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I-‘BRARY |II|IIIIIIIIIIIIIIIIIIIIII IIIIIIIIII IIIIIIIIII
Michlgan State 31293 01766 2556
University

 

 

 

This is to certify that the

dissertation entitled

Development of in vitro Assays to Assess the Effects
of Polychlorinated Diaromatie Hydrocarbons on
Immune Responses of Chinook Salmon

(Oncorpynchus tshawytscha)

 

 

presented by

George E. Noguchi

 

has been accepted towards fulﬁllment
of the requirements for

 

 

Doctor of Philosophy degree in Fish. & Wildl. 5‘ Envir. Tox.

Major pHessor

$11? 7231
0’

Date August 11, 1997

 

MSU is an Affirmative Action/Equal Opportunity Institution 042771

 

PLACE IN RETURN BOX to remove this checkout from your record.
To AVOID FINES return on or before date due.
MAY BE RECALLED with earlier due date if requested.

 

DATE DUE

DATE DUE

DATE DUE

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

1]” Wu

DEVELOPMENT OF IN V] T R0 ASSAYS TO ASSESS THE EFFECTS OF
POLYCl—[LORINIATED DIAROMATIC HYDROCARBONS ON [Ml/TUNE
RESPONSES OF CHINOOK SALMON (ONCORH YNCH US T SHA WY TSCHA)

By

George Edward Noguchi

A DISSERTATION

Submitted to
Michigan State University
in partial fulﬁllment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY
Department of Fisheries and Wildlife and
Program in Environmental Toxicology

1997

RE

diatoms

immune

 

 

 

ABSTRACT

DEVELOPMENT OF IN VIYRO ASSAYS TO ASSESS THE EFFECTS OF
POLYCHLORINIATED DIAROMATIC HYDROCARBONS ON IMMUNE
RESPONSES OF CHINOOK SALMON (ONCORHYNCHUS TSHA WYTSCHA)
By

George Edward Noguchi

In vitro assays were developed to evaluate the direct effects of polychlorinated
diaromatic hydrocarbons (PCDHS) on the chinook salmon immune system. Functional
immune responses including lymphocyte proliferation and antibody production were used as
endpoints for measuring effects on humoral immunity. Conditions which support mitogen-
induced proliferation of chinook salmon lymphocytes, in vitro, were deﬁned. Culture medium
supplemented with homologous salmon plasma was superior to medium supplemented with
either autologous salmon plasma or calf serum for supporting maximal mitogen
responsiveness. Soluble constituents in salmon plasma along with the proportion of
responsive lymphocyte subpopulations were identiﬁed as two factors which may contribute
to variability in mitogen responsiveness among individual ﬁsh.

The in vitro approach was effective for detecting effects of PCDHs on salmon
lymphocyte proliferation. The sensitivity of chinook salmon lymphocytes to 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD) was comparable to dioxin-sensitive mouse strains and
the potency of 3,3',4,4',5-pentachlorobiphenyl (PCB-126) relative to TCDD was similar to
other aryl hydrocarbon receptor (AhR) mediated effects in ﬁsh. Concentrations of Aroclor

1254 that were immunosuppressive (45 and 90 uM) were also cytotoxic. Results from the

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in vitro assays approximated in vivo effects, when effective concentrations were expressed
on a cellular or tissue basis, and suggest that concentrations of PCDHs in Lake Michigan
chinook salmon present a low risk of affecting lympho-proliferative responses.

Conditions are also described for the in vitro production of TNP-speciﬁc antibodies
by chinook salmon splenocytes and for the detection of antibody activity by ELISA and
the detection antibody afﬁnity by afﬁnity ELISA. Concentrations of TCDD that affected
antibody synthesis were comparable to concentrations that suppress lymphocyte
proliferation (3 and 30 nM ). PCB-126 exhibited immunosuppressive tendencies, although
great within-treatment variability diminished the statistical signiﬁcance of these effects.
Non-cytotoxic concentrations of Aroclor 1254 (0.45 - 9 pM) did not depress antibody
activity. The suppressive effects of TCDD and PCB-126 on antibody production were
accompanied by the elimination of antibodies with high affinity for the TNP antigen.
These results suggest B cells expressing high afﬁnity antigen receptor may be particularly

susceptible to TCDD and other AhR agonists.

 

Dedica:

SUP-Port.

 

Dedicated to my wife, Laura, and children, Anna and William, for their encouragement,
support, and love and to my parents, George and Betty, for their unquestioning support,

and appreciation of academic achievement and civic responsibility.

iv

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Work, JC‘.

CW Edt

 

ACKNOWLEDGMENTS

Primary support for this research was provided by the US. Fish and Wildlife Service,
the National Biological Survey and the US. Geological Survey. The author’s salary and
tuition were provided by a training agreement with the Great Lakes Science Center, Ann
Arbor, Michigan, and supported by Drs. Michael Mac and John Gannon. Additional support
was provided by the Aquatic Toxicology Laboratory and the Tissue Typing Laboratory at
Michigan State University, and the Fish Health Laboratory-Leetown Science Center,
Kearneysville, West Virginia.

I am indebted to my academic advisor, Dr. John Giesy for his guidance, support, and
encouragement throughout my graduate program at Michigan State. Other members of my
graduate committee also provided valuable assistance. Dr. Robert Bull, Department of
Medicine, was a valued counselor and provided technical resources at the Tissue Typing
Laboratory. Dr. Norbert Kaminski, Department of Pharmacology and Toxicology, provided
laboratory resources, technical assistance, and guidance. Dr. Frank D’Itri, Department of
Fisheries and Wildlife, provided guidance and helpful review comment on the draﬁ of this
dissertation.

Several other individuals provided invaluable assistance during the course of this
work. Jeff Allen assisted in ﬁsh care and sampling at the Great Lakes Science Center and

Carol Edsall provided technical assistance, moral support and guidance. Peggy Bull, Lori

 

Culham ,
ESSISiang;
King, B

“Else‘s .
Virginia

VaIUabie
Wilder;
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Michiga.

conkersa‘

 

Culham and John Davis at the Tissue Typing Laboratory, Michigan State University, provided
assistance with cell culture techniques, mitogen assays, and ELISA methodology. Dr. Louis
King, Biochemistry Department, Michigan State University, assisted with ﬂow cytometry
analyses and Drs. Stephen Kaattari, Virginia Institute of Marine Sciences, Gloucester Point,
Virginia, and Mary Arkoosh, National Marine Fishery Service, Newport, Oregon, provided
valuable reagents (monoclonal antibodies). The chinook salmon used in this research were
provided by Jim Copeland, Wolf Lake Hatchery, Michigan Department of Natural Resources,
Mattawan, Michigan. The students and staff of the Aquatic Toxicology Laboratory at
Michigan State provided technical assistance, moral support and endless hours of good

conversation.

TABLE OF CONTENTS

LIST OF TABLES ................................................... ix
LIST OF FIGURES ................................................... xi
INTRODUCTION .................................................... 1
Problem Statement ............................................... 1
Research Goals ................................................. 3

CHAPTER I. Literature Review: Immunological Disorders in Fish Associated with
Polychlorinated Biphenyls and Related Halogenated Aromatic Hydrocarbon
Compounds .................................................... 5

CHAPTER 2. In Vitro Assessment of Immunomodulatory Effects of Polychlorinated
diaromatic hydrocarbons (PCDH) on Chinook Salmon, Oncorhynchus

tshawytscha: Optimization of Lymphocyte Proliferation Assays. ........... 35
INTRODUCTION .............................................. 3 5
MATERIALS AND METHODS ................................... 37
RESULTS .................................................... 40
DISCUSSION ................................................. 43

CHAPTER 3. In vitro effects of Polychlorinated Diaromatic Hydrocarbons on Lympho-
Proliferative Immune Responses in Chinook Salmon (Onchorhynchus

tschwytscha). ................................................. 56
INTRODUCTION .............................................. 56
MATERIALS AND METHODS ................................... 58
RESULTS .................................................... 62
DISCUSSION ................................................. 67
CONCLUSIONS ............................................... 73

CHAPTER 4. Development of an Assay to Assess Effects of Polychlorinated Diaromatic
Hydrocarbons on In Vitro Antibody Production and Antibody Afﬁnity in

Chinook Salmon (Onchorhynchus tsclrwytscha) Splenocyte Cultures. ....... 93
INTRODUCTION .............................................. 93
MATERIALS AND METHODS ................................... 95
RESULTS ................................................... 100

vii

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BlBLlO‘I

 

DISCUSSION ................................................ 103

BIBLIOGRAPHY ................................................... 125

viii

 

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LIST OF TABLES

Table 1. Laboratory studies investigating the effects of HAT-Is on functional immune
responses in ﬁsh. Eﬁ‘ects are indicated by following symbols: (-) no statistically
signiﬁcant diﬁ‘erence between chemically treated and non-treated ﬁsh, (t)

signiﬁcant decrease in response, (1) signiﬁcant increase in response. ........

Table 2. Pathology of lymphoid tissues from ﬁsh exposed to HAHs. ..............

Table 3. Proliferative response of splenocytes ﬁ'om 10 salmon assayed under optimal

mitogen concentration (100 pg LPS mL") and incubation time (ﬁve days). . .

Table 4. Eﬂ‘ectiveness of autologous and pooled homologous salmon plasma at
supporting splenocyte proliferation. Cultures were stimulated with LPS (100 pg
mL") and incubated for ﬁve days. Results from assays using splenocytes from

ﬁve individual salmon (mean :t SE). .................................

Table 5. Toxic equivalency factors (TEFs) calculated for the salmon splenocyte
proliferation assay and other Ah-R mediated eﬂ'ects in ﬁsh. TEF = PCB 1C3,o

ITCDD 1cm. ..................................................

Table 6. Analysis of variance table for one-way ANOVA tests for eﬂ‘ects of PCDH
treatment on lymphocyte proliferation when replicate ﬁsh are analyzed together
(it = 4 for each compound). The percent of total variability explained by diﬂ‘erent

sources is the ratio of source sums-of-squares : total sums-of-squares. .....

Table 7. Analysis of variance table for randomized block ANOVA to test for effects of
PCDH treatment on lymphocyte proliferation when replicate ﬁsh are analyzed
together (it = 4 for each compound). Fish number was used as the blocking
factor. The percent of total variability explained by diﬁ‘erent sources is the ratio

of source sums-of-squares : total sums-of-squares. .....................

Table 8. Distribution of PCB 126 in cell cultures 24 hours after dosing with 33 ng PCB

126/mL. Results are ﬁ'om duplicate cultures, mean (SE). ................

ix

25

27

48

49

74

75

76

77

 

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Table 9. Comparison of in Vitro and in viva approaches to evaluating the effect of PCBs
on lymphocyte proliferation. Cellular concentrations of Aroclor 1254 were
estimated from the in vitro assay at the IC_,.,, and compared to the concentration
of Clophen A50 measured in rainbow trout that were dosed (IP) with 80 mg/kg
(Thuvander et al., 1993). Proliferation of rainbow trout splenocytes to LPS in
Clophen A50-treated ﬁsh was depressed by 31% compared to controls, although
the effect was not statistically signiﬁcant in that study. The assimilation efficiency
for Clophen A50 was calculated, based on data from the Thuvander study, as
the ratio of Clophen A50 concentratio measured tissues : dose of Clophen A50
administered. .................................................. 78

Table 10. Comparison of environmental concentrations of PCDHs in Lake Michigan
chinook salmon with estimated effect concentrations based on the in Vitro
splenocyte proliferation assay. ..................................... 79

Table I l. Weighted-average affinity constants (K) for TNP-speciﬁc antibodies in
supematants from TNP-LPS stimulated chinook salmon splenocyte cultures and
serum from TNP-KLH immunized rainbow trout. (aK - apparent affinity
constant; p- proportion of total TNP-speciﬁc antibody with a given aK) ..... l 10

Table 12. TNP-speciﬁc antibody activity of serum from TNP-KLH immunized rainbow
trout. Values from different 11 standard curves. ...................... 111

Table 13. Weighted-average affinity constants for TNP-speciﬁc antibodies in
supematants from chinook salmon splenocyte cultures that were stimulated with
TNP-LPS and dosed with TCDD or PCB 126. (aK - apparent affinity constant;

p- proportion of total TNP-speciﬁc antibody with a given aK) ............ 1 12

 

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LIST OF FIGURES

Figure 1. Halogenated aromatic hydrocarbons (HAHS). General structure of dioxins,
dibenzofurans and polychlorinated biphenyls along with representative planer
congeners. .................................................... 29

Figure 2. Schematic representation of the certain aspects of the teleost immune systems
(Sources: Ainsworth, 1992; Evans and Jaso-Friedmann, I992; Kaattari, 1992;
Sakai, 1992; Secombes and Fletcher, 1992). Abbreviations: APC (antigen
presenting cell), Ag (antigen), C (complement), IL (interlukins), MB (memory
B-cell), NCC (natural cytotoxic cell), PMN (polymorphonuclear granulocytes
also referred to as neutrophils). .................................... 31

Figure 3. Proposed mechanism for Ah receptor-mediated toxins. Modiﬁed from Richter,
1995 (Sources: Whitlock, 1993; Matsumura, 1994). .................... 33

Figure 4. Inﬂuence of plasma concentration on the proliferative response (mean + SE)
of splenocytes from an individual ﬁsh (ﬁlled bars) and pooled splenocytes from
three ﬁsh (hatched bars) to LPS. Background proliferation in control wells is
indicated by the open bars. ........................................ 50

Figure 5. Effect of mitogen concentration on the proliferative response (mean i SE) of
spleen (closed circles) anterior kidney lymphocytes (open circles) to LPS. . . . . 51

Figure 6. Kinetics of LPS-induced lymphocyte proliferation (mean in SE) of spleen
(closed circles) and anterior kidney lymphocytes (open circles). ............ 52

Figure 7. Effect of heterologous salmon plasma on splenocyte proliferation.
Splenocytes from an individual salmon (#14) were cultured with TCM
supplemented with plasma from 11 different salmon, autologous plasma (P14),
no plasma (NP), and stimulated with LPS (100 ug mL"). ................. 53

Figure 8. Characterization of splenocyte subpopulations using ﬂow cytometry. A)
cytogram (forward scatter vs. side scatter) of chinook salmon spleen cell
suspensions. B) F ITC fluorescence histograms for four gated regions ﬁom the
splenocyte cytogram. Mature B-cells (Ig’) are distinguished from [g cells by
FITC ﬂuorescence, as seen in region 2 (R2). .......................... 54

xi

 

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Figure 9. Proliferation of salmon splenocyte cultures stimulated with LPS and treated
with graded doses of TCDD (A), PCB-126 (B) and Aroclor 1254 (C). Results
are from in Vitro assays with cells from four individual ﬁsh (denoted by lower
case letters a, b, c, etc.) Values are meansiSE for 4 replicate cultures per
treatment. Signiﬁcant differences (a=0.05) between PCDH-treated cultures and
controls are indicated by an asterisk (*). Proliferation in control cultures (vehicle
alone) was not signiﬁcantly different from positive control cultures (no vehicle).
Background proliferation (cpm) in non-stimulated cultures (no LPS) from
individual ﬁsh were 562 (a), 252 (b), 238 (c), 683 (d), 769 (e), 906 (f), 847 (g),
1438 (h), 4670 (i), 1232 (j), 550 (k), and 951 (l). ....................... 80

Figure 10. Cytotoxicity of TCDD, PCB-126 and Aroclor 1254 to chinook salmon
splenocytes. Numbers of viable splenocytes were counted after 24 hr incubation.
Values are meansiSE for 4 replicate cultures per treatment. Signiﬁcant
differences (a=0.05) between PCDH-treated cultures and controls are indicated
by an asterisk (*). .............................................. 84

Figure 11. Characterization of the effects of Aroclor 1254 on splenocyte subpopulations
using ﬂow cytometry. A) cytograms (forward scatter vs. side scatter) and B)
FITC ﬂuorescence in the lymphocyte region (R2) of cell suspensions dosed with
0, 9 and 45 uM Aroclor 1254. C) Percentage of Ig’ lymphocytes (B cells) in the
lymphocyte region, based on FITC ﬂuorescense. ....................... 86

Figure 12. Dose-response curves for the effects of TCDD (O ), PCB 126 (A ) and
Aroclor 1254 (I) on chinook salmon splenocyte proliferation. Responses,
expressed as percent of control proliferation (cpm), from 4 replicate ﬁsh were
combined and plotted (meaniSE) against log PCDH concentration (M) ...... 88

Figure 13. Direct effects of TCDD (A), PCB 126 (B) and Aroclor 1254 (C) on
splenocyte proliferation with replicate ﬁsh combined. Values are meansiSE for
the 4 replicate ﬁsh shown in Figure 9. ............................... 90

Figure 14. Comparison of in vivo and in vitro effects of TCDD on lymphocyte
proliferation (CPM). The dose-response curve for in vitro effects (O) is based
on in Vitro assays with chinook salmon splenocytes. Data from Spitsbergen et
al., (1986) were used to construct the dose-response curve for in vivo effects of
TCDD on splenocyte proliferation in rainbow trout (A). ................ 92

Figure 15. Layout for affinity ELISA. Each supernatant sample was analyzed on an
individual ELISA plate. Duplicate wells are reserved for each combination of
coating antigen (TNP-B SA) concentration and competitor (TNP-lysine)
concentration. Shading represents the expected magnitude of antibody bound
to stationary antigen (TNP-BSA); I-greatest concentration of bound antibody,
III-least concentration of bound antibody. ............................ 113

xii

 

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Figure 16. Afﬁnity ELISA: Proportion of total TNP-speciﬁc antibody bound at different
concentrations of TNP-BSA coating antigen; A) serum from TNP-KLH
immunized rainbow trout. B) supernatant from TNP-LPS stimulated splenocyte
cultures, C) cumulative proportion of TNP-speciﬁc antibody bound at different
TNP-BSA antigen concentrations (0 rainbow trout serum, A splenocyte culture
supernatant). .................................................

Figure 17. Competitive-inhibition curves from affinity ELISA with A) supernatant from
TNP-LPS stimulated chinook salmon splenocyte cultures and B) serum from
TNP-KLH immunized rainbow trout ................................

Figure 18. Aﬁinity distribution of TNP—speciﬁc antibodies in A) supernatant from TNP-
LPS stimulated chinook salmon splenocyte cultures and B) serum from TNP-
KLH immunized rainbow trout. ...................................

Figure 19. A representative standard curve for TNP-speciﬁc antibody activity in
standard rainbow trout serum (I 650 nm OD; CI logit 650 nm OD), and dose-
response curves for supernatants from two TNP-LPS stimulated splenocyte
cultures (0 & A - 650 nm OD, O & A logit transformed 650 nm OD). . . . .

