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thesis entitled

EGG INJECTIM STUDIES TO DETERMINE
THE EFFECTS OF POLYHALOGENATED ENVIRNM-
MENTAL CONTAMINANTS M AVIAN SPECIES

presented by
DEBRA CURTIS POWELL

has been accepted towards fulfillment
of the requirements for

Masters degree in Ailing] Science

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Major professor

Date July 24, 1995

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MSU Is An Affirmative Action/Equal Opportunity IMUIAIOH

 

 

EGG INJECTION STUDIES TO DETERMINE THE EFFECTS
OF POLYHALOGENATED ENVIRONMENTAL CONTAMINANT S ON
AVIAN SPECIES

By

Debra Curtis Powell

A THESIS

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

MASTER OF SCIENCE

Department of Animal Science

1995

ABSTRACT

EGG INJECTION STUDIES TO DETERMINE THE EFFECTS
OF POLYHALOGENATED ENVIRONMENTAL CONTAMINANTS 0N
AVIAN SPECIES
By

Debra Curtis Powell

Great Lakes waterbird populations have experienced less than expected egg
hatchability and greater than expected developmental abnormalities. Such deleterious
effects have been attributed to polyhalogenated diaromatic hydrocarbons such as
polychlorinated biphenyls (PCBS), polychlorinated dibenzo-p—dioxins (PCDDS), and
polychlorinated dibenzofurans (PCDFS). PCBS are of primary concern since they are
present in significant quantities in the environment. Specific PCB congeners, 3,3' ,4,4' ,5-
pentachlorobiphenyl (IUPAC #126), 3,3',4,4'-tetrachlorobiphenyl (IUPAC #77), and
2,3,3',4,4'-pentachlorobiphenyl (IUPAC #105), were injected (singly and in combination)
into the yolks of White Leghorn chicken (Gallus domesticus) eggs prior to incubation.
In a subsequent experiment, congener 126 was injected into chicken and double-crested
connorant (Phalacrocorax aun’tus) eggs prior to incubation. Finally, chicken eggs were
injected with one of four concentrations (0.001, 0.01, 0.1, and 1.0 egg equivalent) of an
extract derived from double-crested connorant eggs collected from Green Bay, WI. PCB-
induced increases in mortality were observed in all experiments. Incidences of
abnormalities, body weights, and relative weights of the brain, bursa, heart, liver, and

Spleen were also significantly affected.

To my husband, Jon
and
my parents, David and RuthAnn Curtis

iii

ACKNOWLEDGEMENTS

I would like to thank my major professor, Dr. Steven Bursian, for his assistance,
encouragement, and friendship throughout my Masters program. Next, I would like thank
the rest of committee members: Dr. Richard J. Aulerich for his support, advice, and help
with necropsies, as well as his good humor, Dr. John P. Giesy, for his willingness to assist
me with analysis of data and his critical review of this thesis, Dr. Richard Balander for his
advice regarding embryo development and incubation, and Dr. Ken Stromborg for funding
the project and providing cormorant eggs.

I am very grateful to my fellow graduate student, Janelle Restum, for her
friendship and willingness to help me early in the morning and into the night on the
cormorant experiment as well with all necropsies. I would like to thank Angelo
Napolitano for his help in making sure the incubators kept running, providing my chicks
with good housing, and for helping with necropsies. I would like to thank the rest of my
necropsy team, Chris Bush, Phil Summer, and Lorin Lewis.

Finally, I would like to thank my entire family (parents, grandparents, and in-laws)
for their love and support. In particular, I would like to thank my dad, David Curtis, for
making a stand which made injecting easier and more steady and my mother, RuthAnn
Curtis, for her assistance in proofreading this thesis. I would especially like to thank my
husband, Jon, for his love and patience throughout my masters program.

iv

TABLE OF CONTENTS

PAGE

LIST OF TABLES ................................................................................... vii
LIST OF FIGURES ................................................................................ xiii
LIST OF ABBREVIATIONS ...................................................................... xiv
GENERAL INTRODUCTION ..................................................................... 1
BACKGROUND AND LITERATURE REVIEW ............................................ 3
Great Lakes Contamination ................................................................ 4
Polychlorinated Biphenyls .................................................................. 5
Waterbird Effects ............................................................................ 13

Egg Injections and Incubation .............................................................. 18
References ................................................................................ 23
CHAPTER 1 .......................................................................................... 29
Abstract ................................................................................... 30
Introduction .................................................................................. 31
Materials and Methods ..................................................................... 32
Results ..................................................................................... 37
Discussion ................................................................................ 64
Summary and Conclusions ................................................................ 73
References ................................................................................ 75
CHAPTER 2 .......................................................................................... 79
Abstract ..................................................................................... 80
Introduction ................................................................................ 81
Materials and Methods .................................................................... 82
Results ......................................................................................... 85
Discussion ................................................................................... 100
Summary and Conclusions ............................................................... 105
References ................................................................................... 107

CHAPTER 3 ........................................................................................ 110

Abstract ................................................................................. l 11
Introduction .............................................................................. 1 12
Materials and Methods .................................................................... 114
Results ....................................................................................... 116
Discussion ............................................................................... 125
Summary and Conclusions ............................................................... 128
References .................................................................................. 130
APPENDIX ......................................................................................... 133

vi

TABLE

10.

11.

LIST OF TABLES

PAGE
Effect of congener 126 injected into the yolks of White Leghorn
chicken eggs on day 0 of incubation on lutchability ........................... 40
Effect of congener 77 injected into the yolks of White Leghorn
chicken eggs on day 0 of incubation on hatchability ........................... 40
Effect of congener 105 injected into the yolks of White Leghom
chicken eggs on day 0 of incubation on hatchability ........................... 41
Effect of congener combinations injected into the yolks of White
Leghorn chicken eggs on day 0 of incubation on hatchability ................ 41
LDsoS of congeners 126, 77, and 105 injected into the yolks of
White Leghorn chicken eggs on day 0 of incubtion .......................... 42
Abnormalities caused by congener 126 in White Leghorn
chicken embryos ................................................................. 44
Abnormalities caused by congener 77 in White Leghorn
chicken embryos ................................................................... 45
Abnormalities caused by congener 105 in White Leghorn
chicken embryos ................................................................... 46
Abnormalities caused by congener combinations in White
Leghom chicken embryos ......................................................... 47
Effect of congener 126 on the body weights of White
Leghorn chickens through three weeks of age .................................. 48
Effect of congener 77 on the body weights of White
Leghorn chickens through three weeks of age .................................. 48

vii

26.

27.

28.

29.

30.

31.

32.

33.

34.

35.

36.

37.

38.

39.

Effect of congener 105 on the actual and relative brain
weights of three-week-old White Leghorn chicks .............................. 56

Effect of congener 105 on the actual and relative bursa
weights of three-week-old White Leghorn chicks .............................. 56

Effect of congener 105 on the actual and relative heart
weights of three-week-old White Leghorn chicks .............................. 57

Effect of congener 105 on the actual and relative liver
weights of three-week-old White Leghorn chicks .............................. 58

Effect of congener 105 on the actual and relative spleen
weights of three-week-old White Leghorn chicks .............................. 59

Effect of congener 105 on the actual and relative testes
weights of three-week-old White Leghorn chicks .............................. 60

Effect of congener combinations on the actual and relative
brain weights of three-week-old White Leghorn chicks ....................... 61

Effect of congener combinations on the actual and relative
bursa weights of three-week-old White Leghorn chicks ....................... 61

Effect of congener combinations on the actual and relative
heart weights of three-week-old White Leghorn chicks ........................ 62

Effect of congener combinations on the actual and relative
liver weights of three-week-old White Leghorn chicks ........................ 62

Effect of congener combinations on the actual and relative
spleen weights of three-week-old White Leghorn chicks ...................... 63

Effect of congener combinations on the actual and relative
testes weights of three-week-old White Leghorn chicks ....................... 63

Effect of injection of congener 126 into the yolks of double-
crested cormorant eggs on day 0 of incubation on hatchability ............... 87

Effect of injection of congener 126 into the yolks of White
Leghorn chicken eggs on day 0 of incubation on hatchability ................ 88

ix

40.

41.

42.

43.

45.

46.

47.

48.

49.

50.

51.

52.

53.

LDSOS for congener 126 injected into the yolks of White Leghorn
chicken or double-crested cormorant eggs on day 0 of incubation ........ 88

Effect of congener 126 on the hatch body length and body
weight of double-crested cormorants ............................................ 90

Effect of congener 126 on the hatch body weights of White
Leghorn chickens .................................................................... 90

Effect of congener 126 on the actual and relative brain weights
of hatching double-crested cormorants ........................................... 91

Effect of congener 126 on the actual and relative bursa weights
of hatching double-crested cormorants ........................................... 91

Effect of congener 126 on the actual and relative heart weights
of hatching double-crested cormorants .......................................... 92

Effect of congener 126 on the actual and relative liver weights
of hatching double-crested cormorants .......................................... 93

Effect of congener 126 on the actual and relative spleen weights
of hatching double-crested cormorants .......................................... 94

Effect of congener 126 on the actual and relative brain weights
of hatching White Leghorn chickens ............................................. 95

Effect of congener 126 on the actual and relative bursa weights
of hatching White Leghorn chickens ............................................. 96

Effect of congener 126 on the actual and relative heart weights
of hatching White Leghorn chickens ............................................. 97

Effect of congener 126 on the actual and relative liver weights
of hatching White Leghorn chickens ............................................. 98

Effect of congener 126 on the actual and relative Spleen weights
of hatching White Leghorn chickens ............................................. 99

Concentrations of selected polyhalogenated aromatic hydrocarbons
and TEQ in cormorant egg extract .............................................. 118

54.

55.

56.

57.

58.

59.

61.

62.

63.

Effect of injection of cormorant egg extract into the yolks of White
Leghorn chicken eggs on day 0 of incubation on hatchability ............... 119

Effect of cormorant egg extract on the body weights of White
Leghorn chickens through three weeks of age ................................. 119

Effect of cormorant egg extract on the actual and relative brain
weights of three-week-old White Leghorn chicks ............................ 120

Effect of cormorant egg extract on the actual and relative bursa
weights of three-week-old White Leghorn chicks ............................ 120

Effect of cormorant egg extract on the actual and relative heart
weights of three-week-old White Leghorn chicks ............................. 121

Effect of cormorant egg extract on the actual and relative liver
weights of three-week-old White Leghorn chicks ............................. 121

Effect of cormorant egg extract on the actual and relative spleen
weights of three-week-old White Leghorn chicks ............................. 122

Effect of cormorant egg extract on the actual and relative comb
weights of female three-week-old White Leghorn chick ..................... 123

Effect of cormorant egg extract on the actual and relative comb
weights of male three-week—old White Leghorn chicks ...................... 123

Effect of cormorant egg extract on the actual and relative left
testes weights of three-week-old White Leghorn chick ...................... 124

Effect of cormorant egg extract on the actual and relative right
testes weights of three-week—old White Leghorn chicks ..................... .124

xi

APPENDIX TABLES

1.

Concentrations of polychlorinated dibenzo-p—dioxins (PCDDS) in
cormorant egg extract ............................................................ 134

Concentrations of polychlorinated dibenzofurans (PCDFS) in
cormorant egg extract ............................................................ 135

Concentrations of non-ortho—polychlorinated biphenyls in
cormorant egg extract ............................................................. 136

Concentrations of mono-ortho—polychlorinated biphenyls in
cormorant egg extract ............................................................. 136

Concentrations of organochlorine pesticides in cormorant egg extract. . .. 137

Concentrations of polychlorinated biphenyls in cormorant egg extract... 138

xii

LIST OF FIGURES

FIGURE PAGE
1. The structures of PCBS and TCDD ................................................ 8
2. The structure of three coplanar PCBS .............................................. 9
3. Effect of PCB congeners 126, 77, and 105 on mortality ..................... 43
4. Effect of congener 126 on the mortality of White Leghorn
chicken and double-crested cormorant embryos when exposed
prior to incubation ................................................................. 89

xiii

AHH
Ah-r
EROD
EVD
FWS
LOAEL
MSU
NOAEL
PCB
PCDD
PCDF
PHDH
TCDD
TCDD-EQ
TEF
TEQ
WQC

LIST OF ABBREVIATIONS

Aryl hydrocarbon hydroxylase

Ah receptor

Ethoxyresorufin 0-deethylase

Embryo viability detector

US. Fish and Wildlife Service

Lowest observable adverse effect level
Michigan State University

No observable adverse effect level
Polychlorinated biphenyl

Polychlorinated dibenzo-p-dioxin
Polychlorinated dibenzo furan
Polyhalogenated diaromatic hydrocarbon
2,3,7,8-tetrachlorodibenzo-p—dioxin
TCDD-equivalents (H4IIE bioassay derived)
Toxic equivalency factor

TCDD toxic equivalents (chemical analysis)
Water quality criteria

xiv

GENERAL INTRODUCTION

The Great Lakes have been contaminated as a result of industrial activities
throughout the region. These contaminants are threatening many wildlife species that
reside in the Great Lakes basin. In particular, fish-eating colonial waterbirds, such as the
double-crested cormorant (Phalacrocorax auritus) , have experienced less than expected
hatchability of eggs with an accompanying greater than expected incidence of
developmental abnormalities in both hatchlings and unhatched embryos (Fox et al.,
1991b). Such deleterious effects have been primarily attributed to polyhalogenated
diaromatic hydrocarbons (PHDHS) such as polychlorinated biphenyls (PCBS),
polychlorimted dmenzo-p-dioxins (PCDDS), and polychlorinated dibenzofurans (PCDFS).
PCBS are of primary concern among these toxicants since they have been shown to exist
in significant quantities in the environment (Fox et al., 1991a; Tillitt et al., 1992).
Conclusions about the cause/effect relationship between exposure to chemicals such as
PCBS and the occurrence of deleterious effects have, thus far, been based on field
observations of aquatic birds living near industrial areas. A controlled laboratory study
has been needed to substantiate the link between contaminant exposure and effects
observed in the field. Thus, the hypothesis tested by this project is that environmentally-

derived contaminants are responsible for the decline in egg hatchability and increase in

2

teratogenic effects observed in colonial fish-eating waterbirds.

The first objective of this project was to determine the effects of 3,3',4,4'-
tetrachlorobiphenyl (IUPAC #77), 3,3',4,4',5-pentachlorobiphenyl (IUPAC #126), and
2,3,3',4,4'-pentachlorobiphenyl (IUPAC #105) on developing White Leghorn chicken
(Callus domesticus) embryos. These PCB congeners are considered to be among the most
toxic congeners, and they constitute a significant proportion of the PCB congeners found
in the environment (Yamashita et al., 1993).

The second objective involved determining the difference in species sensitivity
between the White Leghorn chicken and the double-crested cormorant to the most toxic
of these PCB congeners, congener 126. The chicken has been widely used in laboratory
toxicity studies and is an excellent species for egg injection Studies since successful
incubation and hatching conditions have been established (McLaughlin et al., 1963;
BrunstrOm, 1989, 1990). It would be ideal if the chicken could serve as an animal model
for wild avian species like the cormorant in PHDH toxicity studies.

The third objective was to determine the effects of environmentally—derived
contaminants on the chicken embryo. Environmentally—derived contaminants were
extracted from cormorant eggs collected in 1988 from Green Bay, Wisconsin. Because
these contaminants have undergone weathering in the environment and metabolism in the
food chain, the mixture injected into the eggs was a more accurate representation of what

the animals encounter in the field.

BACKGROUND AND LITERATURE REVIEW

Great Lakes Contamination

The Great Lakes basin has become polluted by industrial activity for over a
century. Despite cleanup efforts in recent decades, the Great Lakes still face the problem
of persistent toxic contaminants. Large populations, heavy industry, and the fact that the
Great Lakes have a slow flushing rate for contaminants make this region particularly
susceptible to these persistent contaminants. For example, it takes 182, 106, and 21 years
for water entering Lakes Superior, Michigan, and Huron, respectively, to be flushed out
again. The primary concern with these contaminants is their effect on aquatic organisms,
since some of these compounds bioaccumulate. Phytoplankton, in the process of filtering
for nutrients, scavenge persistent contaminants. They in turn are eaten by zooplankton
which are preyed upon by fish which in turn are consumed by piscivorous birds and
mammals (Hileman, 1988).

In areas of heavy pollution, there has been a high incidence of tumors, goiters,
lack of secondary sex characteristics, and high embryo mortality in fish which are believed
to be a result of contamination in the Great Lakes (Hileman, 1988). There is evidence
that contaminated fish from the Great Lakes have deleterious effects on piscivorous
mammals and birds. Mink (Mustela vison) experienced reproductive failure and/or high
kit mortality when fed fish from the Great Lakes (Aulerich et al., 1971). Fish-eating
waterbirds residing in the Great Lakes have also experienced reproductive failure in

addition to teratogenic effects (Gilbertson, 1974; Weseloh et al., 1983; Fox et al., 1991a).

Polychlorinated Biphenyls

Polychlorinated biphenyls (PCBS) are a group of chlorinated aromatic organic
compounds which are of major concern due to their potential toxicity and also their wide
distribution and persistence in the environment (Tanabe, 1988). PCBS are one group of
compounds that are still found in greater concentrations than the concentrations specified
in water quality agreements established between the United States and Canada in the
19705. In 1986, the concentration of total PCBS in the Great Lakes ranged from 0.337 to
1.8 ng/L. These concentrations are much greater than the water quality criteria (WQC)
for six of the major wildlife species studied in the Great Lakes (Ludwig et al., 1993). A
WQC for total PCBS of 0.001 ng/L has been suggested to protect the most sensitive of the
wildlife species affected by PCBS (Ludwig et al., 1993). Concentrations of PCBS and
DDE in the eggs of piscivorous colonial waterbirds in Canada indicated that the Great
Lakes basin has the greatest concentration of PCBS and DDE in the country (Vermeer and
Peakall, 1977).

PCBS, commercially introduced in 1929, were widely used in industrial products.
Commercial mixtures of PCBS, such as the Aroclor' mixtures sold by Monsanto Chemical
Company, were prepared by the chlorination of biphenyls and sold by the amount of
chlorine they contained. For example, Aroclor' 1260 contained 60% chlorine by weight
(Safe, 1991).

PCBS are chemieally stable. If they contain more than four chlorine atoms they are

not affected by boiling with sodium hydroxide or nitric acid and are nonflammable

6
(Reynolds, 1969). They also have little water solubility (Huckins et al., 1988) and are

good electrical insulators. Due to these properties, they were used for a variety of
purposes, most notably as heat transfer fluids in capacitors and transformers, and as flame
retardants. In 1966, they were first identified in the environment. They are quite possibly
one of the largest groups of environmental pollutants due to widespread use and careless
disposal methods (Richardson et al., 1982). U.S. production of PCBS ended in 1977
(Safe, 1991).

Environmental contamination by PCBS occurs through a variety of means. Some
primary routes of contamination are leaks or spills from landfills, Storage facilities, or
equipment utilizing PCBS, and incinerator or engine emissions (Richardson et al. , 1982).
It lms also been suggested that PCBS may have been used in pesticides to increase their kill
life because PCBS can hold volatile ingredients (Reynolds, 1969). If this was done, it
represents a significant source of the PCBS found in the environment. Once in the
environment, PCBS can be transported via the air or the water. In water, PCBS are
primarily adsorbed to particles since they are insoluble in water (Haque and Schmedding,
1976).

Studies of PCBS are more complicated now that it has been discovered that
different congeners vary in toxicity and in their uptake and elimination by organisms.
Over 200 chlorobiphenyls can theoretieally exist by changing the number and location of
chlorine atoms on the biphenyl molecule. Approximately 100 of these can be found in
PCB technical preparations (Albro et al., 1981). However, after weathering in the

environment and biochemical alterations by organisms, the commercial mixtures first

7

released in the environment are rarely reflected in the environmental mixtures to which
humans and wildlife are now exposed. Thus, the individual congeners have increasingly
become the focus of research rather than the commercial PCB mixtures. Coplanar PCBS
are the most toxic of the PCBS. They are substituted in both para positions and two or
more of the meta positions, and are structurally similar to 2,3,7,8-tetrachlorodibenzo-p-
dioxin (TCDD), the most toxic of the man-made chemicals (Figure 1) (Safe, 1991).
TCDD is an unwanted by-product of chemical manufacturing and industrial processes.
TCDD binds to a cytosolic receptor called the Ah receptor (Ah-r), and it is this binding
event that is thought to elicit the effects associated with TCDD and structurally similar
”dioxin-like" compounds. The compound and receptor form a ligand complex that can
interact with specific areas of the genome and lead to the induction of gene transcription
(Safe, 1991).

Three of these coplanar PCBS in order of decreasing toxicity are 3,3',4,4',5-
pentachlorobiphenyl (IUPAC# 126), 3,3',4,4'-tetrachlorobiphenyl (IUPAC# 77), and
3,3',4,4',5,5'-hexachlorobiphenyl (IUPAC# 169) (Figure 2). All have been shown to be
present in significant quantities in commercial PCB preparations (Tanabe et al., 1987).
Each has been found in humans, marine and terrestrial mammals and birds, and marine
and freshwater fish (Tanabe, 1988). When a variety of animal species were tested for
residual concentrations of these PCBS, congener 77 was found in the greatest concentration

followed by congeners 126 and 169 (Tanabe et al. , 1987).

orcho

meta 3 2 2' 3'

5 6 6t 5t

Polychlorinated Biphenyl

 

2,3,7 ,8-tetrachlorodibenzo-p—dioxin

Figure l. The structures of PCBS and TCDD.

