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Cytochrome P450 1A as a biomarker of contaminant exposure in free-ranging marine mammals Miller, Kelsey A. 2003

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C Y T O C H R O M E P450 1A AS A BIOMARKER OF CONTAMINANT EXPOSURE IN FREE-RANGING MARINE M A M M A L S by KELSEY A. MILLER B.Sc, University of British Columbia, Canada, 1997  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF T H E REQUIREMENTS FOR T H E DEGREE OF MASTER OF SCIENCE in T H E F A C U L T Y OF GRADUATE STUDIES F A C U L T Y OF PHARMACEUTICAL SCIENCES  We accept this thesis as conforming to the required standard  T H E UNIVERSITY OF BRITISH COLUMBIA March 2003 © Kelsey A. Miller, 2003  In presenting  this  thesis in partial  degree at the University of  fulfilment of  the  requirements  for  an advanced  British Columbia, 1 agree that the Library shall make it  freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of department  or  by  his  or  her  representatives.  It  is  understood  that  copying  my or  publication of this thesis for financial gain shall not be allowed without my written permission.  Department of  PHA^ACfeuiTicAt SCi6/vCS<>  The University of British Columbia Vancouver, Canada Date  DE-6 (2/88)  /WAfrlrl  )  ^005  .  ABSTRACT  Marine mammals are exposed to high concentrations  of organochlorine  contaminants that have been linked to adverse health effects. Hepatic cytochrome P450 1A (CYP1A) is a widely used biomarker of organochlorine exposure. C Y P 1 A catalyzes the biotransformation of xenobiotic compounds and is specifically induced by planar aromatic and halogenated hydrocarbons.  C Y P 1 A analysis has traditionally involved  harvesting liver samples from dead animals.  However, because of legal and ethical  constraints, the use of liver for biomarker studies in free-ranging marine mammals has become increasingly unacceptable. The objective of this study was to determine whether C Y P 1 A in skin biopsies, obtained using minimally-invasive techniques, could be used as a biomarker of organochlorine exposure in wild harbour seals and killer whales. This study consisted of three groups: (1) 20 free-ranging harbour seal pups were captured in southern British Columbia (BC) and temporarily housed in captivity. Skin-blubber and liver biopsies were collected, and three seals were orally treated with B-naphthoflavone (BNF), a known C Y P 1 A inducer. (2) Skin-blubber biopsies were collected in the field from 42 seals (pups and adults) in B C and Washington State. (3) Skin-blubber biopsies were collected from 13 free-ranging killer whales in BC. C Y P 1 A enzyme activity and protein levels were quantified in both liver and skin biopsies from seals using the ethoxyresorufin O-deethylase (EROD) and immunoblot assays, respectively. Cutaneous C Y P 1 A expression was near the detection limit and did not correlate with hepatic •r  CYP1A.  However, in both tissues, C Y P 1 A protein levels were induced by B N F  treatment, and C Y P 1 A expression increased during three weeks in captivity. Hepatic  ii  C Y P 1 A expression correlated with blubber contaminant levels, and cutaneous C Y P 1 A protein levels were higher in pups from Washington State than in pups from B C . EROD activity was not detected in killer whale skin, though a possible C Y P 1 A protein band was detected by immunoblot analysis. In this study, C Y P 1 A was quantified in small liver and skin biopsies obtained from free-ranging marine mammals using minimally-invasive methods.  However, further studies are needed to validate the use of C Y P 1 A as a  biomarker of organochlorine exposure in marine mammal skin.  iii  T A B L E OF CONTENTS  Page Title Page Abstract Table of Contents List of Figures List of Tables List of Abbreviations Appendices Acknowledgements 1.  :  :  INTRODUCTION 1.1 Persistent Organic Pollutants 1.1.1 POPs in the marine environment 1.1.2 POPs in marine mammals 1.1.3 Toxicity of organochlorines 1.1.3.1 The Aryl hydrocarbon receptor 1.1.3.2 Metabolism of organochlorines 1.2 Biomarkers 1.2.1 Cytochrome P450 1A 1.2.1.1 Background 1.2.1.2 C Y P 1 A as a biomarker of organochlorine exposure 1.2.1.3 C Y P l A i n s k i n 1.2.1.4 Species specificity of C Y P 1 A 1.2.1.5 Natural factors affecting C Y P 1 A expression 1.3 Study Overview 1.3.1 Description of study species 1.3.1.1 Harbour seals (Phoca vitulina) 1.3.1.2 Killer whales (Orcinus orca) 1.3.2 Challenges associated with wildlife toxicology studies 1.3.3 Overall goals of the present study 1.4 Research Hypothesis 1.5 Experimental Hypotheses 1.6 Specific Objectives  2.  M A T E R I A L S A N D METHODS 2.1 Overview 2.2 Chemicals 2.3 Sample Collection ! 2.3.1 Study 1: Captive harbour seals (Fraser River pups) 2.3.1.1 Animal capture and care 2.3.1.2 Tissue biopsies 2.3.1.3 B-naphthoflavone treatment  i ii iv viii xi xiii xiv xv 1 1 3 4 7 8 12 14 15 15 17 20 22 24 26 26 26 27 27 30 31 31 32 33 33 33 37 38 38 39 40  iv  2.3.2 Study 2: Free-ranging harbour seals 2.3.3 Study 3: Free-ranging killer whales 2.4 Tissue Preparation 2.4.1 Liver 2.4.2 Skin 2.5 Total Protein Determination 2.6 Enzyme Activity 2.6.1 Determination of resorufin stock concentration 2.6.2 Antibody inhibition of EROD activity 2.6.3 B N F inhibition 2.7 SDS-PAGE and Immunoblots 2.8 Quantitation of Immunoblots 2.9 Contaminant Analysis 2.10 Statistical Analysis 2.11 Assay Validation and Optimization 2.11.1 EROD Assay 2.11.1.1 Intra-assay variation 2.11.1.2 Inter-assay variation 2.11.1.3 Substrate concentration for the EROD assay 2.11.1.4 Total protein concentration for the EROD assay 2.11.1.5 N A D P H concentration for the EROD assay 2.11.1.6 Effect of skin biopsy size on total protein yield and EROD activity 2.11.1.7 Reaction time 2.11.2 Immunoblot Assay 2.11.2.1 General protocol for the immunoblot assay 2.11.2.2 Total protein concentration for the immunoblot assay 2.11.2.3 Cross-reactivity of antibodies with harbour seal C Y P 1 A 2.11.2.4 Protein transer from gel to membrane 2.11.2.5 Intra-assay variation 2.11.2.6 Inter-assay variation RESULTS 3.1 Study 1: Captive Harbour Seal Study 3.1.1 Induction of C Y P 1 A by in vivo BNF treatment 3.1.1.1 Immunoblot analysis 3.1.1.2 EROD activity 3.1.1.3 B N F inhibition of EROD activity 3.1.2 C Y P 1 A expression in tissue biopsies obtained two days post-capture 3.1.2.1 Liver 3.1.2.2 Skin 3.1.2.3 Comparison of C Y P 1 A expression in skin and liver 3.1.2.4 Effect of body weight on C Y P 1 A expression 3.1.2.5 Effect of sex on C Y P 1 A expression 3.1.2.6 Effect of contaminant levels in blubber on C Y P 1 A expression 3.1.2.7 Relationship between body weight and contaminant levels in blubber 3'. 1.3 Increase in C Y P 1 A expression over time in captive harbour seals  41 42 44 44 44 46 49 51 51 52 52 54 55 55 55 56 56 56 59 60 62 62 64 66 66 66 67 67 68 68 71 71 71 71 75 75 78 78 83 86 86 90 90 94 94  3.1.3.1 C Y P 1 A protein levels 3.1.3.2 EROD activity 3.1.4 Antibody inhibition of EROD activity 3.2 Study 2: Field Study of Free-Ranging Harbour Seals 3.2.1 Hornby Island harbour seal pups 3.2.1.1 C Y P 1 A expression inskin 3.2.1.2 Effect of body weight on C Y P 1 A expression in skin 3.2.1.3 Effect of sex on C Y P 1 A expression in skin 3.2.1.4 Effect of contaminant levels in blubber on CYP1A expression 3.2.1.5 Relationship between body weight and contaminant levels in blubber... 3.2.2 Puget Sound harbour seal pups 3.2.2.1 C Y P 1 A expression in skin 3.2.2.2 Effect of body weight on C Y P 1 A expression in skin 3.2.2.3 Effect of sex on C Y P 1 A expression in skin 3.2.3 Adult harbour seals from British Columbia 3.2.3.1 C Y P 1 A expression in skin 3.2.3.2 Effect of body weight on C Y P 1 A expression in skin 3.2.3.3 Effect of sex on C Y P 1 A expression in skin 3.3 Harbour Seal Inter-Population Comparisons 3.3.1 Relationship between C Y P 1 A protein levels and EROD activity 3.3.2 Effect of body weight on C Y P 1 A expression 3.3.3 Variability of cutaneous C Y P 1 A expression between pups and adults 3.3.4 Effect of sex on C Y P 1 A expression 3.3.5 Comparison of C Y P 1 A expression in pups from B C and Puget Sound 3.3.6 Effect of contaminant levels in blubber on CYP1A expression 3.4 Study 3: Study of Free-Ranging Killer Whales 3.4.1 C Y P 1 A expression in killer whale skin DISCUSSION 4.1 Study 1: Captive Harbour Seal Study 4.1.1 Analysis of liver samples obtained two days post-capture 4.1.1.1 Liver biopsy approach 4.1.1.2 EROD activity 4.1.1.3 Immunoblot analysis 4.1.2 Analysis of skin samples obtained two days post-capture 4.1.3 Comparison of C Y P 1 A expression levels between liver and skin 4.1.4 Induction of C Y P 1 A by in vivo BNF treatment 4.1.5 Relationship between environmental contaminant levels and C Y P 1 A expression 4.1.6 Relationship between age and contaminant concentrations in blubber 4.1.7 Relationship between body weight (age) and CYP1A expression in harbour seals 4.1.8 Relationship between sex and C Y P 1 A expression in harbour seals 4.1.9 Increase in C Y P 1 A expression over time in captive harbour seals 4.1.9.1 C Y P 1 A development in young seals  95 97 98 102 102 102 104 104 104 104 106 106 108 108 108 108 108 108 111 Ill 111 111 112 112 112 114 114 115 116 116 116 116 118 119 119 123 124 127 128 129 130 130  4.1.9.2 Factors affecting variability of C Y P 1 A expression within and between tissues 4.1.9.3 Possible effects of stress and captivity-related factors on C Y P 1 A expression 4.2 Study 2: Field Study of Free-Ranging Harbour Seals 4.2.1 Inter-population comparisons 4.2.1.1 Relationship between environmental contaminant levels in blubber and cutaneous C Y P 1 A expression 4.2.1.2 Relationship between physiological factors and cutaneous C Y P 1 A expression 4.2.2 Significance of biomarker studies in free-ranging harbour seals 4.3 Study 3: Free-Ranging Killer Whales 4.4 Improvement of C Y P 1 A Detection in Skin 4.5 Effects of Storage Conditions on C Y P 1 A 4.6 Conclusions 4.7 Summary and Future Studies  131 133 135 135 135 135 136 137 138 140 143 144  5.  REFERENCES  147  6.  APPENDICES  164  vii  LIST OF FIGURES Page  Figure 1.1  Organochlorine chemical structures  2  Figure 1.2  The C Y P cycle  16  Figure 1.3  Simplified mechanism of C Y P 1A induction  18  Figure 1.4  'Weight of evidence approach' used to assess the risk of toxic injury in free-ranging marine mammals  29  Figure 2.1  Marine mammal sampling sites in British Columbia and Washington State  37  Figure 2.2  Study 1: Captive harbour seals  38  Figure 2.3  Chemical structure of B-naphthoflavone  40  Figure 2.4  Schematic for preparation of S9 fractions from seal liver and skin  47  Figure 2.5  EROD reaction  50  Figure 2.6  Inter-assay calibration curve for the EROD assay  58  Figure 2.7  Effect of substrate concentration on resorufin formation in harbour seal liver Effect of total protein concentration on resorufin formation in harbour seal liver  59  Figure 2.8 Figure 2.9  Figure 2.10  Figure 2.11  60  Effect of total protein concentration on resorufin formation in harbour seal skin  61  Effect of biopsy size and buffer dilution factor on resorufin formation in harbour seal skin  63  Effect of biopsy size on total protein yield for harbour seal skin  63  Figure 2.12  Resorufin formation over time in harbour seal liver and skin  65  Figure 2.13  Inter-assay calibration curve for immunoblots  69  Figure 3.1  Figure 3.2  Figure 3.3  Figure 3.4  Figure 3.5  Immunoblot of BNF-treated and control harbour seal pup liver samples probed with polyclonal rabbit anti-rat CYP1A2 serum  72  Immunoblot of BNF-treated and control harbour seal pup skin samples probed with polyclonal rabbit anti-rat CYP1A2 serum  72  Immunoblot of both BNF-treated liver and skin samples probed with polyclonal rabbit anti-rat CYP1A2 serum  73  Effect of oral B N F treatment on C Y P 1 A protein levels in liver and skin of captive harbour seals  74  Effect of oral B N F treatment on EROD activity in liver and skin of captive harbour seals  76  Figure 3.6  Inhibition of hepatic EROD activity by in vitro B N F treatment  77  Figure 3.7  Immunoblot of captive harbour seal pup liver S9 fractions probed with polyclonal rabbit anti-rat C Y P 1A2 serum Relationship between hepatic C Y P 1 A protein levels and EROD activity for Fraser River harbour seal pups  Figure 3.8  Figure 3.9  Figure 3.10  Figure 3.11  Figure 3.12  Figure 3.13  Figure 3.14  Figure 3.15  Figure 3.16  80 82  Immunoblot of captive harbour seal pup skin S9 fractions probed with polyclonal rabbit anti-rat CYP1A2 serum  85  Relationship between C Y P 1 A protein levels and EROD activity in skin of Fraser River harbour seal pups  85  Relationship between C Y P 1 A expression in liver and skin of Fraser River harbour seal pups  87  Relationship between body weight at capture and hepatic EROD activity  88  Relationship between body weight and cutaneous EROD activity and C Y P 1 A protein in Fraser River harbour seal pups  89  Relationship between hepatic EROD activity and contaminants in blubber  91  Relationship between hepatic C Y P 1A protein levels and contaminants in blubber  92  Relationship between cutaneous C Y P 1 A expression and contaminants in blubber  93  Figure 3.17  Figure 3.18  Figure 3.19  Figure 3.20  Figure 3.21  Figure 3.22  Figure 3.23  Figure 3.24  Figure 3.25  Figure 3.26  Figure 3.27  Figure 3.28  Figure 3.29  Total TEQ for blubber was negatively correlated with body Weight in Fraser River harbour seal pups  94  Immunoblot of harbour seal pup liver S9 fractions probed with polyclonal rabbit anti-rat CYP1A2 serum from biopsies obtained 2 days post-capture and after 3 weeks in captivity  95  C Y P 1 A protein levels increased in liver and skin of captive harbour seals during 3 weeks in captivity  96  EROD activity increased in liver and skin of captive harbour seals during 3 weeks in captivity  97  Antibody inhibition of EROD activity in harbour seal liver and skin  99  Immunoblot of Hornby Island harbour seal pup skin S9 fractions probed with polyclonal rabbit anti-rat CYP1A2 serum  103  Relationship between blubber total TEQ and cutaneous C Y P 1 A protein levels in Hornby Island harbour seal pups  105  Correlation between body weight and total TEQ for blubber in Hornby Island harbour seal pups  105  Immunoblot of Puget Sound harbour seal pup skin S9 fractions probed with polyclonal rabbit anti-rat CYP1A2 serum  106  Immunoblot of adult harbour seal skin S9 fractions probed with polyclonal rabbit anti-rat C Y P 1A2 serum  110  Relationship between C Y P 1 A protein levels and EROD activity in skin of harbour seal pups from all populations  111  Comparison of cutaneous C Y P 1 A protein levels between harbour seal pups from British Columbia and Puget Sound, Washington  113  Immunoblot of killer whale skin S9 fractions probed with polyclonal rabbit anti-rat C Y P 1A2 serum  114  X  LIST OF T A B L E S  Page Table 1.1  Organochlorine levels in blubber of pinnipeds and cetaceans from various locations  6  Table 1.2  Toxic Equivalency Factors  11  Table 1.3  Summary of the most common methods used to measure C Y P 1 A induction  19  Table 2.1  Buffers and reagents used in this study  36  Table 2.2  Summary of marine mammal samples used in this study  43  Table 2.3  Mean total protein yields obtained from liver and skin biopsies  48  Table 2.4  Reagents and volumes used in microtiter plates for the EROD assay Amount of total protein used per sample for the EROD and immunoblot assays  Table 2.5  Table 2.6  Table 2.7  Table 2.8  Table 2.9  Table 3.1  Table 3.2  Table 3.3  50 54  Intra-assay variation of fluorescence readings for EROD assay standards  57  Inter-assay variation of fluorescence readings for EROD assay standards  58  Intra-assay variation of relative contour quantities for immunoblot standards "  68  Inter-assay variability of relative contour quantity values for immunoblot standards  69  C Y P 1 A expression in hepatic S9 fractions from Fraser River harbour seal pups  79  C Y P 1 A expression in cutaneous S9 fractions from Fraser River harbour seal pups  84  Comparison of CYP1A expression between male and female harbour seal pups from the Fraser River estuary  90  Table 3.4  Cutaneous C Y P 1A expression in Hornby Island harbour seal pups  102  Table 3.5  Cutaneous C Y P 1A expression in Puget Sound harbour seal pups  107  Table 3.6  Cutaneous C Y P 1A expression in adult harbour seals  109  xii  LIST OF A B B R E V I A T I O N S  AHH AhR AP BCIP BNF BPMO DDT DMSO EROD HEPES LOD LOQ MC NADPH NBT OC PAH PBS PCB PCDD PCDF POP PVDF RCQ RT-PCR SDS TEF TEMED TEQ  aryl hydrocarbon hydroxylase aryl hydrocarbon receptor ammonium persulphate 5-bromo-4-chloro-3-indoyl phosphate,/7-toluidine salt B-naphthoflavone benzo[a]pyrene monooxygenase dichlorodiphenyltrichloroethane dimethylsulfoxide ethoxyresorufin O-deethylase N-2-hydroxyethylpiperazine-N-2-ethanesulfonic acid limit of detection limit of quantitation methylcholanthrene nicotindiamide adenine dinucleotide phosphate tetrasodium salt nitro blue tetrazolium organochlorine polycyclic aromatic hydrocarbon phosphate buffered saline polychlorinated biphenyl polychlorinated dibenzo-p-dioxin polychlorinated dibenzofuran persistent organic pollutant polyvinyldenfluoride relative contour quantity reverse transcription polymerase chain reaction sodium dodecyl sulphate toxic equivalency factor N ,N ,N ' N ' -tetramethy lenediamine toxic equivalent  xiii  APPENDICES  Page Table 1.1 Individual harbour seal data for the present study  164  Table 1.2 Contaminant levels in blubber samples from harbour seals and killer whales in the present study  167  xiv  ACKNOWLEDGEMENTS  Thank you to my supervisors,T>r. Peter Ross and Dr. Stelvio Bandiera, for their support and understanding and to Neil Dangerfield and Marta Assuncao for their help both in and out of the lab. Thank you especially to my mom and dad for their love and support and to those rare true friends who have helped me through it all.  XV  1  1.  1.1  INTRODUCTION  Persistent Organic Pollutants Anthropogenic chemical compounds are present in the air, water, sediments, and  in the tissues of humans and wildlife worldwide ( A M A P , 1998). Large-scale production of organochlorine (OC) compounds such as polychlorinated biphenyls  (PCBs),  dichlorodiphenyltrichloroethane (DDT), and other chlorinated pesticides began during the 1940's and 1950's, and many of these compounds are still in use today ( A M A P , 1998). PCBs are one of the most ubiquitous groups of anthropogenic compounds. Due to their chemical stability, PCBs were synthesized for use in heat transfer mixtures, hydraulic fluids, and as electrical insulators.  Production, importation, and most non-  electrical uses of PCBs were banned in Canada in 1977; however, some PCBs remain in use or in storage (Pierce et al., 1998). Other  OC compounds  routinely  detected  in the  environment  include  polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs). PCDDs and PCDFs are by-products of incomplete combustion processes involving chlorine and phenolic compounds. PCDDs and PCDFs are also by-products of herbicide manufacture (Pierce et al., 1998). Pulp mills used to be a major source of PCDDs in Canada, but changes to the bleaching process and bans on the use of pentachlorophenol as a wood preservative have greatly decreased PCDD output (Hagen et al, 1997; Pierce et al., 1998).  Municipal incinerators and old landfill sites are the major sources of  PCDDs and PCDFs today (Pierce et al, 1998).  2  A  B meta  5 meta  ortho  6 ortho  6'  5'  Figure 1.1: Organochlorine chemical structures: (A) PCB showing the designations for substituted chlorine atoms. (B) 2,3,7,8-tetrachlorodibenzodioxin (TCDD). (C) 2,3,7,8tetrachlorodibenzofuran (TCDF). (D) dichlorodiphenyltrichloroethane (DDT).  Organochlorines belong to a larger group of chemicals, collectively called Persistent Organic Pollutants (POPs).  In 2001, the United Nations adopted the  Stockholm Convention, which aims to reduce or eliminate twelve of the most toxic POPs, including PCBs, PCDDs, PCDFs, DDT, and other pesticides worldwide. Although industrialized countries such as Canada and the United States have banned or restricted the use of all twelve POPs targeted by the Stockholm Convention (Schmidt, 1999), some of these compounds are still produced or used in other parts of the world (Walker, 2000). Additionally, unregulated compounds, such as polybrominated diphenyl ethers (PBDEs), polybrominated biphenyls (PBBs), and newer POPs are being released into the environment (Meerts et al., 2000). Unlike PCBs and DDT, the environmental effects of these compounds are poorly understood.  3  1.1.1  POPs in the marine environment The world's oceans are particularly susceptible to large inputs of industrial and  agricultural pollutants. Pollutants enter the marine environment as a result of run-off, leakage, dumping, or atmospheric transport ( A M A P ,  1998).  Areas immediately  surrounding point sources such as pulp mills can be directly contaminated, whereas areas remote from human activity are contaminated indirectly by sediment transport, and more significantly, by atmospheric transport and deposition. PCDFs associated with sediments can be transported by ocean currents over distances of greater than 30 kilometers (Macdonald et al., 1992), and organochlorines are transported thousands of kilometers by circulating air masses in the atmosphere.  Atmospheric transport partially explains the  continued presence of organochlorines in regions where these compounds are no longer produced, and it is largely responsible for the occurrence of OCs in remote regions such as the Arctic ( A M A P , 1998; Muir et al, 1999). After initial deposition to the Earth's surface, OC compounds may re-enter the atmosphere by volatilization or resuspension of dust or snow by wind. This cycle of transport and deposition may occur multiple times (Macdonald et al, 2000). The main factor responsible for the persistence of OCs in the marine environment is  their  relatively  high  resistance  to  chemical  and  biological  degradation.  Organochlorines can remain buried in sediments for long periods of time and can be reexposed by human activity or bottom-dwelling organisms, or by other natural factors such as tides. Additionally, hydrophobic OCs accumulate in the lipid stores of organisms and are subsequently biomagnified through aquatic food webs. Biomagnification is the  4  successive increase in concentration of a chemical with increasing trophic level (i.e. as contaminants are passed up the food chain) (Muir et al., 1999).  1.1.2  POPs in marine mammals  As top predators  in the marine food chain with relatively large energy  requirements, marine mammals are particularly susceptible to the accumulation of lipophilic chemicals in their blubber (Muir et al., 1999). The first scientific reports of contaminants in marine mammals were published in 1966. B y the end of the 1960s, PCBs, dieldrin, and DDT and its metabolites had been documented in the tissues of several pinniped (seal, sea lion, walrus) and cetacean (whale, dolphin, porpoise) species (O'Shea and Tanabe, 1999). After regulations were implemented for P C B use and production in North America and Europe in the 1970s, PCB concentrations decreased in several environmental compartments, including marine mammals. For example, in Eastern Canada, mean total DDT concentrations declined in adult female grey seal (Halichoerus grypus) tissue from 13.2 to 4.0 mg/kg (lipid weight basis) between 1974 and 1982 (Addison et al., 1984a).  PCB concentrations did not  change in grey seals and decreased slightly in harp seals {Pagophilus groenlandicus) over the same period (Addison et al,  1984a).  Concentrations of D D T , PCBs, and  hexachlorobenzene (HCB) also decreased in blubber of western Arctic ringed seals {Phoca hispida) between 1971 and 1991 (Addison and Smith, 1998).  In general,  organochlorine concentrations in marine mammals leveled off during the late 1980s and early 1990s (Addison and Smith, 1998; Ross and Troisi, 2001). This stabilization is  5  attributed to leakage of existing PCB stores, continued use of DDT in some regions of the world, and environmental cycling (Walker, 2000; Ross and Troisi, 2001). PCBs and D D T continue to dominate the contaminant profiles of marine mammals in most locations, and high levels of OC pollutants continue to be measured in marine mammal species worldwide (Ross and Troisi, 2001). PCB and D D T in a number of marine mammal species are shown in Table 1.1. Recently, killer whale (Orcinus orca) populations in British Columbia were identified as being among the most P C B contaminated marine mammals in the world (Ross et al, 2000). PCB concentrations in killer whales often exceed those of other highly contaminated cetacean populations, such as the St. Lawrence beluga whales (Delphinapterus leucas), which researchers suspect have suffered reproductive and immunological effects as a result of P C B exposure (Beland et al, 1993; De Guise et al., 1998) (refer to Table 1.1). The contaminant levels shown in Table 1.1 were measured in blubber and are expressed on a lipid weight basis. However, several factors such as different methods of PCB quantification (eg. number and identity of congeners measured), the use of different tissues for PCB analysis (eg. blubber, liver, blood), natural confounding factors (eg. age, sex, condition), and different ways of expressing contaminant levels (eg. per gram wet weight or lipid weight), often preclude direct comparison of PCB concentrations between studies.  -1 ~<  :  o  "> TO ~  OO  re  m  & TO  a a^ ^ s.  