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dissertation entitled

AN ECOLOGICAL RISK ASSESSMENT OF FISH-EATING
BIRDS EXPOSED TO POLYCHLORINATED
DIBENZOFURANS AND DIBENZO-P-DIOXINS WITHIN THE
TITTABAWASSEE RIVER FLOODPLAIN, MI, USA

presented by
RITA MARIE SESTON

has been accepted towards fulﬁllment
of the requirements for the

Doctoral degree in Zoology-Environmental ToxicologL

(74.5“ f ”Loam
/ Major Professpy’é Signatﬁ

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AN ECOLOGICAL RISK ASSESSMENT OF FISH-EATING BIRDS EXPOSED TO
POLYCHLORINATED DIBENZOFURANS AND DIBENZO-P-DIOXINS WITHIN
THE TITTABAWASSEE RIVER FLOODPLAIN, MI, USA

by

Rita Marie Seston

A DISSERTATION

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

DOCTOR OF PHILOSOPHY
Zoology-Environmental Toxicology

2010

ABSTRACT
AN ECOLOGICAL RISK ASSESSMENT OF F ISH-EATING BIRDS EXPOSED TO
POLYCHLORINATED DIBENZOFURANS AND DIBENZO-P-DIOXIN S WITHIN
THE TITTABAWASSEE RIVER F LOODPLAIN, MI, USA
by
Rita Marie Seston

Concentrations of polychlorinated dibenzofurans (PCDFs) and polychlorinated
dibenzo-p-dioxins (PCDDs) in the Tittabawassee River (TR) and associated ﬂoodplains
‘ downstream of Midland, MI, USA are greater than at upstream locations and regional

background concentrations. Sediments and ﬂoodplain soils in downstream study areas
(SAs) contain total concentrations of the seventeen 2,3,7,8—substituted PC DD/DF
congeners (ZPCDD/DFS) ranging from 1.0x102 to 5.4x104 ng/kg dry wt, respectively. In
contrast, concentrations of EPCDD/DFS in sediments and soils from upstream reference
areas were 10- to 20-fold less. The majority of the contaminant mixture is composed of
2,3,7,8-tetraehlorodibenzofuran and 2,3,4,7,8-pentachlorodibenzofuran, which are likely
present is the result of historical chemical production and associated waste management
practices. Concerns about potential ecological impacts of the elevated concentrations of
PCDD/DFs within the TR floodplain led to a site-speciﬁc multiple lines of evidence
study was executed including dietary- and tissue-based exposures assessments and
measurements of population health. Two ﬁsh-eating bird species that breed along the TR
[great blue heron (Ardea herodias; GBH) and belted kingﬁsher (Ceryle alcyon;BKF)]
were monitored both upstream and downstream of the putative source in order to
elucidate the potential for contaminant driven adverse population-level effects.

Additionally, measured exposures were compared to toxicity reference values (TRVs),

and reproductive parameters were compared to literature values. During the 2005—2007
breeding seasons, a total of 37 BKF nest chambers were excavated for sample collection
and monitored for reproductive effort and success. For GBH, three breeding colonies
located within the SA were monitored during the 2006 and 2007 breeding seasons. Nests

within each colony were also accessed for sample collection. Concentrations of
XPCDD/DF 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents (TEQWHo—Avian) in both
eggs and nestlings of BKF from the SA were 5- to 21afold greater than in those from
upstream reference areas (RAs). Concentrations of TEQWHQAvgan blood plasma of adult

GBH from the SA was 4- to 8-fold greater compared to those from the RA. Contaminant
concentrations in GBH eggs and nestlings were similar among all studied breeding
colonies. Predicted dietary exposures followed this same spatial trend in both species,
being 150— to l90-fold greater along the TR compared to upstream RAs. Comparison of
the predicted daily dietary dose to the TRV suggested there was the risk of adverse
effects as a result of exposure to PCDD/DFs. This is in contrast to the conclusions drawn
from both the tissue—based exposure and effects assessments and site—speciﬁc measures of
individual and population health. This inconsistency is likely the result of the dietary
exposure and effects assessments being more conservative, based on the greater number
of assumptions that must be made and the greater uncertainty associated with the dosing
methodology from which the TRV was derived. Therefore, the overall conclusion of the
research presented herein is that the populations of BKF and GBH breeding along the TR

are not at risk despite elevated concentrations of PCDD/DFs in the diet and tissues.

To my big sister _ for blazing the trail

ACKNOWLEDGEMENTS

There are a countless number of people whose friendship and guidance have helped me
make it through this degree program. Dr. John Giesy provided an environment which
allowed me to grow professionally and personally through trial and error. All of my
fellow colleagues in the Wildlife Toxicology Laboratory (Tim Fredricks, Dusty Tazelaar,
Emily Koppel, Lori Williams, Will Folland, Mike Nadeau, Casey Bartrem, Dave
Hamman, Patrick Bradley, Mick Kramer, Nozomi Ikeda, and many others) provided
invaluable support in the pursuit research, recreational, and social objectives. I have the
utmost gratitude toward the local landowners and parks in the area of the Tittabawassee
River for allowing us to traipse around their property at all hours of the day and for
sharing their local wisdom with us. Their kindness and openness was pivotal to the
success of the research project. I truly appreciate my family for not giving up on me even
though it seems I could never remember to call home to let everyone know what I was up
to. Last, but certainly not least, thanks to Dr. Matt Zwiernik. His guidance, support,
encouragement, and belief in me helped me to succeed in spite of myself. Because of

him, I now believe I am capable of just about anything. Thank you.

TABLE OF CONTENTS

LIST OF TABLES ........................................................................................................... viii

LIST OF FIGURES ......................................................................................................... xiii

KEY TO ABBREVIATIONS .......................................................................................... xvi

Chapter 1

INTRODUCTION .............................................................................................................. 1
Overview ....................................................................................................................... 2
Site Description ............................................................................................................. 3
Contaminant Description .............................................................................................. 8
Toxicology of Dioxin-Like Compounds ..................................................................... 12
Selection of Receptor Species ..................................................................................... l7
Site-Speciﬁc Receptor Species ................................................................................... 19
Research Objectives .................................................................................................... 23
Permits, Approvals, and Funding ................................................................................ 23
References .................................................................................... ‘ ............................... 25

Chapter 2

UTILIZING THE GREAT BLUE HERON (ARDEA HERODIAS) IN ECOLOGICAL

RISK ASSESSMENTS OF BIOACCUMULATIVE CONTAMINANTS ...................... 35
Abstract ....................................................................................................................... 36
Introduction ................................................................................................................. 37
Species Applicability .................................................................................................. 40
Methods ....................................................................................................................... 41
Results ......................................................................................................................... 49
Discussion ................................................................................................................... 55
Acknowledgements ..................................................................................................... 59
Animal Use ................................................................................................................. 60
References ................................................................................................................... 61

