Pholis gunnellus BIOINDICATOR IN SAINT JOHN HARBOUR, NEW BRUNSWICK, CANADA by

Pholis gunnellus BIOINDICATOR IN SAINT JOHN HARBOUR, NEW BRUNSWICK, CANADA by
THE POTENTIAL USE OF ROCK GUNNEL (Pholis gunnellus) AS A
BIOINDICATOR IN SAINT JOHN HARBOUR, NEW BRUNSWICK, CANADA
by
Lottie E. Vallis
B.ScH. Memorial University of Newfoundland 2000
A thesis Submitted in Partial Fulfillment of
the Requirements for the Degree of
MASTER OF SCIENCE
In the Graduate Academic Unit
of Biology
Supervisors:
Deborah L. MacLatchy, Ph.D., Department of Biology
Kelly R. Munkittrick, Ph.D., Department of Biology
Examining Board:
Kenneth Sollows, Ph.D., Department of Engineering
Jeffery Holahan, Ph.D., Department of Biology
This thesis is accepted.
________________________
Dean of Graduate Studies
THE UNIVERSITY OF NEW BRUNSWICK
December, 2003
© Lottie Vallis, 2004
ABSTRACT
The harbour at Saint John, New Brunswick (SJH) receives wastes from both municipal
and industrial sources. We were interested in determining whether rock gunnel (Pholis
gunnellus) would be a suitable sentinel species for documenting contamination levels
around SJH. The growth and reproductive development of rock gunnel were examined
along three shores of SJH where past records indicate differing levels of contamination.
There were no differences in steroid levels between sites, nor was there a difference in
EROD activity levels between sites, but rock gunnel collected at sites with the highest
contamination were larger, had higher condition factors and larger relative liver sizes.
Liver size differences between sites were not evident in spring when fish returned from
offshore spawning. There was also a decrease in the number of juveniles present at the
most contaminated site. Larger sizes could be an indication of eutrophication due to
sewage inputs at the contaminated sites.
ii
ACKNOWLEDGEMENTS
I would like to express my appreciation to the many people who have contributed to this
project. First, I would like to thank my co-supervisors, Dr. Kelly Munkittrick and Dr.
Deborah MacLatchy for their continued support and helping me see the light at the end of
the tunnel. I would also like to thank all of the past and current members of the
MacLatchy and Munkittrick labs and the Canadian Rivers Institute. Without their help
this project would not have been possible. I would also like to thank students from other
UNB labs for their help and advice as well as my family for assisting me even though
they did not always know what they were doing or why they were doing it. Finally, I
would like to thank Barb Dowding, Rainie Sharpe, and Rebecca Ibey who supported
through the many highs and lows over the past several years.
Thank you to my supervisory committee members, Dr. Matt Litvak and Dr. Simon
Courtenay, for providing prompt and encouraging editing marks; to ACAP Saint John,
especially Sean Brillant, who always had time to answer my questions.
I would also like to thank Peter Perley (Department of Fisheries and Oceans) for aging
my otoliths and Dr. Kevin Halcrow for showing me the art of histology.
Funding for this project was provided through the Canadian Water Network, Canadian
Research Chairs, ACAP Saint John, Science Horizons and a NSERC grant to DR. Kelly
Munkittrick.
iii
TABLE OF CONTENTS
ABSTRACT....................................................................................................................... ii
ACKNOWLEDGEMENTS ............................................................................................ iii
TABLE OF CONTENTS ................................................................................................ iv
LIST OF TABLES ........................................................................................................... vi
LIST OF FIGURES ....................................................................................................... viii
1.0
INTRODUCTION ................................................................................................ 1
1.1
Statement of Problem.......................................................................................... 2
1.2
Study Area - Saint John Harbour ........................................................................ 2
1.3
Previous Studies on Saint John Harbour............................................................. 3
1.4
Environmental monitoring in marine areas......................................................... 4
1.4.1
Bioindicators in pollution assessment......................................................... 5
1.4.2
Whole organism indicators ......................................................................... 8
1.4.3
Biochemical Indicators ............................................................................... 9
1.4.3.1 Mixed-Function Oxygenases ................................................................ 10
1.4.3.2 Reproductive Steroids........................................................................... 13
1.6
Previous Studies................................................................................................ 17
1.7
Organization of Thesis...................................................................................... 19
1.8
Objectives and Hypothesis................................................................................ 19
2.0
MATERIAL and METHODS............................................................................ 20
2.1
Sample Sites...................................................................................................... 20
2.2
Sample Collection............................................................................................. 24
2.3
In Vitro Gonadal Incubation ............................................................................. 25
2.4
Ethoxyresorufin-O- Deethylase Assay Protocol............................................... 26
2.4.1
Homogenate Preparation........................................................................... 26
2.4.2
EROD Activity.......................................................................................... 26
2.4.3
Protein Determination............................................................................... 27
2.5
Age Estimation.................................................................................................. 28
2.6
Statistical Analysis............................................................................................ 28
3.0
RESULTS ............................................................................................................ 28
3.1
Reference Comparison...................................................................................... 29
3.1.1
Late Summer and Prespawning Periods at Saints Rest Beach.................. 29
3.1.2
Seasonal Changes for 2002....................................................................... 35
3.1.3
Size at Maturity......................................................................................... 38
3.1.4
Pre and Post Spawning.............................................................................. 41
3.2
Reference Site Comparisons ............................................................................. 43
3.2.1
Early Summer ........................................................................................... 43
3.2.2
Late Summer............................................................................................. 44
3.2.3
Prespawning.............................................................................................. 44
3.3
In-city comparisons........................................................................................... 46
3.3.1
Late Summer and Prespawning at Little River Marsh and Red Head ...... 47
3.3.1.1 Comparisons within sites ...................................................................... 47
iv
Red Head........................................................................................................... 47
Little River Marsh............................................................................................. 47
3.3.1.2 Comparisons among Sites..................................................................... 49
3.3.2
Seasonal Site Differences within Saint John Harbour in 2002 ................. 50
3.3.2.1 Spring.................................................................................................... 50
3.3.2.2 Late Summer......................................................................................... 52
3.3.2.3 Prespawning.......................................................................................... 53
3.4
Production of Gonadal Sex Steroids ................................................................. 53
3.4.1
Late Summer and Prespawning at Red Head............................................ 53
3.4.2
Reference and In City Comparisons ......................................................... 54
3.4.2.1 Among Sites 2001................................................................................. 54
3.4.2.2 Between Incubation Media 2001 .......................................................... 54
3.4.2.3 Among Sites 2002................................................................................. 55
3.4.2.4 Between Incubation Media 2002 .......................................................... 58
3.4.3
Female Comparisons Between 2001 and 2002......................................... 59
3.5
Induction of Hepatic Detoxification Enzymes (EROD levels)......................... 64
3.5.1
Prespawning at Little River Marsh, Red Head and Saints Rest Beach in
2001........................................................................................................... 64
3.5.2
Pre and Post Spawning.............................................................................. 64
3.5.3
Seasonal Changes in 2002 ........................................................................ 68
3.5.3.1 EROD.................................................................................................... 68
3.6
Age Estimation.................................................................................................. 70
3.7
Size at Age ........................................................................................................ 70
4.0
DISCUSSION ...................................................................................................... 74
4.1
General Biology ................................................................................................ 74
4.2
Comparisons within Saint John Harbour .......................................................... 77
4.3
Comparisons Among Reference Sites .............................................................. 78
4.4
Sex Steroid levels.............................................................................................. 79
4.5
EROD Activity.................................................................................................. 81
4.6
Adequacy of Reference Sites ............................................................................ 84
4.7
Recommendations............................................................................................. 85
4.8
Conclusions....................................................................................................... 86
5.0
References............................................................................................................ 88
6.0
Appendix.............................................................................................................. 97
v
LIST OF TABLES
Table 1. Monthly collections at sampling areas in 2001 and 2002…………..…………22
Table 2. Latitude and longitude coordinates of sampling areas……………………..…23
Table 3a. Regression table of liver size in female Pholis gunnellus in 2001 and
2002……………………………………………………………………….……..………30
Table 3b. Regression table of gonad size in female Pholis gunnellus in 2001 and
2002……………………………………………………………………………...….……31
Table 4a. Regression table of liver size male Pholis gunnellus in 2001 and
2002………………………………………………………………….…………..…….…32
Table 4b. Regression table of gonad size in male Pholis gunnelus in 2001 and
2002……………………………………………………….……………………..…….…33
Table 5. Mean weight, length, liver size, gonad size and condition for male and female
Pholis gunnellus during late summer and prespawning at Saints Rest Beach in 2001 and
2002…………………………………………………………….………………..…….…34
Table 6. Anova table of seasonal differences at Saints Rest Beach in 2001 and
2002………………………………………………………………….……………...……35
Table 7. Anova table of seasonal differences at Saints Rest Beach in 2002……...…….36
Table 8. Mean weight, length, liver size, gonad size and condition of male and female
Pholis gunnellus at Saints Rest Beach during 2002…………...………………………....37
Table 9. Mean weight, length, liver size, gonad size and condition of male and female
Pholis gunnellus pre and post spawning at Saints Rest Beach……..……………………42
Table 10. Anova table of pre and post spawning fish at Saints Rest Beach……..……..43
Table 11. Mean weight, length, liver size, gonad size and condition of male and female
Pholis gunnellus at reference areas in 2002………………………………………...……45
Table 12. Anova table of reference site comparisons in 2002…………………..……...46
Table 13. Mean weight, length, liver size, gonad size and condition of male and female
Pholis gunnellus collected in Saint John Harbour in 2001…………………………..…..48
Table 14. Anova table of within site comparison in Saint John Harbour in 2001…..….49
Table 15. Anova table of between site comparisons in Saint John Harbour in 2001..…50
vi
Table 16. Seasonal changes in weight, length, liver size, gonad size and condition of
male and female Pholis gunnellus in Saint John Harbour during 2002……………...…..51
Table 17. Anova table of seasonal site differences in Saint John Harbour during
2002…………………………………………………………………………………...….52
Table 18. Anova table of steroid differences among sites in 2001………………..……54
.
Table 19. Anova table of steroid differences between incubation media in 2001…..….54
Table 20. Anova table of steroid differences among sites in 2002………………..……55
Table 21. Anova table of steroid differences between incubation media in 2002..…….59
Table 22. Anova table of steroid differences between 2001 and 2002……..…………..60
Table 23. Total in vitro estradiol levels of female Pholis gunnellus collected in Saint
John Harbour…………………………………………………………….……..………...63
Table 24. Total in vitro testosterone levels of female Pholis gunnellus collected in Saint
John Harbour…………………………………………………………….………...……..63
.
Table 25. Anova table of pre and post spawning EROD activity………….…….……..64
Table 26. Anova table of seasonal EROD activity during 2002…………….….………68
Table 27. Mean seasonal EROD activity of male and female Pholis gunnellus during
2002……………………………………………………………………………..….…….69
Table 28. Anova table of age estimation among sites…………………………...……...70
Table 29. Anova table of size at age………………………………………….…..…….70
.
vii
LIST OF FIGURES
Figure 1. Map of area and study locations in and around Saint John Harbour, New
Brunswick, Canada…………………………………………………………..…………..23
Figure 2. Gonadal development of female Pholis gunnellus collected at Saints Rest
Beach during 2001 and 2002………………………………………………..…………...39
Figure 3. Gonadal development of male Pholis gunnellus collected at Saints Rest Beach
during 2001 and 2002…………………………………………….……………………...40
Figure 4. Total in vitro testosterone and estradiol production in female Pholis gunnellus
incubated with 25-OH in 2002 over 24 hours. Each value represents the mean + SEM.
Values sharing an alphabetical script within each steroid are not statistically different
(p>0.05)………………………………………………….……………………………….56
Figure 5. Total in vitro testosterone and 11-ketotestosterone production in male Pholis
gunnellus incuated with 25-OH in 2002 over 24 hours. Each value represents the mean +
SEM. Values sharing an alphabetical script within each steroid are not statistically
different (p>0.05)………………………………………...………………………………57
Figure 6. Total in vitro estradiol production in female Pholis gunnellus incubated in
forskolin and 25-OH in 2002 over 24 hours. Each value represents the mean + SEM.
Values sharing an alphabetical script within each media are not statistically different
(p>0.05)……………………………………………………………………...…………...61
Figure 7. Total in vitro testosterone production in female Pholis gunnellus incubated in
forskolin and 25-OH in 2002 over 24 hours. Each value represents the mean + SEM.
Values sharing an alphabetical script within each media are not statistically different
(p>0.05)…………………………………………………………………………………..62
Figure 8. EROD activity in male and female Pholis gunnellus in 2001. Each value
represents the mean + SEM. Values sharing an alphabetical script within each sex are not
statistically different (p>0.05)……………………………………..……………………..65
Figure 9. EROD activity of female Pholis gunnellus pre and post spawning. Each value
represents the mean + SEM. Values sharing an alphabetical script within each site are
not statistically different (p>0.05)……………………………..………….……………..66
Figure 10. EROD activity of male Pholis gunnellus pre and post spawning. Each value
represents the mean + SEM. Values sharing an alphabetical script pre and post spawning
between sites are not statistically different (p>0.05)………………..……..…………….67
Figure 11. Age of Pholis gunnellus in and around Saint John Harbour. Each value
represents the mean + SEM. Values sharing an alphabetical script are not statistically
different (p>0.05)………………………………………………………...………………71
viii
Figure 12. Total length at age of Pholis gunnellus in and around Saint John
Harbour………………………………………..…………………………………………72
Figure 13. Body weight at age of Pholis gunnellus in and around Saint John
Harbour………………………………………..…………………………………………73
.
ix
1.0 INTRODUCTION
In the early 1990’s, a series of amendments to the Pulp and Paper Effluent Regulations
under the Fisheries Act were developed to improve the quality of effluents from
Canadian pulp and paper mills. The amendments included a provision for environmental
effects monitoring (EEM) at pulp and paper mills (Munkittrick et al., 2002). The
objective of the EEM program is to evaluate the effects of effluent on fish, fish habitat
and the use of fisheries resources (Walker et al., 2002). The program was designed as a
series of monitoring cycles where the requirements of one cycle are dependent on the
findings in the previous cycle (McMaster et al., 2002).
The original study designs were developed in riverine situations, and have worked well in
freshwater situations (Munkittrick et al., 2002). They have not been as successful when
applied to marine or estuarine situations (Courtenay et al., 2002). In marine and
estuarine systems, there are difficulties related to tidal fluctuations, trapped plumes, and
uncertain exposure (Munkittrick et al., 2002). While there may be species that are less
mobile in marine systems (Larsson et al., 2002), the traditional species of interest in
marine systems are relatively mobile and may not utilize a single area of the estuary for
all of their life cycle.
A cumulative effects assessment was initiated in the upper reaches of Saint John River in
1999 (Gray et al., 2002; Galloway et al., 2003). The studies are progressing downstream
and will eventually focus on trying to understand issues affecting the estuary. The Saint
John River estuary is a very complex environment, receiving wastes from a variety of
industrial and municipal sources and having mean daily tidal fluctuations of 6.7 m. The
purpose of this project was to initiate a monitoring program in the estuary to evaluate
study designs and potential sentinel species.
1.1
Statement of Problem
Although a variety of pollutants are released into estuarine and marine waters, most
effects monitoring programs that have used fish have focused on freshwater species
(Oberdorster and Cheek, 2000). Studies in marine and estuarine environments face a
number of challenges, including a lack of fish, high dilution of effluents, trapped plumes,
migratory sentinel species, confounding discharges, shipping, dredging and tidal
fluctuations (FSEWG, 1997). As a result, demonstrating the same degree of exposure in
marine species is typically more difficult than in freshwater species. Some of these
difficulties are magnified because marine systems are typically more open and
challenging to sample, and because of the non-homogenous distribution of habitat and
fish within the habitat (Courtenay et al., 2002).
1.2
Study Area - Saint John Harbour
Saint John Harbour is located at the mouth of the Saint John River estuary on the Bay of
Fundy, New Brunswick, Canada (latitude 45º 16’N, longitude 66º 3’ W). For many
years, a wide variety of industrial, commercial and residential waste streams have been
discharged into the harbour. Some of the present industries in the area include three
industrial parks, a pulp mill, tissue plant, paper mill, an oil refinery, New Brunswick
Power generating station, brewery and the Saint John Port Authority which exports and
2
receives cargo from around the world. In addition, Saint John residents also have an
impact on the harbour. The population of the Saint John is approximately 70,000 and
currently just over 50% of the municipal sewage is treated (Statistics Canada, 2003). In
2001, for example, 8.1 million m3 liters of untreated sewage flowed directly into the Saint
John Harbour and surrounding area.
1.3
Previous Studies on Saint John Harbour
Past studies in the region have focused on measuring the contaminants in harbour water
columns, sediments and animal tissue. Monitoring of contaminants within harbour water
and sediment indicate the presence of alcohols, fatty and resin acids, metals, petroleum
residues, phenolics, polychlorinated biphenyls (PCBs) and polycyclic aromatic
hydrocarbons (PAHs) hydrocarbons (Godfrey Associated Ltd. et al., 1993; Wasburn and
Gillis, 1993; Brillant, 1999). In addition, both blue mussel (Mytilus edulis) and winter
flounder (Pseudopleuronectes americanus) tissue samples indicated heavy metal
contamination and increased levels of hepatic mixed function oxygenase activity
(Wasburn and Gillis, 1993). In the 1993 Saint John Harbour Environmental Quality
Study, however, biota were only briefly addressed. Brillant (1999) revealed that Saint
John Harbour had levels of copper, lead, zinc and PCB in bioindicator invertebrates
significantly correlated with locations receiving elevated municipal sewage discharges.
Many of the contaminants identified in the Harbor are suspected to be endocrine
disrupting substances (EDSs), however, only a few studies have examined the effects of
contamination on fish and invertebrates in Saint John Harbour. High organochlorine,
PCB and DDT levels in gonadal tissue were believed to be potentially responsible for
3
decreased reproductive success of adult striped bass (Morone saxatilis) populations in the
Saint John River (Dadswell, 1975). Prouse and Ellis (1997) have demonstrated the
masculinization of resident dogwhelks (Nucella lapillus) as a result of the
bioaccumulation of tributylin (TBT), a compound used in antifouling paints. LeBlanc
(1998) demonstrated the potential for reproductive dysfunction in wild fish living in Saint
John Harbour. Both male and female mummichog (Fundulus heteroclitus) exposed to
Saint John Harbour waters displayed significant reductions in plasma steroid levels and
gonadal steroid production, which LeBlanc (1998) suggested were a result of xenobiotics
present in the harbour. It is unknown whether other species show reproductive problems
associated with exposure to contaminants in the harbour, but potential reproductive issues
were associated with previous discharges from the pulp mill located on the estuary
(Dubé, 2000).
1.4
Environmental monitoring in marine areas
There are a variety of challenges associated with monitoring fish in marine environments,
including issues of residency, reference sites, exposure estimation, and the life histories
of marine species (FSEWG, 1997). The priority factors to be considered for choosing a
sentinel species include that they represent the local environment, are abundant, are
exposed to the stressors of concern, and allow endpoints of interest to be measured
(FSEWG, 1997).