P  Si  tr  CD  P  to 3 w O ft - • (Op O 0  7T  -i  Q  CO re'  cr  B •I  oo" • r-tt-J CD  Z  P  re" re  <=> 2  1  ET  oq  CO m  *  re oo  &  re P  re 00  re P  pj  re  £L  icr fa  re r-t-  P  o re  6  K P  •-1  cr o P i-i  CO  re  P i-i  cr o P i-i  CO  re  EP  X p  p  p  cr o  cr o  cr o  cr o  i-i P i-i  CO  re  1-1  P i-i  CO  re  i-i  P i-i  i-i P  P  oq  re CL  CO  CO  CO  p  P  re  re  P  P  P  p  P  C  K I  w  w  re  P  00 oo o  re  cr cr o P  CO  re p  0O T3  P  i-t  oq  tr P  CO  re P  re  t3 re  CL oo  o_ re'  CO  CO.  r—  ^  re  co  re K- p. a n.  o  00  DO  re cr EL re"  n  a.  0o r-t-  vo  1' P  >.°° o" era " a 01 .  -s CO  2? < re  oo a - i w r - fa • p to O O to - •  ^  p  re P o re  re o  i —  GO  OQ CD " CD X  ro  ro ro 0O  cr  o  B* 3 cr  cr O o B*  cr O o B"  3 cr  P'  0Q  JC-  o  p co  J» ^. OT  cr  %  1  P  °H ^°  P P CL P  r  re c5  o re  •  o  o  on  j - ^*  MC MC  oo o  MC MC U-)  MC MC U)  MC MC U)  MC MC Os  MC MC ON  MC MC  4^  4^  MO  MC  MC  MC  MC  MC  \D <3\  o>  ro  ro  O  o>  Os  as  -O.  r-t-  Pre  P  P  s  i-i P  o  .p  r^-  o' p  o  i-i  re j-i  oo  MC  P  P  1  P CL P  CO  SS. 3 a 09.  re P o re  re core  ro  O  co  ST C L fa  1  P xt 3 P  P  re >-i  ! ± ! CD  0  rP  P  —J  MD OO OO  MD MD OO  MD  MC as  MO ON  re 8  MD  MD tr: MD "T3 - •  CL  u  < < < H 8 S B. B. O O  ro  <  1  o  2 cr c  a 2  P  00  so so  co  XS  P  o  o  co  ro t3  P CL P  P CL P  co  o  r-t-  g. o  P CL P  •§  0Q  p  re  P  co  CO  °  oo  to  1—»  as  to  Lo  1>j  4^-  oo  to oo I oo to  to  4^  4^  bo i  4^  io  ON  L/l  I  e  si  VO  OO  O  o o  MD  3 5'  M  M  o  oo  cr  MC  to  ro  re  CD CL  at >  <: p B. o  <  o  p  oo re  CO  i—'  o  O  o  H si  si  ON  Os  <~r\  OJ  to  ^3  re  7  1.1.3  Toxicity of organochlorines Organochlorines are known to produce several toxic responses in living  organisms. Toxic responses observed in laboratory animals and wildlife species include immunosuppressive effects, reproductive and developmental toxicity, carcinogenesis, endocrine disruption, and neurotoxicity (Colborn et al., 1993; Safe, 1994). In marine mammals, OCs have been implicated in reproductive, developmental, and immunological deficiencies (Ross et al., 1996). PCBs have been associated with decreased reproductive success in pinnipeds (DeLong et al., 1973; Helle et al., 1976a,b; Reijnders, 1986). In California  sea  lions  (Zalophus  californianus  calif or nianus),  P C B and  DDT  concentrations in blubber were 6.6 and 8.0 times higher in females that gave birth to premature pups compared to those that gave birth to full-term pups (DeLong et al., 1973). Although these results are suggestive of an association between PCB blubber levels and reproductive success, they may be partly artifactual. Females that successfully reproduce have lower OC levels than those that abort because OCs are off-loaded during lactation. In Baltic ringed seals (Phoca hispida baltica), pathological changes of the uterus were associated with higher contaminant levels (Helle et al, 1976a,b). In one study, higher levels of PCBs and DDT were found in non-pregnant females compared to pregnant seals, and half of non-pregnant seals had enlarged uteri and scars in the uterine wall, suggesting that abortion or resorption had occurred following implantation (Helle et al, 1976a). Beland et al (1993) found a number of abnormalities in beluga whales in the St. Lawrence River that may have been associated with high contaminant levels. These whales had a high prevalence of tumours and lesions to the digestive system and other  8  glandular structures. No such lesions were found in Arctic belugas, which had P C B and DDT concentrations that were 18 and 12 times lower than St. Lawrence whales (Beland et al,  1993).  relevant  In samples from the same contaminated population, environmentally  concentrations  of organochlorines were  shown to reduce  lymphocyte  proliferation in vitro (De Guise et al, 1998). Impaired immunological function was observed in captive harbour seals (Phoca vitulina) with chronic exposure to environmental contaminants accumulated through the diet. A number of immunological effects, including diminished natural killer cell activity and T-cell function, were observed in seals fed a diet of herring from a relatively contaminated region (De Swart et al, 1994; Ross et al, 1996).  A n immunotoxic  threshold of 17 mg/kg lipid weight was established for PCBs in these captive harbour seals. As shown in Table 1.1, blubber PCB concentrations exceed this threshold in a number of marine mammal species. 1.1.3.1  The Aryl hydrocarbon receptor Many of the biochemical and toxic effects of PCBs, PCDDs, and PCDFs are  mediated by binding to the aryl hydrocarbon receptor (AhR). The resulting receptorligand complex acts as a nuclear transcriptional enhancer for gene expression of metabolic enzymes, such as cytochrome P4501A (Safe, 1994; Whitlock and Denison, 1995).  The unligated receptor consists of two 90 kDa heat-shock protein molecules  (hsp90) and other proteins. Binding of a ligand to the receptor results in dissociation of hsp90 and association with an AhR nuclear translocator protein (ARNT). Activation of transcription of TCDD-responsive genes occurs via a dioxin-responsive enhancer (DRE) sequence (Hahn, 1998).  9  2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) has the highest known affinity for AhR binding and is considered to be the most potent compound among PCBs, PCDDs, and PCDFs (Safe, 1994). A mechanistic link between TCDD-AhR binding and toxicity has been shown using A h R knockout mice (Fernandez-Salguero et al., 1996).  No  significant toxic or pathologic effects were observed in AhR knockout mice at a dose of T C D D ten-fold higher than that shown to cause pathological changes in the liver and thymus of wild-type mice (Fernandez-Salguero et al., 1996). Aryl hydrocarbon receptors have been detected in all vertebrate classes (fish, reptiles, amphibians, birds, mammals) and have been cloned in a number of vertebrate species. In marine mammals, full length AhR c D N A sequences have been reported for beluga whales, harbour seals, and Baikal seals (Phoca sibirica) (Jensen and Hahn, 2001; K i m and Hahn, 2002; K i m et al., 2002). Harbour seal and beluga whale AhR proteins are 82% identical in overall amino acid sequence (Kim and Hahn, 2002). It has been shown that the A h R in beluga whales and harbour seals exhibit specific binding of [ H]TCDD 3  and bind [ H]TCDD with an affinity at least as high as mouse A h R (Jensen and Hahn, 3  2001; K i m and Hahn, 2002). This information suggests that the AhR is highly conserved among vertebrates and that dioxin-mediated toxicities also occur in marine mammals. Toxic equivalency factors (TEFs) compare the ability of coplanar PCBs and 2,3,7,8-substituted PCDDs and PCDFs to bind to the A h receptor and to elicit AhRmediated toxic responses relative to 2,3,7,8-TCDD. 2,3,7,8-TCDD has been assigned a TEF of 1.0 (Van den Berg et al., 1998), and TEFs have been assigned to other planar PCB, PCDD, and PCDF congeners for mammals, birds, and fish, based on their potency relative to 2,3,7,8-TCDD (Table 1.2). These TEFs have been established by the World  10  Health Organization based on dose-response studies in the literature (Van den Berg et al., 1998). For mammals, TEF values are largely based on in vivo studies in rodents (Van den Berg et al, 1998). Where TEF values have been determined for several different AhR-mediated responses (eg. enzyme induction, carcinogenicity, reproductive toxicity), a mean TEF is calculated and an overall TEF is assigned (Safe, 1994). The contribution of an individual compound to the overall toxicity of a contaminant mixture is calculated by multiplying the concentration of the compound in an environmental sample (eg. water or tissue) by its TEF. The resulting value is the toxic equivalent (TEQ). Toxic equivalents for individual congeners are added to give an overall TEQ for a contaminant mixture. The equation for calculating the TEQ of a chemical mixture is as follows, where n = number of congeners: T E Q = Z [PCDDi x TEFj] + £ ([PCDF; x TEF,]) „ + I ([PCB; x TEFj]) n  n  There are many other halogenated compounds besides PCDDs, PCDFs, and coplanar PCBs that meet the criteria for inclusion in the TEF concept; however, there is currently insufficient data to be able to assign TEFs to those compounds (Van den Berg et al., 1998). TEFs also do not account for non-AhR-mediated toxicities or possible synergistic or antagonistic interactions between congeners.  11  Table 1.2: Toxic Equivalency Factors (TEFs). The most recent World Health Organization TEFs for mammals, as published by Van den Berg et al. (1998). International Union of Applied Chemistry (IUPAC) numbers are included for PCBs. Organochlorine PCDDs  Structure (PCB IUPAC #) 2,3,7,8-tetraCDD  1.0  1,2,3,7,8-pentaCDD  1.0  1,2,3,4,7,8-hexaCDD  0.1  1.2.3.6.7.8- hexaCDD  0.1  1.2.3.7.8.9- hexaCDD  0.1  1,2,3,4,6,7,8-heptaCDD  0.01  octaCDD PCDFs  Non-ort/zo PCBs  Mono-or^o PCBs  Mammalian TEFs  0.0001  2,3,7,8-tetraCDF  0.1  2,3,4,7,8-pentaCDF  0.5  1,2,3,7,8-pentaCDF  0.05  1,2,3,4,7,8-hexaCDF  . 0.1  2,3,4,6,7,8-hexaCDF  0.1  1.2.3.6.7.8- hexaCDF  0.1  1.2.3.7.8.9- hexaCDF  0.1  1.2.3.4.6.7.8- heptaCDF  0.01  1.2.3.4.7.8.9- heptaCDF  0.01  octaCDF  0.0001  3,3',4,4'-tetraCB (77)  0.0001  3,4,4',5-tetraCB (81)  0.0001  3,3',4,4',5-pentaCB(126)  0.1  3,3',4,4',5,5'-hexaCB(169)  0.01  2,3,3',4,4'-pentaCB (105)  0.0001  2,3,4,4',5-pentaCB (114)  0.0005  2,3',4,4',5-pentaCB (118)  0.0001  2',3,4,4',5-pentaCB (123)  0.0001  2,3,3',4,4',5-hexaCB (156)  0.0005  2,3,3',4,4',5'-hexaCB (157)  0.0005  2,3',4,4',5,5',-hexaCB (167)  0.00001  2,3,3',4,4',5,5'-heptaCB (189)  0.0001  12  1.1.3.2  Metabolism of organochlorines  Marine mammals are able to metabolize certain organochlorine compounds. In pinnipeds and cetaceans, P C B congeners that attain a planar configuration are more readily metabolized (Tanabe et al., 1988; Boon et al, 1997). The first indication that marine mammals selectively metabolize OC congeners came from studies in which P C B patterns were observed to differ between marine mammals and their prey (fish). Boon et al. (1997) concluded that differences in congener patterns between predator and prey can be  explained by the  availability  of PCBs  from  food  and their  subsequent  biotransformation by metabolic enzymes (the cytochrome P450 enzyme system, discussed in section 1.2.1) (Boon et al, 1987; Boon et al, 1994). Boon et al (1997) also observed concentration-dependent  changes in PCB patterns within various marine  mammals species. This observation indicated that metabolic enzymes were induced by increased P C B concentrations, resulting in increased metabolism of certain congeners in these animals (Boon et al, 1997). Another important finding was that P C B congener patterns differed between marine mammals and terrestrial animals that feed at the top of the food chain, suggesting that the metabolic capacity for specific congeners varies between species (Tanabe et al, 1988). Lower metabolic enzyme activities observed in marine mammals compared to terrestrial mammals, for example, may lead to higher bioaccumulation of certain PCB congeners in marine mammals (Tanabe et al, 1988). Studies with captive harbour seals have provided more direct evidence for metabolism of PCDDs and PCDFs in this species. During a 15-day fasting experiment, body burdens of 2,3,7,8-TCDD and 2,3,7,8-TCDF were less than 10% of the estimated cumulative intake of these compounds from herring. The half-life of 2,3,7,8-TCDD was  13  estimated to be less than one week, as blood concentrations decreased from 2.4 ng/kg lipid to less than half that after a week of fasting, presumably due to metabolism and excretion of this contaminant (De Swart et al, 1995). However, in free-ranging pinniped populations, relatively high levels of PCDDs and PCDFs have been found (eg. in British Columbia), likely reflecting continued contamination of prey from point sources such as pulp mills and urban environments (Ross and Troisi, 2001). For example, mean total P C D D and PCDF concentrations in blubber of nursing harbour seal pups from British Columbia were 187 and 20 ng/kg lipid respectively (Simms et al, 2000); whereas, in grey seals from the east coast of Canada, PCDD and PCDF concentrations were 12 and 8 ng/kg lipid, respectively, in mothers, and 7 and 2 ng/kg, respectively, in pups (Addison et  al, 1999). There is also evidence to suggest that killer whales metabolize dioxin-like compounds. Ross et al. (2000) found that PCDDs and PCDFs were present at low levels in blubber compared to PCBs (eg. mean total P C B concentration was 37 mg/kg lipid in male northern residents; mean total combined PCDD and PCDF concentrations in males and females were 1050 ng/kg and 258 ng/kg lipid, respectively), and there were no differences in TEQ concentrations for these compounds between immature and adult animals or between adult males and adult females.  14  1.2  Biomarkers Biomarkers are biological responses to one or more chemicals that give a measure  of exposure and sometimes of toxic effect (Peakall, 1994). In contrast to the classical approach to environmental toxicology, in which contaminants measured in environmental compartments (i.e. water and sediments) are related to adverse effects observed in experimentally treated laboratory animals (Peakall, 1994), biomarkers can provide a measurement of response to contaminant exposure that is biologically relevant in the species of interest. Characteristics of an ideal biomarker include the following (WHO, 1993; Oikari etal, 1993): •  reflects interaction of the host biological system with the chemical(s) of interest;  •  has known (high) specificity and sensitivity to the interaction;  •  is common to individuals in a population and across species, and is of defined variability within the normal, non-exposed group of interest;  •  has a validated method that is simple, inexpensive, and reproducible, and measurements have defined and appropriate accuracy and precision; and  •  is non-lethal and minimally-invasive Biomarkers may be biochemical, physiological, histological, morphological, or behavioural. Commonly studied biomarkers include:  •  induction of metallothionein in response to cadmium exposure (Addison, 1996);  •  inhibition of acetylcholinesterase by exposure to organophosphates and carbamates (Addison, 1996);  15  •  disruption of vitamin A homeostasis (Simms et al., 2000b);  •  disruption of the thyroid hormone system (Brouwer, et al., 1989); and  •  induction of monooxygenases in response to organochlorine and P A H exposure (Addison, 1996)  In this study, we will look at cytochrome P450 1A (CYP1A), a monooxygenase, as a biomarker of organochlorine exposure.  1.2.1 1.2.1.1  Cytochrome P450 1A Background CYP1A  belongs to a superfamily of monooxygenases  that catalyze the  metabolism of lipophilic endogenous and xenobiotic compounds to more water-soluble products. Due to its wide range of substrates, the C Y P enzyme system is one of the most important xenobiotic metabolizing systems in vertebrates (Schenkman, 1999).  A  substrate binds to the heme-containing C Y P enzyme, and the enzyme then accepts an electron from N A D P H via NADPH-cytochrome P450 reductase.  Oxygen binds to the  reduced hemoprotein, forming the oxycytochrome P450, and then accepts a second electron from the reductase, or from ferrous cytochrome bs. Pro to nation and cleavage of the 0 - 0 bond produces a molecule of water, and the remaining oxygen atom that is bound to iron is transferred to the substrate.  The oxidized substrate is subsequently  released from the C Y P enzyme (Schenkmann, 1999; Bandiera, 2001).  16  Figure  1.2: T h e C Y P c y c l e , w h e r e S = substrate ( S u b s t r a t e + N A D P H + H  Substrate-0 + N A D P  +  +  + 0  2  ->  + H 0 ) ( f r o m L e w i s , 1996). 2  A l t h o u g h the p h y s i o l o g i c a l r o l e o f C Y P 1 A  is to d e t o x i f y  potentially  harmful  c o m p o u n d s , there is e v i d e n c e to s u g g e s t that h i g h C Y P 1 A a c t i v i t y i n i t s e l f c a n l e a d to adverse health effects  (Barouki  established between C Y P 1 A 1984).  and M o r e l , 2001), and a mechanistic link has  induction and toxicity  (Poland and Knutson, 1982; Safe  T w o m e c h a n i s m s b y w h i c h C Y P 1 A c a n l e a d to t o x i c i t y a r e t h e p r o d u c t i o n  t o x i c metabolites (eg. h y d r o x y l a t e d and m e t h y l s u l f o n y l B e r g m a n et al, (ROS)  been  1 9 9 7 ; L e t c h e r et al,  ( P a l a c e et al,  1996).  PCBs)  ( B r o u w e r et  of  1989;  al,  2 0 0 0 ) a n d the p r o d u c t i o n o f reactive o x y g e n s p e c i e s  H y d r o x y l a t e d P C B s c a n interfere w i t h endocrine transport  s y s t e m s i n l a b o r a t o r y a n i m a l s , s u c h as b y c o m p e t i n g w i t h t h y r o x i n e for the b i n d i n g site o n the thyroxine-transporting  p r o t e i n i n t h e c i r c u l a t i o n ( B e r g m a n et al,  m e c h a n i s m is also suspected i n c o n t a m i n a n t - a s s o c i a t e d d i s r u p t i o n  1997).  of vitamin  Such a A  t h y r o i d h o r m o n e h o m e o s t a s i s i n c a p t i v e a n d f r e e - r a n g i n g h a r b o u r s e a l s ( S i m m s et  and al,  17  2000b; Brouwer et al, 1989; De Swart et al, 1994). Methylsulfonyl-PCBs have been shown to interfere with enzyme activity and endocrine-related processes in rats and mice (Letcher et al, 2000). In addition, induction of C Y P activity has been shown to result in the proliferation of oxyradicals, which can cause cellular damage, such as the breakdown of lipid membranes (Palace et al, 1996). 1.2.1.2  CYP1A as a biomarker of organochlorine exposure  C Y P 1 A induction is probably the most widely used biomarker of organochlorine exposure, as it is specific for a limited group of organic chemicals, and its mechanism of response to contaminant exposure is relatively well understood (Addison, 1996) (refer to Figure 1.3). Synthesis of certain C Y P enzymes is known to be induced by exposure to certain classes of environmental contaminants, and C Y P 1 A induction has been used as a biomarker of exposure to dioxin-like chemicals (the planar PCCDs, PCDFs, and PCBs) in fish, birds, and mammals (Rattner et al, 1989; Bosveld and Van den Berg, 1994). The C Y P 1 A enzymes are regulated by the A h receptor and catalyze the metabolism of planar aromatic and chlorinated hydrocarbons such as non-ortho- and mono-ort/zo-substituted PCBs and benzo[a]pyrene (Van den Berg et al, 1998). C Y P 1 A induction is typically measured in terms of enzyme activity, protein concentration, or mRNA concentration (Table 1.3).  C Y P catalytic activity can be  measured using several different mixed function oxidase (MFO) enzyme systems. The MFOs typically used to measure C Y P 1 A activity are ethoxyresorufin (9-deethylase (EROD) and aryl hydrocarbon hydroxylase (AHH), also referred to as benzo[a]pyrene hydroxylase or benzo[a]pyrene monooxygenase (BPMO).  Figure 1.3: Simplified mechanism of C Y P 1 A induction. Signal transduction by dioxinlike ligands is mediated by the A h receptor, which forms a transcription factor complex with A R N T . This heterodimer binds to specific D N A sequences called dioxin responsive elements (DRE), resulting in transcription of the C Y P 1 A gene and synthesis of C Y P 1A protein (Whyte et al, 2000).  19  Table 1.3: Summary of the most common methods used to measure C Y P 1 A induction Measurement C Y P 1 A protein concentration  Method Immunoblotting Immunohistochemistry  C Y P 1 A enzyme activity  Ethoxyresorufin O-deethylase (EROD) Aryl hydrocarbon hydroxlyase (AHH)  C Y P 1 A m R N A concentration  Northern blotting Reverse transcription polymerase chain reaction (RT-PCR)  Hepatic C Y P 1 A catalytic activity and C Y P 1 A protein levels have been measured both in pinnipeds (Goksoyr et al., 1995; Troisi and Mason, 1997; Mattson et al., 1998; Nyman et al., 2000) and cetaceans (Watanabe et al., 1989; Goksoyr et al., 1995), and several studies have found correlations between catalytic activity, C Y P 1 A levels, and contaminant burdens (Goksoyr et al., 1995; Troisi and Mason, 1997; Mattson et al., 1998; Nyman et al., 2000).  Correlations between contaminant levels and C Y P 1 A  expression suggest that C Y P 1 A induction occurs in these animals. Mattson et al. (1998) compared C Y P 1 A activities and apoprotein levels in Baltic ringed seals from an area with high organochlorine pollution («=17) versus seals from a relatively unpolluted area (n=\9). EROD activities were three times greater in seals from the heavily polluted region than in seals from the less polluted region (-800 pmol/mg protein/min versus -250 pmol/mg/min). Polluted and reference regions were chosen based on contaminant data from previous studies. In another study, a five-fold difference in hepatic EROD activity was observed in grey seals (-1200 versus 250 pmol/mg/min) and a three-fold difference in ringed seals (-1500 versus 500 pmol/mg/min) between  20  heavily polluted and reference areas (Nyman et al., 2000). In the same study, EROD activity and CYP1A1 protein levels were positively correlated in liver. This correlation has also been observed in other seal species, including harbour and hooded seals (Goksoyr et al., 1995). Troisi and Mason (1997) investigated cytochrome P450 enzymes and mixed function oxidase activity as biomarkers of PCB exposure in harbour seal pups that had died of bacterial infections and physical injuries. Hepatic EROD activity and total P C B blubber levels were positively correlated for these animals (n=5), though the poor health condition of these seals may have been a confounding factor. It is often not possible to make direct inter-study comparisons of C Y P 1 A expression in marine mammals. Different protocols for measuring C Y P 1 A levels exist, and these have not been standardized among laboratories. Immunoblotting results are dependent on the antibodies used and the methods used to quantify protein bands. C Y P 1 A enzyme activities cannot be easily compared between studies i f different enzyme assays are used (eg. EROD vs. AHH), or if activities are reported in different units. Additionally, comparison of C Y P 1 A expression in marine mammals is often problematic because factors such as species, age, sex, and health status (condition) also need to be considered. 1.2.1.3  CYP 1A in skin  Due to its relatively large size and high monooxygenase activity (Rattner et al., 1989), the liver has been the focus of most C Y P induction studies in both laboratory animals and wildlife. Liver samples used in the marine mammal studies cited above were obtained from stranded animals, from animals killed during commercial or subsistence harvests, or from animals killed for scientific purposes. Killing wild animals in order to  21  obtain tissue samples is generally unacceptable from an ethical standpoint, and is often illegal.  Samples from stranded animals may not provide accurate information due to  degradation of tissues after death, or the fact that deceased animals comprised a nonnormal cross-section of the population (i.e. old, diseased).  For these reasons, the  development of non-lethal (minimally-invasive) tissue sampling techniques is essential for biomarker research. For marine mammals, skin and blubber are the only tissues that can be readily obtained from live, free-ranging animals. To date, C Y P 1 A expression in marine mammal skin and its usefulness as a biomarker of organochlorine exposure has not been well described. Fossi et al. (1997) and Marsili et al. (1998) reported detection of C Y P 1 A induction in skin biopsies of southern sea lions (Otaria flavescens) and Mediterranean fin whales (Balaenoptera physalus), respectively. Fossi et al. compared epidermal CYP-mediated benzo[a]pyrene monooxygenase (BPMO) activity in free-ranging male sea lions from a region contaminated by oil and organic and chemical materials, and a control region. B P M O activity was approximately three to four times higher in skin samples from the heavily polluted area (n=\2) compared to those from the control site («=4). In one stranded sea lion, B P M O activity in skin was found to be one-fifth of that in liver (Fossi et al., 1997). Unfortunately, enzyme activity was measured as B P M O activity and expressed as fluorescence units/hour/g tissue, which precluded a comparison of activity values with most other studies.  Marsili et al. (1998) reported a positive correlation between  epidermal B P M O activity and blubber DDT and PCB concentrations in male fin whales (n=\4). However, the ages of the individual male fin whales were unknown. Age may  22  be a confounding factor in this study, as OC concentrations in blubber generally increase in males with age. Cutaneous induction of C Y P 1 A after topical application of known C Y P inducers has been demonstrated in laboratory rodents (Ichikawa, 1989; Khan et al, 1992; Jugert, 1994; Agarwal et al, 1994). However, our study represents the first such attempt in vivo in a marine mammal species. In river otters (Lontra canadensis) treated with crude oil, CYP1A1  expression  measured  by  immunohistochemical  staining  increased  approximately four-fold in skin biopsies (Ben-David et al, 2001). Godard et al (2002) reported induction of C Y P 1 A in sperm whale (Physeter macrocephalus) skin biopsies using immunohistochemistry following in vitro BNF treatment. These data have yet to be published. 1.2.1.4  Species specificity of CYP1A Certain forms of C Y P enzymes are species-specific (Gonzalez, 1989). However,  cloning and expression studies have shown that some C Y P genes are conserved across species, and the major microsomal C Y P gene families (1, 2, 3, and 4) are found in all vertebrates studied, from fish to mammals (Stegeman and Livingstone, 1998). Some C Y P families, such as CYP2, are quite diverse, and it is difficult to determine homologous gene lineages between vertebrate species. Others, such as the CYP1 family, are well-conserved among species, and antibodies prepared against purified C Y P isozymes in one vertebrate species can be used to identify and characterize CYPs in other vertebrate species (Ronis et al,  1989a,b; Bandiera et al,  1995; L i n et al, 1998).  