Chapter 3

DIETARY EXPOSURE OF GREAT BLUE HERON (ARDEA HERODIAS) TO

PCDD/DFS IN THE TITTABAWASSEE RIVER FLOODPLAIN, MI, USA ......... ' ...... 65
Abstract ....................................................................................................................... 66
Introduction ................................................................................................................. 67
Methods ....................................................................................................................... 69
Results ......................................................................................................................... 79
Discussion ................................................................................................................... 91
Acknowledgements ..................................................................................................... 99

vi

Animal Use ............................................................................................................... 100
References ................................................................................................................. I 1 1

Chapter 4

TISSUE-BASED RISK ASSESSMENT OF GREAT BLUE HERON (ARDEA
HERODIAS) EXPOSED TO PCDD/DF S IN THE TITTABAWASSEE RIVER

FLOODPLAIN, MI, USA ............................................................................................... 1 18
Abstract ..................................................................................................................... 119
Introduction ............................................................................................................... 1 20
Methods ..................................................................................................................... 123
Results ....................................................................................................................... 134
Discussion .............................................................................................. 7 ................... I 54
Conclusions ............................................................................................................... 1 64
Acknowledgements ................................................................................................... 165
Animal Use ............................................................................................................... 166
Supplemental Information ....................................................................... 167

Chapter 5

MULTIPLE LINES OF EVIDENCE RISK ASSESSMENT OF BELTED KINGFISHER
EXPOSED TO PCDFS AND PCDDS IN THE TITTABAWASSEE RIVER

FLOODPLAIN, MIDLAND, MI, USA .......................................................................... 195
Abstract ..................................................................................................................... 196
Introduction ............................................................................................................... I97
Methods ..................................................................................................................... 199
Results ....................................................................................................................... 212
Discussion...............- .................................................................................................. 229
Conclusions ............................................................................................................... 241
Acknowledgements ................................................................................................... 242
Animal Use ............................................................................................................... 242
Supplemental Information ........................................................................................ 244
References ................................................................................................................. 25 1

Chapter 6

CONCLUSIONS ............................................................................................................. 259
Comparison of Receptor Species .............................................................................. 260
Overall Conclusions............’ ...................................................................................... 273
References ............................................................................................................. 276

vii

LIST OF TABLES

Table 1.1. Avian toxic equivalency factors (TEFs) from the World Health Organization
(WHO) for the 17 2, 3, 7, 8-chlorine substituted PCDD/DF congeners and 12 PCB

congeners. ......................................................................................................................... I I

Table 2.1. Description of sampling effort and summary of great blue heron tissues
collected through utilizing described methodologies. Note that there was no rookery
located in the reference area to sample. Necessary adult plasma sample size in reference
area obtained after 2006 ﬁeld season. ............................................................................... 51

Table 2.2. Total TEanOmian (pg/mL) in adult GBH blood plasma from Tittabawassee
River study area collected during 2005 ﬁeld season. .................................. g ..................... 54

Table 3.1. TEQWHO.A,,,am in prey items of great blue herons collected during 2004-2006
from the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.
Values (ng/kg wet wt) are given as the geometric mean and sample size in parentheses
(11) over the 95% conﬁdence interval and range (min-max). ............................................ 80

Table 3.3. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
sediment collected during 2003-2006 from the Chippewa/Pine, Tittabawassee, and
Saginaw Rivers, Midland, MI, USA. Values (ng/kg w) are given as the arithmetic
mean :1: 1 SD over the range. ........................................................................................... 101

Table 3.4. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
frogs collected during 2005- 2006 from the within the Chippewa, Tittabawassee, and
Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg w) are given as the
arithmetic meanb :1: 1 SD over the range. ........................................................................ 103

Table 3.5. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
crayﬁsh collected during 2005-2006 from the within the Chippewa, Tittabawassee, and
Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg w) are given as the
arithmetic mean :I: 1 SD over the range. ......................................................................... 105

Table 3.6. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
forage ﬁsh composites collected during 2004-2007 from the within the Chippewa,
Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg w)
are given as the arithmetic mean i 1 SD over the range. ............................................... 107

Table 3.7. Concentrations of twelve dioxin-like polychlorinated biphenyl congeners in
forage ﬁsh composites collected during 2004-2007 from the within the Chippewa,
Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values (pg/kg w)
are given as the arithmetic mean i 1 SD over the range. ............................................... 109

Table 3.8. Concentrations of twelve dioxin-like polychlorinated biphenyl congeners in
frogs and crayﬁsh collected during 2004-2007 from the within the Chippewa,

viii

Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values (pg/kg w)
are given as the arithmetic mean :i: 1 SD over the range. ............................................... 1 10

Table 4.1. Total concentrations of 2,3,7,8-substituted furan and dioxin (EPCDD/DF) and
TEanamm in blood plasma of great blue herons collected during 2005-2007 as a
result of trapping along the Chippewa and Tittabawassee River ﬂoodplains, Midland, MI,
USA. Values (pg/mL wet wt) are given as the geometric mean with sample size in
parentheses (n) over the 95% conﬁdence interval and range (min-max). ...................... 135

Table 4.2. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF),
TEQWH0_Avian , and co-contaminants in great blue heron eggs collected during 2006-
2007 from the Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Values
(ng/kg wet wt) are given as the geometric mean (n) over the 95% conﬁdence intervals
and range in parentheses. DDX values are reported in ug/kg wet wt. ........................... 140

Table 4.3. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF) and

TEanaAviaU in great blue heron nestling tissues collected during 2006-2007 from the
Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Values are given as
the geometric mean with sample size given in parentheses (12) over range (min-max).
PCDD/DF and PCB values are reported as ng/kg, wet wt in tissues and pg/mL, wet wt in
plasma. DDX values are reported in ug/kg wet wt in tissues and ng/mL in plasma. 142

Table 4.4. Summary statistics for correlations of great blue heron nestling tissue
contaminant concentrations. Spearman rank correlations of wet weight contaminant

concentrations; including TEQWHO-Avian (DF-TEQ), ZPCDD/DF, 2,3,4,7,8—
pentachlorodibenzofuran (2,3,4,7,8-PeCDF), 2,3,7,8—tetrachlorodibenzoﬁ1ran(2,3,7,8-
TCDF), 2,3,7,8-tetrachlorodibenzo—p-dioxin (2,3,7,8-TCDD), and EPCB. ................... 146

Table 4.5. Measured and predicted and measured concentrations of XPCDD/DFS and

DF- TEQWH0.Avian in great blue heron eggs collected from rookeries within the
Tittabawassee and‘Saginaw river ﬂoodplains during 2006-2007. Predicted concentrations
were calculated using plasma to egg relationship developed from eggs and nestling
plasma collected from the same nest. Egg concentrations are reported in ng/kg, wet wt.
......................................................................................................................................... 149