Many marine species use very different habitats in different seasons and are not locally
resident for long periods of time. Measures of exposure may only give an estimate of
4
recent history (Environment Canada, 1997). Measuring effluent concentrations can be
difficult in marine situations because water movement can be bi-directional, effluent
concentrations can change rapidly, mapping of effluent rarely takes account of variability
resulting from lunar cycles and seasonal variability of river discharge, interaction(s) of
salt and fresh water may cause effluents to float or be trapped at depth, and processes
such as particle transport may result in concentrated areas of contaminants depositing in
zones some distance from the discharge point (Courtenay et al., 2002).
The absence of life history information on many species of fish often prevents adequate
design or interpretation of field studies. The normal range of values for a particular
population of fish and the factors that influence these values are often not known
(Bucheli and Fent, 1995). Without these baseline measurements it is difficult to
determine variability within a population and therefore provide a comparison for
exposure studies.
1.4.1
Bioindicators in pollution assessment
Fish populations can become stressed as a result of direct contaminant impacts, natural
environmental stressors, or a combination of natural and anthropogenic perturbations
(Adams, 1992). Streams, rivers, lakes, and estuaries have been subjected to increasing
anthropogenic stress associated with industrial, agricultural, and residential xenobiotics
over the past several decades (Adams and Greeley, 2000). The potential of these
xenobiotics to affect aquatic habitats and animal populations has been of concern (Kime,
1995). Contaminants pose a significant danger to water quality and to the health of
5
aquatic systems not only because of the varied types of pollutants that impact these
systems, but also due to the many pathways by which pollutants can affect the health of
aquatic systems (Adams and Greeley, 2000). It is difficult to establish causal or
correlative relationships between environmental contaminants and organism health
primarily due to a multitude of environmental or ecological factors that likely influence
the response of organisms to particular stressors (Adams et al., 1999).
A variety of approaches have been used to evaluate the effects of environmental stress on
fish. While manipulative approaches such as laboratory studies provide information
about mechanistic links between a particular toxicant and a specific response, most
controlled studies provide limited information regarding how organisms respond to
environmental stressors in their natural environment (Cairns, 1981). As a consequence,
extrapolation of results from laboratory tests to the natural environment may lead to
inaccurate predictions of stress effects on fish (Adams, 1990).
Field approaches are required to provide an integrated framework for addressing
cumulative and/or synergistic impacts on aquatic systems (Adams and Greeley, 2000).
One approach to addressing this problem is through the use of biological indicators.
Bioindicators have been defined as variations in biochemical or physiological
components or functions that have either been correlated or linked to biological effects
(McCarty and Munkittrick, 1996). Bioindicators can provide evidence of exposure to
compounds that are not measurable, do not bioaccumulate, or, are rapidly metabolized
and eliminated. Furthermore, bioindicators integrate the interaction resulting from
6
exposure to complex mixtures of contaminants and present a biologically relevant
measure of toxicant action at target tissues and cumulative adverse effect of the exposure
(McCarthy et al., 1990).
The bioindicator approach uses sentinel species and specific physiological responses of
these organisms at several levels of organization to assess the effect of a particular
stressor response on exposed organisms (Adams and Greeley, 2000). Bioindicators can
be used to provide a signal of damage at the population and community level and to
construct hypotheses concerning the dependence of the observed biological effects upon
the toxic variable under investigation (Adams, 1990; Adams et al., 1999). By measuring
both rapidly responding exposure biomarkers such as biochemical responses and slower
responding bioindicators such as population responses, a range of sensitivities and a
variety of specific effects can be monitored simultaneously (Adams, 1990).
An ideal assessment program includes indicators at various levels of biological
organization, but there have been few studies aimed at an integrated evaluation of
environmental health by means of a concomitant use of biomarkers at different levels of
biological organization (Bucheli and Fent, 1995). Further, biological validation is
required before indicators can be reliably used to detect environmental deterioration. The
greatest limitation of most indicators is that the biological significance for the integrity
and survival of individuals and populations is unclear.
7
The underlying assumption of the bioindicator approach is that manifestations of stress at
each of the lower levels of organization will be evident before disturbances at the
population, community, or ecosystem level are realized (Adams, 1990). Such initial
effects are observed at the molecular level primarily with the induction of cellular
defense systems. However, if a stressor continues to be of sufficient duration and
magnitude, effects at lower levels will eventually be manifested at increasingly higher
levels of biological organization (Adams and Greeley, 2000). When these systems fail,
higher-level damage may occur, potentially causing histological or physiological
impairment. If these systems are permanently altered during vulnerable periods of
development, then reproduction and/or survival may eventually be affected. This might
lead to changes at the population and possibly community levels of organization
(Schlenk, 1999).
No single indicator or group of indicators can provide all the critical information needed
to assess long-term stress on fish (Mayer et al., 1992). The ideal assessment design of a
polluted system, then, would involve measurement of selected indicators for each major
level of organization so that causal relationships between them can be determined. Once
such relationships are recognized, predictions of long-term effects from measurements of
short-term indicators can be made with increased dependability and certainty (Adams,
1990).
1.4.2
Whole organism indicators
Condition indices can be used to relate the consequences of biochemical and
physiological alterations to observed changes in individual fish and the population.
8
The condition factor is a generalized indicator of overall fitness and reflects the
integrated effect of (1) nutrition and (2) metabolic cost (Adams et al., 1990). Ratios of
organ weight to body weight have been used in various studies. Two ratios frequently
used are liver weight or gonad weight to body weight, or the liver-somatic index (LSI)
and gonadosomatic index (GSI). The LSI reflects short-term nutritional status and
metabolic demands (Everaats et al. 1993). The LSI is also sensitive to toxicant stress,
with liver enlargement being reported in fish exposed to toxic substances (Adams et al.
1990). The simplest measure of gonadal dysfunction is the GSI. GSI is utilized as a
morphological indication of sexual status and maturation in fish populations. Decreased
GSI can be indicative of decreased hypothalamic, pituitary or gonadal activity (Kime,
1999). While these measures are prone to the effects of nonpollutant factors, they may
serve as a first tier indicator of effect and possibly provide information on energy
reserves (Adams et al., 1990).
1.4.3
Biochemical Indicators
Environmental contaminants are known to induce measurable physiological changes in
exposed aquatic organisms (Eufemia et al., 1997). Because many signs of exposure
initially occur at the subcellular level, there has been an increased use of biochemical
alterations as indicators of chemical exposure (Schlenk et al., 1996). Studies on the
biological effects of contaminants at molecular and cellular levels can provide rapid and
sensitive indicators of stress and have been recommended in monitoring water pollution
(Livingstone, 1993). A major advantage of specific biochemical indicators is that they
can often indicate the nature of the environmental stressor. Consequently, these
9
biological indicators can be useful for detecting changes in environmental levels of
certain classes of pollutants (Goldfarb et al., 1998).
The potential value of reproductive variables for the evaluation of the effects of stressors
on fish has been recognized for many years. Reproduction in aquatic animals is
potentially the most critical function affected by chronic toxicant stress (Birge et al.,
1985). It is clear that reproduction is essential to the continued existence of a population;
thus, it follows that any stressor that interferes with reproduction at the individual level
can potentially affect the survival of the population.
1.4.3.1 Mixed-Function Oxygenases
Many inducible enzymes have been used as biochemical indicators of exposure in aquatic
environments (Schlenk et al., 1996). Induction of the mixed function oxygenase system
(MFO), which includes cytochrome P450-dependent monooxygenase and conjugating
enzymes, has been widely used to determine contaminant exposure in marine and fresh
water aquatic ecosystems (Ciccotelli et al., 1998). The hepatic cytochrome P450
monooxygenase system is important for the catabolism and transformation of steroid
hormones and other endogenous compounds such as vitamins, but also plays a critical
role in the detoxification of organic contaminants. A series of oxidation reactions are
carried out whereby relatively insoluble organic compounds are converted into watersoluble metabolites that may be further conjugated and excreted (Jimenez and Stegeman,
1990). Among the monooxygenases, 7- ethoxyresorufin- o -deethylase (EROD) activity
is particularly responsive to contaminants (Fammarion, 1997). Induction of MFO is
10
therefore commonly associated with an increase in EROD levels. Mixed function
oxygenase induction, in contrast to other biochemical responses, is a primary
detoxification response. Thus induction can serve as an integrator of response for many
classes of organic pollutants in aquatic environments because this response is rapid
(Williams et al., 1997).
Increased hepatic EROD activity levels have been shown to be a useful tool for indicating
the extent of exposure to pulp and paper mill effluent (PPME) and complex mixtures of
planar, organic chemicals, specifically congeners of PAHs and PCBs (Munkittrick et al.,
1992; Eggens et al., 1995; Bucheli and Fent, 1995). Induction is initiated when aromatic
compounds bind to an aryl hydrocarbon (Ah) receptor. This triggers expression of the
gene coding for the P450 enzyme messenger RNA (mRNA) and thus enzyme synthesis is
increased. The ability of a compound to induce MFO activity, however, appears to be
related to the molecular shape or the co-planarity of connected aromatic rings and the
distribution of chlorine atoms which determines the ability of a particular congener to fit
into the Ah receptor (Whyte et al., 2000).
Studies demonstrating increases in cytochrome P450 activity have also documented
changes in fish performance, including altered steroid hormone profiles, and impairment
of the reproductive and immune systems. To date, however, there has been no
demonstration of causal links between altered cytochrome P450 activity and other
responses in fish (Hodson et al., 1991; Munkittrick et al., 1992; Karels et al., 1999).
11
Payne and Penrose (1975) and Payne (1976) were among the first to use this enzyme as a
biomarker, reporting elevated P450 activity in fish from a petroleum-contaminated lake
in Newfoundland. Since then, extensive research has supported the use of the
cytochrome P450 system in different fish species as a monitoring tool in field and
laboratory studies.
Numerous studies have shown significant effects of PPME on individual fish health with
one of the most consistent effects being the activation of the liver detoxification system
(Karels et al., 1998). Previous studies in North America and Scandinavia have reported
induced EROD activities in fish collected downstream of pulp and paper mills when
compared with fish from reference sites (Munkittrick et al., 1992; Karels et al., 1998).
The measurement of EROD activity has also been shown to be a useful tool for
monitoring areas consisting of complex mixtures of organic chemicals (Fenet et al., 1998;
Kirby et al., 1999). In fish, PCBs have a strong bioconcentration capacity and are readily
absorbed during exposure to contaminated food, water, and sediment (Pacheco and
Santos, 1998).
In several flatfish species, for example, strong correlations of hepatic
EROD activity with pollution gradients related to riverine inputs of organic chemicals
have been described (Sulaiman et al., 1991; Eggens et al., 1995).
The various studies which have used monooxygenase activity as a bioindicator of
pollution in fresh and saltwater environments, as well as the wide range of species tested,
confirm the value of the monooxygenase system as a monitoring tool for point and nonpoint sources of pollution.
12
1.4.3.2 Reproductive Steroids
Fish maturation and reproduction are complex processes that are regulated by
endogenous hormones along the hypothalamic-pituitary-gonadal axis and are
synchronized by exogenous factors (Arukwe and Goksoyr, 1998). Environmental cues,
such as temperature and photoperiod, are integrated within the brain to stimulate the
hypothalamus to produce gonadotropin releasing hormone (GnRH), which promotes the
production and release of gonadotropins (GtH) from the pituitary gland. GtHs further
stimulate the development and maturation of gonads as well as the production of steroids.
The major reproductive steroid hormones are estrogen and testosterone in females and
11-ketotestosterone and testosterone in males. These steroids regulate metabolic
processes, the promotion of secondary sex characteristics and reproductive behaviour
(Redding and Patino, 1993).
Over recent years, a wide range of synthetic and natural substances used for industrial,
agricultural and municipal activities have been shown to disrupt hormonal systems of
fish. One important class of endocrine disruptors acts by binding to the estrogen
receptor. These estrogenic substances include natural steroidogenic substances as well
as, 17α-ethinylestradiol, the synthetic steroid commonly used in the contraceptive pill
(Hashimoto et al., 2000). Other hormone active compounds include organochlorine
pesticides, industrial pollutants such as PCBs, detergents, paints, and surfactants.
Furthermore, natural chemicals such as phytoestrogens commonly found in PPME are
also capable of binding to this receptor. Changes in circulating levels of reproductive
steroids in fish exposed to a number of these environmental contaminants have been used
13
as an indicator of potential reproductive impairment (Munkittrick et al., 1992; Kime,
1995).
Several studies have shown that adverse reproductive effects observed in fish are linked
to the presence of hormone analogs present in the environment. Field studies using a
number of species have demonstrated that reductions in circulating levels of sex steroids
reliably indicate exposure to compounds impacting the reproductive system. Recently,
studies have included the measurement of sex steroid levels in an attempt to correlate
reproductive problems with exposure to inorganic and organic compounds, sewage
effluent and PPME (Collier et al., 1998; Routledge et al., 1998; Escher et al., 1999).
Some Scandinavian and Canadian studies of fish captures near sources of PPME have
shown dramatic impacts on reproduction when compared with fish from reference sites.
Studies conducted in the early 1990’s on white sucker (Catostomus commersoni) at
Jackfish Bay, Lake Superior have shown that fish exhibit an array of altered responses
including altered plasma and gonadal steroid levels, reductions in gonadal size, delayed
sexual maturation, reduced fecundity with age in females, and a reduction in male
secondary sexual characteristics (McMaster et al., 1991; Munkittrick et al., 1992; Van
Der Kraak et al., 1992). Van Der Kraak et al. (1992) assessed the underlying cause of
reproductive dysfunction and reduced circulating steroid levels in fish exposed to PPME
and demonstrated that multiple sites within the pituitary-gonadal axis were affected. This
included reduced secretion of GtH and depressed ovarian steroid synthesis. Additional
studies indicate that the levels of in vitro steroid production paralleled differences in
14
circulating hormone levels suggesting that impacts on ovarian steroid biosynthetic
capacity represent a major site of PPME actions (Munkittrick et al., 1992).
A close correlation between reduced circulating levels of gonadal steroids and elevated
P450 activities in fish exposed to PPME raised interest in the relationship between
hormone levels and P450 activities. As a result, a number of studies have investigated
the impact of hormonal exposure on steroid metabolism and P450 expression in fish.
McMaster et al. (1991) demonstrated there was no relationship between MFO activity
and steroid hormone levels in white sucker (Catostomus commersoni). Steroid changes
were still evident in white sucker during spawning at an uncontaminated site, while MFO
levels were low. Likewise, Munkittrick et al. (1992) demonstrated that during a mill
operational shutdown the persistence of steroid reductions in female white sucker was
coupled with reduced MFO activity; this suggests that the two indicators are not directly
related. In addition, EROD activity could not be related to indices of reproduction in
yellow perch (Perca flavescens) exposed to oil sands-related compounds (van den Heuvel
et al., 1999).
Studies have also indicated that industrial and domestic sewage effluents interfere with
hormonal processes and maturation. Industrial effluents and sewage treatment water
contain a large number of substances that have been found to interact with the endocrine
system. The major estrogenic compounds of sewage treatment works effluent include
natural compounds, 17β-estradiol and estrone, and the synthetic 17α-ethinylestradiol
(Rodgers-Gray et al., 2000). Sewage effluents also contain many chemicals other than
15
steroid estrogens with estrogenic activity, several of which are highly lipophilic and able
to bioaccumulate in organisms. Substances with weaker effects include PCBs, PAHs and
alkylphenols, the breakdown products of industrial surfactants (Bjerselius et al., 2001).
One of the first examples of the effects of estrogenic chemicals in the aquatic
environment was found in the UK in the early 1990’s. Anglers reported the presence of
hermaphrodite roach (Rutilus rutilus) in lagoons of sewage treatment wastewater (STW).
Subsequent studies using caged rainbow trout (Onchorynchus mykiss) have demonstrated
that substances in the final effluent of domestic STWs are estrogenic (Harries et al.,
1996; Purdom et al., 1994). An extensive study of wild roach in U.K. rivers found
widespread feminization of males correlated with the proximity of these fish to estrogenic
treated sewage effluent discharges (Jobling et al., 1998). Johnson et al. (1988) and
Casillas et al. (1991) have reported effects on ovarian development in English sole
(Parophrys vetulus) from Puget Sound. Female fish collected from sites heavily
contaminated with PCBs and PAHs had decreased levels of circulating estradiol levels
and were significantly less likely to undergo gonadal recrudescence than females from
less contaminated sites within Puget Sound. Recently, Collier et al. (1998) have also
reported premature juvenile sexual maturation and inhibited gonadal development in
female flatfish in Hylebos Waterway, an area known to be severely contaminated by
organic and inorganic contaminants.
1.5 Species Selection
16
The use of fish as indicators of pollution effects can permit early detection of
environmental problems (Powers, 1989). Fish are frequently used as environmental
monitors because dysfunction in fish and lower vertebrates may be used to make
inferences of potential biological risk to higher vertebrates (Colborn et al., 1993). They
are highly sensitive to many human disturbances, they can be used in studies at all levels
of biological organization and they can be used to integrate ranges of ecological
processes (Harris, 1995).
Benthic fish from urbanized and industrialized rivers and coastal regions are commonly
exposed to elevated levels of contaminants. Benthic species in immediate contact with
sediment are thought to provide a more direct assessment of sediment than pelagic
species. The rock gunnel (Pholis gunnelus) is an intertidal teleost species endemic to the
eastern and western Atlantic shores. It lives in close contact with the sediment in the
intertidal zone and inshore waters and feeds on benthic mollusks and worms. In addition,
prespawning fish have a low mobility and can be easily collected under rocks at low tide.
These combined characteristics suggest the rock gunnel may be a useful species in
pollution studies (Qasim, 1957; Bombail et al., 2001).
1.6
Previous Studies
The basic biology of the rock gunnel has been described (Qasim, 1956; Qasim, 1957;
Sawyer, 1967) but very little is known of its physiology, ecology and behaviour even
though it is a common component of shallow water fish faunas. To date the majority of
17
studies assessing the rock gunnel have occurred on the eastern side of the Atlantic,
mainly in the United Kingdom.
Pholis gunnellus is widely distributed on both sides of the Atlantic. They grow to 25 cm
total length and have been reported to live up to 14 years (Proudfoot, 1975). The rock
gunnel frequently inhabits tide pools and intertidal areas where it hides under rocks, in
crevices and under seaweed. Its diet consists mainly of polychaetes, amphipods and
crustaceans (Scott and Scott, 1988). Spawning occurs during winter in offshore waters.
Sawyer (1967) noted that female rock gunnel leaving intertidal areas during November
were ripe, while fish returning in March were spent.
Several studies have looked at the distribution and movement of Pholis gunnellus. Kruuk
et al. (1988) investigated fluctuations and activity of inshore fishes in Shetland while
Koop and Gibson (1991) examined the distribution and movement of rock gunnel on the
west coast of Scotland. In the experimental areas, the density of rock gunnel increased
with distance from shore while size of the fish did not change (Koop and Gibson, 1991).
Kruuk et al. (1988) found similar densities of fish but did not mention their intertidal
distribution. When areas were depopulated of fish rock gunnel moved into these areas
within two tidal cycles, although the numbers of repopulating fish were only half of the
original (Koop and Gibson, 1991). In a similar experiment Kruuk et al. (1988) found that
rock gunnel could completely repopulate an area with the same number of fish within two
tidal cycles. Disturbance played an important role in determining the degree of
18
movement, as there was a decline in numbers in regularly visited areas (Koop and
Gibson, 1991).