Figure 20. Direct effects of TCDD (A), PCB-126 (B) and Aroclor 1254 (C) on in vitro
antibody production by chinook salmon splenocytes. Values are means iSE for
the 4 replicate wells. Responses from in vitro cultures from 2 ﬁsh are shown for
each compound. Asterisk denotes treatments that are signiﬁcantly different than
vehicle control (a = 0.05). N - nonstimulated control, C - TNP-LPS stimulated
control, V - vehicle control. ......................................

Figure 21. Competitive-inhibition curves from afﬁnity ELISAs with supematants from
chinook salmon splenocyte cultures treated with TCDD or PCB-126; A) TNP-
LPS stimulated controls B) vehicle control C) 3 nM TCDD and D) 300 nM
PCB-126. ...................................................

Figure 22. Afﬁnity distribution of TNP-speciﬁc antibodies in A) TNP-LPS stimulated
controls B) vehicle control C) 3 nM TCDD and D) 300 nM PCB-126. .....

xiii

114

116

117

118

119

123

 

Problem .

Re

 

Lake \licl

|
salmonmt.

BKD app
role in r
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Lake .\1?,
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(Smith 6’
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1986) c

INTRODUCTION

Problem Statement

Recent declines in the chinook salmon, Oncorhynchus tshawytscha, population in
Lake Michigan have been attributed to massive die-offs of ﬁsh infected with Rem‘bacterium
salmoninamm, the causative agent of bacterial kidney disease (BKD; Smith, 1992). Although
BKD appears to be the ultimate cause of death, other factors are thought to have played a
role in rendering salmon susceptible to this otherwise ubiquitous bacteria. Organic
contaminants are among the group environmental factors which have been considered by the
Lake Michigan Chinook Salmon Mortality Subcommittee as possibly contributing to the
salmon mortality. Detectable levels of contaminants including, dioxins and dibenzofurans
(Smith et al., 1990) and polychlorinated biphenyls (PCB; Masnado, 1987; Smith et al., 1990;
Williams et al., 1992) have been measured in chinook salmon collected from Lake Michigan
waters. These polychlorinated diaromatic hydrocarbons (PCDH) are potent mammalian
immunotoxins (Vos and Luster, 1989) which reduce host resistance to infection and disease.
The efficacy of these compounds to suppress immune responses in ﬁsh is less deﬁnitive
because of the limited number of studies reported and apparent differences in sensitivity
among species.

Most laboratory studies evaluating the immunotoxicity of dioxins (Spitsbergen et al.,

1986) or PCBs (Cleland et al., 1988a & 1988b; Spitsbergen et al., 1988c; Thuvander et al.,

 

10931 hax
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howex'er.

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2

1993) have been conducted using rainbow trout, Oncorhynchus mykiss, and indicate that
immune responses of this species are not greatly affected at sublethal doses. Recent studies,
however, suggest that chinook salmon may be more sensitive to this class of chemicals. The
primary and secondary humoral immune responses of chinook salmon were signiﬁcantly
reduced in ﬁsh injected with PCB at a sublethal dose (54 mg Aroclor 1254 kg" salmon;
Arkoosh et al., 1994a). In addition, sublethal concentrations of PCBs and PAHs (polycyclic
aromatic hydrocarbons) present in Puget Sound, an urban estuary, were associated with
suppression of the secondary humoral immune response of juvenile salmon inhabiting the area
(Arkoosh et al., 1991). Concentrations of PCBs in Lake Michigan chinook salmon range
from 1.2 to 4.1 mg kg" in whole ﬁsh (Noguchi, unpublished data) and concentrations in
speciﬁc tissues range from 1 mg kg" in muscle (Williams et al., 1992) to as great as 13.6 mg
kg" in gonads (Williams and Giesy, 1992). PCBs accumulated in Lake Michigan salmon are
enriched in toxic (planer) PCB congeners such that the toxic potency is estimated to be about
3.3 times greater than Aroclor 1254, a technical PCB mixture (Williams and Giesy, 1992).
Therefore, the toxic potency attributable to the dioxin-like compounds in PCB mixtures in
Lake Michigan chinook salmon are equivalent to a concentration of Aroclor 1254 that is 3.3
times greater or 4 to 13.5 mg kg" whole body and up to 44.9 mg kg" in selected tissues.
These concentrations approach the dose (54 mg kg") that has been shown to be
immunosuppressive and indicate that concentrations of PCBs in Lake Michigan chinook
salmon may be high enough to suppress immune ﬁrnctions and potentially decrease the
resistance of salmon to disease.

Although concentrations of PCBs in Great Lakes ﬁsh have decreased signiﬁcantly

over the past two decades they appear to be stabilizing at concentrations that still exceed

 

 

mam 513k
estimated
endpoints

because c

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3

many state and federal action levels (Cunningham et al., 1990) and are above concentrations
estimated to be protective for sensitive wildlife species (Ludwig et al., 1993). Immunological
endpoints may help to identify ﬁsh species that are at greater risk of infection and disease

because of their sensitivity to PCBS and other xenobiotics.

Research Goals

The salient questions that need to be answered are whether PCBS and other PCDHS
are immunosuppressive to chinook salmon at concentrations observed in Lake Michigan and
if their effects contribute to the susceptibility of salmon to BKD. This is a large task and no
single study will completely answer the question. My research was designed to answer some
fundamental questions regarding the sensitivity of chinook salmon to PCDHs and allow for
more accurate predictions of the immunological consequences of environmental PCDH
concentrations. In Vitro assays were adapted to evaluate the direct effects of PCDHs on
target cells of the salmon immune system. Functional responses such as lymphocyte
proliferation and antibody production were used as endpoints for measuring
immunosuppressive action. The efficacy of non-ortho (dioxin-like) substituted PCB
congeners were assessed to determine if structure-activity relationships indicative of an Ah-
receptor mediated mechanism could be observed. Cellular concentrations of PCBs were
measured to provide tissue-speciﬁc doses. As with any in Vitro approach, effects measured
in cell cultures do not necessarily measure whole animal responses. However, in vitro tests
are critical for identifying sensitive cell-types and responses, and mechanisms of action.
These assays could be adapted for assessing effects following in vivo exposure to

contaminants, for monitoring the immune status of feral salmon and they could also be used

4

to test effects of weathered complex mixtures of residues extracted from ﬁsh tissues.

CHAPTER I

Literature Review: Immunological Disorders in Fish Associated with
Polychlorinated Biphenyls and Related Halogenated Aromatic Hydrocarbon

Compounds

(Submitted for publication as a chapter in Fish Diseases and Disorders, Volume 11]: Non-

Infectious Disorders, eds: J .F. Leatherland and P.T.K. Woo)

INTRODUCTION

The immune system protects the body from disease by detecting and neutralizing
disease-causing pathogens (viruses, bacteria, fungi, and parasites) and transformed
(neoplastic) cells. In order for the immune system to be effective it must be capable of
discriminating between what is foreign and what is not foreign, i. e., ”self”. The process
of self-nonself discrimination involves intricate interactions between target cells (e.g.,
pathogens and tumor cells) and both cellular and humoral (soluble) elements of the
immune system. Once foreign agents are detected, they are subjected to a vast array of
effector cells (phagocytes, granulocytes, cytotoxic cells and natural killer cells) and soluble
factors (antibodies, complement) that facilitate neutralizing, killing and clearing of the

inducing agent. Disruption or modulation of these interactions by drugs or chemical

 

 

 

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6

contaminants is the subject of immunotoxicology. Exposure to immunotoxic chemicals
may result in a variety of disorders including immunosuppression, immunopotentiaiton,
immuodeﬁciency, hypersensitivity or autoimmuity (Dean and Murray, 1991). Although
most of what is known about the action of immunotoxic compounds is based on the
mammalian immune system there is increasing interest in assessing effects on lower
vertebrates, some of which may accumulate high concentrations of immunomodulating
chemicals in the environment.

Fish immunotoxicology is an emerging ﬁeld of study. Recent reviews (Weeks et al.,
1992; Dunier and Sivvicki, 1993; Wester et al., 1994; Zelikoff, 1994, Anderson and Zeeman,
1995) and symposia (Stolen and Fletcher, 1994) report on the manner in which immune
functions in ﬁsh may be modulated by toxic xenobiotic compounds especially mammalian
immunotoxins or by pollutants associated with contaminated habitats where ﬁsh health is
impaired. However, compared to mammalian immunotoxicology, where efforts have been
focused on relatively few, well characterized and extensively investigated animal models,
much less is known about the effects in ﬁsh. This is due, in part, to the large number of ﬁsh
species studied, the lack of many ﬁsh-speciﬁc reagents (e.g., monoclonal antibodies that
detect cell-surface markers on ﬁsh leucocytes and secretory products), and fewer researchers
in the ﬁeld. Nevertheless, there are published reports describing lesions in lymphoid tissues,
altered immune functions, or increased disease susceptibility in toxicant-exposed ﬁsh or in ﬁsh
collected from contaminated areas (Tables I & 2). This review characterizes immunological
disorders in ﬁsh associated with the widespread environmental contaminant, polychlorinated

biphenyls (PCBS), and related halogenated aromatic hydrocarbons (HAHS). Special attention

 

 

 

 

 

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7

is devoted to comparing the sensitivity of ﬁsh species, identifying sensitive immunological

endpoints and postulating mechanisms of action.

TOXICITY OF HALOGENATED AROMATIC HYDROCARBONS (HAHS)

Halogenated aromatic hydrocarbons comprise a class of chemicals that induce
pleiotropic effects in mammals including immunomodulation (Vos and Luster, 1989).
Polychlorinated dibenzofurans, polychlorinated dibenzo-p-dioxins (dioxins) and PCBS are
among the most toxic HAHs (Figure l) and are also ubiquitous environmental contaminants.
Because of their resistance to degradation and high lipophilicity, HAHS tend to be
biomagniﬁed in aquatic food chains. As a result, detectable concentrations of HAHs have
been measured in ﬁsh throughout the North America (Smith et al., 1990). The most toxic
member is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Several ﬁsh species are very
sensitive to the lethal effects of TCDD (LD50 3-16 pg kg"; Kleeman et al., 1988), particularly
when compared to sensitive mammalian species. In fact, the early life stages of salmonids are
most sensitive to TCDD (LDso 0.065 - 0.230 pg kg"; Walker et al., 1991; Walker and
Peterson, 1991).

The mechanism by which TCDD exerts many of its toxic and biochemical effects is
believed to require binding to the cytosolic aryl hydrocarbon receptor (AhR; Poland and
Knutson, 1982). Among the sublethal effects associated with AhR binding is the induction
of cytochrome P4501A1, a mixed ﬁmction oxygenase responsible for HAH metabolism. PCB
congeners that are structurally similar to TCDD, in that they can attain a planar conﬁguration
and are chlorinated in meta and para positions, also bind the AhR and induce P4501Al

activity. Of the 209 PCB congeners, relatively few have high afﬁnity for the AhR (Safe,

 

 

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1987). In ﬁsh only, nonortho-substituted tetrachloro (3,3',4,4'-tetrachloro biphenyl, 3,3',4,5'-
tetrachloro biphenyl), pentachloro (3,3',4,4',5-pentachloro biphenyl) and hexachlor
(3,3',4,4',5,5'-hexachloro biphenyl) congeners are known to induce AhR-mediated responses
(Janz and Metcalfe et al., 1991; Walker et al., 1991; Newsted et al., 1995). AhR-active PCB
congeners are a small percentage of the total mass of commercial PCB formulations such as
Aroclor® 1254, which contains over 50 different congeners (Ballschmiter and Zell, 1980).
Thus, compared to TCDD, greater doses of commercial PCB mixtures are required to
produce similar effects (e. g., chinook salmon, Oncorhynchus tshawystscha, LDSO: 270 mg

Aroclor® 1254 kg"; Arkoosh et al., 1994a).

OVERVIEW OF THE TELEOST IMMUNE SYSTEM

The detection of sublethal effects of PCBs and other HAHs on the ﬁsh immune system
has evolved along with the ﬁrndamental understanding of immunological processes in ﬁsh.
The teleost immune system including non-speciﬁc and speciﬁc immunity, humoral or antibody
producing and cell-mediated responses, is shown in Figure 2. The piscine immune system as
it relates to protective immunity (innate and acquired) is comprehensively reviewed earlier
(van Muiswinkel, 1995; Iwama and Nakanishi, 1996). The intent is to provide a framework
with which to consider the implications of immunotoxic effects and not to describe in great
detail all aspects of the ﬁsh immune system.

Phagocytic cells analogous to mammalian monocytes, macrophages and neutrophils
(Ellis, 1977; Fange, 1992) confer nonspeciﬁc immunity by detecting, engulﬁng, killing and
clearing pathogens. Phagocytes serve as both the ﬁrst line of defense against infection and

as effector cells in the humoral immune response. Natural cytotoxic cells (NCC) detect and

 

 

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9

lyse transformed target cells and protozoan parasites (Evans and Jaso-Friedmann, 1992).
Like their mammalian counterpart, natural killer cells (NK), NCC induce death in target cells
by necrotic and apoptotic mechanisms (Greenlee et al., 1991) and are believed to play an
important role in the surveillance of tumor cells. Antigen-presenting cells (APC) are
phagocytic cells, typically macrophages, that internalize and process antigen and present
processed antigen to T-cells (Vallejo et al., 1992). This results in T-cell activation. The
existence of T-cells in ﬁsh has been based on functional criteria including responses to
mammalian T-cell mitogens (Sizemore et al., 1984; Tillitt et al., 1988), mixed lymphocyte
reactions (Kaattari and Holland, 1990), and delayed type hypersensitivity reactions
(Stevenson and Raymond, 1990); mammalian T-cells are identiﬁed by the presence of speciﬁc
T-cell receptors. Such receptors have yet to be characterized for ﬁsh lymphocytes
(Chilmonczyk, 1992; Manning and Nakanishi, 1996). T-cells along with macrophages
function as accessory cells in the humoral immune response by secreting soluble factors, such
as interleukins (ILs), that are required for B-cell activation, proliferation and differentiation
(Kaattari, 1992). B lymphocytes express antigen receptors, i. e. membrane immunoglobulins
(DeLuca et al., 1983), and are capable of binding free (non-processed) antigen. Some
polymeric antigens and rrritogens can activate B-cells without the participation of T-cells, and
are referred to as thymus independent or T-I antigens. Antigens that require T-cell
involvement to activate B-cells are termed thymus dependent antigens, T-D. B-cells activated
by T-D antigens proliferate and differentiate into either plasma cells, which produce and
secrete antibodies, or memory B-cells. Antibodies circulate in the blood stream and bind to
speciﬁc antigenic features (epitopes) on pathogens that activated the B-cells. These antibody

coated (opsonized) pathogens are targeted for deletion by phagocytic cells (macrophages) or

 

 

 

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destroyed by complement-mediated cell lysis (Saki, 1992). Memory B-cells do not participate
in the initial or priming exposure to antigen but respond to secondary and subsequent
encounters with the speciﬁc antigen (Arkoosh and Kaattari, 1991). Secondary humoral
responses to antigen occur more rapidly and with greater intensity (more antibody producing
cells and higher antibody titers) than the primary response (Arkoosh et al., 1991).

The major lymphoid tissues in teleosts include the anterior kidney (pronephros),
thymus and spleen. The anterior kidney is the principal hemopoietic tissue and also functions
as a primary lymphoid tissue for B-cell maturation (Kaattari, 1992). The thymus is the
primary lymphoid tissue in mammals where T-cell differentiation and maturation occurs. In
ﬁsh, the thymus is believed to play a similar role, although the precise ﬁrnction is not as well
understood (Chilmonczyk, I992). Mature T and B lymphocytes migrate from primary
lymphoid tissues into the blood stream and concentrate in secondary lymphoid tissues (e. g.,
spleen). The spleen contains high numbers of lymphocytes and macrophages and it also
functions as a ﬁlter to trap antigens and allow maximal contact between antigen and

immunoreactive cells.

EFFECTS OF HAHS ON HUMORAL IMMUNITY
Humoral immune responses, particularly the antibody-forming cell response (AFC),
are among the most sensitive indicators of HAH immunotoxicity in higher vertebrates (Davis
and Safe, 1988; V05 et al., 1989; Kerkvliet and Burleson, 1994). The AFC response is a
measure of the number of antibody forming cells (plasma cells) that are produced in response
to immunization with antigen and, therefore, represents an integrated measure of B-cell and

accessory cell (macrophage and T-cell) ﬁrnction. The degree to which HAHS affect humoral

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responses appears to be influenced by many factors which include fish species, type of antigen
(T-D or T-I) and mode of immunization (in mm or in VIII’O).

Primary humoral responses to T-l antigens are more affected by HAHS than primary
responses to T—D antigens (Table I). In rainbow trout, ()ncorhynchus mykiss, PCB treatment
(500 mg Clophen® A50 kg" diet) signiﬁcantly reduced the humoral response (antibody titer)
to Vibrio anguillamm 0 antigen, a T-I antigen (Thuvander et al., 1991). Whereas, humoral
responses in trout to T-D antigens were not affected by PCBS (Cleland et al., 1988a;
Thuvander et al., 1993) or TCDD (Spitsbergen et al., 1986). Similarly, the primary in vitro
AFC response to a T-I antigen (TNP-LPS), but not a T-D antigen (TNP-KLH), was
depressed in juvenile chinook salmon receiving a single dose of Aroclor® 1254 (54 mg kg";
Arkoosh et al., l994a). Low doses of PCB 126 (0.01 mg kg") actually enhanced the AFC
response in channel catﬁsh, Iota/urns punctams, to a TD antigen, yet, the response was not
signiﬁcantly affected by higher doses (0.1 and 1 mg kg"; Rice and Schlenk, 1995). The
differential effect of HAHs on humoral responses to T-D and T-I antigens in ﬁsh may reﬂect
differences in the sensitivity of lymphocyte subpopulations. A discussion of the cellular
targets of HAH-induced immunotoxicity is in a later section.

The effect of HAHs on B-cell mediated immunity in chinook salmon indicates that
secondary or amnestic responses may be more sensitive than the primary response. The
primary in vitro AFC response of juvenile salmon to TNP-KLH (a T-D antigen) was not
affected by PCB treatment (54 mg Aroclor® 1254 kg"), however, the secondary response
was reduced by more than 90% compared to untreated controls (Arkoosh et al., l994a).
Because this effect occurred at a dose that was less than half of the ED50 (118 mg Aroclor®

1254 kg") for HAH-sensitive mice (C57BL/6; Davis and Safe, 1989) it would appear that

 

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12

chinook salmon is one of the more sensitive species in terms of PCB-induced

immunosuppression.

