish with sex ratios, fat pad prominence, nuptial
tubercle number and prominence, sperm abundance, and GSI (figures 1.3, L4, L5) (Barber
et al., 2012). Therefore, the biological observations of these fish exposed to effluent
water reflected the significant shifts in water chemistry from reduced estrogens and NPEs
after 2007 (figure 1.1).
Testing for Discrete Changes
Because the most potent estrogenic compounds, EE2, E2 and NPE, greatly
declined in effluent water after 2007 and remain scarce todayit could be assumed that
disruption is no longer occurring. Howeverhundreds of anthropogenic compounds
including industrial waste byproducts and pharmaceuticals, have been found for decades,
13


and are still present today (USGS, 2012). Because these toxicants remain present in
effluent waterand because either very specific measurements of vitellogenin or coarse-
scale biological measurements such as fish length, gonad size, and gonad histology were
only considered up until this pointit cannot be assumed that no reproductive disruption
is occurring. Effects of toxicants at the molecular, cellular, and tissue level may not
always be obvious at higher levels of organization and would not be reflected in
measurements such as gonad weight or histology. In addition, exposures of 28 days might
not affect physiological processes such as spermatogenesis, which require over 4 months
for completion (Alves et al.2012)and measurements such as sperm abundancegonad
weight, GSI, and histology could remain normal if fish were not exposed for most of this
time. Therefore, because evidence of disruption is subtleand complex mixtures in
effluent water have diverse mechanisms of action that disrupt physiological processes at
multiple sites (figure L6), biomarkers must be general enough to determine end points of
disruption from multiple sites and via multiple mechanismsbut specific enough to
examine subtle changes at the cellular level. Ultimately, these tests must be accurate in
answering the question of whether or not effluent contaminants are adversely affecting
reproduction, and whether broad ranges of species that are exposed to these compounds
are being affected as well.
14


Figure 1.6 Disruption of the Hypothalamus-Pituitary-Gonadal Axis.
This generic figure summarizes potential disruption of the HPG axis in most
vertebrate species from industrial, pharmaceutical, and chemical toxicants found in
effluent water.
15


End points of Disruption
To evaluate whether effluent toxicants are affecting reproductive processes, the
testes of male fathead minnow from the exposure experiments were examined. These
organs were analyzed because they are the ends points of the reproductive pathway, and
also because they are the responsible for production of gametes, and ultimately fertility of
the fish. More specifically, the gonads are the most downstream organ in the
hypothalamic-pituitary-gonadal axis (HPG axis), which is responsible for initiating and
maintaining spermatogenesis and steroidogenesis, and which also can be affected by
many different effluent compounds through multiple sites and via multiple mechanisms
(figure 1.6).
To analyze discrete disruption in the gonads from exposure to environmental
toxicants in effluent watercellular changes in the testes were measured. More
specifically, cellular proliferation and programmed cell death, or apoptosis, were
analyzed, both of which are normal and essential physiological processes during
spermatogenesis (Nagahama, 1994).
Spermatogenesis
Spermatogenesis in vertebrates is a process in which haploid motile cells are
produced from diploid germ cells and mature to form viable gametes within the
seminiferous tubules of the testis (Nagahama, 1994). During development and into
adulthood, the process of spermatogenesis relies on several factors including hormones
and other paracrine factors. These signals allow for sperm production to be increased and
cells to proliferate at sexual maturity and during reproductive seasons, and also for cells
to be decreased and undergo apoptosis during a reproductive hiatus. It is the dynamic
16


balance between cellular proliferation and apoptosis in the gonads that maintains
homeostasis and allows for spermatogenesis to occur (Nagahama, 1994). Therefore,
proliferation and death of cells in the gonad are necessary, and the dynamic balance
between both of these processes is required for reproductive success of male individuals.
Proliferation and Spermatogenesis
Proliferation in testicular tissue constantly occurs as germ cells undergo meiosis
to produce spermatogonia. These cells mature in the epithelium to become
spermatocytes, and travel toward the lumen as they mature into spermatids, and finally,
spermatozoa. This process is dependent on the release of gonadotropin releasing hormone
(GnRH) from the hypothalamus, which acts on the adenohypophysis where it binds to
gonadotrophs and stimulates the release of FSH and LH (Nagahama, 1994). Once
released, LH and FSH travel through the blood to the testicular tissue, where LH
stimulates leydig cells to produce testosterone, and FSH stimulates sertoli cells to initiate
meiosis of spermatogonia. It is the synergistic effects of FSH and LH that promote cell
division and proliferation of spermatogonia in the gonad (Nagahama, 1994). These
hormones also act to negatively feedback on both the adenohypophysis and the
hypothalamus.
PCNA
Effluent compounds such as EE2, NPE, BPA, and pharmaceuticals can directly or
indirectly interfere with regulation of the HPG axis through ERs or interaction with
diverse neurotransmitter systems (figure 1.6). For example, neuroactive pharmaceuticals
that increase dopamine in the brain inhibit the GnRH releasing neurons in the
hypothalamus, and therefore cause decreased LH, FSH, and ultimately a decrease in
17


cellular proliferation (Corcoran et al.2010). In addition, exogenous estrogens such as
EE2 also inhibit GnRH neurons in the brain. Estrogens play an essential role in regulating
the HPG axis in males, because testosterone is aromatized to estrogens locally in the
brain, which then negatively feedback on the GnRH neurons (Jobling et al., 2006). From
what is understood about estrogen mimicking compounds that are found in effluent
waters, ERs are activated, which leads to suppression of the HPG axis, and as a
consequenceoverall decreased reproduction in male animals (Jobling et al.2006).
Apoptosis and Spermatogenesis
Programmed cell death, or apoptosis, is a normal and essential process in all
living tissue including the gonads, where it is fundamental for maintaining the correct
ratio between sertoli cells and sperm cells (Andreu-Vieyra C. et. al, 2005). Initiation of
apoptosis can occur via the intrinsic or extrinsic pathways in instances such as DNA
damage, incorrect replication of DNA, phosphokinase C inhibition, or by the Fas ligand
or tumor necrosis factor a (TNF a) (Andreu-Vieyra C. et. al, 2005). It is the activation of
the intrinsic or extrinsic pathway, or even a combination of both that can influence cell
death, and if chemical or non-chemical stressors occur, these pathways can be activated
to initiate apoptosis.
TUNEL
The effects of compounds such as EE2, NPE, BP A, and pharmaceuticals on the
brain and pituitary may not only influence proliferation, but apoptosis as well. Increasing
dopamine in the brain and depression of the HPlt axis can lead to less FSH and LH, and
therefore less stimulation of gonadal cells (Trudeau et al.2005). Gonadotropin
18


withdrawal not only causes less proliferation, but also decreases the factors maintaining
cell function and cell survival (McClusky2012). Thereforemany effluent compounds
that cause inhibition of hormone release from the hypothalamus and/or pituitary may
result in increased apoptosis.
PCNA and TUNEL
To test for discrete changes from toxicant exposure, two assays were used to
identify cellular proliferation and apoptosis in the testes of fathead minnow. Terminal
deoxynucleotidyl transferase biotin-dUTP nick end labeling, or TUNEL, is an assay that
allows for the location and amount of apoptosis to be measured in tissue. This assay binds
to the 3 hydroxyl ends of DNA fragments that are created only during apoptosis
(DAndrea et al.2010). Cells undergoing necrosis from other causes will not stain
positive for the TUNEL assay, because DNA is sheared haphazardly during necrosis, and
is not fragmented in the same manner as during apoptosis (DAndrea et al.2010).
The TUNEL assay was chosen because it is sensitive and reliable, and has been
applied to a wide range of vertebrate species for analyzing cellular changes from toxicant
exposure. For example, abundance of TUNEL in rat testes was measured after exposure
to the toxicants phthalates, methylmercury, and nonylphenols (Moffit et al., 2007;
Fujimura et al.2012; Han X. et al.2004)and was also used to measure apoptosis in the
gills of brown trout in various exposure experiments (Rojo and Gonzalez, 1998). TUNEL
has also been used in numerous experiments measuring sperm apoptosis in human men
(Delbes et al., 2010). In addition to being a sensitive, reliable, and widely used assay for
measuring biomarkers in many different vertebrates, TUNEL stains cells undergoing a
process that cannot be reversedand therefore shows cellular disruption from many hours
19


or even days of toxicant exposure. Unlike measuring hormones such as FSH or LH,
which can recover hours after a toxicant exposureTUNEL shows apoptosisa process
from which cells cannot recover (DAndrea et al.2010).
Proliferating cell nuclear antigen, or PCNA, is an assay that binds to a
homotrimeric nuclear protein associated with cells undergoing mitosis or meiosis, and
therefore allows for the location and abundance of proliferation to be measured in tissue.
Like TUNEL, PCNA is a sensitive, widely used, and reliable assay that has been used in
a wide range of vertebrates for analyzing changes in cellular proliferation from exposure
to toxicants. For example, PCNA has been used to measure proliferation in testes of rats
exposed to benzenescyclophosphamidesand ethanol (Chandra et al.1997, DAndrea et
al., 2008). In addition, PCNA has been used on zebrafish tissue exposed to acrylamide,
and has been used in numerous studies measuring proliferation of cells after exposure to
carcinogens (Parng et al.2007; Girod et al.1994). Similar to TUNELPCNA binds to
cells that are undergoing a process that cannot reverse in a short period of time. Unlike
measuring hormones that can fluctuate over the course of a day, or even within the hour,
cells undergoing mitosis and meiosis will remain in that state for longer periods of time
(Bennet1971). Therefore, PCNA is a highly effective assay for not only measuring
proliferating cells in a wide range of vertebrate species, but also for measuring
cumulative effects after days or weeks of exposure to effluent toxicants.
Expected Outcomes
From increased amounts of estrogens and NPEs in Boulder Creek effluent in
years 2005 and 2006, it was expected that TUNEL be more abundant in testes of effluent
exposed fish. This was hypothesized because these exogenous compounds bind to ERs,
20


and can cause the downregulation of the HPG axis and ultimately, gonadotropin
withdrawal. In addition, from previous measurements of decreased primary and
secondary sex characteristics, it was predicted that apoptosis was occurring more
frequently at the cellular level. In 2008, however, TUNEL was expected to be similar
between effluent and reference exposed fish because of changes in Boulder Creek
effluent water chemistry, and also from previous observations revealing no significant
changes in fish morphology or gonad histology.
PCNA was predicted to be less abundant in 2005 and 2006 effluent exposed fish
from the depression of the HPG axis, which could have occurred from many effluent
contaminants found in Boulder Creek effluent. In 2008, howeverPCNA was expected to
not be significantly different between reference and effluent exposed fish because certain
EDCs were removed from effluent waterand also because previous observations
revealed no significant changes in morphology or gonad histology in effluent exposed
fish.
21