Additionally, the induction of C Y P 1 A by xenobiotic compounds appears to represent a common response among vertebrates (Stegeman and Livingstone, 1998).  23  In mammals, the C Y P 1 A subfamily consists of two isozymes, CYP1A1 and CYP1A2 (Nebert et al., 1991). These two enzymes have been identified in rats, mice, rabbits, humans, and some marine mammals, but there is uncertainty as to whether the two forms exist in all mammals. Wolkers et al. (1998) observed cross-hybridization of only one C Y P 1 A mRNA band in ringed seal liver with human C Y P 1 A cDNA probes, and one C Y P 1 A protein band cross-reacted with an anti-rat C Y P 1 A antibody on immunoblots. Goksoyr et al. (1992) detected two C Y P 1 A protein bands in both harp (Phoca groenlandica) and hooded seal {Cystophora cristata) liver using an anti-cod CYP1A1 antibody. However, Nyman et al. (2000) were unable to distinguish between CYP1A1 and CYP1A2 protein bands in ringed and grey seal liver. Chiba et al. (2002) detected two hepatic C Y P 1 A protein bands in largha seals {Phoca largha) and ribbon seals {Phoca fasciata), using a polyclonal anti-rat CYP1 A l antibody. In whales, White et al. (1994) reported a single hepatic C Y P 1 A band in beluga using a polyclonal anti-mouse CYP1A1 antibody, Goksoyr et al. (1995) detected two C Y P 1 A bands in minke whale {Balaenoptera acutorostrata) liver using anti-cod C Y P 1 A antibodies, and Boon et al. (2001) reported cross-reactivity of anti-mouse and anti-rat antibodies with a single hepatic C Y P 1 A protein band in sperm whale. Given the results cited above, it is likely that both C Y P 1 A forms (1A1 and 1A2) exist in marine mammals, but that the specificity of C Y P 1 A probes and the sensitivity of the techniques used may not be adequate to detect both forms in all cases. The first C Y P 1 A protein sequences in marine mammals have only recently been identified (Teramitsu et al., 2000; Tilley et al., 2002). Antibodies for marine mammal C Y P proteins are therefore not yet commercially available, and immunodetection of  24  marine mammal C Y P 1 A remains dependent on cross-reactivity with anti-CYP antibodies from other animal species.  To date, hepatic CYP1A1 cDNA fragments from minke  whale, dall's porpoise (Phocoenoides dalli), largha seal, ribbon seal, grey seal, and harp seal, and CYP1A2 c D N A fragments from steller sea lion (Eumetopias jubatus), grey seal, and harp seal have been cloned (Termamitsu et al, 2000; Tilley et al., 2002). CYP1A1 sequences were >99% identical between grey and harp seals and identical between largha and ribbon seals, but seals and cetaceans belong to separate genetic lineages, with gene sequences in seals being most similar to those of dogs (Termamitsu et al., 2000; Tilley et al, 2002). 1.2.1.5  Natural factors affecting CYP1A expression  Although contaminants have been shown to influence C Y P 1 A expression, numerous physiological factors, such as age, diet, reproductive status, and health status also influence levels and activities of C Y P enzymes (Mattson et al, 1998). Age can influence C Y P 1 A expression if genes coding for C Y P 1 A enzymes are  under  developmental control. Developmental control of hepatic C Y P 1 A expression has been demonstrated in rats, rabbits, fish, and humans (Mukhtar and Bickers, 1983; Strom et al, 1992; Rich et al, 1993; Sonnier and Crested, 1998; Sarasquete et al, 2001). In general, hepatic C Y P 1 A is absent or occurs at low levels in fetal and neonatal animals and increases with age. In marine mammals, limited information is available on the developmental regulation of C Y P expression. In minke whales, hepatic EROD activity was about 80% lower in fetal samples than in juveniles and adults, but turnover of ethoxyresorufin (activity/nmol P450) was found to be higher in fetal samples than in most of the adult  25  samples (Goks0yr, 1986). In grey seals, liver EROD activity was not detectable in seals less than one day old; EROD activity increased with age up to 5-12 days post-partum and was lower than that of adults (Addison and Brodie, 1984b). In harbour seals, lower EROD activity was observed in newborn pups and fetuses than in adults (Addison et al., 1986). Mattson et al. (1998) found that hepatic EROD activity in Baltic ringed seals showed a weak positive correlation with age. A variety of other physiological factors have been shown to affect C Y P expression. Dietary constituents such as carotenoids and flavenoids have been shown to induce C Y P 1 A enzymes in rodents (Gradelet et al., 1996; Breinholt et al., 1999). Health status may also affect C Y P 1 A expression, as hepatic C Y P activities are affected (generally suppressed) by infection and by various inflammatory stimuli (Morgan, 1997). These factors will be addressed more fully in the discussion section. In adult marine mammals, C Y P 1 A expression may be influenced by the length of time an animal has had to accumulate lipophilic contaminants (i.e. age).  However,  accumulation of contaminants is also affected by sex and reproductive status. In marine mammals, tissue concentrations of contaminants increase with age in males, but reproductive females offload contaminants to their offspring through lactation (Ross and Troisi, 2001). Conflicting results between studies and a lack of data in marine mammals often makes interpretation of C Y P 1 A expression difficult and speculative. A n association of contaminant levels with effects is often precluded in marine mammal studies due to a lack of control over confounding factors, particularly age and condition. It is, therefore, extremely important to eliminate as many potential confounding factors as possible (eg.  26  age, sex, condition, reproductive status) when evaluating C Y P 1 A expression as a response to contaminant exposure. 1.3 1.3.1  Study Overview Description  1.3.1.1  of study species  Harbour seals (Phoca vitulina)  Harbour seals are widely distributed in temperate coastal waters of the Northern Hemisphere. In southern British Columbia and Washington State, harbour seals are nonmigratory and often inhabit shallow areas where sandbars and beaches are uncovered at low tide. They are generally solitary but loosely aggregate at haul-out sites while resting. Since 1970 harbour seals have been protected in British Columbia, and numbers increased from an estimated 10,000 animals in 1970 to 80-90,000 in 1988. The current harbour seal population in B.C. is estimated at 108,000 (Olesiuk, 1999).  There are  approximately 15,000 harbour seals in the inland waters of Washington State (Jeffries et al., in press). Hake and herring dominate the harbour seal diet in B.C. and Washington State, but other fish, molluscs, and crustaceans are also consumed (Olesiuk, 1993). Females bear one pup per year with a peak pupping season in July or August, depending on location. Lactation lasts for four to six weeks, after which time the pup is abandoned by its mother. At birth, the average weight of seals is about 11 kg, and this weight is doubled by weaning time (Cottrell et al, 2002). Most females mature at three to four years of age, and most males by  five  years of age. The average weight of fully grown  females is 65 kg and males average 87 kg (Bigg, 1969).  27  1.3.1.2  Killer whales (Orcinus orca)  Two sympatric populations of killer whales, residents and transients, frequent the coastal waters of British Columbia.  Resident killer whales feed on fish, principally  salmonids. Transients feed almost exclusively on pinnipeds and cetaceans (Ford et al., 1998). Male killer whales live an estimated fifty to sixty years and females eighty to ninety years. Females bear one calf every three to five years (Ford et al., 1998). The resident population consists of two distinct subpopulations referred to as the northern and southern residents.  The northern and southern resident subpopulations are currently  estimated at 200 and 81 individuals, respectively (Ellis, pers. comm.)  The transient  population is estimated at 220 individuals (Ellis, pers. comm). The southern resident population has declined from a peak of 99 individuals, and there is concern about the possible causes of this decline. The Committee on the Status of Endangered Wildlife in Canada (COSEWIC) has identified three major threats to southern resident killer whales: boat traffic (noise and physical interference), a diminishing food supply, and high levels of toxic contaminants.  1.3.2  Challenges associated with wildlife toxicology studies In contrast to working with laboratory animals in a controlled environment, there  are considerable challenges associated with studying free-ranging wildlife species. Obtaining samples from marine mammals is costly and requires a great deal of time, effort and logistical support. Sampling is also dependent on uncontrollable factors such as weather, ocean conditions, and the behaviour of the animals. Because of these factors, it may not always be possible to obtain the desired sample size or samples from a desired  28  cohort.  Additionally, many marine mammals are protected under Canadian law, and  therefore, must be handled under the scrutiny of Animal Care and Scientific Permitting bodies. The development and application of minimally-invasive sampling approaches is, therefore, necessary in order to minimize the impact that research has on these animals. Adequate preservation of tissue samples in the field is also important. Supplies, such as liquid nitrogen, must be taken into the field, and extra care must be taken to maintain relatively sterile conditions when handling samples. The amount of information that can be obtained from these animals in the field is also limited. Body measurements such as length and weight can be determined for live-captured harbour seals, but specific ages can only be determined if a tooth is removed, and aging cannot be done until postsampling. After sampling is completed, tissue samples must be used efficiently.  Samples  are often divided up for several different studies, and due to the small amount of tissue available, it is not always possible to use study samples for validation and optimization experiments. It may be necessary to use archived samples from dead animals instead. This is not ideal, as there may be differences between archived samples and fresh samples (eg. archived samples may not have been preserved immediately after death of the animal or degradation of tissue components may have occurred during prolonged storage). Toxicological studies in which the aim is to identify a cause-effect relationship between contaminant exposure and a biological response (biomarkers) are especially challenging when working with marine mammals. In general, the animals' environment cannot be controlled (eg. diet, sources of environmental contaminant exposure), and  29  therefore, it is more difficult to identify specific agents responsible for the observed biological responses.  Captive studies of wild marine mammals are expensive and  logistically challenging. Permits must be obtained, and individuals trained in the care of these animals must be employed. In order to evaluate a particular biomarker, it is helpful, if not necessary, to have baseline data for this response in the species of interest and such information is usually not available for marine mammals.  Additionally, it is more  difficult to obtain positive controls for toxicological studies in marine mammals, as it is usually not acceptable to treat these animals with toxic compounds. Toxicological studies in free-ranging marine mammals are extremely valuable to understanding the effects of contaminant exposure in these animals. Such studies provide 'real-world' relevance (eg. exposure to complex contaminant mixtures) that laboratory animal studies lack (Ross, 2000). Figure 1.4 summarizes the various approaches that contribute to our understanding of contaminant effects in marine mammals. j big .£ c 1 -a a> 1 C XI roc £o  mech  n  ogical asing  <D o o  <D O  stic und reasing  evance; ifidence  1  "5 2 9, o 0) <u  'w  c  '(/>  ro <p_  o> c  rea  cnf  o  Effects  Exposure  Free-ranging marine mammals extrapolation  C a p t i v e harbour s e a l s extrapolation  Laboratory rodents (marine m a m m a l surrogates) extrapolation  o  c  f  Laboratory rodents  Figure 1.4: A 'weight of evidence' approach is required to assess the risk of toxic injury in free-ranging marine mammals. This involves extrapolation between single chemical laboratory rodent studies, surrogate studies using rodents as seals, captive feeding studies of seals fed complex contaminant mixtures, and observational and minimally-invasive biomarker studies of free-ranging marine mammal populations (from Ross, 2000).  30  1.3.3  Overall goals of the present study The importance of ecotoxicological studies in marine mammals is twofold. First,  information from such studies may provide information about the health of marine mammal populations.  Causal links between contaminants and events such as virus-  related mass mortalities of marine mammals may be established, or alternatively ruled out. Second, toxicological information obtained from marine mammals can provide an indication of the overall health of the marine environment. The objectives of this study were: •  to develop techniques to measure CYP1A protein levels and enzyme activity in harbour seal and killer whale skin biopsies  •  to determine if C Y P 1 A could be used as a biomarker of organochlorine exposure in harbour seals and killer whales  Our study will focus on the development of minimally-invasive sampling techniques for free-ranging marine mammals.  No C Y P studies have been published on marine  mammals in British Columbia, and to our knowledge, no studies have been published on C Y P 1 A in skin of harbour seals or killer whales from any location. Only one report exists of C Y P 1 A expression in killer whales (CYP1A enzyme activity was detected in liver of killer whales from Japanese waters) (Watanabe et al, 1989).  We used the  harbour seal as a model for method development in marine mammals. Relative to other marine mammal species, harbour seals are locally abundant and easy to sample, and they are the subject of ongoing toxicological studies in our laboratory (Simms et al, 2000a,b; Mos and Ross, 2002).  31  1.4  Research Hypothesis Skin provides a reliable measure of systemic (hepatic) C Y P 1 A in harbour seals  and can be used as an effective biomarker of organochlorine exposure in harbour seals and killer whales.  1.5  Experimental Hypotheses  1) C Y P 1 A can be measured in both liver and skin biopsies from harbour seals and in skin biopsies from killer whales. 2) C Y P 1 A induction in harbour seal skin is positively correlated with C Y P 1 A induction in liver. 3) C Y P 1 A induction in skin is positively correlated with organochlorine concentrations in blubber.  32  6  Specific Objectives to establish protocols for homogenizing small liver and skin samples obtained by biopsy from harbour seals; to measure ethoxyresorufin O-deethylase (EROD) activity and C Y P 1 A protein levels in microsample homogenates of liver and skin from harbour seals, using a plate reader assay and immunoblotting, respectively; to measure C Y P 1 A induction in liver and skin of fi-naphthoflavone (BNF; a C Y P 1 A inducer)-treated harbour seals; to assess whether C Y P 1 A expression in harbour seal skin correlates with that in liver; to measure EROD activity and C Y P 1 A protein levels in skin biopsies from killer whales using techniques validated in harbour seal skin; to determine if C Y P 1 A measurements in liver of harbour seals, and in skin of harbour seals and killer whales, correlate with organochlorine concentrations in blubber.  33  2.  2.1  MATERIALS AND METHODS  Overview  Three biomarker studies were conducted in marine mammals: (1) a study of captive harbour seals with analysis of liver and skin-blubber biopsies; (2) a study of free-ranging harbour seals with analysis of skin-blubber biopsies; (3) a study of free-ranging killer whales with analysis of skin-blubber biopsies. Tissue biopsies from harbour seals and killer whales were analyzed for C Y P 1 A enzyme activity and C Y P 1 A protein concentration using the ethoxyresorufin 0deethylase (EROD) assay and immunoblotting, respectively. Experiments were carried out to optimize protein extraction from tissue biopsies, while minimizing the amount of tissue used, and to validate and optimize the EROD and immunoblot assays for our marine mammal tissue samples (described in section 2.11). The primary methodological goal was to measure C Y P 1 A in small tissue samples collected using minimally-invasive techniques. 2.2  Chemicals The contents of buffers and reagents used in this study are listed in Table 2.1.  Chemicals and reagents used in this study were obtained from the following sources: Aldrich Chemical Company Inc. (Milwaukee, Wl, USA) Resorufin BDH Chemicals (Toronto, Ontario, Canada):  34  Sodium chloride (KC1); sodium hydroxide (NaOH); sodium carbonate ( N a C 0 ) ; 2  3  disodium hydrogen orthophosphate (Na HP04); potassium dihydrogen orthophosphate 2  ( K H P 0 ) ; magnesium chloride (MgCl -6H 0); dimethylsulfoxide (DMSO), cupric 2  4  2  2  sulphate pentahydrate (CuS04"5H 0); isopropanol 2  Bio-Rad (Richmond, CA, USA): Bromphenol blue BIOSOURCE International (Camarillo, CA, USA) Goat (Fab') anti-rabbit immunoglobulins, alkaline phosphatase conjugated 2  Fisher Scientific (Fair Lawn, NJ, USA): Trizma base (Tris); acrylamide; N,N'-methylene-bis-acrylamide (Bis), N,N,N',N'~ tetramethylenediamine  (TEMED);  ammonium  persulphate (AP); sodium dodecyl  sulphate (SDS); glycine; B-mercaptoethanol; reagent grade methanol; HPLC grade methanol; sodium chloride (NaCl); bovine serum albumin (BSA); nitro blue tetrazolium (NBT);  5-bromo-4-chloro-3-indolyl phosphate, p-toluidine salt  Sodium/potassium (Na/K) tartrate tetrahydrate Gibco Goat serum ICNBiochemicals, Inc. (Cleveland,, OH, USA): N , N-dimethylformamide (DMF) MERCK Coomassie blue R250  (BCIP); glycerin;  35  Oxford Biomedical Research Inc. (distributed by Cedarlane Laboratories, ON, Canada) Polyclonal goat-anti rabbit 1 A l and 1A2 IgG Pacific (Vancouver, BC, Canada): Skim milk powder SIGMA Chemical Co. (St. Louis, MO, USA): Ethylenediaminetetra-acetate disodium salt (EDTA); pyronin Y ; polyoxyethylene sorbitan  monolaurate  (Tween  20);  1-butanol; N-2-hydroxyethylpiperazine-N-2-  ethanesulfonic acid (HEPES); dimethylsulfoxide (DMSO); B-naphthoflavone (BNF); nicotindiamide  adenine  dinucleotide  phosphate  tetrasodium  salt  (NADPH);  ethoxyresorufin Dr. S.M. Bandiera (Faculty of Pharmaceutical Sciences, University of British Columbia, Vancouver, Canada): Purified rat cytochrome P450 1A1; purified rat cytochrome P450 1A2; liver microsomes prepared from 3-MC treated adult male Long Evans rats; rabbit anti-rat cytochrome P450 1A1 IgG; rabbit anti-rat cytochrome P450 1A2 serum; rabbit anti-trout cytochrome P450 1A1 IgG prepared against a 15 amino acid peptide of the trout CYP1A1 sequence; control rabbit IgG; control rabbit serum Dr. P. Thomas (Rutgers, The University of New Jersey, Piscataway, NJ, USA): Monoclonal mouse anti-rat cytochrome P450 1A1 IgGs  36  Table 2.1: Buffers and reagents used in this study Buffer or reagent  Contents  Tris-HCl buffer (for homogenizing tissues)  50 m M Tris-HCl, pH 7.4 at room temperature (r.t); 150 m M KC1; 2 m M E D T A  Lowry assay reagents  (a) 2% N a C 0 anhydrous in 0.1 M N a O H 3  (b) 2% Na/K tartrate tetrahydrate (c) l % C u S 0 - 5 H 0 4  2  Sample dilution buffer  62.5 m M Tris-HCl, pH 6.8 at r.t; 10.8% glycerol; 0.001% bromphenol blue; 1% SDS; 5% Bmercaptoethanol  Separating gel  0.375 M Tris-HCl, pH 8.8 at r.t; 0.1% SDS; 7.5% acrylamide bis; 0.042% AP; 0.03% T E M E D  Stacking gel  0.125 M Tris-HCl, pH 6.8 at r.t; 0.1% SDS; 3% acrilyamide bis; 0.08% A P ; 0.05% T E M E D  Separating gel wash buffer  0.68 M Tris-HCl, pH 8.8 at r.t; 0.18% SDS  Stacking gel wash buffer  0.23 M Tris-HCl, pH 6.8 at r.t; 0.18% SDS  Electrophoresis buffer  50 m M glycine; 0.38 M trizma base; 0.2% SDS  Transfer buffer  0.13 M glycine; 25 m M trizma base, 0.1% SDS  Membrane blocking buffer  50 m M NaCl; 10 m M Tris-HCl, pH 7.4 at r.t.; 1 m M E D T A , pH 7.4 at r.t; 5% skim milk powder  Modified PBS (1 Ox)  1.3 M NaCl; 26 m M KC1; 81 m M N a H P 0 ; 15 m M K H P 0 ; 2 mM EDTA 2  2  4  4  Antibody dilution buffer  0.05% Tween 20; 2% B S A ; 5% goat serum in modified PBS  Membrane wash buffer  0.05% Tween 20 in modified PBS  Gel staining solutions  (a) Fixative: 50% methanol; 10% glacial acetic acid (b) Stain: 50% methanol; 10% glacial acetic acid (c) Destain I: 50% methanol; 10% glacial acetic acid (d) Destain II: 5% methanol; 7% glacial acetic acid  Substrate solution r.t. = r o o m t e m p e r a t u r e  0.03% N B T ; 0.015% BCIP; 0.05 m M M g C l in 0.1 M Tris-HCl, pH 9.5 at r.t. 2  37  2.3  Sample Collection  Figure 2.1: Marine mammal sampling sites in British Columbia and Washington State Superscript numbers indicate study groups: (1) captive harbour seal study; (2) study of free-ranging harbour seals; (3) study of free-ranging killer whales.  38  2.3.1  Study 1: Captive harbour seals (Fraser River pups)  Days in Captivity 12  3 4  14  21  23 24  Recovery  Recovery  H-  v  20 pups captured  H—H  Release of 7 seals  Liver & skinblubber biopsies obtained («=20)  Oral BNF treatment (3 treated, 3 controls)  33 ^.  1  Liver & skin biopsies obtained from BNF study seals («=6) Release of remaining seals  Figure 2.2: Study 1: Captive harbour seals. Twenty harbour seal pups were captured from the Fraser River estuary, BC, temporarily held in captivity, and then released. Liver and skin-blubber biopsies were taken from all twenty seals one or two days post-capture. Six seals were used for an in vivo induction study, and liver and skin biopsies were taken a second time from these animals. Seals were treated with 50 mg/kg BNF for three days. Skin biopsies were taken from the hind flipper rather than from the pelvic region during the induction study.  2.3.1.1 Animal capture and care Twenty 3-6 week-old harbour seal (Phoca vitulina) pups were live-captured from the middle arm of the Fraser River estuary, British Columbia (49°10'N, 123° l l ' W ) on August 17-18, 2000. Beached seals were caught with a specialized seine net as they entered the water (Redden Net Company, Vancouver, Canada) (Jeffries et al., 1993), and the ages of the pups were estimated by body weight, as described elsewhere (Cottrell et al., 2001). Seals were transported to the Department of Fisheries and Oceans Cultus  39  Lake Salmon Research Laboratory in canine transport cages and were kept in four outdoor freshwater pools (5 m diameter x 1.7 m deep) with 3 m diameter haul-out islands. Flow-through freshwater was maintained at 12°C. Seals were fasted for two days upon arrival and then subjected to liver biopsy surgery under general anaesthetic. Following surgery, seals were fed herring twice daily, supplemented with multivitamins and sodium chloride twice weekly. Because seals were not yet consuming fish in the wild, they were initially fed herring by hand. Seals were weighed at capture and then again every four to five days during their time in captivity. Seals were captured under Fisheries and Oceans Canada Animal Care Committee and Scientific Use Permits issued to P.S. Ross and were cared for by a professional animal care technician (J. Sicree) from the Marine Mammal Center, Sausalito, C A , under the supervision of veterinarians (D. Huff, M . McAdie, and D. Clegg). Seals were held in captivity for either two or five weeks and were subsequently released at the capture site. Sample size reflected logistical and housing constraints, the relative success of previous sampling efforts, and the requirements for the different biological endpoints being estimated. 2.3.1.2 Tissue biopsies Liver and skin/blubber biopsies were obtained by veterinary staff two days postcapture, while animals were under general anaesthetic. Ten minutes prior to induction of anaesthesia, seals were given an intramuscular injection of 0.02 mg/kg atropine sulphate in the gluteal muscles to maintain constant heart rate. Anaesthesia was induced with 45% isofluorine gas, using an open-ended 2 L plastic beverage bottle as a mask. Seals were then intubated with a lubricated endotracheal tube (5 mm internal diameter) fitted with an inflatable cusp. Oxygen flow rate was 2 L oxygen per minute. Biopsy sites were  40  shaved with clippers and scrubbed with betadine, hibitane, and isopropyl alcohol. Liver biopsies were taken through a 1 cm incision using a 14 gauge x 10 cm E Z Core biopsy needle (Products Groups International, Inc., Lyons, CO, USA). Skin and blubber samples were taken from the side of the body anterior to the pelvis using an 8 mm biopsy punch (Acuderm, Ft. Lauderdale, FL, USA). Xylocaine topical anaesthetic was sprayed on the skin/blubber biopsy site post-biopsy. Liver biopsies were rinsed in phosphatebuffered saline (PBS) solution, weighed, and placed in cryovials in liquid nitrogen. Skin/blubber biopsies were wrapped in aluminum foil, placed in cryovials and stored in liquid nitrogen. Seals were kept indoors overnight in dry, sterilized fiberglass pools to recover following tissue sampling.  