Fewer studies have used Pholis gunnellus in ecological assessments. Bombail et al.
(2001) applied two indicators of genotoxic effects, nuclear anomaly frequency and the
comet assay, to Pholis gunnellus sampled along a pollution gradient in the Firth of Forth,
Scotland. Results indicate that while the analysis of DNA strand breakage using the
comet assay did not reveal any significant differences, the measurement of micronuclei
and nuclear anomaly frequencies were elevated in the contaminated portion compared to
the reference area. This suggests that while the measurement of micronuclei and nuclear
anomaly frequencies may be an appropriate means of detecting exposure to
environmental genotoxins, the comet assay may not be a suitable bioindicator in this fish.
1.7
Organization of Thesis
This thesis has been written in traditional thesis format. Sections of the thesis detail the
introduction, methodology, results and discussion.
1.8
Objectives and Hypothesis
Despite the chemical compounds and various effluents to which animals in the harbour
are exposed, few studies have investigated the effects of chemical contaminants on fish
living in Saint John Harbour. The present study will assess if contaminants discharged
into the harbour affect the status of rock gunnel by examining: 1) whether rock gunnel in
Saint John Harbour show whole animal responses in growth and reproductive
19
development between sites; 2) whether fish in Saint John Harbour show biochemical
differences. Specifically, the null hypothesis (Ho) is that rock gunnel in Saint John
Harbour will not show differences between sites while the alternative (Ha) is that rock
gunnel in SJH will show differences between sites.
2.0 MATERIAL and METHODS
2.1
Sample Sites
Saint John Harbour is located at the mouth of the Saint John River estuary on the Bay of
Fundy in New Brunswick, Canada (45o 15ٰ
N, 66o 03ٰ
W). During the 2001 field season,
fish were collected from August to October at five locations in and around Saint John
Harbour (Table 1). These sites represent different levels of exposure ranging from direct
exposure to sewage wastewater, to sites with no direct inputs that are therefore used as
reference sites. The five harbour sites include: 1. Red Head (RH); 2. Little River Marsh
(LRM); 3. Saints Rest Beach (SRB); 4. Duck Cove (DC); and 5; Bay Shore (BS). Duck
Cove and Bay Shore were excluded in the 2002 field season due to low adult fish
densities. During April to October of 2002, fish were collected at three locations within
the harbour, Little River Marsh, Red Head, Saints Rest Beach and two reference locations
outside harbour limits, Garnett Settlement (GS) and Maces Bay (MB) (Table 2, Figure 1).
Saints Rest Beach was chosen as the in-city reference site because previous studies of the
harbour have indicated lower contamination levels on the west side of the city (Washburn
and Gillis 1983, Brillant 1999). In addition to the presence of ideal habitat for rock
20
gunnel, there are no direct chemical inputs into the area and contamination levels should
therefore represent background harbour levels.
Little River Marsh and Red Head were chosen as exposed areas within the city. An
average of 4000 cubic meters of secondary-treated sewage flows directly into the Bay of
Fundy at Little River Marsh (Godfrey Associates 1993). To date however, plume
delineation or a chemical analysis of the sewage effluent has not been completed.
Studies have also indicated invertebrate exposure to tributylin, an antifouling compound
in marine paint at Red Head. Prouse and Ellis (1997) and Delaney (2001) demonstrated
that female Nucella lapillus at Red Head developed non-functional male reproductive
organs. Prouse and Ellis (1997) reported the presence of a penis and vas deferens while
Delaney (2001) reported the alterations in the pallial oviduct inhibiting reproduction in
female Nucella lapillus.
21
Table 1: Monthly collections at sampling areas in 2001 and 2002
Year
2001
Year
2002
Site
Month
August
September
October
November
Saints Rest
Juveniles Adults
20
0
11
22
4
19
0
6
Red Head
Juveniles Adults
4
0
2
13
11
17
0
0
Little River Marsh
Juveniles Adults
0
0
1
38
5
16
2
5
Duck Cove
Juveniles Adults
14
0
7
2
2
11
0
0
Bay Shore
Juveniles Adults
0
0
26
6
0
4
0
0
Site
Month
April
May
June
July
August
September
October
Saints Rest
Juveniles Adults
0
4
8
28
0
0
7
14
9
15
3
14
5
37
Red Head
Juveniles Adults
0
0
14
50
0
0
0
0
24
12
27
63
0
0
Little River Marsh
Juveniles Adults
0
0
9
38
0
0
0
0
0
0
14
35
0
0
Garnett
Settlement
Juveniles Adults
0
0
0
0
0
0
0
34
0
0
3
25
0
0
Maces Bay
Juveniles Adults
0
0
0
0
3
15
0
7
0
0
0
0
1
27
22
Table 2: Latitude and longitude coordinates of
sampling areas.
Site
Latitude
N 45.226
N 45.242o
N 45.260o
N 45.293o
N 45.246o
N 45.258o
N 45.122o
W 66.118o
W 66.000o
W 66.017o
W 66.091o
W 66.076o
W 65.762o
W 66.475o
Sa
in
t
Jo
hn
R
iv
er
Saints Rest Beach
Red Head
Little River Marsh
Duck Cove
Bay Shore
Garnett Settlement
Maces Bay
Longitude
o
N
Little River Marsh
Duck Cove
Saints Rest
Garnett Settlement
Red Head
Bay Shore
Maces Bay
___
10.8
Figure 1: Map of area and study locations in and around Saint John Harbour, New
Brunswick, Canada
23
Previous studies have also reported contamination in areas near Little River Marsh and
Red Head. Ray and MacKnight (1984) reported portions of the outer harbour, including
the Bay and the mudflats outside the Courtenay Bay breakwater, had elevated
concentrations of metals attributed to industrial sources onshore and annual dredging of
the area. Environmental Consultants (2001) further noted elevated metal concentrations
in the inner harbour. While metal concentrations are elevated against background
concentrations, they are below the probable effects levels of the Canadian Council of
Ministers of the Environment (CCME) marine sediment quality guidelines. In addition,
Courtney Bay and Liver River which empty into Saint John Harbour receive untreated
sewage with Little River also receiving discharges from an oil refinery and paper mill
(Washburn and Gillis 1983, Godfrey Associates 1993).
2.2
Sample Collection
Rock gunnel were collected with hand nets by manually turning rocks for
approximately two hours before and after low tide along rocky portions of the intertidal
zone. Once collected, fish were placed in 8-L plastic containers filled with seawater,
seaweed and rocks. Prior to being transported back to the laboratory, fish were placed
in an aerated cooler to ensure adequate oxygenation.
At the laboratory, fish were anaesthetized by immersion in 0.05% tricaine
methanesulfonate (TMS, Syndel Labs, Vancouver, BC, Canada) and killed by spinal
severance. Each adult rock gunnel was measured for total length (1mm), body weight
(0.01g), gonad weight (0.01g) and liver weight (0.01g). Gonadosomatic index (GSI)
and liver somatic index (LSI) were calculated as tissue weight/ total body weight *100
24
while condition factor was calculated as ((total body weight/total body length)3) * 100.
Aging structures (otoliths) were collected from frozen specimens and analysed by
counting annuli. Fish larger than 90 mm, samples were frozen for biochemical analysis
During the late fall and prespawning of 2001, and during 2002, samples were collected
for hepatic MFO analysis and sex steroid production. For MFO analysis, each liver was
placed in a cryovial, frozen immediately in liquid nitrogen, and stored at –80ºC pending
analysis. For in vitro sex steroid production measurement, gonad tissue was removed
and placed in excess Medium 199 buffer solution (Medium 199 containing Hanks Salt
without bicarbonate; Sigma Chemical Co., St. Louis, MO, USA) for the subsequent
measurement of gonadal sex steroids.
2.3
In Vitro Gonadal Incubation
For gonadal steroids, in vitro production of testosterone (T) and 11-ketotestosterone
(11KT) was measured in male fish and 17 β-estradiol (E2) and testosterone in female
fish. Following dissection, gonadal tissue was placed in 1 mL of Medium 199 buffer.
Two pieces of either ovarian or testis tissue, weighing a total of 18-25 mg, were
incubated in polystyrene tissue culture plates (Fisher Scientific Co., Toronto, ON, CA)
at 18ºC for 24 hours in 1 mL of Medium 199 buffer solution (McMaster et al., 1995).
Immediately prior to the incubation period, the medium was replaced with fresh
Medium 199. To stimulate in vitro steroid production, 10 µM forskolin (Sigma-Aldrich
Canada Ltd., Oakville, ON, CA) or 10µg 25-hydroxycholesterol (25-OH) (SigmaAldrich Canada Ltd., Oakville, ON, CA) were added to the incubations (McMaster et
25
al., 1995). Forskolin and 25-OH were dissolved in ethanol and added to the tissue in
5µL aliquots to avoid ethanol toxicity. After incubation, the media was collected and
stored at –20ºC for later radioimmunoassay (RIA) analysis.
2.4
Ethoxyresorufin-O- Deethylase Assay Protocol
Measurement of MFO activity was based on the catabolism of 7-ethoxyresorufin-odeethylase (EROD) ( Sigma Aldrich Canada Ltd., Oakville, ON, CA) as described by
Efler et al. (1998). Due to the small size of the livers, measurements of MFO activity
were assayed using the whole homogenate in an effort to ensure that all microsomes
were present in the sample.
2.4.1
Homogenate Preparation
Liver samples were thawed on ice from – 80oC storage to facilitate homogenization.
Homogenates were prepared in ice-cold Hepes Grinding Buffer (Sigma Aldrich,
Oakville, ON, CA), pH 7.5, using a Kontes hand tissue grinder. Liver homogenates
were centrifuged at 9,000 x g for 20 minutes at 4º C. Each S9 supernatant was pipetted
into 2.0 mm cryovials and stored at – 80oC while precipitated pellets were discarded
(Hodson and Wilson, 1998).
2.4.2
EROD Activity
Ethoxyresorufin-o-deethylase activity was determined using the methods outlined in
Efler et al. (1998). A 50 µL aliquot of the S9 fraction was added in triplicate to a 96well plate. Standards (10 µL) with an additional 40 µL of Hepes grinding buffer were
added in duplicate to one row of a 96-well plate. 7ER/Hepes (50 µL) buffer was added
26
to each well and the plate was incubated for 10 min at room temperature. During this
time the plate was covered to protect 7ER from degradation from the light. The
catalytic effect was started with the addition of 10 µL of NADPH (ICN Biochemicals,
Aurora, OH, USA). Resorufin fluorescence was measured using a FLx 800 microplate
reader at 590 nm with an excitation wavelength of 530 nm. Enzyme activity increases
linearly with time and protein concentration. The amount of enzyme activity in
pmol/mg/min was obtained by comparing fluorescence against standard concentrations
of resorufin (Sigma Aldrich, Oakville, ON, CA).
2.4.3
Protein Determination
Protein concentrations were determined by the Micro Bio-Rad Assay (Efler et al.,
1998). The Bio-Rad reagent mixture consisted of 20 mL of Bio-Rad (Bio-Rad
Laboratories, Hercules, CA, USA) and 80 mL of distilled water. Liver homogenates
were diluted with distilled water from a 1:2 to a 1:10 ratio to maintain homogenate
concentrations. A series of bovine serum albumin (BSA) (Sigma Aldrich, Oakville,
ON, CA) concentrations, ranging from 0 to 0.5 mg/mL, were used as standards and
blanked against distilled water. Dilute liver homogenate and standards were added in
duplicate to a 96 well plate. A 200 µL aliquot of dilute Bio-Rad reagent was added to
each well and incubated at room temperature for 5 min. Absorbance was read by a LS
800 Lowry spectrophotometer at 595 nm. A linear standard curve for protein
concentrations versus absorbance was produced from the BSA standards and used to
calculate S9 protein concentrations in mg/mL.
27
2.5
Age Estimation
Otoliths from each sampling site were extracted in pairs from frozen carcasses. To
facilitate clearing, one otolith per specimen was placed in propylene glycol for 24 h, and
checked at 12 h for clarity. Otoliths were observed under a compound microscope at 1
to 6.3 magnification. The pith was arbitrarily designated as the point of origin and was
counted as year 1. Ages were reported to the nearest whole number. To ensure
accuracy growth rings were counted a minimum of two times and 10% of the otholiths
were externally aged
2.6
Statistical Analysis
Significant differences were assessed separately for each sex using either a one-way
ANOVA or ANCOVA followed by a Tukey post hoc test. Weight, length, age, size at
age, EROD activity and steroid levels were tested using ANOVA while gonad and liver
size, condition and size at age were tested using ANCOVA analyses of organ weight
versus body weight (by site), or weight versus length (by site). Prior to parametric
analysis, assumptions of normality and homogeneity of variances were examined using
probability plots and normality tests. Data that did not meet parametric assumptions
were log10 transformed and retested. All statistical tests were conducted using Systat
9.0 software (SPSS, SYSTAT, Chicago, IL, USA).
3.0 RESULTS
Saints Rest Beach was chosen as the in-city reference site because previous studies of
the harbour have indicated lower contamination levels on the west side of the city. In
28
addition to the presence of ideal habitat for rock gunnel, there are no direct chemical
inputs into the area and contamination levels should therefore represent background
harbour levels. The initial analysis examined whether there were differences in the
characteristics of rock gunnel between season and year during the late summer and fall
periods. Fish were collected at this site during August to November in 2001 and during
March to October in 2002. Due to occasionally low catch rates, the data were combined
within seasons. Fish caught in April-May were classified as spring, June-July fish were
classified as early summer, August-September fish were classified as late summer and
fish caught in October-November were classified as prespawning.
3.1
3.1.1
Reference Comparison
Late Summer and Prespawning Periods at Saints Rest Beach
Prespawning females in 2002 had significantly larger gonads (p=0.003), livers
(p=0.001) and condition (p=0.007) than late summer females (Table 3; Table 5; Table
6). However, these differences were not observed among females in 2001 (p=0.064;
p=0.16; p=0.090). There were no significant differences in the weight (p=0.40;
p=0.057) or length (p=0.11; p=0.36) of late summer and prespawning females in 2001
or 2002.
29
Table 3a: Regression table of liver sizes in female Pholis gunnellus in 2001 and 2002
Saints Rest Beach
Equation
-0.034(weight) + 0.015
-0.21(weight) + 0.058
-0.0006(weight) + 0.006
-0.008(weight) + 0.003
-0.004(weight) + 0.007
-0.043(weight) + 0.018
DF
1,5
1,8
1,8
1,6
1,9
1,15
p
0.001
0.012
0.009
0.22
<0.001
<0.001
R2
0.90
0.56
0.59
0.25
0.78
0.83
Red Head
Season
Variable
Equation
Late Summer Liver Weight 0.010(weight) + 0.09
Prespawning Liver Weight -0.018(weight) + 0.15
Spring
Liver Weight -0.012(weight) + 0.012
Late Summer Liver Weight -0.15(weight) + 0.012
Prespawning Liver Weight -0.92(weight) + 0.026
DF
1,6
1,6
1,15
1,10
1,23
p
0.21
0.014
0.003
<0.001
<0.001
R2
0.25
0.66
0.47
0.86
0.8
Duck Marsh
Equation
-0.044(weight) + 0.022
-0.021(weight) + 0.019
-0.0002(weight) + 0.008
-0.033(weight) + 0.015
DF
1,20
1,13
1,17
1,23
p
<0.001
<0.001
<0.001
<0.001
R2
0.83
0.75
0.78
0.87
Year
2002
Garnett Settlement
Season
Variable
Equation
Early Summer Liver Weight -0.027(weight) + .011
Late Summer Liver Weight -0.036(weight) + 0.017
DF
1,19
1,11
p
<0.001
<0.001
R2
0.72
0.94
Year
2002
Maces Bay
Season
Variable
Equation
Early Summer Liver Weight -0.031(weight) + 0.012
Prespawning Liver Weight -0.081(weight) + 0.021
DF
1,10
1,14
p
<0.001
<0.001
R2
0.95
0.96
Year
2001
2002
Year
2001
2002
Year
2001
2002
Season
Late Summer
Prespawning
Spring
Early Summer
Late Summer
Prespawning
Season
Late Summer
Prespawning
Spring
Late Summer
Variable
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Variable
Liver Weight
Liver Weight
Liver Weight
Liver Weight
30
Table 3b: Regression table of gonad size in female Pholis gunnellus in 2001 and 2002
Year
2001
2002
Year
2001
2002
Year
2001
2002
Year
2002
Year
2002
Variable
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Saints Rest Beach
Equation
0.003(weight) + 0.006
-0.16(weight) + 0.36
0.006(weight) + 0.001
-0.0002(weight) + 0.002
-0.009(weight) + 0.004
-0.057(weight) + 0.024
d.f.
1,7
1,19
1,15
1,6
1,8
1,16
p
0.10
<0.001
0.17
0.046
0.007
<0.001
R2
0.34
0.80
0.12
0.51
0.62
0.54
Season
Late Summer
Prespawning
Spring
Late Summer
Prespawning
Variable
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Red Head
Equation
0.001(weight) + 0.005
0.021(weight) + <0.0001
0.007(weight) + 0.002
-0.005(weight) + 0.003
-0.12(weight) + 0.007
d.f.
1,7
1,10
1,27
1,11
1,19
p
0.24
0.99
0.030
0.010
0.006
R2
0.19
<0.001
0.16
0.47
0.34
Season
Late Summer
Prespawning
Spring
Late Summer
Variable
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Duck Marsh
Equation
-0.12(weight) + 0.007
-0.81(weight) + 0.22
0.0006(weight) + 0.003
-0.009(weight) + 0.005
d.f.
1,16
1,2
1,17
1,19
p
<0.001
0.12
0.002
<0.001
R2
0.65
0.78
0.42
0.69
Variable
Gonad Weight
Gonad Weight
Garnett Settlement
Equation
0.003(weight) + 0.001
-0.16(weight) + 0.031
d.f.
1,11
1,10
p
0.16
<0.001
R2
0.17
0.69
Variable
Gonad Weight
Gonad Weight
Maces Bay
Equation
-0.002(weight) + 0.002
-0.018(weight) + 0.007
d.f.
1,8
1,9
p
<0.001
0.027
R2
0.95
0.44
Season
Late Summer
Prespawning
Spring
Early Summer
Late Summer
Prespawning
Season
Early Summer
Late Summer
Season
Early Summer
Prespawning
31
Table 4a: Regression table of liver size in male Pholis gunnellus in 2001 and 2002
Year
2001
2002
Year
2001
2002
Year
2001
2002
Year
2002
Year
2002
Variable
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Saints Rest Beach
Equation
0.011(weight) + 0.005
0.0002(weight) + 0.008
0.001(weight) + 0.006
0.007(weight) + 0.003
-0.0004(weight) + 0.006
0.002(weight) + 0.006
d.f.