CHAPTER II
METHODS
On-site Fish Laboratory and Experimental Design
Reproductively stimulated 12 month adult male fathead minnow were exposed to
either 100% effluent or 100% reference in 2005, 2006and 2008, using a flow-through
system mobile laboratory located on-site at the city of Boulder WWTP (Vajda et al.,
2011).For 28 days, controlled variables included temperature, lighting, brine shrimp diet,
aerationflowand effluent concentration dilution (Vajda et al.2011).After 28 daysfish
were sacrificed and measurements were recorded, including fish length, fish weight,
nuptial tubercle number and prominence, dorsal fat prominence, hematocrit, and plasma
vitellogenin (Vajda et al.2011).Gonads and other organs were dissected and weighed
and the gonadosomatic index was calculated (Vajda et al., 2011). All tissue samples were
preserved in 10% neutral-buffered formalin for later histological preparations (Vajda et
al., 2011).
Double Labeling Procedure
Gonad tissue from the 20052006and 2008 experiments was embedded in
paraffin wax and sectioned at 5 |im ribbons that were then mounted onto glass slides.
Slides were then dewaxed and rehydrated in xylene (3x5 minutes each),100% ethanol,
95% ethanol, and 70% ethanol (all 2x3 minutes each). Slides were then immersed in
deionized water for 3 minutes. For antigen retrieval, slides were microwaved for 5
minutes in 0.1 M citrate buffer then immediately placed in 0.1 M Tris-HCl solution for
30 minutes. To quench endogenous peroxidase activity, slides were immersed in a 0.3%
H2Og methanol solution for 15 minutes. The TUNEL reaction mixture was prepared
22


following the Roche In Situ Cell Death Detection Kit, AP, and 100 |iL of TUNEL
reaction solution was added to each slide. After slides were coverslipped, they were
incubated for 60 minutes in a 37 C humidified dark chamber, then rinsed 3x5 minutes
in 0.1 M PBS. The Roche Fast Red Solution was prepared by dissolving one fast red
tabled into 5 mL of substrate buffer and vortexing until tablet was dissolved, and after
slides were blotted dry, 50-100 |iL was added to each slide until the tissue was covered in
solution. Slides were incubated for 10 minutes at 15-25 C in the dark, and after this first
incubation, excess liquid was removed and fast red substrate solution was added again.
Slides were incubated for another 10 minutes in a dark chamber. Slides were then rinsed
in 0.1 M PBS 3x5 minutes.
The primary PCNA antibody was prepared in a 1:200 dilution in 0.1 M PBS.
After slides were blotted dry the antibody was added, and each slide was coverslipped
and left to incubate over night at 4 C in a humidified chamber. After this incubation,
slides were then washed in 0.1 M PBS 3x2 minutes, and the secondary PCNA antibody
was prepared following the Vectastain ABC Kit protocol. After the secondary antibody
was applied to each sample, the slides were coverslipped and incubated for 60 minutes in
a humidified chamber. After this incubation, slides were washed in 0.1 M PBS 3x2
minutes, the ABC reagent from the Vectastain ABC Kit was applied to each sample, and
slides were coverslipped. The slides were placed in a humidified chamber and incubated
with the ABC reagent for 30 minutes. Slides were then rinsed in PBS 3 x 2 minutes. The
DAB, or peroxidase substrate solution, was prepared from the Vector Laboratories
protocol, and 500 |L was added to each slide for approximately 5 minutes. Slides were
tapped to remove excess DAB, and rinsed in tap water. Slides were then placed in
23


hematoxylin for approximately 45 seconds, and rinsed in tap water until the solution ran
clear. After excess liquid was removed, copious amounts of simpomount was added to
each slide, and then suctioned off using a transfer pipette to remove air bubbles.
Fish Sample
Name &
Number
ribbon 1 ribbon 2
V
gonad
tissue
/\
Figure 11.1 Example of Sectioned Slide Used For Analysis.
Each slide contained two 5 |im thick wax ribbons, each containing tissue sections
from different areas of the testis.
24



Figure 11.2 Blind Analysis of Slides.
To analyze the slides blindly without knowing the treatment group of each fish,
the labels on each slide were covered and randomly assigned numbers.
Slide Analysis
Testes from reference and effluent exposed fish in 2005, 2006and 2008, were
embedded in wax, and sectioned into 5 |im thick wax ribbons. Specifically, the wax
ribbons were mounted onto glass slides in a random order, so that the two ribbons on
each slide contained different sections of the gonad (figure 11.1).Therefore, every slide
contained multiple sections of the gonad from each fish, and was more representative of
the gonad as a whole, not just one particular area. For double labeling, five fish out of
approximately thirty fish samples were selected at random from each treatment group
(reference and effluent) in years 20052006and 2008giving a total of thirty fish
samples for double labeling with PCNA and TUNEL. The labels on each slide were then
25


covered with tape, and randomly assigned a number so that analysis was performed
blindly (figure 11.2). From these slides, five tubules were selected from each of the two
wax ribbons, giving a total of ten tubules for each slide. Tubules were selected at random,
however, there was a bias for round tubules because they represent cross sections of the
gonad, unlike oval or irregular shaped tubules, which represent oblique sections of the
gonad. Tubules that were not cross sections often contained varying densities of sperm in
the lumen. Therefore this bias was applied when selecting tubules to avoid miscounting
cell density, and misrepresenting amount of TUNEL or PCNA positive cells.
Using Image J software, the tubule area, lumen area, and sperm area (figure 11.3)
were measured for each tubule. The lumen area was subtracted from the tubule area to
give the area of the epithelium, and sperm area was measured carefully, with areas
without sperm subtracted from the total sperm area (figure 11.3). PCNA and TUNEL
positive cells were counted in the epithelium, and were divided by the epithelium area to
give the number of cells per |im2 (figure 11.4). TUNEL positive cells were counted
individually in the sperm area and divided by the total sperm area. PCNA positive cells
were counted by randomly selecting multiple sample areas within the sperm and counting
the cells within these areas, because most of the sperm cells stained positive for
proliferation, and this expedited the counting process (figure 11.4). The sample areas were
added together then multiplied by the total sperm area to give total number of PCNA
positive cells per |im2 of sperm area.
26


Figure 11.3 Tubule Morphometries.
The tubule area was outlined (yellow) following the basement membrane of the
epithelium, and the lumen area was outlined (red) following the cells bordering the
lumen. The lumen area (red) was subtracted from the tubule area (yellow) to find the area
of the epithelium. Last, the sperm area was measured (green) outlining the sperm area,
and areas inside without sperm cells were measured and subtracted from the total sperm
area.
27


Figure 11.4 TUNEL and PCNA Positive Cell Count.
TUNEL positive cells stained red in the nucleus and PCNA cells stained dark
brown in the nucleus. TUNEL and PCNA positive cells were counted individually in the
epithelium, and were divided by the area of the epithelium (figure 11.3). The number of
TUNEL positive cells were counted individually in the sperm area, and divided by area
of the sperm. Samples sections of sperm were circled at random, and the number of
PCNA positive cells was counted inside. The samples were added together and multiplied
by the sperm area to give total PCNA positive cells per area of sperm. In this figure,
there are several TUNEL positive cells in the epithelium, and hundreds of TUNEL
positive cells in the sperm, which is evident from the dark red staining. Likewise, in the
epithelium, there are dozens of PCNA positive cells in the epithelium, and many
hundreds of PCNA positive cells in the sperm area, evident by the dark brown staining in
the nucleus.
Statistical Analysis
Using the statistical program R, data were analyzed using a two way analysis of
variance and Tukey post hoc test. To correct for homoscedasticity and reduce probability
of witnessing one or more type I errors, the Holmes-Bonferroni correction was also
28


applied. Statistical differences between PCNA positive cells in the epithelium and lumen,
and TUNEL positive cells in the epithelium and lumen were found between effluent and
reference exposed fish in years 2005, 2006, and 2008, with a significance level of 0.05.
29


CHAPTER III
RESULTS
Testicular Morphometries of 2005, 2006, and 2008 Tissue
Sperm, Lumen, and Tubule Areas
14000
12000
10000
8000
6000
4000
2000
0
Sperm
Tubule
Epithelium
Figure 111.1 Averages of Sperm, Lumen, and Epithelium Areas.
The average sperm area, tubule area, and epithelium area of all fish samples in
each treatment group each year was calculated.
30


PCNA and TUNEL Immunoreactivity
Figure 111.3 Epithelial and Luminal PCNA Reactive Cells in 2005
Epithelial (p<0.05) and sperm cells (p<0.01) positive for PCNA was significantly
higher in effluent exposed fish.
2005 TUNEL
0.006 -
0.005
N a 0.004
T, 0.003
o 0.002
0.001
o -
Reference
Effluent
Epith TUN
Sperm TUN
Figure 111.2 Epithelial and Luminal TUNEL Reactive Cells in 2005
Abundance of epithelial TUNEL cells was significantly higher in effluent exposed
fish (p<0.01). Likewise, abundance of sperm TUNEL cells was higher in effluent fish
(p<0.001) compared to reference exposed fish.
2005 PCNA
9 6 3
o o o
o
Nluri/SIP0
31


0.0048
0.004
0.0032
N
I 0.0024
0.0016
0.0008
0
2006 TUNEL

Epith TUNEL Sperm TUNEL
2006 Reference
2006 Effluent
Figure 111.4 Epithelial and Luminal TUNEL Reactive Cells in 2006
Epithelial cells positive for TUNEL in effluent exposed fish were statistically
higher than reference exposed fish (p<0.001). Likewise, sperm cells positive for TUNEL
were higher in effluent fish compared to reference (p<0.01).
2006 PCNA
0.18
0.15
0.12
a
0.09
1
0.06
0.03
0

k
2006 Effluent
Epith PCNA Sperm PCNA
Figure 111.5 Epithelial and Luminal PCNA Reactive Cells in 2006
Epithelial cells positive for PCNA in effluent exposed fish were not statistically
different between reference and effluent exposed fish, however, sperm PCNA positive
cells were higher in reference compared to effluent (p<0.001).
32


2008 TUNEL
0.004
0.003
0.002
2008 Reference
2008 uent
Epith TUNEL Sperm TUNEL
Figure 111.6 Epithelial and Luminal TUNEL Reactive Cells in 2008
Both epithelial and sperm cells positive for TUNEL in effluent exposed fish were
not statistically different than reference exposed fish.
2008 PCNA
1
0.15
0.12
0.09
0.06
0.03
0
2008 Reference
J 2008 Effluent
Epith PCNA Sperm PCNA
Figure 111.7 Epithelial and Luminal PCNA Reactive Cells in 2008
Epithelial cells positive for PCNA in effluent exposed fish were not statistically
different than reference exposed fish, however, PCNA positive cells in sperm of effluent
fish were statistically higher than reference fish (p<0.05).
33


Table 111.1 Previous Measurements, Water Chemistry, and Biomarkers
GSI Vitell. Nup.Tub. Sperm Estrog. Epi TUN Sperm TUN Epith PCNA Sperm PCNA
2005 +++ +++ +++ +++ +++ Eff Eff Eff Eff
2006 ++ ++ ++ Eff Eff Ref
2008 Eff
In 2005, GSI was significantly lower, vitellogenin was increased, nuptial
tubercles were less pronounced, and sperm abundance was lower in effluent exposed fish,
all reflecting an increased amount of estrogens in the effluent water (Vajda et al.2009;
Barber et al., 2012). With discrete testing using PCNA and TUNEL, it was found that
effluent exposed fish had higher amounts of apoptosis and higher amounts of
proliferation than reference exposed fish (p<0.05).
In 2006, GSI was not significantly between the two treatment groups, vitellogenin
was still abnormally higher in effluent fish, nuptial tubercles were less pronounced but
not as diminished as the previous yearwhich all reflected a decrease in effluent
estrogens (Vajda et al.2009; Barber et al.2012). Similar to 2005, effluent exposed fish
showed higher amounts of TUNEL positive cells, but PCNA was more abundant in the
sperm of reference-exposed fish (p<0.05).
In 2008, GSI, vitellogenin, nuptial tubercles, and sperm abundance were all
similar between the effluent and reference exposed fish, reflecting the removal of
estrogens and NPEs from the effluent water (Vajda et al.2009; Barber et al.2012). This
resulted in TUNEL being similar in both reference and effluent exposed fish, but PCNA
being more abundant in sperm of effluent exposed fish compared to reference fish
(p<0.05).
34