2.3.1.3 fi-Naphthoflavone treatment  Figure 2.3: Chemical structure of B-naphthoflavone (BNF) Beta-naphthoflavone (BNF), a CYP1A inducer, was used in this study as a positive control for C Y P 1 A expression and was orally administered to three seals at twenty-three days post-capture, while three additional seals served as negative controls. Gelatin capsules containing pure BNF were prepared and administered in herring.  41  Control seals received only herring. Seals were dosed with B N F at 50 mg/kg body weight/day for three days.  This dose regime was based on dosages reported in the  literature for laboratory animals (Boobis et al., 1977; Vyas et al., 1983). A l l seals were fed an equal number of fish per day, and feedings were done at the same time each day. Seals that were not yet feeding on their own were chosen for the treatment group and were hand-fed to ensure that they received the full dosage of BNF. Control seals were free-feeding.  A l l six seals were males.  Liver and skin biopsies were obtained  approximately 24 hours after the last B N F dose while seals were under general anaesthetic.  2.3.2  Study 2: Free-ranging harbour seals Skin biopsies were obtained from young and adult harbour seals in the field from  several sites around Hornby Island, B C (August 8-9), the Fraser River estuary, B C (August 14), Boundary Bay, B C (August 22), and from Puget Sound, Washington (September 18-19) (see Table 2.2 for sample summary). Seals were captured using the same technique described in section 2.3.1.1 at the Fraser River, Boundary Bay, and Puget Sound sites. Due to the rocky terrain around Hornby Island, seals were captured by hand or using a 60 cm diameter fish landing net (Cottrell et al., 2002). Seals were manually restrained, weighed, and the biopsy site was shaved using clippers or a straight-edge razor and was cleaned as described above (section 2.3.1.2).  Lidocaine (0.7 cc) was  injected intradermally to anaesthetize the biopsy site. Two skin/blubber biopsies were taken from the side of the body anterior to the pelvis using an 8 mm biopsy punch (Acuderm). Xylocaine anaesthetic spray was used post-biopsy. Biopsies were wrapped  42  in foil, placed in cryovials, and immediately frozen in liquid nitrogen.  Seals were  released immediately at the site of capture after sampling was completed. For the purposes of this study, adult harbour seals from the Fraser River and Boundary Bay were considered as a single population. These groups are combined in subsequent sections and are referred to as 'adults'. 2.3.3  Study 3: Free-ranging killer whales Skin/blubber biopsies were obtained from three transient and ten northern resident  killer whales (Orcinus orcd) by G . M . Ellis , J.K.B. Ford , and L . Barret-Lennard 1  1  2  ('Fisheries and Oceans Canada, Vancouver Aquarium) during the summer of 2000. 2  Biopsies were obtained from live, free-ranging whales using a light-weight pneumatic dart system described elsewhere (Barrett-Lennard et al., 1996). A skin/blubber biopsy was also obtained from an orphaned southern resident killer whale using an 8 mm biopsy punch during its relocation from Washington State to British Columbia waters in July, 2002. A l l biopsies were immediately stored in liquid nitrogen. The sex and approximate age of almost all resident killer whales and a large proportion of the transient killer whales that frequent B C coastal waters are known due to a photo-identification study that has been ongoing since 1973 (Ford et al., 1994; Ford and Ellis, 1999).  <P 2. o >  3* 1  §§ r-•  ro  5  '  cd |  -d  cn  ro o  Hi o oP  S - 2. •  3 fa. ro CO  H CD P  o 3* P  3  fa 3 "EL ro*  D.  £2. £  cn  "1  ro'  ^  3  ro -t  S-  C  *° sr  i  3  CL.  ?<3 ro*  3" ro  a- S.  sH r-+  cn  <  2  3 ^  u  w  ?0 « ro  oo_  j  ^  p . 3" ro fa  0  c CO  ro  to o o o  to o o o  03  Od b  03 h  to 0 0  -a  c  OQ ro r-r  c  ro  3  ON  CO  ro fa  to o  P> 3  X p  cr  cr  c  c  *-t O  ro p  cr o  ro  ro p  ro p  to  to o o  to o o o  0  CO  EL  £L  ro  X p  P —i  cr o c  cn  to 0 0  0 0  X o  03  3 cr  3  •I  o  cn  O  X P  C/)  T3  ro o ro' cn  ro  to o o to  o  p cr  c  < g 5 c  T3 C T3  p 3  c 'a  o  c  ro p  T! ro  ro  Q » •-t *<  03  ro  > c  ac  ON  1 —  C/3 p 3  P cn  P co  'ELS'  < ro  Oq  p  c  > OQ ro o p  ro  OQ o p  T3  ro  00 o  4^  GO  00 ro  4^  4^  X  £L  *n  ro to o  p  o'  2;  SO 4^  O O  4^  4^  03  -0  OO  ,5-111  21 >  .75-29  2; >  o  to  t  o  |0q Oq % **+ o  Oq"  o p  c  OP  o  3-  0 0  >—  c o  o  #  T3  fa .V ro 3 ro o o 3  O.  2* >  2;  2;  2;  2;  0 0  o  o o 3  T3  o o  T3  3  O  o o 3 •o ro* ro  3 O  3 o  P  3  "El ro" to  Os  * P  ON  +  r+ •-t ro  3 i-f  C/3  0  NFed rol:  2;  to  a Od +  NFed rol:  3  133  a 03 +  ro  +  p 3  o o P  n  3 C  ^  ro O  r-r,  fa 3. 3 P  3  44  2.4  Tissue Preparation Liver and skin biopsies were transferred from liquid nitrogen to aluminum weigh  boats and were kept frozen on a dry ice/isopropanol slurry during sub-sampling. If a sample size was sufficiently large, a piece of unthawed tissue was archived in liquid nitrogen. Tissue samples were weighed and homogenized, as described in detail below (sections 2.4.1, 2.4.2 ). Liver and skin homogenates were spun for 20 minutes at 9000 x g (5°C). Post-mitochondrial supernatant or S9 fraction was separated from pellets using a Hamilton syringe, and pellets were discarded. The S9 fractions were stored at -80°C. Microsomal fractions were not prepared due to small tissue sample sizes. Although microsomal fractions are more concentrated than S9 fractions with respect to C Y P enzymes, we would not have been able to recover a volume of microsomes sufficient to conduct our experiments. 2.4.1  Liver Liver biopsies were homogenized in 1.5 ml polypropylene mini-centrifuge tubes  (SIGMA Chemical Co., St. Louis, MO, USA) with ice cold 50 m M Tris-HCl buffer (150 m M KC1, 2 m M E D T A , pH 7.4 at 4°C) using a hand-held pellet pestle mixer (Kontes, 1.5 ml). Buffer was added in an approximate 2:1 ratio (v/w), and samples were homogenized for a constant amount of time (5 bursts for 10 seconds each with 10 seconds between bursts). Samples were kept in an ice bath during homogenization. 2.4.2  Skin Blubber was removed from skin biopsies using a straight-edge razor blade. The  skin-blubber interface was defined by a change in tissue consistency (blubber is softer) and the visible difference in colour between the two tissues. As much hair as possible  45  was removed from harbour seal skin where samples had been obtained from unshaven biopsy sites (Study 1). Frozen skin samples were sliced into fine sections using a straight edge razor. Harbour seal skin biopsies for the oral B N F study were homogenized with Tris-HCl buffer in an 8:1 ratio (v/w), and untreated skin biopsies from 2000 were homogenized using a 4:1 v/w ratio. A higher v/w ratio was used for B N F study samples because these biopsies were smaller, and the homogenizing generator had a minimum volume requirement. Skin biopsies from the BNF study were smaller (thinner) because they were taken from the flipper; whereas, all other skin samples were taken from the side of the body. For homogenization of captive harbour seal skin samples, half of each sample was kept frozen in a mortar with liquid nitrogen, while the other half was placed in a 5 ml polypropylene test tube (BD Falcon Canada, Oakville, ON) in an ice bath. The first half of the sample was homogenized in the test tube for 3 cycles (10 seconds homogenize/45 seconds rest on ice), and any large pieces of tissue remaining were then refrozen with liquid nitrogen and homogenized for another 2 cycles. The other half of the sample was then added to the test tube, homogenized for 1 cycle, refrozen, homogenized for 3 cycles, refrozen, and homogenized for a final 10 seconds. A relatively large piece of tissue remained after homogenization. A Polytron tissue homogenizer with a stainless steel generator (7 mm diameter) was used to homogenize skin samples from 2000. Sonication and grinding of skin in liquid nitrogen with a mortar and pestle were also tried, but these methods were not successful. Killer whale and harbour seal skin samples from 2001 were homogenized using a PowerGen 125 tissue homogenizer with a stainless steel saw-tooth generator (Fisher Scientific, Pittsburgh, PA).  This generator allowed for smaller volumes than the  46  generator used previously (250 pi vs. 500 ul minimum) and, therefore, allowed for homogenization of smaller pieces of tissue. A 4-6:1 buffer volume: tissue wet weight ratio was used, depending on the size of the tissue sample. Samples were homogenized for 15 cycles (10 seconds homogenize/20 seconds rest), with addition of the second half of the biopsy after the fifth cycle. The test tube was kept in an ice bath throughout the procedure. 2.5  Total Protein Determination Total protein content of S9 fractions was measured by the method of Lowry et al.  (1951).  Bovine serum albumin (BSA) was used to generate a standard curve. A l l  samples were measured in duplicate at an absorbance of 695 lira, using a Spectronic 20 spectrophotometer England).  (Thermo Spectronic Bausch & Lomb Analytical, Cambridge,  A quadratic polynomial curve fit equation was used to calculate the total  protein content of samples using Microsoft® Excel 97 (Microsoft Corp., Redmond, WA).  47  30 mg liver  A d d 80 u l buffer Homogenize Centrifuge  Extract ~80 ul S9 fraction containing 1.6 mg total protein  -20 ul total for EROD assay (0.1 mg total protein/well) x 3 wells  ~3 ul total for immunoblotting (20 u.g total protein/lane) x 2 lanes  B. 50 mg skin A d d 300 u l buffer Homogenize Centrifuge  Extract -270 ul S9 fraction containing 1.2 mg total protein  -90 ul total for EROD assay (0.1 mg total protein/well) x 3 wells  -25 pi total for immunoblotting (45 ug total protein/lane) x 2 lanes  Figure 2.4: Schematic for preparation of S9 fractions from seal liver (A) and skin (B).  O X3  fa  >  f=  O  a.  n>  >-t  cr  I-I Co  m  CO C/5  r-f-  cr EL  3  O  cn  FT  5'  CS  cs  &.  co  <  5'  CD  oo  Oh  5'  o o 1 to o o  4^  rt  S  oo  -• >  Oh Oh  HOh  HOh  o  Oh  oo  00  bo  ^-J  H- H- HK> — io—. O i> 00 4^  OJ  K>  o  oh OJ  — i»  ^_ oh Ho  4^  4^  OJ  wo  L> Ho  Ko o\  4^  bo H-  k> Ho  b\  K> O Ho on  cr o o OQ  n cs  K>  b\ Ho  4^  OJ  Ho b\  K> io Ho  OJ  — i*  Of!  o Ho  bo Ho  b\ HO  4^  fl> O  3 fJQ  817  n> g CJ  49  2.6  Enzyme Activity Ethoxyresorufin t9-deethylase (EROD) was measured based on the methods of  Burke and Mayer (1974) and Kennedy and Jones (1994) using 96-well flat-bottom microtiter plates (Microfluor 2 Black, ThermoLabsystems, Franklin, M A ) . Fluorescence was measured using a Fluoroskan Ascent F L fluorescence plate reader (Labsystems, Helsinki, Finland). Excitation and emission filter wavelengths were 544 run (half band width = 15 ± 2) and 584 nm (half band width = 16 ± 2), respectively, and a gain of 1.0 was used. Skin and liver S9 fractions, which were initially prepared in Tris-HCl buffer, were thawed, and dilutions in 50 m M Tris-HCl buffer (pH 7.4) were made as necessary. E R O D assays were conducted under low, yellow light conditions to minimize the chance of degradation of resorufin and ethoxyresorufin, although studies have shown that photolysis should not be a problem under fluorescent room lighting (Pohl and Fouts, 1980). Samples were loaded in triplicate whenever possible. Reagents and volumes for the respective standards and samples were added as shown in Table 2.4 (total volume=200 ul/well). Contents of wells were mixed manually with a multi-channel pipette after addition of HEPES buffer and again after addition of S9 fractions. Plates were placed in the plate reader and incubated at 37°C with shaking. After a pre-incubation time of 5 minutes, a fluorescence reading was taken. This was the time zero or blank reading. N A D P H was then dispensed automatically to sample wells, and after 10 seconds of shaking, fluorescence readings were taken every minute for 30 minutes (liver) or for 60 minutes (skin). The amount of resorufin formed was determined from a standard curve of resorufin concentration versus fluorescence intensity. EROD activity was calculated  50  by dividing the amount of resorufin formed by the amount of total protein loaded/well and by the reaction time (pmoles/mg/min). A reaction time that produced a resorufin concentration above the L O Q for the assay and that fell on the linear portion of the resorufin versus reaction time curve was used.  C H 0 2  HO  NADPH, O,  5  Ethoxyresorufin Odeethylase (CYP1A)  Ethoxyresorufin  Figure 2.5 : EROD reaction  Table 2.4: Reagents and volumes used in microtiter plates for the E R O D assay Reagent  Standard wells  Sample wells (liver)  Sample wells (skin)  10 pi (blank: 0 ul)  none  none  10 ul  10 pi  10 pi  180 ul (blank: 190 ul)  160 pi  140 pi  S9 fraction  none  10 ul  30 pi  2.5 m M N A D P H in HEPES buffer  none  20 ^1  20 pi  Resorufin in D M S O 25 u M ethoxyresorufin in DMSO 0.1 M HEPES buffer, 5 m M M g C l , pH 7.8 2  Total volume  200  ul  200 ul  200  ul  51  2.6.1  Determination of resorufin stock concentration The purity of resorufin can be highly variable (Stagg and Addison, 1995), and this  can result in inaccuracies in reported concentrations of resorufin standard solutions. Pure resorufin has an absorption maximum of 572 nm at pH 8.0, with an extinction coefficient of 73 mlvr'cm" (Klotz et al., 1984). By measuring the absorbance of the resorufin stock 1  solution diluted in phosphate buffer, pH 8.0 at 572 nm, and using the formula (absorbance = extinction coefficient x pathlength x concentration), an accurate resorufin concentration can be calculated.  This correction was made for the resorufin stock  solution used in this study. 2.6.2  Antibody inhibition of EROD activity To determine whether EROD activity represented a specific measure of C Y P 1 A  enzyme activity in harbour seal liver and skin, antibody inhibition experiments were conducted. The EROD assay was performed as described above, except that S9 fractions were preincubated with purified rabbit anti-rat CYP1A1 IgG, rabbit anti-rat CYP1A2 serum, or control rabbit IgG or serum prior to addition of substrate.  Samples were  preincubated with antibody for 10 minutes at room temperature. Antibody concentrations were chosen based on the amount of total S9 protein loaded per well and antibody was added to wells in a constant volume of phosphate-buffered saline, pH 7.4. Because the IgG concentration in rabbit anti-rat CYP1A2 serum was not known, an estimated concentration of 10 mg IgG/ml was assumed. This was the typical IgG concentration of serum reported in the literature (Harlow and Lane, 1988; Coding, J.W., 1986). It was not possible to use multiple antibody concentrations for all experiments due to insufficient quantities of samples.  52  The following samples were used in antibody inhibition experiments. A pooled liver sample from BNF-treated seals («=2) and a pooled liver sample from control (nonBNF-treated) seals (n=3) were incubated with rabbit anti-rat CYP1A1 IgG at three different concentrations. A pooled liver sample from BNF-treated animals («=3) was also incubated with rabbit anti-rat CYP1A2 serum at two different IgG concentrations. A pooled skin sample from control animals (n=3) was incubated with a n t i - C Y P l A l IgG at a single IgG concentration. A liver sample from one seal pup that was not part of the study group (captured near Sidney, B.C.) was incubated with rabbit anti-rat C Y P 1 A 2 serum at three different concentrations. A pooled skin sample («=8) was incubated with rabbit anti-rat CYP1A2 serum at three different concentrations. 2.6.3  BNF inhibition The EROD assay was performed as described in section 2.6, except that different  concentrations of B N F dissolved in D M S O were added to S9 fractions prior to the addition of substrate. Final B N F concentrations were chosen based on the final substrate (ethoxyresorufin) concentration (four times higher to twenty times lower than the final substrate concentration). A liver sample from a BNF-treated harbour seal was used for this experiment. 2.7  SDS-PAGE and Immunoblots Sodium dodecyl sulphate polyacrylamide gel electrophoresis (SDS-PAGE) was  performed essentially as described by Laemmli (1970), using a Bio-Rad MiniPROTEAN® 3 Cell (Bio-Rad Laboratories, Hercules, CA).  A 0.75 mm discontinous  SDS-polyacrylamide gel was used (3% stacking gel, 7.5% separating gel). Samples were diluted in sample dilution buffer (62.5 m M Tris-HCl, pH 6.8; 10% glycerol; 0.001%  53  bromphenol blue; 1% SDS; 5% B-mercaptoethanol), boiled for 3 minutes, and loaded at 20 ul/well (see Table 2.5 for total protein concentrations). Electrophoresis was carried out at constant current (30 mA) with cooling for approximately one hour.  Proteins  resolved by SDS-PAGE were transferred electrophoretically to 7 cm x 9 cm Immobilon-P polyvinyldenfluoride (PVDF) membranes (Millipore Corporation, Bedford, M A ) using a modification of the procedure described by Towbin et al. (1979). A Bio-Rad Mini TransBlot® electrophoretic transfer cell was used, and transfers were run at constant voltage (100 V ) for one hour with cooling. After the transfer, gels were fixed for 1 hour in a solution containing 25% isopropanol and 10% acetic acid to remove SDS, stained for 1 hour in a 0.05% Coomassie blue solution, and destained for approximately 3 hours in a solution containing 10% isopropanol, 10% acetic acid. Gels were dried in a slab gel drier with vacuum aspiration for approximately 2 hours. Membranes were left overnight in blocking buffer (10 m M Tris-HCl, pH 7.4; 50 m M NaCl; I m M E D T A ; 5% skim milk powder) at 4°C. The next day, membranes were incubated with polyclonal rabbit anti-rat CYP1A2 serum (1:500 dilution) in antibody dilution buffer (0.05% Tween 20, 2% B S A , 5% goat serum in modified PBS) for 2 hours at 37°C with shaking, followed by three 510 minute incubations with wash buffer (0.05% Tween 20 in modified PBS). Membranes were then incubated with alkaline phosphatase-conjugated goat anti-rabbit IgG (1:3000 dilution) for 2 hours at 37°C with shaking, followed again by three incubations with wash buffer. Immunoreactive proteins were detected colorimetrically by reaction with a substrate solution containing 0.01% NBT, 0.05% BCIP, and 0.05 m M MgCb in 0.1 M Tris-HCl buffer, pH 9.5 at room temperature. The reaction was stopped with the addition of distilled water when protein bands were sufficiently stained.  54  2.8  Quantitation of Immunoblots Staining intensities of protein bands were quantified with a pdi 320 oe scanning  densitometer  using Quantity One® version 4.2.0 software (Bio-Rad Laboratories,  Hercules, CA). Staining intensity was measured as a contour quantity (CQ), calculated by the software program as optical density x contour area (OD x mm ).  Staining  intensities of protein bands were then normalized to a purified rat hepatic CYP1A1 standard that was included on each gel. Immunoblot data are shown as relative contour quantities (RCQ)/mg total protein. RCQs for liver and skin samples were calculated using internal standards of 0.1 pmoles rat C Y P l A l / l a n e  and 0.01 pmoles rat  C Y P l A l / l a n e , respectively. Average RCQs for C Y P 1 A protein were based on duplicate or triplicate determinations for each sample. When both upper and lower C Y P 1 A protein bands were quantifiable, these values were added together.  C Y P 1 A protein levels  determined this way represent total C Y P 1 A protein (i.e. CYP1 A l and CYP1A2). Table 2.5: Amount of total protein used for the EROD and immunoblot assays Sample  EROD assay (mg protein/well)  Immunoblots (ug protein/lane)  Oral BNF study harbour seal liver (treated)  0.01, 0.025, 0.05  2  Oral B N F study harbour seal liver (controls)  0.1,0.05  10  Oral B N F study harbour seal skin (treated)  0.075  10  Oral BNF study harbour seal skin (controls)  0.075  25  Fraser River harbour seal pup liver  0.1  20  Fraser River harbour seal pup skin  0.12  20-60  Puget Sound harbour seal pup skin  0.1  Adult harbour seal skin  0.1  Hornby Island harbour seal pup skin  0.12  Killer whale skin  0.12  1  \J/  when possible, 60 ug total protein/lane was used. Many skin samples did not have sufficient total protein content to load 60 ug/lane. For these samples, the maximum amount of total protein was loaded by diluting S9 fractions with sample dilution buffer in a 1:1 ratio.  t  55  2.9  Contaminant Analysis Blubber biopsies from seal pups from the Fraser River and Hornby Island and  from killer whales were analyzed by the Regional Dioxin Laboratory (Institute of Ocean Sciences, Sidney, B.C.) for congener-specific PCBs, PCDDs, and PCDFs using high resolution gas chromatography/high resolution mass spectrometry.  Total Toxic  Equivalents (TEQs) were calculated using the most recent toxic equivalency factors (TEFs) defined by Van den Berg et al, 1998. 2.10  Statistical Analysis Correlation analyses and t-tests were performed using Microsoft®  Excel.  Unpaired two-tail t-tests were used for comparisons of means, except for analysis of the change in C Y P 1 A expression over time, in which a paired t-test was used. The Grubbs' test was used to identify outliers (GraphPad Software, Inc., San Diego, CA). Normality tests were not conducted due to small sample sizes (InStat version 3.00, GraphPad Software, Inc.), and parametric tests were used.  A p value < 0.05 was considered  statistically significant for all analyses. 2.11 Assay Validation and Optimization Intra- and inter-assay experiments were conducted to determine the limit of detection (LOD) and limit of quantitation (LOQ) for the EROD and immunoblot assays. Because we adapted previously published methods and wanted to use as little tissue as possible, additional experiments were conducted to determine the optimal conditions for analysis of our tissue samples.  56  2.11.1  EROD Assay  2.11.1.1  Intra-assay variation  Intra-assay variability of the EROD assay was measured in a single assay using six resorufin standard concentrations, each analyzed in triplicate.  Coefficients of  variation (CV) were <15% at all standard resorufin concentrations (Table 2.6). 2.11.1.2  Inter-assay variation  Multiple between-day standard curves were obtained with five different resorufin concentrations, each analyzed in triplicate. The limit of detection (LOD) for the E R O D assay was defined as the lowest standard concentration having a fluorescence reading above background, where background was the fluorescence measured in wells containing HEPES buffer and ethoxyresorufin. The limit of quantitation (LOQ) was defined as the standard  concentration  at which fluorescence  readings  background with a coefficient of variation (CV) <15%.  were consistently  above  A l l standard concentrations  tested had C V ' s <15%. At 0.65 pmoles resorufin/well, fluorescence readings were not consistently above background, but at  1.3 pmoles resorufin/well, readings were  consistently above background. Based on these results, the L O D was determined to be between 0.65 and 1.3 pmoles resorufin/well. The LOQ was determined to be 1.3 pmoles resorufin/well (Table 2.7). This LOQ value was used for EROD activity in harbour seal samples.  Table 2.6: Intra-assay variation of fluorescence readings for EROD assay standards Resorufin standard (pmoles/well)  Average fluorescence  CV (%)  .0  7.4 ± 0.2  2.2  0.65  7.4 ± 0 . 4  5.0  1.3  8.9 ± 0 . 2  2.8  2.6  11.1 ± 0 . 3  2.4  5.2  17.5 ± 0 . 9  5.3  10.4  26.5 ± 0.6  2.4  20.8  46.0 ± 1.5  3.4  V a l u e s are p r e s e n t e d as the m e a n ± S D o f t r i p l i c a t e f l u o r e s c e n c e r e a d i n g s  58  Table 2.7: Inter-assay variation of fluorescence readings for EROD assay standards Resorufin standard (pmoles/well)  Mean fluorescence  CV (%)  0  5.9±0.4  a  6.7  0.65  6.9±0.5  b  7.7  1.3  8.7±0.4  b  5.0  2.6  10.9±0.6  a  5.5  5.2  17.3 ± 1.2  7.0  10.4  27.7 ± 1.2  a  4.8  20.8  47.7 ± 2.2  a  4.7  a  Values are presented as the mean ± SD of triplicate fluorescence readings Calculated for six different assays calculated for three different assays b  Figure 2.6: Inter-assay calibration curve for the EROD assay. Fluorescence readings were measured in triplicate on six different days for each resorufin standard concentration, with the exception of 0.65 and 1.3 pmole/well concentrations, which were measured on three different days.  59  2.11.1.3  Substrate concentration for the EROD assay  The effects of substrate concentration on product formation were investigated for harbour seal liver. E R O D activity was measured in S9 fractions from four seals using a range of substrate concentrations. The seals used were PV00101 (pup from near Sidney, B.C.), PV01-50 (adult female from the Fraser River estuary), PV0018 (pup from Fraser River estuary), and PV98102 (pup from near Sidney, B.C.). Other than PV0018, samples from these seals were only used in optimization experiments, but not as part of the overall study. The effect of substrate concentration on resorufin formation is shown in Figure 2.7. The results suggested that a substrate concentration of 25 u M was optimal for seal liver. At concentrations above 25 uM, resorufin formation decreased.  Figure 2.7: Effect of substrate concentration on resorufin formation in harbour seal liver. Liver S9 fraction was from one harbour seal pup (PV00101). Data points are the averages of triplicate fluorescence readings.  60  2.11.1.4  Total protein concentration for the EROD assay  The effect of total protein concentration on product formation was measured in liver of the same four seals listed in section 2.11.1.3. There was a relatively linear relationship between resorufin production and total protein concentration up to a concentration of 0.5 mg/well. A representative plot is shown in Figure 2.8. A protein concentration that falls on the linear portion of the curve should be used for EROD assays.  