EFFECTS ON NON-SPECIFIC AND CELLULAR IMMUNITY

Although relatively few studies on the immunotoxicity of HAHS included non-speciﬁc
and cellular immunity, there is evidence which suggests species-speciﬁc differences in the
sensitivity to these compounds (Table 1). The phagocytic activity of peritoneal macrophages
is a measure of non-speciﬁc immunity and this was not affected in rainbow trout treated with
a lethal dose of TCDD (10 pg kg"; Spitsbergen et al., 1986). However, the oxidative burst
activity in stimulated phagocytes, another indicator of immune competence, was signiﬁcantly
reduced in channel catﬁsh treated with sublethal amounts of PCB 126 (0.1-1.0 mg kg"; Rice
and Schlenk, 1995). In the same study, the activity of natural cytotoxic cells (NCC) was also
suppressed in PCB 126-exposed catﬁsh (1.0 mg kg"). In contrast, NCC activity was not
inhibited in rainbow trout receiving prolonged dietary exposure to Aroclor® 1254 (3-3 00 mg
kg"; Cleland and Sonstegard, 1987). Although these studies examined the effects of different
HAHs it would appear that the non-speciﬁc and cellular immune responses in rainbow trout
are more resistant to HAHs compared to channel catﬁsh.

The proliferative response of lymphocytes to T-cell mitogens is another measure of
cellular immunity. Neither TCDD (Spitsbergen et al., 1986) nor Clophen® A50 (Thuvander
et al., 1993) signiﬁcantly affect the response of rainbow trout lymphocytes to T-cell nritogens.
However, in rainbow trout previously immunized with KLH, the responses to both
phytohaemagglutinin (PHA; a mammalian T-cell mitogen) and lipopolysaccharide (LPS; a

mammalian B-cell mitogen) were signiﬁcantly enhanced following exposure to Clophen® A50

 

 

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(80 mg/kg; Thuvander et al., 1993). These results suggest that HAHs differentially affect
lymphocyte activity and it depends on the immune status of the ﬁsh prior to chemical
exposure. Enhanced mitogen responsiveness was also observed by Faisal et al. (1991a) in
contaminant-exposed spot, Leiostomus xanthurus. The authors suggested that greater LPS
responsiveness of spot leucocytes may have been due to contaminant-induced inhibition of
suppressor T-cell activity. [PCBS have been shown to decrease T suppressor activity of
murine leucocytes (Kerkvliet and Baecher-Steppan, 1988).] Lymphocytes with T suppressor
activity are believed to participate in the regulation of immune ﬁrnctions in ﬁsh (Kaattari et
al., 1986), however, the role of suppressor T cells in mediating HAH-induced

immunomodulation has not been ﬁrlly explored.

PATHOLOGY OF LYMPHOID TISSUES

Thymic involution, or reduction in size and cellularity of the thymus, is an indication
of TCDD toxicity in mammals (Vos et al., 1989). TCDD-induced lesions in the lymphoid
tissues of ﬁsh have been detected but they usually occur at lethal or near-lethal doses (Table
2). Thymic lesions, characterized by multiple invaginations of the thymic epithelium
extending into a lymphoid-depleted cortex, were described by Spitsbergen et al. (1988a) in
rainbow trout receiving a lethal dose of TCDD (10 pg TCDD kg"; the 80-d LDSO). These ﬁsh
also exhibited splenic lymphoid depletion as well as depletion of lymphomyloid elements in
the pronephros and midkidney. No lesions were found in trout dosed with sublethal amounts
of TCDD. Splenic lymphoid depletion was detected by van der Weiden et al. (1992) in
rainbow trout dosed with lower levels of TCDD (0.6 and 3.06 pg TCDD kg"). These doses

were near or below the lethal threshold (20% mortality at 3.06 pg TCDD kg") and in the

 

 

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14

range where moderate hepatic EROD activity (ECSU 0.79 pg TCDD kg") was induced.
Differences in the sensitivity of various rainbow trout strains to TCDD have been reported
for other toxicological endpoints (early life stage mortality; Walker and Peterson, 1991) and
could have contributed to the differences in sensitivity. Percids are also sensitive to TCDD.
Mild to moderate splenic lymphoid depletion in yellow perch, Percaﬂavescens, occurred at
lower doses of TCDD (5 pg kg") than thymic involution and pronephric lymphoid depletion
(>25 pg TCDD kg"; Spitsbergen et al., 1988b), however, these lesions were not detected at
doses below the 80-d LD50 (3 pg TCDD kg").

In studies in which ﬁsh were exposed to PCBS, lesions in thymic and/or splenic tissues
were not always observed. Splenic lesions were found in rainbow trout exposed to dietary
levels of PCBS ranging from 10 to 500 mg Aroclor® 1254 kg" (Nestel and Budd, 1975;
Hendricks et al., 1977; Spitsbergen et al., 1988c). These levels were not reported to be lethal
over the course of these studies (75 days to 12 months). Thymic and splenic hypocellularity
were noted in rainbow trout injected with a sublethal dose (80 mg/kg) of Clophen® A50
(Thuvander et al., 1993). However, no lesions were detected in rainbow trout fed 500 mg
Clophen® A50 kg" for 10 weeks; although signiﬁcant effects on humoral immunity were
observed (Thuvander et al., 1991). Similarly, humoral immune responses were suppressed
in chinook salmon injected with 54 mg Aroclor® 1254 kg", but no lesions in lymphoid tissues
were detected (Arkoosh et al., 1994a). Thus, lesions in lymphoid tissues are not always

associated with HAH exposure or HAH-induced effects on immune ﬁrnctions.

EFFECTS ON DISEASE RESISTANCE

The ultimate manifestation of immunotoxicity is the ability of a toxicant to increase

 

 

 

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disease susceptibility. However, relatively little is know about the effects of HAHS on disease
resistance in teleosts other than in rainbow trout. In this species, disease resistance has not
been compromised by exposure to HAHS. Neither median time to death (MTD) nor
cumulative mortality in rainbow trout challenged with infectious hemopoietic necrosis virus
(IHNV) was affected by exposure to TCDD (0.01 - 1 pg kg" body weight) or PCB (5 - 500
mg Aroclor® 1254 kg" diet; Spitsbergen et al., 1988c). However, lesions characteristic of
IHNV-induced disease were more severe in ﬁsh treated with PCBS or TCDD, which indicates
that HAHs may enhance progression of the disease without hastening mortality. Similarly,
MTD in rainbow trout challenged with Yersinia ruckeri was not shortened following 90-d
water-home exposure to PCBs (0.23 - 2.9 ug Aroclor® 1254:1260 L"; Mayer et al., 1985).
In addition, resistance of rainbow trout to Vibrio anguillarum was not compromised in ﬁsh
fed HAH-contaminated diets consisting of Paciﬁc or Great Lakes coho salmon (0.02 - 2.3 pg
PCB g"; Cleland et al., 1988b). These ﬁndings are consistent with the relative ineffectiveness
of HAHs at altering humoral and cellular immunity in this species. However, impaired disease
resistance associated with HAH exposure has been reported for other ﬁsh species.
lrmnunization against Aeromonas hydrophi/a was ineffective at protecting PCB-treated (70
mg Aroclor® 1232/kg) channel catﬁsh from a challenge with the virulent bacterium (Jones
et al., 1979). More recently, Arkoosh et al. (l994a) reported that juvenile chinook salmon
retrieved from an urban estuary contaminated with PCBs and polycyclic aromatic
hydrocarbons (PAHs) suffered higher mortality following exposure to Vibrio anguillarum
than salmon from a noncontaminanted estuary or salmon held in a hatchery. The humoral
immune response was depressed in salmon form the same contaminated estuary (Arkoosh et

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FIELD OBSERVATIONS

Establishing cause-effect relationships between a suspected chemical agent and effects
observed in wild ﬁsh populations (i. e., epizootiology) can be complicated by uncontrollable
factors that may potentiate, or mask or independently induce the effect(s). Nevertheless,
detection of strong associations between chemical contaminants and biological effects can
strengthen the argument for causality when the same effects have been demonstrated in
controlled laboratory studies.

Altered immune ﬁinctions have been detected in feral ﬁsh from ﬁeld locations known
to be contaminated with HAHS and other organic and inorganic contaminants. Carlson and
Bodammer (1994) found that the humoral immunity was compromised in winter ﬂounder,
Pleuronectes americaims, inhabiting an area of Long Island Sound (Morris Cove - New
Haven Harbor) that was contaminated with PCBS, PAHs and heavy metals. The authors
measured the in vitro AFC response of splenic lymphocytes and observed that the response
to both T-I (TNP-LPS) and TD (TNP-KLH) antigens in ﬁsh from the Morris Cove site was
about 50% lower than the response in ﬁsh from a less contaminated reference site. Humoral
immunity was also depressed in juvenile chinook salmon that were collected from an HAH-
PAH contaminated urban estuary in Puget Sound (Arkoosh et al., 1991). Although no effects
were observed in the primary response, the secondary in vitro AFC response of anterior
kidney leukocytes in salmon from the contaminated urban estuary was signiﬁcantly less than
the response in hatchery salmon or in salmon collected from a non-urban estuary. Several
reports have also documented altered immune ﬁrnctions in ﬁsh from sections of the Elizabeth
River (Virginia) that are heavily contaminated, primarily with PAHs but also with HAHs

(Huggett et al., 1992). The immunological disorders in ﬁsh from that system include,

'11" A.

 

 

 

 

 

l7

diminished natural cytotoxic cell activity (Faisal et al., 1991b), reduced phagocytic and
chemotactic activity of kidney macrophages (Weeks et al., 1990), and altered responsiveness
of pronephric lymphocytes to mitogenic stimulation (Faisal et al., 1991a). The abundance of
macrophage aggregates in wild ﬁsh has been positively correlated with concentrations of
HAHS and other contaminants in bottom sediments (Blazer et al., 1994). Macrophage
aggregates are accumulations of pigmented macrophages, in the spleen, kidney and liver with
normal physiological and immunological ﬁinctions (Wolke, 1992). Changes in abundance of
macrophage aggregates may be due to contaminant-induced stress. Establishing causal
relations between immunological disorders and environmental exposure to HAHS requires an

understanding of the mechanisms by which HAHS modulate the immune system.

MECHANISMS OF IMMUNOMODULATION

Many of the pleiotropic effects attributed to HAHS are mediated by a process that
requires initial binding of ligand to the AhR (Figure 3). Support for the essential role of AhR-
ligand binding is based primarily on two lines of evidence 1) quantitative structure-activity
relationships between AhR binding afﬁnity and toxic potency and 2) the differential sensitivity
of mouse strains possessing alleles encoding high and low affinity AhR. TCDD is the
prototypical AhR agonist. The AhR binding affinity of other HAHS is greatest for planer
congeners that are stnicturally most similar to TCDD (Poland et al., 1976; Safe et al., 1986).
Toxic responses (weight loss, thymic atrophy, and immunomodulation) and biochemical
responses (enzyme induction) to HAHS are correlated with AhR binding afﬁnity (Poland et
al., 1976; Safe, 1987; Davis and Safe, 1988; Kerkvliet et al., 1990a). Thus, TCDD-like

toxicity is observed with HAH congeners that bind with high affinity to the AhR. Similarly,

18

mouse strains possessing the AhR allele that expresses a receptor with high TCDD binding
afﬁnity are much more sensitive to biologically active HAHS than mouse strains that express
receptor with low binding affinity (Silkworth and Gaberstein, I982; Vecchi et al., 1983;
Tucker et al., 1986; Birnbaum et al., 1990; Kerkvliet et al., 1990b). The TCDD binding
afﬁnity of AhR in responsive mouse strains (C57BL/6J mice) is 10 fold greater than in
DBA/2J mice, a low responsive strain (Okey et al., 1989). Binding of HAHs to the AhR is
a prerequisite for many physiological and biochemical effects.

The most well studied of the TCDD-related effects is the induction of cytochrome
P4501Al (a mixed function oxygenase ) that is encoded by the CYP1A1 gene (Figure 3).
P4501A1 induction requires initial binding of ligand (TCDD or other active HAH congeners)
to the AhR followed by a transformation of the receptor and translocation of the ligated AhR
to the nucleus and binding with ARNT, the aryl hydrocarbon nuclear translocator protein
(Nebert and Jones, 1989; Whitlock, 1990; Hankinson, 1995). In the nucleus, the AhR-ligand
heterodimer binds to dioxin responsive enhancer (DRE) regions in the 5' ﬂanking region of
the CYP1A1 gene. Binding of the AhR-ligand complex to DREs enhances transcription of
the downstream gene, CYP1A1. Thus, the DRE binding form of the AhR-ligand complex
ﬁinctions as a transcription factor for CYP1A1, resulting in elevated CYP1A1 transcription
and increased levels of the P4501Al protein.

Induction of detoxiﬁcation enzymes such as P4501A1 is an adaptive response and not
necessarily a measure of toxicity. Whether TCDD acts through this same mechanism to
induce toxic responses has yet to be demonstrated unequivocally. However, CYP1A1 is not
the only gene that is responsive to the AhR. Sutter and Greenlee (1992) have classiﬁed a

number of genes that belong to the Ah gene battery. Members of this family include growth

l9

factors (interlukin-l and transforming growth factor-or) and intracellular proteins involved in
signal transduction (phospholipase A2, protein kinase C and tyrosine kinases). It is possible
that some of the TCDD-related toxic effects may involve direct interactions with DREs that
regulate the transcription of growth factors or other regulators of cellular activity. There is
also evidence that the AhR-ligand complex can modulate the phosphorylation of cytosolic
proteins that are involved in signal transduction pathways (Puga et al., 1992; Matsumura,
1994). Such alterations could affect the responsiveness of cells to extracellular stimuli.
Recent studies by Masten and Shiverick (1995) suggest that the suppressive effect of
TCDD on B-lymphocyte activation and antibody production may involve a direct effect of the
TCDD-AhR complex on gene expression. CD19 is a membrane receptor expressed on the
surface of mammalian B lymphocytes and participates in B-cell activation and differentiation
(Kehrl et al., 1994; Tedder et al., 1994). Treatment of a human B-Iymphocyte cell line (IM-9)
with TCDD resulted in a 67% decrease in CD19 mRNA, indicating that TCDD may affect
CD19 gene expression. The promoter region for the CD19 gene contains a binding site for
BSAP, the B cell lineage-speciﬁc activator protein (Kozmik et al., 1992). BSAP regulates
CD19 gene expression and is believed to play a role in early neurological development as well
(Urbanek et al., 1994). The DNA binding site for BSAP contains a ﬁve base sequence
identical to the DRE consensus sequence. These results suggest that binding sites for the
AhR-ligand complex exist in regulatory regions for genes that modulated B-cell activation and
differentiation. In the case of CD19, the AhR-ligand complex may compete with the
endogenous ligand (BSAP) for binding to the BSAP binding site, resulting in reduced CD19
transcription. Fewer CD19 transcripts may result in reduced expression of CD19 on the cell

surface and a diminished capacity to bind and respond to extracellular stimulation. Thus, the

20

DNA binding activity of the TCDD-AhR complex may not only act to ”turn genes on" but
may also interfere or compete with other transcription factors thereby reducing gene
expression and altering cellular ﬁrnctions

Despite the substantial body of evidence supporting AhR involvement in numerous
HAH-induced effects there are some notable exceptions which indicate that HAHS may act
through other mechanisms. One particular dioxin congener which lacks AhR binding affinity,
2,7-dichlorodibenzo-p-dioxin (Poland and Glover, 1976), suppresses the AFC response of
mouse splenocytes both in viva (Holsapple et al., l986a) and in vitro (Holsapple et al.,
l986b). Unlike TCDD, the immunosuppressive effects of 2,7-dichlorodibenzo-p-dioxin are
not accompanied by elevated levels of hepatic P4501Al. Other dichlorinated dioxin
congeners that have low AhR binding affinity, such as 2,8—dichlorodibenzo-p-dioxin, do not
suppress the AFC response (Tucker et al., 1986). Thus, 2,7-dichlorodibenzo-p-dioxin appears
to act through a unique mechanism which does not require AhR binding in order to suppress
B-cell immunity.

Results from studies with high and low AhR responsive mouse strains also suggest
that some immunosuppressive effects of HAHS may be mediated by AhR-independent
mechanisms. As previously mentioned, the immunosuppressive effects of HAHS have been
shown segregate with the AhR alleles. However, Morris and co-workers (1992) have
demonstrated that the exposure regime can greatly inﬂuence the responsiveness of low AhR-
responsive mice. DBA/2 mice that received subchronic doses of TCDD exhibited a 10-fold
enhancement in humoral immune suppression compared to DBA/2 mice that received the
same cumulative dose of TCDD but in an acute exposure. In addition, the severity of

immunosuppression in subchronically exposed DBA/Z mice was comparable to the

2l

suppression observed in B6C3F1 (AhR responsive) mice. These ﬁndings are supported by
results from in vitro exposures in which TCDD was equally effective at suppressing the AFC
response in splenocytes from both high and low AhR responsive mouse strains (Holsapple et
al., 1986b). The mechanism by which HAHS induce immunotoxic effects, independent of the
AhR, is believed to involve modulation of intracellular Ca2+ (Holsapple et al., 1991a). Taken
together, these ﬁndings indicate that several factors can modulate the immunosuppressive
activity of HAHS and that AhR involvement may be critical for many but not all toxic
responses.

The role of the AhR in HAH-induced immunodepression in ﬁsh is not well
understood. Appreciable amounts of AhR have only recently been detected in ﬁsh cells (20
fmol/mg protein; Lorenzen and Okey. 1990). However, cytochrome P4501AI induction has
been measured in ﬁsh liver, kidney, and gill (Miller et al., 1989; Goksoyr and Forlin, 1992).
Structure-activity relationships in ﬁsh for P4501Al induction (Janz and Metcalfe, 1991;
Newsted et al., 1995) and early life stage mortality (Walker and Peterson, 1991) suggest that
these effects are mediated through the AhR. Although AhR agonists have been shown to
affect various immune responses in ﬁsh, as discussed previously, there is insufﬁcient
information at present to determine whether these effects on dependent on AhR mediated
processes.