CHAPTER IV
DISCUSSION
Cellular Proliferation
From the higher load of effluent contaminants in years 2005 and 2006it was
predicted that fish maintained in effluent waters would show less proliferation within
their gonads compared to reference fish. In 2005 it was found that proliferation measured
by amount of PCNA positive cells was significantly higher in the sperm and epithelium
of effluent fish compared to reference fish (p<0.01)(figure 111.3). In 2006however
proliferation was lower in the sperm of effluent fish (p<0.001)and was not different in
the epithelium between the two treatment groups (figure 1115).
In fish exposed to effluent in 2008, when water chemistry revealed a decrease in
estrogens and NPEsand when almost no disruption in either secondary characteristics or
gonad histology occurred, it was expected that no difference would occur between
effluent and reference exposed fish. It was found that the amount of PCNA positive cells
was not significantly different in the epithelium between the two treatment groups in
2008, but more PCNA positive cells were found in the sperm of effluent fish compared to
reference exposed fish (p<0.05) (figure 111.7).
The insignificant difference of PCNA positive cells in the epithelium between the
two treatment groups in 2006 and 2008, and the inconsistent trends of PCNA in the
sperm between all years, suggests that PCNA is not as reliable as some previous studies
have found (Chandra et al., 1997). However, other studies have concluded that PCNA
might not be as sensitive or reliable as TUNEL. For example, D5 Andrea et al.found that
when analyzing rat testes, PCNA might not be as sensitive as TUNEL, especially when
35


using the double labeling technique (2010). Likewise, from findings of this study, it is
suggested that when staining testes of fathead minnow, this assay is not as beneficial as
using TUNEL, because it does not provide more information than what is provided by
secondary sex characteristics, gonad histology, or plasma vitellogenin (Table 1111).
Two possible reasons for PCNA not being consistent are that this assay might be
less sensitive than TUNEL, and human error might have occurred when counting cells.
Due to artifacts in the tubules, counting number of cells was at times difficult, which
most likely introduced human error. In addition, most of the many hundreds or even
thousands of sperm cells stained positive for PCNA, which made counting cells even
more difficult. Perhaps this assay would be more easier to count and thus more
informative in less proliferative tissue such as hepatic or renal tissue. Organs that
normally undergo little mitosis might show more profound differences in cellular
proliferation when the organism is stressed from exposure to environmental toxicants.
TUNEL
Increased concentrations of estrogenic effluent compounds in 2005 and 2006 were
predicted to increase TUNEL reactive cells in the gonad of effluent exposed fish
compared to reference fish. It was found that epithelial and sperm TUNEL positive cells
were significantly higher in 2005 and 2006 effluent exposed fish (p<0.001) compared to
reference fish (figures 111.2 and 111.4). In 2008, when water chemistry revealed a decrease
in estrogens and NPEsand when no disruption in either secondary characteristics or
gonad histology occurred, it was hypothesized that TUNEL would not be significantly
higher in effluent exposed fish compared to reference exposed fish. It was found that
epithelial and sperm TUNEL positive cells were not significantly different between
36


effluent and reference exposed fish in 2008. However, there were more TUNEL positive
sperm cells in effluent exposed fish in 2008, which although is not statistically
significantmay be biologically relevant (figure 111.6).
From the immunohistochemistry results of 2005, 2006, and 2008 tissue, TUNEL
proved to be the more informative and sensitive biomarker. TUNEL in the sperm of
effluent fish was significantly higher than in the sperm of reference fish, which is
consistent with previous findings of low sperm abundance and reduced GSI in 2005
(table 111.1).However, TUNEL also revealed that sperm apoptosis was occurring in
effluent exposed fish in 2006 and 2008. TUNEL revealing apoptosis when other
measurements failed to respond suggests that it is a sensitive and valuable biomarker for
detecting cellular disruption (Table 111.1).This is consistent with previous studies finding
that TUNEL is more sensitive and more reliable then PCNA (DAndrea et al.2010).
Therefore, this is a useful biomarker when analyzing testes of fathead minnow, because it
allows discrete changes from effluent compound exposure to be seen in the gonad when
other measurements show no changes.
The higher abundance of TUNEL and therefore apoptosis in spermatic cells of
effluent exposed fish in 2006 and 2008 indicates that disruption is still occurring from
effluent compounds. The physiological and functional significance of these findings is
that sperm production declines when fish are exposed to the complex mixtures of effluent
waters, but whether or not this decline has significant impacts on individual fitness
depends on a sperm threshold for fertility. This is difficult to estimate in fish, because
even in humans where sperm counts have continuously been monitored in thousands of
men from many different countries, interpretation of how much sperm is required for
37


fertility is in fact difficult to define and controversial (Sharp and Skakkebaek, 2003). This
difficulty is due to factors such as differing subject recruitmentdiffering types of semen
analysis, and even differing seasons (Sharp and Skakkebaek, 2003). These factors are
even more difficult to control in fish, compounding the difficulty in estimating how much
sperm is required for male fish to be fertile. However, the aim of this study was only to
evaluate whether or not effluent compounds were dismpting reproductive pathways to
cause any changes in gonad cells. Moreover, the aim was to find sensitive tests that could
indicate physiological disruption in fish when coarse-scale measurements would
otherwise fail to respond. Further testing is required to correlate the amount of sperm
positive for TUNEL with overall fish fertility and reproductive success.
Use as Biomarkers
Use of Biomarkers in Fathead Minnow Testes
From results of this study, the higher prevalence of TUNEL in the gonads of
effluent exposed fish with no significant decreases in sperm abundance, suggests that this
assay should be used in combination with hematoxylin and easin staining (H & E) for
histological observations. When staining only with H & E, sperm abundance may look
high and similar to that of fish exposed to no toxicants (figure IV.1).This is because H &
E only stains the nucleus and cytoplasm, and will therefore stain even apoptotic cells.
Howeverwhen stained with TUNELone can see that significant disruption is occurring
in the tubule, evident by abundant TUNEL staining (figure IV.1)
38


Figure IV.1 Use of TUNEL in Combination with Hematoxylin and Eosin Staining.
When analyzing gonad tubules with only hematoxylin and eosin through light
microscopy, sperm abundance looks high and similar to that of fish exposed to no
toxicants. However, when stained with TUNEL, it is evident that sperm apoptosis is
abnormally high especially within the sperm cells.
Use of Biomarkers in Vertebrates
PCNA and TUNEL as biomarkers in any vertebrate species can be informative of
discrete cellular changes in tissue. This is because these assays bind to elements that
almost all eukaryotic cells have in common. PCNA binds to a nuclear protein found in
eukaryotic proliferating cells, and TUNEL binds to nicked 35 OH ends on DNA, which
are created in eukaryotic cells as a result of apoptosis (DAndrea et al.2010). However
even though both of these assays can be used in most tissue in a wide range of vertebrate
species, it was found that when analyzing gonad tissue in fathead minnow, TUNEL was a
more dependable, sensitive, and informative biomarker for analyzing cellular changes
from multiple pathways of disruption.
The translational ability of PCNA and TUNEL from a vertebrate model organism
to another vertebrate may depend on many factors, for example, differences in the tissue
39


being examined, the neural or hormonal pathways acting on that tissue, and the
differences in how intracellular and extracellular signals act not only in the animal but on
that particular tissue as well (Hess and Carnes, 2004). In the testes, for example,
proliferation occurs via activation of the HPG axis, which is mostly conserved in
vertebrate species including fish and humans (Trudeau et al.2003). Howeverthere are
some slight differences between the HPG axis in fish and humans that might alter the
translational ability of PCNA from one species to the other (see Biomarkers for Use in
Humans). In addition, apoptosis occurs from a combination of the intrinsic or extrinsic
pathways in eukaryotic cells, but stressors that induce apoptosis in one organism might
not reach a specific threshold to induce apoptosis in another organism. In addition,
differences in organismal-level phenomena such as life-stages, metabolism, uptake of
compounds, etc. can influence events at the cellular level, and can therefore result in
differences in cellular response from one organism to another. For these reasons, the
differences between species are necessary to know and understand when interpreting and
predicting how proliferation and apoptosis in one organism will compare to another.
Use of Biomarkers in Humans
When translating cellular proliferation from fathead minnow to humans, there are
several factors to note. Not only are there anatomical differences, but physiological
differences as well. For example, neurons in the hypothalamus of fathead minnow
directly innervate the pituitary and release GnRH directly onto gonadotrophs, whereas
humans have a portal system that moves GnRH from the hypothalamus to the pituitary
via two capillary networks (Trudeau et al.2003). In additiondistributions of ERs in the
hypothalamus and pituitary differ between fish and humans, as do specific pathways of
40


estrogen feedback regulation (Trudeau et al.2003). But ER distribution is not only
different in the brain, for example, estrogen receptors a (ER a) are located in different
areas of the fish testes compared to the humans (Hess et al.2004). These differences in
ER distribution between humans and fish could be significant because of the potential
differing roles in regulating spermatogenesis, and because exogenous estrogenic
compounds may have different effects if they are binding to receptors in different
locations and in different densities (Hess and Carnes, 2004). However, differences in ER
distribution in the brain do not affect the proportion of negative feedback from estrogens
between fish and mammalsand therefore do not have an impact on pathways leading to
cellular changes at the gonad (Hess and Cames2004). In additionER a has been found
to be irrelevant in maintaining or initiating spermatogenesis in almost every species (Hess
and Carnes, 2004). ER P, on the other hand, is highly implicated in initiating and
maintaining spermatogenesis, and is abundant in both fish and humans. More
specifically, ER P is found in almost every cell type of both human and fish testis such as
leydig, sertoli, germ cells, and peritubular cells (Hess and Carnes, 2004). Therefore, from
the distribution of critical ERs in humans and fish, and from the HPG axis being
conserved in both, effluent contaminants that increased apoptosis in fathead minnow are
predicted to also increase apoptosis in testis of humans.
41


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Full Text

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CELLULAR APOPTOSIS AND PROLIFERATION IN TESTES OF FATHEAD MINNOW EXPOSED TO WASTEWATER TREATMENT PLANT EFFLUENT by Andrea Elizabeth Geddes Hallagin B.A., University of Colorado, 2009 A thesis submitted to the Faculty of the Graduate School of the University of Colorado in partial fulfillment of the requirements for the degree of Master of Science Integrative Biology 2013

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ii This thesis for the Master of Science degree by Andrea Elizabeth Geddes Hallagin has been approved for the Department of Int egrative Biology by Alan Vajda Chair Michael Greene Amanda Charlesworth April 16, 2013