Table 4.6. Reproductive parameters for great blue heron rookeries located in the
Tittabawassee and Shiawassee river ﬂoodplains from 2006-2007. Values are given as the
arithmetic mean i 1 SD over the sample size given in parentheses (n). ........................ 155

Table 4.7. Concentrations of 2,3,7,8-TCDD equivalents (TEQs) in eggs and nestling
tissues of great blue herons collected during 2006-2007 from rookeries within the
Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg wet

Wt) are given as geometric mean (sample size) over the 95% conﬁdence interval. ....... 167

ix

Table 4.8. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
blood plasma of great blue herons collected during 2005-2007 as a result of trapping
within the Chippewa and Tittabawassee River ﬂoodplains, Midland, MI, USA. Values
(ng/kg wet wt) are given as the arithmetic mean :t 1 SD over the range. ....................... 168

Table 4.9. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
eggs of great blue herons collected during 2006-2007 from rookeries within the
Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values
(ng/kg wet wt) are given as the arithmetic mean :1: 1 SD over the range. ....................... 170

Table 4.10. Concentrations of selected co-contaminants in eggs of great blue herons
collected during 2006-2007 from rookeries within the Chippewa, Tittabawassee, and
Saginaw River ﬂoodplains, Midland, MI, USA. Values (pg/kg wet wt) are given as the
arithmetic mean :I: 1 SD over the range. ......................................................................... 172

Table 4.11. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners

in blood plasma of great blue herons nestlings collected during 2006-2007 from rookeries
within the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.
Values (ng/kg wet wt) are given as the arithmetic mean i 1 SD over the range. ........... 174

Table 4.12. Concentrations of selected co-contaminants in blood plasma of great blue
herons nestlings collected during 2006-2007 from rookeries within the Chippewa,

Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg wet
wt) are given as the arithmetic mean i 1 SD over the range. ......................................... 176 .

Table 4.13. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners
in adipose of great blue heron nestlings collected during 2006-2007 from rookeries within
the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.

Values (ng/kg wet wt) are given as the arithmetic mean i 1 SD over the range. ........... 177

Table 4.14. Concentrations of selected co-contaminants in adipose of great blue heron
nestlings collected during 2006-2007 from rookeries within the Chippewa,

Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values (pg/kg wet
wt) are given as the arithmetic mean i 1 SD over the range. ......................................... 179

Table 4.15. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners
in liver of great blue heron nestlings collected during 2006-2007 from rookeries within
the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.

Values (ng/kg wet wt) are given as the arithmetic mean i: 1 SD over the range. ........... 181

Table 5.1. TEQWHGAvian in prey items of belted kingﬁsher collected during 2004-2006
from the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.
Values (ng/kg w) are given as the geometric mean TEQWHOMan and sample size in
parentheses (11) over the 95% conﬁdence interval and range (min—max). ...................... 213

Table 5.2. Predicted daily dietary dose of ZPCDD/DFS and TEQSWHO-AVian (mg/kg body
weight/d) for adult belted kingﬁsher breeding during 2004-2006 within the Chippewa,
Tittabawassee, and Saginaw river ﬂoodplains, Midland, Michigan, USA, based on the
geometric mean (95% conﬁdence interval) of site-speciﬁc dietary items ...................... 217

Table 5.3. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF) and

TEanaAvian belted kingﬁsher eggs and nestlings collected during 2005-2007 from the
Chippewa, Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Values
(mg/kg w) are given as the geometric mean (n) over the 95% conﬁdence intervals and
range in parentheses. ZPCB and ZDDX values are reported in ug/kg ww. .................. 218

Table 5.4. Parameters of reproductive effort and success of belted kingﬁshers breeding
along the Chippewa and Tittabawassee river ﬂoodplains during 2005-2007. ................ 228

Table 5.5. Concentrations of 2,3,7,8-TCDD equivalents (TEQs) in eggs and nestlings of
belted kingﬁsher collected during 2005—2007 from within the Chippewa and
Tittabawassee River ﬂoodplains, Midland, MI, USA. Values (ng/kg w) are given as the
geometric mean (sample size) over the 95% conﬁdence interval ................................... 244

Table 5.6. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
eggs and nestlings of belted kingﬁsher collected during 2005-2007 within the Chippewa
and Tittabawassee River ﬂoodplains, Midland, MI, USA. Values (ng/kg w) are given
as the arithmetic mean :1: 1 SD over the range. ............................................................... 245

Table 5.7. Concentrations of selected co—contaminants in eggs and nestlings of belted
kingﬁsher collected during 2005-2007 within the Chippewa and Tittabawassee River
ﬂoodplains, Midland, MI, USA. Values (pg/kg w) are given as the arithmetic mean i 1
SD over the range ........................................................................................................... -. 247

Table 5.8. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
soils collected during 2006 from nest chambers of belted kingﬁshers within the
Chippewa and Tittabawassee River ﬂoodplains, Midland, MI, USA. Values (ng/kg w)
are given as the arithmetic mean :1: 1 SD over the range. ............................................... 249

Table 6.1. TEQWHO.,5,,,ian in prey items of belted kingﬁsher and great blue heron
collected during 2004-2006 from the Chippewa, Tittabawassee, and Saginaw River
ﬂoodplains, Midland, MI, USA. Values (ng/kg 1w) are given as the geometric mean
TEQWHO.AV;an and sample size in parentheses (11) over the 95% conﬁdence interval and
range (min—max) .......................................................................................................... 263

Table 6.2. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF) and
TEQWH0-Avian in the diet of BKF and GBH collected during 2005-2006 from the
Tittabawassee River ﬂoodplain, Midland, Michigan, USA, expressed with varying

dietary compositions. Values (ng/kg 1w) are given as the geometric mean over the 95%
conﬁdence intervals. ....................................................................................................... 267

xi

Table 6.3. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF) and
TEQWH0-Av;an in eggs of belted kingﬁsher and great blue heron collected during 2005~
2007 from the Chippewa, Tittabawassee and Saginaw River ﬂoodplains, Midland, MI,
USA. Values (ng/kg 1w) are given as the geometric mean (21) over the 95% conﬁdence
intervals and range in parentheses. ................................................................................. 268

Table 6.4. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF) and

TEQWH0-AV;an nestlings of belted kingﬁsher and great blue heron collected during 2005-
2007 from the Chippewa, Tittabawassee and Saginaw River ﬂoodplains, Midland, MI,
USA. Values (ng/kg 1w) are given as the geometric mean (n) over the 95% conﬁdence
intervals and range in parentheses. ................................................................................. 269

xii

LIST OF FIGURES

Figure 1.1. Study site locations within the Chippewa, Tittabawassee, and Saginaw River
ﬂoodplains, Michigan, USA. Reference Areas (R-l to R-2), Tittabawassee River Study
Areas (T-3 to T-6), and Saginaw River Study Areas (S-7 and S-9). Direction of river
ﬂow is designated by arrows; suspected source of contamination is enclosed the dashed
oval ...................................................................................................................................... 6