Based on these results, 0.1 mg total protein/well was chosen as the loading  concentration for liver samples.  0 -I 0.0  1  1  1  ,  r  0.1  0.2  0.3  0.4  0.5  0.6  Total protein (mg/well)  Figure 2.8: Effect of total protein concentration on resorufin formation in harbour seal liver. Liver S9 fraction was from one harbour seal pup (PV00101). Data points are the averages of triplicate fluorescence readings.  61  Serial dilutions of skin S9 fractions from two harbour seals, PV01-11 (pup from Fraser River) and PV01-34 (adult from Boundary Bay), were used to look at the effect of protein  concentration  on product  formation in skin  homogenates.  Resorufin  concentration increased linearly with protein concentration and then reached a plateau (Figure 2.9).  It should be noted, however, that readings for the three lowest protein  concentrations were below the limit of detection. Based on these results, a minimum of 0.1 mg total protein/well was chosen as the loading concentration for skin samples.  B  a  o  a •a  a  a is  0.1  0.2  0.3  0.4  Total protein (mg/well)  Figure 2.9: Effect of total protein concentration on resorufin formation in harbour seal skin. Skin S9 fraction was from one harbour seal pup (PV01-11). Data points are the averages of triplicate fluorescence readings. 'Data points' that were below the LOQ are shown only for comparative purposes.  62  2.11.1.5  NADPH concentration for the EROD assay  The concentration of N A D P H , a cofactor in the EROD reaction, was also varied to determine i f it affected EROD activity in harbour seal liver. Increasing the initial N A D P H concentration above 2.5 m M did not have an effect on E R O D activity, indicating that N A D P H was in excess at this concentration. A n initial concentration of 2.5 m M was used for subsequent EROD assays. This concentration is comparable to concentrations used in other EROD assay protocols. 2.11.1.6  Effect of skin biopsy size on total protein yield and EROD activity  To determine whether the amount of tissue homogenized and the buffer dilution factor had an effect on total protein yield and/or EROD activity, different amounts of skin from two harbour seals, PV01-26 (pup from Boundary Bay) and PV01-32 (adult from Boundary Bay), were homogenized in a fixed volume of buffer. The effect of biopsy size on product formation was variable (Figure 2.10).  Different pieces of skin likely  contained variable amounts of CYP1A, and this factor likely produced some of the scatter observed.  Figure 2.11 shows data from the same two seals together with data from  PV01-11 and PV01-34. Total protein yield increased linearly as the amount of tissue homogenized was increased (Figure 2.11).  I  1  1  1  0  20  40  60  80  1  1  1  100  120  140  A m o u n t of tissue homogenized (mg)  Figure 2.10: Effect of biopsy size and buffer dilution factor on resorufin formation harbour seal skin.  10 Skin •  PV01-11  v  PV01-34  •  PV01-26  O  PV01-32  6H  O  O  o  0  100  200  300  400  A m o u n t of tissue homogenized (mg)  Figure 2.11: Effect of biopsy size on total protein yield for harbour seal skin.  64  2.11.1.7  Reaction time  Resorufin formation was linear with respect to time for both harbour seal liver and skin (Figure 2.12). EROD activity was calculated using data points from the linear portion of the product vs. time curve for each sample. EROD activity was calculated using a reaction time of 40 minutes for skin samples from Fraser River pups versus 20 minutes reaction time for skin samples from the other populations (Figure 2.12C), as the reaction profile was noticeably different for the Fraser River pups (Figure 2.12B).  65  B  Skin (captive pups) • O T  £ 6  JQ  "3  PVOOIO PV002 PV00I7  •  E  •  •  T  • LOQ  T  « T t  T T  o o  o  O  o 30  40  Time (min)  3.0  c  Skin  (adults) T  "33  T T •  • •O  •  T  O  •  • o T  T o T  0  8  a 0  O  •  LOQ  0.0  T  T  1  10  PV01-I7 PV01-18 PV01-23  1  20  30  40  50  60  70  Time (min)  Figure 2.12: Resorufin formation over time in harbour seal liver (A) and skin (B,C). Reaction profiles for three different seals are shown in (B) and (C).  66  2.11.2  Immunoblot Assay  2.11.2.1  General protocol for the immunoblot assay  Experiments involving minor modifications of the Laemmli (1970) and Towbin (1979) protocols were conducted to optimize the detection of C Y P 1 A bands in harbour seal samples. Both a large gel system (Hoeffer) and mini-gel system (Bio-Rad) were used.  Although protein bands were detected by both systems, the mini-gel protocol  appeared to be more sensitive. This may have been due to the membranes that were used with the different systems (nitrocellulose for large gels vs. PVDF for mini-gels), as P V D F membranes have a higher binding capacity for proteins (170-200 ug/cm vs. 802  100 ug/cm ) (Bio-Rad Blotting Guide). However, PVDF membranes may also have been 2  responsible for the higher degree of non-specific protein staining obtained with mini-gels. We varied the composition of the electrophoresis, transfer, and sample dilution buffers using mini-gels in an attempt to increase band resolution and staining intensity for the skin samples. Varying the buffer composition had no effect. Chemiluminescence was also tested in this study. A few harbour seal and killer whale skin samples were run on large gels, transferred to nitrocellulose membranes, and visualized using chemiluminescence.  Chemiluminescence was expected to be more  sensitive for detecting proteins on immunoblots than the colorimetric technique used in this study; however, it did not improve our ability to detect C Y P 1 A in harbour seal or killer whale skin and was not further pursued. 2.11.2.2  Total protein concentration for the immunoblot assay  We varied the amount of total protein loaded per lane for harbour seal S9 fractions in order to obtain an optimal response.  In general, skin samples (with the  67  exception of B N F study samples) for both harbour seals and killer whales were loaded at the highest total protein concentration possible by diluting S9 fractions with sample dilution buffer in a 1:1 ratio, up to a maximum of 60 ug total protein/lane. This was necessary because the total protein content of skin S9 fractions was low relative to liver (approximately four times lower), and protein bands were often faint or undetectable for skin samples. Sample loading concentrations are summarized in Table 2.5. 2.11.2.3  Cross-reactivity of antibodies with harbour seal CYP1A  To investigate the specificity and sensitivity of different antibodies with putative harbour seal C Y P 1 A proteins, four different primary antibodies (rabbit anti-trout CYP1A1 IgG, rabbit anti-rat CYP1A2 serum, mouse anti-rat CYP1A1 IgG, goat antirabbit 1A1/1A2 IgG) were tested with harbour seal liver S9 and microsomal fractions. Mouse anti-rat CYP1A1 monoclonal IgG and rabbit anti-rat CYP1A2 polyclonal serum were most cross-reactive.  The rabbit anti-rat CYP1A2 serum was used for the  immunoblot analyses of harbour seal samples. Killer whale skin samples were probed with rabbit anti-rat CYP1A2 serum and rabbit anti-trout CYP1A1 IgG, and similar results were obtained with these two antibodies. Rabbit anti-rat CYP1A2 serum was used for subsequent assays of killer whale skin. 2.11.2.4  Protein transer from gel to membrane  To determine if proteins in the molecular weight range of interest (~ 55,000 Da) were being transferred from gels to membranes, four gels were stained with Coomassie blue after the transfer step. Protein bands on stained gels were compared with bands on immunoblots by measuring the distance that bands had migrated from the top of the gel or membrane respectively. Although the protein bands that were due to non-specific  68  staining were not transferred completely from gels to membranes, it appeared that the putative C Y P 1 A proteins completely transferred. 2.11.2.5  Intra-assay variation  Three purified rat CYP1A1 concentrations were loaded in five lanes on two gels. Only three concentrations could be assayed due to the limited number of wells in the mini-gels. The 0.01 pmole standard was used as the internal standard to calculate relative contour quantity (RCQ) values for the 0.005 and 0.025 pmole standards. The coefficient of variation was < 20% for both of these concentrations (Table 2.8).  Table 2.8: Intra-assay variation of relative contour quantities for immunoblot standards a  [Rat CYP1A1 standard]  Mean relative contour quantity  CV  (pmoles/lane)  (RCQ)  (%)  0.005  0.6 ± 0 . 1  12.2  0.01  1.0  -  0.025  2.0 ± 0 . 4  19.0  Values are presented as the mean ± SD Each standard was loaded in 5 lanes. Only three standard concentrations could be assayed due to the limited number of lanes available for loading. The 0.01 pmol/lane standard was used as the internal standard for intra-assay R C Q calculations. a  2.11.2.6  Inter-assay variation  Standard curves consisting of five different purified rat CYP1A1 concentrations were obtained from assays conducted on five different days (two curves/day).  The  coefficient of variation for all concentrations was approximately 20% (Table 2.9). The inter-assay calibration curve is shown in Figure 2.13.  69  Table 2.9: Inter-assay variability of relative contour quantity (RCQ) values for immunoblot standards 3  [Rat CYP1A1 standard]  Mean relative contour quantity  CV  (pmoles/lane)  (RCQ)  (%)  0.0025  0.3 ± 0.06  20.9  0.005  0.5 ±0.09  18.1  0.01  1.0  -  0.025  2.3 ± 0 . 6  24.4  0.05  4.1 ± 0 . 7  16.5  Vales are presented as the mean ± SD Standard curves were determined on five different days (2 curves/day). The 0.01 pmol/lane standard was used as the internal standard for intra-assay R C Q calculations. The inter-assay C V for the 0.01 pinole/lane standard, using 0.05 pmoles/lane as the internal standard was 17.8%. a  Figure 2.13: Inter-assay calibration curve for immunoblots. R C Q - relative contour quantity of each CYP1A1 band relative to that of the internal standard (0.01 pmole rat standard).  70  Purified harbour seal C Y P 1 A is not yet available, therefore rat C Y P 1 A was used as the internal standard for all sample C Y P 1 A protein bands. Because the standards were from a different species than the study samples, the standards could not be used to obtain an absolute measure of C Y P 1 A in harbour seal S9 fractions. In order to determine if harbour seal protein bands were quantifiable, R C Q values of sample bands were compared to those of rat standards.  As 0.0025 pmoles was the lowest rat standard  concentration analyzed, a value could not be reported for the limit of detection (LOD). The limit of quantitation (LOQ) was defined as the mean inter-assay relative contour quantity (RCQ) value obtained for the 0.0025 pmole rat CYP1A1 standard (RCQ=0.3), where 0.01 pmoles rat C Y P 1 A/lane was the internal standard.  71  3.  3.1 3.1.1  RESULTS  Study 1: Captive Harbour Seal Study Induction of CYP1A by in vivo BNF treatment Liver and skin biopsies were obtained from six of the captive harbour seal pups  for an induction study (3 BNF-treated, 3 controls). Two C Y P 1 A protein bands having approximately the same mobility as a purified CYP1A1 rat standard were detected by immunoblot analysis.  Protein bands are shown in Figures 3.1 (liver) and 3.2 (skin).  C Y P 1 A protein bands for liver and skin of BNF-treated harbour seals are shown on the same immunoblot in Figure 3.3.  These protein bands had approximately the same  mobility as C Y P 1 A in liver microsomes from a 3-MC-treated rat. EROD activity was quantified in all samples, except for a skin sample from one BNF-treated seal (below LOQ). 3.1.1.1 Immunoblot analysis Oral B N F treatment caused induction of CYP1A protein levels in both liver and skin of harbour  seals (Figures 3.1-3.4).  Hepatic C Y P 1 A protein levels were  approximately five times higher in treated seals than in controls, as assessed by combining upper and lower bands (p=0.0001). Individually, both upper (p=0.00003) and lower (p=0.003) C Y P 1 A bands were also significantly induced in treated seals compared to control seals. A t-test could not be performed for skin as data were obtained for only one control skin sample. However, cutaneous CYP1A protein levels were approximately sixteen times higher in BNF-treated seals («=3) than in one control seal.  72 CYP1A1?  1  CYP1A2?  3  2  4  6  5  7  8  Figure 3.1: Immunoblot of BNF-treated and control harbour seal pup liver samples probed with rabbit anti-rat CYP1A2 polyclonal serum (1:500 dilution). Lanes 1 and 8 are purified rat CYP1A1 standards, applied at 0.1 pmoles and 0.05 pmoles per lane respectively. Lanes 2-4 are liver S9 fractions from control (untreated) seals PV001, 07, and 16, each loaded at 10 ug per lane. Lanes 5-7 are liver S9 fractions from BNF-treated seals PV0012, 14, and 15, each loaded at 2 ug per lane. The membrane was developed with substrate solution for 2 minutes.  CYP1A1?  1  2  CYP1A2?  3  4  5  Figure 3.2: Immunoblot of BNF-treated and control harbour seal pup skin samples probed with rabbit anti-rat CYP1A2 polyclonal serum (1:500 dilution). Lane 1 is a skin S9 fraction from control (untreated) seal PV0016. Lanes 2-4 are skin S9 fractions from BNF-treated seals PV0012, 14, and 15, respectively. Lane 5 is a purified rat CYP1A1 standard, applied at 0.01 pmoles per lane. Control and treated S9 fractions were applied at final concentrations of 25 ug and 10 ug total protein, respectively. The membrane was developed with substrate solution for 3 minutes.  73  1  2  3  4  Figure 3.3: Immunoblot showing both BNF-treated liver and skin samples probed with rabbit anti-rat CYP1A2 polyclonal serum (1:500 dilution). Lanes 1 and 2 are skin S9 fractions from BNF-treated seals PV0012 and 14, applied at a final concentration of 10 ug total protein per lane. Lane 3 is a 3-MC-treated rat liver microsomal sample, loaded at 0.07 pg per lane. Lane 4 is a liver S9 fraction from BNF-treated seal PV0012, loaded at 2.5 pg per lane. The membrane was developed with substrate solution for 5 minutes.  n = ^ cro c  rt W w  CO  Hepatic C Y P 1 A protein ( R C Q / m g protein)  p cTm CS CS til 3 fD  ^ -+ j" CD  cr >-s  o  «K  ft. OT  o  cn o o  n> o  II  o o o o  cr P3  < «  n> i—  o  co  1  o  r-t-  s  n CS  0Q ft)  o o o  Ul o  P.  cs i-t-  cs OCS CO  to  O •<  > i-i  O  2 5"  e cr  < a . ro CO  CS o  N  Cutaneous C Y P 1 A protein ( R C Q / m g protein)  >  o o  cr ro &. CS  co  05  5' 3'  5  W  a  o  * s. O •-+> o ^  CJ. <: fl> cr  ^& 3 cT CO  O • © ^  3 °  °£  00 O O  NJ O O  w  O  75  3.1.1.2 EROD activity Oral B N F treatment did not cause an increase in EROD activity in either liver or skin relative to control seal tissues (Figure 3.5). 3.1.1.3 BNF inhibition of EROD activity Although B N F is a known inducer of C Y P 1 A , BNF treatment has been found to inhibit E R O D activity (Haasch et al., 1993). A n experiment was carried out to determine whether in vitro B N F treatment could inhibit EROD activity in harbour seal liver, as has been shown previously in other species (Haasch et al., 1993). inhibited in vitro at all B N F concentrations tested.  EROD activity was  The lowest concentration of B N F  (0.0625 uM) caused more than 80% inhibition of EROD activity at a reaction time of 20 minutes (Figure 3.6).  Hepatic EROD activity (pmoles/mg/min) K> O  CO O  o  o  re  ^  Cutaneous E R O D activity (pmoles/mg/min) o o  o cn i  - - * - - > • o cn i i  NJ o i  NJ cn i CZ>  H  77  Figure 3.6:  Inhibition o f hepatic E R O D  activity b y  in vitro  B N F treatment.  A liver S9  fraction f r o m o n e BNF-treated seal (PV0015) w a s preincubated w i t h B N F d i s s o l v e d i n DMSO  a n d l o a d e d at 0.01 m g t o t a l p r o t e i n / w e l l .  reaction time o f 20 minutes.  EROD  activity was measured using a  78  3.1.2  CYP1A expression in tissue biopsies obtained two days post-capture  3.1.2.1 Liver Liver biopsies were analyzed for twelve of twenty harbour seal pups captured. Liver from the remaining seals was either insufficient for C Y P 1 A analysis or was used for parallel biomarker studies (vitamin A ) . EROD activity was detected in all liver biopsy samples, and activity in all but two samples was quantifiable. For most liver biopsy samples, two C Y P 1 A bands were detected on immunoblots (Figure 3.7). When both an upper and lower band were quantifiable, these values were added together. However, for some blots it was only possible to quantify one band (i.e. only one band stained strongly on immunoblots). Average values (relative contour quantities) for C Y P 1 A protein were based on duplicates or triplicates for each sample, except for PV0012, PV0017, and PV0018. Relative contour quantities (RCQs) for liver samples were calculated using an internal standard of 0.1 pmoles rat C Y P l A l / l a n e . This concentration was more appropriate for the staining intensity of protein bands in liver.  79  Table 3.1: C Y P 1 A expression in hepatic S9 fractions from Fraser River harbour seal pups Sample  Average C Y P 1 A protein levels  EROD activity  (RCQ/mg total protein)  (pmoles/mg/min)  3  PV004  16.3  1.15  PV007  18.2  <LOQ  PV009  31.3  3.90  PV0011  18.9  1.68  PV0012  16.1  1.80  PV0013  19.9  2.31  PV0014  16.3  2.28  PV0016  19.6  2.97  PV0017  29.5  2.60  PV0018  8.6  <LOQ  PV0019  24.9  4.36  PV0020  20.9  3.23  L O Q (EROD) = 0.65 pmoles/mg/min L O Q (immunoblots) = 0.3/mg total protein loaded R C Q (relative contour quantity) is the staining intensity of sample bands normalized to a 0.1 pmole/lane rat C Y P 1A1 standard. EROD activity was calculated at 20 minutes reaction time for all samples. a  80  CYP1A1?  CYP1A2?  Figure 3.7: Immunoblot of captive (Fraser River) harbour seal pup liver S9 fractions probed with rabbit anti-rat polyclonal CYP1A2 serum (1:500 dilution). Lanes 1 and 9 are purified rat C Y P 1 A 2 standards, applied at 0.05 pmoles (lane 1) and 0.1 pmoles (lane 9) per lane. Lanes 2-7 are samples from seals PV0014, 16, 17, 18, 19, and 20, respectively. Lane 8 is a liver S9 fraction from a BNF-treated seal (PV0012). The S9 fractions were applied at a final concentration of 20 ug total protein per lane (lanes 2-7) and 5 ug (lane  81  Hepatic C Y P 1 A protein levels were positively correlated with hepatic EROD activity (r =0.481, /?=0.026, n=lO) when upper and lower putative C Y P 1 A bands were 2  combined (Figure 3.8A). C Y P 1 A protein and EROD activity were also correlated when only the upper protein band was quantified (r = 0.538, /?=0.025) (Figure 3.8B). The 2  lower C Y P 1 A protein band did not correlate with EROD activity (p=0.836) (data not shown).  82  18  12  24  Hepatic C Y P 1 A protein (RCQ/mg protein)  B 5.0 c | "Hi  4.5  A  4.0  A  E  Liver (upper C Y P 1 A band only)  E  a.  3.5  u 3.0 A O  o w  2.5 r =0.538 ,0=0.025 2  2.0  1.5 8  10  12  14  16  1E  20  Hepatic C Y P 1 A protein (RCQ/mg protein)  Figure 3.8: Relationship between hepatic C Y P 1 A protein levels and EROD activity for Fraser River harbour seal pups. C Y P 1 A protein levels shown in (A) are the combined RCQ values of upper and lower C Y P 1 A bands. RCQ values for only the upper C Y P 1 A band are shown in (B).  83  3.1.2.2 Skin E R O D activity was quantified in captive harbour seal skin S9 fractions; however, activities were below the limit of quantitation for 10 of 20 samples. Two C Y P 1 A protein bands with the same relative mobility as C Y P 1 A in seal liver were detected in skin samples. Average RCQs were based on duplicate or triplicate readings when possible. There were no replicates for PV006, PV0016, PV0017, or PV0018 (replicate values were either undetectable or < LOQ).  C Y P 1 A protein levels in skin correlated with E R O D  activity in skin when upper and lower protein bands were combined (r =0.494, /»=0.023, 2  n=\0) (Figure 3.10). Individually, upper (p=0.399) and lower (p=0.222) protein bands did not correlate with EROD activity (data not shown). Both upper and lower bands were not detected in all samples.  84  Table 3.2: C Y P 1 A expression in cutaneous S9 fractions from Fraser River harbour seal pups Sample  Average C Y P 1 A protein levels  EROD activity  (RCQ/mg protein)  (pmol/rng/min)  3  PV001  11.9  0.55  PV002  8.3  0.30  PV003  20.3  <LOQ  PV004  17.1  <LOQ  PV005  17.8  <LOQ  PV006  6.4  0.64  PV007  10.6  <LOQ  PV008  14.5  <LOQ  PV009  9.0  <LOQ  PV0010  15.8  1.00  PV0011  14.6  <LOQ  PV0012  <LOQ  <LOQ  PV0013  11.2  0.83  PV0014  16.6  0.78  PV0015  5.2  0.48  PV0016  7.4  0.45  PV0017  8.4  0.54  PV0018  16.4  <LOQ  PV0019  15.1  <LOQ  PV0020  6.7  0.49  b  L O Q (EROD) = 0.27 pmoles/mg/min The EROD assay L O Q was defined in terms of pmoles resorufin formed. Because EROD activity was calculated at a different time point for Fraser River pups than the other seal groups (40 vs. 20 minutes), the L O Q , when expressed as EROD activity, is much lower for Fraser River pups. L O Q (immunoblot) = 0.3/mg total protein loaded " R C Q (relative contour quantity) is the staining intensity of sample bands normalized to a 0.01 pmole/lane rat CYP1 A l standard. E R O D activity was calculated at 40 minutes reaction time for all samples. b  85  1  2  3  4  5  6  7  8  Figure 3.9: Immunoblot of captive (Fraser River) harbour seal pup skin S9 fractions probed with rabbit anti-rat CYP1A2 polyclonal serum (1:500 dilution). Lanes 1 and 8 are purified rat CYP1A1 standards, applied at 0.01 pmoles (lane 1) and 0.05 pmoles (lane 8) per lane. Lanes 3-7 are samples from seals PV0015, 16, 01, 02, and 05, respectively. Lane 2 is a skin S9 fraction from a Puget Sound seal pup (PV01-56). The S9 fractions were applied at final concentrations of 31 ug, 68 ug, 69 pg, 57 pg, 65 ug, and 48 pg (lanes 2-7), respectively.  Skin (both CYP1A bands)  r=0.49 p=0.023 2  4  8  12  16  20  Cutaneous CYP1A protein (RCQ/mg protein)  Figure 3.10: Relationship between CYP1A protein levels and EROD activity in skin of Fraser River harbour seal pups. CYP1A levels shown here are the combined RCQ values of both C Y P 1 A bands (upper and lower).  86  3.1.2.3 Comparison of CYP1A expression in skin and liver E R O D activity in skin was negatively correlated with EROD activity in liver (r =0.782, ;?=0.046, n-5) (Figure 3.11 A). There was no significant correlation between 2  C Y P 1 A protein levels in the two tissues (p=0.227, n=\ 1) (Figure 3.1 IB). Comparison of C Y P 1 A protein band migration in liver and skin samples relative to each other and to the rat C Y P 1 A standard suggested that putative C Y P 1 A protein bands detected in skin corresponded to the same bands detected in liver (Figures 3.7, 3.9). 3.1.2.4 Effect of body weight on CYP1A expression In order to determine whether physiological factors were affecting C Y P 1 A expression in harbour seal tissues, I also looked at relationships between C Y P 1 A and body weight and differences in C Y P 1 A expression between males and females. There was a significant negative correlation between body weight at capture and hepatic EROD activity in Fraser River pups (r =0.572, p=0.011, «=10) (Figure 3.12A). However, there 2  were no significant correlations between body weight and C Y P 1 A protein levels in liver (/?=0.303, «=12) (Figure 3.12B ) or CYP1A expression in skin (EROD p=0.541, rc=10; CYP1A/?=0.975, «=19) (Figure 3.13).  87  0.9 E R O D activity (skin vs. liver)  c  |  "SJD 0.8  E  r =0.782 p=0.046 2  0.7  V a a  0.6  OS  0.5  o o  H  w  s ©  s s  •S  0.4 H  U  0.3 2.2  2.4  2.6  2.8  3.0  3.2  3.4  Hepatic E R O D activity (pmoles/mg/min)  o  u  a.  16  C Y P 1 A protein (skin vs. liver)  H  p=0.221  M  E  a u os  14  12 o  u  a.  10  CM  >U  (/3  3  O c  « 3  u  10  15  20  25  30  35  Hepatic C Y P 1 A protein (RCQ/mg protein)  Figure 3.11: Relationship between C Y P 1 A expression in liver and skin of Fraser River harbour seal pups. (A) EROD activity, (B) C Y P 1 A protein levels.  88  Body weight (kg)  B 40  Liver  ^=0.303  CYP1A protein vs. body weight •K o a.  35  01)  £  30  u B5  25  o u a.  20  >H  U  15  a. 10 <u  X  16  20  22  24  26  28  30  Body weight (kg)  Figure 3.12: Relationship between body weight at capture and hepatic EROD activity (A) and C Y P 1 A protein levels (B) for Fraser River harbour seal pups. C Y P 1 A protein values shown in (B) are the combined RCQ values of upper and lower C Y P 1 A protein bands.  89  1.2 Skin  "Si  p=0.541  E R O D vs. body weight  e 1.0  "o E  a.  -—-  0.8  u  « O 05  0.6  H  W  tw 3 o a> 0.4 C  S3  -w 9  u  0.2 18  20  —i— 22  —i—  24  26  32  30  28  B o d y weight (kg)  10  12  14  16  18  20  Body weight (kg)  Figure 3.13: Relationship between body weight and cutaneous EROD activity (A) and C Y P 1 A protein (B) in Fraser River (captive) harbour seal pups.  90  3.1.2.5 Effect of sex on CYP1A expression There was no significant difference in hepatic EROD activity (p=0.371) or hepatic C Y P 1 A protein (p=0.644) between males and females.  There were also no sex  differences in cutaneous EROD activtity (p=0.776) or CYP1A protein (p=0.854). C Y P 1 A expression was not significantly different between males and females. Table 3.3: Comparison of C Y P 1 A expression between male and female harbour seal pups from the Fraser River estuary Mean CYP1A protein levels  Mean EROD activity  (RCQ/mg protein)  (pmoles/mg/min)  Males  Females  Males  Females  Liver  19.5 ± 7 . 0 n=9  21.6 ± 3.1 n=3  2.43 ±0.87 n=l  3.09 ± 1 . 3 5 n=3  Skin  12.2 ± 4 . 2 n=l4  12.6 ± 6 . 0 n=5  0.56 ±0.19 n=8  0.56 n=2  Values are presented as the mean ± SD  3.1.2.6 Effect of contaminant levels in blubber on CYP1A expression There were no significant correlations between CYP1A expression in Fraser River harbour seal pup tissues and total TEQ for blubber (Figures 3.14A, 3.15A, 3.16). However, when mono-ortho PCBs were removed from the total TEQ value, there were significant positive correlations between EROD (p=0.037, «=10) and C Y P 1 A protein (p=0.016, n=\2) in liver and contaminants in blubber (Figures 3.14B, 3.15B).  