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iii Hallagin, Andrea, Elizabeth Geddes (M.S. Integrative Biology) Cellular Apoptosis and Proliferation in Testes of Fathead Minnow Exposed to Was tewater Treatment Plant Effluent Thesis directed by Assistant Professor Alan Vajda ABSTRACT Environmental toxicants and their effects on both wildlife and human populations have become more c oncern ing as worldwide human populations grow, water sources become scarce, and anthropogenic w aste becomes more prevalent. M unicipal wastewater treatment plants (WWTPs) are designed to conserve w ater by removing nutrients and pathogens from wastewater and releas e efflu ent water back into local streams. Unfortunatel y many wastewater compounds are not completely removed from the effluent and are released into the environment sometimes causing deleterious effects even in minute qu antities on biological life. T esting of anthropogenic compounds and effects on wildlife was conducted at the City of Boulder WWTP in 2005, 2006, 2008, and 2011 Biological assays were performed by maintaining adult male fathead minnow in either referenc e or effluent water for 28 days, and by measuring secondary sex characteristics, plasma vi t ellogenin, and gonad histology. I t was found that significant disruption in gonad weight, sperm abundance, and vitellogenin occurred in effluent exposed fish in 2005, but little or no significant disruption occurred in any later year because of a major WWT P upgrade that improved effluent water chemistry T he aim of this study was to examine biological effects further by test ing for discr et e cellular changes of the same experimental fish from 2005, 2006, and 2008 This approach involved analyzing biomarkers which identified cell ular proliferation via proliferating cell nuclear antigen ( PCNA ) and apoptosis via terminal deoxynucleotidyl transferase

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iv biotin dUTP nick end labeling (TUNEL) It was found that epithelial PCNA was higher in effluent exposed fish in 2 005 only, but was lower in sperm cells of effluent fish in 2006, and higher in sperm of reference fish in 2005 and 2008 (p< 0 .05). E pithelial and sperm TUNEL was greater in effluent exposed fish in 2005 and 2006 (p< 0 .05) but was not significantly different in 2008. P CNA was found t o be inconsistent between years, but TUNEL was consisten t with previous measurements, proved to be a more reliable biomarker and showed that disruption is still occurring in 2008 fish even after the WWTP upgrade F uture use of TU NEL is recommended as it responds to dis crete cellular disruption early and shows disruption when other measurements fail to respond. The form and content of this abstract are approved. I recommend its publication. Approved: Alan Vajda

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v ACKNOWLEDG MENTS I would like to thank members of the Vajda lab including Zia Faizi and Ethan Cabral for all of their help I also want to thank all of my friends and family for all of their support and for putting up with me during this process

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vi TABLE OF CONTENT S CHAPTER I. INTRODUCTION1 Worldwide Water Concerns ... ..1 Endocrine Disruption from Water Contaminants .. ..2 Endocrine Disrupting Compounds ... 3 Natural and Synthetic Steroid .. 3 Industrial Waste Compounds ... 4 Pharmaceutical Compounds .5 Boulder Creek .7 Stream Background 7 Effluent Source and Contaminants ..8 Fish Exposure in Boulder Creek .. 9 Biological Findings from Fish Exposures ... ..12 Testing for Discrete Changes .13 Endpoints of Disruption 16 Spermatogenesis 16 P roliferation and Spermatogenesis 17 PCNA 17 Apoptosis and Spermatogenesis ... .18 TUNEL.. 18 PCNA and TUNEL .. ..19 Expected Outcomes ... 20

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vii II. M ETHODS....22 On Site Laboratory and Experimental Design ... 22 Double Labeling Procedure ... 22 Slide Analysis 25 Statistical Analysis 28 III. RESULTS30 Testicular Morphometrics of 2005, 2006, and 2008 Tissue .. 30 PCNA and TUNEL Immunoreactivity .. 31 IV. DISCUSSION..35 Cellular Proliferation 35 TUNEL .36 Use as Biomarkers 38 Use of Biomarkers in Fathead Minnow Testes .38 Use of Biomarkers in Vertebrates .. 39 Use of Biomarkers in Humans .. 40 R EFERENCES ..42

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viii List of Tables Table III .1 Previous Measurements, Water Chemistry, and Biomarkers 34

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ix L IST OF FIGURES Figure I.1 Boulder Creek EEq .. .. 9 I.2 Fathead Minnow ( Pimephales promelas ) .. .. 10 I.3 Plasma Vitellogenin in Reference and Effluent Exposed Fish .. .. 11 I.4 Nuptial Tubercles in Effluent and Reference Exposed Fish .. .. 11 I.5 Gonadosomatic Index of Reference and Effluent Exposed Fish .. ...12 I.6 Disruption of the Hypothalamus Pituitary Gonadal Axis .15 II.1 Example of Sectioned Slide Used for Analysis .24 II.2 Blind Analysis of Slides 25 II.3 Tubule Morphometrics ...27 II.4 TUNEL and PCNA Positive Cell Count 28 III.1 Averages of Sperm, Lumen and Epithelium Areas ... ...30 III.2 Epithelial and Luminal TUNEL Reactive Cells in 2005 ... ...31 III.3 Epithelial and Luminal PCNA Reactive Cells in 2005 ... ..31 III.4 Epithelial and Luminal TUNEL Reactive Cells in 2006 ... ...32 III.5 Epithelial and Luminal PCNA Reactive Cells in 2006 ... ..32 III.6 Epithelial and Luminal TUNEL Reactive Cells i n 2008 ... ...33 III.7 Epithelial and Luminal PCNA Reactive Cells in 2008 ... ..33 IV.1 Use of TUNEL in Combination with Hematoxylin and Eosin Staining .. 39

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x LIST OF ABBREVIATIONS BPA bisphenol A EDC endocrine disrupting compoun d EDTA e thylenediaminetetraacetic acid EE2 ethinyl estradiol ER estrogen receptor FSH follicle stimulating hormone GnRH gonadotropin releasing hormone HPG hypothalamic pituitary gonadal LH luteinizing hormone NPE nonylphenol ethycarboxylates SSRI selective serotonin reuptake inhibitor WWTP wastewater treatment pla nt

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1 CHAPTER I INTRODUCTION Worldwide Water Concerns We live in a world o f finite resources, yet increasing worldwide consumption of these resources is decreasing our supply at an alarming r ate. One of the most rapidly diminishing essential resource s is fresh water, which is a result from increased consumption over the past century from dramatic increase s in human populations expanding urbanization, and increased agricultural use (Hinrichsen and Tacio, 201 2) Not only is there a worldwide shortage of water, but increased pollution and the presence of increasingly varied e nvironmental toxicants pose additional threats as ant hropogenic waste compounds are contaminating what little fresh water i s available (Hinrichsen and Tacio, 2012) One effort to conserve water and minimize waste is the implementation and continual improvement of m unicipal wastewater treatment plants (WWTPs) which are principally designed to effectively remove nutrients and pathogens from urban wastewater and release the effluent into local streams and rivers. This process occurs by first filtering wastewater from debris such as trash, branches, and rocks and then removing large particulate matter with grit chambers and sett ling tanks ( Lee et al., 2006 ). Smaller compounds are then absorbed or biochemically converted and removed using trickli ng filters and/or activated sludg e. After these processes, advanced treatment of the water occurs where p athogens are destroyed, disinfec tion processes occur, and harmful

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2 nutrients are removed (Lee et al., 2006) After these WWTP processes, the effluent water is then released into a local stream where it provides necessary stream flow for the survival of many ecosystems ( Vajda and Norris, 2 011 ). Although most compounds are removed at the WWTP many others such as steroids, pharmaceuticals, and other inorganic and organic compounds are not completely removed, and are therefore released into the environment with effluent water Most of these compounds are relatively i nnoc u ous but unfortunately, some have been found to have deleterious effects on biological life, even in minute quantities (Vajda and Norris 2011) In addition, effluent compounds are found in complex mixtures and sometimes can act synergistic ally when consumed compounding and multiplying the effects of each other resulting in physiological systems being affected at multiple sites and via multiple mechanisms ( Ankley et al., 2009 ) Endocrine Disruption from Water Contaminants Al though effluent compounds can affect multiple physiological processes in many vertebrate species, a concerning affect these compounds can have is disruption of reproductive pathways because not only is the fitness and health of the individual af fected, bu t reproductive disruption can decrease fecundity of individuals and can ultimately af fect populations of the species (Ankley et al., 2008) In addition, reproductive and endocrine pathways are mostly conserved amongst vertebrate species, so the negative im pact s effluent compounds have on one species can also occur in a wide range of vertebrate populations (Nagahama, 1994) Therefore because reproductive pathways are conserved in vertebrate species, and because disruption of reproductive

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3 pathways by effluen t compounds can have populations level effects, i t is imperative that the impact of effluent compounds on reproduction be studied to understand not only individual health affects, but individual fitness, population growth, and ecosystem impacts as well. E ndocrine Disrupting Compounds Natural and Synthetic Steroids The most notable and concerning endocrine and reproductive disrupting compounds (EDCs) found in effluent water include natural and synthetic steroids, organic and inorganic industrial waste, and pharmaceutical compounds. Steroidal compounds such as the natural estrogens estrone (E1), 17! estradiol (E2), and estriol (E3) and the synthetic estrogen 17 ethynyl estrad iol (EE2) are found in wastewater downstream from populated areas where widespread use and improper disposal of birth control occurs ( Barber et al., 2012 ) Although t hese compounds are present in significantly smaller quantities than industrial byproducts or pharmaceuticals, they are potent endocrine disrupting compounds because of their strong affinity for both and estrogen receptors (ERs) in vertebrate species (T horpe et al., 2003 ) More specifically, EE2 has substantial estrogen disrupting potential because of its identical structure to endogenous steroid hormones in vertebrates When this exogenous steroid is present at just ng/L concentrations and able to bind to ERs, biological effects such as behavioral changes, gonadal intersex, increased plasma vitellogenin levels, and decreased sperm abundance occur in male fish (Vajda and Norris 2011). However, the binding of exogenous steroids to and ERs throughout the body are not the only sites of reproductive disruption.