Cl Cl

cr© CG).

c:

3,3',4,4',S-pentachlorobiphenyl

3,3',4,4'-tetrachlorobiphenyl

or or '
.. o o ..
C1 C1
3,3',4,4',5,5'-hexachlorobiphenyl

Figure 2. The structure of three coplanar PCBS.

10

Mono-ortho PCB congeners, chlorinated in one of the ortho positions in addition
to the para and meta positions required of coplanar congeners, can also bind to the Ah-r.
They are often present in greater environmental concentrations than non-ortho congeners
(Brunstrom, 1990). Congener 105 (2,3,3‘,4,4'-pentachlorobiphenyl), while less toxic
than the coplanar congeners, may be responsible for a significant portion of the toxicity
attributed to PCBS because of its concentration in the environment (Kubiak et al., 1989;
Smith et al., 1990; Yamashita et al., 1993).

In accessing complex environmental mixtures, toxic equivalency factors (TEFS)
have been developed to evaluate these mixtures based on a common measure of toxicity.
The intent is to compare the toxicity of a particular compound or mixture of compounds
to a standard compound, which is usually TCDD. The test compound or mixture must
have a similar mode of action as the standard, thus TEFS can be derived for dioxin-like
compounds when the standard is TCDD. The standard is given a TEF of 1.0 and all other
compounds generally have TEFS less than 1. To calculate the TEF of a compound, the
endpoint or measured response must be known or determined for TCDD and then
determined for the test compound. The TEF for the compound can then be obtained by
dividing the dose which results in the response at a Specified magnitude for TCDD by the
dose for the same response for the test compound. If a variety of endpoints are measured,
one ean obtain a range of TEFS for a particular compound (EPA, 1991). The toxicity of
a mixture of compounds can be expressed in toxic equivalents ('TEQs) by multiplying the
TEF of each compound in the mixture by the measured concentration of each compound

in the total mixture (Safe, 1991; Bosveld et al., 1992). The toxicity of environmental

11

extracts can also be tested using the H4IIE bioassay. TCDD toxic equivalents (TCDD-
EQS) are determined by comparing the hepatic enzyme induction in extract-treated cells
with TCDD-treated cells (Tillitt et al., 1991).

Samples from double-crested cormorant (Phalacrocorax auritus) eggs indicate that
PCB congeners 126, 77, and 105 account for over 90% of the TEQS associated with the
egg (Yamashita et al., 1993). When fish (salmon from Lake Michigan) were examined,
congeners 126, 77, 105, and 118 (2,3',4,4',5-pentachlorobiphenyl) accounted for 92% of
the TEQS in the fish (Williams et al., 1992).

PCBS are known to eause a variety of effects that include body weight loss, atrophy
of the thymus, liver injury, skin cell disorders (eg. chloracne in humans), reproductive
dysfunction, depressed immune response, and teratogenic effects. In addition, some PCBS
are believed to be carcinogenic and possibly mutagenic in some species (Safe, 1989).
Some studies (Jacobson et al., 1993) have also shown intellectual deficits in children as
a result of in utero and/or lactational exposure. Laboratory studies with PCBS have
demonstrated many toxic effects in avian species. Liver lesions, hydropericardium,
shortened beak, microphthalmia, and subcutaneous edema have been observed in PCB-
exposed chick embryos (Brunstrom, 1988). Dioxin—like compounds have also been shown
to decrease the number of B-cells in the bursa of Fabricius (Andersson et al., 1991).
Besides the previously mentioned effects, dioxin-like compounds are known to be strong
inducers of cytochrome P-450 1A1 and 1A2 (Brunstrom, 1991; Sanderson et al, 1994) as
indieated by increased hepatic microsomal ethoxyresorufin O—deethylase (EROD) activity.

There have been significant positive correlations between TCDD concentrations in the

12

environment and hepatic EROD activity in blue heron chicks (Bellward et al., 1990).

There are differences in sensitivity to the toxic effects of PCBS among species. Of
the avian species, the White Leghorn chicken (Callus domesticus) is one of the most
sensitive. Turkeys (Meleagris gallopavo) are approximately 50 times less sensitive to
congener 77 than chickens when this congener is injected into the yolk (Brunstrom and
Lund, 1988). Pheasant (Phasianus colchicus), goldeneye (Bucephala clangula), black-
headed gull (Larus ridibundus), mallard (Anas platyrhynchos), goose (Amer anser), and
herring gull (Law-us argentarus) embryos are also much less sensitive to congener 77 than
domestic chicken embryos (Brunstrom and Reutergardh, 1986; Brunstrom, 1988). These
species differences may be due to differences in the concentration of or affinity for the Ah-
r or differences in the metabolism of PCBS. Another explanation for such differences in
toxic effects could be the greater genetic variation in wild species when compared to
domestic species. (Brunstrom and Reutergardh, 1986).

Nonetheless, the domestic chicken is often used as a laboratory model for the toxic
effects of compounds to which wild avian species are exposed. The chicken was used as
a model Species for colonial waterbirds in a Study where carp from Saginaw Bay was fed
to adult hens. Eggs laid by these hens had a dose-dependent increase in the incidence of
teratogenic effects similar to the effects seen in colonial waterbird populations (Summer,

1992).

13

Waterbird Effects

As indicated previously, some populations of fish-eating waterbirds in the Great
Lakes have experienced increased embryo mortality, changes in reproductive behavior,
increased incidence of abnormalities (including edema), histopathological anomalies,
suppression of the immune system, and changes in biochemical homeostasis. These avian
populations have been located primarily in areas where contamination is greatest (Tillitt
et al., 1991). GLEMEDS, Great Lakes embryo mortality, edema, and deformities
syndrome, is an acronym often used to refer to the effects observed in these waterbird
populations (Gilbertson et al., 1991). The developmental abnormalities include club feet,
missing eyes, defective feathering, and bill deformities (Yamashita et al., 1993). Other
effects include enlargement, necrosis, and porphyria of the liver, as well as subcutaneous,
peritoneal, and pericardial edema. These symptoms are also characteristic of a poultry _
condition called chick edema disease which is caused by organochlorine compounds that
are structurally Similar to TCDD. Clinical signs of this disease include reduction in body
weight gains, unsteady gait, labored breathing, ruffled feathers, and increased morbidity.
Observed lesions include hydropericardium, subcutaneous and peritoneal edema, swollen
and pale kidney, and swollen liver (Gilbertson et al., 1991).

Colonial fish-eating birds are often used as indicators of environmental
contamination. Significant amounts of environmental contaminants like PCBS can be
found in these birds Since they bioaccumulate, and their diet consists almost entirely of fish

and other food items which bioconcentrate contaminants. As an illustration, PCBS can be

14

bioconcentrated in the eggs of these birds by as much as 2.5 x 107 times the concentration
of PCBS in the water (Fox et al., 1991b).

In assessing the effects of environmental contamination on colonial waterbirds,
reproductive outcome (the production of viable offspring) is the primary monitoring
method. Of the species of colonial fish-eating waterbirds, herring gulls, Forster's terns
(Stemaforsten’), and double-crested cormorants have been the most studied species (Fox

et al., 1991b).

Herring Gulls (Larus argentatus)

Studies of herring gulls in the Great Lakes have provided information on
contamination trends (Gilman et al., 1977; Bishop and Weseloh, 1990). Lake Ontario and
Lake Michigan herring gulls are the most contaminated with TCDD and PCBS while Lake
Superior herring gulls are the least contaminated (Bishop and Weseloh, 1990). An
association was found between residual concentrations of contaminants in the eggs and
embryo lethality. The compounds identified as the primary contaminants in Lake Ontario
eggs were PCBS and mirex (Gilman et al., 1977).

A study with extracts of organochlorine compounds from herring gull eggs injected
into uncontaminated gull eggs produced no increase in mortality over controls (Gilman et
al., 1978). That study suggested either that organochlorine compounds in eggs are not the
direct cause of the high embryo mortality observed in Lake Ontario or that the effect of
contaminants may occur prior to laying of the egg in the form of alterations in genetics or

Structure of the egg.

15
Studies on gulls from Lake Ontario have Shown that abnormal behavior on the part

of the adult can contribute significantly to egg loss and subsequent low reproductive
outcome. Egg exchange experiments were performed between colonies on Prince Edward
Island and Lake Ontario which indicated that adults in Lake Ontario colonies incubating
eggs from Prince Edward Island had no better success than Lake Ontario gulls incubating
their own eggs. These birds showed an obvious lack of nest defense and attentiveness
(Fox et al., 1978). Inattentiveness leads to significant cooling and overheating of eggs
which results in increased egg lethality. Nests in Lake Ontario were unattended longer

than nests in less contaminated areas (Bishop and Weseloh, 1990).

Forster's Terns (Stema forsten)

In a 1983 study, Forster's tern eggs were collected from Green Bay, WI, a Lake
Michigan colony. Birds hatched from this contaminated region had relative liver weights
(% body weight) that were, on average, 26% greater than relative liver weights of birds
from an uncontaminated area. PCB congeners 105 and 126 accounted for over 90% of the
total PCB concentration. From this study, contamination was strongly suggested as being

the eausal factor in decreased reproductive outcomes (Kubiak et al. , 1989).

Double-Crested Cormorants (Phalacrocorax aun’tus)
It has been suggested that cormorants are more sensitive to some contaminants than
other species (Fox et al., 1991c). This species is long-lived and entirely piscivorous,

yellow perch being a major prey item (Weseloh et al., 1983). It is also an abundant

16

species which breeds throughout the Great Lakes region. Thus, it has been extensively
Studied as a bioindicator for contamination (Fox et al. , 1991c).

These birds experiemd a population decline in the Lake Huron region during the
1960s (Ludwig, 1984). Their colonies were still small in 1972, and they were unable to
reproduce successfully due to egg loss and breakage (Weseloh et al., 1983). This was
primarily attributed to p,p'-DDE which caused eggshells to become thinner (Peakall,
1988). Egg concentrations of DDE were 14.5 :1; 5.6 ppm and PCB concentrations were
23.8 :1: 9.6 ppm. The cormorant populations began to recover in the late 1970S and their
numbers are elevated over those reported in the early 1970s (Scharf and Shugart, 1981;
Weseloh et al., 1983; Ludwig, 1984).

Cormorant chicks from the upper Great Lakes region were studied from 1986-1989
and were found to have the following developmental abnormalities: crossbill, club foot,
hip dysplasia, dwarf appendages, subcutaneous edema, unabsorbed yolk-sac
(gastroschisis), and deformed as well as diseased eyes. These abnormalities have been
correlated with concentrations of PCBS in cormorant eggs (Tillitt et al., 1991). Since these
birds have been increasing in abundance and distribution, neither the quantity of food nor
its nutritioml quality are responsible for the previously described malformations in chicks.
There is also no evidence for the presence of infectious disease. Evidence in support of
a chemical contaminant as the causative agent includes reduced survival and growth of
embryos, elevated hepatic microsomal aryl hydrocarbon hydroxylase (AHH) activity,
congenital abnormalities, and increased liver to body weight ratios. The induction of AHH

activity indicates that polyhalogenated diaromatic hydrocarbons (PHDHS) may be the

17

chemical agent responsible (Fox et al., 1991a).

The bioaccumulation factor for PCBS from fish to cormorants is approximately 53
(Weseloh et al., 1983). The bioaccumulation factor is the concentration of PCBS in the
predator (cormorant) divided by the concentration of PCBS in the prey (fish). Fat in
cormorants was found to have concentrations of PCBS 10-100 times greater than marine
fishes and 100-1,000 greater than freshwater fishes (Scharenberg, 1991). Of the chemicals
analyzed in cormorant egg samples, PCBS (3.6—7.3 lag/g) and p,p‘-DDE (2.3-6.3 pg/g)
were found in the highest concentrations (Yamashita et al. , 1993). According to a Study
done in the late 19805, there was a signifieant correlation between cormorant egg mortality
and PCB concentrations in eggs taken from the same colony (Tillitt et al.,
1991). Cormorant eggs from Green Bay, WI have some of the greatest concentrations of
PCBS and other contaminants found in the Great Lakes region (Yamashita et al. , 1993).
PCBS in eggs from these birds were also found to be similar to congeners comprising
Aroclor' 1254 (Stalling et al., 1985). A significant, positive regression has been found
between toxic equivalents in cormorant chicks from different locations in Canada and the
induction of liver enzymes (Sanderson et al., 1994). This was supported by Similar
research in the Netherlands (Van den Berg et al., 1994)

In a 1973 study, contaminant concentrations were determined in cormorants from
South Dakota (Greichus et al., 1973). Average PCB concentrations in an adult cormorant
body, cormorant egg, and cormorant nestling were 4.6 ppm, 5.9 ppm, and 0.39 ppm,
respectively. Concentrations of PCBS in cormorants were 60 fold greater than those in

fishes. The PCB concentrations in eggs of cormorants also appeared to reflect the PCB

18

concentrations in the adult bodies.

Egg Injections and Incubation

The avian embryo is exposed to toxins present in the egg from fertilization through
hatching and even beyond Since the yolk is not entirely absorbed until several days after
hatching. Thus, the chicken egg is a practical way of observing toxicological effects (Fox
et al., 1991b). The injection of substances into an egg is a fairly inexpensive and simple
way to study toxicity. According to McLaughlin and associates (1963), the earliest
published work in this area appeared in the 18903 by Fere. Fere's interest was primarily
the teratogenic effects produced by chemicals. He injected test substances prior to
incubation and opened the eggs after a few days of incubation.

The toxicity of several test substances has been examined by injecting them into the
yolks of chicken eggs prior to incubation (McLaughlin et al. , 1963). The injection method
consisted of what ean be called a "yolk Shake”. An egg was given a quick shake prior to
injection to free the germinal disk such that it was able to float freely which reduced the
risk of injury to the embryo during injection. The needle was inserted perpendicular to
the long axis of the egg. Injections were done in volumes of up to 0.10 ml. Of the
substances injected, the ones with no or little toxicity were water, propylene glycol, corn
oil, peanut oil, isotonic saline solution, and isotonic glucose solution. Since eggs injected
with these solutions had near normal rates of hatching (95%, 95%, 90%, 90%, 90%, and

90%, respectively), these chemicals can be considered to be potential vehicles for testing

19

non-liquid substances.

Recent work involving egg injectiom has focused on this technique as an alternative
to the long and costly mammalian studies required for toxicity testing. Hashizume and
associates (1992) examined differences in toxicity among chemicals depending on whether
the Site of injection was the air cell or yolk. Injection of distilled water and physiological
saline produwd greater mortality when injected into the yolk as compared to air cell
(18.4% and 16.0% vs. 10.2% and 11.7%). However, sesame oil produced greater
mortality (26.3%) when injected into the air cell than it did when injected into the yolk
(13.3%). Some differences in malformations were noticed between the two injection sites,
but, as with mortality, the differences varied with the substance being tested. They also
noted greater mortality the earlier the time of administration.

A considerable amount of research involving the injection of PCBS into chicken
eggs has been done by Brunstr6m and associates. The primary obstacle to successful
injection studies is finding a proper vehicle (Brunstrfim and Orberg, 1982). One objective
of their research was to devise a vehicle which could be used for lipophilic substances such
as PCBS. After much experimentation, an emulsion of lecithin, peanut oil, and water was
chosen as the vehicle. They examined how well a lipophilic substance in this vehicle
would be taken up by the embryo. This was accomplished by using l“C-labelled 2,2',4,5'-
tetrachlorobiphenyl and assessing distribution by autoradiography (Brunstrdm and
Darnerud, 1983). The autoradiographs indicated a continual uptake of the injected PCBS
by the embryo. Thus, in this experiment, the emulsion vehicle provided early uptake of

the test substance. The injection procedure as well as the vehicle were apparently harmless

20

to the embryo.

The relative toxicity of PCB congeners 126, 77, and 169 has been determined in
chickens when injected into the yolk on day 4 of incubation (Brunstrom and Andersson,
1988). The most toxic of the three, congener 126, is approximately 50 times more potent
than the least toxic of the three, congener 169, based on mortality. All three were shown
to induce EROD activity and produced abnormalities such as subcutaneous edema,
hydroperieardium, microphthalmia, and Shorter upper beaks. In addition to the above
coplanar PCB congeners, Brunstrbm (1990) has demonstrated the toxicity of several mono-
ortho PCB congeners in chickens. He found that PCB congeners with an ortho chlorine
adjacent to a meta chlorine were more toxic than congeners with an ortho chlorine adjacent
to a meta hydrogen (Figure 1). These congeners also induced EROD in a dose-dependent
fashion suggesting that their toxicity is mediated via the Ah-r. Similar abnormalities were
also eaused by these congeners.

The toxic effects of congener 77 have been compared in chickens and turkeys
(Brunstrém and Lund, 1988). Mortality of both species was proportional to dose, but
turkeys were about 50 times less sensitive than chickens. Brunstrbm (1990) noted that
while non~orflro PCB congeners are extremely toxic in chick embryos, they seem to be less
toxic in embryos of many other avian species (Brunstrbm et al. , 1990). Brunstrbm (1989)
suggests that it is difficult to predict the effects of such compounds on reproduction in wild
avian embryos given the great differences between species.

While injectiom into the chicken egg are done routinely, injections into wild avian

eggs, particularly those of the cormorant, have been limited. Incubation techniques have

21

been a complicating factor in the success of such experiments. Larson (1991) injected
”undeveloped“ cormorant eggs from Bachelor's Island in Lake Winnipegosis (Manitoba,
Camda) with TCDD and PCB congener 126. Analysis of 10 uninjected eggs revealed a
total PCB concentration of 910 ppb. The same three doses (1.84 pg/g, 184 pg/g, and
18,400 pg/g) were used for both TCDD and PCB congener 126. The vehicle used was
1,4—dioxane. Chicken eggs were also injected with the same dose of TCDD and congener
126 (15,100 pg/g). All chicken eggs died within a week of injection. The hatching rate
of chicken eggs injected with the vehicle alone was 66.7% . Cormorant egg hatchability
was poor in all groups. Thus, Larson based analysis on embryos alive on days 7 and 23.
Cormorant mortality at any dose of congener 126 or TCDD was not significantly greater
than the vehicle on either day. Larson's results suggested that the NOAELs (no observable
adverse effect level) were the greatest doses of both TCDD and congener 126 (18,400
pg/g).

There has been limited success in the artificial incubation of cormorant eggs from
day 0. The incubation conditions for cormorant eggs in Larson's Study (1991) were
37.5 °C and 65.5% relative humidity. The relative humidity was raised to 67% on day 26
of incubation. Eggs were placed with their blunt end up and automatically rotated six
times a day. These conditions did not afford proper development of the aircell and as a
result only 3 % of the cormorant eggs hatched. Unincubated cormorant (Phalacrocorax
carbo) eggs from the Netherlands incubated at 375°C with 50% relative humidity for 24
to 25 days and 80% relative humidity for the last one to two days resulted in less than half

the eggs hatching (Van den Berg et al. , 1994). A more successful incubation procedure

22

for the double-crested cormorant (Phalacrocorax auritus) was developed by Sotherland
(unpublished data). The incubation temperature used was 37.2°C (99°F). Eggs were
placed horizontally and rotated automatically every two hours in addition to being rotated
by hand twice a day. Pipped eggs were transferred to a hatching room which was
maintained at 37°C and 70% relative humidity. Sotherland also noted that hatching
success was greater if the eggs had been incubated for about a week by the parents. This
has been supported by Sanderson et al. (1994), who obtained 77% hatchability of
cormorant eggs naturally incubated for two weeks by the parents prior to artifical
incubation (354°C, 46% relative humidity). However, natural incubation is not practical
for egg injection studies if one wants to determine the effects of compounds which would

be present from fertilization through hatch.

23
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CHAPTER 1

THE EFFECTS OF 3,3',4,4'-TETRACHLOROBIPHENYL,
2,3,3',4,4'-PENTACHLOROBIPHENYL,
AND 3,3 ' ,4,4' ,S-PENTACHLOROBIPHENYL ON DEVELOPING
CHICKEN EMBRYOS WHEN INJECTED
PRIOR TO INCUBATION

29

30
ABSTRACT

Great Lakes waterbird populations have experienced less than expected hatchability
of eggs and a greater than expected incidence of developmental abnormalities. Such
deleterious effects have been attributed to polyhalogenated hydrocarbons such as
polychlorinated biphenyls (PCBS). PCBS are of primary concern since they are present in
significant quantities in the environment. Specific PCB congeners, 3,3',4,4',5-
pentachlorobiphenyl (IUPAC #126), 3,3',4,4'-tetrachlorobiphenyl (IUPAC #77), and
2,3,3',4,4'-pentachlorobiphenyl (IUPAC #105), were injected (singly or in combination)
into the yolks of White Leghorn chicken (Gallus domesticus) eggs prior to incubation.
Teratogenicity was assessed in dead embryos and in hatchlings. Hatchlings were raised
for three weeks to assess body weight gain and mortality. At the end of the three week
period, chicks were subjected to necropsy and the brain, bursa, heart, liver, spleen, and
testes were removed and weighed. All three congeners caused increased mortality with
approximately 50% mortality occurring at 0.6, 4.9, and 5,592 rig/kg eg for congeners
126, 77, and 105, respectively. All three congeners also produced Significantly more
abnormalities than the vehicle. Chicks from PCB-injected eggs had lower body weights
at the second and third weeks of age. Congener 126 caused lower relative bursa weights,
congener 77 caused greater relative spleen weights and lower relative liver weights, and

all three congeners eaused relalive heart weights to be greater when compared to control.