Several approaches have been used to identify cellular targets in HAH-induced
immunotoxicity. Results from in vitro and ex vivo recombination studies with inbred mice
indicate that suppression of the AFC response by TCDD is due to an alteration in the ﬁrnction
of B-cells, and not T-cells nor antigen presenting cells (Dooley and Holsapple, 1988). TCDD

has been shown to directly affect B-lymphocyte differentiation under in vitro conditions

22

(Tucker et al., 1986; Luster et al., 1988). However, T-cells appear to be more sensitive than
B-cells when the effects of dioxins on the AFC response are tested in vivo (Kerkvliet and
Brauner, 1987). This conclusion is based on the ﬁnding that mice immunized with T-D
antigens are more sensitive to the suppressive effects of dioxin (1,2,3,4,6,7,8-
heptachlorodibenzo-p-dioxin; HpCDD) than mice immunized with T-I antigens. Because the
AFC response to T-D antigens requires greater T-cell involvement, the logical explanation
for the antigen-dependent sensitivity to HpCDD is impaired T-cell function. It is not clear
why differences in dosing and immunization schemes would result in differential sensitivity
of B and T-cells, although Kerkvliet and Burleson (1994) suggested that dioxins might affect
activated T-cells, in vivo, through indirect mechanisms. Indirect effects are known.
Depletion of thymocytes associated with TCDD-induced thymic atrophy is believed to occur
indirectly through cell-cell contact with TCDD-affected thymic epithelial cells (Greenlee et
al,1985)

In ﬁsh, it seems B-cells are a target of HAH-induced depression of the primary AFC
response because signiﬁcant effects have been demonstrated with T-I antigens. Surprisingly,
responses to T-D antigens are less affected. Perhaps stimulation provided by T lymphocytes
in some way protects ﬁsh B-cells from the modulatory effects of HAHS. If this is so, then the
activation of naive T-cells would have to be less affected by HAHS. Lower T-cell sensitivity
may be inferred from the study by Spitsbergen et al. (1986). TCDD treatment depressed the
proliferative response of rainbow trout splenocytes to poke weed mitogen, a stimulator of B
and T lymphocytes, but did not signiﬁcantly affect the response to Con A (a mammalian T-

cell nritogen).

23

The heightened sensitivity of the secondary AFC response to T-D antigens observed
in PCB-treated chinook salmon, indicates that T-cell mediated events may be affected in the
memory response. Arkoosh et a1. (l994a) suggest that if ﬁsh have a requirement for memory
T-cells similar to that of mammals then PCBS may affect the transition of naive T-cells to
memory T-cells. Such an effect would reduce the pool of memory cells available to
participate in the secondary AFC response. Further progress in identifying the mechanisms
of HAH immunotoxicity will undoubtedly require both in vivo and in vitro approaches given

the complexity of immune responses and the multiplicity of HAH-associated effects.

CONCLUSIONS

HAHS can disrupt normal immune ﬁrnctions in ﬁsh, but some species are more
severely affected than others. For example, rainbow trout, one of the more thoroughly
studied species, seems to be less sensitive than chinook salmon or channel catﬁsh. Humoral
immunity, particularly the secondary AFC response, is one of the more sensitive indicators
of HAH immunotoxicity. Nonspeciﬁc and cell-mediated responses have not been as
thoroughly investigated, although some effects have been reported. Histological lesions in
lymphoid tissues, similar to those described in mammals, have been observed in HAH-treated
ﬁsh but the incidence and severity of these lesions has not always coincided with impaired
immune ﬁrnction. Immunodepression has been reported in wild ﬁsh inhabiting areas
contaminated with HAHS and other organic and inorganic pollutants. However, a better
understanding of the mechanisms underlying HAH-induced immunomodulation and of the
sensitivity of ﬁsh species in aquatic communities will be required to more accurately assess

the risk posed by environmental exposure to HAHs.

24

FUTURE CONSIDERATIONS

One of the major limitations in identifying sensitive immunological endpoints of HAH
immunotoxicity has been the ﬁsh-to-ﬁsh variability often encountered in measuring immune
responses. In some studies the coefficient of variation (a measure of within-group variability)
has far exceeded 50% (Spitsbergen et al., 1986; Thuvander et al., 1993). This tremendous
variation increases the probability of type II error (i. e., accepting the null hypothesis when in
fact there were real differences). Mammalian immunotoxicologists have the advantage of
working with inbred and syngeneic strains of animals that respond more consistently. This
permits greater sensitivity in detecting subtle differences. Inbred ﬁsh strains are being
developed (Komen et al., 1990) and this will improve the sensitivity of these studies.
Alternatively, in vitro techniques using tissue sections (Anderson, 1992) or primary cell
cultures (Noguchi et al., 1994, 1996) from an individual ﬁsh will allow the effects to be
measured in genetically identical cell populations. In vitro approaches are valuable for
studying mechanisms of action, assessing the intrinsic sensitivity of individual ﬁsh, and to help
identify factors that may account for variability in immune responses between ﬁsh.

HAHS and other contaminants represent only one of the many environmental factors
that may affect the immune status of wild ﬁsh. Identiﬁcation of HAH-speciﬁc immunological
perturbations (perhaps effects on the secondary AFC response) may help to distinguish
chemical-induced effects from other contributing factors such as nutrition (Blazer, I992),
temperature (Clem et al., 1992), or season (Zapata et al., 1992). Currently, it is necessary to
employ a battery of immunological and other tests (e.g. enzyme induction) to generate a

proﬁle of immunomodulation that characterizes a chemical etiology.

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CHAPTER 2

In Vitro Assessment of Immunomodulatory Effects of Polychlorinated diaromatic
hydrocarbons (PCDH) on Chinook Salmon, Oncorhynchus tshawytscha:

Optimization of Lymphocyte Proliferation Assays.

(For Submission to Fish and Shellﬁsh Immunology)

INTRODUCTION

Recent declines in the chinook salmon, Oncorhynchus tshawytscha, population in
Lake Nﬁchigan have been attributed to massive die-offs of ﬁsh infected with Renibacterium
salmoninarum (Smith, 1992), the causative agent of bacterial kidney disease (BKD).
Although BKD appears to be the ultimate cause of death, other factors are thought to have
played a role in rendering salmon susceptible to this otherwise ubiquitous bacteria. Organic
contaminants are among the group environmental factors which have been considered as
possibly contributing to the salmon mortality. Detectable levels of contaminants including,
dioxins and dibenzofurans (Smith et al., 1990) and polychlorinated biphenyls (PCB)
(Masnado, 1987; Smith et al., 1990; Williams et al., 1992) have been measured in chinook
salmon collected from Lake Michigan waters. These polychlorinated diaromatic

hydrocarbons (PCDHs) are potent mammalian immunotoxins (Vos and Luster, 1989) which

35

36

reduce host resistance to infection and disease. The potency of these compounds at
suppressing immune responses in ﬁsh is less deﬁnitive because of the limited number of
studies reported and apparent differences in sensitivity among species (Noguchi, 1997). An
in Vitro approach was used to assess the sensitivity of chinook salmon to the
immunomodulatory effects of PCDHs. These results were used to conduct a hazard
assessment of the potential contribution of these compounds to suppressed immune responses
in feral salmon of the Great Lakes and potential contribution to the susceptibility of these ﬁsh
to epizootics.

In vitro assays have been effective for testing the effects of drugs (Siwicki et al., 1990)
and chemical contaminants (Anderson, 1992; Voccia et al., 1994) on ﬁmctional immune
responses in ﬁsh. Lymphocyte proliferation is one of the early events in adaptive immunity
that can be measured in Vitro by stimulating primary lymphocyte cultures with B- and T- cell
mitogens. The responsiveness of lymphocytes to mitogen stimulation is a general measure
of the organisms ability to mount a humoral or cell-mediated immune response. Concomitant
exposure of cell cultures to mitogens and immunomodulating chemicals provides a model for
assessing direct effects on lymphocyte proliferation. In order for such assays to be responsive
to toxicant-related effects and not attenuated by other factors, it is critical that assay
conditions which support maximal responsiveness are determined. Previous attempts to
measure mitogen-induced proliferation of salmonid lymphocytes have indicated that the type
of mitogen, mitogen concentration, incubation time, and source and quantity of plasma used
to supplement culture media, can greatly affect the proliferative response (DeKonning and

Kaattari, 1991; Leith et al., 1987; Tillitt et al., 1988).

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37

The objective of this study was to deﬁne in vitro assay conditions for chinook salmon
lymphocytes that support maximal mitogen-induced proliferation. Responsiveness of spleen
and anterior kidney lymphocytes to mitogen stimulation was compared. The effectiveness
of different media supplements (autologous and homologous salmon plasma and calf serum)
and optimal mitogen concentrations and incubation times were also evaluated. In addition,
factors that likely contribute to the ﬁsh-to-ﬁsh variability in mitogen responsiveness were

identiﬁed.

MATERIALS AND METHODS

Fish Holding and Sampling

Juvenile chinook salmon, Oncorhynchus tshawytscha, were obtained from the
Michigan Department of Natural Resources Wolf Lake Hatchery, Mattawan, MI and held
at the National Biological Service, Great Lakes Science Center in Ann Arbor, MI. Salmon
were maintained in 12° C well water and fed Biodiet, commercial salmon diet. All assays
were performed using lymphoid tissues from sexually immature 1-2 year old chinook salmon.
On sampling days, ﬁsh were quickly netted and sacriﬁced by immersion in a lethal
concentation of anesthetic solution (200 mg MS-222 L“). Blood and lymphoid tissues were
harvested shortly thereafter. Blood was sampled from the caudal vein by severing the caudal
peduncle and was collected in heparinzed vacutainer tubes. Lymphoid tissues (spleen,
anterior kidney and thymus) were removed aseptically and placed in tubes containing ice cold,

RPM-1640. Blood and tissue samples were held on ice until ﬁthher processing.

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38

Reagents

Culture medium (RPMI-1640), antibiotics (gentamycin sulfate), mitogens
(lipopolysaccharide; [LPS] and phytohemagglutinin; [PHA]) and 2-mercaptoethanol (2-Me)
were purchased from Sigma, St. Louis, MO. 3H-thymidine was purchased from Amersham,
Arlington Heights, IL and provided by the Tissue Typing Laboratory at Michigan State
University. The Warr 1.14 monoclonal antibody used for ﬂow cytometry was obtained from
the National Marine Fisheries Service, Newport, Oregon. FITC-conjugated goat anti-mouse

antibody was purchased from Kirkigard and Perry, Gaithersburg, MD.

Cell Cultures

Pronephric and splenic cell suspensions were prepared by repeatedly expelling tissues
from a one mL syringe (Arkoosh and Kaattari, 1992). Tubes were placed on ice for 20 min.
to allow connective tissue particles to settle. The cell suspensions (supematants) were then
washed twice in RPMI-1640. After the ﬁnal wash, the cells were resuspended in tissue
culture medium (TCM). The basic TCM contained RPM-1640 with L-glutamine, NaHCO3
(0.2%), gentamycin sulfate (50 ug mL"), 2-Me (50 pM) and pH adjusted to 7.2. Various
TCM supplements, including autologous or homologous salmon plasma or newborn calf
serum, were assessed for their ability to sustain cell cultures and support mitogen
proliferation. Salmon plasma was prepared by centrifuging heparinized blood samples (500x
g at 4° C) for 10 min. Plasma from individual ﬁsh was heat-treated at 56° C for 30 min (to
inactivate complement), diluted to 20% with RPMI, ﬁlter sterilized and stored at -20° C.
Lymphocytes were enumerated by microscopic examination in a hemocytometer and viability

was determined by trypan blue exclusion. The volume of cell suspension was adjusted to

39

achieve a ﬁnal cell concentration of 5 x 106 cells mL".

Lymphocyte Proliferation Assay

The proliferation assay was adapted from methods described by Leith et al. (1987) and
DeKoning and Kaattari (1991). Ten uL of mitogen solution (LPS or PHA) was dispensed
into appropriate wells of a round bottom 96 well microtiter plate (Limbro) followed by 100
ul of cell suspension. Four wells were assayed per treatment unless otherwise stated. The
plates were covered with mylar plate sealers, placed in an sealed gas chamber that was purged
with a 10% C02, 80% N2, 10%02 gas mixture, and incubated at 17° C. Mitogen
concentrations and incubation times were varied in order to determine conditions for maximal
proliferation. 3H-thymidine was added to each culture well (1.0 uCi per well) 20-24 hr before
cell were harvested. The incorporation of 3H-thymidine is a measure of mitotic activity. Cells
were harvested using an automated cell harvester (Skatron) and the 3H activity was measured
by liquid scintillation and reported as counts per minute (CPM) as well as stimulation index
(SI; equation 1):

S‘] : cpm mitogen—stimulated culture (I)

 

cpm control (unstimulated) culture

Flow Cytometry

Splenocyte cultures were characterized by ﬂow cytometry by use of a modiﬁcation
of the method described by Thuvander et al. (1992). Cell suspensions were prepared as
described previously and Ig’“ cells (B-cells) were labeled by incubating one mL of cell

suspension for 45 min with Warr 1.14 monoclonal antibody (1 : 1000 dilution). The cells were

40

then wash twice and incubated for 45 min with an FITC conjugated secondary antibody (goat
anti-mouse @ 5 pg mL). Cells were washed twice then resuspended in one mL RPMI-1640
a 1 x 10° cells mL". Splenocyte subpopulations were distinguished based on forward and side
scatter (a measure of cell surface roughness and granularity) and FITC ﬂuorescence in a
Becton Dickinson FACS-III flow cytometer. Cell suspensions were also prepared from
salmon thymus in order to check the speciﬁcity of the Warr 1.14 monoclonal antibody for
the Ig ligand on salmon lymphocytes and to aid in identifying the lymphocyte region of the

cytograms.

Statistics

All statistical tests were performed using SYSTAT statistical software (Version 5.1).
SYSTAT Inc., Evanston, IL. Effects of plasma concentration, mitogen concentration and
incubation time were analyzed by one-way analysis of variance (AN OVA) on log transformed
data. The effect of heterologous plasma source on lymphocyte proliferation was analyzed
using the Kruskal-Wallis test. The paired sample T-test was used to analyze the effect of
autologous versus homologous plasma on proliferation and Bartletts test was used to test for

homogeneity.

RESULTS
Plasma Concentration
Plasma concentration signiﬁcantly inﬂuenced (P<0.001) mitogen-induced proliferation
of salmon lymphocytes (Figure 4). Maximal proliferation of splenic lymphocytes, incubated

for 5 d with 100 ug LPS mL", was observed in cultures supplemented with 1% salmon

41

plasma. Increasing plasma concentration from 1 to 10% resulted in a dose-related decrease
in proliferation. The proliferative response was abrogated in cultures incubated with 10%
plasma. Initial assays using splenocyte cultures from three salmon indicated that 1%
autologous plasma supported greater proliferation (81:48.0 :h 4.2) than culture medium

supplemented with 5% newborn calf serum (SI=12.3 i 4.0).

Mitogen Concentration

Mitogen titer was detennined for both spleen and anterior kidney lymphocytes (Figure
5). Cultures were incubated for ﬁve days in TCM supplemented with 1% autologous plasma
and stimulated with 12.5, 25, 50 and 100 pg LPS mL". Mitogen responsiveness was
signiﬁcantly affected by LPS concentration (P<0.001) with lymphocytes from both tissues
exhibiting maximal proliferation at 100 pg LPS mL“. Doubling the LPS concentration from
100 to 200 pg mL'l did not enhance the response (data not shown). Splenic lymphocyte
cultures consistently generated higher maximal responses than cultures prepared from

pronephric lymphocytes.

Incubation Time

Kinetics of the proliferative response were studied by measuring 3H-thymidine
incorporation in splenic and pronephric cell cultures incubated for 5, 7, 9, 11 and 13 d and
stimulated with 100 pg LPS mL" (Figure 6). Incubation time signiﬁcantly affected 3H-
thymidine uptake in mitogen-stimulated cultures (P<0.001). The greatest stimulation occurred
after 5 d, and the response decreased with continued incubation. Results from earlier tests

indicated that 3H-thymidine incorporation in cultures incubated for shorter periods (3 d) was

42

less than after 5 d.

Response to PHA

Proliferative responses to the mammalian T-cell mitogen PHA were less than the
response to LPS and were also more variable. Maximum stimulation of splenic lymphocytes
ranged from 2.0 - 28.9 (SI) in cultures treated for 5 d with 2.5 - 5 pg PHA mL". Anterior
kidney lymphocytes were less responsive to PHA. Maximum stimulation ranged from 0.8 to
6.26 with appreciable stimulation occurring in only in 2 of 5 ﬁsh (SI: 3.53 and 6.26 at 1.25
pg PHA mL'l ). Relatively great background proliferation in pronephric lymphocyte cultures

(567 - 5752 cpm) contributed to the small 515 observed.

F ish-to-Fish Variability in Proliferative Response to LPS

Splenic lymphocytes from 10 individuals were cultured with optimal concentrations
of autologous plasma (1%) and LPS (100 pg mL'l ) and incubated for the optimal time period
(5 d). Even under these optimal conditions there was considerable varibility among
individuals (CV 86%) with the $15 for responsive ﬁsh (8 of 10) ranging from 9.9 to 100.1
(Table 3). Two factors were identiﬁed that may contribute to the variability in mitogen
responsiveness among individual ﬁsh; the source of plasma and the proportion of Ig‘
lymphocytes (B-cells). To test for effects of plasma source, splenocytes were cultured with
plasma from 12 different salmon and stimulated with LPS (100 pg mL'l ). The proliferative
response was signiﬁcantly affected (P<0.05) by the source of plasma with $15 ranging from
2 to 220 (Figure 7). These results indicate that soluble factors in salmon plasma can modulate

responsiveness of lymphocytes and that concentrations of these factors can vary greatly

43

among ﬁsh. In order to reduce variability due to plasma-associated factors, plasma from
several salmon were pooled and this homologous plasma was used to supplement the culture
medium. The proliferative responsivness of splenocytes cultured in homologous plasma
exhibited signiﬁcantly less (P<0.05) ﬁsh-to-ﬁsh variability (CV 39%) compared to the same
cells cultured in autologous plasma (CV 120%; Table 4). In addition, lymphocyte
proliferation was signiﬁcantly enhanced in cultures containing homologous plasma (P<0.05).
Use of homologous rather than autologous plasma standardizes culture conditions for in vitro

assays and reduces variability in the proliferative response of salmon lymphocytes.

Flow Cytometry

The proportion of Ig‘ lymphocytes (mature B-cells) in spleen cell suspensions from
6 individuals ranged from 31 to 62% with an average of 43% (Figure 8). Only 5.2% of
thymic lymphocytes were Ig‘, which indicates little nonspeciﬁc binding of the Warr 1.14
antibody. Differences in the abundance of lymphocyte subpopulations capable of responding
to a speciﬁc stimulant will likely contribute to ﬁsh-to-ﬁsh variability in lympho-proliferative

responses.