1,7
1,19
1,15
1,6
1,8
1,16
p
0.001
<0.001
<0.001
0.032
<0.001
<0.001
R2
0.79
0.70
0.89
0.56
0.94
0.65
Season
Late Summer
Prespawning
Spring
Late Summer
Prespawning
Variable
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Red Head
Equation
0.026(weight) + 0.001
0.007(weight) + 0.004
0.0015(weight) + 0.007
-0.001(weight) + 0.006
-0.025(weight) + 0.010
d.f.
1,7
1,10
1,27
1,11
1,19
p
0.43
0.001
<0.001
<0.001
<0.001
R2
0.090
0.66
0.56
0.92
0.81
Season
Late Summer
Prespawning
Spring
Late Summer
Variable
Liver Weight
Liver Weight
Liver Weight
Liver Weight
Duck Marsh
Equation
-0.011(weight) + 0.011
0.016(weight) + 0.005
-0.009(weight) + 0.009
-0.007(weight) + 0.008
d.f.
1,16
1,2
1,18
1,19
p
<0.001
0.095
<0.001
<0.001
R2
0.85
0.82
0.86
0.89
Season
Early Summer
Late Summer
Variable
Liver Weight
Liver Weight
Garnett Settlement
Equation
-0.014(weight) + 0.01
-0.003(weight) + 0.007
d.f.
1,11
1,10
p
<0.001
0.003
R2
0.76
0.59
Variable
Liver Weight
Liver Weight
Maces Bay
Equation
-0.003(weight) + 0.009
0.012(weight) + 0.003
d.f.
1,8
1,9
p
<0.001
<0.001
R2
0.95
0.82
Season
Late Summer
Prespawning
Spring
Early Summer
Late Summer
Prespawning
Season
Early Summer
Prespawning
32
Table 4b: Regression table of gonad size in male Pholis gunnellus in 2001 and 2002
Year
2001
2002
Year
2001
2002
Year
2001
2002
Year
2002
Year
2002
Variable
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Saints Rest Beach
Equation
0.003(weight) + 0.006
-0.16(weight) + 0.36
0.006(weight) + 0.001
-0.0002(weight) + 0.002
-0.009(weight) + 0.004
-0.057(weight) + 0.024
d.f.
1,7
1,19
1,15
1,6
1,8
1,16
p
0.10
<0.001
0.17
0.046
0.007
<0.001
R2
0.34
0.80
0.12
0.51
0.62
0.54
Season
Late Summer
Prespawning
Spring
Late Summer
Prespawning
Variable
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Red Head
Equation
0.001(weight) + 0.005
0.021(weight) + 0.0001
0.007(weight) + 0.002
-0.005(weight) + 0.003
-0.12(weight) + 0.007
d.f.
1,7
1,10
1,27
1,11
1,19
p
0.24
<0.001
0.030
0.010
0.006
R2
0.19
0.99
0.16
0.47
0.34
Season
Late Summer
Prespawning
Spring
Late Summer
Variable
Gonad Weight
Gonad Weight
Gonad Weight
Gonad Weight
Duck Marsh
Equation
-0.12(weight) + 0.007
-0.81(weight) + 0.22
0.0006(weight) + 0.003
-0.009(weight) + 0.005
d.f.
1,16
1,2
1,17
1,19
p
<0.001
0.12
0.002
<0.001
R2
0.65
0.78
0.42
0.69
Season
Variable
Early Summer Gonad Weight
Late Summer Gonad Weight
Garnett Settlement
Equation
0.003(weight) + 0.001
-0.16(weight) + 0.031
d.f.
1,11
1,10
p
0.16
<0.001
R2
0.17
0.69
Season
Variable
Early Summer Gonad Weight
Prespawning Gonad Weight
Maces Bay
Equation
-0.002(weight) + 0.002
-0.018(weight) + 0.007
d.f.
1,8
1,9
p
<0.001
0.027
R2
0.95
0.44
Season
Late Summer
Prespawning
Spring
Early Summer
Late Summer
Prespawning
33
Table 5: Mean weight, length, liver size, gonad size and condition for male and female Pholis gunnellus
during late summer and prespawning at Saints Rest Beach in 2001 and 2002.
Sex
Year
Time
Weight (g)
Length (mm)
Liver
Gonad
Condition
2001
Late Summer
Prespawning
Late Summer
Prespawning
6.91 + 0.80 a
5.95 + 0.80 a
5.21 + 0.70 b
7.12 + 0.60 b
135.3 + 4.7 a
125.4 + 3.6 a
124.6 + 5.8 b
130.7 + 3.6 b
0.94 + 0.09 a
1.27 + 0.20 a
0.63 + 0.05 b
1.13 + 0.10 c
1.02 + 0.10 a
1.91 + 0.70 a
0.44 + 0.07 b
1.59 + 0.30 c
0.27 + 0.10 a
0.29 + 0.10 a
0.25 + 0.10 b
0.30 + 0.10 c
Late Summer
Prespawning
Late Summer
Prespawning
4.00 + 0.90 a
4.77 + 0.50 a
7.15 + 1.20 b
7.90 + 0.60 b
111.9 + 5.4 a
118.3 + 3.1 a
136.0 + 6.4 b
138.3 + 3.8 b
0.83 + 0.07 a
0.76 + 0.06 a
0.56 + 0.03 b
0.61 + 0.04 b
0.73 + 0.3 a
0.67 + 0.2 a
0.20 + 0.04 b
1.57 + 0.2 c
0.26 + 0.01 a
0.27 + 0.01 a
0.26 + 0.01 b
0.29 + 0.01 b
Female
2002
2001
Male
2002
a
Each value represents the mean + SEM. Organ size for summary purposes is expressed as percentage of body weight.
Values sharing an alphabetical superscript within each sex are not statistically different (p>0.05). Liver size, gonad size
and condition were analyzed by ANCOVAs (organ weight versus body weight or body weight versus body length) but
are represented by summary indices.
34
Prespawning males in 2002 had significantly larger gonads (p=0.005) than late summer
males, however, this difference was not observed in 2001 (p=0.31) (Table 4; Table 5;
Table 6). There were no differences in the weight (p=0.42; p=0.56), length (p=0.47;
p=0.74), liver size (p=0.93; p=0.48) or condition (p=0.47; p=0.29) of late summer and
prespawning males within 2001 or 2002.
Table 6: Anova table of seasonal differences at Saints Rest Beach in 2001 and 2002
Females
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.40
0.11
0.16
0.064
0.090
2001
F
0.74
2.88
2.26
4.03
3.31
DF
1,15
1,15
1,14
1,14
1,14
Sign.
0.057
0.36
0.001
0.003
0.007
2002
F
3.98
0.883
13.98
11.31
8.70
DF
1,26
1,26
1,24
1,24
1,24
Sign.
0.56
0.74
0.48
0.005
0.29
2002
F
0.35
0.11
0.51
9.69
1.16
DF
1,26
1,26
1,25
1,24
1,25
Males
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
3.1.2
Sign.
0.42
0.47
0.93
0.31
0.47
2001
F
0.65
0.54
0.008
1.05
0.53
DF
1,28
1,28
1,27
1,26
1,27
Seasonal Changes for 2002
Fishing attempts earlier than April were unsuccessful. Fish were collected from April
to November of 2002 at Saints Rest Beach. Fish were collected during spring when
they returned from offshore (April to May), early summer (June to July), late summer
(August to September) and prespawning (October to November) to look at the seasonal
changes over one full reproductive cycle.
35
Prespawning females had significantly larger gonad, liver and condition than fish
collected in the spring (p=0.015; p=0.009; p=0.025), early summer (p=0.005; p=0.001;
p=0.026) or late summer (p=0.025; p=0.001; p=0.007) (Table 7; Table 8). There were
seasonal differences in weight, with significant differences between prespawning and
early summer (p=0.002) fish and also in length with spring and prespawning fish being
significantly longer than fish collected in early summer (p=0.043; p=0.029). The
reduced size in early summer was related to the increased number of smaller fish as
younger animals returned.
Prespawning male fish had significantly larger gonad sizes than fish collected in spring
(p<0.001), early summer (p<0.001) and late summer (p<0.001) (Table 7; Table 8).
Liver sizes of prespawning fish were significantly larger than in fish collected during
early summer (p=0.039). There were also significant increases in condition between
spring and prespawning fish (p=0.006). There were no seasonal differences in weight
(p=0.064) or length (p=0.13).
Table 7: Anova table of seasonal differences at Saints Rest Beach in 2002
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.003
0.027
<0.001
<0.001
0.008
Female
F
5.36
3.37
9.42
8.03
4.52
DF
3,42
3,42
3,38
3,38
3,38
36
Sign.
0.064
0.13
0.13
<0.001
<0.001
Male
F
2.59
2.00
1.97
9.42
11.37
DF
3,49
3,49
3,45
3,45
3,45
Table 8: Mean weight, length, liver size, gonad size and condition for male and female Pholis gunnellus during
2002 at Saints Rest Beach.
Sex
Time
N
Weight (g)
a,b
Length (mm)
a
Liver
Gonad
Condition
Female
Spring
Early Summer
Late Summer
Prespawning
10
8
11
17
5.31 + 0.37
3.71 + 0.30 a
5.21 + 0.70 a,b
7.12 + 0.60 b
131.5 + 3.7
112.6 + 2.5 b
124.6 + 5.8 a,b
130.7 + 3.6 a
0.58 + 0.05
0.51 + 0.04 a
0.63 + 0.05 a
1.13 + 0.10 b
0.48 + 0.10
0.38 + 0.01 a
0.44 + 0.07 a
1.59 + 0.30 b
0.23 + 0.01 a
0.26 + 0.01 a
0.25 + 0.01 a
0.30 + 0.01 b
Male
Spring
Early Summer
Late Summer
Prespawning
17
8
10
18
5.68 + 0.60 a
5.03 + 1.1 a
7.15 + 1.20 a
7.90 + 0.60 b
130.6 + 4.4 a
118.9 + 9.8 a
136.0 + 6.4 a
138.3 + 3.8 a
0.59 + 0.02 a,b
0.49 + 0.06 a
0.56 + 0.03 a,b
0.61 + 0.04 b
0.22 + 0.04 a
0.18 + 0.03 a
0.20 + 0.04 a
1.57 + 0.20 b
0.24 + 0.01 a
0.26 + 0.01 a,b
0.26 + 0.01 a,b
0.29 + 0.01 b
a
a
a
Each value represents the mean + SEM. Organ size for summary purposes, is expressed as percentage of body weight.
Values sharing an alphabetical superscript within each sex are not statistically different (p>0.05). Liver size, gonad size
and condition were analyzed by ANCOVAs (organ weight versus body weight or body weight versus body length) but are
represented by summary indices.
37
3.1.3
Size at Maturity
In 2001 and 2002 significant increases in gonad size with body weight were observed in
prespawning female gunnel greater than 5.5 g body weight (Figure 2). Below that
weight, gonad size was independent of body weight. These fish were classified as
immature because their gonads were small and undeveloped. Immature fish were
removed from analysis. A similar difference was seen in 2002. There were no
differences in gonadal development between 2001 and 2002 for either sex.
Mature female gunnel were longer (p=0.044; p=0.011) and heavier (p=0.039, p=0.003)
than immature gunnel at Saint Rest in 2001 and 2002. Mature gunnels in 2002 had
significantly larger liver sizes (p=0.005) and condition (p=0.032) than immature gunnels,
however, these differences were not observed in 2001 (p=0.30; p=0.17).
Male gunnel showed a similar size of maturity, with fish above 5.5 g showing increased
gonad weight with body weight (Figure 3). In 2001, mature male gunnel were longer
(p<0.001) and heavier (p<0.001) than immature gunnel, however, this difference was not
observed in 2002 (p=0.083; p=0.081). Liver sizes of mature fish were larger than
immature fish in 2002 (p=0.031), this difference was not observed in 2001 (p=0.20).
There were no significant differences in condition for either year (p=0.67; p=0.76).
38
0.0
2001: Log Gonad Weight = 2.77(Log Weight) - 3.37 (df =1,8, r2 =0.67, p =0.004)
2002: Log Gonad Weight = 3.19 (Log Weight) - 3.59 (df =1,15, r2 =0.88, p<0.001)
Log Gonad Weight (g)
-0.5
-1.0
-1.5
-2.0
-2.5
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1.0
1.1
Log Body Weight (g)
2001
2002
Figure 2: Gonadal development of female Pholis gunnellus collected at Saints Rest Beach
during 2001 and 2002.
39
0.0
2001: Log Gonad Weight = 2.97(Log Weight) - 3.69 (df =1,15, r2 =0.77, p<0.001)
2002: Log Gonad Weight = 2.42(Log Weight) - 3.20 (df =1,17, r2 = 0.44, p =0.003)
Log Gonad Weight (g)
-0.5
-1.0
-1.5
-2.0
-2.5
-3.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
Log Body Weight (g)
2001
2002
Figure 3: Gonadal development of male Pholis gunnellus collected at Saints Rest Beach
during 2001 and 2002
40
3.1.4
Pre and Post Spawning
Little is known about the reproductive biology of rock gunnel in the Bay of Fundy; thus,
spring collections in 2002 compared fish size and organ size after wintering at the
offshore spawning areas.
Prespawning females had significantly larger livers (p=0.016), gonads (p=0.034) and
condition (p<0.001) than fish returning in the spring (Table 9; Table 10). There were no
seasonal differences in the weight (p=0.29), and length (p=0.23) of prespawning females
in 2001 and females returning in the spring of 2002.
Prespawning males also had significantly larger livers (p=0.004), gonads (p=0.011), and
condition (p=0.005) than fish returning in the spring of 2002 (Table 9; Table 10).
However, fish returning in the spring were longer (p=0.037) than prespawning fish.
There were no differences in the weight (p=0.25) of prespawning males of 2001 and male
gunnel returning in spring of 2002.
41
Table 9: Mean weight, length, liver size, gonad size and condition of male and female Pholis gunnellus pre and post spawning
at Saint Rest Beach.
Sex
Year
Time
N
Weight (g)
Length (mm)
Liver
Gonad
Condition
2001
Prespawning
10
5.95 +/- 0.8 a
125.4 +/- 3.6 a
1.27 +/- 0.2 a
1.91 +/- 0.7 a
0.29 +/- 0.01 a
2002
Spring
10
5.31 +/- 0.5 a
131.5 +/- 3.7 a
0.58 +/- 0.05 b
0.48 +/- 0.06 b
0.23 +/- 0.01 b
2001
Prespawning
21
4.77 +/- 0.5 a
118.3 +/- 3.1 a
0.76 +/- 0.06 a
0.67 +/- 0.2 a
0.27 +/- 0.01 a
2002
Spring
17
5.68 +/- 0.6 a
130.6 +/- 4.4 b
0.59 +/- 0.02 b
0.22 +/- 0.04 b
0.24 +/- 0.01 b
Female
Male
a
Each value represents the mean + SEM. Organ size for summary purposes, is expressed as percentage of body weight. Values
sharing an alphabetical superscript within each sex are not statistically different (p>0.05). Liver size, gonad size and condition were
analyzed by ANCOVAs (organ weight versus body weight or body weight versus body length) but are represented by summary
indices.
42
Table 10: Anova table of pre and post spawning fish at Saints Rest Beach
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
3.2
Sign.
0.47
0.25
0.017
0.04
0.003
Female
F
0.52
1.38
7.02
4.98
12.68
DF
1,18
1,18
1,18
1,18
1,18
Sign.
0.26
0.037
0.004
<0.001
0.005
Male
F
1.29
4.69
9.81
62.73
9.10
DF
1,35
1,35
1,35
1,35
1,35
Reference Site Comparisons
Saints Rest Beach was selected as the least contaminated site within the harbour, but it
was necessary to compare performance of fish at Saints Rest Beach with fish from
additional reference sites outside of Saint John Harbour. Two additional sites outside the
harbour, Garnett Settlement and Maces Bay, were chosen for comparison. Fish were
collected from all sites during early summer, from Garnett Settlement and Saints Rest
Beach during late summer and Maces Bay and Saints Rest Beach at prespawning.
3.2.1
Early Summer
Female gunnel from Maces Bay were heavier, longer and had a higher condition than fish
collected at Garnett Settlement (p<0.001; p=0.020; p<0.001) and Saints Rest (p=0.023;
p=0.003; p<0.001) (Table 11; Table 12). There were no differences in female liver size
(p=0.93) or gonad size (p=0.10) among the three sites.
Male gunnel from Maces Bay were heavier, longer, had higher liver sizes and condition
than fish collected at Garnett Settlement (p<0.001; p=0.012; p=0.012) and Saints Rest
43
(p<0.001; p<0.001; p=0.034) (Table 11; Table 12). There were no differences in gonad
size (p=0.69) among all three sites.
3.2.2
Late Summer
Female fish collected at Garnett Settlement were heavier (p=0.040), had larger livers
(p<0.001), gonads (p=0.005) and condition (p=0.007) than fish from Saints Rest. There
were no significant differences in length (p=0.24) of female fish collected from Garnett
Settlement or Saints Rest Beach.
Male gunnel from Garnett Settlement had larger gonads and condition than fish collected
at Saints Rest (p=0.011; p=0.015). There were no differences in the weight (p=0.93),
length (p=0.62) or liver size (p=0.25) of male fish collected during late summer at
Garnett Settlement or Saints Rest Beach.
3.2.3
Prespawning
Prespawning female fish from Maces Bay were heavier (p=0.013), longer (p=0.013) and
had a higher condition (p<0.001) than fish from Saints Rest. There were no differences
in the liver (p=0.093) or gonad sizes (p=0.10) in fish from either site.
Male fish collected at Maces Bay had larger gonad sizes (p=0.011) and condition
(p=0.016) than fish collected at Saints Rest. There were no differences in the weight
(p=0.54), length (p=0.78) or liver size (p=0.11) of male fish collected during
prespawning.
44
Table 11: Mean weight, length, liver size, gonad size, and condition of male and female Pholis gunnellus at reference areas in
2002.