3 1
INTRODUCTION

Years of industrial activity in the Great Lakes region have led to environmental
contamination which is threatening many wildlife species that reside in the Great Lakes
basin (Peakall, 1988). Fish-eating colonial waterbirds, such as double-crested cormorants
(Phalacrocorax auritus), have experienced reduced hatchability of eggs with a
concomitant greater rate of developmental abnormalities in both hatched and unhatched
chicks (Weseloh et al., 1983). Such deleterious effects have been primarily attributed to
aromatic hydroearbons such as polychlorinated biphenyls (PCBS), polychlorinated dibenzo-
p—dioxins (PCDDS), and polyhalogenated dibenzofurans (PCDFS). PCBS are of primary
concern among these industrial contaminants as they have been shown to exist in
Significant quantities in the environment (Tillitt et al., 1991; Yamashita et al. , 1993; Jones
et al., 1994).

The non-ortho PCB congeners, 3,3',4,4',5—pentachlorobiphenyl (IUPAC #126) and
3,3',4,4'-tetrachlorobiphenyl (IUPAC #77), have been shown to cause abnormalities and
lethality when injected into chicken eggs (Brunstrom and Darnerud, 1983; Brunstrom,
1988; Brunstrom and Andersson, 1988; Brunstrcm and Lund, 1988; Brunstrom, 1991).
These compounds are structurally similar to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)
and ean thus bind to the Ah-receptor (Ah-r) (Brunstrbm and Andersson, 1988; Brunstrbm
and Lund, 1988; Brunstrém, 1991). The toxicity of these compounds observed in chick
embryos is thought to be mediated through this Ah-r (Brunstrtim, 1990). Mono-ortho

congeners, such as 2,3,3',4,4'-pentachlorobiphenyl (IUPAC #105), can also bind to the

32

Ah-r and are present in greater concentrations in some commercial preparations of PCBS
than non-ortho PCBS (Brunstrdm, 1990). Congener 105 together with congener 126
accounted for more than 90% of the toxic equivalents (TEQS) in Forster's tern eggs
collected at Green Bay, WI in 1983 (Kubiak et al., 1989).

The objective of the present Study was to determine the effects of three
environmentally relevant PCB congeners (Kubiak et al., 1989; Yamashita et al., 1993)
when injected into the yolks of White Leghorn chicken (Gallus domesticus) eggs prior to
incubation. Yolk injection was chosen since the majority of environmental contaminants
are lipophilic and hence will be deposited into the yolk. Injection on day 0 of incubation
was significant since wild avian embryos are exposed to contaminants from the time of
fertilization, and laboratory exposure just prior to incubation ensures that the potential
adverse effects induced by contaminants during early development will be included in the

analysis of toxicity of the contaminant tested.

MATERIALS AND METHODS

Doses of PCB congeners 126, 77 , and 105 were chosen to bracket the smallest and
greatest concentrations detected in cormorant eggs collected from various sites in the Great
Lakes (Yamashita et al., 1993) as well as previously reported LDsos for seven-day-old
chicken embryos injected on day 4 (Brunstrom and Andersson, 1988). These three
congeners were also injected in all possible combinations at the smallest environmental

concentration detected in cormorant eggs collected from various Great Lakes sites

33

(Yamashita et al., 1993).

Doses and Design: The PCB congeners were purchased from AccuStandard (New
Haven, CT). Doses for congener 126 were 0.1, 0.3, 0.9, 2.7, or 8.1 ag/kg egg.
Congener 77 was injected at doses of 1.0, 3.0, 9.0, 27.0, or 81.0 lag/kg egg and congener
105 was injected at doses of 100, 300, 900, 2700, or 8100 ug/kg egg. The doses chosen
for the congener combinations were the lowest environmental concentrations detected in
cormorant eggs collected from five colonies around the Great Lakes. Eggs from
Tahquamenon Island in Lake Superior had the lowest concentrations of these congeners;
0.80, 0.98, and 110 ug/kg egg for congeners 126, 77, and 105, respectively (Yamashita
et al., 1993). There were four congener combinations: 105 plus 77, 105 plus 126, 77
plus 126, and 105 plus 77 plus 126. There were 20 eggs/treatment group per replicate and
three replications of each congener and combination of congeners. Due to factors
unrelated to the experiment, a fourth replicate was added to increase the sample size of
chicks exposed to congener 77. In addition to the dose groups, there were non-injected
and vehicle control groups in each replicate.

Vehicle: The injection vehicle was an emulsion of lecithin from egg yolk (Merck:
Darmstedt, Germany), peanut oil, and water. The lecithin was initially dissolved in
dichloromethane (Brunstrbm and Darnerud, 1983). The lecithin and peanut oil were
combined (1:10, w/v) and then placed in a waterbath (60°C) for approximately one hour
to allow the dichloromethane to evaporate. The lecithin/peanut oil (L/PO) solution was
autoclaved for 30 min and then mixed with sterile distilled water (2:3, v/v) as described

by Brunstrdm and Orberg (1982).

34

Dose Solutions: Dosing solutions were prepared from an initial stock solution of
PCB congener in L/PO. Congeners 126 and 77 (5 mg) readily dissolved in L/PO (10 ml)
following sonication for 10 min. Congener 105 was first dissolved in dichloromethane
since it was uncertain if 202 mg could be dissolved in 10 ml L/PO. Using a nitrogen
evaporator, dichloromethane was then evaporated. Appropriate proportions of stock
solution and UPO were combined to provide two ml total L/PO solution. Three ml sterile
distilled water were then added as previously described. New dosing solutions were
prepared for each replicate.

Egg preparation: Eggs from White Leghorn chickens were obtained from the
Michigan State University (MSU) Poultry Science Teaching and Research Center. All
eggs were candled to loeate the air cell and then weighed to determine the injection volume
(1 ul/g egg). After eggs were weighed and labeled, they were positioned horizontally and
placed in a cooler (58°F) until injection on the following day. Eggs were laid horizontally
overnight to allow the germ spot to float away from the site of injection thus reducing
potential injury during injection.

Injections: The injections were made prior to incubation (day 0) and the site of
injection was the yolk sac. Eggs were removed from the cooler and left horizontal until
after being injected. A 100,21 syringe (228 gauge) (Hamilton Company, Reno, NV) was
used for the injections. It was sterilized with 70% alcohol prior to insertion into the vial
and before injection into the egg. The needle length was marked at approximately 29 mm
to ensure injection into the yolk (this length was chosen after a series of injections of a

colored solution into the yolks of infertile eggs). The blunt end of the egg was surface

35
sterilized with 70% alcohol before a small hole was made in the shell over the air cell with

a sterile pin. After horizontal insertion of the needle, the solution was Slowly injected, and
the needle was left in for a moment following injection. Upon removal of the needle, the
hole was sealed with melted paraffin. Following injections, all eggs were placed in the
incubator.

Incubation: Eggs were incubated in a Petersime (Gettysburg, OH) poultry
incubator for up to 24 days with their blunt end up. Conditions in the incubator were
Standard for commercial operations, 37.5-37.7°C (99.5-99.75°F) with wet bulb readings
of 85-87°F to yield 65 % relative humidity. Eggs were automatieally rotated every two
hours. Embryo viability was determined by candling on days 4 and 11. On day 18,
viability was assessed with an embryo viability detector (EVD) which was provided by the
U.S. Fish and Wildlife Service (FWS). The EVD detects vibrations within the egg and
changes these Vibrations to sound waves which can be heard in headphones attached to the
EVD (Mineau and Pedrosa, 1986). Any non-viable eggs on days 11 or 18 were opened
to estimate the stage of development and to determine the presence of any anomalies.
Viable eggs were then transferred to hatching baskets in a second incubator with higher
humidity (70-75 %) than the incubator used for incubation prior to day 18.

Post-hatch: Upon hatching, each chick was weighed and uniquely identified with
a wing band. Chicks were also examined for abnormalities. Once chicks were sufficiently
dry, they were moved to a floor pen with pine and spruce wood chips (Pestell, Agri-
Products, New Hamburg, Ontario, Canada) and acclimated to food and water by dipping

their beaks in water and feed (Purina Chick Starter) which were both provided ad libitum.

36

All chicks were raised for a period of 21 days with body weights being recorded on a
weekly basis. In studies involving herring gulls, a three-week survival period was used
to estimate reproductive success since mortality of chicks was observed during the first
three weeks after hatch (Gilman et al., 1977). Any chick which was determined incapable
of acquiring food, water, or warmth on its own was euthanized.

Necropsies: At the end of the three weeks, chicks were killed by cervical
dislocation. The organs taken for weighing included the brain, bursa, heart, liver, spleen,
and testes.

Histopathology: The bursa, liver, Spleen, and testes from five (where possible)
three-week—old chicks per dose were saved in 10% formalin. Organs from the greatest
dose group of each congener were examined by a veterinary pathologist.

Data Analysis: All comparisons were made with respect to the vehicle control.
Due to the relatively low rate of hatching in the vehicle control group, comparisons with
the non-injected controls alone or in combination with vehicle controls was not considered
to be appropriate. Mortality was evaluated using a 2x2 contingency table and Bonferroni
Chi-Square ("infertiles' were considered early deads Since it was often difficult to
distinguish between the two). The same analysis was used to test the occurrence of
abnormalities. LD,o values were obtained using probit analysis. Analysis of body weight
was conducted using univariate analysis of variance for Split-plot repeated measures (Gill,
1986) followed by Dunnett's test for comparisons with control. Organ weights and
differences between left and right testes were analyzed using a one-way ANOVA and

Dunnett's test. Analyses of variance were conducted with the statistical software SAS

37

(SAS Institute Inc., 1992). The level of statistical Significance was 0.05 unless otherwise

stated.

RESULTS

Modality: The vehicle control group had statistically greater mortality than the
non-injected control group. Despite great mortality in the vehicle control group, all three
congeners (singly or in combination) produced significantly greater mortality than the
vehicle control at the greater doses (Tables 1-4). The LDsos (95% confidence limits)
(Table 5) were 0.6 lug/kg egg (0.4 - 0.7) for congener 126, 4.9 tag/kg egg (3.6 - 6.1) for
congener 77, and 5,592 ug/kg egg (2,256 - 7,637) for congener 105. Injection of
combinations containing congener 126 produced significantly greater mortality compared
to the vehicle. There was no statistically significant mortality post-hatch induced by any
of the congeners or combinations. Chicks that died after transfer to floor pens were
usually those that hatched late or were weak at hatch. These chicks occurred in both the
vehicle control and treated groups.

Teratogenicity: The types of morphological abnormalities observed included
edema; skull and brain abnormalities; crossed, missing, and shortened beaks; curled toes;
leg and hip abnormalities; and missing or small eyes. Congener 126 (Table 6) produced
Significantly more abnormalities at 0.9 rig/kg egg (22%) when compared to the vehicle
control group (7%). The two greatest doses (2.7 and 8.1 rig/kg egg) resulted in such high

mortality that Statistical assessment of abnormalities was not possible. The predominant

38

types of abnormalities observed in the 0.9 jug/kg egg dose group were miscellaneous
abnormalities which included unusually small embryos, embryos with abnormal head/jaw
area, or embryos with abnormal organs. Congener 77 (Table 7) produced significantly
more abnormalities at 27 ,ug/kg eg (15 %) than the vehicle control group (4%). Ninety
percent of the eggs injected with 81 ug/kg egg died by day 11, thus, the incidence of
abnormalities in surviving embryos was not statistically significant. The predominant
abnormalities which occurred in the 27 rig/kg egg group were beak deformities such as
crossed or shortened upper or lower beaks. Congener 105 (Table 8) caused greater
abnormalities relative to the vehicle control group (5 %) only at the greatest dose, 8100
lag/kg eg (22 %). The predominant abnormalities included unusually small embryos,
embryos with abnormal head/ jaw area, or embryos with abnormal organs. None of the
combimtions (Table 9) of congeners produced any significant differences in the occurrence
of abnormalities. Edema was the predominant type of abnormality observed in embryos
receiving the combination of all three congeners.

Body Weights: The greater doses of all three of the congeners tested produced
significantly lower chick body weights than the vehicle control (Tables 10—13). Congener
126 caused reduced body weights beginning at two weeks of age at doses of 0.9 and 2.7
rag/kg egg. Chicks hatching from eggs injected with 3 ug/kg egg of congener 77 were
Significantly lighter than vehicle controls at three weeks of age while chicks dosed with 9
rag/kg egg were significantly lighter than controls beginning at two weeks of age. A dose
of 300 ,ug/kg egg of congener 105 caused lower body weights at two weeks of age, 900

and 2700 lug/kg egg produced significantly lower body weights at three weeks of age, and

39

8100 lug/kg egg resulted in lower body weights after only one week post hatch. Congener
126 in combination with congener 105 and/or congener 77 resulted in lower body weights
beginning at two weeks.

Organ Weights: All organ weights are expressed as actual and relative weights.
Since body weights were affected, statements of Significance are in reference to relative
organ weights (% body weight). Congener 126 caused an increase in brain (Table 14)
weight and a decrease in weights of the bursa of Fabricius (Table 15) at the 0.9 rig/kg egg
dose This dose also caused an increase in heart (Table 16) weights when compared to the
vehicle control group. Congener 77 caused an increase in brain (Table 20) weight at 9.0
rig/kg egg. This congener caused an increase in heart (Table 22) weight and a decrease
in liver (Table 23) weight at a dose of 3.0 rag/kg egg. Spleen (Table 24) weights were
greater in the 9.0 lug/kg egg dose group than in the vehicle control. Congener 105 caused
an increase in brain (Table 26) and heart (Table 28) weights at 8100 jug/kg egg. There
were no significant differences in any of the lower dose groups. Only the combination of
congeners 105, 77, and 126 caused an increase in brain (Table 32) weight and a decrease
in the weight of the bursa of Fabricius (Table 33). Relative heart weight (Table 34) was
increased by the combination of congeners 77 and 126, as well as the combination of
congeners 105, 77, and 126. The combination of 105 and 126 was the only combination
treatment to affect the liver (Table 35) which was decreased in weight.

Histopaflwlogy: There were no significant histological lesions associated with PCB

treatment in any of the tissues examined.

4o

 

 

 

 

Dose Dose # dead!
1.4;; 0'1; nmo e #eggs % Mortality
Non-injected 0 7/60 11.7
0‘ 0 22/59 37.3
0.1 0.3 12/60 20.0
0.3 0.9 23/60 38.3
0.9 2.8 50/60” 83.3
2.7 8.3 59/60” 98.3
8.1 24.8 60/60” 100

 

  
 
   

' Vehicle (lecithin/pear“ oil and water ermlsion)
” Significantly differem from comol at p<0.05

Table 2. Effect of congener 77 injected into the yolks of White Leghorn chicken eggs on day 0 of

 

 

 

 

 

 

on hatchability. g fl
Dose Dose # dead/
um nmol/kgi # eggs % Mortality

Non-injected o 13/30 16.3
I o- o 32/80 40.0
1.0 3.4 21/80 26.3

3.0 10.3 29/80 36.3
9.0 30.8 54/80” 67.5

27.0 92.5 80/80” 100

81.0 277.4 80/80” 100

 

‘ Vehicle (lecithin/peanut oil and water ermlsion)
” Si differem comm! at . <00

41

    

1
Table 3. Effect ofcongener105 injectcdirtotheyolks ofWhite Leghornchickeneggs onday 0
ofrncubatrononhatchability. ,

  

   
 

 

 

 

 

 

0- 0 24/60 40-0
100 306 28/60 46.7
300 919 25/60 41.7
900 2,757 31/58 53.5
2700 8,271 31/60 51.7
8100 24,813 50/59” 84.8
" Vehicle (lecithin/pearnrt oil and water emulsion)
” Significamly different from control at p<0.05

 

 

 

Table 4. Effect of congener combinations injected into the yolks of White Leghorn chicken eggs
on day 0 of incubation on hatchability.

 

 

 

Total (PCBS) # dead/
Combination” nmol/kg egg Ill eggs % Mortality

Non-injected 0 8/58 13.8 J

0” 0 31/60 51.7

105/77 340.4 25/60 41.7

105/126 339.5 47/59° 79.7

77/126 5.9 50/60° 83.3

105/77/126 342.9 55l60° 91.7

 

' Doses in #8(nmol)kg eg for congeners 126, 77, and 105 are 0.80 (2.5), 0.98 (3.4), and 110

(337), respectively
” Vehicle (lecithin/peanut oil and water enulsion)

 

 

42

 

- L

Table 5. LDws of congeners 126, 77, and 105 injected into the yolks of White Leghorn chicken

 

 

 

 

 

 

 

eggs on day 0 oMon.
LDso 95% Confidence Limits
Congener ag(nmol)/kg egg Lower Upper
126 0.6 (1.8) 0.4 (1.2) 0.7 (2.3)
77 4.9 (16.8) 3.6 (12.4) 6.1 (21.0)
105r 5,592 (17,130) 2,256 (6,912) 7,637 (23,394)

 

 

 

 

 

43

 

 

 

 

 

 

 

 

 

 

 

 

 

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l x 10’ l x 10’ 1 x10‘
log(dose +1)

 

-- Congener 126 -o- Congener 77 + Congener 105]

 

Figure 3. Effect of PCB congeners 126, 77, and 105 on mortality. One was added
to each dose to allow inclusion of the vehicle control on a log scale. For
example, vehicle control (0 rig/kg egg) = 1 and the highest PCB congener
dose (8100 ag congener 105/kg egg) = 8101.

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Table 10. Effect ofcongener126 on the body weights ofWhite Leghorn chickem through three
weeks of age.

 

 

 

 

Dose 11‘ SE” Hatch‘ 1 Week 2 Weeks 3 Weeks
Vehicle 28 2.37 45.4 69.7 123.3 184.8
0.1 rig/kg 35 2.12 47.2 69.8 122.6 183.7
I 0.3 units 26 2.46 47.1 67.2 117.8 180.1
0.9 pg/kg 9 4.19 45.7 61.1 98.7‘ 146.7‘
2.7 pm 2 8.88 45.7 58.3 94.21I 142.8“
' Sanple size
” Standard error (Not appropriate for conpan'sons across time)
° Data expressed as mean (g)
" Significantly different from vehicle comrol at p<0.05 I I

 

—.-——.__ _ . __. __ .__________. ___— _ _. . __.- _ _ — __ _. _ _____—________——_—___“

Table 11. Effect ofcongener 77 on the body weighs ofWhite Leghorn chickens through three

 

 

 

 

,___g__ _____ _wceks 0f 383- _—
Dose 11‘ SE” Hatch° 1 Week 2 Weeks 3 Weeks
Vehicle 36 2.53 46.2 68.3 121.4 189.8
1 jag/kg 44 2.29 45.6 67.9 122.4 191.8
3 pg/kg 35 2.57 45.1 65.0 114.4 177.5“
9 ”fig 16 3.80 44.9 64.1 108.6‘1 168.3‘l
‘ Sample size
” Standard error (Not appropriate for comparisons across time)
° Data expressed as mean (g)
" Significamly differem from vehicle control at p<0.05
“

 

49

Table 12. Effect of congener 105 on the body weights of White Leghorn chickens through mree
weeks of age. I

SE”

 
  
    
      
 

  

1 Week 2 Weeks

   

 

   

Vehicle 2.32 74.8 132.0

            

    

100 AIS/kg 30 2.65 45.4 73.1 129.3 191.7

      
 

 

300 um 31 2.61 44.7 70.5 123.4:1 186.0“

 

     
 

900 ug/kg 19 3.33 43.7 70.3 124.0 183.1“

    
 

 

2700 #g/kg 20 3.24 44.8 69.8 123.4 187.1“

       
  
 

8100 rig/kg 7 5.49 42.7 60.9“ 103.5“ 153.3“

 

 
  

‘ Sanple size
” Standard error (Not appropriate for comparisons across time)
° Data expressed as mean (g)

“ Significamly differem from vehicle control at p < 0.05

 
   
   

Table 13. Effect of congener confirmations on the body weighs of White Leghorn chickem
_ _through three weeks of age.

Treatments SE” 1 Week 2 Weeks
Vehicle 3.38 . 69.8 124.5 190.2
105/77 2.88 . 70.1 124.8 192.9
105/126 5.52 . 60.7 104.8“ 161.3“
77/126 5.52 . 63.7 108.6“ 166.4“

105/77/126 8.28 . 56.4 96.6‘l 149.3“

‘ Sample size
” Standard error (Not appropriate for comparisons across time)
° Data expressed as mean (g)

 

50

 

 

 

 

   
    
 

 

 
  
    

Table 14. Effect of congener 126 on the actual and relative brain weigh of three-week-old White
Legrhom chicks.
Dose‘ n” Weiw % Body Weight‘
Vehicle 28 1.64 :1: 0.019 0.88 :1: 0.017
0.1 34 1.66 i: 0.018 0.90 a: 0.016
0.3 26 1.59 :1: 0.020 0.89 :t 0.018
0.9 9 1.51 1; 0.034“ 1.03 3: 0.031“
2.7 2 1.43 :1; 0.072“ 1.00 :t 0.065
‘ Doses expressed as rig/kg egg
” Sanmle Size

°Dataexpressedasmean i standarderror
“S' 'camlydifferemfromvehiclecontrolatp<0.05

Table 15. Effect of congener 126 on the actual and relative bursa weight of three-week-old

 

 

 

 

White Leghorn chicks.