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4 Their ability to modulate the hypothalamic pituitary gonadal (HPG) axis by negative feed back on the hypothalamus reduc es follicle stimulating hormone (FSH) and luteinizing hormone (LH) release from the pituitary, which ultimately decreases gametogenesis and sex hormone production (Trudeau et al., 1997) The strong affinity these exogenous steroids have for ERs induces wid espread disruption of endocrine and reproductive processes, ultimatel y causing reduced fecundity and decreased population s in aquatic vertebrates (Ankley et al., 2008). Industrial Waste Compounds In addition to estrogenic steroidal compounds found in eff luent waters, industrial waste compounds have been present in large quantities for decades, and also affect reproductive and endocrine systems. The most prevalent compounds include the organic metal complexing agent EDTA, the organic nonionic surfactants n onylphenol ethycarboxylates (NPEs), and the organic plastic byproduct bisphenol A (BPA). EDTA in concentrations similar to that found in effluent water is cytotoxic and has been found to cause reproductive and developmental effects in zebrafish (Lanigan an d Yamarik, 2002). Likewise, in a wide range of vertebrate species, NPEs are known to disrupt reproductive and developmental pathways, because of their small affinity for both and ERs. For example, in a study by Sumpter and Jobling, male juvenile fish e xposed to NPEs produced vitellogenin, a hepatic egg yolk precursor protein that is normally produced by mature female fish in response to estrogen (2005) Because the immature fish produced vitellogenin, NPEs were found to have, albeit weak, estrogenic eff ects in vivo and although the affinities of NPEs for estrogen receptors are weak compared to endogenous estrogens, the high concentrations of these compounds in effluent water

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5 increases their overall estrogenicity and potential for reproductive disruption (Sumpter and Jobling, 2005). BPA, another industrial byproduct with estrogenic effects, has a small affinity for both and ERs throughout tissue in fish and mammalian species (Hatef et al., 2012). In a study by Hatef et al., male goldfish exposed to en vironmentally relevant concentrations of BPA ( 0.2 and 20 #g/L) had increased levels of estrogen receptor mRNA, brain and testis specific aromatase, and vitellogenin (2012) In addition, sperm abundance and motility were reduced, furth er proving its ability to disrupt reproduction and reduce fitness of an animal (Hatef et al., 2012). From observations such as these, it is clear that adverse effects of industrial waste products and their ability to affect a wide range of species, implicates them as agents of biological disruption Not only is their ability to disrupt cellular processes at many sites and through multiple mechanisms concerning, but their presence in such large quantities worldwide renders them as major contributors to the negative effects anthr opogenic compounds are having on aquatic and mammalian species (Paterson and Metacalfe, 2008 ) Pharmaceutical Compounds In addition to steroids and industrial waste products, o ther compounds that are found in WWTP effluent worldwide and adversely affect a quatic and mammalian species are pharmaceuticals (Corcoran et al., 2010) Although many pharmaceuticals act on evolutionarily conserved pathways disrupting fitness and life histories in diverse organisms neuroactive drugs such as selective serotonin reupt ake inhibitors (SSRIs) and monoamine oxidase inhibitors (MAOIs) are most concerning This is because their mechanism of action is only known at the cellular level, but how they affect neural pathways overall is still unknown ( Corcoran et al., 2010 ) In add ition, they are engineered

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6 to have potent and long lasting effects on neural tissue, and because anatomy is conserved in neuroendocrine systems such as the serotonergic system across vertebrates, they have affinities for binding at evolutionarily conserved target sites in a wide range of non target species (Corcoran et al., 2010 ) The effects of psychopharmaceuticals vary between species as these drugs act at multiple sites, and can disrupt physiological processes through multiple mechanisms. For example, f luoxetine, a widely prescribed SSRI, affects the hypothalamus of fish, and was shown to greatly disrupt feeding and metabolism in male goldfish (Mennigen et al., 2010). In addition, disruption of aggressive and sexual behavior, and also modulation of the H PG axis has been reported from fish exposure to psychopharmaceuticals (Kreke and Dietrich, 2008). In a study by Kreke and Dietrich, fish exposed to SSRIs had increased levels of serotonin, which stimulated release of gonadotropins from the hypothalamus, an d therefore stimulated release of FSH and LH from the pituitary (2008) Meanwhile fish exposed to pharmaceuticals that increased presence of dopamine had a decrease in the release of gonadotropins, and therefore a depression in all aspects of the HPG axis (Kreke and Dietrich, 2008). SSRIs and other neuroactive drugs not only affect the central nervous system, they also directly affect organs such as the gonads. In a study by Cerd‡ et al., female killifish ( Fundulus heteroclitus ) exposed to 0.05 M of 5 HT hydrochloride, an SSRI, demonstrated a significant reduction in fecundity from increased serotonin levels blocking meiotic maturation of the oocytes (1998) From studies involving pharmaceutical exposure and observations of bio logical disruption, it is clear that neuroactive compounds pose a complex problem because they are present in effluent waters worldwide, and are chronically acting on native species with known and

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7 potentially unknown adverse effects. Therefore, pharmaceuti cals must be further studied to understand how chronic exposure affects species reproduction and overall survival. Boulder Creek Stream Background Locally, the impact of treated effluent water containing compounds such as estrogens, industrial waste, and pharmaceuticals was examined using a fi sh model in a long term study at the City of Boulder, CO WWTP (Barber et al., 2012, Vajda et al., 2006, 2011) Boulder Creek receives its pristine water source from snow melt at the continental divide, and flows thro ugh the mountain foothills relatively unmodified besides minor nonpoint source pollution from human activity such as fishing and swimming (City of Boulder, 2012). The native stream water flows until it reaches the 75 th St. WWTP, where 30 80% of the stream flow is then composed of effluent water depending on the season and amount of snowmelt (Barber et al., 2012). The change of season and subsequent variations in snowmelt create substantial fluctuations in stream flow, where low flow occurs during fall and w inter months from October to March, and high flow during spring and summer months from April to July. The unpolluted snowmelt source, relatively undisturbed flow through the city of Boulder, single effluent source from the WWTP, and the fluctuations in wat er flow leading to a dynamic effluent load, has provided an ideal site for examining effects of anthropogenic effluent compounds on organisms.

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8 Effluent Source and Contaminants Wastewater from Boulder and Gunbarrel urban populations enters the City of Boul der WWTP where water is filtered from debris, compounds are absorbed or biochemically converted and removed, and pathogens are destroyed by purification and disinfection processes before the eff luent water is released into Boulder Creek (City of Boulder, 2 012) Although the WWTP has undergone massive upgrades to improve removal efficiency and purification methods, many compounds have still not been completely removed and have been released into the creek. From over a decade of water chemistry analyse s of th e Boulder Creek, efflu ent water has included hundreds of compounds, including natural and synthetic steroids, industrial wastes, and pharmaceuticals (Barber et al., 2012). It is clear from this Boulder Creek study that the anthropogenic compounds have var ied over time, reflecting a change in demographics, chemical usage, and WWTP treatment processes (USGS, 2012). The most notable change s took place between 2006 and 200 7 when a sharp decline of NPEs occurred due to the establishment of governmental policie s limiting the commercial use of these compounds (City of Boulder, 2012) In addition, a fifty million dollar upgra de of the WWTP in 2007 occurred that implemented the activated sludge process, which significantly improved the removal of steroidal estrogen compounds, including the potent EE2 ( Figure I.1) ( USGS, 2012 ) Although t hese changes improved effluent water chemistry immensely, other compounds that have been found in effluent waters for decades are still being found today including compounds that ar e known to affect vertebrate species such as pharmaceuticals, EDTA, and BPA (Barber et al., 2012).

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9 Figure I.1 Boulder Creek EEq. E strogenicity in effluent waters from 2005, 2006, 2008, and 2011 from estrogen compounds such as E1, E2, E3, and EE2 show s that s ignifican t reductions occurred with a major WWTP upgrade in 2007. Fish Exposure in Boulder Creek The biological effects of the City of Boulder WWTP effluent were evaluated in 2005, 2006, 2008, and 2011 with controlled flow through fish exposure ex periments by pumping water just upstream (reference) and just downstream (effluent) from the Boulder Creek 75 th WWTP into a mobile lab housing several tanks of fathead minnow ( Pimephales promelas ) purchased from Aquatic Biosystems in Fort Collins Colorado Fathead minnow were chosen as an animal model for several reasons, most notably because they have short generation times, they are large enough for adequate tissue samples, their geographical distributions are broad, they swim throughout the entire water column, encountering all contaminants regardless of contaminant partitioning behavior their steroid dependent secondary sex characteristics are distinct and easy to measure, and because there are over forty years of basic biology and established paramete rs in their neuroendocrine and reproductive processes ( Ankley et al., 2001; Vajda et al., 2011 ) 0 10 20 30 40 50 60 2005 2006 2008 2011 EEq (ng/L) Year Effluent Estrogenicity ( EEq )

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10 Figure I.2 Fathead minnow ( Pimephales promelas ). The fathead minnow was used for the effluent exposure experiments. Secondary sex characteristics were measu red including the dorsal fatpad and nuptial tubercles on the snout. Gonads were removed for later histology and immunohistochemical analysis. In the mobile lab, reproductively stimulated male fathead minnow were randomly selected and exposed for 28 days in effluent dilutions of 100% effluent, 50% effluent or 0% effluent, with controlled temperature, photoperiod, diet, aeration, and flow. After this 28 day period, fish were sacrifice d, sexed, measured, and weighed. The number of nuptial tubercles on the s nout of the fish and the prominence of the dorsal fat pad were then measured, and blood plasma was collected from the caudal vein for plasma vitellogenin measurement (Vajda et al. 2004) Organs including the gonads were then removed, weighed, and preserve d in neutral buffered formalin (10% NBF) (Vajda et al., 2004 ). Gonads were later embedded, sectioned, and stained with hematoxylin and eosin for analysis of morphology.

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11 Figure I.3 Plasma Vitellogenin in Reference and Effluent Exposed Fish. Plasma vite llogenin was measured after fish were exposed for 28 days in either Boulder Creek's effluent or reference waters. This egg yolk precursor protein is a biomarker for estrogen exposure. Figure I.4 Nuptial Tubercles in Effluent and Reference Exposed Fis h. Secondary sex characteristics including nuptial tubercles were measured on fish exposed to either the reference or effluent waters in 2005, 2006, 2008, and 2011. Changes in secondary sex characteristics are indicative of endocrine disruption. 0 1000 2000 3000 4000 5000 6000 7000 8000 9000 2005 2006 2008 2011 ng/mL Year REF 50% Eff EFF Plasma Vitellogenin (day 28 ) 0 2 4 6 8 10 12 14 16 18 20 2005 2006 2008 2011 Number Year REF 50% EFF EFF Nuptial Tubercles (day 28 )

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12 Figu re I. 5 Gonadosomatic Index of Reference and Effluent Exposed Fish. The ratio of the gonad weight to the fish weight was compared as a gonadosomatic index in fish exposed to either reference or effluent waters in 2005, 2006, 2008, and 2011. Biological Fin dings from Fish Exposures In 2005, although fish exposed to effluent water had no increased mortality, biological observations included diminished primary and secondary sexual characteristics evident by less abundant and prominent nuptial tubercles (figur e I.4 ) less prominent dorsal fat pads, reduced gonadosomatic index (GSI) (figure I.5 ) decreased sperm abundance, and the complete absence of sperm in several male fathead minnow ( Vajda et al., 2011 ). In addition, plasma vitellogenin was greatly elevated in male fish exposed to any amount of effluent in this year (figure I.3 ) ( Vajda et al., 2006 ). From the biological findings of the effluent exposed fish, it was concluded that complex mixtures in effluent water downstream from the WWTP was significantly di srupting reproductive processes. In 2006, fish from the 28 day exposure maintained in 100% effluent showed sign ificant disruption of secondary sexual characteristics with less abundant and prominent nuptial tubercles (figure I.4 ) and less prominent dorsal fat pads. E ffluent 0 0.2 0.4 0.6 0.8 1 1.2 1.4 2005 2006 2008 2011 Year REF 50/50 EFF Gonadosomatic Index (day 28)