Sex
Site
Saints Rest Beach
Female
Garnett Settlement
Maces Bay
Saints Rest Beach
Male
Garnett Settlement
Maces Bay
Time
N
Weight (g)
Length (mm)
a
112.6 +/- 2.5
a
Liver
Gonad
0.51 +/- 0.04
a
Condition
a
0.26 +/- 0.01 a
0.44 +/- 0.07 a
0.25 +/- 0.01 a
0.38 +/- 0.01
Early Summer
8
3.71 +/- 0.3
Late Summer
11
5.21 +/- 0.7 a
124.6 +/- 5.8 a
Prespawning
17
7.12 +/- 0.6
a
a
Early Summer
21
5.94 +/- 0.6 a
126.8 +/- 3.1 a
0.66 +/- 0.05 a
0.43 +/- 0.03 a
0.27 +/- 0.01 a
Late Summer
13
8.31 +/- 1.1 b
134.5 +/- 5.7 a
1.11 +/- 0.1 b
1.21 +/- 0.2 b
0.31 +/- 0.01 b
Early Summer
12
13.30 +/- 1.7 b
158.6 +/- 5.7 b
0.94 +/- 0.05 a
0.44 +/- 0.03 a
0.31 +/- 0.01 b
Prespawning
16
13.91 +/- 2.3 b
150.4 +/- 6.6 b
1.26 +/- 0.1 a
1.57 +/- 0.3 a
0.34 +/- 0.01 b
Early Summer
8
5.03 +/- 1.1 a
118.9 +/- 9.8 a
0.49 +/- 0.06 a
0.18 +/- 0.03 a
0.26 +/- 0.003 a
Late Summer
10
7.15 +/- 1.2 a
136.0 +/- 6.4 a
0.56 +/- 0.03 a
0.20 +/- 0.04 a
0.26 +/- 0.01 a
Prespawning
18
7.90 +/- 0.6
a
a
a
Early Summer
13
7.51 +/- 0.6 a
135.8 +/- 3.5 a
0.58 +/- 0.4 a
0.14 +/- 0.02 a
0.29 +/- 0.005 a
Late Summer
12
7.04 +/- 0.6 a
132.4 +/- 3.5 a
0.63 +/- 0.05 a
0.73 +/- 0.2 b
0.29 +/- 0.01 b
Early Summer
10
19.9 +/- 3.6 b
169.9 +/- 11.5 b
0.81 +/- 0.04 b
0.14 +/- 0.01 a
0.35 +/- 0.02 b
Prespawning
11
8.86 +/- 1.6 a
136.4 +/- 6.4 a
0.72 +/- 0.05 a
1.57 +/- 0.2 b
0.31 +/- 0.01 b
a
130.7 +/- 3.6
138.3 +/- 3.8
0.63 +/- 0.05 a
1.13 +/- 0.1
a
0.61 +/- 0.04
1.59 +/- 0.3
0.42 +/- 0.1
a
a
0.30 +/- 0.01 a
0.29 +/- 0.01 a
Each value represents the mean + SEM. Organ size for summary purposes, is expressed as percentage of body weight. Values
sharing an alphabetical superscript within each sex are not statistically different (p>0.05). Liver size, gonad size and condition were
analyzed by ANCOVAs (organ weight versus body weight or body weight versus body length) but are represented by summary
indices.
45
Table 12: Anova table for reference site comparisons in 2002.
Early Summer
Sign.
<0.001
<0.001
0.03
0.66
<0.001
Male
F
13.86
9.07
4.03
0.42
22.99
DF
2,28
2,28
2,25
2,27
2,25
F
4.78
1.46
18.63
9.79
8.94
Late Summer
DF
Sign.
1,20
0.93
1,20
0.61
1,20
0.25
1,20
<0.001
1,20
0.097
F
0.007
0.26
1.42
21.74
3.04
DF
1,20
1,20
1,19
1,18
1,19
F
6.88
7.01
3.01
2.82
13.45
Prespawning
DF
Sign.
1,31
0.54
1,31
0.78
1,30
0.11
1,30
0.011
1,29
0.015
F
0.39
0.078
2.80
7.57
6.84
DF
1,27
1,27
1,26
1,25
1,25
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Female
Sign.
F
<0.001
20.36
<0.001
26
0.93
0.77
0.101
2.44
<0.001
15.97
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.04
0.24
>0.001
0.005
0.007
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.013
0.013
0.093
0.1
0.001
3.3
DF
2,38
2,38
2,37
2,37
2,35
In-city comparisons
Little River Marsh and Red Head were chosen as exposed areas within the city. Primarytreated sewage wastewater flows directly into Little River Marsh and previous studies
have indicated invertebrate exposure to tributylin, an antifouling compound in marine
paint at Red Head (Prouse and Ellis 1997, Delaney 2001).
46
3.3.1
Late Summer and Prespawning at Little River Marsh and Red Head
3.3.1.1 Comparisons within sites
Red Head
Prespawning female fish had significantly larger gonads than late summer fish (p=0.006)
(Table 13; Table 14). However, there were no differences in the weight (p=0.12), length
(p=0.10), liver sizes (p=0.36) or condition (p=0.81) of late summer and prespawning fish.
Male fish collected in late summer had significantly larger liver sizes (p=0.008) and
condition (p=0.016) than prespawning fish (Table 13; Table 14). There were no
differences in the weight (p=0.16), length (p=0.97) or gonad sizes (p=0.50) of fish
collected in late summer or prespawning.
Little River Marsh
Prespawning female gunnel had significantly larger gonads (p<0.001) than late summer
fish, however, condition (p<0.001) was significantly higher in fish collected during late
summer. There were no differences in weight (p=0.19), length (p=0.89) or liver sizes
(p=0.76) of late summer and prespawning fish.
Male fish collected during late summer had significantly larger liver sizes (p=0.040) and
condition (p=0.005) than prespawning fish, however, fish collected during prespawning
had significantly larger gonads (p=0.004). There were no significant differences in
weight (p=0.48) or length (p=0.14) of late summer and prespawning males.
47
Table 13: Mean weight, length, liver size, gonad size and condition of male and female Pholis gunnellus collected in Saint
John Harbour during 2001
Sex
Female
Male
Site
Saints Rest
Beach
Time
N
Weight (g)
Length (mm)
Liver
Gonad
Condition
Late Summer
7
6.91 + 0.8 a,A
135.3 + 4.7 a,A
0.94 + 0.09 a,A
1.02 + 0.10 a,A
0.27 + 0.01 a,A
Red Head
Prespawning
Late Summer
Prespawning
10
8
8
5.95 + 0.8 a,A
5.97 + 0.5 a,A
4.74 + 0.6 a,A
125.4 + 3.6 a,AB
124.6 + 2.8 a,A
115.5 + 4.4 a,B
1.27 + 0.2 a,A
0.95 + 0.1a,A
1.03 + 0.2 a,A
1.91 + 0.70 a,A
0.95 + 0.1 a,A
1.69 + 0.3 b,A
0.29 + 0.01 a,A
0.31 + 0.02 a,A
0.30 + 0.01 a,A
Little River
Marsh
Late Summer
22
9.47 + 1.2 a,A
131.7 + 5.3 a,A
1.56 + 0.1 a,A
1.26 + 0.1 a,A
0.38 + 0.01 a, B
Prespawning
15
7.38 + 0.7 a,A
130.8 + 3.4 a,A
1.67 + 0.1a,A
2.94 + 0.3 b,B
0.31 + 0.01 b, B
Saints Rest
Beach
Late Summer
9
4.0 + 0.9 a,A
111.9 + 5.4 a,A
0.83 + 0.07 a,A
0.73 + 0.30 a,A
0.26 + 0.01 a,A
Red Head
Prespawning
Late Summer
Prespawning
21
9
12
4.77 + 0.5 a,A
4.83 + 0.6 a,A
3.78 + 0.4 a,A
118.3 + 3.1 a,AB
113.1 + 3.2 a,A
112.9 + 3.8 a,A
0.76 + 0.06 a,A
0.81+ 0.10 b,A
0.59 + 0.03 a,A
0.67 + 0.20 a,A
0.46 + 0.1 a,A
0.58 + 0.2 a,A
0.27 + 0.01 a,A
0.32 + 0.02 b,AB
0.26 + 0.003 a,A
Little River
Marsh
Late Summer
18
6.33 + 0.7 a,A
119.3 + 4.7 a,A
0.89 + 0.05 a,B
0.44 + 0.07 a,A
0.35 + .007 a,B
Prespawning
4
7.46 + 1.2 a,B
135.5 + 7.3 a,B
0.73 + 0.07 b,A
1.06 + 0.3 b,B
0.29 + 0.01 b,A
a
Each value represents the mean + SEM. Organ size for summary purposes, is expressed as percentage of body weight. Values
sharing an alphabetical superscript within each sex are not statistically different (p>0.05). Liver size, gonad size and condition were
analyzed by ANCOVAs (organ weight versus body weight or body weight versus body length) but are represented by summary
indices. Smaller case letters represent within site comparison. Capital letters represent between site comparisons.
48
Table 14: Anova table of within site comparisons in Saint John Harbour in 2001
Red Head
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.12
0.1
0.36
0.006
0.82
Sign.
0.19
0.89
0.76
<0.001
<0.001
Female
F
2.79
3.01
0.92
11.04
0.057
F
1.76
0.019
0.094
33.43
15.81
Sign.
0.16
0.97
0.008
0.97
0.017
Male
F
2.18
0.001
8.73
0.002
7.00
DF
1,19
1,19
2,18
2,18
2,17
Little River Marsh
DF
Sign.
1,35
0.49
1,35
0.14
2,34
0.041
2,34
0.004
2,34
0.006
F
0.51
2.34
4.81
11.15
9.65
DF
1,20
1,20
2,19
3,18
2,19
DF
1,14
1,14
2,13
2,13
2,13
3.3.1.2 Comparisons among Sites
Late summer and prespawning females at Little River Marsh had significantly larger
condition than fish collected at Red Head (p=0.006; p=0.023) and Saints Rest (p<0.001;
p=0.005) (Table 13; Table 15). Prespawning fish at Little River Marsh were also longer
than fish colleted at Red Head (p=0.021) and had a larger gonad sizes than fish collected
at Red Head (p=0.015) and Saints Rest Beach (p=0.003), however, this difference was
not observed during late summer (p=0.49). There were no differences in body weight
(p=0.13; p=0.061), liver sizes (0.083; p=0.24) and gonad sizes (p=0.83; p=0.67) among
late summer and prespawning fish collected in Saint John Harbour in 2001.
Male gunnel collected at Little River Marsh during late summer had significantly larger
livers than fish collected at Red Head (p=0.003) and Saints Rest (p=0.004) and also had
higher condition than fish at Saint Rest (p=0.001) (Table 13; Table 15). There were no
49
differences in weight (p=0.10), length (p=0.49) or gonad sizes (p=0.47) of fish from
either site.
During prespawning fish collected at Little River Marsh were heavier and had larger
gonads than fish collected at Red Head (p=0.006; p=0.001) and Saints Rest (p=0.038;
p=0.027). Little River Marsh fish were also longer than fish collected at Red Head
(p=0.021). There were no differences in liver weight (p=0.063) or condition (p=0.18) of
prespawning fish.
Table 15: Anova table of between site comparisons in Saint John Harbour in 2001.
Late Summer
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.13
0.58
0.083
0.83
<0.001
Female
F
2.13
0.56
2.68
0.18
15.88
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.061
0.032
0.24
0.67
0.028
F
3.08
3.86
1.52
0.41
4.08
3.3.2
DF
2,34
2,34
2,33
2,33
2,33
Sign.
0.1
0.49
<0.001
0.47
0.002
Male
F
2.46
0.72
9.85
0.76
7.73
DF
2,33
2,33
2,32
2,32
2,32
Prespawning
DF
2,30
2,30
2,29
2,29
2,27
Sign.
0.009
0.028
0.064
<0.001
0.18
F
5.47
3.98
2.99
9.37
1.78
DF
2,34
2,34
2,33
2,31
2,33
Seasonal Site Differences within Saint John Harbour in 2002
3.3.2.1 Spring
Females collected during the spring at Little River Marsh were longer and heavier than
fish collected at Red Head (p<0.001; p=0.001) and Saints Rest (p=0.001; p=0.004) (Table
16; Table 17). There were no differences in liver (p=0.051) and gonad sizes (p=0.42) or
condition (p=0.35) of female fish among the sites.
50
Table 16: Seasonal changes in weight, length, liver size, gonad size and condition of male and female
Pholis gunnellus in Saint John Harbour in 2002.
Sex
Female
Male
Site
Time
N
Weight (g)
a
Length (mm)
131.5 + 3.7
a
Liver
0.48 + 0.06
Condition
a
0.23 + 0.01 a
Spring
10
5.31 + 0.5
Saints Rest
Beach
Late Summer
11
5.21 + 0.7 ab
124.6 + 5.8 a
0.63 + 0.05 a
0.44 + 0.07a
0.25 + 0.01 a
Red Head
Prespawning
Spring
Late Summer
Prespawning
17
17
12
25
7.12 + 0.6 a
5.68 + 0.5 a
4.54 + 0.5 a
5.93 + 0.3 a
130.7 + 3.6 a
132.8 + 3.5 a
114.2 + 4.3 a
123.1 + 1.6 b
1.13 + 0.1 a
0.92 + 0.1 a
0.82 + 0.1 b
0.99 + 0.1 b
1.59 + 0.3 a
0.64 + 0.08 a
0.66 + 0.1 a
1.04 + 0.1 a
0.30 + 0.01 a
0.24 + 0.01 a
0.29 + 0.01 a
0.31 + 0.01 a
Little River
Marsh
Spring
19
9.59 + 0.9 b
155.3 + 5.1 b
0.81 + 0.03 a
0.63 + 0.03 a
0.24 + 0.01 a
Late Summer
25
7.11 + 0.6 b
126.9 + 2.8 a
0.99 + 0.06 b
0.66 + 0.05 a
0.33 + 0.01 b
Spring
17
5.68 + 0.6 a
130.6 + 4.4 a
0.59 + 0.02 a
0.22 + 0.04 a
0.24 + 0.01 a
Saints Rest
Beach
Late Summer
10
7.15 + 1.2 a
136.0 + 6.4 a
0.56 + 0.03 a
0.20 + 0.04 a
0.26 + 0.01 a
Red Head
Prespawning
Spring
Late Summer
Prespawning
18
29
13
21
7.90 + 0.6 a
4.47 + 0.4 a
4.73 + 0.6 a
6.79 + 0.1 a
138.3 + 3.8 a
120.4 +- 3.2 a
118.4 + 4.0 a
128.2 + 3.3 b
0.61 + 0.04 a
0.78 + 0.6 b
0.54 + 0.02 a
0.62 + 0.05 a
1.57 + 0.2 a
0.36 + 0.04 a
0.21 + 0.04 a
0.46 + 0.07 b
0.29 + 0.01 a
0.24 + 0.01 a
0.27 + 0.01 a
0.31 + 0.01 a
Little River
Marsh
Spring
20
5.54 + 0.7 a
125.1 + 5.2 a
0.74 + 0.05 ab
0.30 + 0.04 a
0.26 + 0.01 a
Late Summer
21
6.84 + 0.7 a
125.3 + 4.6 a
0.68 + 0.03 b
0.30 + 0.03 a
0.32 + 0.01 b
a
0.58 + 0.05
Gonad
a
Each value represents the mean + SEM. Organ size for summary purposes, is expressed as percentage of body weight. Values
sharing an alphabetical superscript within each sex are not statistically different (p>0.05). Liver size, gonad size and condition were
analyzed by ANCOVAs (organ weight versus body weight or body weight versus body length) but are represented by summary
indices.
51
Red Head males had significantly larger livers than males collected at Saint Rest
(p=0.023). There were no differences in the weight (p=0.22), length (p=0.23), gonad
sizes (p=0.061) or condition (p=0.35) of male fish in Saint John Harbour.
Table 17: Anova table of seasonal site differences in Saint John Harbour during 2002
Spring
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
<0.001
<0.001
0.052
0.43
0.35
Female
F
10.91
9.39
3.18
0.87
1.08
Sign.
0.22
0.23
0.022
0.061
0.14
Male
F
1.53
1.5
4.08
2.93
2.06
DF
2,63
2,63
2,63
2,63
2,63
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.013
0.072
0.002
0.122
0.005
F
4.78
2.78
7.26
2.2
6.14
Late Summer
DF
Sign.
2,45
0.11
2,45
0.1
2,44
0.021
2,44
0.2
2,44
0.002
F
2.36
2.41
4.3
1.69
7.43
DF
2,41
2,41
2,40
2,40
2,40
Variables
Weight
Length
Liver Weight
Gonad Weight
Condition
Sign.
0.055
0.035
0.033
0.091
0.78
F
3.9
4.77
4.87
3.00
0.079
Prespawning
DF
1,40
1,40
1,39
1,39
1,39
F
1.65
4.13
7.56
8.61
3.15
DF
1,37
1,37
1,36
1,36
1,36
DF
2,43
2,43
2,43
2,43
2,43
Sign.
0.21
0.049
0.009
0.006
0.084
3.3.2.2 Late Summer
Little River Marsh females were heavier than fish from Red Head (p=0.018). Female fish
collected from Little River Marsh and Red Head had had significantly larger livers than
fish from Saints Rest (p=0.002; p=0.028), while the condition of fish from Little River
Marsh was significantly higher than fish from Red Head (p=0.010) and Saints Rest
52
(p=0.012). There were no differences in the length (p=0.072) or gonad sizes (p=0.12) of
females from the three sites.
Male fish collected at Little River Marsh had larger livers and condition than fish
collected at Red Head (p=0.016; p=0.017) and Saints Rest (p=0.029; p=0.005). There
was no difference in the weight (p=0.11), length (p=0.10) or gonad sizes (p=0.20) of
male fish.
3.3.2.3 Prespawning
Female fish from Saints Rest Beach were longer and had larger livers than fish collected
from Red Head (p=0.034; p=0.033). There were no differences in female weight
(p=0.055), gonad weight (p=0.091), or condition (p=0.78).
Prespawning male fish at Saints Rest were longer (p=0.049) and had larger gonads
(p=0.006) than fish collected at Red Head. Male fish displayed no difference in body
weight (p=0.21), liver weight (p=0.90) or condition (p=0.084).
3.4
3.4.1
Production of Gonadal Sex Steroids
Late Summer and Prespawning at Red Head
Fish collected at Red Head after September 13 were considered prespawning fish because
the GSI was significantly larger than fish collected earlier in the year. There were no
significant differences between sites in female estradiol (p=0.63) and testosterone
(p=0.47) or male 11-ketotestosterone (p=0.73) and testosterone (p=0.53) steroid levels in
gonads incubated in forskolin. Prior to September 13 there was not enough gonadal tissue
53
to incubate in both forskolin and 25-OH, therefore gonads were only incubated in
forskolin.
3.4.2
Reference and In City Comparisons
3.4.2.1 Among Sites 2001
There were no significant differences in female estradiol or testosterone levels of gonadal
tissue incubated in forskolin (p=0.11; p=0.36) or 25-OH (p=0.33; p=0.10) between sites
(Table 18).
Table 18: Anova table of steroid differences among sites in 2001
Female
Steroid
Estradiol
Testosterone
Sign.
0.11
0.36
Forskolin
F
2.47
1.08
DF
2,24
2,24
Sign.
0.33
0.10
25-OH
F
1.16
2.55
DF
2,21
2,21
3.4.2.2 Between Incubation Media 2001
Estradiol levels of gonadal tissue from Saints Rest (p=0.005), Red Head (p<0.001) and
Little River Marsh (p<0.001) females were higher when incubated in forskolin than
25OH (Table 19). In contrast testosterone levels (p<0.001; p=0.013; p<0.001) were
higher when gonadal tissue was incubated in 25OH as opposed to forskolin.
Table 19: Anova table of steroid differences between incubation media in 2001
.
Female
Steroid
Saints Rest
Red Head
Little River Marsh
Sign.
0.005
<0.001
<0.001
Estradiol
F
13.67
61.56
51.30
DF
1,9
1,9
1,27
54
Sign.
<0.001
0.013
<0.001
Testosterone
F
30.21
8.98
31.97
DF
1,9
1,10
1,27
3.4.2.3 Among Sites 2002
Estradiol levels of female gonadal tissue incubated in 25OH from Saints Rest were
significantly lower than fish collected at Little River Marsh (p<0.001), Garnett
Settlement (p=0.039) and Maces Bay (p=0.015) and Red Head (p=0.008) (Figure 4;
Table 20). Female testosterone levels of Saint Rest fish were significantly lower than
fish collected Little River Marsh (p=0.038), Garnett Settlement (p<0.001) and Maces Bay
(p<0.001). There were no significant differences in female estradiol (p=0.27) or
testosterone (p=0.90) levels of gonadal tissue incubated in forskolin in late summer or
prespawning fish.