I Dose” n” Weigfig)° % Body Weight“
Vehicle 28 1.07 i 0.052 0.572 :1: 0.0221
0.1 34 1.06 :1; 0.048 0.567 :1: 0.0201
0.3 26 1.05 1; 0.054 0.567 :1: 0.0229
0.9 9 0.62 i 0.096“ 0.414 :1; 0.0390“
2.7 2 0.56 i 0.196 0.389 :1: 0.0827

' Doses expressed as rig/kg egg

” Sample size

 

°Dataexpressedasmean j: standarderror
“S' 'camlydifferemfromvehiclecomrolatp<0.05 _

51

Table 16. Efl‘ectofcongmerlfiontheaaualandmlaiveheanweighofthree-week-old

 

 

 

 

White Leghorn chicks.
Dose‘ n” Weiglli)‘ % Body Weigh;
Vehicle 28 1.12 :1: 0.035 0.602 :1: 0.0155
0.1 34 1.12 :1: 0.032 0.604 a: 0.0141
0.3 26 1.14 :1: 0.037 0.624 :2 0.0161
0.9 9 1.24 :t 0.062 0.851 :1: 0.0274“
2.7 2 0.77 :1: 0.132“ 0.545 :t 0.0581
‘ Doses expressed as [4m egg
” Sanple size

°Dataexpressedasmean :1: standardermr
“ Significantly differem from vehicle control at p<0.05

[— .

 

 

 

 

 

Table 17. Effect of congener 126 on the actual and relative liver weight of three-week-old
White Leghorn chicks.
Dose“ n” Weigfig)° % Body We' 3.
Vehicle 28 4.95 :1: 0.158 2.65 a: 0.058
0.1 34 5.00 i 0.143 2.69 :1: 0.053
0.3 26 5.10 i 0.164 2.81 :1: 0.060
0.9 9 3.64 i 0.279“ 2.47 a: 0.103
2.7 2 4.25 :t 0.591 2.99 :t 0.217

 

‘Dosesexpressedasug/kgegg
”Sanplesize

° Data expressed as mean :t standard error

 

‘ mew-05

52

Table 18. Effect of congener 126 on the actual and relative spleen weigh of three-week-old

 

 

 

 

 

  

White Leghorn chicks.
Dose' n” Weigflg)c % Body Weight°
Vehicle 28 0.24 i 0.019 0.131 :1: 0.0091
0.1 34 0.27 i 0.017 0.147 t 0.0083
0.3 26 0.28 :1: 0.020 0.151 :1: 0.0095
0.9 9 0.26 i 0.033 0.174 :1: 0.0161
2.7 2 0.18 i 0.071 0.122 :t 0.0341
‘ Doses expressed as rig/kg egg
” Simple size

°Datae ressedasmeanistandarderror

 

 

Table 19. Effect of congener 126 on the actual and relative testes weigh of three-week-old

 

 

 

 

White Leghorn chicks.
Dose' n” WeigELg)‘ % Body Weight°
Vehicle 14 0.046 :t 0.0030 0.023 :1: 0.0015
0.1 20 0.045 :t 0.0026 0.023 :1: 0.0013
0.3 11 0.037 :1: 0.0034 0.020 :1: 0.0017
0.9 7 0.034 t 0.0043 0.023 :t 0.0022
2.7 0 - --
‘ Doses expressed as rig/kg egg
” Sample size

° Data expressed as mean i standard error

Table 20. Effect of congener 77 on the actual and relative brain weigh of three-week-old White

53

 

 

 

 

 

Leghorn chicks.
Dose‘ 11” Wow? % Body Weightc
Vehicle 36 1.62 :1: 0.019 0.86 :1: 0.015
1 44 1.62 :t 0.018 0.86 :1: 0.014
3 35 1.58 :1: 0.020 0.90 :1: 0.016
9 16 1.56 :1; 0.029 0.93 :1: 0.023“
‘ Doses expressed as rig/kg egg
” Sanple size

°Dataexpressedasmeanistandarderror

“ Significamly differeh from vehicle cohrol at p<0.05

 

 

 

 

Table 21. Effect of congener 77 on the actual and relative bursa weigh of three-week-old

 

 

 

 

 

White Leghorn chicks.
Dose' n” Weigh (g)° % Body Weightc
Vehicle 36 0.98 :t 0.050 0.510 1 0.0202
1 44 1.00 i 0.046 0.512 :t 0.0182
3 35 0.90 :t 0.051 0.495 :1: 0.0205
9 15 0.81 :t 0.078 0.466 :1: 0.0313
' Doses expressed as rig/kg egg
” Sample Size
°Dataexpmssedasmeanistandardgnor fl

 

54

Table 22. Effect of congener 77 on the actual and relative heart weigh 0f three-week-okl White

 

     
 
      
 
  

 

 

 

 

 
  
    

Leghorn chicks.
Dose' n” Weigh (g)° % Body Weighc
Vehicle 36 1.24 i 0.032 0.657 :1: 0.0149
1 44 1.26 :1: 0.029 0.661 i 0.0135
3 34 1.27 i: 0.033 0.727 :1: 0.0153“
9 16 1.46 :1: 0.048“ 0.870 :1: 0.0223“
‘ Doses expressed as rag/kg egg
” Sanmle size
° Data expressed as mean :1: Standard error
“ S' ' differeh from vehicle cghrol at p<0.05

 

 

 

 

 

 

 

 

Table 23. Effect of congener 77 on the actual and relative liver weigh of three-week-old White
Leghorn chicks.
Dose‘ n” Weig (g)° % Body Weig
Vehicle 36 5.76 :1: 0.138 3.04 :1: 0.059
1 44 5.73 i 0.125 3.01 :1: 0.054
3 35 5.03 :1: 0.140“ 2.76 3: 0.060“
9 16 4.56 :1: 0.207“ 2.74 i 0.089
' Doses expressed as rig/kg egg
” Sample size
°Dataexpressedasmean :1: standarderror
“ Significahly different from vehicle control at p <=0.05

 

 

 

55

 
    
  
 
   
  
 

 

 

 

 

  

 

Table 24. Effect of congener 77 on the actual an! relative spleen weigh of three-week-old
White Leghorn chicks.
Dose' n” Weigflf % Body Weight°
Vehicle 36 0.29 :1: 0.014 0.154 :1: 0.0068
1 44 0.27 i 0.013 0.141 :1: 0.0061
3 35 0.29 :1: 0.014 0.158 i 0.0069
9 16 0.35 i 0.021 0.211 :t 0.0102“
° Doses expressed as rig/kg egg
” Sample size

° Data expressed as mean :1: Standard error
“ Significamly different from vehicle comm] at p<0.05

 

 

 

 

 

 

 
    

=
Table 25. Effect of congener 77 on the actual and relative testes weigh of three-week-old
White Leghorn chicks.
Dosell n” Weigfl)‘ % Body Weight°
Vehicle 16 0.048 1 0.0034 0.024 a: 0.0018
1 21 0.043 :1: 0.0030 0.021 :1: 0.0015
3 18 0.048 :1: 0.0032 0.025 :t 0.0017
9 6 0.044 :1; 0.0055 0.024 i 0.0029
‘ Doses expressed as rig/kg egg
” Sample size

°Datae ressedasmeanistandarderror

1

Table 26. Effect of congener 105 on the actual and relative brain weigh of three-week-old White

56

 

 

 

 

 

 

 

 

 

 

 

 
  
 
 
 
 
 
 
   
 
  

 

 

Leghorn chicks.
Dose‘ n” Weiflf % Body Weightc
Vehicle 39 1.67 :1: 0.017 0.86 :1: 0.017
100 30 1.62 i 0.020 0.87 i 0.020
300 31 1.65 :t 0.019 0.90 :1: 0.019
900 19 1.65 :1: 0.025 0.91 :t 0.025
2700 20 1.67 3: 0.024 0.90 :1: 0.024
8100 7 1.53 :t 0.041“ 1.01 i 0.041“
‘ Doses expressed as rig/kg egg
” Sample size
°Dataexpresscdasmean :1; standarderror
“ S’ ' camly different from vehicle control at p<0.05
_
Table 27 . Effect of congener 105 on the actual and relative bursa weigh of three-week-old
White Leghorn chicks.
Dose‘ n” Weigh(_g_)° % Body Weigh° I
Vehicle 39 1.23 i 0.044 0.573 i 0.0192
100 30 1.08 i 0.051 0.553 :1; 0.0219
300 31 1.07 i 0.050 0.577 :1: 0.0215
900 19 1.02 :1: 0.064 0.556 :1: 0.0275
2700 20 1.01 :t 0.062 0.535 i 0.0268
8100 7 0.69 :t 0.105“ 0.450 t 0.0453
‘ Doses expressed as rig/kg egg
” Sanple size

° Data expressed as mean :1: standard error
“ Si ' different from vehicle control at p < 0.05
‘

  

Table 28. Effect of congener 105 on the actual and relative heart weigh of three-week-old
White Leng chicks.

Dose' 11"
Vehicle 39
100 30
300 3 1
900 19
2700 20
8100

' Doses expressed as ug/kg egg
" Sample size

Weiglg)c
1.14 :1: 0.031
1.13 :1; 0.036
1.12 :1; 0.035
1.11 :1: 0.045
1.15 i: 0.043

1.09 :t 0.074

° Data expressed as mean 1 standard error
“ Significantly differem from vehicle control at

% Body wgng‘
0.579 3: 0.0119
0.586 i 0.0136
0.603 :1: 0.0134
0.603 :1: 0.0171
0.617 :1: 0.0167

0.714 :t 0.0282“

 

 

 

58

 

Table 29. Efiect of congener 105 on the actual and relative liver weigh of three-week-old

 
 

 

 

 

 

 

White Leghom chicks.
Dose“ n" Weigfl)c % Body Weight“
Vehicle 39 5.28 :1: 0.129 2.68 :1: 0.059
100 30 5.18 :t 0.147 2.75 :1: 0.067
300 31 5.01 :t 0.144 2.70 :t 0.066
900 19 5.00 :t 0.185 2.73 i 0.084
2700 20 5.23 :1: 0.180 2.81 :1: 0.082
8100 7 3.96 1; 0.304“ 2.60 :1: 0.138
' Doses expressed as pg/kg egg
" Sample size

° Data expressed as mean i standard error

“ Significamly different from vehicle comrol at p < 0.05

59

 
 

Table 30. Efiect of congener 105 on the actual and relative spleen weigh of three-week-old
White Leghorn chicks.

 
        
      

Dose“ n" Weigh (g)° % Body Weitht‘

 

         

Vehicle 39 0.26 :1; 0.013 0.133 :1: 0.0071

       
    

100 30 0.24 :1: 0.015 0.122 :1: 0.0081

         
    

300 31 0.26 :1: 0.015 0.141 :1: 0.0079

         
    

900 19 0.25 :1: 0.019 0.136 1: 0.0101

       
    

2700 20 0.29 :1: 0.018 0.158 1: 0.0099

  
   

    

8100 7 0.26 :1: 0.031 0.167 :t 0.0167

 

  
   
  

' Doses expressed as ug/kg egg

" Sample size
° Data expressed as mean :1: standard error

  
 

 

Table 31. Effect of congener 105 on the actual and relative testes weigh of three-week-old

 

 

 

 

White Leghorn chicks.
Dose“ n" Weigh (_g_)c % Body Weightc
Vehicle 18 0.044 :1: 0.0031 0.021 :1: 0.0014
100 15 0.041 .4; 0.0034 0.020 :t 0.0016
300 15 0.042 t 0.0034 0.021 :t 0.0016
900 7 0.040 :1; 0.0050 0.020 i 0.0023
2700 9 0.043 i 0.0044 0.021 :t 0.0020
8100 4 0.038 :1; 0.0067 0.023 i 0.0030
‘ Doses expressed as film egg
" Sample size

° Data expressed as mean 1 standard error

Table 32. Effect of congener combinations on the actual and relative brain weigh of three-week-

61

 

 

 

 

 

 

 

old White Leghorn chicks.
Dose“ n" Weigh (g)° % Body We'ggi
Vehicle 24 1.64 :t 0.022 0.87 :t 0.021
105/77 33 1.62 i 0.019 0.85 :1: 0.018
105/126 9 1.54 :1: 0.036 0.96 :1: 0.034
77/126 9 1.49 1: 0.036“ 0.92 :t 0.034
105/77/126 4 1.53 i 0.054 1.03 i 0.051“
“ Doses expressed as ,ug/kg egg
" Sample size
° Data expressed as mean :1; standard error
“ Sig cahly different from vehicle control at p<0.05

 

 

 

 

 

 

 

° Data expressed as mean :1; standard error
J “ Significantly differem from vehicle carnal at p<0.05

 

Table 33. Effect of congener combinations on the actual and relative bursa weigh of three-
week-old White Leghorn chicks.
Dose“ n" Weigh—(g)° % Body Weigi
Vehicle 24 0.92 1: 0.060 0.481 :1: 0.0227
105/77 33 1.01 :t 0.051 0.515 1 0.0193
105/126 9 0.63 i 0.099 0.384 3: 0.0370
77/126 9 0.67 d: 0.099 0.396 :t 0.0370
105/77/126 4 0.48 :t 0.148“ 0.321 i 0.0555“
“ Doses expressed as pg/kg egg
“ Sample size

 

 

62

 

Table 34. Effect ofcongener continatiom on the actual and relative heart weigh ofthree-

 

 

 

 

week-old White Leghorn chicks.
Dose“ n“ Weigh—(g)c % Body Weight“
Vehicle 24 1.31 :t 0.046 0.690 :1: 0.0218
105/77 33 1.33 i 0.039 0.693 1: 0.0186
105/126 9 1.21 :1: 0.075 0.769 :1: 0.0357
77/126 9 1.41 :1: 0.075 0.857 :1: 0.0357“
105/77/126 4 1.36 :t 0.112 0.916 :t 0.0535“
“ Doses expressed as lag/kg egg
" Sanple size

‘Dataexpressedasmean i standarderror
“ Significantly differem from vehicle control at p < 0.05

 

 

Table 35. Effect of congener continatiom on the actual and relative liver weigh of three-

 

 

 

 

° Data expressed as mean i standard error
“ Significahly differeh from vehicle control at p < 0.05

week-old White Leghorn chicks.
Dose“ n" Weigh (g)c % Body Weightc
Vehicle 24 5.68 i 0.204 2.98 :I: 0.064 I
105/77 33 5.62 :1; 0.174 2.92 :I: 0.055
105/126 9 4.33 :1; 0.333“ 2.66 :1: 0.105“
77/126 9 4.75 :1: 0.333 2.85 :t 0.105
105/77/126 4 4.32 :I: 0.500 2.91 :1: 0.157
“ Doses expressed as 11ng eg
" Simple size

 

63

Table 36. Effect of congener combinations on the actual and relative spleen weigh of three-

 

 

 

 

 

week-old White Leghorn chicks.
Dose“ n" Weig_h_(g)“ % Body Weightc
Vehicle 24 0.30 i 0.029 0.157 :1: 0.0139
105/77 33 0.27 :1: 0.025 0.139 :t 0.0119
105/126 9 0.25 :I: 0.048 0.155 :t 0.0227
77/126 9 0.41 :1: 0.048 0.242 :1: 0.0227
105/77/126 4 0.30 :t 0.072 0.205 :1: 0.0341
“ Doses expressed as 11ng egg
" Sanple size

expressed as mean :t error _

Table 37. Effect of congener combinations on the actual and relative testes weigh of three-

 

 

 

 

 

week-old White Leghorn chicks.

F Dose“ nb Weijg_h;£)c % Body Weight“
Vehicle 13 0.042 :1: 0.0042 0.022 :t 0.0020
105/77 17 0.043 :1: 0.0036 0.021 :1: 0.0017
105/126 3 0.037 i 0.0087 0.022 :1: 0.0041
77/126 5 0.046 1 0.0067 0.026 :t 0.0032
105/77/126 3 0.049 i 0.0087 0.032 :1: 0.0041

“ Doses expressed as lag/kg egg

" Sample size

° M__‘f “3 m“. i SW “"0“ _

 

DISCUSSION

Populations of fish-eating waterbirds such as double-crested cormorants (Weseloh
et al., 1983) have been adversely affected by contaminants in the Great Lakes (Peakall,
1988). These effects have included increased embryo mortality and increased incidence
of abnormalities (Peakall, 1988; Gilbertson et al., 1991; Giesy et al., 1994). Based on
concentrations of specific PCB congeners in cormorant eggs collected from Green Bay,
WI and toxic equivalency factors (TEFS) derived by Safe (1991, 1992), congeners 126,
77, and 105 were shown to account for 83%, 1.2%, and 13%, respectively, of the toxic
equivalents (TEQS) in these cormorant eggs or over 97% of the toxicity attributed to
dioxin-like compounds specifically quantified by instrumental analysis (Yamashita et al.,
1993).

Mortality: All three congeners produced significantly lower hatchability of eggs
than the vehicle despite significant mortality in the vehicle control group. The lower rate
of hatching (60%) in the vehicle control group was believed to be a result of injection on
day 0 as opposed to day 4. Injections are generally made on day 4 of incubation in studies
of this type (Brunstrém and Darnerud, 1983; Brunstrom and Reutergardh, 1986;
Brunstrom, 1991) for two reasons: the first reason is that fertility of the eggs can easily
be determined at this time, thus only fertile eggs are injected; the second reason is that
embryo lethality due to the injection is less when injection is done on day 4 as compared
to earlier injection (Brunstrbm, personal communication; Hashizume et al., 1992).

However, lipophilic environmental contaminants like PCBS are deposited in the yolk by

65

the female and are thus present from the time the egg is laid. This early exposure to
contaminants may affect the critical developmental stages of the embryo during the first
nine days (organogenesis) and cause early mortality (Carlson and Duby, 1973). As PCB
concentrations in the yolk increase, embryonic development ceases at earlier stages
(Tumasonis et al.,l973). For this reason, injection on day 0 was thought to best mimic
the exposure situation in the wild. This factor was important enough to outweigh possible
negative consequences associated with early injection.

In the present study, all three PCB congeners had an adverse effect on the hatching
rate of chicken embryos. Injection of congener 126 into the yolk on day 0 produced
almost 100% mortality at 2.7 leg/kg egg. These results agree with those reported by
Brunstrbm and Andersson (1988) in which 2.0 jug/kg egg of congener 126 injected into the
yolk of chicken eggs on day 4 of incubation resulted in 90% mortality by day 18 of
incubation. Injection of 3 jug/kg egg of congener 77 into the yolk on day 0 resulted in
60% hatchability. This is comparable to the 64% hatchability of chicken eggs injected on
day 4 with 4 jug/kg egg of congener 77 by Brunstrdm and Darnerud (1983). Congener 105
resulted in only 85% mortality (day 21) at the greatest dose used in this study, 8100 ug/kg
egg. In contrast, 2500 pg/kg egg of congener 105 injected on day 4 resulted in 85%
mortality by day 18 of incubation (Brunstrdm, 1990). One explanation for this
discrepancy is in the preparation of the stock solution of congener 105. Unlike the other
two congeners which dissolved in the UPC directly, this congener had to be placed in
methylene chloride first, and it is possible that not all the solvent evaporated off. Thus,

the stock solution may have been less concentrated than desired.

66

The present investigation confirms studies establishing the relative toxicity of these
three congeners with congener 126 being the most toxic followed by congeners 77 and 105
(Tanabe et al., 1987; Brunstrbm, 1989, 1990). Since these congeners have been shown
to have detrimental effects on the hatchability of chickens, it is possible that environmental
concentrations of these congeners could be having negative effects on the hatchability of
wild avian species. In particular, congener 126 has been found at greater concentrations
(3.6 leg/kg egg) in the eggs of double-crested cormorants fro rn Green Bay, WI (Yamashita
et al., 1993) than the dose (2.7 ug/kg egg) which caused nearly 100% mortality in chicken
embryos in this study. While no significant differences were observed in mortality post-
hatch, one study in which chicks of hens fed 50 mg total PCBs weekly for 17 weeks did
report increased mortality between hatching and six weeks of age (Dahlgren and Linder,
1971). It is conceivable that increased mortality in those birds with lower body weights
would have been observed had this study continued for another three weeks.

Teratogenicity: All three congeners caused greater incidences of abnormalites than
the vehicle control. Congener 126 produwd significantly (p < 0.10) more abnormal
embryos than the vehicle at 0.9 ug/kg egg. While miscellaneous abnormalities were the
most frequently observed type of abnormality in this dose group, when all PCB dose
groups were combined, the most frequently observed abnormality was beak deformities.
Other predominant abnormalities included edema and eye deformities. These types of
abnormalities were also observed by Brunstrom and Andersson (1988). Congener 77
produced significantly (p <0.10) more abnormal embryos than the vehicle at 27.0 ug/kg

egg. The predominant abnormalities in this group as well as when all congener 77 dose

67

groups were considered together were beak deformities. Eye deformities and edema were
also commonly observed. Brunstrbm and Darnerud (1983) observed these abnormalities
in their injection studies with congener 77. Congener 105 produced significantly
(p <0.05) more abnormal embryos than the vehicle at 8100 ng/kg egg. Abnormally small
embryos, edema, and beak deformities were the most frequently observed abnormalities
in all dose groups combined. None of the congener combimtions produced any significant
differences in the number of abnormal embryos when compared to the vehicle control.
Combining all die congener combination groups, edema of the head and neck was the most
common type of abnormality followed by miscellaneous abnormalities. The lack of a
consistent pattern in most frequently observed types of abnormalities is not surprising.
Inhibition of cell-cell communication (Loch-Caruso and Trosko, 1985; Kavanagh et al. ,
1987) is one potential mechanism for teratogenesis. The location of this interrupted
intercellular communication will determine the resulting types of abnormalities.

This study has demonstrated the teratogenic capabilities of each of these congeners
when injected singly. Many of the abnormalities observed, such as edema and bill
deformities which included crossed and/or shortened bills, have been commonly observed
in double-crested cormorants (Yamashita et al., 1993). All abnormalities were observed
in either dead embryos or chicks upon hatching. No new abnormalities were observed
during the three weeks post-hatch, suggesting that all abnormalities resulting from these
congeners occur early in the development of the embryo and do not arise in response to

continual uptake of the yolk following hatch.