91  30  40  50  Total T E Q (ng/kg lipid)  B  s  Liver E R O D vs. planar T E Q  I  •  "Si E o E  a.  •  •  o w  • •  Q O «  • 12  r =0.437 p=0.037 2  16  20  24  T E Q value (total T E Q - mono-ortho P C B s ) (ng/kg lipid)  Figure 3.14: Relationship between hepatic EROD activity and contaminants in blubber. Total T E Q values are shown in (A). Total TEQ values, excluding mono-ortho PCBs are shown in (B).  92  Liver e '3  35  C Y P 1 A vs. total T E Q  o  u  Q.  u> E C>  30  Pi  25  H  o >-  20  H  u  a  < PH  >U _o <s  a,  15  10  A  p=0.247  20  10  30  —i—  —i—  40  50  60  70  T o t a l T E Q (ng/kg lipid)  B 40 Liver C Y P l A v s . planar T E Q  o —  Q.  30  H  E  a u  'a 20 o  u C  PM  u  10  r =0.454 p=0.016  8  12  16  20  24  T E Q value (total T E Q - mono-ortho P C B s ) (ng/kg lipid)  Figure 3.15: Relationship between hepatic C Y P 1 A protein levels and contaminants in blubber. (A) shows total TEQ values. (B) shows total TEQ excluding mono-ortho PCBs C Y P 1 A levels are the combined RCQ values of both C Y P 1 A bands (upper and lower).  93  1.2 Skin E R O D vs. total T E Q  c  E  i.o  p=0.642  H  "o  E o.  a o  0.8  0.6  oi 3 O  <u 0.4 c es  s U  0.2  —i—  20  40  —i—  60  100  80  120  140  T o t a l T E Q (ng/kg lipid)  B 22 Skin C Y P 1 A vs. total T E Q  .S3  -t-t  20 1  Q. W)  18  '3 '2  o u  p=0.976  E  d o * o  16 H 14 12  u.  a. 10 A  < >< U to  3 O  •S  4 —— i  20  40  60  80  100  120  140  T o t a l T E Q (ng/kg lipid)  Figure 3.16: Relationship between cutaneous C Y P 1 A expression and contaminants in blubber. C Y P 1 A protein levels shown in (A) are the combined RCQ values of upper and lower C Y P 1 A protein bands.  94  3.1.2.7 Relationship between body weight and contaminant levels in blubber Contaminant levels in blubber (total TEQ) were negatively correlated with body weight of Fraser River pups (Figure 3.17).  When a significant outlier (TEQ=129.4)  (p<0.05) was removed from the data set, total TEQ was still negatively correlated with body weight 0=0.0002, «=19).  70  60  r =0.55 p=0.0002 2  •  'a. 50  •  M  •  •  •5- 40  O w H  ~S  Captive seal pups  •  •  30  o H  •  • • \.  20  •  10 16  18  20  22  24  26  28  30  32  Body weight (kg)  Figure 3.17: Total TEQ for blubber was negatively correlated with body weight in Fraser River harbour seal pups. A significant outlier (total TEQ =129.4) (p<0.05) was removed from this data set.  3.1.3  Increase in CYP1A expression over time in captive harbour seals A n increase in C Y P 1 A expression in both liver and skin was observed in control  seals (n=3) in the oral B N F study after three weeks in captivity. C Y P 1 A expression in tissue biopsies that were taken during the B N F study was significantly higher than in biopsies taken three weeks earlier. Liver biopsies were obtained at both time points for  95  two of three control seals, and skin biopsies were obtained for all three seals at both time points. 3.1.3.1 CYP1A protein levels C Y P 1 A protein levels in liver were approximately seven times higher after three weeks in captivity (p=0.04), and there was a six-fold increase in C Y P 1 A protein levels in skin from one seal (Figures 3.18, 3.19). However, due to limited sample volumes and an air bubble on one of the immunoblots, I did not obtain skin data from the second set of biopsies for the other two seals.  CYP1A1?  1  2  3  4  CYP1A2?  5  6  Figure 3.18: Immunoblot of harbour seal pup liver S9 fractions probed with rabbit antirat CYP1A2 polyclonal serum (1:500 dilution) from biopsies obtained 2 days postcapture and after 3 weeks in captivity. Lane 1 is a purified CYP1 A l rat standard, applied at 0.1 pmoles per lane. Lane 2 is a 3-MC-treated rat liver microsomal sample, loaded at 0.07 ug per lane. Lanes 3 and 4 are liver S9 fractions from PV007 (biopsies taken 2 days and 3 weeks post-capture, respectively). The S9 fractions were applied at 20 ug and 10 ug total protein per lane, respectively. Lanes 5 and 6 are liver S9 fractions from PV0016 (biopsies taken 2 days and 3 weeks post-capture, respectively). The S9 fractions were loaded at 20 ug and 10 ug total protein per lane. The immunoblot was developed with substrate solution for 3.5 minutes.  96  200 Liver — • —  •o-•  o  a  —-T—  •  PV001 PV007 PV0016  P  ••>  140  u  'S o  u a. <  H  u  •S  a.  .••/ //  120  .-'/ ••'/  100  //  •/ .7 .7 •/  80 60 /  40  /  /  f  /  4  20  1st biopsy  2nd biopsy  B 40 Skin w •S  o  35  a. OX  E  30  CM  25  a u  • PV001 O • PV007 - - T PV0016  O 20 A a. CM  u 3  10 ^  /  O c  « 3  8/  5  4  u 1st biopsy  2nd biopsy  Figure 3.19: C Y P 1 A protein levels increased in liver (A) and skin (B) of captive harbour seals during three weeks in captivity.  97  3.1.3.2 EROD activity Hepatic EROD activity was approximately ten times higher in control seals after three weeks in captivity (p=0.005). In skin, there was a four to five-fold increase in EROD activity (p=0.001) (Figure 3.20).  B 35  2.5 •  Liver 30 25  T  / /P / / / .' / ••' / ••' / .•• / / // / / / •' / ••'  —•—  PV001 • - o - PV007 — • - PV00I6  D.  >  15  o  10  W  /  /  I: i • i:  Skin  i E  a.  a  o «  / •• V/  2 c  d  2.0  • PV001 • O •• PV007 - • T - PV0016  1.5  1.0 A  0.5  3  u 0.0  1st biopsy  2nd biopsy  1st biopsy  2nd biopsy  Figure 3.20: EROD activity increased in liver (A) and skin(B) of captive harbour seals during three weeks in captivity.  98  3.1.4  Antibody inhibition of EROD activity To determine whether EROD activity represented a specific measure of C Y P 1 A  enzyme activity in harbour seal liver and skin, antibody inhibition experiments were conducted.  In liver, anti-rat CYP1A1 IgG inhibited EROD activity by approximately  80% at the maximum antibody concentration.  Inhibition occurred in pooled samples  from BNF-treated and control seals (Figure 3.21 A,B). Rabbit anti-rat C Y P 1 A 2 serum inhibited EROD activity in a liver sample from a BNF-treated seal (Figure 3.2IC). Rabbit anti-rat CYP1A2 serum inhibited EROD activity in a liver sample obtained 2 days post-capture from a captive seal pup (Figure 3.2ID). In skin, anti-rat CYP1A1 IgG may have inhibited EROD activity slightly in B N F control seals (Figure 3.2IE).  Anti-  CYP1A2 serum did not appear to inhibit EROD activity in skin biopsies taken immediately post-capture (Figure 3.21F).  99  Figure 3.21: Antibody inhibition of EROD activity in harbour seal liver and skin. EROD activity was calculated at 20 minutes reaction time for all except (F). Serum IgG concentrations were estimated, based on information in the literature. (A) A pooled liver S9 fraction from BNF-treated seals (n-2) was incubated with anti-rat CYP1A1 IgG and loaded at 0.01 mg total protein/well. EROD activity decreased from approximately 64 pmol/mg/min (0 IgG) to 11 pmol/mg/min (0.02 mg IgG/well). (B) A pooled S9 fraction from BNF study control seals («=3) was incubated with anti-rat CYP1A1 IgG and loaded at 0.01 mg total protein/well. EROD activity decreased from approximately 42 pmol/mg/min (0 IgG) to 5 pmol/mg/min (0.02 mg IgG/well). (C) A liver S9 fraction from one BNF-treated seal was incubated with anti-rat CYP1A2 serum and loaded at 0.01 mg total protein/well. EROD activity decreased from approximately 74 pmol/mg/min (0 IgG) to 28 pmol/mg/min (0.02 mg IgG/well). (D) A liver S9 fraction from one seal was incubated with anti-rat CYP1A2 serum and loaded at 0.1 mg total protein/well. EROD activity decreased from approximately 5 pmol/mg/min (0 IgG) to 0.9 pmol/mg/min (0.05 mg IgG/well). (E) A pooled skin S9 fraction from BNF study control seals (n=3) was incubated with anti-rat CYP1A1 IgG and loaded at 0.075 mg total protein/well. Only one IgG concentration was used due to limited sample volume. EROD activity decreased slightly from approximately 2 pmoles/mg/min (0 IgG) to 1.7 pmoles/mg/min (0.15 mg IgG/well). However, EROD activity of S9 fraction incubated with control serum also decreased to 1.8 pmol/mg/min (0.15 mg control IgG/well). (F) A pooled skin S9 fraction (rc=8) was incubated with anti-rat C Y P 1 A 2 serum and loaded at 0.12 mg total protein/well. At 20 minutes reaction time, E R O D activity was < L O Q for all concentrations. Results shown are at a reaction time of 40 minutes.  EROD activity (% of control activity)*  001  EROD activity (% of control activity)*  E R O D activity (% of control activity)*  101  102  3.2 3.2.1  Study 2: Field Study of Free-Ranging Harbour Seals Hornby Island harbour seal pups  3.2.1.1 CYP IA expression in skin Both C Y P 1 A protein concentration and EROD activity were quantified in skin biopsies from Hornby Island seal pups.  Two faint C Y P 1 A bands having the same  mobility as C Y P 1 A protein bands in BNF-treated liver were detected on immunoblots (Figure 3.22). The lower C Y P 1 A band was not detectable in all samples. EROD activity was below the limit of quantitation for six of ten samples. Thus, the sample size was too small to determine i f EROD activity and C Y P 1 A protein concentration were correlated in these samples (refer to Table 3.4). Table 3.4: Cutaneous C Y P 1 A expression in Hornby Island harbour seal pups Sample  Average C Y P 1 A protein levels  EROD activity  (RCQ/mg protein)  (pmol/mg/min)  3  PV01-01  13.4  0.64  PV01-02  12.7  <LOQ  PV01-03  14.0  0.91  PV01-04  16.0  <LOQ  PV01-05  15.1  <LOQ  PV01-06  14.0  <LOQ  PV01-07  8.6  <LOQ  PV01-08  13.1  0.66  PV01-09  9.5  0.75  PV01-10  10.1  <LOQ  LOQ (EROD) = 0.54 pmoles/mg/min LOQ (immunoblots) = 0.3/mg total protein loaded RCQ (relative contour quantity) is the staining intensity of sample bands normalized to a 0.01 pmole/lane rat CYP1A1 standard. EROD activity was calculated at 20 mins reaction time for all samples. 3  103  CYP1A1?  1  2  CYP1A2?  3  4  5  6  7  8  Figure 3.22: Immunoblot of Hornby Island harbour seal pup skin S9 fractions probed with rabbit anti-rat CYP1A2 polyclonal serum (1:500 dilution). Lane 1 is purified CYP1A1 rat standard, applied at 0.01 pmoles per lane. Lane 2 is a 3-MC rat liver microsomal sample, loaded at 0.07 pg per lane. Lanes 3-5 are skin S9 fractions from seals PV01-05, 06, and 07, loaded at 42 pg, 60 pg, and 60 pg, total protein per lane, respectively. Lanes 6 and 7 are skin S9 fractions from Puget Sound seal pups PV01-54 and 55, loaded at 26 pg and 45 pg, respectively. Lane 8 is a liver S9 fraction from B N F treated seal PV0012, loaded at 2.5 pg total protein per lane. This immunoblot was developed with substrate solution for 5 minutes.  104  3.2.1.2 Effect of body weight on CYP IA expression in skin Body weight was not correlated with C Y P 1 A protein levels (p=0.834,  n-\0)  (significant outlier removed) or EROD activity (^=0.851, w=4) for Hornby Island pups (data not shown). 3.2.1.3 Effect of sex on CYP1A expression in skin C Y P 1 A protein levels (/?=0.207) and EROD activity (only one male above LOQ) were not significantly different between Hornby Island males («=6) and females («=4) (data not shown). 3.2.1.4 Effect of contaminant levels in blubber on CYP1A expression E R O D activity and C Y P 1 A protein levels (Figure 3.23) were not significantly correlated with total TEQ for blubber.  There were no correlations between C Y P 1 A  protein levels and contaminants when mono-ortho PCBs were removed from the total TEQ 0=0.227) (data not shown). 3.2.1.5 Relationship between body weight and contaminant levels in blubber Contaminant levels in blubber, expressed as total TEQ, were positively correlated with body weight (r =0.48, p=0.026) (Figure 3.24). A single data point (39.0 kg) was 2  largely responsible for the positive correlation, however.  This data point was a  significant outlier (p<0.05), and no correlation was observed when it was removed (p=0.536).  105 18  H o r n b y Island pups  S k i n CYP1A vs. T E Q 16  14  A  12  10  H  12  p=0.146  —-—i 14  1  1  16  18  1—  22  20  24  26  —i 28  1  30  32  T o t a l T E Q (ng/kg lipid)  Figure 3.23: Relationship between blubber total TEQ and cutaneous C Y P 1 A protein levels in Hornby Island harbour seal pups.  —i—  16  20  24  28  32  36  40  Body weight (kg)  Figure 3.24: Correlation between body weight and total TEQ for blubber in Hornby Island harbour seal pups.  106  3.2.2  Puget Sound harbour seal pups  3.2.2.1 CYP IA expression in skin Both C Y P 1 A protein concentration and EROD activity were quantified in skin biopsies from Puget Sound seal pups. Two C Y P 1 A protein bands with approximately the same mobility as the rat CYP1A1 standard were detected in skin (Figure 3.25). E R O D activity was below the limit of quantitation for most samples. Thus, the sample size was too small to determine i f EROD activity and C Y P 1 A protein concentration were correlated in these samples (refer to Table 3.5).  CYP1A1?  1  2  CYP1A2?  3  4  5  6  7  8  Figure 3.25: Immunoblot of Puget Sound harbour seal pup skin S9 fractions probed with rabbit anti-rat CYP1A2 polyclonal serum (1:500 dilution). Lanes 1 and 8 are purified CYP1A1 rat standards, applied at 0.01 pmoles and 0.05 pmoles per lane, respectively. Lanes 2-7 are skin S9 fractions from seals PV01-46, 47, 48, 49, 51, and 53, loaded at 31 pg, 37 pg, 60 pg, 60 pg, 37 pg, and 22 pg, respectively. The immunoblot was developed with substrate solution for 5 minutes.  107  Table 3 . 5 : Cutaneous C Y P 1 A expression in Puget Sound harbour seal pups Sample  Average C Y P 1 A protein levels  EROD activity  (RCQ/mg protein)  (pmoles/mg/min)  3  -  PV01-40  16.6  <LOQ  PV01-41  14.7  0.65  PV01-42  17.0  <LOQ  PV01-43  11.2  <LOQ  PV01-44  15.2  <LOD  PV01-45  19.6  <LOQ  PV01-46  21.8  <LOQ  PV01-47  23.1  <LOQ  PV01-48  5.9  0.79  PV01-49  7.8  1.10  PV01-51  22.2  <LOQ  PV01-52  11.6  <LOQ  PV01-53  <LOQ  <LOQ  PV01-54  25.0  <LOD  PV01-55  21.8  <LOD  PV01-56  21.3  <LOD  b  L O Q (EROD) = 0.65 pmol/mg/min LOQ (immunoblots) = 0.3/mg total protein loaded R C Q (relative contour quantity) is the staining intensity of sample bands normalized to a 0.01 pmo le/lane rat C Y P 1A1 standard. E R O D activity was calculated at 20 mins reaction time for all samples.  a  108  3.2.2.2 Effect of body weight on CYP IA expression in skin C Y P 1 A protein levels were not correlated with body weight for Puget Sound pups (p=0.832, «=14) (data not shown). Correlation analysis was not performed for EROD activity, as only three values were above the LOQ. 3.2.2.3 Effect of sex on CYP IA expression in skin C Y P 1 A protein levels (p=0.744) and EROD activity were not significantly different between males («=8) and females (n=7) (data not shown). 3.2.3  Adult harbour seals from British Columbia  3.2.3. J  CYP1A expression in skin  Both C Y P 1 A protein levels and EROD activity were quantified in skin biopsies from adult seals. Two C Y P 1 A protein bands having the same mobility as C Y P 1 A in BNF-treated liver were detected on immunoblots (Figure 3.26).  EROD activity was  below the limit of quantitation for 12 of 16 samples (see Table 3.6). 3.2.3.2 Effect of body weight on CYP1A expression in skin C Y P 1 A protein levels (p=0.296, «=14) and EROD activity (p=0.431, «=4).'were not correlated with body weight for adult harbour seals (data not shown). 3.2.3.3 Effect of sex on CYP IA expression in skin C Y P 1 A protein levels were not significantly different between males in=T) and females («=9) (p=0.439) (data not shown). EROD activity was below the L O Q for all males; therefore, a t-test was not performed.  109  Table 3 . 6 : Cutaneous C Y P 1 A expression in adult harbour seals Average C Y P 1 A protein levels  EROD activity  (RCQ/mg protein)"  (pmol/mg/min)  PV01-17  8.2  0.69  PV01-18  17.8  <LOQ  PV01-20  <LOQ  <LOQ  PV01-22  13.7  0.84  PV01-23  <LOQ '  0.81  Sample  PV01-25  9.1  <LOQ  PV01-29  6.6  <LOQ  PV01-30  10.2  <LOQ  PV01-31  13.6  <LOQ  PV01-32  11.6  <LOQ  PV01-33  18.4  <LOQ  PV01-34  48.6  <LOQ  PV01-36  24.7  <LOQ  PV01-37  12.4  <LOQ  PV01-38  12.8  <LOQ  PV01-39  5.6  1.14  L O Q (EROD) = 0.65 pmol/mg/min LOQ (immunoblots) = 0.3/mg total protein loaded R C Q (relative contour quantity) is the staining intensity of sample bands normalized to a 0.01 pmole/lane rat CYP1A1 standard. E R O D activity was calculated at a reaction time of 20 minutes for all samples. a  110  CYP1A1?  1  2  CYP1A2?  3  4  5  6  7  8  Figure 3.26: Immunoblot of adult harbour seal skin S9 fractions probed with rabbit antirat CYP1A2 polyclonal serum (1:500 dilution). Lanes 1 and 8 are purified CYP1A1 rat standards, applied at 0.01 pmoles and 0.05 pmoles per lane, respectively. Lane 2 is a liver S9 fraction from BNF-treated PV0012, loaded at 2.5 pg per lane. Lanes 3-5 are skin S9 fractions from seals PV01-32, 34, and 37, loaded at 60 pg, 55 pg, and 60 pg, total protein per lane, respectively. Lanes 6 and 7 are skin S9 fractions from Puget Sound seal pups (PV01-54 and 55), loaded at 26 pg and 45 pg total protein per lane, respectively. This immunoblot was developed with substrate solution for 4 minutes 45 seconds.  Ill  3.3  Harbour Seal Inter-Population Comparisons  3.3.1  Relationship between CYP1A protein levels and EROD activity There was no correlation between C Y P 1 A protein levels and EROD activity in skin  when data from all harbour seal pup populations were combined (Fraser River, Hornby Island, Puget Sound) (p=0.122) (Figure 3.27).  Skin  .5 1-2 E  •  "3D  E  1  o  1.0  >. •t? 0.8 Q O en  • •  0.4  •  3  0.2 *  •  • •  0.6  w  u  •  •  E c  •  • •  •  ••  p=0.122  •  •  •  i  1  I  6  8  10  12  14  16  1  Cutaneous CYP1A protein (RCQ/mg protein)  Figure 3.27: Relationship between C Y P 1 A protein levels and EROD activity in skin of harbour seal pups from all populations (Fraser River, Hornby Island, Puget Sound) 3.3.2  Effect of body weight on CYP1A  expression  Body weight was not correlated with C Y P 1 A protein levels (p=0.821, «=43) or EROD activity (p=0.588, n=\6) in skin when data from all seal pup populations were combined (data not shown). 3.3.3  Variability of cutaneous CYP1A  expression between pups and adults  There was no difference in cutaneous C Y P 1 A protein levels (p=0.529) or EROD activity (/?=0.114) between harbour seal adults and pups (data not shown).  112  3.3.4  Effect of sex on CYP1A expression There was no difference in C Y P 1 A expression in skin between male and female  harbour seal pups when all populations were combined (EROD p=0.398; C Y P 1 A p=0.346) (data not shown). 3.3.5  Comparison of CYP1A expression in pups from BC and Puget Sound C Y P 1 A protein levels were significantly higher in skin from Puget Sound harbour  seal pups than in pups from British Columbia (combined data of Fraser River pups and Hornby Island pups) (p=0.003) (Figure 3.28). In each group, one sample was below the LOQ.  The LOQ value was substituted for these two samples. However, if these two  values were not included in the analysis, the result was still significant. When the B C populations were considered individually, C Y P 1 A protein levels were significantly higher for Puget Sound pups than for Fraser River pups (p=0.03), but were not significantly higher than for Hornby Island pups (p=0.09). Only two Puget Sound skin samples had EROD activity above the LOQ; therefore, a statistical comparison was not performed. 3.3.6  Effect of contaminant levels in blubber on CYP1A expression Contaminant data was only available for Fraser River and Hornby Island pups.  When data from these two groups were combined, there was no correlation between contaminants in blubber and cutaneous CYP1A protein levels (p=0.643) (data not shown).  113  25  n  0 J  1  ,  BC  '  ,  —  Puget Sound  */?=0.003 Error bars=SD  Figure 3.28: Comparison of cutaneous C Y P 1 A protein levels between harbour seal pups from British Columbia (BC) and Puget Sound, Washington  114  3.4 3.4.1  Study 3: Study of Free-Ranging Killer Whales CYP1A expression in killer whale skin E R O D activity was not detected in killer whale skin, but a possible C Y P 1 A protein  band was detected on immunoblots for a few killer whale skin samples. A protein band having the same mobility as the most darkly stained protein band in beluga whale liver microsomes can be seen in lane 2 (Figure 3.29). However, the band was very faint and absent in most samples.  1  2  3  4  5  6  7  8  Figure 3.29: Immunoblot of killer whale skin S9 fractions probed with rabbit anti-rat C Y P 1 A 2 polyclonal serum (1:500 dilution). Lane 1 is a liver microsomal sample from beluga whale, loaded at 2 pg (estimated) total protein per lane. Lanes 2-4 are killer whale skin S9 fractions from southern resident killer whale A73 (60 pg), and transient killer whales KW00-11 (60 pg) and KW00-03 ( 54 pg), respectively. Lanes 5 and 6 are skin S9 fractions from Fraser River harbour seal pups PV0019 and 13, loaded at 55 pg and 60 pg, respectively. Lane 7 is a skin S9 fraction from BNF-treated seal PV0015, loaded at 10 pg total protein per lane. Lane 8 is a purified CYP1A1 rat standard, applied at 0.01 pmoles per lane. This immunoblot was developed with substrate solution for 1 minute. A short reaction time was used in order to minimize non-specific bands and to visualize the most immunoreactive protein bands.  115  4.  DISCUSSION  The overall objective of the present study was to obtain liver and skin biopsies from marine mammals using minimally-invasive means and to validate the use of C Y P 1 A as a biomarker of contaminant exposure in these biopsy samples. This study consisted of three groups: (1) 20 free-ranging harbour seal pups that were live-captured and then temporarily held in captivity; (2) 42 free-ranging harbour seals, including both pups and adults, sampled in British Columbia and Washington; and (3) 14 free-ranging killer whales sampled in coastal waters of British Columbia.  Each of these groups  offered particular advantages in the development of biomarker methods for marine mammal populations. The captive seal study allowed for biopsy of both liver and skin using general anaesthetic and provided a relatively controlled environment in which to investigate the inducibility of CYP1A in these tissues. This provided a foundation for conducting and validating field studies in which less invasive sampling was conducted (i.e. only skin-blubber biopsies). Field sampling of harbour seals provided skin biopsy samples from individuals of different age categories inhabiting both industrialized (contaminated) and more remote (relatively uncontaminated) areas. This provided an opportunity to investigate whether natural factors (eg. age) or anthropogenic factors (eg. differences in contaminant exposure) affected C Y P 1 A expression in skin. Samples from both the captive and free-ranging harbour seal groups were used to develop and validate laboratory methods, providing an important basis for the study of killer whales, which could not be captured or sampled under controlled conditions.  The availability of  individual-based information for each of the killer whales sampled provided a unique  116  opportunity to investigate cutaneous C Y P 1 A expression in the context of age, sex, diet, and contaminant levels. 4.1 4.1.1  Study 1: Captive Harbour Seal Study Analysis of liver samples obtained two days post-capture  4.1.1.1 Liver biopsy approach Our approach for measuring hepatic C Y P 1 A in harbour seals was novel because we used small tissue biopsies from healthy animals that were live-captured and temporarily held in captivity. Whereas previously published studies of C Y P 1 A activity in seals or other marine mammals used liver from dead animals (stranded or sacrificed), we demonstrated that an amount of tissue sufficient for C Y P 1 A analysis could be obtained by liver biopsy under general anaesthetic. Seals appeared to recover rapidly and fully from the biopsy procedure and remained healthy while in captivity. 4.1.1.2 EROD activity The present study is the first to report hepatic C Y P 1 A data for marine mammals on the west coast of Canada and the first using a liver biopsy approach. EROD activity measured in S9 fractions prepared from liver biopsies obtained two days post-capture ranged from < 0.65 to 4.4 pmoles product formed/min/mg protein. Comparison of data from the present study with previously published studies shows differences in EROD activity values between seal species, as well as between age classes.  Addison et al.  (1986) reported a mean EROD activity of 6.3 ± 4.6 nmoles product formed/min/mg protein for newborn harbour seal pups of variable ages (including two fetuses) from the east coast of Canada, 1000 times higher than activities measured in the present study. EROD activities reported for other seal species (Wolkers et al, 1998; Hyyti et al., 2001;  117  Chiba et al., 2002; Ruus et al., 2002) were also several orders of magnitude higher than those obtained in this study. Wolkers et al. (2002), however, reported activities that are similar to those measured in the present study, with mean EROD activity of 9.6 ± 9.6 pmol/min/mg protein for hepatic microsomes from unweaned harp seal pups, and 11.5 ± 5.4 pmol/min/mg for harp seal mothers. In the present study, C Y P 1 A was measured in the S9 fraction rather than the microsomal fraction, representing one possible reason for the relatively low hepatic E R O D activity observed. The S9 fraction contains many other cytosolic proteins that contribute to the total protein determination (Munkittrick et al., 1993), thereby decreasing EROD activity values when expressed per mg total protein. Munkittrick et al. (1993) reported that EROD activity was 3.5 times higher in hepatic microsomal fractions compared to S9 fractions from white sucker (fish). Direct comparisons of EROD activity among studies is problematic, however, if different protocols, equipment, and reagents are used.  