11-Ketotestosterone levels of gonadal tissue incubated in forskolin were significantly
higher in Maces Bays than Little River Marsh (p=0.011) or Red Head (p=0.005) males.
11-Ketotestosterone levels of gonadal tissue incubated in 25OH however, were
significantly higher at Garnett Settlement (p=0.003), Maces Bay (p=0.007), and Saints
Rest (p=0.015) than Red Head (Figure 5). There were no significant site differences in
testosterone levels of male gonadal tissue incubated in forskolin (p=0.095) or 25-OH
(p=0.85).
Table 20: Anova table of steroid differences among sites in 2002
Female
Steroid
Estradiol
Testosterone
Sign.
0.90
0.27
Forskolin
F
0.26
1.31
DF
4,98
4,98
Sign.
<0.001
<0.001
25-OH
F
6.72
7.42
DF
4,44
4,44
Sign.
0.847
0.002
25-OH
F
0.27
7.08
DF
3,20
3,20
Male
Steroid
Testosterone
11-Ketotestosterone
Sign.
0.095
0.001
Forskolin
F
2.07
4.97
DF
4,67
4,67
55
10
A
A
Steroid Level (pg/mg)
8
A
A
A
A
6
B
AB
4
A
B
2
0
Testosterone
Estradiol
Steroid
Garnett Settlement
Saints Rest Beach
Maces Bay
Red Head
Little River Marsh
Figure 4: Total in vitro testosterone and estradiol production in female Pholis gunnellus
incubated with 25-OH in 2002 over 24hours. Each value represents the mean + SEM.
Values sharing an alphabetical script within each steroid are not statistically different
(p>0.05).
56
6
A
A
Steroid Level (pg/mg)
5
3
A
A
4
A
A
A
B
2
1
0
Testosterone
11-ketotestosterone
Steroid
Garnett Settlement
Saints Rest Beach
Maces Bay
Red Head
Figure 5: Total in vitro testosterone and 11-ketotestosterone production in male Pholis
gunnellus incubated with 25-OH in 2002 over 24 hours. Each value represents the mean
+ SEM. Values sharing an alphabetical script within each steroid are not statistically
different (p>0.05).
57
3.4.2.4 Between Incubation Media 2002
Estradiol levels of female fish collected at Maces Bay were significantly higher when
gonadal tissue was incubated in forskolin (p=0.047) than 25-OH (Figure 6; Table 21).
There were no differences in estradiol levels of fish collected at Saints Rest Beach
(p=0.10), Red Head (p=0.51), Little River Marsh (p=0.81) and Garnett Settlement
(p=0.46) when incubated in forskolin or 25-OH. In contrast, testosterone levels in female
fish collected at Red Head (p<0.001), Little River Marsh (p<0.001), Garnett Settlement
(p<0.001) and Maces Bay (p<0.001) were significantly higher when gonadal tissue was
incubated in 25-OH as opposed to forskolin (Figure 7).
There were no significant differences in gonadal testosterone and 11-ketotestosterone
levels of male fish collected at Saints Rest (p=0.53; p=0.54), Red Head (p=0.16; p=0.36),
Garnett Settlement (p=0.31; p=0.34) or Maces Bay (p=0.51; p=0.88) when incubated in
either forskolin or 25-OH.
58
Table 21: Anova table of steroid differences between incubation media in 2002
Sign.
0.10
0.51
0.81
0.46
0.047
Female
Estradiol
F
2.88
0.44
0.059
0.57
4.27
DF
1,30
1,29
1,28
1,16
1,29
Sign.
0.53
0.16
0.31
0.51
Male
Testosterone
F
0.41
2.05
1.13
0.43
DF
1,31
1,18
1,12
1,12
Steroid
Saints Rest
Red Head
Little River Marsh
Garnett Settlement
Maces Bay
Steroid
Saints Rest
Red Head
Garnett Settlement
Maces Bay
3.4.3
Sign.
0.63
<0.001
<0.001
<0.001
<0.001
Testosterone
F
0.24
15.73
30.48
24.56
35.34
DF
1,30
1,36
1,28
1,16
1,29
11-Ketotestosterone
Sign.
F
DF
0.53
0.38
1,31
0.16
0.92
1,18
0.31
0.98
1,12
0.51
0.024
1,12
Female Comparisons Between 2001 and 2002
Estradiol levels of gonadal tissue incubated in forskolin and 25-OH were significantly
higher in 2002 at Saints Rest Beach (p=0.009; p<0.001), Red Head (p<0.001; p<0.001)
and Little River Marsh (p<0.001; p<0.001) than 2001 (Table 22; Table 23).
Testosterone levels in gonadal tissue incubated in forskolin and 25-OH of female gunnel
collected at Little River Marsh were significantly higher in 2002 (p=0.008; p=0.006) than
2001 (Table 22; Table 24). There were no differences in testosterone levels in fish
collected at Saints Rest Beach (p=0.060; p=0.24) or Red Head (p=0.15; p=0.17) between
2001 and 2002.
59
Table 22: Anova table of steroid differences between 2001 and 2002
Estradiol
Saints Rest
Red Head
Little River Marsh
Sign.
0.26
<0.001
<0.001
Forskolin
F
5.26
23.54
33.51
Sign.
0.060
0.15
0.003
Forskolin
F
3.89
2.18
7.89
DF
1,24
1,40
1,37
Sign.
<0.001
<0.001
<0.001
25-OH
F
19.03
27.88
202.4
DF
1,15
1,11
1,18
25-OH
F
1.51
2.14
9.80
DF
1,15
1,12
1,18
Testosterone
Saints Rest
Red Head
Little River Marsh
DF
1,24
1,34
1,37
60
Sign.
0.24
0.17
0.006
10
A
A
A
A
A
A
Estradiol Level (pg/mg)
A
A
8
A
B
6
4
2
0
Forskolin
25-OH
Incubation Media
Garnett Settlement
Saints Rest Beach
Maces Bay
Red Head
Little River Marsh
Figure 6: Total in vitro estradiol production in female Pholis gunnellus incubated with
forskolin and 25-OH over 24 hours. Each value represents the mean + SEM. Values
sharing an alphabetical script within each media are not statistically different (p>0.05).
61
8
B
B
Testosterone Level (pg/mg)
B
6
B
4
A A
A
A
A A
2
0
Forskolin
25-OH
Incubation Media
Garnett Settlement
Saints Rest Beach
Maces Bay
Red Head
Little River Marsh
Figure 7: Total in vitro testosterone production in female Pholis gunnellus incubated in
forskolin and 25-OH over 24 hours. Each value represents the mean + SEM. Values
sharing an alphabetical script within each media are not statistically different (p>0.05).
62
Table 23: Total in vitro estradiol levels of female Pholis gunnellus
collected in Saint John Harbour
Incubation
Media
Forskolin
Site
Year
Steroid Level (pg/mg)
Saints Rest
2001
2002
3.84 + 0.20 a
7.74 + 1.28 b
2001
3.84 + 0.14 a
2002
2001
7.86 + 0.40 b
4.39 + 0.19 a
2002
8.53 + 0.54 b
2001
2.75 + 0.21 a
2002
4.88 + 0.29
2001
2002
2001
2.42 + 0.11 a
b
7.18 + 0.69
2.73 + 0.11 a
2002
8.25 + 0.52 b
Red Head
Little River
Marsh
Saints Rest
25-OH
Red Head
Little River
Marsh
a
b
Each value represents the mean + SEM. Values sharing an
alphabetical superscript within each site are not statistically
different (p>0.05).
Table 24: Total in vitro testosterone levels of female Pholis gunnellus
in Saint John Harbour
Incubation
Media
Site
Saints Rest
Forskolin
Red Head
Little River
Marsh
Saints Rest
25-OH
Red Head
Little River
Marsh
Year
Steroid Level (pg/mg)
2001
2002
2001
2002
2001
2002
2.33 + 0.12 a
a
2.87 + 0.11
2.77 + 0.22 a
2.92 + 0.87 a
a
2.69 + 0.17
b
3.08 + 0.09
2001
2002
2001
2002
2001
2002
3.61 + 0.02
2.86 + 0.14 a
3.47 + 0.85 a
a
3.79 + 0.53
3.90 + 0.22 a
5.51 + 0.76 b
a
a
Each value represents the mean + SEM. Values sharing an
alphabetical superscript within each site are not statistically
different (p>0.05).
63
3.5
Induction of Hepatic Detoxification Enzymes (EROD levels)
3.5.1
Prespawning at Little River Marsh, Red Head and Saints Rest Beach in 2001
There were no significant differences in EROD levels for either sex (p=0.33) within site
(Figure 8).
3.5.2
Pre and Post Spawning
Since EROD levels are commonly reduced in female fish as a consequence of elevated
estradiol levels, sexes were examined separetly for site differences (Whyte et al., 2000).
There were no significant differences in EROD activity of prespawning female fish at
Saints Rest (p=0.77) or Little River Marsh (p=0.59) in 2001 and fish returning in the
spring of 2002 (Figure 9; Table 25).
Prespawning male fish collected at Saints Rest during 2001 had significantly higher
EROD activity than returning fish in the spring (p=0.018) of 2002 (Figure 10; Table
25).There were no differences in EROD activity (p=0.85) of prespawning male fish
collected at Little River Marsh in 2001 and fish returning the following spring.
Table 25: Anova table pre and post spawning EROD activity
Site
Saints Rest
Little River Marsh
Sign.
0.77
0.59
Female
F
0.89
0.29
DF
1,13
1,28
64
Sign.
0.018
0.85
Male
F
6.09
0.035
DF
1,37
1,18
3000
A
EROD Activity (pmol/mg/min)
2500
A
A
A
A
2000
1500
A
1000
500
0
Male
Female
Sex
Saints Rest Beach
Red Head
Little River Marsh
Figure 8: EROD activity in male and female Pholis gunnellus in 2001. Each value
represents the mean + SEM. Values sharing an alphabetical script within each sex are
not statistically different (p>0.05).
65
2500
B
EROD Activity (pmol/mg/min)
B
2000
1500
A
A
1000
500
0
Pre
Post
Spawning
Saint Rest Beach
Little River Marsh
Figure 9: EROD activity of female Pholis gunnellus pre and post spawning. Each value
represents the mean + SEM. Values sharing an alphabetical script pre and post spawning
within site are not statistically different (p>0.05).
66
3000
A
A
EROD Activity (pmol/mg/min)
2500
A
2000
B
1500
1000
500
0
Pre
Post
Spawning
Saints Rest Beach
Little River Marsh
Figure 10: EROD activity of male Pholis gunnellus pre and post spawning. Each value
represents the mean + SEM. Values sharing an alphabetical script pre and post spawning
between sites are not statistically different (p>0.05).
67
3.5.3
Seasonal Changes in 2002
Fish were not collected from each site during each season because of periodic low catch
rates. As a result fish were collected from Little River Marsh and Saints Rest Beach
during spring, Garnett Settlement and Maces Bay during early summer, Little River
Marsh, Garnett Settlement, Red Head and Saints Rest Beach during late summer and
from Maces Bay, Red Head and Saints Rest Beach during prespawning.
3.5.3.1 EROD
Female gunnel collected at Little River Marsh during the spring had significantly higher
EROD activity than fish from Saints Rest Beach (p=0.033) (Table 26; Table 27). There
were no significant differences in EROD activity in early summer (p=0.15), late summer
(p=0.080) and prespawning (p=0.054).
Prespawning males at Saints Rest had higher EROD activity than fish collected at Maces
Bay (p=0.035) (Table 26; Table 27). There were no differences in EROD activity during
spring (p=0.23), early summer (p=0.96) or late summer (p=0.24).
Table 26: Anova table of seasonal EROD activity during 2002
Spring
Early Summer
Late Summer
Prespawning
Sign.
0.033
0.15
0.078
0.054
Female
F
5.36
2.26
2.44
3.07
DF
1,18
1,18
3,41
2,54
68
Sign.
0.023
0.96
0.24
0.035
Male
F
1.46
0.003
1.55
5.10
DF
1,30
1,13
2,24
2,45
Table 27: Mean seasonal EROD activity of male and female
Pholis gunnellus during 2002
Sex
Garnett Settlement
Saints Rest
Female
Time
EROD
Activity
Early Summer
2396 + 628 a
Late Summer
1996 + 348 a
Spring
Late Summer
Prespawning
750 + 255 a
4291 + 1351 a
1641 + 461 a
Early Summer
1492 + 268 a
Prespawning
927 + 204 a
Late Summer
Prespawning
3910 + 613 a
1383 + 261 a
Spring
1931 + 274 b
Late Summer
2786 + 806 a
Early Summer
1953 + 448 a
Late Summer
4608 + 842 a
Spring
Late Summer
Prespawning
Early Summer
Prespawning
Late Summer
1057 + 255 a
5751 + 1351 a
2428 + 461 a
1981 + 268 a
1055 + 204 b
3280 + 613 a
Prespawning
Spring
1765 + 261 a
1733 + 474 a
Late Summer
4974 + 806 a
Site
Maces Bay
Red Head
Little River Marsh
Garnett Settlement
Saints Rest
Male
Maces Bay
Red Head
Little River Marsh
a
Each value represents the mean + SEM. Values sharing
an alphabetical superscript within each site are not
statistically different (p>0.05).
69
3.6
Age Estimation
There was no difference in the age of male and female fish within a site (p=0.53)
however fish at Little River Marsh were older than fish at any other site (p=0.023) (Table
28; Figure 11).
Table 28: Anova table of age estimation among sites
Sex
Variable
Site
3.7
Sign.
0.023
F
2.88
DF
4,196
Size at Age
There was no difference in the length at age (p=0.68) or weight at age (p=0.98) of male
and female fish within a site (Table 29). There were differences in length at age and
weight at age between sites with fish at Maces Bay being longer (Figure 12) and heavier
(Figure 13) than at the other Saints Rest Beach (length: p<0.001, weight: p<0.001),
Garnett Settlement (length: p<0.001, weight: p<0.001), Red Head (length: p<0.001,
weight: p<0.001) and Little River Marsh (length: p<0.001, weight: p<0.001).
Table 29: Anova table of size at age
Weight at age
Length at age
Sign.
0.033
0.15
Female
F
5.36
2.26
DF
1,18
1,18
70
Sign.
0.023
0.96
Male
F
1.46
0.003
DF
1,30
1,13
B
5
A
Age (Years)
4
A
A
A
3
2
1
0
MB
GS
Maces Bay
Garnett Settlement
Saints Rest Beach
Little River Marsh
Red Head
SRB
LRM
RH
Site
Figure 11: Age of Pholis gunnellus in and around Saint John Harbour. Each value
represent the mean + SEM. Values sharing an alphabetical script are not statistically
different (p<0.05).
71
250
Length (mm)
200
Maces Bay: Length=18.57(age) + 77.2 (df=1,41, r2= 0.82, p<0.001)
Garnett Settlement: Length=14.41(age) + 73.6 (df=1,22, r2=0.76, p<0.01)
Saints Rest Beach: Length= 13.88(age) + 65.9 (df=1,53, r2=0.76, p<0.01)
Red Head: Length= 13.02(age) + 73.9 (df=1,56, r2=0.59, p<0.01)
Little River Marsh: Length=15.78(age)+ 66.6 (df=1,47, r2=0.77, p<0.01)
150
100
50
0+
1+
2+
3+
4+
5+
6+
7+
Age (Years)
Saints Rest Beach
Red Head
Maces Bay
Garnett Settlement
Little River Marsh
Figure 12: Total length at age of Pholis gunnellus in and around Saint John Harbour.
72
50
40
Maces Bay: Weight=4.98(age)- 5.6 (df=1,41, r2=0.79, p<0.01)
Garnett Settlement: Weight=2.21(age)- 1.2 (df=1,22, r2=0.69, p<0.001)
Saints Rest Beach: Weight=1.50(age)+ 0.09 (df=1,53, r2=0.80, p<0.01)
Red Head: Weight=1.50(age)+ 0.07 (df=1,56, r2=0.45, p<0.01)
Weight (g)
Little River Marsh: Weight=2.16(age)-1.8 (df=1,47, r2=0.68, p<0.01)
30
20
10
0
0+
1+
2+
3+
4+
5+
6+
7+
Age (Years)
Saints Rest Beach
Red Head
Maces Bay
Garnett Settlement
Little River Marsh
Figure 13: Body weight at age of Pholis gunnellus in and around Saint John Harbour.
73
4.0 DISCUSSION
4.1
General Biology
There is very little information available on rock gunnel populations. This study collected
data describing the general biology of rock gunnel at reference sites along the New
Brunswick Bay of Fundy coast. At Saint Rest Beach, fish ranged in size from 2.46 g 15.2 g and 97 mm-170 mm, with an average size of 6.2 g and 129.5 mm. Males and
females were similar in body length and weight, except for late summer when females
were slightly larger. Although both Sawyer (1967) and Qaism (1957) reported no
differences in sizes or growth rates between sexes of UK gunnel populations, Proudfoot
(1975) noted that male fish in Newfoundland tend to be larger than female fish after the
first year and reach a larger maximum size.
Proudfoot (1975) reported no difference in the condition of male and female fish.
However, male fish had a higher condition than female fish after returning from
spawning, there were no differences between sexes in late summer, and females had a
higher condition factor in the fall. The fall gonadal maturation period is accompanied by
an increase in liver size in both sexes, although gonad size does not begin to increase
until very late in the year. While Qaism (1957) reported that fish matured from June to
August in the UK, gonad size did not increase until October.
The rock gunnel appeared to move offshore in November, when gonad sizes were 1.91
and 1.59 % (body weight) in females and 0.67 and 1.57 % in males during 2001 and 2002
respectively. When the fish reappeared onshore in the Bay of Fundy, their gonads were
74
0.22 and 0.48 % body weight for males and females respectively. Adult gunnel
reappeared onshore towards the end of April. Qasim (1957) indicated that fish spawn in
the intertidal zone in January and Feburary. Sawyer (1967) reported that the reproductive
season extends from approximately November to March, starting with a move to deeper
water in late November and ending in March with the reappearance of fish onshore.
Condition decreased between pre- and post- spawning fish. Gudger (1927) suggested that
rock gunnel undergo partial starvation during spawning; however, there was no
difference in weight of pre- and post- spawning fish. Kirby et al. (1999) reported that
condition factor reaches a maximum in flounder (Platichthys flesus) just prior to the onset
of vitellogensis after the fish has been feeding during the summer months. In
post-spawning fish, females had the larger drop in condition. During spawning, both
sexes have been reported to guard egg masses during the incubation period, but only
females are thought to remain guarding as the season progresses (Proudfoot 1975). The
decreased condition associated with post-spawning fish is likely a result of sporadic
feeding and the mobilization of body energy stores associated with spawning and
possible migration. Post spawning fish also had a decreased liver size. Janssen et al.
(1995) have shown that liver size increases at the onset of, and during, vitellogensis. The
large decrease in the liver size of female Pholis gunnellus may therefore be attributed to
the lack of vitellogenin, a yolk protein, production in the liver, and the utilization of
energy stores during the guarding period.