68

Body weight: Wasting syndrome, a condition in which animals experience
decreases in body weight which ultimately lead to death, was not observed in this study.
However, signifieantly lower body weight gains were observed at the greater doses of all
three congeners and the combinations containing congener 126. In a previous injection
study with congener 77, a delay in the onset of weight gain was observed in chicks
hatching from eggs injected with 4 ug/kg egg. The lower body weights were significantly
different from the control weights at three to 10 days post-hatch (Brunstrom and Darnerud,
1983). Growth retardation has also been seen in chicks hatching from eggs injected with
Aroclor’ 1242 on day 0 (Carlson and Duby, 1973). It was suggested that this growth
retardation may be due to a combination of effects on the liver and hematopoietic tissue.
Dahlgren and Linder (1971) found that 50 mg Aroclor' 1254 fed to adult hens weekly for
17 weeks resulted in chicks with lower body weights than control chicks at six weeks of
age. Harris and associates (1976) fed White Leghorn chickens 5, 10, and 20 ppm Aroclor'
1242 or 1248 for eight weeks. Body weights of progeny from these PCB-fed hens were
not different at hatch, but three-week-old chicks of hens fed 20 ppm of either Aroclor'
were signifieantly lighter than control progeny. These results support the conclusion
reached in the present study; PCBs have an adverse effect on body weight gain in chicks
exposed in ovo.

Bruin: In the present study, relative brain weights increased in a dose-dependent
manner when compared to the vehicle control group. However absolute brain weights
were either not affected (congener 105) or were lower (congener 126, combination of 126

and 77) when compared to vehicle control absolute brain weights. The changes observed

69

in the brain weight relative to body weight probably do not represent a real effect. As
with congener 105, where the absolute brain weights did not change, the apparent increase
was due to lower body weights rather than greater brain weights. Brain weight was also
unaffected in nestling pelicans fed 100 mg Aroclor' 1254 daily for 10 weeks followed by
two weeks of imposed food stress prior to necropsy (Greichus et al., 1975).

Bursa of Fabricius: Bursa weights, while significantly lower only in chicks
hatching from eggs injected with congener 126 and the combination of all three congeners,
were numerieally lower in birds exposed to congeners 77 and 105 as well as the remaining
combinations containing congener 126. The injection of congener 77 into the air cell of
chicken eggs after 13 days of incubation and killed on day 19 of incubation resulted in
lower bursa weights (73% of control weights at 30 ug/kg egg and 58% of control weights
at 300 jug/kg egg) (Nikolaidis et al., 1988). Histological assessment of the bursae
indieated a dose-dependent decrease in the number bursa] follicles as well as in the number
of lymphoid cells (Nikolaidis et al., 1988). Bursae of lO-day—old chicken embryos
cultured in media containing TCDD for 24 hours and subsequently transplanted to the
chorioallantoic membrane of 10-day-old chicken embryos had lower numbers of lymphoid
cells (Nikolaidis et al., 1990). Thus, dioxin and dioxin-like compounds can inhibit
lymphoid development and perhaps compromise the immune system. While histological
examination of the bursa in the present study indicated no significant alteration when
compared to vehicle control, the trend toward lower relative bursa weights could indicate
a potentially compromised immune system as a result of PCB exposure. However, Harris .

et al. (1976) reported that feeding of Aroclor' 1248 to hens for 8 weeks resulted in lower

70

bursa weights in one-day-old chicks but not ll-week-old chicks. This decrease in bursa
weight was not associated with a decrease in antibody production.

Heart: There was a general trend toward greater relative heart weights with
increasing doses of all congeners and with all combinations containing congener 126. No
gross malformations of the heart were observed, however, cardiovascular teratogenicity
was not specifically assessed as has been done in embryos exposed to TCDD. Chicken
embryos exposed to TCDD via injection into the albumin on day 0 of incubation had
ventricular septal defects, aortic arch anomalies, and conotruncal malformations (Cheung
et al., 1981). Previous studies involving exposure of various avian species to Aroclor'
1254 via the feed reported a PCB-induced increase (finches), decrease (chickens), and no
clunge (pelicans) in heart weight (Prestt et al., 1970; Greichus et al., 1975; Cecil et al.,
1978). Although there were no obvious cases of pericardial edema in the present study,
the increase in relative heart weight observed in PCB-exposed chicks could be due to
interml fluid accumulation. Chicks fed 0.04% chlorinated biphenyls for four weeks had
hydropericardium, hydroperitoneum, and enlarged hearts (McCune et al., 1962). Flick
et al. (1963) fed one-day-old cockerels a fatty by-product known to cause chick edema
disease. Birds consuming this ”toxic fat” had increased fluid accumulation in the
pericardium and decreased hematocrits. The authors suggested that the chick edema factor
may increase the permeability of the vascular bed of the heart. Another possible
explanation could be an increase in the muscle mass of the heart (Walker, personal

communieation).

71

Liver: Changes in liver weight did not show a consistent pattern. Congener 77
eaused a signifieant decrease in relative liver weight at 3 ag/kg egg. While not significant
due to the small sample size, the 9 pig/kg egg group also had lower liver weights. The
combination of congeners 105 and 126 also produwd significantly lower relative liver
weights. In another study in which PCBs were injected into chicken eggs (day 5), a
decrease in actual liver weight was observed (Hatano and Hatano, 1994). In contrast,
liver weights were greater in cockerels fed 50 or 500 ppm of either Aroclor' 1254 or 1268
for 30 days which was associated with induction of hepatic mixed function oxidases (Cecil
et al., 1978). Liver weights were also greater when compared to the liver weights of
controls in nestling pelicans fed 100 mg/day of Aroclor' 1254 (Greichus et al., 1975) as
well as in mice (C57BL/Rij) administered weekly ip injections of 100 mg congener 77/kg
body weight for four weeks (Brouwer and Van den berg, 1984).

Each congener, 126, 77 , or 105, has been shown to cause induction of hepatic 7-
ethoxyresorufin O-deethylase (EROD) in chick embryos. These embryos were exposed
via injection into the air cell on day 7 and their livers were removed 72 hour later for
EROD analysis (Brunstrom and Andersson, 1988; Brunstrbm, 1990). Brunstrom (1986)
reported that five-day—old chicken embryos had measurable activity of hepatic 7-
ethoxycoumarin O-deethylase (ECOD) and aryl hydrocarbon hydroxylase (AHH). This
enzyme activity increased from day 5 to day 10 of incubation, and then plateaued until day
19. Activity peaked at one day following hatch. Livers of five-day-old embryos taken
48 hour after injection of 5 pg congener 77/kg egg into the air cell had induced AHH (14

times control) and ECOD (two times control) activity. It is questionable whether any

72

changes in enzyme activity would have been observed in the chicks in the present study
since these chicks were initially exposed on day 0 before any development of the liver or
enzyme systems. Additionally, these chicks were not subject to necropsy until 21 days
post-hatch (42 days from initial exposure). In the above mentioned studies, livers were
removed two to three days following dosage of the embryos with PCB congeners.

Spleen: Spleen weights also lacked a definitive trend but results seemed to suggest
an increase over the vehicle control with increasing dose. This was only significant with
congener 77 at 9 rig/kg egg. In contrast, spleen weights were lower in chicks of hens fed
various Aroclors' for eight weeks (Harris et al. , 1976). Aroclor’ 1254 at 500 ppm caused
decreased spleen weights in chickens after 30 days of dietary exposure (Cecil et al. , 1978).
However, actual spleen weights were greater in day-18 chicken embryos injected with 60
nmol congener 77 on day 9 of incubation than in embryos injected with vehicle (dioxane)
only (Rifkind and Muschick, 1983). Relative spleen weights were also significantly
greater in pelicans exposed daily to 100 mg Aroclor’ 1254 in the diet for two and half
months (Greichus et al., 1975).

Testes: Testes weights were not significantly different in any of the congener
groups or combinations used in the present study. Similarly, testes were unaffected in
hatchlings of Aroclor'-fed hens (Harris et al., 1976). However, Cecil et al. (1978)
reported that Aroclor' 1242 or 1254 at 500 ppm caused decreased testes weights in
chickens after 30 days of dietary exposure. Similarly, cockerels fed 500 ppm showed
lower testes weights at nine and 13 weeks of feeding (Platonow and Funnell, 1971).

However, rats exposed to Aroclor' 1254 during lactation had testes with weights, absolute

73

and relative, signifieantly greater than controls. This increase in testes weight supported
a hypothesis that PCB exposure neomtally and during lactation (dams treated on days 1-3,
5, 7, and 9 of lactation) results in a hypoandrogenic condition in adult male rats. These

male offspring also experienced impaired mating and fertility (Sager, 1993).

SUMMARY AND CONCLUSIONS

In summary, all three PCB congeners caused adverse effects in chicken embryos
exposed prior to incubation. Congener 126 caused significantly greater mortality,
increased incidence of abnormalities, lower post-hatch body weights, lower relative bursa
weights, and greater relative heart weights at the 0.9 lag/kg egg dose when compared to
the vehicle. Congener 77 caused significantly greater mortality at the 9 ag/kg egg dose
than the vehicle. Increased incidence in abnormalities occurred at the 27 pig/kg egg dose
of congener 77. A dose of 3 jug/kg egg of congener 77 caused lower post-hatch body
weights and greater relative heart and liver weights than the vehicle. Relative spleen
weights were greater at the 9 ag/kg egg dose when compared to the vehicle. Congener
105 caused significantly greater mortality at the 8100 rig/kg egg dose than the vehicle.
There were also increased incidences of abnormalities and greater relative heart weights
at this dose group. Post-hatch body weights were significantly lower beginning with the
300 jug/kg egg dose. The LDws were 0.6, 4.9, and 5 ,592 ug/kg egg for congeners 126,
77, and 105, respectively. Post-hatch body weight was the most sensitive of the

parameters studied. LOAELs and NOAELs were determined for each of the congeners

74

using doses that had an adverse effect on body weight . The LOAELs were 0.5, 5.2, and
173 ag/kg eg for congeners 126, 77, and 105, respectively. The NOAELs were 0.3, 1,

and 100 ag/kg eg for congeners 126, 77, and 105, respectively.

75
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Platonow, LS. and HS. Funnel]. 1971. Anti-androgenic-like effect of polychlorinated
biphenyls in cockerels. Vet. Rec., 88:109-110.

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Safe, S. 1992. Development, validation, and limitations of toxic equivalency factors.
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Sager, DB. 1983. Effect of postnatal exposure to polychlorinated biphenyls on adult male
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Highly toxic coplanar PCBS: occurrence, source, persistency and toxic implications to
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1991. H4IIE rat hepatoma cell bioassay-derived 2,3,7,8—tetrachlorodibenzo-p—dioxin
equivalents in colonial fish-eating waterbird eggs from the Great Lakes. Arch. Environ.
Contam. Toxicol., 21:91-101.

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with embryonic mortality and deformity of hatched chicks. Arch. Environ. Contam.
Toxicol., 12312-324.

Weseloh, D.V., S.M. Teeple, and M. Gilbertson. 1983. Double-crested cormorants of
the Great lakes: egg-laying parameters, reproductive failure, and contaminant residues
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Yamashita, N., S. Tanabe, J.P. Ludwig, H. Kurita, M.E. Ludwig, and R. Tatsukawa.
1993. Embryonic abnormalities and organochlorine contamination in Double-Crested
Cormorants (Phalacrocorax auritus) and Caspian Terns (Hydroprogne caspia) from the
upper Great lakes in 1988. Environ. Pollut., 79:163-173.

CHAPTERZ

INJECTION OF 3,3',4,4',5-PENTACHLOROBIPI-IENYL INTO

THE YOLKS OF DOUBLE-CRESTED CORMORANT

(Phalaemcorax auritus) AND CHICKEN (Gallus domestieus) EGGS

79

80
ABSTRACT

Unincubated double-crested cormorant (Phalacrocorax auritus) eggs were collected
from a relatively uncontaminated colony at lake Winnipegosis, Manitoba, Canada in May,
1994. Eight doses of 3,3',4,4',5-pentachlorobiphenyl (IUPAC #126) were injected into
the yolks of these eggs prior to their artificial incubation. Congener 126 was also injected
into the yolks of White Leghorn chicken (Callus domesticus) eggs on day 0 of incubation.
Upon hatching, chicks of both species were subjected to necropsy. The brain, bursa,
heart, liver, and spleen were removed and weighed. Assessment of the rate of hatching
indieated LD,.,s of 246 and 2.4 ug/kg egg for cormorants and chickens, respectively. No
significant differences in the incidence of abnormalities were observed in the cormorant,
while in the chicken significantly more abnormalities were observed at 3.2, 6.4, and 12.8
tug/kg egg than the vehicle control. There was no effect on body weight in hatchling
cormorants. Relative heart, liver, and spleen weights were increased in the cormorant at
5 rig/kg egg. In the chicken, a dose of 3.2 pig/kg egg caused lower hatchling weights and
greater relative brain, heart, and liver weights when compared to the vehicle control
group. Relative spleen weight was unchanged, and while not significant, there was a trend
toward lower relative bursa weights with increasing dose in both the cormorant and

chicken.

81

INTRODUCTION

Polychlorinated biphenyls (PCBs) have become widespread contaminants in the
Great lakes. They are known to cause a variety of adverse effects in birds: increased
embryo mortality, hydropericardium, bill deformities, subcutaneous edema, liver lesions,
and induction of cytochrome P-450 enzymes (Brunstrom, 1989). Coplanar PCBS are
structurally similar to 2,3,7,8-tetrachlorodibenzo-p—dioxin (TCDD), and thus, can act
through the same mechanism of toxicity as TCDD; binding to the Ah receptor. The most
toxic of these PCB congeners is 3,3',4,4',5-pentachlorobiphenyl (IUPAC# 126) (Tanabe
et al., 1987; Brunstrom, 1989). This congener represents approximately 83 % of the toxic
equivalents in cormorant eggs collected from Green Bay, Wisconsin, one of the most
contaminated sites in the Great lakes basin (Safe, 1991; Yamashita et al., 1993).

Double-crested cormorants (Phalacrocorax auritus) experienced reproductive
problems from the late 1950s through the 1970s (Weseloh et al., 1983; Ludwig, 1984).
While cormorant populations have been increasing in recent years (Weseloh and Ewins,
1994), they are still experiencing reproductive impairment including embryo lethality and
anatomic deformities (Fox et al., 1991).

Research, thus far, has focused on correlations between contaminant concentrations
at colony sites and biological end points such as mortality, incidence of abnormalities, and
induction of liver enzymes (Yamashita et al., 1993; Sanderson et al., 1994). To my
knowledge, there have been no controlled laboratory studies which assessed the effects of

a particular PCB congener injected into relatively uncontaminated cormorant eggs prior

82

to incubation. Additionally, it would also be valuable to determine the difference in
sensitivity to specific PCB congeners between the cormorant and the chicken. The chicken

has been extensively studied and is easy to use in controlled laboratory experiments.

MATERIALS AND METHODS

Cormorant: Double-crested cormorant (Phalacrocorax auritus) eggs, provided by
the FWS, were collected from a relatively uncontaminated colony at lake Winnipegosis
in Manitoba, Canada. Eggs were selected from nests containing only two eggs to ensure
fresh (unincubated) eggs. Upon arrival at MSU, less than 24 hours after collection, the
eggs were weighed and candled for air cell location. The eggs were returned to egg
cartons and placed in coolers such that the eggs were laying horizontally. They were kept
at room temperature for approximately 12 hours prior to injection, during which time they
were rotated 90° about their long axis twice.

The eggs were divided into 11 groups, 60 eggs per group with the exception of the
greatest dose group which contained only 15 eggs due to the amount of PCB congener
required. There were 8 doses of congener 126 (5, 10, 25, 50, 100, 200, 400, and 800
ug/kg egg). There were also three control groups with 60 eggs each: non-injected, mock-
injected (insertion of the syringe needle only), and vehicle control.

Congener 126 (AccuStandard, New Haven, CT) was diluted in triolein (Sigma, St.
Louis, MO). Due to the instability of this oil in the presence of oxygen, all injection vials

were flooded with argon following preparation of the doses. Congener 126 was dissolved

83

directly in the triolein to make stock solutions (5 ug/al for the first seven doses and 10
ag/al for the greatest dose). Each dose was cold filtered with a 0.22 micron syringe filter
to sterilize the solution prior to injection.

All injections were made directly into the yolk prior to incubation. The injection
volume was 0.1 al/g egg. The surface of the egg was cleaned with 75% ethanol and a
small hole was rmde into the air cell with a sterile pin. A sterile 25);] syringe (Hamilton
Company, Reno, NV) (228 gauge) was inserted horizontally approximately 30 mm into
the egg which was positioned on its side. Following injection of the solution and removal
of the needle, the hole was sealed with melted paraffin.

Cormorant eggs were incubated on their sides in a Petersime (Gettysburg, OH)
poultry incubator for up to 29 days. The temperature (99.0°F) and relative humidity (61-
63%) were maintained using 99°F dry bulbs and a 87.5°F wet bulb. Eggs were
automatically rotated every two hours. In addition, they were rotated 180° about their
long axis by hand once a day. Approximately 80 additional eggs were incubated and
weighed every two to three days to evaluate mass loss during incubation. These eggs were
not rotated once a day by hand. All eggs were candled on days 7, 14, and 21 of
incubation. Eggs were transferred to hatching baskets on day 25 or when they began to
pip, whichever came first. All unhatched eggs were opened and assessed for abnormalities
and approximate stage of development.

Upon hatching at approximately 26 days of incubation, cormorants were weighed
and body length (crown of head to tip of tail) was measured. They were killed by

decapitation, and the brain, bursa, heart, liver, and spleen were removed and weighed.

84

Five livers from each group, where possible, were wrapped in foil and put in liquid
nitrogen for later analysis of 7—ethoxyresorufin O-deethylase (EROD) induction.

Chicken: White Leghorn chicken (Gallus domesticus) eggs were provided by the
MSU Poultry Science Teaching and Research Center. They were weighed and candled for
the location of the air cell the day before injection. In addition, the eggs were laid on dteir
sides to allow the germ spot to raise to the top and minimize potential injury during
injection. Eggs were kept in a cooler at 58°F prior to injection.

There were eight doses of congener 126 (0.1, 0.2, 0.4, 0.8, 1.6, 3.2, 6.4, and
12.8 ug/kg egg) as well as non-injected and vehicle controls. Previous injections of
chicken eggs eliminated the need for a mock-injected control group. The doses were
prepared from a stock solution (0.2 ag/al) which was made with the same lot of congener
126 used for the cormorant. The same vehicle and injection techniques were used for both
the chicken and the cormorant. Since availability of chicken eggs was not a limiting
factor, the chicken egg injections were done in a series of three replicates with 20
eggs/group in each replieate to give the same total number of eggs used for the cormorant.

Chicken eggs were also incubated in a Petersime poultry incubator. They were
positioned with their blunt end up and incubated according to standard conditions for this
species, 99.5-99.75°F during the first 19 days, 97-99°F for days 21 and 22, and relative
humidity approximately 60% (North, 1978). Eggs were candled on days 4 and 11.
Viability was assessed at the time of transfer (day 18) with an EVD provided by the FWS.
Any embryos dead on days 11 or 18 were opened and assessed for abnormalities and age

at death was estimated.

85

Hatchling chicks were weighed and killed by decapitation. The brain, bursa, heart,
liver, and spleen were removed and weighed. Five livers from each group were frozen
for EROD comparisons with the cormorant.

Statistics: Mortality and abnormalities were evaluated using a 2x2 contingency
table and Bonferroni Chi-square. The LDws were ealculated by linear regressions on log-
probit transformed data. Analysis of all other parameters was done using one-way
ANOVA followed by Dunnett's test. All comparisons were made relative to the vehicle
control group due to the low rate of hatching experienced by the cormorant vehicle control

group. Statements of significance were based on a level of 0.05, unless otherwise stated.

RESULTS

Congener 126 caused embryo mortality in both the cormorant and chicken and
increased incidence of abnormalities in the chicken. The LDsos of congener 126 were 246
and 2.4 Mg 126/kg egg (Tables 38-40) for the cormorant and chicken, respectively. There
were no significant differences in the incidence of abnormalities between cormorant
embryos in PCB treatment groups and those in the vehicle control group. There were
unhatched cormorant chicks with edema in every group except the greatest dose group
where the majority of embryos died before day 21. One embryo in the 10 ag/kg egg dose
group had no eyes or bill. One embryo in the 400 jig/kg egg dose group had a hooked
upper bill, and at 800 jug/kg egg, one embryo had no lower jaw. In the chicken, no

abnormal embryos were observed in the vehicle control group while 13/60, 19/59, and

86
10/60 embryos had one or more types of abnormalities in the 3.2, 6.4, and 12.8 ug/kg egg

dose groups, respectively. These incidences of abnormalities were significantly greater
than in the vehicle control group (p <0.01). The types of abnormalities observed included
skull, eye, beak, and toe deformities as well as edema.

Congener 126 also caused sublethal effects such as changes in body weights and
various organ weights. There were no differences in body weights (Tables 41-42) of
cormorant hatchlings from the vehicle control group while the body weights of chicks in
the greatest dose group which hatched (3.2 ,ug/kg egg) were significantly lower than those
in the vehicle control group. Relative organ weights, organs expressed as a percent of
body weight, were used for statements of significance. Relative heart, liver, and spleen
weights in cormorant were all significantly greater at the lowest dose (5 ag/kg egg) relative
to the vehicle control group. In the chicken, relative brain, heart, and liver weights were
all significantly greater at 3.2 ag/kg egg than the values of the same parameters in the
vehicle control group. There were no significant differences in spleen weights relative to

the vehicle control group.