DISCUSSION
The proliferative response of chinook salmon lymphocytes was sensitive to the
quantity and quality of plasma used to supplement culture media. Past attempts to induce
proliferation of trout and salmon lymphocytes with mitogens have employed a variety of
plasma or serum supplements (Leigh et al., 1987; Sptisbergen et al., 1986; Thuvander et al.,

1993; Tillitt et al., 1988; Yui and Kaattari, 1987). In studies which have used fetal calf serum

44

(FCS) or human serum to supplement the culture medium the proliferative response has been
relatively small. The use of autologous or homologous trout plasma has been shown to be
superior to FCS for supporting maximal mitogen responsiveness in rainbow trout (DeKonning
and Kaattari, 1991). Similarly, we have demonstrated that salmon plasma was effective at
supporting LPS-induce proliferation of chinook salmon lymphocytes. The level of stimulation
was greater than in cultures supplemented with newborn calf serum and also exceeded levels
previously reported for assays using FCS as a supplement (Leigh et al., 1987; Yui and
Kaattari, 1987). Interestingly, the optimal plasma concentration for proliferation of chinook
salmon lymphocytes (1%) was less than that for rainbow trout (10%; DeKonning and
Kaattari, 1991). In fact, lymphocyte proliferation was completely inhibited in salmon
splenocyte cultures containing 10% salmon plasma. These differences support
recommendations by DeKonning and Kaattari (1992) that assay conditions should be
optimized on a species-speciﬁc basis in order to ensure maximal responsiveness.

The chinook salmon spleen and anterior kidney harbored populations of lymphocytes
that were mitogen-responsive, however, the proliferative response of splenic lymphocytes was
consistently greater and less variable. Differences in mitogen responsiveness may be related
to differences in cell populations. As the principal hemopoetic tissue, the anterior kidney is
populated with a diversity of cells in terms of both lineage and stage of development (Zapata,
1979). Whereas, the spleen contains primarily mature lymphocytes and erythrocytes.
Previous studies have demonstrated that the salmonid spleen possesses B-lymphocytes with
greater afﬁnity for antigen than the B-lymphocytes found in the anterior kidney (Irwin and
Kaattari, 1986). Such lymphocyte heterogeneity indicates that organ-speciﬁc differences in

lymphocyte subpopulations may be responsible for differential antigen-afﬁnity and

45

responsiveness. Thus, greater proportions of mature, antigen-responsive, lymphocytes in the
spleen compared to the anterior kidney, could have contributed to the greater proliferation
observed in splenocyte cultures.

The responsiveness of splenic and pronephric lymphocytes to PHA was less and more
variable than the response to LPS. Previous studies have also indicated that chinook salmon
lymphocytes exhibit little PHA responsivenes compared to LPS and other B-cell mitogens
(Leigh et al., 1987; Yui and Kaattari, 1987). This would suggest either lesser frequency and
greater variability in the number of PHA-responsive lymphocytes or that assay conditions that
are optimal for LPS responsiveness may not be optimal for supporting PHA-induced
proliferation. Under the assay conditions described here, LPS would be superior to PHA for
testing effects of immunomodulating toxins in vitro, because of the greater and more
consistent responses.

The variability in mitogen responsiveness between individual ﬁsh was reduced by
culturing cells in media supplemented with pooled homologous plasma rather than autologous
salmon plasma. The use of pooled plasma helps to standardize culture conditions by
eliminating the potential effect of varying concentrations of immunomodulating factors, such
as cortisol or reproductive steroid hormones, that may occur in autologous plasma from
diﬁ‘erent ﬁsh. Because all of the salmon used in this study were sexually immature, it is less
likely that salmon plasma contained immunomodulating concentrations of reproductive
hormones such as testosterone or estrogen. Although attempts were made to minimize
handling stress it is possible that differences in the physiological responses of individual ﬁsh
to capture resulted in varying concentrations of cortisol or other stress related hormones in

the blood. Stress-induced elevation in plasma cortisol levels and associated

46

immunosuppressive effects is a well documented phenomena in mammals (Pruett et al., 1993)
and ﬁsh (Barton and Iwama, 1991). Addition of physiologically relevant concentrations of
cortisol (100 ng mL" ) to lymphocyte cultures, in Vitro, has been shown to directly suppress
mitogen responsiveness (Tripp et al., 1987). Differences in plasma cortisol may have
contributed to the ﬁsh-to-ﬁsh variability in mitogen responsive of lymphocytes cultured in
autologous plasma (Table 4) as well as the variability in responsiveness among lymphocyte
cultures supplemented with plasma from different ﬁsh (Figure 7). Further improvements in
culture conditions to reduce variability and enhance responsiveness might be achieved by
using carbon-stripped plasma to reduce the concentration of steroid hormones or by
developing a chemically-deﬁned plasma-free medium. Both of these approaches have been
effective in mammalian cell culture.

When cell suspensions used for proliferation assays are prepared and enumerated
using light microscopy it is not possible to distinguish lymphocyte subpopulations i.e. B-cells
and T-cells. By standardizing the number of lymphocytes used in each assay (5 x 105 cells per
well) it is assumed that the proportion of B- and T-cells remains constant. Thus, the number
of B-cells capable of responding to LPS, for example, would be the same for each ﬁsh that
is assayed. This assumption was tested by determining the proportion of Ig’ cells (mature B-
cells) in salmon spleen cell suspensions. The Warr 1.14 monoclonal antibody, previously
shown to cross react with free salmon Ig, was effective for detecting 51g on salmon
lymphocytes. The average proportion of lg+ lymphocytes in the salmon spleen (43%) was
similar to values reported for rainbow trout, 28 - 35% (DeLuca et al., 1983; Thuvander et al.,
1990). However, there was considerable variability among individual ﬁsh for both salmon

(CV 26%) and rainbow trout (CV 26 -3 7%) with a two fold difference between the minimum

47

and maximum values in the present study. Such variability would likely contribute to
differences in mitogen responsiveness among ﬁsh which could obfuscate detection of
contaminant-induced changes in ﬁJnctional immune responses.

The use of in Vitro assays in immunotoxicity testing should enhance the ability to
detect toxicant-related effects by eliminating factors that contribute to between-ﬁsh variability
in immune responses. All treatments such as multiple compounds or multiple concentrations
of a single compound could be tested with cell cultures from an individual ﬁsh. Thus,
treatments would be applied to cultures that are genetically identical, have the same
phenotypic composition and the same physiological history. Differences in immune responses

could, therefore, be more directly associated with the action of immunomodulating chemicals.

48

Table 3. Proliferative response of splenocytes from 10 salmon assayed under optimal
mitogen concentration (100 pg LPS mL") and incubation time (ﬁve days).

 

 

 

Stimulation
Salmon Index
1 37
2 14
3 l
4 100
5 1
6 63
7 71
8 63
9 3 5
10 10
Mean 39
SE 1 1

CV 86%

 

49

Table 4. Effectiveness of autologous and pooled homologous salmon plasma at supporting
splenocyte proliferation. Cultures were stimulated with LPS (100 pg mL") and
incubated for ﬁve days. Results from assays using splenocytes from ﬁve individual
salmon (mean i SE).

 

Stimulation Index

 

 

 

Autologous Homologous
Salmon Plasma Plasma
l 1 a: 0.05 323 :t 35
2 34 :i: 17 204 i 18
3 1 :t 0.01 90 :l: 11
4 1 14 i 10 235 :1: 5
5 182 :1: 24 256 :1: 9
Mean 66 222
SE 36 38

CV l 20% 39%

 

Ellie ,
Slimm 1
(halchf
bars.

 

50

 

125000

100000

75000

CPM

50000

25000

 

 

 

 

1 2 5 1o 1 2 5 1o
% Plasma in TCM

Figure 4. Inﬂuence of plasma concentration on the proliferative response (mean + SE) of
splenocytes from an individual ﬁsh (ﬁlled bars) and pooled splenocytes from three ﬁsh
(hatched bars) to LPS. Background proliferation in control wells is indicated by the open
bars.

5l

 

40000

30000

PM

0 20000

1 0000

 

 

 

0 I l l A

0 25 50 75 100 125

LPS (ug/mL TCM)

Figure 5. Effect of mitogen concentration on the proliferative response (mean t SE) of spleen
(closed circles) anterior kidney lymphocytes (open circles) to LPS.

52

 

80
70 r l
60 "
50 ‘
40 l .

30 r L
20 “
10 ‘

SI

 

 

 

3 5 7 9 11 13 15

Incubation Time (days)

Figure 6. Kinetics of LPS-induced lymphocyte proliferation (mean :1: SE) of spleen (closed
circles) and anterior kidney lymphocytes (open circles).

53

 

400

300 ”

200 ”

SI

100 "

 

 

 

1L
ii ii

NP P1 P2 P3 P4 P5 P6 P7 P8 P11 P12 P13 P14

Source of Plasma

Figure 7. Effect of heterologous salmon plasma on splenocyte proliferation. Splenocytes
from an individual salmon (#14) were cultured with TCM supplemented with plasma from 1 1
different salmon, autologous plasma (P14), no plasma (NP), and stimulated with LPS (100
ug mL“).

54

Figure 8. Characterization of splenocyte subpopulations using ﬂow cytometry. A) cytogram
(forward scatter vs. side scatter) of chinook salmon spleen cell suspensions. B) FITC
ﬂuorescence histograms for four gated regions from the splenocyte cytogram. Mature B-cells
(Ig’) are distinguished from Ig' cells by FITC ﬂuorescence, as seen in region 2 (R2).

55

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CHAPTER 3

In vitro effects of Polychlorinated Diaromatic Hydrocarbons on Lympho-

Proliferative Immune Responses in Chinook Salmon (Onchorhynchus tschrqytscha).

(for submission to Environmental Toxicology and Chemistry).

INTRODUCTION

Concern over the recent epizootic outbreak of bacterial kidney disease (Smith, 1992)
in Lake Michigan chinook salmon (Oncorhynchus tshawytscha) has raised questions about
the role that chemical contaminants may play in modulating immunity and disease resistance
in ﬁsh. Polychlorinated diaromatic hydrocarbons (PCDHs), including polychlorinated
biphenyls (PCBS), dioxins, and dibenzoﬁirans, are a class of anthropogenic compounds that
have been detected in ﬁsh throughout the Great Lakes (Masnado, 1987; Zacharewski et al.,
1989; Smith et al., 1990; Williams et al., 1992). Some PCDHs are potent mammalian
immunotoxins that reduce host resistance to infection and disease (Vos and Luster, 1989).
The potency of some PCDHs in suppressing immune responses in ﬁsh is less deﬁnitive.
Altered immune responses have been observed in some studies (Spitsbergen et al., 1986;
Thuvander et al., 1991, 1993; Rice and Schlenk, 1995) but not in others (Cleland and

Sonstegard, 1987; Cleland et al., 1988a). However, recent evidence suggests that humoral

56

57

(antibody producing) immune responses in chinook salmon may be sensitive to some PCDHs.
Salmon inhabiting an urban estuary in Puget Sound, contaminated with PCDHs and other
anthropogenic compounds, had depressed secondary humoral immune responses (Arkoosh
et al., 1991). Under controlled laboratory conditions studies showed that primary and
secondary humoral immune responses were suppressed in chinook salmon treated with
Aroclor 1254, a commercial polychlorinated biphenyl (PCB) mixture (Arkoosh et al., 1994a).

This study was designed to answer some fundamental questions regarding the
sensitivity of chinook salmon to PCDHs by measuring direct in vitro effects on immune
ﬁmctions. In vitro assays have been used to identify cellular targets of PCDH immunotoxicity
in mammals (Dooley and Holsapple, 1988 ), characterize structure activity relationships for
PCDH immunotoxicity (Harper et al., 1995), and elucidate mechanisms of action (Holsapple
et al., 1991). In vitro assays have also been effective for testing the effects of drugs (Siwicki
et al., 1990) and chemical contaminants (Anderson, 1992; Voccia et al., 1994) on functional
immune responses in ﬁsh. Effects measured in vitro may not be predictive of whole animal
responses if the toxicant indirectly affects immune ﬁinction by modulating other physiological
systems, e.g. the neuroendorcrine system. However, in vitro assays offer several beneﬁts
including the ability to directly assess the sensitivity of target cells and immune processes.
Within and among-ﬁsh variability in immune response can be estimated by studying
subsamples of cells from individual ﬁsh. In this way the ﬁsh serves as its own control, thereby
improving detection. In vitro tests also eliminate the need for containment and exposure
systems that are necessary for in vivo studies, and require fewer ﬁsh. In addition, cellular or
tissue-based effect concentrations can be predicted if toxicant concentrations in cells are

measured.

58

Lymphocyte proliferation is one of the early events in adaptive immunity that can be
measured in vitro by stimulating primary lymphocyte cultures with mitogens (Leith et al.,
1987; Tillitt et al., 1988; DeKonning and Kaattari, 1991; Noguchi et al., 1997a). The
responsiveness of lymphocytes to B-cell mitogens is a general measure of the organism’s
ability to mount a humoral immune response. Concomitant exposure of cell cultures to
mitogens and immunomodulating chemicals provides a model for assessing direct effects on
lymphocyte proliferation. The objectives of this study were to determine whether the
proliferative responses of chinook salmon lymphocytes were affected by direct exposure to
PCDHs (in vitro), to determine the immunotoxic potency of PCBs relative to 2,3,7,8-
tetrachlorodibenzo-p-dioxin (TCDD), and to determine whether results from in vitro tests are

consistent with, or predictive of, effects on whole ﬁsh.

MATERIALS AND METHODS
Fish Holding and Sampling
Juvenile chinook salmon were obtained from the Michigan Department of Natural
Resources Wolf Lake Hatchery, MI, and held at the National Biological Service, Great Lakes
Science Center in Ann Arbor, Mattawan, MI. Fish holding and sampling information have
been described previously (Noguchi et al., 1997a). All assays were performed using lymphoid

tissues from juvenile (220 - 420 mm) salmon.

Reagents and Chemicals
TCDD and PCB congener 126 (3,3',4,4',5-pentachlorobiphenyl) were purchased from

Accurate Chemical, New Haven, CT. Aroclor 1254 was obtained from Monsanto, St Louis,

59

MO. Culture medium (RPMI-1640), antibiotics (gentamycin sulfate), mitogens
(lipopolysaccharide-LPS), 2-mercaptoethanol (2-Me) and dimethylsulfoxide (DMSO) were
purchased from Sigma Chemical, St. Louis, MO. 3H-thymidine was purchased from
Amersham, Arlington Heights, IL and provided by the Tissue Typing Laboratory at Michigan
State University. The Warr 1.14 monoclonal antibody used for ﬂow cytometry was obtained
from the National Marine Fisheries Service, Newport, OR. FITC-conjugated goat anti-mouse
antibody was purchased from Kirkigard and Perry, Gaithersburg, MD. Stock solutions of
TCDD and PCBs were prepared in DMSO at concentrations 1000x greater than the nominal

dosing concentration.

Cell Cultures

Splenic cell suspensions from individual salmon were prepared as described by
Noguchi et al. (1997a). Brieﬂy, the spleens were disrupted by repeated expulsion from a one
mL syringe and the resulting suspension was held on ice for 20 min to allow connective tissue
particles to settle. The suspended splenocytes were washed twice in RPMI-1640 and
resuspended in tissue culture medium (TCM). The TCM contained RPMI-1640 with L-
glutamine, NaHCO3 (0.2%), gentamycin sulfate (50 pg mL“), 2-Me (50 pM), 1% pooled
homologous salmon plasma and pH adjusted to 7.2. Lymphocytes were enumerated by
microscopic examination in a hemocytometer and viability was determined by trypan blue
exclusion. The volume of cell suspension was adjusted to achieve a ﬁnal cell concentration

of5 x 10° cells mL".

60

Lymphocyte Proliferation Assay

The proliferation assay optimized for chinook salmon lymphocytes (Noguchi et al.,
1997a) was used to test for effects of PCDHs on in vitro mitogen responsiveness. One mL
aloquots of cell suspension were dosed with l pL of PCDH or vehicle (DMSO). The
suspension was mixed by aspirating with a repeat-dispensing pipetter. One hundred pl. of cell
suspension was dispensed into appropriate wells of a round-bottom 96-well microtiter plate
(Limbro) containing 10 pL of mitogen solution (100 ug LPS mL“). All treatments were
assayed in quadruplicate. Negative controls (cell suspension containing no LPS) and positive
controls (cell suspensions containing LPS and no PCDH or vehicle) were assayed along with
the treated cultures. With this dosing scheme the difference in proliferative response between
replicate wells is a measure of the variability in responsiveness due to natural variability in
splenocyte sensitivity and methodological factors including differences in PCDH uptake and
culture conditions between wells; it does not include variability associated with dose
administration, 1'. e. pipetting error. The plates were covered with mylar plate sealers, placed
in a sealed gas chamber that was purged with a 10% C02, 80% N2, 10%O2 gas mixture, and
incubated for four days at 17° C. 3H-thymidine was added to each culture well (1.0 pCi per
well) and incubated for an additional 20-24 hr before cells were harvested. Cells were
harvested and the and 3H activity was measured by liquid scintilation and reported as counts

per minute (CPM).

Cytotoxicity
Frank cytotoxicity was determined after incubating splenocyte cultures with PCDHs

for 24 hr. Cell cultures were prepared as described previously and dosed at the same

61

concentrations used for the proliferation assay. The number of splenocytes surviving after 24
hr was enumerated microscopically using a hemocytometer and vibility was determined by
trypan blue exclusion. Cytotoxicity was determined by comparing the number of viable
splenocytes in PCDH-treated cultures with untreated controls. Treatments were administered

to quadruplicate cultures.

Flow Cytometry

Salmon spleocyte cultures were characterized by ﬂow cytometry as described
priviously (Noguchi et al. 1977a). Ig’ cells (B-cells) were detected by labelling with a
primary, anti-ﬁsh Ig antibody (Wart 1.14 monoclonal antibody) and a FITC conjugated
secondary antibody. Splenocyte subpopulations were distinguished based on forward and side

scatter and FITC ﬂuorescence by a Becton Dickinson FACS-III ﬂow cytometer.

Cellular Uptake of PCBs

The concentration of PCB-126 accumulated by cell cultures in vitro was estimated
by measuring the concentration in the cell pellet. A cell suspension was prepared from salmon
spleen and anterior kidney cells and adqusted to 5 x 106 cells mL". Two-2.5 mL aliquots of
cell suspension were transferred to 10 mL culture tubes dosed with PCB-126, at a nominal
concentration of 100 nM. The cultures was incubated for 24 hr at 17° C then centriﬁiged at
500x g for 10 min. The supernatant was removed and the cell pellet was washed with fresh
culture medium. The cultures were centrifuged a second time and the culture medium was
removed and combined with the original supernatant. The supematants were extracted with

hexane:diethyl ether (1:1). The cell pellet was saponiﬁed with 2M alcoholic KOH and

62

extracted 3x with hexane:diethyl ether (1:1). Coextracted biogenic compounds, such as
lipids, were removed by silica gel column chromatography. Extracts were concentrated by
rotary evaporation and analyzed by GC-ECD. Cell culture dosing was performed in duplicate
and culture medium without cells was dosed as a spike control. The mass of the cell pellet
was estimated by multiplying the number of cells in the pellet by the cell volume and the

speciﬁc gravity which was assumed to be 1.