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13 exposed fish also had significantly increased plasma vitellogenin (figure I.3 ) However, there was no significant effect at the g onad, which was evident by the GSI and sperm abundance being similar to reference exposed fish (figure I.5 ) ( Vajda et al., 2006 ). In 2006, g overnmental policies limited the commercial use o f the extremely potent compounds NPEs and a major WWTP upgrade occurred during this time, which significantly reduced estrogens. Therefore, it was expected that biological ob servations of the effluen t exposed fish would be similar to reference exposed fish (Vajda et al., 2011) However, from the effluent exposed fish measurements and observations, it was found that although effluent water in 2006 did not contain compounds estr ogenic enough to induce changes in gross morphology of the gonad it was estrogenic enough to induce plasma vitellogenin and affect secondary sex characteristics in male fish (figures I.1 and I.3 ) From observations in 2008 and 2011, effluent exposed fish had statistically similar observations to the reference fish with sex ratios, fat pad prominence, nuptial tubercle number and prominence, sperm abundance, and GSI (figures I.3, I.4, I.5 ) (Barber et al., 2012). Therefore, the biological observations of the se fish exposed to effluent water reflected the significant shifts in water chemistry from reduced estrogens and NPEs after 2007 (figure I.1) Testing for Discre t e Changes Because the m ost potent estrogenic compounds, EE2, E2 and NPE, greatly decli ned in effluent water after 2007 and remain scarce today, it could be assumed that disruption is no longer occurring. However, hundreds of anthropogenic compounds, including industrial waste byproducts and pharmaceuticals have been found for decades,

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14 and are st ill present today ( USGS, 2012 ) Because these toxicants remain present in effluent water, and because either very specific measurements of vitellogenin or coarse scale biological measurements such as fish length, gonad size, and gonad histology were only c onsidered up until this point, it cannot be assumed that no repro ductive disruption is occurring. E ffects of toxicants at the molecular, cellular, and tissue level may not always be obvious at higher levels of organization and would not be reflected in mea surements such as gonad weight or histology In addition, exposures of 28 days might not affect physiological processes such as spermatogenesis which require over 4 months for completion ( Alves et al., 2012 ) and measurements such as sperm abundance, gona d weight, GSI, and histology could remain normal if fish were not exposed for most of this time. Therefore, because evidence of disruption is subtle, and complex mixtures in effluent water have diverse mechanisms of action that d isrupt physiological proces ses at multiple sites (figure I.6 ) biomarkers must be general enough to determine end points of disruption from multiple sites and via multiple mechanisms but specific enough to examine subtle changes at the cellular level. Ultimately, these tests must b e accurate in answering the question of whether or not effluent contaminants are adversely affecting reproduction, and whether broad ranges of species that are exposed to these compounds are being affected as well.

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15 Figure I. 6 Disruption of the Hypothal amus Pituitary Gonadal Axis This generic figure summarizes potential disruption of the HPG axis in most vertebrate species from industrial, pharmaceutical, and chemical toxicants found in effluent water.

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16 End points of Disruption To evaluate whether ef fluent toxicants are affecting reproductive processes, the testes of male fathead minnow from the exposure experiments were examined These organs were analyzed because they are the ends points of the reproductive pathway, and also because they are the res ponsible for production of gametes, and ultimately fertility of the fish More specifically, t he gonads are the most downstream organ in the hypo thalamic pituitary gonadal axis (HPG axis), which is responsible for initiating and maintaining spermatogenesis and steroidogenesis, and which also can be affected by many different effluent compounds through multiple sites and via multiple mechanisms (figure 1.6). To analyze discrete disruption in the gonads from exposure to environmental toxicants in effluent w ater, cellular changes in the testes were measured. More specifically, cellular proliferation and programmed cell death, or apoptosis, were analyzed, both of which are normal and essential physiological processes during spermatogenesis (Nagahama, 1994). S permatogenesis Spermatogenesis in vertebrates is a process in which haploid motile cells are produced from diploid germ cells and mature to form viable gametes within the seminiferous tubules of the testis (Nagahama, 1994) During development and into adu lthood, the process of spermatogenesis relies on several factors including hormones and other paracrine factors. These signals allow for sperm production to be increased and cells to proliferate at sexual maturity and during reproductive seasons, and also for cells to be decreased and undergo apoptosis during a reproductive hiatus. It is the dynamic

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17 balance between cellular proliferation and apoptosis in the gonads that maintains homeostasis and allows for spermatogenesis to occur (Nagahama, 1994). Therefor e, proliferation and death of cells in the gonad are necessary, and the dynamic balance between both of these processes is requir ed for reproductive success of male individuals Proliferation and Spermatogenesis Proliferation in testicular tissue constan tly occurs as germ cells undergo meiosis to produce spermatogonia. These cells mature in the epithelium to become spermatocytes, and travel toward the lumen as they mature into spermatids, and finally, spermatozoa. This process is dependent on the release of gonadotropin releasing hormone (GnRH) from the hypothalamus, which acts on the adenohypophysis where it binds to gonadotrophs and stimulates the release of FSH and LH (Nagahama, 1994) Once released, LH and FSH travel through the blood to the testicular tissue, where LH stimulates leydig cells to produce testosterone, and FSH stimulates sertoli cells to initiate meiosis of spermatogonia. It is the synergistic effects of FSH and LH that promote cell division and proliferation of spermatogonia in the gonad (Nagahama, 1994). These hormones also act to negatively feedback on both the adenohypophysis and the hypothalamus. PCNA Effluent compounds such as EE2, NPE, BPA, and pharmaceuticals can directly or indirectly interfere with regulation of the HPG axis thr ough ERs or interaction with diverse neurotransmitter systems (figure I.6) For example, neuroactive pharmaceuticals that increase dopamine in the brain inhibit the GnRH releasing neurons in the hypothalamus, and therefore cause decreased LH, FSH, and ulti mately a decrease in

PAGE 28

18 cellular proliferation (Corcoran et al., 2010). In addition, exogenous estrogens such as EE2 also inhibit GnRH neurons in the brain. Estrogens play an essential role in regulating the HPG axis in males, because testosterone is aromatiz ed to estrogens locally in the brain, which then negatively feedback on the GnRH neurons (Jobling et al., 2006). From what is understood about estrogen mimicking compounds that are found in effluent waters, ERs are activated, which leads to suppression of the HPG axis, and as a consequence, overall decreased reproduction in male animals (Jobling et al., 2006) Apoptosis and Spermatogenesis Programmed cell death, or apoptosis, is a normal and essential process in all living tissue including the gonads, whe re it is fundamental for maintaining the correct ratio between sertoli cells and sperm cells (Andreu Vieyra C. et. al, 2005). Initiation of apoptosis c an occur via the intrinsic or extrinsic pathways in instances such as DNA damage, incorrect replication o f DNA, phosphokinase C inhibition, or by the Fas ligand or tumor necrosis factor (TNF ") ( Andreu Vieyra C. et. al, 2005). It is the activation of the intrinsic or extrinsic pathway, or even a combination of bot h that can influence cell death, and if chem ical or non chemical stressors occur, these pathways can be activated to initiate apoptosis. TUNEL The effects of compounds such as EE2, NPE, BPA, and pharmaceuticals on the brain and pituitary may not only influence proliferation, but apoptosis as well. Increasing dopamine in the brain and depression of the HPG axis can lead to less FSH and LH, and therefore less stimulation of gonadal cells (Trudeau et al., 2005). Gonadotropin

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19 withdrawal not only causes less proliferation, but also decreases the factors maintaining cell function and cell survival (McClusky, 2012). Therefore, many effluent compounds that cause inhibition of hormone release from the hypothalamus and/or pituitary may result in increased apoptosis. PCNA and TUNEL To test for discr et e change s from toxicant exposure, two assays were used to identify cellular proliferation and apoptosis in the testes of fathead minnow. T erminal deoxynucleotidyl transferase biotin dUTP nick end labeling, or TUNEL, is an assay that allows for the location and amo unt of apoptosis to be measured in tissue. This assay binds to the 3' hydroxyl ends of DNA fragments that are cre ated only during apoptosis (D'Andrea et al., 2010 ). Cells undergoing necrosis from other causes will not stain positive for the TUNEL assay, be cause DNA is sheared haphazardly during necrosis and is not fragmented in the same manner as during apoptosis ( D'Andrea et al., 2010 ). The TUNEL assay was chosen because it is sensitive and reliable, and has been applied to a wide range of vertebrate spe cies for analyzing cellular changes from toxicant exposure. For example, abundance of TUNEL in rat testes was measured after exposure to the toxicants phthalates, methylmercury and nonylphenols (Moffit et al., 2007; Fujimura et al., 2012 ; Han X. et al., 2 004 ), and was also used to measure apoptosis in the gills of brown trout in various exposure experiments ( Rojo and Gonzalez, 1998). TUNEL has also been used in numerous experiments measuring sperm apoptosis in human men ( Delbs et al., 2010). In addition t o being a sensitive, reliable, and wide ly use d assay for measuring biomarkers in many di fferent vertebrates, TUNEL stains cells undergoing a process that cannot be reversed, and therefore shows cellular disruption from many hours

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20 or even days of toxicant e xposure. Unlike measuring hormones such as FSH or LH, which can recover hours after a toxicant exposure TUNEL shows apoptosis, a process from which cells cannot recover ( D'Andrea et al., 2010 ) P roliferating cell nuclear antigen, or PCNA, is an assay th at binds to a homotrimeric nuclear protein associated with cell s undergoing mitosis or meiosis and therefore allows for the location and abundance of proliferation to be measured in tissue. Like TUNEL, PCNA is a sensitive, widely used, and reliable assay that has been used in a wide range of vertebrates for analyzing changes in cellular proliferation from exposure to toxicants. For example, PCNA has been used to measure proliferation in testes of rats exposed to benzenes, cyclophosphamides, and ethanol (Ch andra et al., 1997, D'Andrea et al., 2008). In addition, PCNA has been used on zebrafish tissue exposed to acrylamide, and has been used in numerous studies measuring proliferation of cells after exposure to carcinogens (Parng et al., 2007; Girod et al., 1 994). Similar to TUNEL, PCNA binds to cells that are undergoing a process that cannot reverse in a short period of time. Unlike measuring hormones that can fluctuate over the course of a day, or even within the hour, cells undergoing mitosis and meiosis wi ll remain in that state for longer periods of time (Bennet, 1971) Therefore, PCNA is a highly effective assay for not only measuring pro liferating cells in a wide range of vertebrate species, but also for measuring cumulative effects after days or weeks o f exposure to effluent toxicants Expected Outcomes From increased amounts of estrogens and NPEs in Boulder Creek effluent in years 2005 and 2006, it was expected that TUNEL be more abundant in testes of effluent exposed fish This was hypothesized becau se these exogenous compounds bind to ERs,

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21 and can cause the downregulation of the HPG axis and ultimately, gonadotropin withdrawal In addition, from previous measurements of decreased primary and seco ndary sex characteristics, it was predicted that apopto sis was occurring more frequently at the cellular level. In 2008, however, TUNEL was expected to be s imilar between effluent and reference exposed fish bec ause of changes in Boulder Creek effluent water chemistry, and also from previous observations reveal ing no significant changes in fish morphology or gonad histology. PCNA was predicted to be less abundant in 2005 and 2006 effluent exposed fish from the depression of the HPG axis, which could have occur red from many effluent contaminants found in Boulder Creek effluent. In 2008, however, PCNA was expected to not be significantly different between referen ce and effluent exposed fish because certain EDCs were removed from effluent water, and also because previous observations revealed no significant changes in morphology or gonad histology in effluent exposed fish