 
  

 

 

Table 38. Effect of injection of congener 126 iho the yolks of double-crested cormoram eggs on

 

 

 

day00fincubationonhatchability. _ ___
Dose Dose l dead/
‘ jug/kg egg nmol/kgegg # eggs % Mortality
Non-injected 0 18/60 30.0
0' 0 32/60 53.3
5 15 17/60 28.8
10 31 17/59 28.8
25 77 23/60 38.3
50 153 20/60 33.3
100 306 24/59 40.7
200 613 36/60 60.0
400 1225 52/60" 86.7
800 2451 15/15" 100
“ Vehicle (triolein)
‘ __‘ggnificahly differeh from the vehicle control at u___<5_____ _«fi ‘#_ __E _

 

 

88

 

 
      
 

Table 39. Effect of injection of congener 126 iho the yolks of White Leghorn chicken eggs on
day 0 of incubatrglon hatchability.

 

 

 

 

 

 

Dose Dose
118/kg egg nmol/kgg
Non-injected 0
0' 0 5/59 8.5
0.1 0.3 11/60 18.3
0.2 0.6 18/59" 30.5
0.4 1.2 15/59 25.4
0.8 2.5 12/59 20.3
1.6 4.9 13/60 21.7
3.2 9.8 55/60" 91.7
6.4 19.6 59/59” 100
12.8 39.2 60/60" 100
“ Vehicle (triolein)
" Significamly differem from the vehicle control at p<0.05 I

 

 

 

 

Table 40. was for congener 126 injected into the yolks of White Leghorn chicken or double-
crested cormorah eg_gs on day 0 of incubation.

 

 

 

 

Chicken Cormorant
LD,0 ag(nmol)/kg egg 2.4 (7.4) 246 (754)
95% Confidence limits 1.7 (5.2) - 2.7 (8.3) 179 (548) - 302 (925)
Intercept i Standard Error 1.3 :1; 1.30 - 4.7 i 2.36
95% Confidence limits -1.29 - 3.82 -9.31 - -0.04
Slope 1; Standard Error 9.9 :1: 2.71 4.1 :l: 0.96
95% Confidencelimits 4.59 - 15.22 2.17 - 5.92

 

 

 

 

89

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

% Mortality
100 J - /
90 g f
80 .. /
70 g
60 /
50
40 M
30 '1’ H
20
10
0 1 id“; t 1111:??? : ::+:+.r
1 1 x 101 l x 10“ 1 x 103
log(dose +1)
—-— Chicken -o- Cormorant I
Figure 4. Effect of congener 126 on the mortality of White Leghorn chicken and

double-crested cormorant embryos when exposed prior to incubation. One
was added to each dose to allow inclusion of the vehicle control on a log
scale. For example, vehicle control (0 ag/kg egg) = 1 and the highest dose
of congener 126 in the cormorant (800 jug/kg egg) = 801.

90

Table41. Effeaofcmgarerl260ndielntchbodylengthandbodyweighofdum1e-creswd

 

 

 

cormorams.
Dose“ 11" Length (cm)° n" Body Weigh (g)°
Non-injected 49 11.22 i 0.063 53 34.99 i: 0.479
Vehicle 28 11.01 :1: 0.084 29 34.63 :1: 0.647
5 37 11.33 :t 0.073 38 34.14 :t 0.565
10 42 11.15 :1: 0.069 44 33.09 :1: 0.525
25 32 11.21 :t 0.078 34 34.70 :1: 0.598
50 39 11.16 :1: 0.071 39 34.50 i: 0.558
100 31 11.15 i 0.080 31 34.94 :1: 0.626
200 24 11.02 :I: 0.091 24 33.84 :1: 0.711
400 8 10.98 :I: 0.157 8 36.40 i 1.232
“ Doses expressed as pig/kg egg
° Sample size
°Wms°dflm _ l l - a l_

 

Table 42. Effect of conggner 126 on the hatch body weighs of White Leghorn chickens.

 

 

 

    
   
   
   
   
   
  
  
  

 

   
 
  

 

 

Dose“ n" Body Weighigg)
Non-injected 50 43.0 i: 0.49
Vehicle 54 41.4 i 0.47
0.1 49 42.4 i 0.50
0.2 41 42.4 i 0.54
0.4 42 41.9 i 0.54
0.8 45 41.6 d: 0.52
1.6 47 41.8 :t 0.51
3.2 8 36.1 i 1.23“
“ Doses expressed as am egg
° Sample size

° Data expressed as mean :1: standard error
“ Significantly different from vehicle control at p < 0.05

91

Table 43. Effect of congener 126 on the actual and relative brain weigh of hatchling double-

 

 

 

 

_W‘sicm

 

 

crested cormorahs.
I Dow 11" MM)“ % Bgy Wei- ..
Non-injected 48 0.788 1 0.0105 2.27 :1: 0.048
Vehicle 29 0.780 :1; 0.0135 2.30 i 0.061
5 36 0.785 :1; 0.0121 2.34 :1: 0.055
10 42 0.783 1 0.0112 2.41 i 0.051
25 32 0.769 1: 0.0128 2.23 i: 0.058
50 37 0.748 :t 0.0119 2.19 :t 0.054
100 31 0.775 i 0.0130 2.23 :I: 0.059
200 24 0.769 :t 0.0148 2.29 i: 0.067
400 8 0.762 i 0.0257 2.11 :t 0.116
“ Doses expressed as pig/kg egg
' Sanple size

Table 44. Effect of congener 126 on the actual and relative bursa weigh of hatchling double-

 

 

 

 

 

 

crested cormorants.
Dose“ n" Weigh;__(g)° % Body Weigh°
Non-injected 48 0.030 1 0.0011 0.087 :1: 0.0032
Vehicle 28 0.030 3: 0.0014 0.088 :1: 0.0042
5 36 0.031 :1; 0.0013 0.092 :t 0.0037
10 41 0.026 1 0.0012 0.080 i 0.0035
25 32 0.027 :1: 0.0013 0.078 :t 0.0040
50 36 0.029 1 0.0013 0.084 :1; 0.0037
100 30 0.028 1; 0.0014 0.080 t 0.0041
200 24 0.025 i 0.0016 0.074 :t 0.0046
400 8 0.028 i 0.0027 0.075 i 0.0079
“ Doses expressed as film egg
" Sample size
°Dataegressedasmean i standarderror

 

 

 

Table 45. Effect ofcongener126 onthe actual and relative heart weigh ofhatchling double-

 
 

 

 

 

 

crested cormorants.
Dose“ n" Weifl)“ % Body Weigh“
Non-injected 48 0.244 :1: 0.0050 0.702 :1; 0.0150
Vehicle 29 0.223 :1; 0.0065 0.653 :1: 0.0192
5 36 0.251 :t 0.0058“ 0.744 :1: 0.0173“
10 42 0.239 :1: 0.0054 0.725 :1: 0.0160
25 32 0.242 :1: 0.0062 0.700 :1: 0.0183
50 37 0.235 :t 0.0057 0.686 :1: 0.0170
100 31 0.226 3; 0.0063 0.646 i 0.0186
200 24 0.228 i 0.0071 0.679 :1; 0.0211
400 8 0.211 1 0.0123 0.581 i 0.0366
“ Doses expressed as rig/kg egg
“ Sample size

cData expressed as mean i standard error

 

.1 . . . .
Srgmficahly drffereh from vehicle cortrol at p<0.05

 

 

Table 46. Effect of congener 126 on the acural and relative liver weigh of hatchling double-

  

 

 

 

 

crested cormorants.
Dose“ r1b Weig;ht_(_g)c % Body Weighc
Non-injected 48 0.709 :I: 0.0147 2.04 :1: 0.047
Vehicle 29 0.659 :t 0.0189 1.92 :t 0.060
5 36 0.763 i 0.170“ 2.26 i 0.054“
10 42 0.676 :t 0.157 2.06 :1: 0.050
25 32 0.691 :t 0.0180 2.00 :1: 0.057
50 37 0.703 :1: 0.0168 2.06 :t 0.053
100 31 0.666 :t 0.0183 1.92 :t 0.058
200 24 0.672 t 0.0208 2.00 :t 0.066
400 8 0.684 :1: 0.0361 1.88 :t 0.114
“ Doses expressed as rig/kg egg
° Sample size
° Data expressed as mean 1 standard error
“ 812m differ-em from vehicle control at p < 0.05

 

 

 

94

 

 

 

 

 

 

. — — 1
Table 47. Effect of congener 126 on the actual and relative spleen weigh of hatchling double-
crested comrorants.
H— Dose“ n" Weig=h%(g)c % Body Weigh°
Non-injected 48 0.070 i 0.0040 0.202 :I: 0.0119
Vehicle 29 0.066 :1: 0.0051 0.192 :t 0.0153
5 36 0.092 :1: 0.0046“ 0.272 i 0.0138“
10 41 0.081 :1: 0.0043 0.248 :1: 0.0129
25 32 0.070 :t 0.0049 0.203 :t 0.0146
50 37 0.081 :1; 0.0045 0.237 1 0.0136
100 30 0.072 :1: 0.0050 0.207 :1: 0.0151
200 24 0.056 :1: 0.0056 0.163 i 0.0168
400 8 0.068 :1: 0.0097 0.186 :I: 0.0292
“ Doses expressed as pg/kg egg
" Sanple size

° Data expressed as mean :t standard error

“ Significamly different from vehicle control at p<0.05

 

 

 

95

 

Table48. Effectofcongener 126ontheactualandrelativebrainweighofhatchlingWhite

 

 

 

 

 

 

—

“Dataexpressedasmean i standarderror
“ Significahly differeh from vehicle cohrol at p<0.05

Leghorn chickens
Dose“ n" Weiflg)° n" % Body Weigh“;
Non-injected 50 0.871 :1: 0.0090 50 2.04 :t 0.030
Vehicle 54 0.846 :1; 0.0087 54 2.05 :1: 0.029
0.1 49 0.851 1; 0.0091 49 2.02 :t 0.030
0.2 41 0.853 i 0.0099 41 2.02 :I: 0.033
0.4 43 0.853 :1: 0.0097 42 2.04 :t 0.033
0.8 45 0.843 i 0.0095 45 2.04 :I: 0.032
1.6 47 0.853 :1: 0.0093 47 2.05 :I: 0.031
3.2 10 0.847 j: 0.0201 8 2.37 i: 0.075“
“ Doses expressed as rig/kg egg
“ Sample size

 

 

 

96

 

 

 

2‘1“ 01 ".r

 

 

 

° Data expressed as mean :1; standard error

i,

differed from vehicle comm] at p<0.05

Table49. Effeaofcongener1260ndreacmalandrelafivehrrsaweighoflntchlmgWhite
Leghom chickens.
Dose“ nb Weigh (g)° % Body Weigh°
Non-injected 50 0.061 :1: 0.0025 0.14 :1: 0.006
Vehicle 53 0.059 1 0.0024 0.14 :1: 0.006
0.1 49 0.055 3: 0.0025 0.13 :I: 0.006
0.2 40 0.060 :1; 0.0028 0.14 :1: 0.007
0.4 42 0.059 :1; 0.0027 0.14 :1: 0.006
0.8 45 0.053 :t 0.0026 0.13 i 0.006
1.6 47 0.055 :1; 0.0026 0.13 :1; 0.006
3.2 9 0.036 :1: 0.0058“ 0.11 i 0.016
“ Doses expressed as rig/kg egg
° Simple size (For % Body Weigh n=41 for 0.4 [.4ng egg and n=7 for 3.2 Hm egg)

 

 

97

Table50. Effectofcongener126ondreacmalandreladveheanweighofhatchlingWhite

 

 

 

 

 

Leghorn chickens.
Dose“ n" Weigh—(g)° % Body We'
Non-injected 50 0.26 :1: 0.004 0.61 :1: 0.011
Vehicle 54 0.25 :I: 0.004 0.60 :1: 0.011
0.1 49 0.23 :1: 0.005 0.55 :1: 0.011
0.2 41 0.24 :t 0.005 0.56 i: 0.012
0.4 43 0.24 d: 0.005 0.58 :1; 0.012
0.8 45 0.25 i 0.005 0.60 :1: 0.012
1.6 47 0.25 :I: 0.005 0.59 :1: 0.012
3.2 10 0.25 :1: 0.010 0.70 i: 0.028“
“ Doses expressed as rig/kg egg

° Sanmle size (For % Body Weigh n=42 for 0.4 rig/kg egg and n=8 for 3.2 rig/kg egg)
° Data expressed as mean :1: standard error

“ Sim differeh from vehicle cohrol at <0.05

98

Table 51. Effect ofcongener126 on the actual and relative liver weigh ofhatchling White

 

 

 

 

Leghom chickens.
Dose“ n“ Weigfl)“ 95 Body Wei 3:
Non-injected 50 0.87 i 0.015 2.03 i 0.039
Vehicle 54 0.86 d: 0.014 2.09 :t 0.037
0.1 49 0.88 1: 0.015 2.07 :1: 0.039
0.2 41 0.88 :t 0.016 2.09 t 0.043
0.4 43 0.88 :t 0.016 2.10 i 0.042
0.8 45 0.88 i 0.016 2.13 :I: 0.041
1.6 47 0.89 i 0.015 2.14 :t 0.040
3.2 10 0.90 i 0.033 2.71 :I: 0.096“

“ Doses expressed as Hm egg

° Sanple size (For % Body Weigh n=42 for 0.4 pg/kg egg and n=8 for 3.2 1.43/k8 egg)
° Data expressed as mean 1: standard error
“ Significamly differem from vehicle cohrol at p< 0.05

 

—=- ———- --1———————-—- — -- -- — —- — - --

 

Table 52. Effect of congener 126 on the actual and relative spleen weigh of hatchling White

 
 

 

 

 

 

 

Leghorn chickens.
Dose“ n" Weigflg)‘ % Body Weighc
Non-injected 50 0.015 1: 0.0005 0.034 :1: 0.0012
Vehicle 54 0.014 :t 0.0005 0.034 :1: 0.0012
0.1 49 0.013 :1: 0.0005 0.031 :t 0.0012
0.2 41 0.014 :1: 0.0006 0.033 :1; 0.0013
0.4 43 0.014 3: 0.0006 0.033 :1: 0.0013
0.8 45 0.013 i 0.0006 0.032 :1: 0.0013
1.6 46 0.014 t 0.0006 0.033 :1: 0.0013
3.2 10 0.013 :1; 0.0012 0.036 t 0.0030

“ Doses expressed as rig/kg egg

" Sanple size (For % Body Weigh n=42 for 0.4 [lg/kg egg and n=8 for 3.2 rig/kg egg)

° Data expressed as mean :1: standard error

 

 

 

100

DISCUSSION

Mortality: Seventy percent hatchability of cormorant eggs was considered
successful since the eggs were artificially incubated from day 0. Previous attempts to
incubate flesh cormorant eggs have met with limited success. No cormorants hatched with
an incubation temperature of 37.5°C and relative humidity of 67 % near the day of hatch
(larson, 1991). It has been noted that laboratory incubation of cormorant eggs is more
successful if eggs are collected after approximately a week of natural incubation by the
parents (Sodrerland, unpublished data). Cormorant eggs naturally incubated for two weeks
prior to their artificial incubation at 35.4°C with relative humidity of 46% had 77%
hatchability (Sanderson et al. , 1994).

Eggs collected in excess of those required for the PCB injection experiment were
used to collect additional information on the effects of artificial incubation on the
comorant. The average mass lost during incubation under the previously described
conditions was about 8.5 %. Cormorant eggs naturally incubated lose approximately 11%
of their initial mass (Sotherland, unpublished data). Only 54% of the eggs which were not
rotated by hand once a day hatched. Thus, this additional rotation improved the rate of
hatching. None of the fifteen eggs incubated widr their blunt end up throughout incubation
hatched. Sixty percent of these embryos died after 23 days, which indicates that
positioning or rotation is particularly important in the later stages of development.

Greater mortality in the vehicle control group may be due to random occurrence

of infertile eggs. Hatchability of the mock-injected eggs (70%) was equal to that of the

 

 

101

non-injected group thus suggesting that the injection procedure itself was not harmful. The
two smallest doses also had good hatchability (71%) which would indicate that the vehicle
was not harmful. The vehicle had no toxic effects in the chicken since the rate of hatching
in the vehicle control group was almost 92%. Despite great mortality in the cormorant
vehicle control, significant increases in mortality were observed at 400 and 800 jig/kg egg.
In another study there were no significant differences in mortality in cormorant eggs
injected into the yolk on day 0 with 18.4 pg congener 126/kg egg which is comparable to
results in the present study (Larson, 1991).

Mortality induced by exposure to congener 126 has been assessed in other chicken
egg injection studies. An LD,0 of 3.1 ag/kg eg for congener 126 in chickens was
calculated when this congener was injected into the air cell on day 7 and the eggs were
assessed for mortality 72 hours later (Brunstrom and Andersson, 1988). When 0.2 ag/kg
egg was injected into the yolk on day 4, they observed only 10% mortality by day 18.
Ninety percent mortality was observed when 2 reg/kg egg of congener 126 was injected the
yolk on day 4 (Brunstrbm, 1991), whereas the LD,0 determined in the present study was
2.4 rig/kg egg. In Chapter 1, the LD,0 for congener 126 was reported as 0.6 jug/kg egg.
The only differences in methods described in Chapter 1 and the methods reported in this
chapter are the vehicle and injection volume. In the initial trial, the vehicle used was that
described by Brunstrdm and colleagues (Brunstrt‘rm and Orberg, 1982; Brunstrém and
Darnerud, 1983) at a volume of 1 al/g egg as opposed to triolein at 0.1 ul/g egg which
was used in the present trial. Thus, both vehicle and injection volume can significantly

affect the LD,0 of a given compound. Given the greater rate of hatching in the present

 

102

trial, 2.4ag/kg egg is a more appropriate LDso for congener 126 in the chicken than the
LD,0 determined in chapeter l.

Teratogenicity: The predomimnce of edema in cormorants from all groups is most
likely due to humidity rather than PCB exposure. High humidity reduces evaporation from
the eggs and chicks that do not evaporate water fast enough are larger than normal. This
could explain the swollen appearance of most of the cormorants upon hatching. Sotherland
(unpublished data) also observed that dead embryos with edema had significantly less
fractional mass losses than dead embryos without edema or live hatchlings. Fox and
associates (1991) observed 52.1 bill defects per 10,000 cormorant chicks from Green Bay,
WI. Crossed or abnormal bills were also the most frequently reported abnormalities in
other colonial waterbirds (Fox et al., 1991). The present study did not show a causal link
between congener 126 and abnormalities such as bill deformities. However, given such
a low rate of occurrence in the wild, it is difficult to show the same effects in the
laboratory with the relatively small number of eggs injected. Injecting concentrations
greater than those observed in the environment would seem to produce detectable increases
in abnormalities, but, greater doses increased early embryo mortality thus reducing the
number of embryos in which abnormalities may be observed.

Despite the lack of the abnormalities in the cormorant under laboratory conditions,
significant increases in abnormalities were observed in the chicken at the three greatest
doses. Edema, microphthalmia, and beak deformities were common types of
abnormalities observed. The teratogenic capabilities of congener 126 have also been

reported for the chicken by Brunstr6m and Andersson (1988).

 

103
Body weight: Body weights and crown/rump lengths were unaffected in hatchling

cormorants. Several parameters have been compared in cormorant hatchlings from two
different colonies in the Netherlands, each with different levels of organochlorine
contamination (Van den Berg er al., 1994). Hatchlings from the more contaminated site
did not have body weights significantly different than chicks from the less contaminated
site although there was a tendency for them to be slightly lower. These authors reported
an inverse correlation between lower body weights and increased respiration and energy
metabolism and suggested drat these effects inhibit growth. The authors further suggested
that wasting syndrome may be responsible for this increase in respiration.

Chickens hatching from the 3.2 rig/kg egg dose group had significantly lower body
weights than those in the vehicle control group. This is in contrast to results in the first
trial in which body weights of hatchling chicks injected with congener 126 on day 0 did
not show any differences when compared to the vehicle control group (Chapter 1).
However, the highest hatching dose in first trial was 2.7 ag/kg egg which was lower than
the dose which caused a change in hatch body weights in the present trial.

Bruin: Relative brain weights were affected in hatchling chicks but not in hatchling
cormorants. Relative brain weights were greater in chicks hatchling from the 3.2 ug/kg
egg dose group. However, this group of chicks also had lower hatchling body weights.
Thus, the apparent effect observed in relative brain weight was influenced by the change
in body weight. Actual brain weights were not different when compared to the vehicle
control group. Brain weights were also unaffected in pelicans fed 100 mg of PCBs for 10

weeks (Greichus etal., 1975).

-..- _‘hrr

h‘kYBI .

 

 

104

Bursa of Fabricius: Relative bursa weights were not significantly different in
either hatchling cormorants or chickens. Actual bursa weights were significantly lower
in hatchling chickem in the 3.2 rig/kg egg dose group. Bursa weights were also lower in
hatchling chicks of hens fed PCBs for eight weeks (Harris et al., 1976). Both chicken and
cormorant hatchlings did show a trend toward lower relative bursa weights with increasing
dose indieating a potential effect on the immune system. Van den Berg and associates
(1994) recorded bursa weights in hatchling cormorants from contaminated and
uncontaminated colonies in the Netherlands, however they excluded them from the study
since the small size of the bursa did not allow for reliable measurement.

Heart: Congener 126 caused an increase in relative heart weights in both
cormorants (5 rig/kg egg) and chickens (3.2 tag/kg egg). The greater heart weights could
be due to an internal accumulation of fluid in the heart. Other studies with chlorinated
biphenyls have reported pericardial edema in exposed chicks (Mc Cune et al. , 1962; Flick
etal., 1963; Brunstrbm, 1988; Brunstrom, 1990; Brunstrbm etal., 1990). An alternate
explanation is an increase in the heart muscle itself (Walker, personal communication),
which seems to be the more plausible explanation since there were no obvious cases of
pericardial edema in the present study.