Different laboratories have  reported different absolute activities when analyzing the same samples (Munkittrick et al, 1993; Stagg and Addison, 1995). In one study, for example, mean EROD activities determined by eleven different laboratories for the same contaminant-exposed fish ranged from 20 to 298 pmol/min/mg protein (Munkittrick et al, 1993). Different methods of measuring total protein content of samples also produced variable results.  Individual  samples assayed using the same method had up to 33% less or 63% greater total protein content than corresponding samples assayed with another method (Munkittrick et al., 1993), and this subsequently affected EROD activities.  118  4.1.1.3 Immunoblot analysis In the present study, two immunoreactive C Y P 1 A protein bands, tentatively identified as C Y P l A l - a n d CYP1A2, were detected in S9 fractions prepared from liver. This result was expected, as CYP1A1 and CYP1A2 gene fragments have recently been cloned in four different seal species (Teramitsu et al., 2000; Tilley et al., 2002). Tilley et al. (2002) assigned molecular weights to CYP1A1 and CYP1A2 in grey and harp seals based on deduced amino acid sequences. Molecular weights were too close together to be able to definitively distinguish between CYP1 A l and CYP1A2 (58 K D a and 57 KDa, respectively), although the greater molecular weight of CYP1A1 suggests that the upper protein band is CYP1A1 and the lower band is CYP1A2 for these two seal species. Given that the two C Y P 1 A enzymes in grey, harp, ribbon, and largha seals were >90% identical (Teramitsu et al., 2000; Tilley et al., 2002), it is likely that upper and lower protein bands correspond to the same C Y P 1 A enzymes in harbour seals. Moreover, in the present study, levels of putative CYP1A1 protein corresponding to the upper band correlated with E R O D activity in liver, but the lower protein band did not correlate with EROD activity. This result suggests that the upper protein band is CYP1A1 in harbour seals, as CYP1A1 has been shown to be the primary catalyst of the EROD reaction in other species (Burke et al., 1994). Additionally, in liver and skin samples from B N F treated seals (discussed in section 4.14), the upper C Y P 1 A band was induced to a greater extent than the lower band, providing further evidence that the upper (higher molecular weight) protein band is CYP1A1. In rats, B N F is known to induce CYP1A1 more than CYP1A2 (Thomas et al., 1983).  119  4.1.2  Analysis of skin samples obtained two days post-capture The present study is the first to measure C Y P 1 A expression in skin preparations  from harbour seals and is the first to report the presence of two C Y P 1 A proteins in skin of a marine mammal. These results provide an important foundation for the future development of less-invasive approaches to biomarker research.  The immunoblot  analysis clearly shows that there are two C Y P 1 A enzymes in skin. EROD activity was also detected in cutaneous S9 fractions. However, EROD activity and C Y P 1 A protein levels were close to their respective LOQs in all skin samples, and E R O D activity, in particular, was often below the LOQ. C Y P 1 A protein values had high inter-assay variability because bands on immunoblots were often faint and were, therefore, more difficult to quantify. 4.1.3  Comparison of CYP1A expression levels between liver and skin In the present study, EROD activity was four to five times higher in liver than  skin of harbour seal pups, reflecting the general tissue distribution of C Y P enzymes and r  the central role that the liver plays in enzymatic detoxification processes. Fossi et al. (1997) reported that BaP hydroxylase activity was five times higher in microsomal fractions prepared from liver than in homogenates prepared from skin of a sea lion, although sample size (n=l) clearly limits interpretation of this result. It is not clear i f the skin homogenates analyzed by Fossi et al. (1997) were S9 fractions. As microsomes are more concentrated with respect to C Y P 1 A protein than pre-microsomal homogenates, the five-fold difference in EROD activity between liver and skin reported by Fossi et al. (1997) was likely due in part to the analysis of different fractions. To date, the report by Fossi et al. (1997) is the only published comparison of hepatic and cutaneous C Y P 1 A -  120  mediated enzyme activity in a marine mammal species. In our study, C Y P 1 A protein levels were estimated to be sixteen times higher in liver than skin, but direct comparison was not possible because different internal standards were used to calculate relative staining intensities in the two tissues. Total C Y P content was not measured in liver or skin S9 fractions due to limited sample volumes, but in other species it is known that total C Y P concentration is much less in skin than liver. Moreover, C Y P 1 A makes up a small fraction of the total C Y P content in uninduced animals. In liver microsomes from untreated male rats, C Y P 1 A represents 1-3% of the total C Y P content (Bandiera, 2001). In rats, the total C Y P content of liver was reported to be at least 4.5 times greater than that of skin, and in mice, total C Y P content was at least seventeen times greater in liver than skin (Lewis, 1996). Pham et al. (1989) similarly found that total C Y P content was approximately seventeen times greater in microsomal fractions of liver than skin from rats. In a T C D D dosing study in mice, Diliberto et al. (2001) reported a 40-fold difference between EROD activity in liver and skin of control animals. Bickers et al. (1974) found that basal BaP hydroxylase activity was approximately 100 times greater in rat liver than skin, and Mukhtar and Bickers (1981) reported that skin homogenates had 7-28% of the activity of liver in BaPand Arocolor 1254-treated rats, depending on the treatment group. In rats, CYP1A1 was not detected in untreated skin and was only faintly detected in skin from 3-MC-treated rats by immunoblot analysis (Pham et al, 1989). Although C Y P 1 A expression levels in skin and liver were not positively correlated for captive harbour seal pups, the apparent molecular weights of the two immunoreactive C Y P 1 A protein bands for skin corresponded to those for liver on  121  immunoblots. Additionally, C Y P 1 A induction was observed in both skin and liver in the B N F induction study (section 4.14).  If sample size was too small, or i f C Y P 1 A  expression in skin was too low to detect differences among individuals, no correlation between liver and skin would be expected. However, a significant negative correlation between hepatic and cutaneous EROD activity was observed.  A negative correlation  between C Y P 1 A in liver and skin suggests that C Y P 1 A expression may be either tissuedependent (eg. due to toxicokinetic or developmental differences between tissues), or that other proteins in cutaneous S9 fractions interfere with C Y P 1 A activity in skin.  It is  possible that seals with higher hepatic C Y P 1 A levels also had elevated levels of other cytosolic proteins in skin. These proteins may have 'diluted' the already low C Y P 1 A levels in skin homogenates. To determine i f C Y P 1 A catalyzed the EROD reaction in both liver and skin, in vitro antibody inhibition experiments were conducted using rabbit anti-rat CYP1A1 monoclonal IgG and rabbit anti-rat CYP1A2 polyclonal, serum. Hepatic EROD activity was inhibited by up to 85%, indicating that C Y P 1 A is responsible for catalyzing the EROD reaction in harbour seal liver samples. EROD activity was not inhibited in skin samples, however, suggesting that different enzymes may be catalyzing EROD in skin compared to liver. C Y P forms other than C Y P 1 A have been shown to contribute to EROD activity in untreated rats (Burke et al., 1994). Alternatively, EROD activity in skin may have been too low (i.e. too close to the detection limit) to detect a decrease in EROD activity with the addition of antibody. Control serum IgG concentrations also may have been too high, masking actual differences between control and antibody-treated skin samples. The serum IgG concentration was not known and was, therefore, estimated  122  based on literature values. The serum IgG concentration may have been underestimated, resulting in greater IgG concentrations in reaction wells than expected. In fact, EROD activity in liver was significantly reduced by high concentrations of control serum (estimated concentration of 0.025 mg IgG/well) (Figure 3.21D). Antibody inhibition experiments indicated that C Y P 1 A catalyzed the EROD reaction in liver, but results in skin were inconclusive. Studies in rodents have provided evidence to suggest that hepatic and epidermal CYP1A1 enzyme structure and inducibility are identical. Raza et al. (1992) found that in rats, hepatic and epidermal CYP1A1 had common substrate specificity and similar responsiveness  to  inducers  (both were induced by B N F treatment),  common  immunocross-reactivity ( A H H activity was inhibited by monoclonal and polyclonal antiCYP1A1 antibodies in both tissues, and to a similar degree), and the same H P L C elution pattern, molecular mass (SDS-PAGE), monoclonal antibody epitopes on fingerprint analysis, N-terminal amino acid sequences, and tryptic peptides (Raza et al., 1992). Further evidence to support the identity of C Y P 1 A genes in liver and skin has been provided by Shimizu et al. (2000). Bt?P treatment induced CYP1A1 expression in both liver and skin of mice, but did not induce CYP1A1 in either tissue of AhR knockout mice. The results of rodent studies and the analagous response of C Y P 1 A in harbour seal liver and skin to oral B N F treatment provide support for the idea that C Y P 1 A enzymes in harbour seal skin are the same as those in liver. However, in the present study, relatively small sample sizes and low activities and protein levels in skin limited interpretation of the results.  123  4.1.4  Induction of CYP1A by in vivo BNF treatment To my knowledge, the present study is the first to show that in vivo B N F  treatment induces C Y P 1 A protein levels in any tissue from a marine mammal species. This finding provides evidence that an inducible C Y P 1 A protein is present in harbour seals and supports the development of C Y P 1 A as a biomarker of environmental contaminant exposure in these animals. C Y P 1 A protein levels were induced in both liver and skin, indicating that the same mechanistic response to contaminant exposure occurs in these tissues. Unexpectedly, EROD activity was not induced in either liver or skin in B N F treated seals. These results were confirmed for liver samples by repeating the assay, as well as by measuring EROD activity of a treated and a control sample using a different EROD assay protocol. Chemical inhibition by B N F (in vivo) is one possible explanation for the apparent lack of induction of EROD activity in BNF-treated harbour seals, despite increases in C Y P 1 A protein levels.  This observation is consistent with reports that  EROD activity is not always induced concurrent with the induction of C Y P 1 A protein or mRNA levels and has been attributed to competitive inhibition by the inducing agent (Gooch et al, 1989; Haasch et al, 1993; Goksoyr and Husoy, 1998; Petrulis and Bunce, 1999). Haasch et al. (1993) found that in vitro B N F treatment inhibited EROD activity in liver microsomes from rainbow trout and suggested that inhibition of EROD activity after in vivo B N F treatment may be due to the presence of B N F in the microsomal preparations.  However, the amount of B N F , if any, retained in the microsomal  preparation was not known. In the present study, liver biopsies were taken 24 hours after  124  the last B N F treatment, and this may have been sufficient time for BNF to be eliminated from the liver. Haasch et al. (1993) noted that measurement of B N F tissue levels was difficult because the bioconcentration factor of BNF was not known, and radiolabeled B N F was not commercially available. Gooch et al. (1989) similarly suggested that a lack of E R O D induction in fish given high doses of 3,3',4,4'-tetrachlorobiphenyl (TCB) could result from inhibition by T C B retained in microsomes during subcellular fractionation. Petrulis and Bunce (1999) suggested that competitive inhibition would not be a factor when assessing environmental exposure to dioxin-like compounds, however, because these compounds would not be carried over from the intact animal to the microsomal fraction. In the present study, a subsequent inhibition of EROD activity by in vitro B N F treatment supported the possiblity that in vivo BNF treatment might result in inhibition of hepatic microsomal EROD activity. Another suggested but less likely mechanism by which EROD activity can be inhibited is oxidative inactivation of the enzyme. Some contaminants, such as 3,3',4,4'tetrachlorobiphenyl (TCB) have been shown to stimulate production of reactive oxygen species (ROS) by liver microsomes, which can inactivate C Y P 1 A by attacking its active site (Schlezinger et al, 1999; Schlezinger and Stegeman, 2001). 4.1.5  Relationship between environmental contaminant levels and CYP1A expression I hypothesized that C Y P 1 A expression in liver and skin would be positively  correlated with blubber organochlorine levels, expressed as total TEQ. Although C Y P 1 A expression did not correlate with total TEQ (representing the sum of non-ortho and mono-ortho PCBs, PCDDs, and PCDFs), a significant relationship was apparent when mono-ortho PCBs were not included in this analysis (significant correlation between  125  contaminants and hepatic, but not cutaneous, CYP1A). Mono-ortho substituted PCBs tend to be mixed-type inducers (i.e. induce both C Y P 1 A and CYP2B enzymes) and are weaker C Y P 1 A inducers than non-ortho substituted PCBs, 2,3,7,8-substituted PCDDs, and 2,3,7,8-PCDFs. Mammalian TEFs are largely based on rodent studies and may not accurately reflect the contribution of certain congeners to AhR-mediated endpoints in all species. Mono-ortho PCBs contributed approximately half of the total T E Q for captive harbour seal pups in the present study because their absolute concentration was the highest of the contaminant classes measured. However, it is possible that current TEFs overestimate the binding affinity of mono-ortho PCBs for the A h receptor in harbour seals, and therefore a correlation between TEQ and hepatic C Y P 1 A induction was only observed when mono-ortho congeners were excluded. Results of the present study suggest that mono-ortho PCBs are not inducing C Y P 1 A in harbour seal liver. High levels of mono-ortho PCBs in the blubber relative to PCDDs, PCDFs, and non-ortho PCBs may reflect the inability of harbour seals to preferentially metabolize mono-ortho congeners. Like rodents, marine mammals can more easily metabolize nonortho (coplanar) congeners than or/Tzo-substituted congeners, and the availability of adjacent unsubstituted carbon atoms at the ortho- and meta-positions, facilitates metabolism (Boon et al., 1992). Studies to date have shown that relationships between C Y P 1 A expression and OC levels in marine mammals are variable.  Although several studies reported positive  correlations between hepatic C Y P 1 A expression and contaminant levels in seals (Chiba et al., 2002) or differences in hepatic C Y P 1 A expression between seals from  126  contaminated and less contaminated areas (Mattson et al., 1998; Nyman et al., 2000) (previously discussed in section 1.2.1.2), other studies found no correlations between these variables (Chiba et al., 2002; Wolkers et al., 2002; Ruus et al., 2002). Chiba et al. (2002) found positive correlations between hepatic CYP1A (EROD activity and C Y P 1 A protein levels) and contaminant levels expressed as total TEQ in blubber of adult largha seals (9 males, 7 females), but did not find correlations in adult ribbon seals (6 males, 9 females). In another study, there was no correlation between hepatic EROD activity in harp seal mothers and pups and contaminant levels in blubber or milk (6 mother-pup pairs) (Wolkers et al., 2002). Ruus et al. (2002) did not find a correlation between hepatic EROD activity and levels of OCs in blubber (total PCB concentrations and total TEQ) in harbour seals (5 males, 5 females; 0-11 years). A n important consideration in evaluating results of these studies is the potential contribution of confounding factors (eg. age and sex of study animals). Contaminant burdens increase in males with age, whereas reproductive females off-load a large proportion of their contaminant burdens to their offspring during lactation. Many marine mammal studies do not control for these confounding factors (samples are from both male and female adults of variable ages), and spurious correlations between contaminant levels and C Y P 1 A expression may be reported. In many cases, contaminant exposure may not be high enough to elicit measurable C Y P 1 A induction responses, especially in extra-hepatic tissues, in which constitutive C Y P 1 A levels are relatively low. Additionally, sample sizes may often be too small to detect trends. In the present study, total PCB levels in blubber of harbour seal pups were comparable to levels measured in blubber of adult largha seals, in which correlations  127  between contaminants and hepatic EROD activity were reported (Chiba et al., 2002), but were relatively low compared to levels in harbour seals at other locations (see Table 1.1). Contaminant levels may not have been high enough in captive harbour seal pups to elicit a strong C Y P 1 A induction response in skin. It is also possible that a specific contaminant or group of contaminants not measured in the present study significantly contributed to the C Y P 1 A induction pattern in harbour seal skin. Planar contaminants not included in the TEF/TEQ scheme, such as PAHs and brominated hydrocarbons, can induce C Y P 1 A in animals (Brack et al, 2000; Van den Berg et al., 1998; Varanasi et al., 1992). This could partially explain the lack of correlation between C Y P 1 A expression in skin and TEQ. 4.1.6  Relationship between age and contaminant concentrations in blubber With the exception of lactating females, we expect age and organochlorine levels  to be positively correlated in adult seals (Ross and Troisi, 2001).  The relationship  between age and contaminant levels is strong for PCBs, but inconsistent for PCDDs and PCDFs as a result of relatively rapid metabolic elimination (Ross and Troisi, 2001). In young seals, however, the relationship between age and contaminants is less welldocumented. In the present study, body weight was negatively correlated with contaminant levels in blubber (total TEQ) for captive harbour seal pups. Addison and Stobo (1993) found that OC residue concentrations did not significantly change in grey seal pups during the nursing period. It is likely that an accompanying increase in the blubber layer effectively 'diluted' the contaminants that were also accumulating.  Once pups were  weaned, blubber weight decreased, and blubber residue concentrations  increased  128  significantly (Addison and Stobo, 1993).  Such a dilution effect may be partially  responsible for the negative correlation observed between body weight and total T E Q in captive harbour seal pups. 4.1.7  Relationship between body weight (age) and CYP1A expression in harbour seals Body weight can be used to approximate the age of nursing harbour seals pups  (Cottrell et al., 2002). Seal pups in the present study were likely at the end of their nursing period but were probably still nursing at the time of capture. It is not surprising that a correlation between body weight and CYP1A protein levels was not observed for captive harbour seal pups because ages were similar among individuals. Seals of the same approximate size were deliberately selected to try to eliminate age as a potential confounding factor (larger seal pups are typically older).  However, there was a  significant negative correlation between body weight and hepatic EROD activity («=10). This result contrasts with Wolkers et al. (2002), in which a positive correlation between body weight and hepatic EROD activity was found for nursing harp seal pups (n=6). This correlation was attributed to larger (older) pups having a more fully developed C Y P enzyme system. However, small sample sizes limit the interpretation of results in both Wolkers et al. (2002) and in the present study. Although correlations between body weight or age and contaminant levels are frequently reported for marine mammals (Addison and Stobo, 1993; Bernt et al., 1999), few studies reported correlations between body weight and CYP I A. During the nursing period, as body weight increases, developmental (age)-related changes in C Y P 1 A enzymes may take place in seal pups (Addison and Brodie, 1984).  However, such  129  developmental changes have not been well-studied in seals, and results from the present study highlight the importance of further research into some of the natural factors that affect the expression of C Y P 1 A enzymes. 4.1.8  Relationship between sex and CYP1A expression in harbour seals In some animal species, hepatic C Y P levels and C Y P enzyme activities are higher  in males than in females (Goksoyr, 1995). This topic has been well-studied in fish, in which sex differences are particularly pronounced during the reproductive period. Hepatic E R O D activity, for example, is lower in reproductively active females, and this suppression of C Y P 1 A is partially attributed to increased plasma levels of the steroid hormone, 17p-estradiol, in pre-spawning females (Whyte et al., 2000). Sex differences in total C Y P content and EROD activity have also been reported in rats (Pham et al., 1989). The evidence for sex differences in EROD activity in adult marine mammals is conflicting.  Goks0yr (1995) found that male hooded seals had higher EROD activity  than females, and Chiba et al. (2000) found that in ribbon seals, hepatic EROD activity was higher in males than females. However, no difference was detected between male and female largha seals (Chiba et al., 2000), or between male and female ringed seals (Mattson et al, 1998). In the present study, C Y P 1 A expression did not differ between male and female harbour seals. However, harbour seal pups were sexually immature.  130  4.1.9  Increase in CYP1A expression over time in captive harbour seals E R O D activity and C Y P 1 A protein levels increased significantly in both liver and  skin of harbour seals during their three-week stay in captivity. EROD activity was approximately ten times higher in liver samples (p=0.005) and four to five times higher in skin samples (p=0.001) from control seals than in samples taken immediately postcapture (three weeks prior to the BNF study). C Y P 1 A protein levels increased seven-fold in liver (p=0.04), and although there were insufficient data to conduct a t-test for skin, C Y P 1 A protein increased six-fold in one seal. 4.1.9.1  CYP1A development in young seals A plausible explanation for the increase in CYP1A expression observed in liver  and skin samples from control seals is that CYP1A levels and activities changed as seal pups aged. Past studies detected EROD activity in liver samples from newborn harbour seal pups, with increases in hepatic EROD activity occurring up to 12 days post-partum in grey seal pups (Addison and Brodie, 1984b; Addison et al, 1986). The developmental profile of C Y P enzymes in harbour seals is not known. In general, studies in rodents and pigs have shown that development of hepatic microsomal drug-metabolizing enzymes occurs during the first four to six weeks after birth, with the phase of most rapid development occurring between birth and the third to fifth week post-partum.  This  developmental pattern was observed regardless of species, gestation period, or the degree of maturation at birth (Dickerson and Basu, 1975). Previous studies of grey seals and harp seals showed that hepatic EROD activity was higher in adults than pups (Addison and Brodie, 1984b; Addison el al, 1986). In the present study, EROD activity in skin samples was similar in adults and pups, and after  131  three weeks in captivity, EROD activity in skin samples from untreated seals was higher than that of adults. Studies in rats and rabbits suggest that young animals can actually have higher C Y P expression than adults. In rats, total hepatic C Y P levels were low six days after birth, increased rapidly for three to five weeks after birth (three-fold increase when expressed per gram liver weight), and then decreased to adult levels at fifty to seventy days of age (approximately 2.5 times higher than levels at six days of age) (Dickerson and Basu, 1975). A similar age-related development pattern has been shown in rabbits (Fouts and Devereux, 1972). Alterman et al. (1994) reported that EROD activity in liver from rats peaked in very young animals, declined until about three months of age, and thereafter did not change further. These studies raise the possibility that harbour seals experience a similar increase in C Y P expression at about six weeks of age. The fact that harbour seals are known to be highly precocious, having good sight, coordination, and a highly developed immune system (Ross et al., 1994) and a short period of maternal care (Cottrell et al., 2002), may partially explain why C Y P 1 A expression in pups was comparable to that in adult seals. 4.1.9.2  Factors affecting variability of CYP1A expression within and between tissues Within tissues, one potential source of variation in CYP1A expression is protein  localization.  