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22 CHAPTER II M ETHODS On site Fish Laboratory and Experimental Design Reproductively stimulated 12 month adult male fathead minnow were expose d to either 100% effluent or 100% refe rence in 2005, 2006, and 2008, using a flow through system mobile lab oratory located on site at the c ity of Boulder WWTP (Vajda et al., 2011). For 28 days, controlled variables included temperature, lighting, brine shrimp diet, aeration, flow, and effluent concentration dilution (Vajda et al., 2011). After 28 days fish were sacrificed and measurements were recorded including fish length, fish weight, nuptial tubercle number and prominence dorsal fat prominence, hematocrit, and plasma vitellogenin (Vajda et al., 2011). Gonads and other organs were dissected and weighed, and the gonadosomatic index was calculated (Vajda et al., 2011). All tissue samples were preserved in 10% neutral buffered formalin for later histological preparations (Vajda et al., 2011). Double Labeling Procedure Gonad tissue from the 2005, 2006, and 2008 experiments was embedded in paraffin wax and sectioned at 5 m ribbon s that were then mounted onto glass s lides. Slides were then dewaxed and r ehydrated in xylene (3 x 5 minutes each) 100% ethanol, 95% ethanol, and 70% et hanol (all 2 x 3 minutes each). Slides were then immersed in deionized water for 3 minutes. For antigen retrieval, slides were microwaved for 5 minutes in 0.1 M citrate buffer then immediately placed in 0.1 M Tris HCl s olution for 30 minutes. To quench endogenous peroxidase activity, slides were immersed in a 0.3% H O methanol solution for 15 minutes. The TUNEL reaction mixture was prepared

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23 following the Roche In Situ Cell Death Detection Kit, AP, and 100 L of TUNEL reaction solution was added to each slide. After slides were coverslipped, they were incubated for 6 0 minutes in a 37 ¡ C humidified dark chamber, then rinsed 3 x 5 minutes in 0.1 M PBS. The Roche Fast Red Solution was prepared by dissolving one fast red tabled into 5 mL of substrate buffer and vortexing until tablet was dissolved, and after slides were b lotted dry 50 100 L was added to each slide until the tissue was covered in solution. Slides were incubated for 10 minutes at 15 25 ¡ C in the dark, and a fter this first incubation, excess liquid was removed and fast red substrate solution was added again. S lides w ere incub ated for another 10 minutes in a dark chamber Slides were then rinsed in 0.1 M PBS 3 x 5 minutes. The primary PCNA antibody was prepared in a 1:200 dilution in 0.1 M PBS. After slides were blotted dry the antibody was added, and each slide was c overslipped and left to incubate over night at 4 C in a humidified chamber. After this incubation, slides were then washed in 0.1 M PBS 3 x 2 minutes, and the secondary PCNA antibody was prepared following the Vectastain ABC Kit protocol. After the secondary antibody was applied to each sample, the slid es were coverslipped and incubated for 60 minutes in a humidified chamber. After this incubation, slides were washed in 0.1 M PBS 3 x 2 minutes, the ABC reagent from the Vectastain ABC Kit was applied to each sample and slides were coverslipped. The slide s w ere placed in a humidified chamber and incubated with the ABC reagent for 30 minutes. Slides were then rinsed in PBS 3 x 2 minutes. Th e DAB, or peroxidase substrate solution, was prepared from the Vector Laboratories protocol, and 500 L was added to each slide for approximately 5 minutes. S lides were ta pped to remove excess DAB, and rinsed in tap water. Slides were then placed in

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24 hemato xylin for approximately 45 seconds, and rinsed in tap water unti l the solution ran clear. After excess liquid was removed copious amounts of simpomount was added to each slide, and then suctioned off using a transfer pipette to remove air bubbles. Figu re II.1 Example of Sectioned Slide Used For Analysis. Each slide contained two 5 m thick wax ribbons, each containing tissue sections f rom different areas of the testis

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25 Figure II.2 Blind Analysis of Slides. To analyze the slides blindly without knowing the treatment group of each fish, the labels on each slide were covered a nd randomly assigned numbers. Slide Analysis Testes from reference and effluent exposed fish in 2005, 2006, and 2008, were embedded in wax, and sectioned into 5 m thick wax ribbons. Specifically, the wax ribbons were mounted onto glass slides in a rando m order, so that the t wo ribbons on each slide contained different sections of the gonad (figure II.1) Therefore every slide contained multiple sections of the gonad from each fish and was more represent ative of the gonad as a whole not just one partic ular area For double labeling, five fish out of approximately thirty fish samples were selected at random from each treatment group (reference and effluent) in years 2005, 2006, and 2008 giving a total of thirty fish samples for double labeling with PCNA and TUNEL The labels on each slide were then

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26 covered with tape, and randomly assigned a number so that analysis was performed blindly (figure II.2). From these slides, five tubules wer e selected from each of the two wax ribbon s giving a to tal of ten tub ules for each slide Tubules we re selected at random, however t here was a bias for round tubules because they represent cross sections of the gonad, unlike oval or irregular shaped tubules, which represent oblique sections of the gonad. Tubules that were not cross sections often contained varying densities of sperm in the lumen. Therefore this bias was applied when selecting tubules to avoid miscounting cell density, and misrepresenting amount of TUNEL or PCNA positive cells. Us ing Image J software, the t ubule area, lumen area and sperm area (figure II .3 ) were measured for each tubule. The lumen area was subtracted from the tubule area to give the area of the epithelium, and sperm area was measured carefully, with areas without sperm subtracted from t he t otal sperm area (figure II.3 ). PCNA and TUNEL positive cells were counted in the epithelium and were divided by the epithelium area to give the number of cells per m$ (figure II.4 ) TUNEL positive cells were counted individually in the sperm are a and div ided by the total sperm area. PCN A positive cells were counted by randomly selecting multiple sample areas within the sperm and counting the cells within these areas, because most of the sperm cells stained positive for proliferation and this expedited th e counting process (figure II.4 ) The sample areas were added together then multiplied by the total sperm area to give total number of PCNA positive cells per m$ of sperm area

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27 Figure II.3 Tubule Morphometrics. The tubule area was outlined (yellow) following the basement membrane of the epithelium, and the lumen area was outlined (red) following the cells bordering the lumen. The lumen area (red) was subtracted from the tubule area (yellow) to find the area of the epithelium. Last, the sperm area was measured (green) outlining the sperm area, and areas inside without sperm cells were measured and subtracted from the total sperm area.

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28 Figure II. 4 TUNEL and PCNA Positive Cell Count. TUNEL positive cells stained red in the nucleus and PCNA cells stained dark brown in the nucleus. TUNEL and PCNA positive cells were counted individually in the epithelium, and were divided by the area of the epitheli um (figure II.3 ). The number of TUNEL positive cells were counted individually in the sperm area, and divided by area of the sperm. Samples sections of sperm were circled at random, and the nu mber of PCNA positive cells was counted inside. The samples were added together and multiplied by the sperm area to give total PCNA positive cells per area of sperm. In this figure, there are several TUNEL positive cells in the epithelium, and hundreds of TUNEL positive cells in the sperm, which is evident from the da rk red staining. Likewise, in the epithelium, there are dozens of PCNA positive cells in the epithelium, and many hundreds of PCNA positive cells in the sperm area, evident by the dark brown staining in the nucleus. Statistical Analysis Using the statisti cal progr am R, data were analyzed using a two way analysis of vari ance and Tukey post hoc test To correct for homoscedasticity and reduce probability of witnessing one or more type I errors, the Holmes Bonferroni correction was also

PAGE 39

29 applied. Statistical d ifferences between PCNA positive cells in the epithelium and lumen, and TUNEL positive cells in the epithelium and lumen were found between effluent and reference exposed fish in y ears 2005, 2006, and 2008, with a significance level of 0 .05

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30 CHAPTER III RESULTS Testicular Morphometrics of 2005, 2006, and 2008 T is sue Figure III.1 Averages of Sperm, Lumen and Epit helium Areas. The average sperm area, tubule area, and epithelium area of all fish samples in each treatment group each year was calculat ed 0 2000 4000 6000 8000 10000 12000 14000 Reference Effluent Reference Effluent Reference Effluent 2005 2006 2008 m Sperm Tubule Epithelium Sperm, Lumen, and Tubule Areas

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31 PCNA and TUNEL Immunoreactivity Fi gure III.2 Epithelial and Luminal TUNEL Reactive Cells in 2005 Abundance of e pith elial TUNEL cells was significantly higher in effluent exposed fish (p<0.01). Likewise, abund ance of sperm TUNEL cells was higher in effluent fish (p<0.001) compared to reference exposed fish. Figure III.3 Epithelial and Luminal PCNA Reactive Cells in 2005 Epithelial (p<0.05) and sperm cells (p<0.01) positive for PCNA was significantly highe r in effluent exposed fish. 0 0.001 0.002 0.003 0.004 0.005 0.006 Epith TUN Sperm TUN cells/m 2005 TUNEL Reference Effluent ** *** 0 0.03 0.06 0.09 0.12 0.15 Epith PCNA Sperm PCNA cells/ m 2005 PCNA 2005 Reference 2005 Effluent **

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32 Figure III.4 Epithelial and Luminal TUNEL Reactive Cells in 2006 Epithelial cells positive for TUNEL in effluent exposed fish were statistically higher than reference exposed fish (p<0.001). Likewise, sperm cells positive f or TUNEL were higher in effluent fish compared to reference (p<0.01). Figure III.5 Epithelial and Luminal PCNA Reactive Cells in 2006 Epithelial cells positive for PCNA in effluent exposed fish were not statistically different between reference and ef fluent exposed fish, however, sperm PCNA positive cells were higher in reference compared to effluent (p<0.001). 0 0.0008 0.0016 0.0024 0.0032 0.004 0.0048 Epith TUNEL Sperm TUNEL cells/ m $ 2006 TUNEL 2006 Reference 2006 Effluent *** ** 0 0.03 0.06 0.09 0.12 0.15 0.18 Epith PCNA Sperm PCNA cells/ m $ 2006 PCNA 2006 Reference 2006 Effluent ***

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33 Figure III.6 Epithelial and Luminal TUNEL Reactive Cells in 2008 Both e pithelial and sperm cells positive for TUNEL in effluent exposed f ish were not statistically different than referen ce exposed fish Figure III.7 Epithelial and Luminal PCNA Reactive Cells in 2008 Epithelial cells positive for PCNA in effluent exposed fish were not statistically different than re ference exposed fish, however, PCNA positive cells in sperm of effluent fish were statistically higher than reference fish (p<0.05). 0 0.001 0.002 0.003 0.004 Epith TUNEL Sperm TUNEL cells/ m $ 2008 TUNEL 2008 Reference 2008 Effluent 0 0.03 0.06 0.09 0.12 0.15 Epith PCNA Sperm PCNA cells/ m $ 2008 PCNA 2008 Reference 2008 Effluent

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34 Table III.1 Previous Measurements, Water Chemistry, and Biomarkers GSI Vitell Nup.Tub. Sperm Estrog. Epi TUN Sperm TUN Epith PCNA Sperm PCN A 2005 +++ +++ +++ +++ +++ Eff Eff Eff Eff 2006 ++ ++ ++ Eff Eff Ref 2008 Eff In 2005, GSI was significantly lower, vitellogenin was increased, nuptial tubercles were less pronounced, and sperm abundance was lower in effluent exposed fish, all ref lecting an increased amount of estrogens in the effluent water (Vajda et al., 2009; Barber et al., 2012). With discrete testing using PCNA and TUNEL, it was found that effluent exposed fish had higher amounts of apoptosis and higher amounts of proliferatio n than reference exposed fish (p<0.05). In 2006, GSI was not significantly between the two treatment groups, vitellogenin was still abnormally higher in effluent fish, nuptial tubercles were less pronounced but not as diminished as the previous year, whi ch all reflected a decrease in effluent estrogens (Vajda et al., 2009; Barber et al., 2012) Similar to 2005, effluent exposed fish showed higher amounts of TUNEL positive cells, but PCNA was more abundant in the sperm of reference exposed fish (p<0.05). In 2008, GSI, vitellogenin, nuptial tubercles, and sperm abundance were all similar between the effluent and reference exposed fish, reflecting the removal of estrogens and NPEs from the effluent water (Vajda et al., 2009; Barber et al., 2012) This result ed in TUNEL being similar in both reference and effluent exposed fish, but PCNA being more abundant in sperm of effluent exposed fish compared to reference fish (p<0.05).