Liver: Relative liver weights were greater in both the cormorant and chicken at a
dose of 5 jug/kg egg and 3.2 jug/kg egg, respectively. A positive correlation between
concentrations of PCDDs, PCDFs, or mono-ortho PCBs in the yolk sacs of one-day-old
cormorants (Phalacrocorax aun'tus) and increased liver weights was observed in a field

study in Europe (Van den Berg et al., 1994). Greater liver weights have also been

 

 

105
observed in cockerels fed 50 ppm Aroclor' 1242, 1254, and 1268 or 500 ppm Aroclor'

1268 for 30 days (Cecil et al., 1978) and in pelicans fed 100 mg Aroclor’ 1254 for 10
weeks (Greichus et al., 1975). Increases in liver weight have been associated with
induction of detoxifying enzymes (Cecil et al., 1978). Congener 126 is known to induce
hepatic 7-ethoxyresorufin O-deethylase (EROD) in chicken embryos (Brunstrom and
Andersson, 1988; Brunstrbm, 1990).

Spleen: Cormorants hatching from the lowest dose group had greater relative
spleen weights than the vehicle control group, while relative spleen weights were
unaffected in chicken hatchlings. Spleen weights in White Leghorn chicks of hens fed
Aroclor' mixtures for eight weeks were less than spleen weights in control chicks (Harris
et al., 1976). Spleen weights were also lower in cockerels fed PCBs for 30 days (Cecil
et al., 1978). However, like the cormorant, pelicans fed 100 mg PCBs for 10 weeks had

significantly greater spleen weights than controls (Greichus et al., 1975).

SUMMARY AND CONCLUSIONS

Cormorants are less sensitive to the adverse effects of congener 126 than the
domestic chicken. In terms of mortality, the cormorant embryo is approximately 100
times less sensitive to the effects of congener 126 than the domestic chicken embryo.
Cormorants appear to be even less sensitive to the teratogenic effects of this congener.
The LOAEL (NOAEL) for the cormorant based on mortality was 283 (200)ug/kg egg,

while the LOAEL (NOAEL) for the chicken based on mortality, incidence of

106

abnormalities, body weight, and relative heart and liver weight was 2.3 (1.6) ag/kg egg.
TCDD exposure in the last third of incubation can cause induction of hepatic microsomal
EROD in both hatchling chickens and cormorants (Sanderson and Bellward, 1995). The
magnitude of induction indicated that the cormorant is one to two orders of magnitude less
sensitive than the chicken. Relative species sensitivities are helpful in assessing the
potential effects of contaminants on wild populations. However, it would not be advisable
to extrapolate from effects seen in a domestic precocial species to a wild altricial species.
Wild species have much greater genetic diversity than the domestic species. In addition,
the organ systems of the altricial and precocial chick develop at different stages (Dunn,
1975). These differences can lead to different toxicities making it difficult to predict the
outcome of contaminant exposure for one species based on the effects seen in another that

develops at a different rate.

 

107

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breeds to 3,3',4,4'-tetrachlorobiphenyl. Poultry Sci., 67:52-57.

Brunstrém, B. 1989. Toxicity of coplanar polychlorinated biphenyls in avian embryos.
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Brunstrom, B. 1990. Mono-ortho-chlorobiphenyls: toxicity and induction of 7-
ethoxyresorufm O-deethylase (EROD) activity in chick embryos. Arch. Toxicol. , 63:188-
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Fox, G.A., B. Collins, E. Hayakawa, D.V. Weseloh, J.P. Ludwig, T.J. Kubiak, and
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study. Master's Thesis, University of Wisconsin, Madison, WI, 120 pp.

Ludwig, J .P. 1984. Decline, resurgence, and population dynamics of Michigan and Great
lakes double-crested cormorants. Jack-Pine Warbler, 62:90-102.

North, Mack 0. Commercial Chicken Production Manual. Westport, Connecticut: AVI
Publishing Company, Inc., 1978. pp. 692.

McCune, E.L., J.E. Savage, and B.L. O'Dell. 1962. Hydropericardium and ascites in
chicks fed a chlorinated hydorcarbon. Poultry Sci., 41:295-299.

Safe, S. 1991. Polychlorinated dibenzo-p—dioxins and related compounds: sources,
environmental distribution and risk assessment. Environ. Carcinog. & Ecotox. Rev.,
C9(2):261-302.

Sanderson, J.T., R.J. Norstrom, J.E. Elliott, LE. Hart, K.M. Cheng, and GD.
Bellward. 1994. Biological effects of polychlorinated dibenzo-p-dioxins, dibenzofurans,
and biphenyls in double-crested cormorant chicks (Phalacrocorax auritus). J. Toxicol.
Environ. Health, 41:247-265.

Sanderson, J .T., G.D. Bellward. 1995. Hepatic microsomal ethoxyresorufin O-
deethylase-inducing potency in ovo and cytosolic Ah receptor binding affinity of 2,3,7,8-
tetrachlorodibenzo-p-dioxin: comparison of four avian species. Toxicol. Appl.
Pharmacol., (in press).

Tanabe, S., N. Kannan, An. Subramanian, S. Watanabe, and R. Tatsukawa. 1987.
Highly toxic coplanar PCBs: occurrence, source, persistency and toxic implications to
wildlife and humans. Environ. Pollut., 47:147-163.

 

109

Van den Berg, M., B.L.H.J. Craane, T. Sinnige, S. van Mourik, S. Dirksen, T.
Boudewijn, M. van der Gaag, I.J. Lutke-Schipholt, B. Spenkelink, and A. Brouwer. 1994.
Biochemical and toxic effects of polychlorinated biphenyls (PCBs), dibenzo—p-dioxins
(PCDDS) and dibenzofurans (PCDFS) in the cormorant after in ovo exposure. Environ.
Toxicol. Chem., 13:803-816.

Weseloh, D.V. and P.J. Ewins. 1994. Characteristics of a rapidly increasing colony of
double-crested cormorants (Phalacrocorax auritus) in lake Ontario: Population size,
reproductive and band recoveries. J. Great lakes Res. , 20:443-456.

Weseloh, D.V., S.M. Teeple, and M. Gilbertson. 1983. Double-crested cormorants of
the Great lakes: egg-laying parameters, reproductive failure, and contaminant residues
in eggs, lake Huron 1972-1973. Can. J. 2001., 61:427-436.

Yamashita, N., S. Tanabe, J.P. Ludwig, H. Kurita, M.E. Ludwig, and R. Tatsukawa.
1993. Embryonic abnormalities and organochlorine contamination in double-crested
cormorants (Phalacrocorax auritus) and caspian terns (Hydroprogne caspia) from the
upper Great lakes in 1988. Environ. Pollut., 79:163-173.

Hasn‘t}

 

 

CHAPTER 3

THE EFFECTS OF AN EXTRACT DERIVED FROM
CONTAMINATED DOUBLE-CRESTED CORMORANT
(Phalacrocorax auritus) EGGS ON THE CHICKEN

(Gallus domesticus)

110

 

 

111

ABSTRACT

White Leghorn chicken (Callus domesticus) eggs were injected prior to incubation
with one of four concentrations (0.001, 0.01, 0.1, and 1.0 egg equivalent) of an extract
derived from double-crested cormorant (Phalacmcorax aun'tus) eggs collected from Green
Bay, WI. Injection of 1.0 egg equivalent (the concentration of contaminants corresponded
to what was present in an average cormorant egg) resulted in 77 % mortality at hatch. The
incidence of abnormalities was not affected by injection of the extract. Body weight gain
was retarded in the 1.0 egg equivalent dose group in the second and third week post-hatch.
At three weeks of age, chicks were subject to necropsy. Relative brain weights were
greater and bursa weights were lower in the 1.0 egg equivalent dose group when compared
to the vehicle control. There were no significant differences in the relative weights of the

heart, liver, spleen, or comb.

112

INTRODUCTION

Double-crested cormorants (Phalacrocorax aun’tus) are one of the species of
piscivorous colonial waterbirds residing in the Great lakes that have experienced
reproductive problems. Cormorant populations began to decrease in the 1950s and 19605
due to destruction by commercial fisherman and egg breakage. They stopped breeding in
lake Michigan in 1963 and by the 1970s they were no longer breeding in lake Superior
or the Canadian waters of lake Ontario. They were the only species of Great lakes
waterbirds to cease breeding which suggests that they may be more sensitive to some
contaminants than other waterbird species. (Ludwig, 1984; Fox et al., 1991c).

DDT/DDE are environmental contaminants known to have had an adverse effect
on cormorant populations (Weseloh et al., 1983). Cormorants nesting in Lake Huron were
reported to have egg shells 24% thinner than normal (1972) leading to egg breakage
(Weseloh et al., 1983; Fox et al., 1991c). Since then, concentrations of DDT/DDE have
been decreasing, and as a result, egg shell thickness has increased. Thus, beginning in the
early 19805 cormorant populations have been increasing (Ludwig, 1984; Fox et al., 1991a;
Weseloh and Ewins, 1994).

Despite recovering populations, embryo lethality (Tillitt et al., 1991) and
congenital malformations have been observed in cormorant populations of the Great lakes
since the late 19703 (Fox et al., 1991c). These malformations have included deformed feet
and legs, missing or abnormal eyes, and crossed or abnormal bills (Fox et al., 1991b).

Since such malformations are uncommon in most wild avian populations (Fox et al.,

113

1991b), these observations prompted concern as to etiology. It has been suggested that
polyhalogenated diaromatic hydrocarbons (PHDHS) capable of inducing the hepatic
microsomal enzyme, aryl hydrocarbon hydroxylase (AHH), are responsible for these
effects (Fox et al., 1991a; Giesy et al., 1994a and 1994b). Observations of malformations
in cormorants and Forster's terns in Green Bay, WI were correlated with concentrations
of polychlorinated biphenyls (PCBs) in the environment (Gilbertson et al., 1991). PCB
concentrations have also been correlated with decreased hatchability of Forster's tern eggs
(Kubiak et al., 1989).

Toxic equivalency factors (T EFs) are frequently used in evaluating the toxicity of
PHDHs. PCB, polychlorinated dibenzo-p-dioxin (PCDD), and polychlorinated
dibenzofuran (PCDF) isomers are assigned a TEF value based on their relative toxicity
compared to the most toxic of these PHDHs, 2,3,7,8-tehachlorodibenzo—p-dioxin
(TCDD). Environmental extracts can be assessed for potential toxicity by multiplying the
concentrations of the dioxin-like contaminants by their respective TEFs to obtain toxic
equivalents (TEQS). Great concentrations of TEQS have been reported in the eggs of
waterb irds from polluted industrialized areas of the Great lakes where reproductive
impairment has been the most severe (Tillitt et al., 1991; Yamashita et al., 1993). These
areas include Green Bay, lake Michigan; Saginaw Bay, lake Huron; Hamilton Harbor,
lake Huron; and the Detroit River (Fox, 1993).

In this study, the domestic chicken (Callus domesticus) embryo was used to
examine the potential effects of an extract derived from cormorant eggs collected from

Green Bay, Lake Michigan, an area of relatively great contamination by PCDHs with a

114

history of compromised reproduction in resident colonies of double-crested cormorants

(Fox et al., 1993; Yamashita et al., 1993; Giesy et al., 1994a and 1994b).

MATERIALS AND NIETHODS

Egg Collection: Double-crested cormorant eggs were collected in 1988 from
Spider Island in Green Bay, lake Michigan (45°13'N, 86°59'W). All three, four, and
five egg clutches were collected on May 5, 13, and 19. Eggs were refrigerated until
opening. Egg contents were pooled and placed into chemically clean jars (I-Chem,
Hayward, CA). Samples were then frozen and transported on dry ice to the U.S. National
Biological Survey (Columbia, M0) on Feb. 26, 1991 where they were processed.
Samples were compositesof 10 to 20 eggs. These composites were given individual data-
base identification numbers and kept frozen at -20°C until sample preparation.

Extract Preparation: Egg contents were thawed and dehydrated with anhydrous
sodium sulfate. The mixture was homogenized, packed in an extraction column, and
extracted with methylene chloride. This extract was then concentrated by rotoevaporation
and consolidated. Egg lipids were dialyzed in 50 ml aliquants with 80:20 (v/v)
hexane/methylene chloride in 80 cm by 5 cm sections of layflat PE tubing (Brentwood
Plastics, Inc., St. Louis, MO) (Meadows et al., 1993). Dialysates for all extracts were
concentrated and combined. Cholesterol was removed by crystallization and filtration
from cold pentane. Gravity-flow gel permeation chromatography (GPC) was used to

remove any residual lipids in the egg extract. The fmal extract was in a volume of 10 mls.

115

Analysis of Extract: Capillary gas chromatography/electron capture detection
(CGC/ECD) was used to analyze aliquants of purified sample extracts for residues of
organochlorine pesticides (Schmitt et al., 1990) and mono-ortho-PCBs (Feltz et al., 1995).
CGC and mass spectrometry were used to analyze for PCDDs, PCDFs, and non-ortho-
PCBs (Feltz et al., 1995).

Egg Extract Injection Doses: The 1001 cormorant eggs used to prepare the extract
weighed 42,127 grams. Thus, an average cormorant egg weighed 42 g. Doses were
expressed in egg equivalents. One egg equivalent contained the amount of extract equal
to that derived from a cormorant egg weighing 42.085 g. The doses chosen were 1, 0.1,
0.01, and 0.001 egg equivalent. Doses were prepared with cold-filtered triolein as the
vehicle. The injection volume was 10 p1 per egg. Approximately 1.2 pl of the 10 pl
injected (vehicle or extract) was methylene chloride.

Injections: The extract doses and vehicle arrived at MSU in July, 1994. They
were kept frozen between injection days. There were three replications with 20 eggs per
group. Egg injections were performed as described in previous studies (Chapters 1 and
2).

Incubation, Post-Hatch, and Necropsy: Eggs were incubated as described in
Chapter 1. Mortality was assessed at hatching and abnormalities were looked for in all
nonviable embryos and hatchling chicks. Chicks were raised for three weeks as described
in Chapter 1. Body weights were recorded weekly. Three-week-old chicks were killed
by cervical dislocation and the brain, bursa, heart, liver, spleen, testes, and comb were

removed and weighed.

 

116

Data Analysis: All comparisons were made with respect to the vehicle control, due
to the relatively low hatchability of the vehicle control group. Mortality data were
evaluated using a 2x2 contingency table and Bonferroni Chi-Square ("infertiles" were
considered early deads since it was often difficult to distinguish between the two). The
same analysis was used to test the occurrence of abnormalities. Analysis of body weight
was conducted using univariate amlysis of variance for split-plot repeated measures (Gill,
1986) followed by Dunnett's test for comparisons with control. Organ weights were
analyzed using a one-way ANOVA and Dunnett's test. Analysis of body and organ
weights was performed with the statistical software SAS (SAS Institute Inc., 1992). The

level of significance was 0.05, unless otherwise stated.

RESULTS

Concentrations of predominant contaminants and their relative contribution to the
toxicity of the cormorant egg extract based on TEFs derived for the chicken (Bosveld et
al., 1995) are given in Table 53. The major contributors to toxic equivalents were PCB
congener 126 [3,3',4,4',5-PeCB, (79.95 %)]; 1,2,3,7,8-PeCDD (5.11%), PCB congener
77 [3,3',4,4'-TCB, (3.48%)]; 2,3,7,8—TCDD (3.37%); PCB congener 105 [2,3,3',4,4'-
PeCB, (3.03 %)]; and 2,3,4,7,8-PeCDD (2.65%).

One egg equivalent (1.0) was the only dose to produce significantly greater
mortality than the vehicle control (Table 54). Body weights of both males and females

were significantly lower at the second and third week post-hatch in the 1.0 egg equivalent

117

dose group (Table 55). Relative brain weights (Table 56) were greater in this group and
actual and relative bursa weights (Table 57) were less when compared to the vehicle
control values. The heart, liver, and spleen were not significantly affected by injection of
the cormorant egg extract (Tables 58-60). Comb weight (Table 61-62) was also
unaffected, although there was a tendency toward lower relative comb weights in the
extract-exposed males. Relative testes weights (Tables 63-64) showed no statistical

differences.

 

118

 

Table 53. Concernrations of selected polyhalogenated aromatic hydrocarbons and TEQ in

 

 

 

cormorah eg_g extract.
Cohaminah (pg/g)“ TEF“ TEQ“ gagg) at TEQS

2,3,7,8-TCDD 14.0 1.0 14.0 3.37
1,2,3,7,8-PeCDD 17.7 1.2 21.24 5.11
1,2,3,4,7,8-HxCDD 3.4 0.05 0.17 0.04
1,2,3,6,7,8-HxCDD 22.3 0.01 0.223 0.05
1,2,3,7,8,9-HxCDD 5.2 0.1 0.52 0.13

1,2,3,4,6,7,8-I-IpCDD 16.7 0.001 0.017 0
2,3,7,8-TCDF 0.70 0.9 0.63 0.15
1,2,3,7,8-PeCDF 0.27 0.3 0.081 0.02
2,3,4,7,8-PeCDF 10.00 1.1 11.0 2.65

1,2,3,4,7,8-HxCDF 2.00 0.01 0.02 0
2,3,4,6,7,8-HxCDF 1.30 0.1 0.13 0.03
1,2,3,6,7,8-HxCDF 1.60 0.4 0.64 0.15
3,3,4' ,4',5-PeCB (126) 3,320 0.1 332 79.95
3,3,4',4'-TCB (77) 722 0.02 14.44 3.48
3,3',4,4',5,5'-cha (169) 376 0.001 0.376 0.09
2,3,3',4,4'-PeCB (105) 247,000 5.1 x 105 12.60 3.03
2,3',4,4',5-PeCB (118) 711,000 4.1 x 10* 2.92 0.70
2,3,3',4,4',5-HxCB (156) 50,000 6.1 x 105 3.05 0.73
2,3,3',4,4',5'-HxCB (157) 15,000 7.1 x 105 1.07 0.26
2,3',4,4',5,5'-cha (167) 50,000 3.1 x 10* 0.155 0.04
415.269 100%

 

 

“ Vahres are expressed on a wet weigh basis

“Toxic Equivalency Factor (Bosveld et al., 1995) based on cytochrome P450 induction studies in

chicken embryos

° Toxic Etpivalems, relative to 2,3,7 ,8-TCDD

 

 

 

119

Tab1e54. Etfectofinjectionofcormoranteggextractihotheyolks ofWhiteLeghornchicken
eg_gsondayOofincubationonhatchability.