Figure 1.2. Chemical structure of 2,3,7,8~tetrachlorodibenzo—p-dioxin (TCDD), and
general dibenzofuran and biphenyl rings with potential halogenation sites numbered. 10

Figure 2.1. Location of great blue heron rookeries (F RE, SNWR, and CAS) and reaches
within the Tittabawassee River study area, MI, USA where trapping occurred ............... 39

Figure 2.2. Number of great blue herons trapped in the Tittabawassee River study area
between 2005 and 2007 using the bait station and foot-hold trapping method. Number of
GBH normalized to the number of trapping hours. .......................................................... 53

Figure 3.1. Sampling locations for dietary components along the Chippewa,
Tittabawassee, and Saginaw river ﬂoodplains, Michigan, USA. Reference area (R-1 and
R-2; RA); Upper Tittabawassee River (T-3 and T-4; UTR); Lower Tittabawassee River
(T-S, T-6, and S-7; LTR); and Saginaw River (S-8 and S-9; SR). ................................... 71

Figure 3.2. Percent mean contribution of individual 2,3,7,8-substituted congeners to
ZPCDD/DF in dietary items collected from reference (RF), Upper Tittabawassee (UT),
Lower Tittabawassee (LT), and Saginaw River (SR) reaches. ......................................... 84

Figure 3.3. Range of hazard quotients for dietary-based exposure of great blue herons to

DF-TEQth/HQ.AV,an and total TEQSWH0-AVian along the Chippewa, Tittabawassee, and
Saginaw river ﬂoodplains based on assumption of 100% (A) and 75% (B) site-use. ...... 88

Figure 4.1. Location of great blue heron rookeries within the Tittabawassee River
ﬂoodplain study area, Midland, MI, USA. ..................................................................... 124

Figure 4.2 Principle component analysis of PCDD/DF concentration congener proﬁles in
blood plasma of great blue heron collected during 2005-2007 from the Chippewa,
Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Symbols are labeled
by area (RA==Reference Area; SA=Study Area) and age class (HY=hatch year;
AHY=after~hatch year). Ellipses indicate 95% conﬁdence intervals of each group.
Individual PCDD/DF congener loading scores for each principle component is depicted

in the inset. TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzoﬁiran;

TCDD = tetrachlorodibenzo-p-dioxin; PeCDD = pentachlorodibenzo-p-dioxin; OCDD =
octachlorodibenzo-p-dioxin. ........................................................................................... I 37

xiii

Figure 4.3. Pattern of percent mean ZPCDD/DF congeners in blood plasma collected
from hatch year (HY) and after hatch year (AHY) great blue herons trapped within the
Tittabawassee River study area, Midland, MI, USA. ..................................................... 139

Figure 4.4. Patterns of percent mean ZPCDD/DF congeners in great blue heron tissues
collected from the FRE (a), SNWR (b), and CAS (c) rookeries, located within the
Tittabawassee and Saginaw river ﬂoodplains, MI, USA. ............................................... 144

Figure 4.5. Plasma to egg relationship for ZPCDD/DF (a) and DF—TEQw-H()-Avian (b) for
great blue herons from the Tittabawassee and Saginaw river ﬂoodplains (N=11). Line of
best ﬁt with 95% conﬁdence intervals. ........................................................................... 147

Figure 4.6. Modeled probabilistic distribution of expected cumulative percent

frequencies for great blue heron egg TEQWHO-A,,;an concentrations ng/kg wet wt in site-
speciﬁc eggs collected from the river ﬂoodplains near Midland, Michigan in 2005-2007.

Concentrations of DF- TEQWHOAvian and total TEQWHO-Avian indicated by dashed and
solid lines, respectively. NOAEC and LOAEC indicated by vertical bars ................... 150

Figure 4.7. Range of hazard quotients for great blue heron tissues collected from the
Tittabawassee and Saginaw River ﬂoodplains. Upright triangles represent values based
on lowest observed adverse effect levels (LOAECs) and inverted triangles represent
values based on no observed adverse effect levels (NOAECs). 95% conﬁdence limits
also presented. ................................................................................................................. 153

Figure 5.1. Assessment area along the Chippewa, Tittabawassee, and Saginaw river
ﬂoodplains, Michigan, USA. Sampling locations for dietary components of belted
kingﬁsher were located in the Reference area (R-1 and R2); Upper Tittabawassee River
(T-3 and T—4); Lower Tittabawassee River (T—5, T-6, and S7); and Saginaw River (8-8
and S-9). Tissues of belted kingﬁsher were collected in the reference area (RA) and
along the Tittabawassee River (study area; SA). ............................................................ 200

Figure 5.2. Percent mean contribution of individual 2,3,7,8—substituted congeners to
EPCDD/DF in eggs and nestlings of belted kingﬁsher collected from reference (RA) and
study (SA) areas. ............................................................................................................. 220

Figure 5.3. Range of hazard quotients for dietary-based exposure of belted kingﬁsher to

DF-TEQSWHO_A,ian and total TEQSWHO_AVian along the Chippewa, Tittabawassee, and
Saginaw river ﬂoodplains. 223

Figure 5.4. Range of hazard quotients for DF—TEstnonrvm and total TEQSwno-Avian in

eggs of belted kingﬁsher from within the Chippewa and Tittabawassee river ﬂoodplains.
......................................................................................................................................... 224

xiv

Figure 5.5. Modeled probabilistic distribution of expected cumulative percent
frequencies for concentrations of TEQWHO-Avian in eggs of belted kingﬁsher collected
from the river ﬂoodplains near Midland, MI in 2005-2007. Sampling locations included
a reference area (RA) and study area (SA). Vertical bar represents NOAEC (USEPA
2003). LOAEC (11090 ng total TEQWHO.Avian/kg ww) not included on plot. Note break
in x-axis ........................................................................................................................... 225

Figure 5.6. Correlation plot of percent hatching success and DF-TEstnOAvian in eggs
of belted kingﬁsher for nesting attempts with data collected for both variables from the
river ﬂoodplains near Midland, Michigan during 2005-2007. R- and p-values and

sample size indicated; RA=reference area and SA=study area. Nesting attempts with data
for both hatching success and concentrations of total TEQSW’HQ-AVian also plotted but not
included in correlation analysis. ..................................................................................... 227

XV

KEY TO ABBREVIATIONS

ADDpot — potential average daily dose

ARN T — AhR nuclear translocator

AhR — aryl hydrocarbon receptor

AH Y — after hatch year

BKF — Belted kingﬁsher (C eryle a/cyon)