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35 CHAPTER IV DISCUSSION Cellular Proliferation From the higher load of effluent conta minants in years 2005 and 2006 it was predicted that fish maint ained in effluent waters would show less proliferation within the ir gonads compared to reference fish In 2005 it was found that proliferation measured by amount of PCNA po sitive cells was si gnificantly higher in the sperm and epithelium of effluent fish compared to reference fish (p<0.01) (figure III. 3 ). In 2006, however, proliferation was lower in the sperm of effluent fish (p<0.001) and was not different in the epithelium between the two t reatment groups (figure III.5) In fish exposed to effluent in 2008, when water chemistry revealed a decrease in estrogens and NPEs, and when almost no disruption in either secondary characteristics or gonad histology occurred it was expected that no diff erence would occur between effluent and reference exposed fish. It was found that the amount of PCNA positive cells was not significantly different in the epithelium between the two treatment groups in 2008, but more PCNA positive cells were found in the s perm of effluent fish compared to reference exposed fish (p<0.05) (figure III.7) The insignificant difference of PCNA positive cells in the epithelium between the two treatment groups in 2006 and 2008, and the inconsistent trends of PCNA in the sperm bet ween all years, suggests that PCNA is not as reliable as some previous studies have found (Chandra et al., 1997) However, other studies have concluded that PCNA might not be as sensitive or reliable as TUNEL. For example, D'Andrea et al. found that when a nalyzing rat testes, PCNA might not be as sensitive as TUNEL especially when

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36 using the double labeling technique (2010). Likewise, from findings of this study, it is suggested that when staining testes of fathead minnow, this assay is not as beneficial as using TUNEL, because it does not provide more information than what is provided by secondary sex characteristics, gonad histology, or plasma vitellogenin (Table III.1) Two p ossible reasons for PCNA not being consistent are that this assay might be less sensitive than TUNEL, and human error might have occurred when counting cells. Due to artifact s in the tubules counting number of cells was at t imes difficult, which most likely introduced human error. In addition, most of the many hundreds or even thousa nds of sperm c ells stained positive for PCNA which made counting cells even more difficult. P erhaps this assay would be more easier to count and thus more informative in less proliferative tissue such as hepatic or renal tissue. Organs that normally under go little mitosis might show more profound differences in cellular proliferation when the organism is stressed from exposure to environmental toxicants. TUNEL Increased concentrations of estrogenic effluent compounds in 2005 and 2006 were predicted to in crease TUNEL reactive cells in the gonad of effluent exposed fish compared to reference fish. It was found that epithelial and sperm TUNEL positive cells were significantly higher in 2005 and 2006 effluent exposed fish (p<0.001) compared to reference fish (figures III.2 and III.4) In 2008, when water chemistry revealed a decrease in estrogens and NPEs and when no disruption in either secondary characteristics or gonad histology occurred, it was hypothesized that TUNEL would not be significantly higher in effluent exposed fish compared to reference exposed fish. It was found that epithelial and sperm TUNEL positive cells were not significantly different between

PAGE 47

37 effluent and reference exposed fish in 2008. H owever, there were more TUNEL positive sperm cells in effluent exposed fish in 2008, which although is not statistically s ignificant may be biologically relevant (figure III.6) From the immunohistochemistry results of 2005, 2006, and 2008 tis sue, TUNEL proved to be the more informative and sensitive bio marker. TUNEL in the sperm of effluent fish was significantly higher than in the sperm of reference fish, which is consistent with previous findings of low sperm abundance and reduced GSI in 2005 (table III.1). However, TUNEL also revealed that sperm apopt osis was occurring in effluent exposed fish in 2006 and 2008. TUNEL revealing apoptosis when other measurements failed to respond suggests that it is a sensitive and valuable biomarker for detecting cellular disruption (Table III.1) This is consistent wit h previous studies finding that TUNEL is more sensitive and more reliable then PCNA (D'Andrea et al., 2010). Therefore, this is a useful biomarker when analyzing testes of fathead minnow, because it allows discrete changes from effluent compound exposure t o be seen in the gonad when other measurements show no changes. The higher abundance of TUNEL and therefore apoptosis in spermatic cells of effluent exposed fish in 2006 and 2008 indicates that disruption is still occurring from effluent compounds. The phy siological and functional significance of these findings is that sperm production declines when fish are exposed to the complex mixtures of effluent waters, bu t whether or not this decline has significant impacts on individual fitness depends on a sperm th reshold for fertility. This is difficult to estimate in fish, because even in humans where sperm counts have continuously been monitored in thousands of men from many different countries, interpretation of how much sperm is required for

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38 fertility is in fac t difficult to define and controversial (Sharp and Skakkebaek, 2003) This difficulty is due to factors such as differing subject recruitment, differing types of semen analysis, and even differing season s (Sharp and Skakkebaek, 2003). These factors are eve n more difficult to control in fish, compounding the difficulty in estimating how much sperm is required for male fish to be fertile. However, the aim of this study was only to evaluate whether or not effluent compounds were disrupting reproductive pathway s to cause any changes in gonad cells Moreover, t he aim was to find sensitive tests that could indicate p hysiological disruption in fish when coarse scale measurements would otherwise fail to respond. Further testing is required to correlate the amount of sperm positive for TUNEL with overall fish fertility and reproductive success. Use as Biomarkers Use of Biomarkers in Fathead Minnow Testes From results of this study, the higher prevalence of TUNEL in the gonads of effluent exposed fish with no signific ant decreases in sperm abundance suggests that this assay should be used in combination with hematoxylin and easin staining (H & E) for histological observations. When staining only with H & E, sperm abundance may look high and similar to that of fish exp osed to no toxicants (figure IV.1). This is because H & E only stains the nucleus and cytoplasm, and will therefore stain even apoptotic cells. However, when stained with TUNEL, one can see that significant disruption is occurring in the tubule, evident by abundant TUNEL staining (figure IV.1)

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39 Figure IV.1 Use of TUNEL in Combination with Hematoxylin and Eosin Staining. When analyzing gonad tubules with only hematoxylin and eosin through light microscopy, sperm abundance looks high and similar to t hat of fish exposed to no toxicants. However, when stained with TUNEL, it is evident that sperm apoptosis is abnormally high especially within the sperm cells. Use of Biomarkers in Vertebrates PCNA and TUNEL as biomarkers in any vertebrate species can be informative of discrete cellular changes in tissue. This is because these assays bind to elements that almost all eukaryotic cells have in common. PCNA binds to a nuclear protein found in eukaryotic proliferating cells, and TUNEL binds to nicked 3' OH ends on DNA, which are created in eukaryotic cells as a result of apoptosis (D'Andrea et al., 2010) However, even though both of these assays can be used in most tissue in a wide range of vertebrate species it was found that when analyzing gonad tissue in fa thead minnow, TUNEL was a more dependable, sensitive, and informative biomarker for analyzing cellular changes from multiple pathways of disruption. T he translational ability of PCNA and TUNEL from a vertebrate model organism to another vertebrate may dep end on many factors for example, differences in the tissue

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40 being examined, the neural or hormonal pathways acting on that tissue, and the differences in how intracellular and extracellular signals act not only in the animal but on that particular tissue a s well (Hess and Carnes, 2004) In the testes, for example, proliferation occurs via activation of the HPG axis, which is mostly conserved in vertebrate sp ecies including fish and humans (Trudeau et al., 2003) However, there are some slight differences be tween the HPG axis in fish and humans that might alter the translational ability of PCNA from one species to the other (see Biomarkers for Use in Humans) In addition, a poptosis occurs from a combination of the intrinsic or extrinsic pathways in eukaryotic cells but stressor s that induce apoptosis in one organism might not reach a specific threshold to induce apoptosis in another organism. In addition, differences in organismal level phenomena such as life stages, metabolism, uptake of compounds, etc. can influence events at the cellular level, and can therefore result in differences in cellular response from one organism to another. For these reasons the differences between species are necessary to know and understand when interpreting and predicting how proliferation and apoptosis in one organism will compare to another Use of Biomarkers in Humans When translating cellular proliferation from fathead minnow to humans, there are several factors to note. N ot only are there anatomical differences, but phy siological differences as well. For example, neurons in the hypothalamus of fathead minnow d irectly innervate the pituitary and release GnRH di rectly onto gonadotrophs, where as humans have a portal system that moves GnRH from the hypothalamus to the pituit ary via two capillary networks (Trudeau et al., 2003). In addition, distributions of ERs in the hypothalamus and pituitary differ between fish and humans, as do specific pathway s of

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41 estrogen feedback regulation (Trudeau et al., 2003) But ER distribution i s not only different in the brain, f or example, estrogen receptor s (ER ") are located in different areas of the fish t estes compared to the human' s (Hess et al., 2004) The se differences in ER d istribution between humans and fish could be significant because of the potential differing rol e s in regulating spermatogenesi s, and because exog enous estrogenic compounds may have different effects if they are binding to receptors in different locations and in different densities ( Hess and Carnes, 2004). However, differences in ER distribution in the brain do not affect the prop ortion of negative feedback from estrogens between fish and mammals, and therefore do not have an impact on pathways leading to cellular changes at the gonad ( Hess and Carnes, 2004 ) In addition ER has been found to be irrelevant in maintaining or initiating spermatog enesis in almost every species ( Hess and Carnes, 2004) ER on the other hand, is highly implicated in initiating and maintain ing spermatogenesis, and is abundant in both fish and humans More specifically, ER is found in almost every cell type of both human and fish testis such as leydig, sertoli, germ cell s, and peritubular cells (Hess and Carnes, 2004). Therefore, f rom the distribution of critical ERs in humans and fish, and f rom the HPG axis being conserved in both, effluent contaminants that increased apoptosis in fathead minnow are predicted to also increase apoptosis in testis of humans.

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