Localization of C Y P 1 A within the liver and skin could explain both  variations in expression over time when multiple biopsies are obtained from a single animal and potentially, detection of higher levels of CYP1A in an uninduced animal compared to an induced animal. In other words, one biopsy may have higher C Y P 1 A expression than another simply because of differences in C Y P 1 A expression within the same organ. In rat liver, C Y P 1 A expression and the induction response vary between  132  centrilobular, midzonal, and periportal hepatocytes (Baron et al., 1982; Wolf et al, 1984). In livers of untreated rats, centrilobular hepatocytes bound 25% more rabbit antirat C Y P 1 A antibody than midzonal or periportal hepatocytes (Baron et al., 1982), and the degree of induction after BNF treatment also varied between regions (Baron et al., 1982; Wolf et al., 1984). Because our liver biopsy technique involved a 'blind' sampling of liver, it is possible that we obtained variable C Y P 1 A yields among biopsies. However, the fact that similar C Y P 1 A induction responses to B N F treatment and concurrent increases in C Y P 1 A expression over time were observed in both liver and skin of captive seals suggests that these responses are real and likely do not represent an artifact of biopsy sampling. In the present study, skin biopsies taken from harbour seals two days post-capture were from the pelvic region of the body, whereas three weeks later, biopsies were taken from the hind flipper. It is possible that differences in constitutive C Y P 1 A expression between flipper and pelvic regions contributed to the apparent increase in cutaneous EROD activity and C Y P 1 A protein levels over time. In rats, A H H activity was eight-fold higher in epidermis from the ear and groin than from the abdomen, back, and tail (Don et al., 1987). However, there were no differences in C Y P 1 A expression between pelvic and flipper skin samples from an adult harbour seal.  Additionally, there was a parallel  increase in hepatic C Y P 1 A expression during the captive period. These results suggest that the observed increase in cutaneous C Y P 1 A expression over time reflects a real change in C Y P 1 A levels.  133  4.1.9.3  Possible effects of stress and captivity-relatedfactors on CYP1A expression C Y P 1 A expression in seal pups also may have been affected by stress associated  with their capture and time in captivity. Harbour seal pups were at approximate weaning age at the time of capture, meaning that they were still on a milk diet or that they had recently stopped receiving milk but were not yet eating fish.  In captivity, pups were  hand-fed herring, supplemented with vitamins and minerals. A change in diet from milk to whole fish may have elicited a stress or other physiological response. Contaminant levels in herring are unlikely to have had a direct effect on CYP1A, as herring has much lower levels of OCs than is predicted for fat-rich seal milk (Ross, pers. comm.). Compounds such as flavenoids and carotenoids consumed in the diet are known to affect C Y P 1 A expression in rats (Gradelet et al., 1996; Breinholt et al., 1999; Jewell and O'Brien, 1999). Carotenoids induced hepatic EROD activity forty to fifty-fold in rats in one study (Jewell and O'Brien, 1999) and 140-fold in another study (Gradelet et al.,' 1996).  Canthaxanthin (CX) and astaxanthin (AX), two of the carotenoids shown to  induce C Y P 1 A , are found in crustaceans and fish.  Although dietary intake of these  compounds is probably too low to cause CYP1A induction in humans (assuming a reponse in humans that is similar to rats) (Gradelet et al., 1996), C Y P 1 A levels may be affected in animals that consume large quantities of C X - and AX-containing foods. Some C Y P enzymes are suppressed and others are induced by immune and hormonal responses to stress (Morgan, 1997). Stress may be either psychological (eg. social stress caused by group housing) or physical (eg. infection or general anaesthesia) (Hinson and Raven, 1996). Seals in the captive study were subjected to many stressors, such as surgery, blood sampling, force-feeding, handling, and a change in environment.  134  Stress stimulates glucocorticoid secretion, and glucocorticoids are known to affect hepatic C Y P activities (Harvey, 1996).  In most species, Cortisol is the most  physiologically important glucocorticoid. In rainbow trout hepatocyte cultures, induction of  EROD activity by B N F was increased two to three-fold by Cortisol (Devaux et al.,  1992).  In rats, Prough et al. (1989) found that in vivo dexamethasone (a synthetic  glucocorticoid) treatment in adrenalectomized 3-MC-treated rats potentiated induction of CYP1A1 protein content and EROD activity. Down-regulation of C Y P 1 A expression in response to stress has also been reported. In Arctic charr, Jorgensen et al. (2001) found that fish subjected to daily handling and confinement stress had significantly higher plasma  Cortisol  levels than unstressed fish.  However, C Y P 1 A expression was not  significantly different between stressed and unstressed fish. In the same study, in BaPexposed fish, C Y P 1 A protein levels and EROD activity were lower in stressed fish compared to unstressed fish. A direct relationship between stress and C Y P 1 A expression has not been observed in marine mammals, but Engelhardt (1982) found that captivity and rigorous handling caused a four-fold increase in plasma Cortisol levels in juvenile ringed seals. Although harbour seals are known to be readily adaptable and acclimate quickly to captivity (Ross, pers. comm.), it is possible that stress affected gluocorticoid levels, which in turn may have increased C Y P 1 A expression in captive harbour seals.  135  4.2  Study 2: Field Study of Free-Ranging Harbour Seals  4.2.1  Inter-population  comparisons  4.2.1.1 Relationship between environmental contaminant levels in blubber and cutaneous CYP1A expression Although C Y P 1 A  expression i n s k i n d i d not correlate w i t h contaminant  (total T E Q i n c l u d i n g a n d e x c l u d i n g  mono-ortho P C B s )  f r o m H o r n b y Island, higher C Y P 1 A  i n blubber o f harbour seal pups  protein levels were found i n s k i n o f Puget S o u n d  pups c o m p a r e d to B r i t i s h C o l u m b i a ( B C ) pups (Fraser R i v e r + H o r n b y Island), suggests  differences  i n contaminant  levels  exposure.  Although  contaminant  which  data w a s not  a v a i l a b l e f o r P u g e t S o u n d seals s a m p l e d i n the present study, past studies h a v e s h o w n that seals f r o m this r e g i o n are s e v e r a l t i m e s m o r e c o n t a m i n a t e d w i t h o r g a n o c h l o r i n e s t h a n those f r o m  B C ( R o s s et  al.,  submitted;  S i m m s et al.,  2000b).  Contaminant  levels  e x p r e s s e d as total T E Q w e r e three t i m e s greater i n b l u b b e r f r o m r e c e n t l y - w e a n e d h a r b o u r seal p u p s f r o m P u g e t S o u n d c o m p a r e d to pups f r o m B C ( 1 5 4 v e r s u s 4 4 n g / k g ( S i m m s et al., 2 0 0 0 b ) .  lipid)  T h e differences i n C Y P 1 A protein levels detected i n the present  study m a y reflect these k n o w n differences i n contaminant exposure between regions.  4.2.1.2 Relationship between physiological factors and cutaneous CYP IA expression A s w i t h t h e c a p t i v e s e a l s , p u p s s a m p l e d i n t h e field w e r e o f s i m i l a r b o d y and  age.  There w a s no correlation between cutaneous C Y P 1 A  weight  expression and body  weight f o r either H o r n b y Island or Puget S o u n d pups. S e x differences were not observed for either pups or adults.  C Y P 1 A expression was not quantifiable in m a n y skin samples,  i n c l u d i n g those f r o m adults.  I n f a c t , n o n e o f t h e a d u l t m a l e s e a l s (n-1)  and only four o f  nine f e m a l e seals h a dE R O D activities above the L O Q . S m a l l s a m p l e size a n d l o w levels  136  of cutaneous C Y P 1 A expression limited detection of possible relationships between physiological factors and C Y P 1 A . 4.2.2  Significance of biomarker studies in free-ranging harbour seals Chronic exposure to environmentally-relevant levels of organochlorines has been  shown to cause immune and endocrine disruption in harbour seals (Brouwer et al., 1989; De Swart et al., 1994; Ross et al., 1996; Simms et al., 2000b). In a captive harbour seal study, seals fed fish from an area of high P C B contamination had diminished immune function (eg. reduced natural killer cell activity and T-cell function) compared to seals fed fish from a relatively uncontaminated area (De Swart et al., 1994; Ross et al., 1996). A n adverse effect level for immunotoxicity of 17 mg/kg lipid (S PCB) was reported for these harbour seals. Another captive feeding study found that circulatory levels of retinol (vitamin A) were lower in seals fed a relatively more contaminated diet (Brouwer et al., 1989). Additionally, Simms et al. (2000b) reported a correlation between circulatory retinol levels and contaminant exposure in free-ranging harbour seal pups in B C and Washington State, indicating that relatively low levels of contaminants affect retinoid homeostasis in these animals. Vitamin A is a dietary hormone that is important for growth and development and plays a role in resistance to microbial infections (Brouwer et al., 1989). One mechanism of contaminant-associated disruption of vitamin A is the binding of P C B hydroxy metabolites to the plasma protein complex that transports vitamin A (and thyroxine) to target organs.  As C Y P 1 A is thought to catalyze the  formation of OH-PCB metabolites from planar P C B congeners, induction of C Y P 1 A enzymes may in fact facilitate endocrine disruption. Contaminant levels in the range of those suggested to cause immunotoxic effects in captive harbour seals have been  137  measured in free-ranging harbour seal pups in Washington State (Table 1.1) (Ross et al., submitted), and correlations between contaminant levels and circulatory vitamin A have been observed in free-ranging harbour seal pups in B C and Washington (Simms et al., 2000b).  In the present study, correlations between contaminant levels and C Y P 1 A  expression were observed. Considering these combined results, we should be concerned that harbour seals in the present study are being negatively impacted by environmental contaminant exposure, and we should continue to develop and improve biomarker approaches, such as CYP1A, to assess these impacts. 4.3  Study 3 : Free-Ranging Killer Whales  EROD activity was not detected in any of the S9 fractions prepared from killer whale skin.  However, a faint protein band with the expected molecular weight of  C Y P 1 A was detected in some samples by immunoblot analysis. I expected to detect C Y P 1 A in killer whale skin, as cutaneous CYPlA-mediated enzyme activity and immunohistochemical staining have been detected in skin of other whale species (fin whale and sperm whale) (Marsili et al., 1998; Godard et al., 2002). Additionally, killer whales sampled in the present study were highly contaminated with organochlorines relative to other marine mammals in which C Y P 1 A has been measured, including harbour seals in this study (Tables 1.1, II. 1). Whereas BNF-treated harbour seals exhibited a strong C Y P 1 A response to which untreated samples could be compared, there was no positive reference sample available for killer whale CYP1A. Non-specific staining on immunoblots also made it difficult to distinguish a putative C Y P 1 A band in all killer whale samples. Structural differences  138  between killer whale and harbour seal skin (eg. no hair in whales) may also have affected the relative levels of C Y P 1 A expression in these two species. High contaminant levels in killer whales in British Columbia, and concerns about further declines of the endangered southern resident population underlie the importance of continued research into the development of minimally-invasive biomarker approaches for free-ranging cetaceans (Ross et al., 2000). 4.4  Improvement of CYP1A Detection in Skin Measurement of C Y P 1 A in harbour seal and killer whale skin might be improved  by using more sensitive techniques, such as quantitative real-time reverse transcription polymerase chain reaction (RT-PCR).  Quantitative real-time RT-PCR has been  suggested to be the most sensitive and reliable method for detection of low-abundance mRNA (Miller et al, 1999; Bustin, 2000; Bowen et al, 2000). In this process, total R N A is extracted from tissue and complementary D N A (cDNA) is synthesized via reverse transcription and amplified by PCR. Accumulation of the P C R product (DNA) is then quantified in real-time by fluorometric detection using a spectrofluorometric thermal cycler. Because C Y P 1 A gene sequences have not yet been determined for harbour seals or killer whales, primers would have to be designed based on the sequences of other marine mammal species (eg. largha seal and minke whale) (Teramitsu et al, 2000). Immunohistochemistry is another technique frequently used to detect C Y P 1 A expression and induction in tissues of both laboratory animals and wildlife (eg. fish, river otters, cetaceans) (Husoy et al, 1994; Sarasquete et al, 2000; Ben-David et al, 2001; Godard et al, 2002). Immunohistochemistry involves labelling and staining the protein of interest in situ in thin tissue sections and then assessing the degree of staining by  139  visualization with a microscope. One advantage of this method is that tissues do not need to be homogenized, a process that dilutes induced cells with non-induced cells (Husoy et al, 1994). Disadvantages of immunohistochemistry are that a highly specific primary antibody and controls to minimize the effect of non-specific staining are needed, and this method is more qualitative than quantitative (Sarasquete et al., 2000).  Staining is  generally scored for intensity (eg. 0=none to 4=very strong) and occurrence (eg. 0=no cells to 4=all), and these scores are multiplied to obtain a stain index on a scale of 0 to 16 (Ben-David et al, 2001). The sensitivity of immunodetection techniques, such as immunoblotting and immunohistochemistry, could also be improved by using species-specific antibodies. Although C Y P 1 A protein sequences have now been identified for a few marine mammal species, marine mammal-specific antibodies are not yet commericially available, and immunodetection remains dependent on cross-species reactivity of antibodies prepared against C Y P 1 A enzymes from other animal species. Immunohistochemistry could also be used to determine CYP1A localization prior to homogenization. It may be advantageous to isolate a particular skin layer in order to obtain a more CYPlA-concentrated homogenate,  or whole skin may provide a  sufficiently concentrated sample as well as a larger sample volume. Mammalian skin consists of an outer epidermal layer and an inner dermal layer. In humans and rats, the epidermis is the major site of xenobiotic metabolism in skin (Chapman et al, 1979; Bickers et al, 1982). Mukhtar and Bickers (1981) found that in BaP- and Aroclor 1254treated rats, the epidermis had the highest CYP1A enzyme activity when expressed per mg protein.  140  Cutaneous C Y P 1 A activity has been shown to differ between species, and Williams (1995) suggested that these differences may be due to skin thickness and the number of hair follicles present. In rats, topical treatment with methylcholanthrene (MC) caused induction of C Y P 1 A in hair follicles and sebaceous glands but not in the epidermis (Anderson et al, 1989).  Cetaceans (whales, dolphins, porpoises) have an  epidermal layer that is ten to twenty times thicker than that of terrestrial mammals (Geraci et al., 1986) and do not possess hair (or hair follicles).  To my knowledge,  information about C Y P 1 A localization in marine mammal skin and techniques for separating skin layers has not yet been published. However, at least one laboratory has localized C Y P 1 A in dermal endothelial cells, rather than epithelial cells, in cetaceans and in river otters (Wilson et al, 2000; Ben-David et al, 2001; Godard et al, 2002). 4.5  Effects of Storage Conditions on CYP1A Prolonged storage, storage temperature, and repeated freeze-thawing of tissue  samples and homogenates have been shown to affect detection of C Y P 1 A protein, and particularly enzyme activity (Pearce et al 1996; Anulacion et al, 1997; Yamazaki et al. 1997). In the present study, however, extensive precautions were taken to minimize these potential effects.  Liver and skin samples were immediately frozen in liquid nitrogen  following collection and were stored in liquid nitrogen (-196°C) or at -80°C until they were homogenized (from a few months to a year later). Following homogenization, S9 fractions were stored at -80°C, and C Y P 1 A analysis was carried out a few days later (eg. killer whale skin biopsies), several months later, or more than a year later (eg. C Y P 1 A protein analysis of captive seals). Tissue biopsies were kept frozen during sub-sampling, and homogenates were aliquotted into small volumes (eg. 60 pi) so that vials were not  141  repeatedly thawed and refrozen. Previously thawed samples were not used for the EROD assay but were used for the immunoblot assay i f necessary. When harbour seal skin S9 fractions were thawed and refrozen several times, it did not appear that C Y P 1 A detection by immunoblot analysis was affected. In the present study, differences in storage time among sample groups may have introduced some variability to CYP1A expression. However, studies in the literature have shown that duration of storage may or may not affect C Y P 1 A expression. Pearce et al. (1996) found that C Y P content and EROD activity decreased by 20-40% after human liver samples were frozen at -80°C for up to six months.  However, these same  measurements were relatively unaffected by storage of hepatic microsomes at -80°C for one or two years.  In hepatic microsomes from English sole (fish), A H H activity  decreased by 60% after five months of storage at -80°C (Anulacion et al., 1997). Yamazaki et al. (1997) found that there was no difference in C Y P levels and enzyme activity i f microsomes were prepared from fresh tissue and then stored at -80°C or made from tissues that were stored at -80°C prior to homogenization. Anulacion et al. (1997) also studied the effect of storage volume on C Y P 1 A (250 vs. 750 pi microsomal suspension/vial) and found no difference in A H H activity between volumes. In the present study, the use of smaller storage volumes (eg. 60 pi) may have affected sample stability. Additionally, S9 rather than microsomal fractions were used for C Y P 1 A detection, and S9 fractions may be less stable than microsomal fractions during prolonged freezing due to differences in storage solutions (Tris-HCl for S9 fractions versus sucrose for microsomes).  142  Although every effort was made to keep samples cold during laboratory procedures (eg. samples kept on dry ice during sub-sampling and in ice baths during homogenization),  unintentional  warming  of  samples,  especially  during  tissue  homogenization, may have decreased CYP1A enzyme activity. Yamazaki et al. (1997) found that detection of C Y P proteins by immunoblot analysis was not affected by leaving liver samples at 25°C for six hours; however, EROD activity declined to undetectable levels. The latter result highlights the sensitivity of enzyme activity to temperature and the need to keep samples cold. We expect that increasing the homogenization time will increase C Y P 1 A protein yield up to a certain point. However, because heat is generated during homogenization, it is possible that increasing the homogenization time will also result in enzyme degradation, thereby decreasing enzyme activity.  143  4.6  Conclusions  1) Minimally-invasive biopsy sampling techniques were used to obtain high quality liver and skin samples from harbour seals sufficient for C Y P 1 A analysis; 2) Both EROD activity and C Y P 1 A protein levels were quantified in liver and skin biopsies from harbour seals. 3) Two distinct immunoreactive C Y P 1 A protein bands, which likely correspond to CYP1 A l and CYP1A2, were detected in both liver and skin of harbour seals. 4) Oral B N F treatment induced C Y P 1 A protein levels in both liver and skin of harbour seals, but C Y P 1 A expression in skin was not positively correlated with that in liver. 5) C Y P 1 A expression in harbour seal liver was positively correlated with contaminant levels in blubber, suggesting that harbour seals in British Columbia are exposed to contaminant levels that are sufficiently high to elicit induction of hepatic enzymes. 6) Cutaneous C Y P 1 A expression did not correlate with contaminant levels, but these results were limited by small sample size. 7) Confounding factors (eg. age, sex, and condition) may have influenced C Y P 1 A expression in liver and skin from captive harbour seals.  These results further  highlight the importance of characterizing and minimizing natural factors prior to assessing contaminant effects. 8) C Y P 1 A could not be quantified in killer whale skin biopsies, possibly as a result of small sample size and/or low C Y P 1 A expression in this tissue. Immunohistochemical localization of C Y P 1 A and more sensitive techniques, such as RT-PCR, are needed to confirm the presence of C Y P 1 A in killer whale skin.  144  4.7  Summary and Future Studies The importance of biomarker studies in marine mammals is twofold. First, such  studies can provide us with information about the health of marine mammal populations, and second, they can provide an indication of the overall health of the marine environment. The development and application of sensitive, biologically relevant, and minimally-invasive biomarkers in sentinel marine species, such as harbour seals and killer whales, will facilitate conservation-based management of these animals and improve efforts to regulate anthropogenic chemicals to minimize environmental impact. In contrast to most other biomarker-based studies in marine mammals, we obtained tissue samples using minimally-invasive means from healthy, free-ranging animals, and we had a good understanding of the condition of each animal sampled (i.e. in the case of harbour seal pups, all animals were of approximately the same age and condition; in the case of killer whales, the age, sex, and dietary preference of each individual was known). Almost all previous biomarker studies in marine mammals have utilized tissue samples from dead animals (stranded or sacrificed), presenting both ethical concerns (killing animals for scientific purposes) and problems associated with poor sample quality (stranded animals). Sampling live, free-ranging marine mammals using minimally-invasive methods challenges.  presents considerable  logistical  and methodological  However, as conservation of these animals is the primary objective of  toxicological studies, such barriers must be overcome. In the present study, C Y P 1 A was successfully quantified in small liver and skin samples obtained by non-lethal means, providing a foundation for further development of this conservation-driven biomarker approach.  145  A 'weight of evidence' approach is needed to assess the effects of contaminant exposure on free-ranging marine mammals (Ross, 2000). This involves extrapolating between laboratory rodent studies involving single contaminants, captive marine mammal studies, and studies in free-ranging marine mammal populations.  Laboratory-based  rodent studies provide increased mechanistic understanding of biomarker responses and increased confidence of cause-effect relationships, whereas studies in free-ranging populations provide less confidence of cause and effect but increased ecological relevance (Ross, 2000). The present study showed that C Y P 1 A is inducible in harbour seals and provided evidence that ambient contaminant levels affect C Y P 1 A expression in free-ranging seals (ecological relevance).  Combined with the observations of other  ongoing marine mammal biomarker studies in our laboratory (eg. vitamin A) (Simms et al., 2000b) and mechanistic laboratory-based studies in the literature, results from this study suggest that current levels of contaminant exposure are high enough to warrant continued concern and regulatory action. Whereas many biomarker studies in marine mammals do not adequately control for natural confounding factors (eg. age and condition), the present study was designed to minimize the influence of these factors on CYP1A response.  Nevertheless, natural  factors, such as age and stress, may have affected C Y P 1 A expression in harbour seal pups.  This observation provides an important foundation for the design and  interpretation of future biomarker studies and further highlights the importance of characterizing and minimizing the input of natural factors when assessing cause-effect relationships between contaminants and biomarker responses.  146  Future studies of C Y P 1 A in marine mammals should (1) further improve the efficiency of C Y P 1 A extraction from skin biopsies; (2) develop, validate, and apply additional quantitative" methods to measure CYP1A; and (3) further explore the contribution of natural factors to C Y P 1 A expression.  This study demonstrated that  C Y P 1A can be quantified in small liver and skin biopsies using EROD and immunoblot assays.  However, as C Y P 1 A could not be quantified in some of the skin samples,  including samples from killer whales, strong conclusions about cause-effect relationships between contaminants and cutaneous C Y P 1 A expression could not be made.  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