BW — body weight

°C — degrees centigrade

CI —— conﬁdence interval

CNC ~ Chippewa Nature Center

COC — chemicals of concern

COPEC — compound of potential environmental concern
CYP IA — cytochrome P4501A

d ~ day

DEQ r Department of Environmental Quality

DNA ~ deoxyribonucleic acid

DOW ~ The Dow Chemical Company

DDT ~ diehloro-diphenyl-trichloroethane

DDE ~ dichloro-diphenyl-dichloroethylene

DDXs ~— dichlorodiphenyltrichloroethane and related metabolites
DRE - dioxin-responsive element

dW ~— dry weight

xvi

ERA — ecological risk assessment

EROD ~7-ethoxy-resorufn-0—deethylase
g — gram

GBH — Great blue heron (A rdea herodias)

GHO — Great horned owl (Bubo virginianus)

HpCB — heptachlorinated biphenyl
HpCDD — heptachlorodibenzo-p-dioxin

HpCDF — heptachlorodibenzoﬁiran

HQ —- hazard quotient

HxCB — hexachlorinated biphenyl

HxCDD — hexachlorodibenzo-p-dioxin

HxCDF ~ hexachlorodibenzofuran

IACUC — Michigan State University’s Institutional Animal Care and Use Committee
IR — intake rate

km - kilometer

kg _ kilogram

LCSO — lethal concentration for 50% of dosed

In ~ natural log

xvii

LOAEL(C) —— lowest observed adverse effect level (concentration)
In — meter

MDEQ — Michigan Department of Environmental Quality

MI — Michigan

pg - microgram

MSU-WTL ~ Michigan State University-Wildlife Toxicology Laboratory
ng — nanogram

NOAEL(C) ~ no observed adverse effect level (concentration)

OC — organochlorine pesticide

OCDD ~ octachlorodibenzo-p-dioxin

OCDF — octachlorodibenzofuran

PCB ~ polychlorinated bi phenyls

PCDD — polychlorinated dibenzo-p-dioxins

PCDF — polychlorinated dibenzoﬁrrans

PeCB -— pentachlorinated biphenyl

PeCDD —- pentachlorodibenzo-p-dioxin

PeCDF ~ pentachlorodibenzofuran

R-1 and R-2 — speciﬁc reference areas

RA ~ reference area

S-7 to S9 — speciﬁc Saginaw River study areas

SA ~ study area

xviii

SC — scientiﬁc collection
SNWR ~ Shiawassee National Wildlife Refuge

SR — Saginaw River

T—3 to T-6 — speciﬁc Tittabawassee River study areas
TCB — tetrachlorinated biphenyl

TCDD — tetrachlorodibenzo-p-dioxin

TCDF ~ tetrachlorodibenzofuran

TEF — Toxic equivalency factor

TEQ ~ 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalent
TR — Tittabawassee River

TRB — Tittabawassee river basin

TRV ~ toxicity reference value

USA — United States of America

USEPA — United States Environmental Protection Agency
USFWS — United States Fish and Wildlife Service
WHO ~ World Health Organization

Ww ~ wet weight

y — year

xix

CHAPTER 1

Introduction

Rita Marie Seston

£3!

Overview

The Tittabawassee River downstream of Midland, Michigan, USA, has been shown to
contain elevated concentrations of polychlorinated dibenzofurans (PCDFs) and
polychlorinated dibenzo-p—dioxins (PCDDS). This contamination is a result of historical
chemical production processes and waste management techniques performed at the Dow
Chemical Company Mid-Michigan Plant located in Midland. Due to the persistence of
these compounds of potential environmental concern (COPECs), resident wildlife species
may be exposed to concentrations of PCDFs and PCDDS which could potentially affect
their health, at either the individual or population level. To understand both exposure and
associated effects, ﬁeld studies were conducted over several years in support of a site-
speciﬁc ecological risk assessment. This assessment utilizes a multiple lines evidence
approach to reduce the uncertainty which is often inherent in the risk assessment process.
To most accurately assess an ecosystem as complex as the Tittabawassee River
ﬂoodplain, the site-speciﬁc risk assessment employs many different receptor species,
each with multiple lines of evidence utilized. Receptor species were chosen to be
representative of the various feeding guilds that have the greatest potential for exposure
to the COPECs. Due to the tendency of PCDD/DPS to bioaccumulate through trophic
transfer, species located near the t0p of food webs were selected as the most appropriate
receptors. With a primarily aquatic—based exposure pathway, the American mink
(Mustela vison) was chosen as the top-tier mammalian receptor. Several passerine
species were selected as intermediate avian receptor species of aquatic-, terrestrial-, and
combined-based exposure pathways, the great horned owl (Bubo virginianus) was

selected as the top-tier avian receptor of a primarily terrestrial-based exposure pathway,

and the great blue heron (Ardea herodias; GBH ) and belted kingﬁsher (Ceryle alcyon;
BKF) were selected as top-tier avian receptors of an aquatic-based exposure pathway.
Dietary exposure and tissue-based exposure, combined with assessments of individual
and population health of each receptor species, are the multiple lines of evidence to be
employed in assessing the risk present to resident species of the Tittabawassee River
ﬂoodplain. Integrating the data resulting from these multiple assessments provides a
better understanding of the contaminants and their movement within the ecosystem. In
turn, this reduces the uncertainty inherent in the risk assessment process and provides
sound data for use in making decisions regarding the future of the site.

The focus of the research described in this dissertation is the movement of COPECs
through an aquatic—based exposure pathway, monitoring population-health parameters,
and assessing overall risk posed by PCDD/DFs to piscivorous avian species nesting in the
Tittabawassee River ﬂoodplain. The GBH and BKF were chosen as the receptor species

that could best meet this objective.

Site Description

Located in the east-central lower peninsula of Michigan, the Tittabawassee River
(TR) is a tributary of the Saginaw River (SR), which eventually empties into Saginaw
Bay and Lake Huron. The TR runs through The Dow Chemical Company (DOW)
property, which is located on the southern edge of Midland and is the accepted source of
the PCDD/DP contamination. The area henceforth referred to as the study area (SA)
includes approximately 37 km of the TR (sites T-3 to T-6) and associated wetlands from

DOW to the conﬂuence of the TR and SR and 35 km of the SR (sites S-7 to S-9) until it

enters Saginaw Bay (Figure 1.1). Sampling sites located in the SA were chosen to
characterize maximal exposure potential designated as “worst case scenario” locations
based on a previous study that measured soil and sediment concentrations (Hilscherova et
al. 2003) and availability of landowner access to sites. The reference area (RA) is
composed of the TR upstream of Midland, together with the Pine and Chippewa Rivers,
both of which are tributaries of the TR upstream of Midland. Sampling locations in the
RA were on the upstream TR (R-1) and on the Pine River, just upstream of its conﬂuence
with the Chippewa River (R-2). Distinct sampling areas were assessed individually as
well as grouped spatially based on river characteristics. Spatial groupings included
reference (RA) R—1 and R-2, upper Tittabawassee River (UTR) T—3 and T4, lower
Tittabawassee River (LTR) T-5 to S-7, and SR S-8 and S—9. Components of each
receptor species diet were collected from these sampling areas while the area of
collection of the tissue of receptor species was determined by nest location.