I! W
Dose“ TEQ“ TEQS/eggc I 126/egg“ I eggs

Non-injected 0 0 0 3/60
Vehicle 0 0 0 12/60
1““)3 0.415 1.7"‘10l 1.4"'104 16/62
1"'10r2 4.15 l.7"‘102 l.4"'10’3 6/61
1*10‘ 41.5 1.7*10’ 1.4'1'10'2 13/64
1.0 415 1.7*10‘ 1.4“10' 46/60“
“ Doses expressed in units of an egg equivalem (vehicle = triolein)
" Doses expressed in Toxic Equivalems (pg/g)
° Doses expressed in pg/egg (avg. cornmrah egg = 42g)

“ Doses expressed in pg PCB congener 126/egg (avg. cormoram egg = 42g)
° 5mm vehicle ml 0-05

 

Table 55. Effect of cormorant egg extract on the body weighs of White Leghorn chickens

 

 

 

through three weeks of age.
Dose“ n" SE° Hatch“ 1 Week“ 2 Weeks“ 3 Weeks“
Vehicle 47 2.15 43.6 72.2 125.6 187.0
1‘10“ 44 2.22 43.9 72.1 124.3 186.4
1"‘10‘2 54 2.00 43.6 71.4 124.2 185.2
1'10' 51 2.06 44.1 68.9 119.4 181.1
1.0 13 4.08 44.5 62.8 106.5“ 160.0“
“ Doses are expressed in units of an egg equivalent
" Sanple size
° Standard error (Not appropriate for comparison across time)
“ Data expressed as mean (g)

 

 

“ Significahly different from vehicle control at p < 0.05
— a

 

120

 

 

 

 

 

 
  
   

Table 56. Effectofcormorahegg extractontheacmalandrelativebrainweights oftlrree-week-
old White Leghorn chicks.
Dose“ n" Weigh (g)° % Body Weigh‘
Vehicle 47 1.69 i 0.014 0.914 1: 0.0149
1‘10“ 43 1.67 :I: 0.015 0.909 :1: 0.0156
1"10'2 54 1.66 :1: 0.013 0.908 :1: 0.0139
1"10l 51 1.64 :I: 0.013“ 0.913 :1: 0.0143
1.0 13 1.55 3; 0.026“ 0.994 :1: 0.0283“
“Dosesareexpressedintmitsofaneggeqrivaleh
° Sanple size
“Damexpressedasmean :t standarderror
“ Si ' ditfereh from vehicle cohrol at <0.05

 

 

 

 

 

 

 

 

 

 

 

Table 57. Effect ofcormorah egg extract on the actual and relative bursa weighs ofthree-week- 1
old White Leghorn chicks.
Dose“ n“ Weimg)‘ % Body Weigh° ‘
Vehicle 47 1.13 :t 0.047 0.601 :1: 0.0202
1’10“ 43 , 1.09 :t 0.050 0.582 t 0.0211
1"10‘2 54 1.13 i 0.044 0.600 :t 0.0188 W
1*10' 51 1.07 i 0.046 0.584 :1: 0.0194
I 1.0 12 0.62 i: 0.094“ 0.385 i 0.0399“
“Dosesareexpressedintmitsofaneggeqrivaleh
" Sample size
“Dataexpressedasmean :t standarderror
“ Significantly different from vehicle control at p<0.05

 

121

 

Table 58. Effect of cormorah egg extract on the actual and relative heart weighs of three-week-

 

 

 

old White lnghom chicks.
Dose“ n“ Weigh—(g)° % Body Weigh“
Vehicle 47 1.21 :I: 0.026 0.645 :1: 0.0094
1‘10“ 43 1.24 :1: 0.027 0.668 3; 0.0098
1‘10?2 54 1.23 :I: 0.024 0.667 i 0.0088
1'“10‘l 51 1.21 :I: 0.025 0.667 i 0.0090
1.0 13 1.01 :1: 0.049“ 0.642 i 0.0179

 

 

“Dosesareexpressedinimitsofaneggeqrivaleh

“Sample size
“Dataexpresscdasmean :1; standarderror

“ Significahly differeh from vehicle cohrol at p < 0.05

 

Table 59. Effect of cormoram egg extract on the actual and relative liver weighs of three-week-

 

 

 

old White Leghorn chicks.
Dose“ n" Weig_l_n_(g)° % Body Weig_ht°_
Vehicle 47 5.21 i 0.132 2.79 i 0.044
1‘10“ 43 5.11 :t 0.138 2.73 i 0.046
1"‘10‘2 54 5.29 :1: 0.123 2.86 :t 0.041
1'“‘10‘l 51 5.05 :1: 0.126 2.79 :t 0.042
1.0 13 4.54 1: 0.250 2.85 i 0.084

 

 

“ Doses are expressed in units of an egg equivalent

° Sanple size

“Dataexpressedasmean :1; standarderror

 

 

122

 

J

Table 60. Effect of cormorah egg extract on the actual and relative spleen weighs of three-

 

 

 

 

week-old White Leghorn chicks.
Dose“ n“ WeiM)‘ % BodLWeigh‘
Vehicle 47 0.247 1: 0.0118 0.131 :1: 0.0054
1‘10“ 43 0.258 :1: 0.0123 0.137 :I: 0.0057
1‘102 54 0.253 :1: 0.0110 0.136 :1: 0.0051
1’“‘10¢l 50 0.245 t 0.0114 0.134 :1: 0.0053
1.0 13 0.243 :1: 0.0224 0.146 :1: 0.0103

 

 

“Dosesareexpressedintmitsofaneggetpivaleh

“Sailple size
°Dataexpressedasmean i standarderror

123

 

 

 

 

 

—
Table 61. Effect of cormorah egg extract on the actual and relative cont weighs of female
three-week-old White Leghom chicks.
Dose“ n“ Weifl)° % Body WeigL ‘
Vehicle 22 0.052 1 0.0027 0.030 t 0.0015
1*10?3 24 0.050 :1; 0.0026 0.028 at 0.0014
l"'102 28 0.051 :t 0.0024 0.029 :1: 0.0013
1*10‘ 28 0.049 :t 0.0024 0.029 :1: 0.0013
1.0 6 0.050 1: 0.0052 0.032 i 0.0029
“Dosesareexpressed inimitsofaneggequivaleh

 

 

 

 

 

 

 

 

 

-
Table 62. Effect ofcormorah egg extract on the actual and relative comb weighs ofrmle three-
week-old White Eghorn chicks.
Dose“ n" Weigh (g)° % Body WeiL
Vehicle 24 0.376 i 0.0404 0.193 :1: 0.0213
1"‘103 19 0.242 1 0.0454 0.122 :t 0.0240
11'102 26 0.235 :1; 0.0388“ 0.120 :1: 0.0205
1‘10' 21 0.254 :1: 0.0432 0.135 i 0.0228
1.0 7 0.192 i 0.0749 0.115 :1: 0.0395
“Dosesareexpressedinunitsofaneggeqrivaleh
" Sanple size
° Data expressed as mean :1; standard error
a . . . .
Srgmficantly different from vehicle control at p < 0.05

 

 

 

 

Table 63. Effect of cormorant egg extract on the actual and relative left testes weighs of

three-week-old White Leghorn chicks.

 

Dose“ n" Weigh (g)° % Body Weigh“

Vehicle 24 0.028 :t 0.0016 0.015 :1: 0.0008
l"'103 19 0.026 :1: 0.0018 0.013 :1: 0.0009

1"‘10'2 26 0.028 i 0.0015 0.014 :1: 0.0008

1"‘10'l 23 0.025 t 0.0016 0.013 :t 0.0009

1.0 7 0.021 :t 0.0029 0.013 :1: 0.0016

 

 

" Sample size
I “Data expressedasmean :t standarderror

E

“Dosesareexpressedintmitsofaneggequivaleh

 

ll

 

Table 64. Effect of cormorant egg extract on the actual and relative righ testes weights of three-

 

 

 

 

 

week-old White leghom chicks.
I Dose“ n“ WeiM)“ % Body WeigL ‘
Vehicle 24 0.025 :1: 0.0015 0.013 i 0.0008
1"‘10'3 19 0.024 :1: 0.0017 0.012 :1; 0.0008
1"‘10‘2 26 0.025 :1: 0.0015 0.013 3: 0.0007
1 1’10" 23 0.022 :1: 0.0016 0.011 3: 0.0008
1.0 7 0.018 t 0.0028 0.011 :1: 0.0014
“ Doses are expressed in imits of an egg equivalent
" Sanmle size
“ Data expressed as mean :t standard error I

 

 

 

125

DISCUSSION

Total TEQs present in the cormorant egg extract (Table 53) were approximately
415 pg/g. These toxic equivalents were calculated using TEFs for chickens derived by
Bosveld et al. (1995). Total TEQs reported for cormorant eggs collected from Green Bay,
WI in 1988 by Yamashita et al. (1993) were around 1300 pg/g. The majority of this
difference in total TEQs is due to the difference in TEFs used to calculate the TEQ for
congener 126. Yamashita and associates used 0.3 while a TEF of 0.1 was used in the
present study. Total TCDD-EQs in cormorant eggs collected from Green Bay in the
summer of 1989 were approximately 382 pg/g (Jones etal., 1994).

There is an association between abnormalities in live cormorant chicks and areas
with eggs contaminated with 2,3,7 ,8—TCDF and coplanar PCBs (Yamashita etal., 1993).
No correlation was found for other organochlorine contaminants. Of the TEQs reported
by Yamashita et al. (1993), PCB congener 126 accounted for 83% of the TEQs and PCB
congener 105 accounted for 13%. Congener 126 also accounted for the majority of the
TEQs in the present cormorant extract (80%). This congener was present at 3,320 pg/g
egg which is consistent with the concentration (3,600 pg/g egg) determined in cormorant
eggs collected by Yamashita and associates (1993). PCB congeners 77 and 105 each
accounted for approximately 3 % of the TEQs in the present cormorant egg extract and
2,3,7,8-TCDD and 1,2,3,7,8-PeCDD accounted for over 8% of the TEQs.

Mortality: The injection of the cormorant egg extract in the present study produced

increased mortality at the same concentration as found in cormorant eggs in Green Bay,

126

WI in 1988. Since congener 126 was present in the extract at 3,320 pg/g egg and an
average cormorant egg weighed 42 g, each chicken egg received approximately 139,440
pg of congener 126. This corresponds to an injection dose of approximately 2.32 pg
congener 126/kg egg which is very close to the I..D,0 for congener 126 in chickens, 2.42
pg/kg egg (Chapter 2). While greater than 50% mortality occurred at 2.32 pg/kg eg
(77 %), which may be attrlbuted to the remaining components in the cormorant extract, the
primary effect still seems to be due to the concentration of congener 126. It is unlikely
that cormorants would experience increased mortality with this concentration of congener
126 since the LD,0 for this congener in cormorants is approximately 246 pg/kg egg
(Chapter 2).

Teratogenicity: There were no significant differences in the incidences of
abnormalities among any of the dose groups. The highest dose group, 1.0 egg equivalent,
provided a lower concentration of congener 126 (2.32 pg/kg egg) than did the doses of
congener 126 which caused an increased incidence of abnormalities in a previous injection
study, (3.2 - 12.8 pg/kg egg) (Chapter 2).

Body weights: The cormorant extract had a significant effect on body weights at
the second and third weeks post-hatch. Chicks from eggs injected with congener 126 (0.9,
2.7 pg/kg egg) also experienced lower body weights at the second and third week of
rearing (Chapter 1).

Organ and comb weights: Relative brain weights were greater in the 1.0 egg
equivalent dose group, however, this was probably a result of the significantly lower body

weights also observed in chicks of this dose group. The weights of the bursa of Fabricius

127

were significantly lower in chicks exposed to 1.0 egg equivalent prior to incubation.
Chicks exposed to congener 126 prior to incubation also had lower relative bursa weights
at three weeks post-hatch (Chapter 1). Relative heart, liver, and spleen weights were not
significantly different when compared to the vehicle control chicks. While not
significantly different in either extract-exposed males or females, the comb weights of
drree-week-old males were numerically lower with increasing concentration of extract.
This suggests that the extract may have had an anti-androgenic effect. Fry and associates
(1987) have repomd that organochlorine compounds can cause feminization of male germ
cells in the testes. New evidence is suggesting that feminizing effects are the result of
androgen blockage. One organochlorine contaminant, pp'-DDE, has been shown to bind
to androgen receptors, thus blocking the binding of testosterone (Culotta, 1995). When
one-month-old cockerels were fed 500 ppm Aroclor’ 1254 over a 12 week period, they
showed no change in comb weights, while controls showed a significant increase in
weight. These PCB-fed birds also had retarded gain in testes weight (Platonow and
Funnell, 1971). Mature cockerels fed 500 ppm of the same Aroclor’ for 30 days also had
lower comb and testes weights than the control birds (Cecil et al., 1978).

It is questionable whether any adverse effects would be observed if this extract
were to be injected into cormorant eggs at the levels naturally present in a single cormorant
egg. Gilman and associates (1978) injected organochlorine contaminants (DDE, PCBs,
mirex, photomirex, and hexachlorobenzene) extracted from lake Ontario herring gulls
into relatively uncontaminated, unincubated eggs collected from New Brunswick, Canada.

All eggs were injected in the field and naturally incubated. Contaminant-injected eggs had

128

no increase in embryo mortality suggesting that organochlorine contaminants in the egg
are not directly responsible for the increased incidence of embryo mortality observed in
lake Ontario herring gulls. However, it should be noted than the injection procedure
alone reduced the rate of hatching from 91 to 52%. The solvent (10% acetone in DMSO)
further reduced the rate of hatching, and 14 to 29% of all eggs used in the study were
destroyed by predators. In addition, most of the photomirex and about half the mirex
were lost in the preparation of the extracts. Gilman and associates (1978) suggested that
the toxic effects of contaminants may be expressed before the egg is laid. They offered
chromosomal alterations, sex cell viability, yolk quality, and changes in the structural
characteristics of the eg as possible explanations for the reduction in the rate of hatching

observed in herring gull eggs.

SUMMARY AND CONCLUSIONS

Injection of the contents of an average cormorant egg (1.0 egg equivalent) from
Green Bay, WI into White Leghorn chicken eggs prior to incubation had adverse effects
on the developing embryo. Mortality during incubation was greater, post-hatch body
weights were lower, and there was a reduction in the relative weight of the bursa of
Fabricius. Total TEQs in this extract were approximately 415 pg/g. Toxic equivalents
determined by the H4IIE bioassay (TCDD-EQS) have ranged from 49 to 415 pg/g in eggs
from double-crested cormorants and caspian terns residing in the Great lakes (Tillitt et al.,

1991). Tillitt et al. (1992), using the H4IIE bioassay, demonstrated a dose-response

129

relationship between PCBs and mortality of cormorant eggs from the Great lakes. PCBs
accounted for nearly 90% of the TEQs determined in the extract used for this egg injection

study and congener 126 accounted for nearly 80% of this toxicity.

130
REFERENCES

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and M. Van den Berg. 1995. Effects of PCDDs, PCDFs, and PCBs on common terns
(Sterna hirundo), breeding in estuarine and coastal colonies in the Netherlands and
Belgium. Environ. Toxicol. Chem., 14:99-115.

Cecil, H.C., 8.]. Harris, and J. Bitman. 1978. Liver mixed function oxidases in chicken:
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Culotta, E. 1995. St. Louis Meeting Showcases ”Creature Features” Science, 267:330-
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fractionation of PCDDs and PCDFs and plamr and nonplanar PCBs on Cls—dispersed PX-
21 earbon. Environ. Sci. Technol., 29:709-718.

Fox, G.A., B. Collins, E. Hayakawa, D.V. Weseloh, J.P. Ludwig, T.J. Kubiak, and
T.C. Erdman. 1991a. Reproductive outcomes in colonial fish-eating birds: a biomarker
for developmental toxieants in Great lakes food chains. J. Great lakes Res., 17:153-167.

Fox, G.A., M. Gilbertson, A.P. Gilman, and T.J. Kubiak. 1991b. A rationale for the
use of colonial fish-eating birds to monitor the presence of developmental toxicants in
Great lakes fish. J. Great lakes Res., 17:151-152.

Fox, G.A., D.V. Weseloh, T.J. Kubiak, and T.C. Erdman. 1991c. Reproductive
outcomes in colonial fish-eating birds: a biomarker for developmental toxicants in Great
Lakes food chains. J. Great lakes Res., 17:153-157.

Fox, GA. 1993. What have biomarkers told us about the effects of contaminants on the
health of fish-eating birds in the Great lakes? The theory and a literature review. J.
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Fry, D.M., C.K. Toone, S.M. Speich, and R.J. Peard. 1987. Sex ratio skew and
breeding pattems of gulls: demographic and toxicological considerations. Studies in
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Giesy, J.P., J.P. Ludwig, D.E. Tillitt. 1994a. Dioxins, dibenzofurans, PCBs and
Colonial, Fish-Eating Water Birds. In: Schecter, A. (ed) Dioxins and Health. Plenum
Press, New York. pp. 249-307.

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Giesy, J.P., J.P. Ludwig, D.E. Tillitt. 1994b. Deformities in birds of the Great lakes
region: assigning causality. Environ. Sci. Technol., 28:128A-l36A.

Gilbertson, M., T. Kubiak, J. Ludwig, and GA. Fox. 1991. Great lakes embryo
mortality, edema, and deformities syndrome (GLEMEDS) in colonial fish-eating birds:
similarity to chick-edema disease. J. Toxicol. Environ. Health, 33:455-520.

Gill, J.L. 1986. Repeated measurement: sensitive tests for experiments with few animals.
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Gilman, A.P., D.J. Hallett, G.A. Fox, L.J. Allan, W.J. Learning, and DB. Peakall.
1978. Effects of injected organochlorines on naturally incubated herring gull eggs. J.
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Jones, P.D., J.P. Giesy, J.L. Newsted, D.A. Verbrugge, J.P. Ludwig, M.E. Ludwig,
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cormorant (Phalacrocorax auritus, Pelicaniformes) chicks in the North American Great
lakes. Ecotox. Environ. Safety, 27:192-209.

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Sileo, D.E. Docherty, and T.C. Erdman. 1989. Microcontaminants and reproductive
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Contam. Toxicol., 21:91-101.

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Tillitt, D.E., G.T. Ankley, J.P. Giesy, J.P. Ludwig, H. Kurita-Matsuba, D.V. Weseloh,
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Weseloh, D.V., S.M. Teeple, and M. Gilbertson. 1983. Double-crested cormorants of
the Great lakes: egg-laying parameters, reproductive failure, and contaminant residues
in eggs, lake Huron 1972-1973. Can. J. Zool., 61:427—436.

Yamashita, N., S. Tanabe, J.P. Ludwig, H. Kurita, M.E. Ludwig, and R. Tatsukawa.
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upper Great lakes in 1988. Environ. Pollut., 79:163-173.

 

APPENDIX

133

134

Table 1. Concemations of polychlorinated dibenzo-p-dioxins (PCDDs) in cormorah

 

 

 

 

egg extract.
Cohaminah (pg) TEF“ TEQ (p -)“
2,3,7,8-TetraCDD 14.0 1.0 14.0
1,2,3,7,8-PehaCDD 17.7 0.5 8.85
l,2,4,7,8-PemaCDD ND“ - -
l,2,3,4,7,8-HexaCDD 3.4 0.1 0.34
1,2,3,6,7,8-HexaCDD 22.3 0.1 2.23
1,2,3 ,7,8,9-HexaCDD 5.2 0.1 0.02
1,2,3 ,4,6,7,8-HeptaCDD 16.7 0.01 0.133
OctaCDDs 202.3 0.001 0.2023
“ Toxic Equivalency Factor (TEF) (Safe, 1992)
" Toxic-Equivaleh

“ Not Detected at specified detection limit (0.1 pg/g)

135

Table 2. Concemations of polychlorinated dibenzofurans (PCDFS) in cormorah egg

 

 

 

extract.
Cohaminah (ng TEF- TEQ (pg/g):
2,3 ,7,8-TetraCDF 0.70 0.1 0.070
1,2,3 ,7,8-PehaCDF 0.27 0. 1 0.027
2,3 ,4,7,8-PerlaCDF 10.00 0.5 5.00
l,2,3,4,7,8-HexaCDF 2.00 0.1 0.20
1,2,3 ,6,7,8-HexaCDF 1.60 0.1 0.16
1,2,3 ,7,8,9-chaCDF NO“ 0.1 -
2,3 ,4,6,7,8-HexaCDF 1.33 0.1 0.133
I,2,3,4,6,7,8-HeptaCDF 0.80 0.1 0.08
1,2,3 ,4,7,8,9-HeptaCDF 0.67 0.1 0.067
OctaCDFs 1.73 0.001 0.00173

 

“ Toxic Equivalency Factor (TEF) (Safe, 1992)
. " Toxic-Equivalent
‘ ° Not at average concemation (0.2 ngg)

136

 

I I
I Table 3. Concemrations of non-ortho-polychlorinated biphenyl congeners in cormorant

 

 

 

 

 

 

 

 

 

egg extract.
PCB Congener (ng g) TEF“ TEQ (pg/g)“
126 3,320 0.1 322
77 722 0.01 7.22
169 376 0.05 18.8
81 852 - -
“ Toxic Equivalency Factor (TEF) (Safe, 1992) l
° Toxic-Eqrivaleh
Table 4. Concemrations of mono-ortho-polychlorinated biphenyl congeners in J
cormorah eg_g extract.
PCB Congener (ng/g) TEF“ TEQ (us/8)“ ,
123 6 0.001 0.006
118 711 0.001 0.711
114 16 0.001 0.016
105 247 0.001 0.247 I
167 50 0.001 0.050
156 50 0.001 0.050
157 15 0.001 0.015
189 6 0.001 0.006
“ Toxic Equivalency Factor (TEF) (Safe, 1992)
" Toxic-Equivalent +

 

 

 

137

 

Table 5. Concehrations of organochlorine pesticides in cormorant eg_g extract.

 

 

 

 

 

 

Cohaminah (Pg/i) Contamimm (Peg)
Hexachlorobenzene 11.7 cis Chlordane 23.6
alpha-BHC 3 .8 o,p' -DDE < MQL
Lindane <MQL“ Dieldrin 255.1
beta-BHC 4.9 p,p'-DDE 1147.2
Heptachlor < MDL“ o,p' -DDD 4.9
delta-BHC < MQL Entrin 12.8
Dacthal < MDL cis Nonachlor 43.1
Oxychlordane 80.3 o,p' -DDT < MDL
Heptachlor epoxide 88.3 p,p'-DDD 23.8
trans Chlordane < MQL p, p'-DDT 29.8
trans Nonachlor 13.8 Mirex 23.8
Total PCB 8872

 

“ Method Quantitation Limit (2.9 ng/g)
“ Method Detection Limit (1.0 ng/g)

 

 

138

 

Table 6. Concehrations of polychlorinated biphenyl congeners in cormorant egg extract.

 

 

 

 

 

Conggr (nag) Congener
004, 010 <1.1 064 10.6
007, 109 <1.1 040 <2.9
006 <1.1 067 4.8
005, 008 <1.1 063 9.0
019 <1.1 074 206.1
018 <1.1 070, 076 24.4
017, 015 <1.1 066, 095, 088 408.6
024, 027 <1.1 091 14.9
016, 032 <1.1 056, 060 34.0
029 <1.1 092 29.9
026 8.0 084 11.6
025 5.9 101, 090 65.4
031 <2.0 099 450.8
028 81.8 119 18.0
020, 033, 053 <1.1 083 <2.9
051 <1.1 097 5.5
022 <1.1 087 31.1
045 <1.1 110 18.0
046 <1.1 082, 151 48.3
052 45.3 135, 144, 124 6.9
049, 043 51.1 107 25.1
047, 048 221.1 123, 149 36.6
044 24.1 118 251.7
042 <2.0 134 <2.9
041 9.7 146 166.7

 

 

 

(us/g)

 

139

 

 

 

 

 

 

 

Table 6 (coh'd)

Copgener (11$) Congener (ng/ g)
153 466.4 185 5.3
132 9.2 174 5.5
105 34.1 177 69.6
141 9.9 202, 171, 156 74.5
179 <1.1 157, 201 13.9
137 37.5 172 24.2
176 <1.1 180 307.2
130 48.1 193 25.7
138 568.6 191 9.3
158 47.6 200 <1.1

129, 178 46.2 170, 190 127.5

182, 187 249.8 199 69.3
183 122.4 196, 203 65.3
128 172.4 208, 195 22.4
167 3.9 194 47.7

Total PCBs 5352 I

 

 

 

REFERENCES

Safe, S. 1992. Development, validation and limitations of toxic equivalency factors.
Chemosphere, 25:61-64.

 

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