Statistics

All statistical tests were performed using SYSTAT statistical software (Version 5.1),
SYSTAT Inc., Evanston, IL. Effects of PCDHs on the proliferation of splenocytes from
individual ﬁsh were analyzed by one-way analysis of variance (ANOVA) and tukeys test was
used to make pairwise-comparison of treatment means. To evaluate the inﬂuence of ﬁsh-to-
ﬁsh variability on the ability to detect PCDH-related effects, treatment means for assays with
individual ﬁsh were combined (n=4). The combined data sets were analyzed by one-way
ANOVA or by ANOVA with ﬁsh as a blocking factor, in a randomized block design. The
proportion of variability explained by PCDH treatment and ﬁsh (blocking factor) was
calculated by dividing the appropriate sums-of-squares (sums-of squares “PCDH” or sums-of

squares “ﬁsh”) by the total sums-of—squares (Neter and Wasserman, 1974).

RESULTS
Sensitivity of Splenocytes to Direct Exposure to PCDHs
Direct exposure of salmon splenocytes to PCDHs resulted in a dose-related

suppression of LPS-induced proliferation. TCDD was immunosuppressive at concentrations

63

of 3 and 30 nM in splenocyte cultures prepared from four different salmon (Figure 9A).
Because the toxicant effect was compared to unexposed (control) cultures from the same ﬁsh
treatment effects were detectable despite relatively great differences in responsiveness among
ﬁsh (2000 - 95000 CPM). The effect of TCDD at lesser concentrations (0.3 and 0.03 nM)
was less certain. These concentrations tended to be suppressive in cultures that exhibited
moderate to great LPS—responsiveness, whereas, in less responsive cultures proliferation was
either not affected or was slightly enhanced. PCB-126 signiﬁcantly reduced splenocyte
proliferation at 1000 nM (Figure 9B). As with TCDD, the effectiveness of lesser
concentrations of PCB-126 (1 - 100 nM) was more variable. Aroclor 1254 at concentraitons
of 45 and 90 pM was suppressive, reducing splenocyte proliferation by >80% (Figure 9C).
However, at these great concentrations Aroclor 1254 was also cytotoxic to salmon
splenocytes (Figure 10). At non-cytotoxic concentrations of Aroclor 1254 (4.9 - 9 pM)
suppressive effects were not consistently observed.

The cytotoxic activity of Aroclor 1254 on lymphocyte subpopulations was assessed
using ﬂow cytometry. The abundance of lymphocytes in cultures dosed with 45 uM Aroclor
1254 was less than in controls (Figure 11). In addition, the reduction in the number of lg’
lymphocytes (B cells) was proportionately greater than for lg lymphocytes. These results
suggest that lymphocyte subpopulations may be differentially affected by PCBs and that B

cells may be more sensitive than Ig' lymphocytes.

Potency of PCDHs
The relative potency of TCDD, PCB-126 and Aroclor 1254 in modulating splenocyte

proliferation was estimated from dose-response curves. The effects of PCDH treatment,

64

expressed as a percentage of control CPM, from individual ﬁsh were averaged and plotted
against log PCDH concentration (Figure 12). Typically, the toxic potency of PCDHs relative
to TCDD is determined by comparing the concentrations at which the response is reduced to
50% of maximum. Because the maximum suppression by PCB-126 was less than 50%, the
relative potency was calculated for concentrations at which the response was suppressed by
30% (1C30). The 1C30, was calculated by linear interpolation from the dose-respnse curves.
The IC3O values for TCDD, PCB-126 and Aroclor 1254 were 1.5 nM, 1.3 pM and 18.7 pM,
respectively (Table 5). The potency of PCB-126 and Aroclor 1254 relative to TCDD was
expressed the toxic equivalency factor (TEF). TEFs are the ratio of the TCDD IC30 : PCDH
IC30 . Values >1 indicate that the compound is more toxic than TCDD and PCDHs that are
less toxic have TEFs <1. The calculated TEFs indicated that PCB 126 is 1,100 times and
Aroclor 1254 is 900,000 times less potent than TCDD at inhibiting salmon splenocyte

proliferation.

Distinguishing T oxicant from Non- T oxicant Effects

The principal advantage to using an in vitro approach to assess the immunotoxicity
of PCDHs was to increase the potential for detecting toxicant effects by eliminating the
inﬂuence of among-ﬁsh variability. Having demonstrated direct effects of PCDHs on
lymphocytes from individual ﬁsh, the next question was whether toxicant-related effects are
detectable when natural variability is not restricted, in other words, what proportion of the
total variability in immune response is attributable to the toxicant effect? By combining the
data for replicate ﬁsh, the proportion of variability attributed to the toxicant effect and that

which is due to among-ﬁsh variability was estimated. Data from four replicate ﬁsh were

65

combined and the effect of PCDH dose was assessed ﬁrst using one-way ANOVA and
subsequently by using a randomized block design ANOVA, with “Fish” as the blocking
factor. When the proliferative responses (CPM) from replicate ﬁsh were analyzed together
there appeared to be a dose-related depression in immune response, however, the effects were
not signiﬁcant for TCDD (p=0.891) or PCB-126 (p=0.822) because of relatively great
among-ﬁsh variability (Figure 13A&B; Table 6). The proportion of the total variability in the
proliferative response explained by these two PCDHs (sums-of-squares treatment/ total sums-
of-squares) was less than 10%. Thus, greater than 90% of the variability in the proliferative
response was attributed to unexplained variation (“error”; Table 6). However, when the data
for the Aroclor 1254-treated cultures are analyzed together, 55% of the variability was
toxicant related. This effect was statistically signiﬁcant (p=0.027) due to the great efficacy
(complete abrogation of the immune response) of Aroclor 1254 at the greatest doses tested
(30 pM). When the data were blocked according to “Fish”, signiﬁcant toxicant effects were
detected for TCDD, PCB-126 and Aroclor 1254 (Figure 13C; Table 7). Among-ﬁsh
variability accounted for greater than 80% of the total variability in the responses to TCDD
and PCB-126, thereby reducing the “error” to 5.4 and 4.8% respectively. Blocking was very
effective at reducing “error” variance and allowed for the detection of toxicant-related

effects.

Cellular Concentration of PCDHs
The in vitro uptake of PCDHs by salmon lymphocytes was determined by measuring
the concentration of PCB-126 accumulated by pooled cultures of salmon splenic and

pronephric lymphocytes. The concentration of PCB-126 in the cell pellet after 24 hr, 284 ng

66

PCB-126 g wet wt", was 8.5 times greater than the dose administered to the cell suspension,
33.3 ng PCB-126 mL’l (Table 8). The total recovery of PCB-126 from cells plus supernatant
was 46.5%. Recovery of PCB-126 from spiked culture media (45.6%) was comparable to
the recovery in cell cultures and indicates that a consistent proportion of the initial dose is not

recoverable, possibly due to adsorption onto exposed surfaces of the culture tube.

In Vitro Predictions of In Vivo Effects

Results from the in vitro proliferation assay were compared with published reports of
PCDH-related immunotoxic effects in ﬁsh dosed in vivo. In rainbow trout dosed with 80 mg
Clophen A 50 kg" (IP), lymphocyte proliferation was reduced by 31% (Table 9). That dose
is 12.9 times greater than the dose of Aroclor 1254 (6.16 pg mL") which was found to
suppress proliferation by 30% in vitro in this study. This might suggest that effects
demonstrated in vitro greatly overestimate the immunotoxicity of PCBs in vivo. However,
when the effects are related to tissue or cellular concentrations rather than dose, the assays
are more similar. The estimated concentration of Aroclor 1254 accumulated by splenocytes
in vitro (53 ug g‘1 wet wt) was within a factor of 2 of the Clophen A 50 concentration
measured in rainbow trout (30 ug g wet wt") that were dosed with 80 mg kg wet wt" (Table
9). A similar approach was taken to compare in vitro and in vivo effects of TCDD. The
dose-response relationship from a study by Spitsbergen et al. (1986) was compared with the
response curve ﬁom the in vitro assay (Figure 14). TCDD concentrations in tissue from the
in vivo study were estimated by multiplying the TCDD dose by 0.37; the assimilation
efﬁciency derived for Chlophen A 50 (Table 9). Cellular concentrations from the in vitro

assay were estimated by multiplying the TCDD dose by the 8.5 concentration factor. There

67

is considerable overlap between the two curves when the TCDD concentrations are expressed
on a tissue or cellular basis. Although these comparisons are based on tests with different
species and assume similar toxicokinetics for TCDD, PCB-126 and Aroclor 1254 and
Clophen A 50 they suggest that immunotoxic effects measured in vitro may be useﬁil for

predicting effects in whole ﬁsh.

DISCUSSION

The use of in vitro assays allowed for the detection of direct effects of PCDHs on
lymphocyte function. Salmon splenocytes were affected by TCDD over the same range of
concentrations (3 - 30 nM) that suppress proliferative responses in TCDD-sensitive mouse
strains (Karras and Holsapple, 1994). This indicates that in terms of the immunological
processes involved in stimulating lymphocyte proliferation, salmon are as sensitive to TCDD
as mice. However, the sensitivity to PCB-126, a non-ortho substituted (dioxin-like) PCB
congener, was less in salmon (TEF=0.0011) compared to mice (TEF=0.77; calculated from
data in Harper et al., 1995). Other AhR mediated effects in ﬁsh such as enzyme induction and
early life stage mortality (TEFs 0.0014 - 0.027) are also less affected by PCB-126 than would
be expected based on the murine model (Table 5). Thus, the structure-activity relationship
(SAR) for PCDH—induced immunotoxicity in salmon appears to be consistent with other
AhR-mediated effects in ﬁsh and less similar to SARs for mammals.

TCDD and PCB-126 modulated lymphocyte proliferation at concentrations that did
not affect viability, whereas, immunosuppressive concentrations of Aroclor 1254 were also
cytotoxic. It would appear that Aroclor 1254 affects the lympho-proliferative response

through a different mechanism than TCDD or PCB-126. Because only non-ortho PCB

68

congeners have been effective at inducing AhR-mediated effects in ﬁsh, the dioxin-like
activity of Aroclor 1254 would likely be due to the activity of non-ortho substituted
congeners (#77, #81, #126 and #169). These congeners comprise 0.137% of Aroclor 1254
by weight. If all of the non-ortho congeners are considered to be as immunotoxic as PCB-
126 (usually the most potent in terms of AhR-mediated effects) and their toxicity is assumed
to be additive, then the sum of the non-ortho congeners in the Aroclor mixture can be
compared to the potency of PCB-126 alone. This would overestimate the total potency of
the non-ortho substituted PCBs. The concentration of non-ortho congeners in Aroclor 1254
at the IC3O would be 25.7 nM, which is about 50 times less than the IC30 for PCB-126 (1.3
uM). Thus, even with the over-estimate of total TCDD-like activity AhR-mediated effects
induced by non-ortho congeners alone cannot account for the immunotoxicity induced by
Aroclor 1254. However, estimated cellular concentrations of Aroclor 1254 that affected
viability (0.38 and 0.76 mM) are in a range where nonspeciﬁc affects on cellular ﬁinction,
such as nonpolar narcosis, would be predicted (McCarty and Mackay, 1993). Therefore,
direct effects of Aroclor 1254 on splenocyte proliferation seem to be the result of nonspeciﬁc
effects on membrane ﬁinction rather than effects mediated through the AhR. Results from
the FACS analysis indicate that B-lymphocytes may be particularly sensitive to these
nonspeciﬁc effects.

The ability to partition the variance in immune response between toxicant-related
effects and “among-ﬁsh” variability, by the in vitro approach, enhanced the ability to
statistically demonstrate PCDH-induced effects on lymphocyte proliferation. This was
achieved by either analyzing dose-response relationships for individual ﬁsh separately or by

combining ﬁsh but including a blocking factor (“Fish”) in the analysis which increased the

69

homogeniety in the responsiveness of the cell populations to which the treatments were
applied. This approach is particularly useful for immunological responses, such as lymphocyte
proliferation, where toxicant effects may be one of several factors that modulate the response.
The degree of among-ﬁsh variability frequently observed with immune responses (Spitsbergen
et al., 1986; Thuvander et al., 1993: Noguchi et al., 1997a) may mask toxicant effects and
increase the likelihood of Type 11 error, i. e. inability to reject the null hypothesis when in fact
there were real differences.

Distinguishing toxicant-related effects with such high “background variability” is
particularly challenging when treatments are applied to whole animals such as in vivo
laboratory studies or ﬁeld studies. However, there are several ways in which detection could
be improved. Increasing sample size would improve the power of the test and the ability to
detect toxicant effects. If the variance of the response is known or could be estimated then
a simple power analysis can be used to determine the appropriate sample size required to
detect a difference of a particular magnitude. For example, a sample size of 75 would be
required to detect a difference of 50% given a coefficient of variation (CV) of 115%, which
was the variability observed among the four ﬁsh used in the in vitro TCDD assays, and a 0.16
probability of a type 11 error. Depending on the resources available, increasing sample size
in order to increasing the power of the test may not be practical. Alternatively, the immune
response of individual ﬁsh could be measured before and aﬁer treatment. The effect would
be measured as the difference between pre and post-treatment responses for individual ﬁsh,
thus each ﬁsh would serve as its own control. However in some situations, for example ﬁeld
studies with feral ﬁsh or when the ﬁsh must be sacriﬁced in order to perform the assay, it may

not be possible to measure the immune response of an individual before the treatment is

70

applied. In those cases, among-ﬁsh variability may be controlled using analysis of covariance
(ANCOVA). This approach would require the selection of one or more covariates to
measure in addition to the immunological variables. The covariates would be factors that
explain some of the variability in immune response that is unrelated to the toxicant effect.
These could include endogenous factors that are known to affect immune functions such as
glucocorticoid or reproductive hormone concentrations (Tripp et al., 1987; Slater and
Schreck, 1993). The covariate would be included in the ANCOVA model, similar to the
blocking factor in the ANOVA model, to reduce the error variance and increase detection
of toxicant-related effects.

Predicting immunotoxic effects on whole ﬁsh from in vitro assays based simply on the
dose of toxicant administered leads to difficulties in interpretation because of differences
between in vivo and in vitro exposure methods. In other words, it is difficult to directly
compare in vitro doses such as nM toxicant in the medium, with in vivo doses such as nmol
toxicant kg" ﬁsh (injected IP) or nmol toxicant kg'1 in the diet. A more reﬁned approach for
comparing the toxicity of different compounds using critical body residue (CBR) has been
proposed by McCarty and Mackay (1992). The CBR approach uses toxicant concentrations
in the organism, rather than the exposure concentration or dose, to determine potency. While
this method was intended to account for pharmacokinetic differences between compounds it
would also be useful for normalizing effect concentrations derived from in vivo and in vitro
tests. Thus, results derived from in vitro and in vivo studies would be more comparable if the
potency was based on toxicant concentrations in the cells or tissues, rather than dose or

exposure concentrations.

71

When the CBR approach was used to normalize PCDH concentrations that affect
lymphocyte proliferation, the results from the in vitro assays were surprisingly similar to
results from in vivo studies performed in other laboratories. Expressing the toxic potency on
a common basis enabled a more direct comparison of effect concentrations. It should be
noted, however, that the calculated PCDH concentrations in cells (in vitro) and tissues (in
vivo) were based on several assumptions and estimations. For example, the in vitro
concentration factor (8.5) for PCB-126 was detennined using a culture system (culture tubes)
that was greater in scale that the 96-well microplate format used for the in vitro assays.
Differences in the surface area to volume ratios of the two systems may have slightly affected
the proportion of the PCB dose that was available for cellular uptake. Also, the mass of cells
in the culture tube was estimated from the number of cells, their calculated volume and their
speciﬁc gravity which was assumed to be about 1. In addition, cellular concentrations of
TCDD and Aroclor 1254 were calculated using the in vitro concentration factor derived for
PCB-126. The partitioning of PCDHs between culture media and cells, in vitro, would be
inﬂuenced primarily by the octanolrwater partition coefficient (Kow) of the PCDH. Because
the log K0w for PCB-126 (6.89; Hawker and Connell, 1988) is similar to TCDD (6.15 - 7.28;
Jackson et al., 1986) and Aroclor 1254 (6.1 - 6.8; Mackay et al., 1983), the 8.5 value would
be a reasonable estimation of the concentration factors for TCDD and Aroclor 1254. It was
also assumed that the concentration factor, determined at 100 nM, would remain constant
over the range of PCDH concentrations tested (0.3 nM - 90 uM), i.e. the uptake curve is
linear. Departures from linearity may occur at the greater exposure concentrations if uptake
follows the Michaelis-Menton model of saturation kinetics. If this occurs then applying a

constant concentration factor may tend to overestimate cellular. concentrations at greater

72

exposure concentrations. In terms of the tissue concentrations, the assimilation efficiency
derived from the Thuvander et al. (1993) study for Aroclor 1254 (0.37) was used to estimate
TCDD tissue concentrations from the Spitsbergen et al. (1986) study based on the dose of
TCDD administered. Considering the relatively short exposure duration (6 weeks) and similar
half lives of TCDD (105 days; Kleeman et al., 1986) and PCBs (219 days; Nimii and Oliver,
1983) it would appear that the 0.37 assimilation efficiency would provide an adequate
approximation of the expected TCDD tissue concentrations. Overall, the assumptions and
estimations required to calculate cellular and tissue concentrations would not appear to be
extraordinary.

Having determined the potency with which PCDHs modulate lymphocyte proliferation
it was possible to predict whether current concentrations in feral chinook salmon are likely
to be affecting immune responses. The likelihood that immunotoxic effects may occur was
based on the quotient method; the ratio of the PCDH concentration in the ﬁsh divided by
effect concentration derived from the in vitro tests (Table 10). For this calculation, the IC30
for each PCDH was multiplied by the concentration factor (8.5) to estimate the cellular
concentration at the IC,,,. Ratios greater than 1 indicate that the concentrations in the ﬁsh
exceed the effect concentration and there is a risk that the effect will be realized, whereas,
values less than 1 indicate little risk of occurrence. The likelihood that TCDD or PCB—126
alone are affecting lympho-proliferative responses would be negligible because their ratios
were <0.0001 (Table 10). The ratios for Aroclor 1254 (0.019 - 0.076) were about 100 times
greater than for TCDD and PCB-126, but well below 1. The greatest dose of Aroclor 1254
that did not consistently affect the proliferative response was 9 pM; estimated to be 25 p g/g

in the cells. This concentration is 6 fold greater than Aroclor 1254 concentration in salmon

73

tissues and suggests that PCB concentrations are below the threshold for effects on

lymphocyte proliferation.

CONCLUSIONS
The lymphocyte proliferation assay provided a means of measuring direct effects of
PCDHs on salmon immune functions, in vitro, and improved the ability to distinguish toxicant
from non-toxicant effects. The toxic potency of PCDHs derived from the in vitro assays
compared favorably with potencies determined from in vivo studies. Concentrations of
TCDD, PCB-126 and Aroclor 1254 in feral chinook salmon from Lake Michigan are below
the predicted effect concentrations derived from in vitro assays and would present a low risk

in terms of direct effects on lymphocyte proliferation.