Founded in 1897, the Midland plant is DOW’s headquarters and was historically their
primary chemical research and manufacturing facility. Initial operations at DOW were
focused on mining brine and extracting bromine and chlorine to produce brominated and
chlorinated compounds. Using an electrolytic process that generated chlorine from brine,
bleach was Dow’s dominant product until its production stopped in 1914, although the
production of other chlorine-based products continued. Electrolytic processes using
carbon electrodes were used at DOW until the 19805. Although these processes have
ceased, PCDD/DPS were likely released into the environment as unwanted by-prOducts.
Concentrations of PCDD/DPS in sediments and ﬂoodplain soils downstream of Midland

were 10- to 20- fold greater than those upstream, yielding concentrations ranging from

102 to 53,600 pg/g dry weight (dw) (Hilscherova et al. 2003). Rivers in industrialized
areas of the eastern United States, including the Housatonic River and the Passaic River,
have sediment contamination of PCDD/DFs ranging from 160 to 5,400 pg/ g dw (with one
sample at 82,000 pg/g dw) and 370 to 24,000 pg/g dw, respectively (Eitzer 1993;
Wenning et a1. 1992). The Dutch River Rhine has PCDD/DPS concentrations ranging
from 200 to 18,000 pg/g dw (Evers et al. 1988). Sediments collected from Masan Bay,
Korea contained PCDD/DPS concentrations ranging from 122 to 16,729 pg/g dry wt
(Kannan et al. 2007). In Tokyo Bay, Japan, sediment concentrations of PCDD/DPS range
(from 3,150 to 20,300 pg/g dry wt. (Sakurai et al. 2000). These studies indicate that
PCDD/DPS concentrations found in the TR study area are elevated but remain
comparable to those reported in a number of other industrialized locations.

In addition to PCDD/DPS, concentrations of polychlorinated biphenyls (PCBs) and
organochlorine pesticides (OCs) were measured in the soils and sediments of the TR.
Concentrations of total PCBs in sediments of the TR were less than 150 ng/ g dry weight
(Hilscherova et al. 2003; Michigan Department of Environmental Quality 2002).
However, downstream of the TR, in the SR and Saginaw Bay, PCBs have been measured
at relatively great concentrations (Froese et al. 1998; Giesy et al. 1997; Kannan et al.
2008; Ludwig et al. 1993). From approximately 1936 until the early 19705, various
brominated and chlorinated compounds were manufactured at the Michigan
Chemical/Velsicol Corporation, and were subsequently released into the Pine River, a
tributary of the Chippewa River and eventually the TR. Dichlorodiphenyltriehlordethane
(DDT) was released into the Pine River environment due to activities at the Velsicol

Chemical Company and consequently became a primary contaminant of concern in the

Figure 1.1. Study site locations within the Chippewa, Tittabawassee, and Saginaw River
ﬂoodplains, Michigan, USA. Reference Areas (R-l to R-2), Tittabawassee River Study
Areas (T-3 to T-6), and Saginaw River Study Areas (S-7 and S-9). Direction of river

ﬂow is designated by arrows; suspected source of contamination is enclosed the dashed

oval.

 

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area as relatively great concentrations of DDT and its metabolites (DDXs) were measured
in Pine River sediments and various ﬁsh species (Michigan Department of Environmental
Quality 2000). Continued presence of DDXs within this system and the risk they may
pose to wildlife residing in the TR SA led to their inclusion as COPECs. Of all identiﬁed
COPECs, PCDD/DPS in the TR ﬂoodplain remain the primary concern due to their

elevated concentrations in comparison to predicted toxic thresholds.

Contaminant Description

PCDD/DPS and PCBs are chemically classiﬁed as halogenated aromatic
hydrocarbons. There are a total of 75 PCDD congeners and 135 PCDF congeners.
PCDDS consist of two benzene rings with varying numbers and positions of chlorine
atom substitutions, connected by two oxygen atoms. PCDFs are structurally very similar,
differing in that the two benzene rings are joined by only one oxygen atom. PCB
congeners consist of two benzene rings connected by a single C—C bond with varying
numbers and positions of chlorine atom substitutions. Of 209 PCB congeners, 12 are
coplanar congeners that are either mono-ortho or non-ortho substituted and are
structurally and conformationally similar to PCDD/DPS. Those PCDD/DP congeners
with chlorine atoms substituted at the 2,3,7,8—positions exhibit the greatest toxicity, and
are thus of the greatest interest (7 PCDD and 10 PCDF congeners). The congener
thought to be the most potent and thus most widely studied of these compounds is
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). The structure of TCDD and related
compounds are shown in Figure 1.2. This suite of compounds will be referred to as

dioxin-like compounds throughout the remainder of this document.

Despite their structural relatedness, each PCDD/DP and PCB congener has unique
physical-chemical properties that affect its fate, transport, bioavailability, and toxicity
(Eisler 1986). In order to investigate the complex mixtures of these compounds which
occur in the environment, the concept of toxic equivalency factors (TEFs) has been
developed (van den Berg 8! al. 1998).. Speciﬁc TEFs were developed for mammals,
birds, and ﬁsh to account for differences in sensitivities between these taxa. Based on the
assumption that all dioxin—like compounds exhibit their toxicity through the same
mechanism of action, the toxicity of a mixture of these compounds should be additive.
TEFs have been developed for the 17 2,3,7,8-substituted PCDD/DP congeners and 12
structurally related PCB congeners (Table 1.1). Using TEFs, concentrations of each
congener can be converted to TCDD equivalents (TEQs) to assess the overall toxic
potential of a mixture of compounds. Although this scheme is useful in gaining an
understanding of the toxic potential of a mixture of these compounds, it is important that
TEQs are not used to compare between soils/sediments and biota or movement through
trophic levels.

Although natural combustion and geological processes may result in trace quantities
of PCDD/DPS, essentially all PCDD/DPS in the environment are unwanted byproducts
from various anthropogenic activities (Czuczwa er al. 1984; Schecter er al. 1988). The
formation of PCDD/DPS can occur during various combustion processes, including the
incineration of municipal and industrial solid waste (Goovaerts et al. 2008; Lasagni et al.
2009; Lustenhouwer et al. 1980). Dioxinvlike compounds are also formed through
various processes at paper and pulp mills. For example, the use of elemental chlorine in

the bleaching process and the use of wood contaminated with polychlorophenolic-based

 

 

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\/___.\ \, \/

(CI)n 5 6 5' (CI)n

 

 

biphenyl

./ \ ”T \.