75
Both males and females matured at an age of three years, a body size of 5.5 g and a
length of approximately 125 mm. Previous studies suggest that in the UK both sexes
begin to mature when they are approximately two years of age and between 90 to 100
mm in length (Qaism, 1957). Sawyer (1967) observed a difference in growth rate among
rock gunnel collected on the eastern and western portion of the Atlantic with fish
collected in Wales having a higher growth rate than fish collected from the Bay of Fundy.
Sawyer (1967) suggested that the warmer winter temperatures in the eastern Atlantic
could play a role in the different growth rate. However, Qasim (1957) noted that rock
gunnel varied in length with age in eastern and the western Atlantic and found
overlapping size ranges in all age groups. Thus, the difference in observed age to
maturity may be a result of a greater size range within each age group or the colder water
in the Bay of Fundy may lengthen age to maturity.
Few studies had examined physiological endpoints of rock gunnel, and no data have been
reported concerning the reproductive physiology of rock gunnel. Gonads of male and
female gunnel were incubated in forskolin and 25-OH. Because steroid levels had not
previously been reported, gonads were stimulated to try and ensure a detectable level of
steroid production. In both years, female in vitro estradiol levels were higher in tissue
incubated in forskolin while testosterone levels were higher in tissue incubated in 25-OH
mg. In 2002, in vitro levels of testosterone and 11-ketotstosterone did not differ in tissue
incubated in forskolin than 25-OH. It is unclear why there was such a large difference in
estradiol levels between 2001 and 2002, however, continued sampling may provide more
information on year to year variability.
76
4.2
Comparisons within Saint John Harbour
In 2001 and 2002, condition factor and liver size were higher at Little River Marsh in
comparison to fish from Red Head and Saints Rest Beach, and the fish tended to be
larger. Laboratory experiments indicate that exposure to various effluents often results in
increased liver size and the induction or stimulation of EROD (Adams et al., 1999; Kirby
et al., 1999). However, there were no differences in EROD activity among all sites. The
general response, however, suggests an overall increase in energy storage and utilization
at Little River Marsh. There were also differences in size and condition of juvenile rock
gunnel collected within the harbour. Increases in body size, liver size and condition have
been reported in fish exposed to industrial and municipal effluents and may result from
indirect effects of increased nutrient concentration and improved productivity (Galloway
et al, 2003).
In 2001 and 2002, gunnel from Little River Marsh were heavier and longer than gunnel
collected at Saints Rest Beach. Despite increased size, however, in 2001 there were
fewer juveniles at Little River Marsh. Catch per unit effort data for juvenile gunnel
indicate higher catch on the west as opposed to the east side of the harbour. Current
modeling indicates that currents in the harbour are predominately in the east-west
directions. To date, few studies have examined life history characteristics of juvenile
gunnel. Therefore, it is unclear if gunnel actively seek a suitable habitat or whether
dominant currents at the spawning area determine distribution.
77
While condition and liver size were elevated at Little River Marsh, the performance of
fish at Saints Rest Beach was similar to that of fish at Red Head. There were no seasonal
changes in gonad sizes between sites throughout 2002. Previous studies indicate that
contamination levels at Saint Rest Beach are lower than sites located on the eastern side
of the harbour. Brillant (1999) reported higher lead and zinc concentrations on the
eastern side of the harbour, likely associated with sewage discharge. The inner harbour is
highly industrialized while areas in the outer harbour have fewer potential sources of
hydrocarbon exposure. Brillant (1999) also indicated that for several biomonitors,
Mytilus edulis and Gammarus oceanicus, there was a trend of increasing tissue content of
PCBs from the outside toward the inside of the harbour. Prouse and Ellis (1997)
demonstrated that Nucella lapillus sampled from Red Head beach exhibited a high
frequency of imposex and a mean tissue TBT content of 9 ng/g dry weight. In 1995,
Prouse and Ellis examined Littorina littorea from Red Head for intersex. Approximately
5% of the periwinkles were sterile while intersex was not detected in samples collected
on the western side of the harbour (Brillant 1999). Delaney (2001) also indicated
increased imposex in periwinkles collected from Red Head in comparison to the western
side of the harbour.
4.3
Comparisons Among Reference Sites
While the performance of fish from Saints Rest Beach was similar to fish from Red Head,
there was a difference between the local reference sites and sites outside of Saint John
Harbour. Fish from Maces Bay and Garnett Settlement were larger and typically had
higher gonad sizes and condition than fish from Saints Rest Beach. To date no studies
78
have examined potential sources of discharge into Garnett Settlement or Maces Bay.
Potential sources of contaminants are limited at these sites, and would include runoff
from faulty septic systems and anthropogenic inputs associated with local fishing
activities.
Fish at Maces Bay and Garnett settlement reach a maximum size of 37.7 g and 205 mm
and 17.6 g and 174 mm respectively, relative to a maximum of 15.2 g and 170 mm at
Saints Rest Beach. Proudfoot (1975) reported mean weights and condition factor of fish
ranging from 2.03 g to 42.2 g and 0.25 to 0.39 respectively. Size of these fish range from
80 to 230 mm with sampling accomplished through SCUBA diving. Qasim (1957)
reported mean condition factors ranging from 0.30 to 0.43. Mean weights and condition
factors of fish from Saints Rest Beach range from 2.48 g - 13.8 g and 0.23 to 0.30 with
size ranging from 90-170 mm. Size ranges of these samples reach a maximum of 180
and 170 mm for Qasim (1957) and Sawyer (1967). Variation between Sawyer, Qasim,
this study and Proudfoot suggest that larger individuals of the populations sampled in this
study might have been available from deeper waters.
4.4
Sex Steroid levels
In 2001 and 2002, there was a difference in the response of female gonadal tissue to the
incubation buffers. During 2001, forskolin stimulated increased estradiol production
while 25-OH stimulated increased testosterone levels. During 2002 estradiol levels at
Maces Bay were higher when incubated in forskolin while female testosterone levels at
Red Head, Little River Marsh, Garnett Settlement and Maces Bay were higher when
79
incubated in 25-OH. Forskolin, a gonadotropin mimic, acts by bypassing the
gonadotropin receptor and increases cAMP, the intracellular mediator of GtH. 25-OH is
a steroid precursor which increases the amount of substrate available for steroid
conversion.
Increased estradiol levels in tissue incubated in forskolin suggest that estradiol production
in rock gunnel is regulated through an adenylate cyclase (cAMP) mechanism (Amiri et
al., 1999). Aromatase catalyzes the aromatization of androgens to estrogens and is a key
enzyme in estrogen biosynthesis. Forskolin induces steroidogensis via an adenylate
cyclase-cAMP system (Nagahama 1987). Stimulatory effects of forskolin on
intracellular accumulation of cAMP and steroid production have been reported in several
teleost fishes (Chang and Hung 1982, Nagahama 1987).
Incubating ovarian tissue in 25-OH resulted in elevated concentrations of testosterone as
compared to tissue incubated in forskolin. Cholesterol is the precursor to all steroids.
In 2001, there were no differences among sites in female testosterone or estradiol when
tissue was incubated in forskolin or 25-OH. In 2002, however, female fish from Saints
Rest Beach had lower testosterone and estradiol levels than fish at the other sites when
incubated in 25-OH. For steroid production levels, the normal conditions in rock gunnel
are largely unknown making it difficult to draw inferences about significance. More
work is required to better understand the 2002 site differences.
80
Regardless of the incubation medium, female estradiol levels at Saints Rest Beach, Red
Head and Little River Marsh were higher in 2002 than 2001. Testosterone levels in
female fish collected from Little River Marsh were higher in 2001. However, there were
no differences in testosterone levels between years at Saints Rest Beach or Red Head. In
2002, fish were collected up to four weeks earlier than fish from the previous year.
Observations in rainbow trout (Oncorhynchus mykiss) would suggest that levels of
estradiol in females during late vitellogensis should begin to drop gradually, while
testosterone levels remain high until ovulation (Scott and Sumpter, 1983). Decreased
steroid levels in 2001 may have been due to sampling at a later period in the reproductive
cycle.
4.5
EROD Activity
Although 2001 female EROD activities at Saints Rest Beach during prespawning 2001
and 2002 male and female activity at Red Head in late summer were considerably lower
than at other sites, there were no differences in male or female EROD activities at either
in-city sampling site during 2001 or 2002. The Saint John Harbour has a long history of
discharges of industrial and domestic effluents. In-city sampling sites were chosen based
on varying degrees of exposure. Treated sewage effluent is discharged at Little River
Marsh, signs of TBT exposure are evident at Red Head, and Saints Rest Beach was the
in-city reference site. Natural variability is high between the three sites and may be
responsible for the absence of a difference. To date, however, there are no published
studies which report seasonal EROD activity of this species which makes determining
induction difficult. George et al. (1995) reported EROD levels up to 1000 pmol/min/mg
81
in rock gunnel in an area affected by an oil spill in Scotland. Subsequent sampling
showed levels dropped to below 500 pmol/min/mg within four months. EROD activities
in reference areas over the same time period, however, ranged from 50-650
pmol/min/mg, indicating high natural variability. The difference in EROD activity for
fish from reference sites on the east and west coast of the Atlantic could reflect site
differences in other environmental factors such as temperature or food availability.
Induction of EROD activity after exposure to PAHs and PCBs has made this activity a
biomarker of xenobiotic exposure in fish (Sole et al., 2002). In addition to possible
inhibition of catalytic activity by certain pollutants, other factors must be taken into
consideration when using EROD activity as a bioindicator (Livingstone 1993).
A seasonal trend in EROD activity was evident in fish collected from Saints Rest Beach
throughout 2002. The enzyme activities observed in males and females collected from
Saints Rest in late summer during gonadal development in 2002 were approximately five
to six times higher than those observed from the same site in spring subsequent to
spawning. Seasonal variation of monooxygenase activity is one of several important
factors that needs to be taken into account when interpreting EROD induction in the field
(Livingstone, 1993). Seasonal changes are closely related to temperature, sexual
maturation and nutritional status. Variation in EROD activity associated with sexual
maturity in fish is well established (Bucheli and Fenet, 1995). EROD activity is affected
by reproductive state due to its role as a natural clearance mechanism for endogenous
compounds such as steroid hormones (Whyte et al., 2000).
82
This seasonal decrease may explain why the EROD activities of prespawning female fish
collected at Saints Rest as well as male and female fish collected at Little River Marsh in
2001 were not significantly different from post-spawning fish in the spring of 2002. In
addition, this decrease is evident when comparing fish collected during late summer and
prespawning at Saints Rest Beach. EROD activity of male and female fish from Saints
Rest Beach decreased by a factor of approximately 2.5 just prior to the fish moving
offshore for spawning.
In many fish species, EROD activity decreases shortly before or
during the spawning season in reproductively-active mature fish (Doyotte et al., 2001).
While there were no differences in EROD activity before and after spawning at Saints
Rest Beach and Little River Marsh, there were differences in EROD activity between the
sites when they returned post spawn. The absence of rock gunnel from intertidal areas
from late November to March suggest that rock gunnel move offshore to spawn.
Proudfoot (1975) reported that eggs masses were found at depths from 2.5 to 18 metres.
Fish from Little River Marsh may have spawned in offshore area(s) where they were
potentially exposed to PAHs and or PCBs. Ray and MacKnight (1984) reported portions
of the outer harbour, including the Bay and the mudflats outside the Courtenay Bay
breakwater, had elevated concentrations of metals attributed to industrial sources
onshore. Environmental Consultants (2001) further noted elevated metal concentrations
in the inner harbour. While metal concentrations are elevated against background
concentrations, they are below the probable effects levels of the Canadian Council of
Ministers of the Environment (CCME) marine sediment quality guidelines.
83
4.6
Adequacy of Reference Sites
The selection of adequate reference sites is potentially the most crucial aspect of an
assessment. Problems in identifying a potential reference site may be magnified in
marine environments (Munkittrick 1992). Reference sites can be very difficult to assess,
because different habitats may exist between reference and exposure areas. When
reference and exposure areas are near one another with no physical barrier separating
them, the reference site may be difficult to evaluate (Courtenay et al., 2002).
Areas of concern in selecting reference sites include the mobility of sentinel species and
the presence of confounding factors and natural variability. Following cycle 1 of the
EEM program, FSWEG (1997) recommended that selection of reference sites should
depend on the mobility of the sentinel species, that timing of sampling as fish move may
render a site unsuitable during a season, and that there is sufficient availability of the
sentiel species in sufficient numbers at the reference and exposed sites. In addition,
FSWEG (1997) recommended that studies sample multiple reference sites rather than
increasing sample sizes at one site, as a way of decreasing variability.
The similarity of size of rock gunnel at Saints Rest compared to gunnel in previous
studies, and the increased presence of contamination on the eastern portion of the
harbour, suggest that Saints Rest Beach is an appropriate local reference site. It is
unclear why fish collected outside the harbour are larger than fish collected at Saints Rest
Beach, however, variation in size along the Bay of Fundy is something that should be
further investigated. The combined action of the Saint John River and the tides of the
Bay of Fundy produce a well mixed and flushed system.
84
4.7
Recommendations
Despite previous studies which report signs of contamination at Red Head, there were
few differences between fish collected at Red Head and Saints Rest Beach. Sediment
sampling and water quality monitoring have been a common component of
environmental affects monitoring. To date, there are few studies which have done
sediment and water sampling in Saint John Harbour. While the appearance of chemicals
in sediments does not give any information on the degree of exposure or level of effects,
determining the amount of contamination in around the harbour, sediment and water
sampling in addition to fish monitoring may help explain any observed differences.
The use of a second fish species such as flounder or sculpin to determine if the same
patterns are evident would give a better indication of any effects present at harbour sites.
The capture of winter flounder (Pleuronectes americanus) was variable, and sample sizes
were too low to interpret results. Juvenile and young of the year are present in tide pools
but few adults were seen.
The use of rock gunnel in laboratory experiments could provide additional information
not available through field studies. The level of exposure and the impact of confounding
factors are often difficult to determine. However, laboratory experiments would allow
these factors to be separated as well as to measure the response to different levels and
types of contamination. Laboratory experiments would also give an indication of natural
variability in steroid levels and EROD activity, two common indicators in environmental
monitoring.
85
Rock gunnel are not resident in the area throughout year. They move offshore prior to
significant gonadal development. The inability to evaluate reproductive development,
and the reduced exposure at spawning time reduce the suitability of this species for
monitoring. However, the apparent low mobility for an extended period of time suggest
that it would give some valuable information. There have been no studies examining
rock gunnel mobility in the Bay of Fundy. Short term tag and recapture studies in
Scotland showed that recaptured fish moved a mean of 2.1 m in 1.6 days (Koop 1991).
Tagging and recapture studies could provide information regarding the amount
movement between tidal cycles and in fish returning from spawning. Tagging studies
would not determine the location of spawning areas but may help establish the mobility
of fish returning from spawning.
4.8
Conclusions
Rock gunnel are endemic to the British Isles and Northern Atlantic. In the North Atlantic
they are widely distributed and extend from Newfoundland to Massachusetts. They live
in close contact with the sediment and feed on benthic mollusks and worms, both of
which have been shown to accumulate lipophilic contaminants. In addition, rock gunnel
have a low mobility and can be easily collected under rocks and weed at low tides. While
they are easily collected during low tide, the most industrialized portion of the harbour
were either too difficult to sample or did not provide a suitable habitat for the rock
gunnel. Another difficulty in this study was obtaining sufficient sample sizes for each
period within and between sites. The large tidal fluctuations prevented the use of traps
which were found to be effective in other studies.
86
To date, few studies on the western Atlantic have used the rock gunnel as a bioindicator.
In addition, very little background information was available on the biology of rock
gunnel and this study may represent the most extensive work on this species. The
objectives of this study were to determine whether rock gunnel in Saint John Harbour
show whole animal responses in growth and reproductive development, biochemical
differences, and responses consistent with exposure to contaminants among different
sites. Rock gunnel in Saint John Harbour did not show any reproductive or biochemical
differences typically associated with contaminant exposure. There were differences in
growth, with adult and juvenile fish from Little River Marsh being larger than fish from
other in-city sampling sites. Liver size and gonad size were also larger at Little River
Marsh than other sites in Saint John Harbour. Larger sizes could be an indication of site
eutrophication due to sewage inputs at Little River Marsh. Despite the apparent problems
associated with this species further studies would minimize the challenges of sampling
efficiency.
87
5.0 References
Adams SM. 1990. Status and use of biological indicators for evaluating the effects of
stress on fish. American Fisheries Society Symposium 8: 1-8.
Adams SM, Shugart LR, Southworth GR, Hinton DE. Application of bioindicators in
assessing the health of fish populations experiencing contaminant stress. In McCarthy JF
and Shugart (eds.), Biomarkers of environmental contamination. Boca Raton, FL, Lewis
Press: 333-350.
Adams SM, Crumby WD, Greeley MS, Ryon MG, Schilling EM. 1992. Relationships
between physiological and fish population responses in a contaminated stream.
Environmental Toxicology and Chemistry 11: 1549-1557.
Adams SM, Bevelhimer MS, Greeley MS. 1999. Ecological risk assessment in a large
river-reservoir: 6. Bioindicators of fish population health. Environmental Toxicology and
Chemistry 18(4): 628-640.
Adams SM, Greeley MS. 2000. Ecotoxicological indicators of water quality: Using
multi-response indicators to assess the health of aquatic ecosystems. Water, Air and Soil
Pollution 123: 103-115.
Adams SM, Greeley MS. 2000. Evaluating effects of contamination on fish health at
multiple levels of biological organization: Extrapolating from lower to higher levels.
Human and Ecological Risk Assessment 6(1): 15-27.
Amiri BM, Maebayashi M, Adachi S, Moberg GP, Doroshov SI, Yamauchi K. 1999. In
vitro steroidogensis by testicular fragments and ovarian follicles in a hybrid sturgeon,
Bester. Fish Physiology and Biochemistry 21: 1-14.
Arukwe A, Goksoyr A. 1998. Xenobiotics, xenoestrogens and reproduction disturbances
in fish. Sarsia 83: 225-241.
Bjerselius R., Lundstedt-Enkel K, Olsen H., Mayer I., Dimberg K. 2001. Male goldfish
reproductive behaviour and physiology are severely affected by exogenous exposure to
17β-estradiol. Aquatic Toxicology 53: 139-152.
Bombail V, Aw D, Gordon E, Batty J. 2001. Application of the comet and micronucleus
assays to butterfish (Pholis gunnellus) erythrocytes from the Firth of Forth, Scotland.
Chemosphere 44: 383-392.
Brillant S. 1999. The marine food web in relation to the movement and accumulation of
toxins in Saint John Harbour, New Brunswick, Canada. M.Sc. Thesis. University of New
Brunswick. Saint John, New Brunswick. 145pp.
88
Bucheli T, Fent K. 1995. Induction of cytochrome P450 as a biomarker for
environmental contamination in aquatic ecosystems. Critical Reviews in Environmental
Science and Technology 25(3): 201-268.
Cairns J. 1981. Biological monitoring Part VI-Future needs. Water Research 15(8): 941952.