74

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Table 6. Analysis of variance table for one-way ANOVA tests for effects of PCDH treatment
on lymphocyte proliferation when replicate ﬁsh are analyzed together (n = 4 for each
compound). The percent of total variability explained by different sources is the ratio
of source sums-of-squares : total sums-of-squares.

 

 

Sums of Mean Percent of
Source Squares df Square F-Ratio P Total Variability
TCDD 0.0998E10 4 2.49E8 0.272 0.891 6.8
Error 1.37E10 15 9.16E8 93
Total 1.47E10
PCB 126 0.345E10 4 0.862E9 0.377 0.822 9.1
Error 3.44E10 15 2.29E9 91
Total 3.78E10
A1254 3.25E10 2 1.62E10 5.538 0.027 55
Error 2.64E10 9 0.293E10 45
Total 5.88E10

 

76

Table 7. Analysis of variance table for randomized block ANOVA to test for effects of PCDH
treatment on lymphocyte proliferation when replicate ﬁsh are analyzed together (n =
4 for each compound). Fish number was used as the blocking factor. The percent of
total variability explained by different sources is the ratio of source sums-of-squares
: total sums-of-squares.

 

 

Sums of Mean P630331] 0f
Source Squares df Square F-Ratio P Variability
FISH 1.29E10 3 4.31E9 65.1 <0.001 0.88
TCDD 009981310 4 0.249E9 3.77 0.033 0.068
Error 0.0794E10 12 0.0662E9 0.054
Total 1.47E10
FISH 3.25E10 3 1.08E10 71.2 <0.001 0.86
PCB 126 0.345E10 4 0.0862E10 5.7 0.008 0.091
Error 0.183E10 12 0.0152E10 0.049
Total 3.78E10
FISH 1.71E10 3 0.569E10 3.66 0.083 0.290
A1254 3.25E10 2 1.62E10 10.4 0.011 0.552
Error 0.932E10 6 0.155E10 0.158
Total 5.88E10

 

77

Table 8. Distribution of PCB 126 in cell cultures 24 hours after dosing with 33 ng PCB
126/mL. Results are from duplicate cultures, mean (SE).

 

 

n , PCB 12 6 Concentration Ratio:
Matrix 5 . Recovery (%) in Matrix Conc. : Dose
in matrix
Supernatant 27.8 (3.7) 33.4 11.1 ng/mL 0.33
Cells 11.1 (1.3) 13.2 284 ng/g 8.5
Supernatant + Cells 38.8 (4.9) 46.5
Spiked TCM 38.0 (8.5) 45.6

Blank not detected

78

Table 9. Comparison of in vitro and in vivo approaches to evaluating the effect of PCBs on
lymphocyte proliferation. Cellular concentrations of Aroclor 1254 were estimated
from the in vitro assay at the IC30, and compared to the concentration of Clophen A50
measured in rainbow trout that were closed (IP) with 80 mg/kg (Thuvander et al.,
1993). Proliferation of rainbow trout splenocytes to LPS in Clophen A50-treated ﬁsh
was depressed by 31% compared to controls, although the effect was not statistically
signiﬁcant in that study. The assimilation efficiency for Clophen A50 was calculated,
based on data from the Thuvander study, as the ratio of Clophen A50 concentration
measured tissues : dose of Clophen A50 administered.

 

In vivo test
(Thuvander et al., 1993)

In vitro assay
Species: chinook salmon rainbow trout
Dose Administered: 6.2 pg Aroclor 1254/mL 80 mg Clophen A50/mg (IP)
Assimilation Efficiency: 0.37

Concentration Factor: 8.5

Concentration in

cells or tissues: 5" pg/g 3O fig/g

Effect on Proliferation: -30% -31%

 

79

Table 10. Comparison of environmental concentrations of PCDHs in Lake Michigan chinook
salmon with estimated effect concentrations based on the in vitro splenocyte
proliferation assay.

 

 

Concentration in Estimated cellular Ratio:
Lake Michigan concentraiton at the Environmental conc./
PCDH Chinook Salmon IC30 Effect Conc
Aroclor 1254 1 - 4 pg/g' 53 pg/g 0.019 - 0.076
PCB 126 0.5 - 1.3 ng/g' 3048 ng/g 0.0002 - 0.0004
TCDD 3 pg/g2 4624 pg/g 0.0006

 

‘ G. Noguchi (unpublished data)

2 Smith et al., 1990

80

Figure 9. Proliferation of salmon splenocyte cultures stimulated with LPS and treated with
graded doses of TCDD (A), PCB-126 (B) and Aroclor 1254 (C). Results are from in vitro
assays with cells from four individual ﬁsh (denoted by lower case letters a, b, c, etc.). Values
are meansiSE for 4 replicate cultures per treatment. Signiﬁcant differences (a=0.05)
between PCDH-treated cultures and controls are indicated by an asterisk (*). Proliferation
in control cultures (vehicle alone) was not signiﬁcantly different from positive control cultures
(no vehicle). Background proliferation (cpm) in non-stimulated cultures (no LPS) from
individual ﬁsh were 562 (a), 252 (b), 238 (c), 683 (d), 769 (e), 906 (f), 847 (g), 1438 (h),
4670 (i), 1232 (j), 550 (k), and 951 (l).

81

 

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84

Figure 10. Cytotoxicity of TCDD, PCB-126 and Aroclor 1254 to chinook salmon
splenocytes. Numbers of viable splenocytes were counted after 24 hr incubation. Values are
meansztSE for 4 replicate cultures per treatment. Signiﬁcant differences (a=0.05) between
PCDH-treated cultures and controls are indicated by an asterisk (*).

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Figure 1]. Characterization of the effects of Aroclor 1254 on splenocyte subpopulations
using ﬂow cytometry. A) cytograms (forward scatter vs. side scatter) and B) FITC
fruorescence in the lymphocyte region (R2) of cell suspensions dosed with O, 9 and 45 pM
Aroclor 1254. C) Percentage of Ig’ lymphocytes (B cells) in the lymphocyte region, based

on FITC ﬂuorescense.

87

     
  

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88

Figure 12. Dose-response curves for the effects of TCDD (O ), PCB 126 (A ) and Aroclor
1254 (I) on chinook salmon splenocyte proliferation. Responses, expressed as percent of
control proliferation (cpm), from 4 replicate ﬁsh were combined and plotted (meaniSE)
against log PCDH concentration (M).

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Figure 13. Direct effects of TCDD (A), PCB 126 (B) and Aroclor 1254 (C) on splenocyte
proliferation with replicate ﬁsh combined. Values are meansiSE for the 4 replicate ﬁsh
shown in Figure 9.

91

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92

In vivo: ng TCDD/g tissue

 

 

 

 

 

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Figure 14. Comparison of in vivo and in vitro effects of TCDD on lymphocyte proliferation
(CPM). The dose-response curve for in vitro effects (0) is based on in vitro assays with
chinook salmon splenocytes. Data from Spitsbergen et al., (1986) were used to construct the
dose-response curve for in vivo effects of TCDD on splenocyte proliferation in rainbow trout
(A).

“L

CHAPTER 4

Development of an Assay to Assess Effects of Polychlorinated Diaromatic
Hydrocarbons on In Vitro Antibody Production and Antibody Affinity in Chinook

Salmon (Onchorhynchus tschwytscha) Splenocyte Cultures.
(for submission to Developmental and Comparative Immunology)

INTRODUCTION

Polychlorinated diaromatic hydrocarbons (PCDH) comprise a class of chemicals such
as, polychlorinated biphenyls (PCB), polychlorinated dibenzo-p-dioxins (dioxins) and
polychlorinated dibenzofurans (PCDF), that induce pleiotropic effects in mammals including
immunomodulation (Vos and Luster, 1989). The humoral or antibody producing immune
response in adult and juvenile rodents is particularly sensitive to some of these compounds
(Holsapple et al., 1991b) and PCDHs can also affect humoral immunity in ﬁsh (Thuvander
et al., 1991; Arkoosh et al., 1994a; Rice and Schlenk, 1995). The method most frequently
used to measure humoral immune responses in ﬁsh and mammals is the antibody-fonning-cell
(AFC) assay, also referred to as the plaque-forming-cell (PFC) assay. The AFC assay detects
the number of B cells that have been stimulated to produce antibodies in response to contact

with antigen. Another method for detecting humoral responses is to measure the quantity of

93

94

antibody produced. The advent of enzyme-linked immunosorbant assay (ELISA) technology
has lead to the development of sensitive and reproducible quantitative assays for the detection
of serum antibodies (Hombeck et al., 1991). Determination of antibody production by ELISA
has been shown to be as sensitive as the AFC response for detecting effects of
immunomodulating chemicals on humoral immunity in rodents (Temple et al., 1993). ELISA
protocols have been adapted for the detection of antibodies produced by ﬁsh (Arkoosh and
Kaattari, 1990).

ELISAs have also been developed to characterize the binding affinity of antibodies for
antigen (Nieto et al., 1984). An affinity ELISA has recently been adapted for assessing the
affinity of antibodies produced by ﬁsh (Shapiro, 1995). Results from that study have
demonstrated that affinity maturation, a critical element of the memory responses in mammals,
also occurs in ﬁsh. Thus, analysis of serum antibodies by ELISA can be used to detect not
only the magnitude of the humoral response, but to also characterize the nature of the
response, i. e. primary versus secondary responses.

Elements of the humoral immune response including the induction of antibody
production have been reproduced in vitro using primary cultures of human (Wood et al.,
1993), rodent (Morris et al., 1993) and ﬁsh (Kaattari and Yui, 1987; Shapiro, 1995)
lymphocytes. An in vitro assay system for antibody production coupled with quantitative and
affinity ELISA for antibody detection could provide a sensitive and informative system for
assessing effects of PCDHs on humoral immunity in ﬁsh. The objectives this study were to
evaluate the effects PCBs and 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on humoral
immunity in chinook salmon. Lymphocyte culture techniques were adapted to support

antibody production by splenocyte cultures, in vitro. Antibodies produced in culture

95

supernatants were detected by quantitative ELISA and the affinity of antigen-speciﬁc

antibodies was determined by affinity ELISA.

MATERIALS AND METHODS
Fish Holding and Sampling
Juvenile chinook salmon were obtained from the Michigan Department of Natural
Resources Wolf Lake Hatchery, Mattawan, MI, and held at the National Biological Service,
Great Lakes Science Center in Ann Arbor, MI. Fish holding and sampling information were
described previously (Noguchi et al., 1997a). All assays were performed using splenic cell

suspensions from juvenile (200 - 300 mm) chinook salmon salmon.

Reagents and Chemicals

TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin) and 3,3',4,4',5-pentachlorobiphenyl
(IUPAC #: PCB-126) were purchased from Accurate Chemical CO, New Haven, CT.
Aroclor 1254 was obtained from Monsanto, St Louis, MO. Culture medium (RPMI-1640),
antibiotics (gentamycin sulfate), trinitrophenyl-haptenated lipopolysaccharide (TNP-LPS), 2-
mercaptoethanol, dimethylsulfoxide (DMSO), were purchased from Sigma Chemical, St.
Louis, MO. The Warr 1.14 monoclonal antibody used in the quantitative and afﬁnity ELISA
was obtained from the Virginia Institute of Marine Sciences, College of William and Mary,
Gloucester Point, VA and the HRP-conjugated goat anti-mouse secondary antibody was
purchased from Kirkigard and Perry, Gaithersburg, MD. Stock solutions of TCDD and PCBS
were prepared in DMSO at concentrations 1000 times greater than the nominal dosing

concentration. Intermediate dilutions of PCDHs and DMSO were prepared in TCM just prior

96

to dosing cells. TNP(trinitrophenyl)-haptenated bovine serum albumin (TNP-BSA) was

prepared according to the method of Garvy et al. (1977).

Culture Conditions for In Vitro Antibody Assay

Conditions for culturing chinook salmon splenocytes for in vitro antibody production
were adapted from Arkoosh and Kaattari (1992). Spleen cell suspensions from individual
salmon were prepared as described by Noguchi et al. (1997a). However, after the ﬁnal
washing the cells were resuspended in tissue culture medium (TCM) which contained RPMI-
1640 supplemented with non-essential and essential amino acids, sodium pyruvate, L-
glutamine, nucleosides (adenosine, uridine, cytidine, and guanine), 2-mercaptoethanol,
gentamicin sulfate, and homologous chinook salmon plasma (1%). Preparation of
homologous plasma was described previously (Noguchi et al., 1997a). Lymphocytes were
enumerated by microscopic examination in a hemocytometer and viability was determined by
trypan blue exclusion. The volume of cell suspension was adjusted to achieve a ﬁnal cell
concentration of 1 x 107 cells mL".

Cell cultures were dosed with PCDHs by adding 25 pL of PCDH or DMSO solution
(in TCM) and 1 pL of TNP-LPS to 474 pL of cell suspension. The suspension was mixed by
aspirating it with a repeat-dispensing pipetter after which a 100 pL aliquot was dispensed into
4 wells of a ﬂat-bottom 96-well microtiter plate. The ﬁnal concentration of DMSO in PCDH
and vehicle control cultures was 0.01%. Negative controls (cell suspension containing no
TNP-LPS, PCDH or vehicle) and positive controls (cell suspensions containing TNP-LPS but
no PCDH or vehicle) were assayed along with the treated cultures. All treatments were

assayed in quadruplicate as a measure of the variability in responsiveness due to

 

97

methodological factors including differences in PCDH uptake and culture conditions among
wells. The design does not allow assessment of variability associated with dose
administration, i. e. pipetting error. The plates were covered with mylar plate sealers, placed
in a sealed gas chamber that was purged with a 10% C02, 80% N2, 10%O2 gas mixture, and
incubated at 17° C. Every 3 d, 10 pL of a nutrient supplement was added to each cell
culture. After 12 d cultures were transferred to microcentriﬁige tubes and centriﬁiged at
500xg for 10 min. The supematants were removed, transferred to new microcentrifuge tubes

and stored at -200 C.

ELISA for Anti- TNP Antibody Activity

TNP—speciﬁc antibodies were measured by indirect, antigen-excess ELISA. The TNP
antigen was bound to wells of a round-bottom 96-well microtiter plate (Corning) by
incubating 100 pL of TNP-BSA in carbonate buffer (147 ug TNP-BSA mL") overnight at
4° C. The contents were removed by rapidly ﬂicking the inverted plate. The plate was then
washed 3x with a wash buffer (0.05% BSA and 0.1% Tween 20 in PBS). Wells were ﬁlled
with a blocking buffer (3% BSA, 0.5% Tween 20 and 0.025% sodium azide in PBS) and
incubated at 37° C for 1.5 hr. Plates were washed 3x then 50 pL of serum or culture
supernatant was added to appropriate wells and incubated at 17° C for 1.5 hr. The serum or
supematants were removed and the plate was washed 3x, then 50 pL of Warr 1.14 antibody
(1:4000 dilution) was added to each well. The plates were incubated at room temperature
for 1 hr then washed 3x. Next, 50 pL of secondary antibody (HRP-conjugated goat anti-
mouse 1g @ 1 pg mL") was added to each well and incubated at room temperature for 30

min. The plates were washed 3x then 100 pL of HRP substrate (Th/H3 + H202) was added to

 

 

98

each well and incubated at room temperature for 10 min to allow the color to develop. A
Cayman microplate autoreader (Cayman Analytical, Ann Arbor, MI) was used to measure the
optical density (OD) at 650 nm. Serum from TNP-KLH immunized rainbow trout (RBT),
provided by Dr. Mary Arkoosh (National Marine Fisheries Service, Newport, OR), was used
as a standard for TNP-speciﬁc antibody activity. Dilutions of RBT sera, culture supernatants,

Warr 1.14 antibody and the secondary antibody were prepared in wash buffer.

Affinity ELISA

An ELISA technique described by Kaattari and Shapiro (1994) was used to
characterize the anti-TNP binding affinity of antibodies in RBT sera and supematants from
chinook salmon splenocyte cultures. The assay is based on the principal that high antigen
concentrations in the stationary phase (bound to the wells) are required for low affinity
antibodies to bind whereas antibodies with high affinity can bind at low antigen
concentrations. Thus, an affinity proﬁle can be generated by culturing sera or supematants
in wells containing a range of antigen concentrations. The affinity of antibodies binding at
each antigen concentration was determined by measuring the concentration of free antigen
(TNP-lysine) that reduced binding to the stationary phase (plate) by 50%. Binding of high
affinity antibodies to stationary antigen would be depressed by lower concentrations of free
antigen than the binding of lower affinity antibodies.

Seven concentrations of TNP-BSA (0.037 - 147 pg/mL) were prepared in carbonate
buffer and 100 pL of each concentration was added to all wells of one row of a round-
bottom 96-well microtiter plate (Figure 15). The ﬁrst row (A) received carbonate buffer

containing no TNP-BSA. The plate was incubated overnight at 4° C. The plate was washed

 

 

99

and active sites were blocked with blocking buffer as previously described. Five
concentrations of free TNP competitor (TNP-lysine) were prepared in wash buffer (1x10'° -
lxlO'2 M) and 25 pL of each concentration was added to duplicate wells (Figure 15). Wells
in columns 1 and 2 received 25 pL of wash buffer alone. All wells then received 25 pL of
culture supernatant at a dilution which would generate 50% of maximun OD. The
appropriate dilution was determined from prior analysis of supernatant titer. The plate was
incubated at 17° C for 1.5 hr. The remaining steps, i.e. addition of Warr 1.14 antibody,
secondary antibody, HRP substrate (TMB) and measurement at 650 nM, were the same as
described for the TNP-speciﬁc ELISA. Apparent aﬂinity constants (aK) were calculated from
the following equation:
aK = log (“1115111) (1)
where:
[H,,,] = concentration of competitor (TNP-lysine) at 50% of max 650 nm OD

calculated from competitive-inhibition curves.

Weighted average affinities (K) were calculated with equation 2.
1K. = Z aK x p (2)
where:
p = the proportion of total TNP-speciﬁc antibody with a give

binding affinity.

Statistics
All statistical tests were performed using SYSTAT statistical software (Version 5.1),

Evanston, IL. Effects of PCDHs on antibody production in splenocyte cultures from

 

 

100

individual ﬁsh were analyzed by one-way analysis of variance (ANOVA) and tukeys test was

used to make pairwise-comparison of treatment means.