\\ //
(CI)n/2\\\/ \/’/\8/(Cl)n
Cl\2/1\IOIQ\B/m
c1/3\4/ o %7\

6 Cl

 

 

TCDD

Figure 1.2. Chemical structure of 2,3,7,8-tetraehlorodibenzo-p—dioxin (TCDD), and

general dibenzofuran and biphenyl rings with potential halogenation sites numbered.

10

Table 1.1. Avian toxic equivalency factors (TEFs) from the World Health
Organization (WHO) for the 17 2,3,7,8-chlorine substituted PCDD/DP congeners and
12 PCB congeners.

 

 

PCDD/DFsa TEFb PCBsC TEFb
Polychlorinated dibenzmp-dioxins Non-ortho-substituted

2,3,7,8-TCDD 1 3,3’,4,4’-TCB (77) 0.05
1,2,3,7,8-PeCDD I 3,4,4’,5-TCB (81) 0.1
1,2,3,4,7,8-HxCDD 0.05 3,3,’,4,4’,5—PeCB (126) 0.1
1,2,3,6,7,8-HxCDD 0.01 3,3’,4,4’,5,5’-HxCB (169) 0.001
1,2,3,7,8,9-HxCDD 0. 1

1,2,3,4,6,7,8-HpCDD <0.001

1,2,3,4,6,7,8,9-OCDD 0.0001

Polychlorinated dibenzofurans Mono-ortho-substituted

2,3,7,8-TCDF 1 2,3,3’,4,4’-PeCB (105) 0.0001
1,2,3,7,8-PeCDF 0.1 2,3,4,4’,5-PeCB (114) 0.0001
2,3,4,7,8-PeCDF 1 2,3’,4,4’,5-PeCB (118) 0.00001
1,2,3,4,7,8—HxCDF 0.1 2’,3,4,4’,5—PeCB (123) 0.00001
1,2,3,6,7,8-HxCDF A 0.1 2,3,3’,4,4’,5-HXCB (156) 0.0001
l,2,3,7,8,9-HxCDF 0.1 2,3,3 ’,4,4’,5 ’-HxCB(157) 0.0001
2,3,4,6,7,8-HxCDF 0.1 2,3’,4,4’,5,5’-HxCB (167) 0.00001
1,2,3,4,6,7,8-HpCDF 0.01 2,3,3’,4,4’,5,5’-HpCB (189) 0.00001
1,2,3,4,7,8,9-HpCDF 0.01

1,2,3,4,6,7,8,9-OCDF 0.0001

 

 

a TCDD = tetrachlorodibenzo-p-dioxin; PeCDD = pentachlorodibenzo-p- dioxin;
HxCDD= hexachlorodibenzo-p-dioxin; HpCDD= heptachlorodibenzo-p-dioxin;
OCDD == octachlorodibenzo-p-dioxin; TCDF = tetrachlorodibenzofuran; PeCDF =
pentachlorodibenzofuran; HxCDF = hexachlorodibenzofuran; HpCDF =
heptachlorodibenzofuran; OCDF= octachlorodibenzofuran

b

van den Berg etal.1998

CTCB= tetrachlorinated biphenyl; PeCB: pentachlorinated biphenyl; HxCB
hexachlorinated biphenyl; HpCB= heptachlorinated biphenyl

II

11

wood preservatives leads to mill efﬂuents which contain PCDFs and PCDDS (Bright et
al. 1999; Clement et al. 1989; Swanson et al. 1988). PCDFs and PCDDS are
inadvertently created during the production of chlorine and chlorinated compounds, such
as chlorophenols (Hutzinger et al. 1985; Rappe et al. 1991). The persistent and lipophilic
nature of PCDD/DPS leads to their accumulation in soils and sediments, and makes
improper waste storage and the disruption of contaminated sites important sources of
these compounds into the environment (USEPA 1994b). The predominant congeners
found in the TR SA, TCDF and OCDD, suggest the contamination originated from the
production of chlorophenol or chlorobenzene, general chlor-alkali processes, or
associated waste products (Hilscherova et al. 2003; Kannan er al. 1998; Kannan et al.

2008)

Toxicology of Dioxin—Like Compounds

Since the discovery of its potency, TCDD has been an extensively studied compound.
Dioxin—like compounds bind with high afﬁnity to the aryl hydrocarbon receptor (AhR), a
ligand-activated nuclear transcription factor (Safe 1986). Activation of this receptor-
mediated pathway results in a diverse array of effects, including biochemical adaptive
changes such as enzyme induction, developmental deformities, reproductive failure,
hepato—toxicity, immuno-toxicity, carcinogenicity, wasting syndrome, and eventually
death. TCDD is one of the most toxic of the dioxin-like compounds, as it binds with the
greatest afﬁnity to the AhR (Poland and Knutson 1982). Structurally related compounds
bind the AhR with varying degrees of afﬁnity and thus their toxicity varies (Safe 1986).

Free AhR resides in the cytoplasm, but upon binding with a ligand, such as TCDD, the

12

AhR translocates to the nucleus. In the nucleus, AhR dimerizes with its DNA binding
partner, the AhR nuclear translocator (ARNT). This heterodimer then binds speciﬁc
DNA response elements, known as dioxin-responsive elements, leading to an increase in
the transcription of certain genes, such as mammalian CYP1A1 and CYP1A2 and avian
CYP1A4 and CYP1A5. However, the relationship between the induction of these genes
and the toxicity of these compounds is not completely understood (Schmidt and Bradﬁeld
1996)

The toxicity of dioxin-like compounds, particularly TCDD, has been well established
in birds. Egg injection studies using the domestic chicken (Gallus gallus) have calculated
50% lethal dose (LDSO) values that range between 122-297 pg/g wet wt TCDD (Allred
and Strange 1977; Blankenship et al. 2003; Henshel et al. 1997; Powell et al. 1996;
Verrett 1976). Developmental abnormalities observed after in ovo exposure to TCDD
and other dioxin-like compounds in various avian species include edema of the head and
neck, microphthalmia (reduced eye size), liver damage, and skeletal and beak deformities
(Blankenship et al. 2003; Hoffman el al. 1998; Powell et al. 1996; Sanderson and
Bellward 1995). Comparative egg injection studies have shown a large difference in
species sensitivity to the toxic effects of dioxin-like compounds exists. Chickens have
been shown to be up to 250-fold more sensitive than turkeys (Meleagris gallopavo),
pheasants (Phasianus colchicus), ducks (mallard (Anas platyrhynchos) and goldeneye
(Bucephala clangula), domestic geese (Anser anser), herring gulls (Larus argentatus),
and black-headed gulls (Larus ridibundus) (Brunstrom 1988; Brunstrom and Lund 1988;

Brunstrtim and Reutergardh 1986). Other species studied include the American kestrel

13

(Falco sparverius) and common tern (Sterna hirundo), in which EROD activity was 800-
and IOOO—fold less, respectively, than in the chicken (Hoffman et al. 1998).