Casillas E, Mistitano D, Johnson LL, Rhodes LD, Collier TK, Stein JE, McCain BB,
Varanasi U. 1991. Inducibility of spawning and reproductive success of female English
sole (Parophrys vetulus) from urban and nonurban areas of Puget Sound, Washington.
Marine Environmental Research 31: 99-122
Ciccotelli, M, Crippa S, Colombo A. 1998. Bioindicators for toxicity assessment of
effluents from wastewater treatment plant. Chemosphere 37(14-15): 2823-2832.
Chang YS, Hung FL. 1982. The mode of action of carp gonadotropin on the stimulation
of the androgen production by carp testis in vitro. General and Comparative
Endocrinology 48: 147-153.
Colborn T, Vom-Saal FS, Soto AM. 1993. Developmental effects of endocrine disrupting
chemicals in wildlife and humans. Environmental Health Perspectives 101: 378-384.
Collier TK, Johnson LL, Stehr SM, Myers MS, Stein JE. 1998. A comprehensive
assessment of the impacts of contaminants on fish from an urban waterway. Marine
Environmental Research 46(1-5): 243-247.
Courtenay SC, Munkittrick KR, Dupuis HMC, Parker R, Boyd J. 2002. Quantifying
impacts of pulp mill effluent on fish in canadian marine and estuarine environments:
Problems and progress. Water Quality Research Journal of Canada 37(1): 79-99.
Dadswell M J. 1975. Mercury, DDT, and PCB content of certain fishes from the Saint
John River Estuary, N.B. Transactions of Canadian Society of Environmental Biology
(annual meeting): 133-146.
Delaney JL. 2001. The occurrence of imposex in Nucella lapillus within the Saint John
Harbour. Honours Thesis. University of New Brunswick. Saint John, New Brunswick.
30pp.
Doyotte A, Mitchelmore CL, Ronisz D, McEnvoys J, Livingstone D, Peters LD. 2001.
Hepatic 7-ethoxyresorufin-o-deethylase activity in Eel (Anguilla anguilla) from the
Thames Estuary and comparisons with other United Kingdom estuaries. Marine Pollution
Bulletin 42(12): 1313-1322.
Dubé M, MacLatchy DL. 2000. Endocrine responses of Fundulus heteroclitus to effluent
from a bleached-kraft pulp mill before and after installation of reverse osmosis treatment
of a waste stream. Environmental Toxicology and Chemistry 19(11): 2788-2796.
89
Efler S, Hodson PV, Wilson JY. 1998. Bioassays for measuring the potency of effluents
for inducting the activity of ethoxyresorufin-o-deethylase (EROD) in fish liver. National
Water Research Institute 47p.
Eggens M, Bergman A, Vethaak D, Van Der Weiden M, Celander M, Boon JP.1995.
Cytochrome P4501A indices as biomarkers of contaminant exposure: Results of a field
study with plaice (Pleuronectes platessa) and flounder (Platichthys flesus) from the
southern North Sea. Aquatic Toxicology 32: 211-225.
Environment Canada. 1997. Fish monitoring, Fish Survey Section 5.1 technical
guidance document for pulp and paper environmental effects monitoring. Ottawa, ON,
EEM/1997/7.
Environmental Consultants Limited. 2001. Environmental monitoring at the Black Point
ocean disposal site: Assessing long-term impacts of dredge spoil disposal in Saint John
Harbour, New Brunswick. Nova Soctia. Prepared for Enviroment Canada. 90pp.
Escher M, Wahli T, Buttner S, Meler W, Burkhardt-Holm P. 1999. The effect of sewage
plant effect on brown trout: A cage experiment. Aquatic Science 61: 93-110.
Eufemia NA, Collier TK, Stein JE, Watson DE, Di Giulo T. 1997. Biochemical
responses to sediment-associated contaminants in brown bullhead (Ameiurus nebulosus)
from the Niagara River ecosystem. Ecotoxicology 6: 13-34.
Everaarts JM, Shugart LR, Gustin MK, Hawkins WE, Walker WW. 1993. Biological
markers in fish: DNA integrity, hematological parameters and liver somatic index.
Marine Environmental Research 35: 101-107.
Fenet H, Casellas C, Bontoux J. 1998. Laboratory and field - caging studies on hepatic
enzymatic activities in European eel and rainbow trout. Ecotoxicology and
Environmental Safety 40: 137-143.
Flammarion P, Garric J. 1997. Cyprinids EROD activities in low contaminated rivers: A
relevant statistical approach to estimate reference levels for EROD biomarker.
Chemosphere 35(10): 2375-2388.
FSEWG. 1997. Fish survey expert working group: Recommendations from cycle 1
review. Environment Canada, Ottawa, Ont., EEM/1997/6. 262p.
Galloway B J, Munkittrick KR, Currie S, Gray MA, Curry RA, Wood C. 2002.
Examination of the responses of slimy sculpin (Cottus cognatus) and white sucker
(Catostomus commersoni) collected on the Saint John River downstream of pulp mill,
paper mill, and sewage discharges. Environmental Toxicology and Chemistry in press.
90
George SG, Wright J, Conroy J. 1995. Temporal studies of the impact of the Braer
oilspill on inshore feral fish from Shetland, Scotland. Archives of Environmental
Contamination and Toxicology 29: 530-534.
Goldfarb P, Livingstone D, Birmelin C. 1998. Biomonitoring in the aquatic
environment: use of molecular biomarkers. Biochemical Society Transactions 26(4):
690-694.
Gray M, Curry A, Munkittrick KR. 2002. Non-lethal sampling methods for assessing
environmental impacts using a small-bodied sential fish species. Water Quality Research
Journal of Canada 37(1): 195-211.
Godfrey and Associates. 1993. Saint John Harbour Wastewater Strategy. Saint John,
New Brunswick. 86pp.
Gudger EW. 1927. The nest and nesting habits of the butterfish or gunnel, Pholis
gunnellus. Natural History 27: 65-71.
Harries JE, Sheahan DA, Jobling S, Matthiessen P, Neall P, Routledge EJ, Rycroft R,
Sumpter JP, Taylor T. 1996. A survey of estrogenic activity in United Kingdom inland
waters. Environmental Toxicology and Chemistry 15(11): 1993-2002.
Harris JH. 1995. The use of fish in ecological assessments. American Journal of
Ecology 20: 65-80.
Hashimoto S, Bessho H, Hara A, Nakamura M, Iguchi T, Fujita K. 2000. Elevated serum
vitellogenin levels and gonadal abnormalities in wild male flounder (Pleuronectes
yokohamae) from Tokyo Bay, Japan. Marine Environmental Research 49(1): 37-53.
Janssen PAH, Lambert JGD, Goos HGT. 1995. The ovarian cycle and the influence of
pollution on vitellogensis in flounder, Platichthys flesus (L.). Journal of Fish Biology 47:
509-523.
Jimenez B, Stegeman JJ. 1990. Detoxification enzymes as an indicator of environmental
stress of fish. American Fisheries Society Symposium 8: 67-79.
Jobling S, Nolan M, Tyler CR, Brighty G, Sumpter JP. 1998. Widespread sexual
disruption in wild fish. Environmental Science and Technology 32: 2498-2506.
Johnson LL, Casillas E, Collier TK, McCain BB, Varanasi U. 1988. Contaminant effects
on ovarian development in English sole (Parophrys vetulus) from Puget Sound,
Washington. Canadian Journal of Fisheries and Aquatic Sciences 45: 2133-2146.
Karels AE, Soimasuo M, Lappivaara J, Leppanen H, Aaltonen T, Mellanen PJ Oikari
AOJ. 1998. Effects of ECF-bleached kraft mill effluent on reproductive steroids and liver
MFO activity in populations of perch and roach. Ecotoxicology 7: 123-132.
91
Karels AE, Soimasuo M, Oikari A. 1999. Effects of pulp and paper mill effluents on
reproduction, bile conjugates and liver MFO (mixed function oxygenase) activity in fish
at southern lake Saimaa, Finalnd. Water Science Technology 40(11-12): 109-114.
Kime DE. 1995. The effect of pollution on reproduction in fish. Review in Fish Biology
and Fisheries 5: 52-96.
Kime DE. 1999. A strategy for assessing the effects of xenobiotics on fish reproduction.
The Science of the Total Environment 25: 3-11.
Kirby MF, Matthiessen P, Neall P, Tylor T, Allchin CR, Kelly CA, Maxwell DL, Thain
JL. 1999. Hepatic EROD activity in flounder (Platichthys flesus) as an indicator of
contaminant exposure in English estuaries. Marine Pollution Bulletin 38(8): 676-686.
Koop JH, Gibson RN. 1991. Distribution and movements of intertidal butterfish Pholis
gunnellus. Journal of Marine Biological Association of the United Kingdom 71: 127-136.
Kruuk H, Nolet B, French D. 1988. Fluctuations in numbers and activity of inshore
demersal fishes in Shetland. Journal of Marine Biological Association of the United
Kingdom 68: 601-617.
Larsson DGJ, Forlin L. 2002. Male-biased sex ratios of fish embryo near a pulp mill.
Temporary recovery after a short-term shutdown. Environmental Health Perspectives
110(8): 739-742.
LeBlanc K. 1998. Effects on reproductive endocrine function of mummichog (Fundulus
heteroclitus) and goldfish (Carassius auratus) exposed to Saint John, New Brunswick
harbour waters. M.Sc. Thesis. University of New Brunswick. Saint John, New
Brunswick. 118pp.
Livingstone D. 1993. Biotechnology and pollution monitoring: Use of molecular
biomarkers in the aquatic environment. Journal of Chemistry, Technology and
Biotechnology 57: 195-211.
Mayer FL, Versteeg DJ, McKee MJ, Folmar LC, Graney RL, McCume DC, Rattner BA.
1992. Physiological and nonspecific biomarkers. In: Hugget RJ, Kimerle RA, Mehrle
PM, Bergman HL (eds.), Biomarkers: Biochemical, physiological, and histological
markers of anthrogenic stress. Boca Raton, FL, Lewis Press: 5-61.
McCarthy JF, Jimenez BD, Shugart LR, Sloop FV. 1990. Biological markers in animal
sentinels: Laboratory studies improve interpretation of field data. In: Evaluations of
Biological Hazards of Environmental Polluntants. S. S. e. a. Sandhu. New York, Plenum
Press: 163-175.
92
McCarty LS, Munkittrick KR. 1996. Environmental biomarkers in aquatic toxicology:
Fiction, fantasy or fictional? Human and Ecological Risk Assessment 2(2): 268-274.
McMaster ME, van der Kraak GJ, Portt CB, Munkittrick KR, Sibley PK, Smith IR,
Dioxin DG. 1991. Changes in hepatic mixed-function oxygenase (MFO) activity,
plasma steroid levels, and age at maturity of a white sucker (Catostomus commersoni)
population exposed to bleached kraft pulp mill effluent. Aquatic Toxicology 21:199-218
McMaster, M.E., K.R. Munkittrick, J.J. Jardine, R.D. Robinson, G.J. VanDer Kraak.
1995. Protocol for measuring in vitro steroid production by fish gonadal tissue. 1961.
Canadian Technical Report of Fisheries and Aquatic Sciences. Fisheries and Oceans
Canada, Burlington, ON, Canada.
McMaster ME, Frank M, Munkittrick KR, Riffon R, Wood CS. 2002. Follow-up studies
addressing questions identified during cycle 1 of the adult fish survey of the pulp and
paper EEM program. Water Quality Research Journal of Canada 37(1): 133-153.
Munkittrick KR, Van Der Kraak GJ, McMaster ME, Portt CB. 1992. Reproductive
dysfunction and MFO activity in three species of fish exposed to bleached kraft mill
effluent at Jackfish Bay, Lake Superior. Water Pollution Research Journal of Canada
27(3): 439-446.
Munkittrick KR. 1992. A review and evaluation of study design considerations for sitespecifically assessing the health of fish populations. Journal of Aquatic Ecosystem Health
1: 283-293.
Munkittrick KR, Van Der Kraak GJ, McMaster ME, Portt CB. 1992. Response of
hepatic MFO activity and plasma sex steroids to secondary treatment of bleached kraft
pulp mill effluent and mill shutdown. Environmental Toxicology and Chemistry 11:
1427-1439.
Munkittrick KR, McGeachy A, McMaster ME, Courtenay SC. 2002. Overview of
freshwater fish studies from the pulp and paper environmental effects monitoring
program. Water Quality Research Journal of Canada 37(1): 49-77.
Nagahama Y. 1987. Gonadotropin action on gametogensis and steroidogensis in teleost
gonads. Zoological science 4: 209-222.
Nagahama Y. 1987. 17α,20β-Dihydroxy-4-pregnen-3-one: A teleost maturation-inducing
hormone. Development, Growth and Differentiation 29: 1-12.
Oberdorster E, Cheek AO. 2000. Gender benders at the beach: Endocrine disruption in
marine and estuarine organisms. Environmental Toxicology and Chemistry 20(1): 23-36.
93
Pacheco M, Santos MA. 1998. Induction of liver EROD and erythrocytic nuclear
abnormalities by cyclophosphamide and PAHs in Anguilla anguilla L. Ecotoxicology and
Environmental Safety 40: 71-76.
Payne JF, Penrose WR. 1975. Induction of aryl hydrocarbon (Benzo[a]pyrene)
hydroxylase in fish by petroleum. Bulletin of Environmental Contamination and
Toxicology 14(1): 112-116.
Payne JF. 1976. Field evaluation of benzopyrene hydroxylase induction as a monitor for
marine pollution. Science 191: 945-946.
Powers D. 1989. Fish as model systems. Science 246: 352-358.
Proudfoot L A 1975. The biology of the rock gunnel Pholis gunnellus (Linnaeus).
Honours thesis. Memorial University of Newfoundland. St. John’s, Newfoundland.
40pp.
Prouse N J, Ellis DV. 1997. A baseline survey of dogwhelk (Nucella Lapillus) imposex
in eastern canada and interpretation in terms of tributltin (TBT) contamination.
Environmental Technology 18: 1255-1264.
Purdom CE, Hardiman PA, Bye VJ, Taylor NC, Sumpter JP. 1994. Estrogenic effects of
effluents from sewage treatment works. Chemical Ecology 8:275-285.
Qasim S. 1956. The biology of Blennius pholis L. (Teleostei). Proceedings of the
Zoological Society of London 128: 161-208.
Qasim S. 1957. The biology of Centronotus gunnellus (L.) (Teleostei). Journal of Animal
Ecology 26(2): 389-401.
Ray S, MacKnight SD. 1984. Trace metal distributions in Saint John Harbour. Marine
Pollution Bulletin 15(1): 12-18.
Redding MJ, Patino R. 1993. Reproductive physiology. In Evans DH, ed, The
Physiology of Fishes. Marine Science Series. CRC, Boca Raton, FL, USA, pp 503-534.
Rodgers-Gray T P, Jobling S, Morris S, Kelly CA, Kirby S, Janbakhsh A, Harries JE,
Waldock MJ, Sumpter JP, Tyler CR. 2000. Long-term temporal changes in the estrogenic
composition of treated sewage effluent and its biological effects on fish. Environmental
Science and Technology 34: 1521-1528.
Routledge EJ, Sheahan D, Desbrow C, Brighty G, Waldock MJ, Sumpter JP. 1998.
Identification of estrogenic chemicals in STW effluent. 2. In vivo responses in trout and
perch. Environmental Science and Technology 32: 1559-1565.
94
Sawyer P, 1967. Intertidal life-history of the rock gunnel, Pholis gunnellus, in the
Western Atlantic. Copeia 1967: 55-61.
Schlenk D, Perkins EJ, Hamilton G, Zhang YS, Layher W. 1996. Correlation of hepatic
biomarkers with whole animal and population-community metrics. Canadian Journal of
Fisheries and Aquatic Sciences 53: 2299-2309.
Schlenk D. 1999. Necessity of defining biomarkers for use in ecological risk assessment.
Marine Pollution Bulletin 39(1-12): 48-53.
Scott AP, Sumpter JP. 1983. A comparison of the female reproductive cycles of
autumn-spawning and winter-spawning strains of rainbow trout (Salmo gairdneri
Richardson). Gen. Comp. Endocrinol. 52: 79-85.
Scott WB, Scott MG. 1988. Atlantic Fishes of Canada. Canadian Bulletin of Fisheries
and Aquatic Sciences 219: 713 p.
Sole M, Barcelo D, Porte C. 2002. Seasonal variation of plasmatic and hepatic
vitellogenin and EROD activity in carp, Cyprinus carpio, in relation to sewage treatment
plants. Aquatic Toxicology 60: 233-248.
Sulaiman N, George S, Burke MD. 1991. Assessment of sublethal pollutant impact on
flounders in an industrialized estuary using hepatic biochemical indices. Marine Ecology
Progress Series 68: 207-212.
Thomas P. 1990. Molecular and biochemical responses of fish to stressors and their
potential use in environmental monitoring. American Fisheries Society Symposium 8: 928.
van den Heuvel MR, Power M, MacKinnon MD, Dioxin DG. 1999. Effects of oil sands
related aquatic reclamation on yellow perch (Perca flavescens). II. Chemical and
biochemical indicators of exposure to oil sands related waters. Canadian Journal of
Fisheries and Aquatic Sciences 56: 1226-1233.
Van Der Kraak GJ, Munkittrick KR, McMaster ME, Portt CB, Chang JP. 1992. Exposure
to bleached kraft mill effluent disrupts the pituitary-gonadal axis of white sucker at
multiple sites. Toxicology and Applied Pharmacology 115: 224-233.
Walker SL, Hedley K, Porter EL. 2002. Pulp and paper environmental effects
monitoring in Canada: An overview. Water Quality Research Journal of Canada 37(1):
7-19.
Washburn and Gillis Associates Ltd. 1993. Saint John Harbour Environmental Quality
Study. Fredericton, New Brunswick.
95
Whyte JJ, Jung RE, Schmitt CJ, Tillitt DE. 2000. Ethoxyresorufin-o-deethylase (EROD)
activity in fish as a biomarker of chemical exposure. Critical Reviews in Toxicology
30(4): 347-570.
Williams TG, Lockhart WL, Metner DA, Harbicht S. 1997. Baseline studies in the Slave
River, NWT, 1990-1994: Part III. MFO enzyme activity in fish. The Science of the Total
Environment 197: 87-109.
96
6.0 Appendix
1. Hepes Buffer
26.03 g Hepes in 1L of water, cold pH adjusted to 7.8 with 50% concentrated HCl
2. Hepes Grinding Buffer
11.184 g KCl and 5.206 g Hepes Buffer in 1L of water, cold pH adjusted to 7.5 with 50%
concentrated HCl
3. 7ER
~0.022 mg/mL in DMSO. Measure the absorbance at 461.5 nm and dilute until the
absorbance reads between 1.60 and 1.70 absorbance units.
4. 7ER/Hepes Buffer
550 µL 7ER with 4550 µL Hepes buffer
5. Resorufin Standards
A superstock of resorufin is prepared with 5.0 mg resorufin per mL DMSO. A stock
solution of resorufin is prepared by diluting 1 mL superstock in 9 mL DMSO.
97
Was this manual useful for you? yes no
Thank you for your participation!

* Your assessment is very important for improving the work of artificial intelligence, which forms the content of this project

Download PDF

advertising