RESULTS

Optimizing ELISA Conditions

ELISA conditions were initially established for detecting TNP-speciﬁc antibodies in
standard anti-sera, i. e. serum from TNP-KLH immunized rainbow trout (RBT). The RBT
sera contained high anti-TNP antibody activity and could be detected by ELISA even at low
concentrations of coating antigen (0.036 ug TNP-BSA mL"). Preliminary attempts to induce
TNP-speciﬁc antibody production in splenocyte cultures resulted in low antibody activity; less
than a 2 fold increase in antibody activity compared to unstimulated controls. Because of this
low in vitro antibody production, ELISA conditions were reevaluated to determine whether
they were suitable for detecting TNP-speciﬁc antibodies generated in culture supematants.
The concentration of coating antigen was assessed to determine whether concentrations of
TNP-BSA that bind serum antibodies were equally effective at binding antibodies in culture
supernatant. RBT sera and culture supematants were assayed with varying concentrations
of TNP-BSA antigen. At the antigen concentration used in the initial tests (0.036 ug TNP-
BSA mL") about 70% of the of the total TNP-speciﬁc antibody activity in RBT sera was
detected, whereas, less than 5% of the antibody activity in culture supernatant was detectable
(Figure 16). Much greater antigen concentrations were required for detecting maximal
antibody activity in culture supematants (147 ug TNP-BSA mL") than in the RBT sera (0.58
ug TNP-BSA mL"). These greater antigen concentrations suggest that the antibodies

produced in vitro had lower affinity for the TNP antigen that antibodies in the RBT sera.

101

Analysis of TNP-speciﬁc antibodies by affinity ELISA conﬁrmed that antibodies
detected in culture supematants did, in fact, have lower affinities than antibodies in the RBT
sera (Table 11). Apparent affinity constants (aK) were determined from competitive-
inhibition curves for culture supernatant and RBT sera (Figure 17). The antibody affinity
proﬁles indicate that while most of the TNP-speciﬁc antibody activity in supematants was due
to low affinity antibodies (aK 2.98 - 3.69) a small, but detectable, proportion of high affinity
antibody (aK 6.3) was also present (Figure 18). Nevertheless, the weighed average affinity
(1(0) of TNP-speciﬁc antibodies was less in culture supernatant (3.3) compared to RBT sera

(6.3; Table 11).

Quantitation of TNP-Speciﬁc Antibody in Supernatants

Serum from TNP-KLH immunized RBT was used as a standard for quantitation of
TNP-speciﬁc antibody activity in culture supernatants. Standard curves were generated by
analyzing serial (2x) dilutions of RBT sera; from 1:2000 to I:1,024,000. A representative
standard curve is shown in Figure 19. The curves were consistently linear between 132,000
and 1:512,000 dilutions when the responses (ODs) were logit-transfonned and the serum
concentrations were log transformed. The OD at the least dilution within the linear range
(132,000) was considered to be the max OD and it ranged from 0.9 to 1.1. The level of
TNP-speciﬁc antibody activity in RBT sera was standardized by assigning a value of one unit
of antibody activity to the serum concentration at '/2 max OD. The antibody activity of
undiluted RBT sera was then calculated by multiplying the dilution factor at ‘/2 max OD by
the activity at ‘/2 max OD (1 unit). The average antibody activity in RBT sera from 11

standard curves was 1466 units of activity pL" serum (Table 12).

102

TNP-speciﬁc antibody activity of culture supematants was calculated from RBT sera
standard curves. The appropriate dilution of supematants for quantitative ELISA was
determine by screening samples in order to select a dilution that would result in an OD that
was within the linear range of the standard curve. Since antibody activity was based on one
dilution per sample, it was also important to demonstrate that response curves for
supernatant antibodies were parallel to RBT standard curves. Departures from parallelism
would indicate that the response is either attenuated or strengthened by confounding
substances in the supernatant. Supernatant from two in vitro splenocyte cultures were titered
and the response curves were compared to a RBT standard curve (Figure 19). Slopes of the
logit-transformed response curves for culture supematants (-l .98 and -2.14) were within the
95% conﬁdence interval (Neter and Wasserman, 1974) for the slope of the RBT standard
curve (-2.10). Thus, the RBT sera standard curve was appropriate for calculating antibody

activity in the culture supematants.

Effects of PCDH on In Vitro Antibody Production

Exposure of splenocyte cultures to TCDD resulted in a dose-related decrease in TNP-
speciﬁc antibody production (Figure 20A). Statistically signiﬁcant reductions in antibody
activity were detected at concentrations of 3 and 30 nM TCDD. A downward trend in
antibody activity was also observed in cultures exposed to increasing concentrations of PCB-
126 (Figure 208). Antibody activity was depressed by approximately 50% at the greatest
concentration (300 nM PCB-126) , however, this effect was not statistically signiﬁcant at the
0.05 level because of variability within treatments. The response of splenocyte cultures to

Aroclor 1254 exposure was less predictable (Figure 20C). Antibody activity was enhanced

 

103

at the greatest concentration, 9 pM Aroclor 1254, but the effect at lower concentrations was

less certain.
Effects of TCDD and PCB-126 on antibody affinity were evaluated by affinity ELISA.
TNP-speciﬁc antibody affinities were detennined from cultures treated with 3 nM TCDD and
300 nM PCB-126. Supernatants from replicate culture wells were pooled in order to have
sufficient sample volume to perform the assay. Antibody affinities were calculated from
competitive-inhibition curves (Figure 21). The average affinity (K) of antibodies produced
in vitro appeared to be unaffected by TCDD or PCB-126 (Ko 3.5 - 4.0; Table 13). However,
close inspection of affinity proﬁles revealed that high afﬁnity antibodies (aK 5.6 - 6.3) present
in control cultures were absent in supernatant from TCDD-treated or PCB-126 treated
cultures (Figure 22). Elimination of high affinity antibodies by treatment with 3 nM TCDD
or 300 nM PCB-126 may have contributed to the lesser antibody activity observed in cultures

exposed to TCDD and PCB-126 at those concentrations (Figures 20A&B).

DISCUSSION

Conditions are described for the in vitro induction of TNP-speciﬁc antibodies by
chinook salmon splenocytes and for the detection of antibody activity by ELISA. With slight
modiﬁcations, culture conditions that support in vitro AFC responses (Arkoosh and Kaattari,
1992) were also suitable for inducing antibody production in vitro. However, the detection
of antibodies in culture supematants by ELISA required considerably greater concentrations
of coating antigen (TNP-B SA) than for the detection of standard antisera. Greater antigen
concentrations were necessary because of the low affinity of TNP-speciﬁc antibodies

produced in vitro. The difference between the affinity of antibodies in splenocyte cultures and

104

TNP-speciﬁc antibodies in standard antisera (produce in vivo), could be explained by affinity
maturation.

Affinity maturation is a characteristic of the amnastic (memory) humoral response in
which there is an increase in antibody affinity following secondary and subsequent exposure
to antigen (Roitt et al., 1989). In mammals, the increase in affinity is accompanied by a
switch in the predominant form (isotype) of antibody produced, from IgM to IgG (Roitt et
al., 1 989). Secondary exposure to antigen results in an increase in the affinity of IgG
antibodies whereas the affinity of IgM antibodies remains unchanged. Fish possess a more
restricted repertoire of antibody isotypes compared to mammals and the process of isotype
switching does not appear to be a feature of the memory response in teleosts (Kaattari, 1992).
The predominant form of serum antibodies in teleosts is an IgM-like tetrameric
immunoglobulin (Kaattari and Piganelli, 1996). Until recently, there has been no evidence
that afﬁnity maturation occurs in ﬁsh (Arkoosh et al., 1991a). However, an affinity ELISA
technique developed by mammalian immunologists (Nieto et al., 1984) and adapted for ﬁsh

by Kaattari and Shapiro (1994) has enabled the detection of antibodies expressing a range of
afﬁnities for antigen. Using the affinity ELISA, Shapiro (1995) demonstrated that the
rainbow trout receiving a single (priming) dose of antigen produced antibodies with lower
affinity than rainbow trout that were exposed to antigen multiple times. In addition, there
was a time-dependent increase in antibody affinity following a single immunization. Increases
in antibody affinity are attributed to decreasing antigen concentrations over time. Only B
cells with high affinity antigen receptor are capable of responding at low antigen
concentrations, therefore, the proportion of high afﬁnity antibody increases as antigen

concentration diminishes (MacLennan and Gray, 1986). Thus, it should not be surprising that

105

in vitro immunization of splenocytes from naive chinook salmon induced production of
predominantly low affinity antibodies, analogous the primary in vivo response. Antibodies
with much greater affinity for TNP were detected in standard sera from rainbow trout that
had received repeated immunizations. Multiple and prolonged exposure of rainbow trout to
antigen (in vivo) would have provided the necessary conditions for affinity maturation to
occur and could explain the predominance of high affinity antibody that was detected by
affinity ELISA.

It should be noted that the segregation of antibodies into discrete affinity subclasses,
as determined by affinity ELISA, is inﬂuenced by the design of the ELISA. The maximum
number of antibody subclasses that can be detected is limited by the number of different
coating antigen concentrations (TNP-BSA). In the present study seven concentrations of
TNP-BSA were used, therefore, no more than seven antibody affinities could have been
detected when, in fact, there may be a continuum of affinities. Thus, the affinity proﬁles
generated by ELISA are estimations of the actual affinity distributions. The most useﬁil
information that can be derived from the afﬁnity proﬁles is the range of antibody affinities and
the weighted average affinity (K).

The ability to induce antigen-speciﬁc antibody synthesis in vitro combined with the
quantitative ELISA provided a model system for assessing direct effects of PCBs and TCDD
on humoral immunity in chinook salmon. The antibody response integrates all aspects of
humoral immunity including, antigen detection, interactions between B cells and accessory
cells, clonal proliferation, terminal B-cell differentiation and antibody synthesis. The
proliferative response of chinook salmon splenocytes was previously shown to be affected by

direct exposure to TCDD (Noguchi eta1., 1997b) and this study demonstrated that TCDD

 

106

also suppresses antibody production. Based on these results it can be concluded that TCDD
can alter humoral immune responses in chinook salmon. Concentrations of TCDD that
affected antibody synthesis and lymphocyte proliferation (3 and 30 nM ) were comparable to
concentrations that suppress the in vitro AFC response in murine splenocytes (1C50 7 nM;
Harper et al., 1995). This indicates that the intrinsic sensitivity of the humoral immune
response in chinook salmon to TCDD is similar to that of B6C3F1 mice, one of the more
sensitive mammalian models for TCDD-induced immunotoxicity. PCB-126 exhibited
immunosuppressive tendencies, although great within-treatment variability diminished the
statistical signiﬁcance of these effects. Nevertheless, an average reduction in antibody activity
of about 50%, observed at the greatest concentration of PCB 126 (300 nM), may be
biologically signiﬁcant. PCB-126 also suppresses lymphocyte proliferation (Noguchi et al.,
1997b), although the potency was less than that of TCDD (toxic equivalency factor 0.001 1).
Similarly, greater concentrations of PCB 126 were needed to depress antibody production
compared to TCDD. In contrast, the sensitivity of murine humoral responses to PCB 126
(EC50 9.1 nM) and TCDD (EC50 7.0 nM) are similar (Harper et al., 1995). Structure-activity
relationships for immunomodulation as well as other effects such as enzyme induction (Janz
and Metcalfe, 1991; Newsted et al., 1995) and early life-stage mortality (Walker and
Peterson, 1991) indicate that planar PCB congeners are less potent at inducing dioxin-like
effects in ﬁsh than in mammals.

Aroclor 1254 suppresses in vitro lympho-proliferative responses but only at
concentrations that are also cytotoxic (45 and 90 pM; Noguchi et al., 1997b). In the present
study, Aroclor 1254 was tested at non-cytotoxic concentrations (0.45 to 9 pM) and no

signiﬁcant depression in antibody activity was detected. However, at lesser concentrations

 

107

Aroclor 1254 can enhance antibody production (Figure 20C). Up-regulation of B-cell activity
has also been observed in ﬁsh exposed in vivo to PCBs (Thuvander et al., 1993; Rice and
Schlenk, 1995) and to complex mixtures of environmental contaminants containing PCBS
(Faisal et al., 1991a). However, in some studies PCBs have been ineffective at modulating
immune responses (Cleland and Sonstegard, 1987; Cleland et al., 1988a) while in others PCBs
were immunosuppressive (Arkoosh et al., I994a). Whether PCBs stimulate, suppress or have
no effect on the piscine immune system may well be related to the concentration of PCBs that
accumulates in target cells. Aroclor 1254 appears to have a rather steep dose-response
relationship in terms of direct effects on humoral immunity (in vitro), because there was only
a 5-fold difference between concentrations that were cytotoxic and concentrations that were
immunostimulatory. This type of response does not allow for discrimination of effects beyond
the limited range of concentrations that either stimulate or kill susceptible cells and could lead
to the apparent contradictory results mentioned above. Interpretation of results from in vivo
exposure studies is ﬁirther complicated by the fact that tissue or cellular concentrations of
PCBs are rarely reported and differences in exposure method, e.g. IP injection or dietary
exposure, preclude the determination of potency based solely on exposure.

The suppressive effects of TCDD and PCB-126 on antibody production were
accompanied by the elimination of antibodies with high afﬁnity for the TNP antigen. Because
B cells express antigen receptors that are analogues of the antibody they will produce, high
affinity antibodies are produced by B cells with high affinity antigen receptors. B cells
expressing high affinity receptors would be more sensitive to antigen-induced activation than
cells with low affinity receptors. The heightened level of activation attained by clones

expressing high affinity receptor may have made these cells more susceptible to TCDD and

108

PCB-126. There are two lines of evidence which indicate that the effectiveness of TCDD at
inducing biochemical or immunomodulatory effects is dependent on the activation state of the
lymphocytes. Induction of 7-ethoxyresoruﬁn-O-deethylase (EROD) and aryl hydrocarbon
hydroxylase (AHH) synthesis are measures of the biochemical activity of TCDD and other
AhR agonist. TCDD induces EROD synthesis in mitogen-stimulated human lymphocytes but
not in unstimulated lymphocytes (Wood et al., 1993). TCDD also induces AHH activity in
a human lymphoblastoid cell line following mitogen stimulation (Waithe et al., 1991).
Similarly, high density (resting) B cells are refractory to TCDD, whereas, TCDD suppressed
both background proliferation and IgM secretion by low density (activated) B cells (Morris
et al., 1993; Wood et al., 1993). Thus, the state of lymphocyte activity may be an important
factor in regulating the efficacy of TCDD and may render B cells expressing high affinity
antigen receptors particularly susceptible to TCDD and other AhR agonists.

If B cells expressing high affinity receptors are in fact more sensitive to ArR agonists
then secondary humoral responses should be more sensitive than the primary response. As
a result of affinity maturation and clonal proliferation (Kaattari, 1992), B-cell clones
expressing high affinity antigen receptors would be expected to constitute a greater
proportion of the responding cells during the memory response. Thus, preferential
suppression of high afﬁnity clones would diminish a greater proportion of responding cells
during the memory response than during the primary response. Secondary AFC responses
have been reported to be more sensitive than primary responses in chinook salmon treated
with Aroclor 1254 or dimethylbenz[a]anthracene (Arkoosh et al., 1994a) and in chinook
salmon retrieved from a PCB/PAH (polycyclic aromatic hydrocarbon)-contaminated estuary

in Puget Sound (Arkoosh et al., 1991). The mechanism underlying chemical-induced

 

109

immunosuppression of secondary humoral responses has yet to be elucidated, however,
results from the present study indicate that greater sensitivity of B cells expressing high
affinity antigen receptor may contribute to a diminished memory response.

In conclusion, in vitro induction of antibody production combined with the detection
of antibody activity and affinity by ELISA provided a system for assessing the potency with
which PCBs and TCDD modulate humoral immunity in chinook salmon and for revealing

possible mechanisms of action.

 

110

Table 11. Weighted-average affinity constants (1(0) for TNP-speciﬁc antibodies in
supernatants from TNP-LPS stimulated chinook salmon splenocyte cultures and
serum from TNP-KLH immunized rainbow trout. (aK - apparent affinity constant;
p- proportion of total TNP-speciﬁc antibody with a given aK)

 

 

 

 

Culture Supernatant Rainbow Trout Sera
TNP-BSA TNP-BSA
aK p aK * p aK p aK * p
(rug mL") (113 mL")
0.576 6.33 0.057 0.36 0.009 7.48 0.0909 0.68
2.3 3.69 0.305 1.13 0.036 7.31 0.195 1.43
9.2 3.15 0.366 1.16 0.144 6.69 0.26 1.74
36.9 3.18 0.128 0.41 0.576 5.89 0.258 1.52
147 2.98 0.143 0.43 2.3 5.18 0.178 0.92

 

Weighted-Average Affinity (K): 3.3 6.3

 

111

Table 12. TNP-speciﬁc antibody activity of serum from TNP-KLH immunized rainbow trout.
Values from different 1 1 standard curves.

 

Logit at pL Serum per well Units of Anti-TNP Antibody

 

 

Plate V2 Max OD ‘/2 Max OD at '/2 Max OD Activity per pL Serum
1 0.560 -1 .444 0.00071 1418
2 0.565 -1.745 0.00068 1465
3 0.495 -2.019 0.00068 1471
4 0.528 -1.779 0.00065 1536
5 0.473 -1 .337 0.00067 1482
6 0.453 -1.786 0.00068 1470
7 0.467 -1.670 0.00066 1521
8 0.447 -1.939 0.00070 1419
9 0.497 -1.778 0.00067 1501
10 0.479 -2.074 0.00071 1404
11 0.492 -1.857 0.00070 1433

Mean 1466
SD 43

CV 0.03

 

112

 

 

 

 

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(TNP-BSA) concentration and competitor (TNP-lysine) concentration. Shading represents
the expected magnitude of antibody bound to stationary antigen (TNP-BSA); I-greatest
concentration of bound antibody, Cl-least concentration of bound antibody.

114

Figure 16. Aﬁinity ELISA: Proportion of total TNP-speciﬁc antibody bound at different
concentrations of TNP-BSA coating antigen; A) serum from TNP-KLH immunized rainbow
trout, B) supernatant from TNP-LPS stimulated splenocyte cultures, C) cumulative
proportion of TNP-speciﬁc antibody bound at different TNP-BSA antigen concentrations (0
rainbow trout serum, A splenocyte culture supernatant).

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Proportion of Total TNP-Speciﬁc Ab

Proportion of Total TNP-Specific Ab

Cumulative Proportion of TNP-Speciﬁc Ab

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Figure 20. Direct effects of TCDD (A), PCB-126 (B) and Aroclor 1254 (C) on in vitro
antibody production by chinook salmon splenocytes. Values are means :tSE for the 4
replicate wells. Responses from in vitro cultures from 2 ﬁsh are shown for each compound.
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B) vehicle control C) 3 nM TCDD and D) 300 nM PCB-126.

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Figure 22 (continued).

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