In addition to laboratory studies, the toxic effects of dioxin-like compounds have
been observed in ﬁeld studies of avian species. In the 19605, population declines in
colonial ﬁsh-eating birds of the Great Lakes were largely attributed to exposure to high
levels of PCBs (Gilbertson et al. 1991). For instance, Ludwig er al. (1996) determined
there was a relationship between TCDD-EQs and the incidence of embryonic deformities
and death rates in double-crested cormorants (Phalacrocorax auritus) and Caspian terns
(Hydroprogne caspia) nesting in colonies in the Great Lakes. A later study also observed
decreased hatching success and increased incidence of nestling deformities in a colony of
double-crested cormorants located in an area along Lake Michigan contaminated with
PCBS when compared to a reference colony (Larson 8! al. 1996). Impaired reproductive
success of Forster’s terns (Sterna forsteri) was associated with a median egg TCDD
concentration of 37 pg/g wet wt (Kubiak et al. 1989). The concentration of TCDD in
eggs appeared to be related to the severity of reproductive failure observed in colonies of
herring gulls around the Great Lakes, although a casual relationship could not be
established (Gilbertson 1983). The symptoms observed were consistent with those of
chick-edema disease in domestic chickens, which include edemas, hydropericardium,
ascites, liver enlargement, porphyria, liver necrosis with fatty degeneration, and high rate
of mortality, following exposure to AhR active compounds (Gilbertson 1983; V03 1972;
V03 and Koeman 1970). In wildlife exposed to these contaminants, this Suite of
symptoms has been labeled Great Lakes embryo, mortality, edema, and deformities

syndrome (GLEMEDS) (Gilbertson er al. 1991). Common cormorants (Phalacrocorax

14

carbo) with hepatic total-TEQ concentrations ranging between 12-1900 pg/g wet wt
exhibited a correlation between TEQs and CYPIA protein levels (Kubota et al. 2005).
CYPIA protein and EROD induction in bald eagle (Haliaeetus leucocephalus) hatchlings
were also correlated with TCDD, TCDF, and TEQs (Elliott at al. 1996).

Nesting colonies of GBH near paper and pulp mills in British Columbia have been
monitored for exposure to the dioxin-like compounds present in mill efﬂuents. In 1987,
colony failure coincided with a 3-fold increase in TCDD concentrations (210 pg/g wet
wt), but the authors determined the poor productivity was most likely a result of
disturbance due to human or bald eagle activity and subsequent predation by crows
(Corvus caurinus) and ravens (Corvus corax) (Elliott er al. 1989). Paired eggs were
collected from nests within these same breeding colonies, contaminant concentrations
were measured in one and the other was artiﬁcially incubated and allowed to hatch in a
laboratory (Bellward et al. 1990; Hart et al. 1991; Sanderson et al. 1994). Hatching
success was not different between eggs collected from control and contaminated colonies,
but subcutaneous edema was observed in 4 of 12 hatchlings from the most contaminated
colony (211 pg/ g wet wt TCDD) (Hart et al. 1991). Comparisons between hatchlings and
eggs from the same clutch revealed that TCDD concentrations regressed negatively with
growth measures in hatchlings, including yolk-free body weight, tibia length, and organ
weights, and positively with hepatic microsomal EROD activity in GBH hatchlings
(Bellward et al. 1990; Hart 81' a]. 1991). Continued monitoring of these colonies has
shown a decrease in the severity of these effects as environmental concentrations of
PCDDS and PCDFs have decreased due to process changes implemented by the pulp

industry (Sanderson et al. 1994). Deformities in pipping embryos from other GBH

15

breeding colonies near paper and pulp mills were observed with TCDD concentrations in
eggs ranging from 1.7-8.3 pg/g wet wt, but the overall reproductive success of the
colonies was not affected (Thomas and Anthony 1999).

Dioxin-like compounds have also been found in herons from areas not associated
with the paper and pulp industry. Skeletal deformities, namely multiple fractures of the
tarsus and tibia and metacarpal bones, were present in grey heron (Ardea cinerea)
nestlings in the United Kingdom with total TEQs in nestling adipose tissue ranging from
300-640 pg/g wet wt (Thompson et al. 2006), but no statistical relationship between
contaminant concentrations and the frequency of these deformities was reported. Pipping
embryos of black—crowned night-herons (Nycticorax nyelicorax) with PCB-TEQs of
approximately 290 pg/g wet wt had elevated EROD activity compared to those collected
from associated reference areas, but no gross abnormalities were observed. Reproductive
success, measured as clutch size and hatching success, of GBH was not adversely
affected with concentrations of PCB-TEQs in nestling adipose tissue ranging from 13-
100 pg/g lipid weight (1w) (Straub et- al. 2007). Concentrations of PCB-TEQS
(approximately 48 pg/g wet wt) and DF—TEQs (11 pg/g wet wt) in GBH eggs collected
along the Mississippi River were too low to induce EROD activity (Custer er al. 1997).
Although the TEQ scheme is helpful in summing the toxic potential of the PCDD, PCDF,
and PCB congeners, comparisons between studies reporting TEQs must be done
cautiously, as different studies may have reported different or incomplete congener lists
and used different TEFs to arrive at their end values. The best attempt to reduce these
discrepancies was done in the above study summaries by converting reported values with

the TEFs reported by van den Berg (1998) when possible.

16

There is very limited information available on the presence and the potential effects
of dioxin-like compounds in BKF. In China, eggs of a different species of kingﬁsher, the
lesser pied kingﬁsher (Ceryle rudis), collected from an area of PCDD/DP contamination
contained 1.6 pg/g wet wt of DF-TEQS (Fang et al. 2007). Total PCB concentrations in
adult and juvenile BKF collected along the Sheboygan River in Wisconsin, another site
of PCB contamination, had concentrations of total PCBs ranging from 65-220 jig/g
(Heinz et al. 1984). Unfortunately, these studies only reported tissue residues and no
data on potentially associated individual or population health effects were presented.
Eggs of BKF collected from along the Hudson River, which is an area heavily
contaminated with PCBs, contained concentrations of total PCBs ranging from 2 to 80
ug/g and total TEQs ranging from 100 to 5200 pg/g wet wt (geometric mean of 620 pg/g
wet wt) (Custer et al. 2010). No relationship was observed between these contaminant

concentrations and reproductive success.

Selection of Receptor Species

Selection of an appropriate species is a key element of effective ecological risk
assessments (ERA), especially when site-speciﬁc ﬁeld studies are to be employed. When
selecting a species to serve as a receptor in a site—speciﬁc ecological risk assessment, the
intensity of that species exposure to the COPEC(s) must be considered (USEPA 1994a).
In general, PCDD/DPS in the environment are predominately associated with particulate
matter, such as sediments, suspended material, and soils. Although the uptake of
PCDD/DPS from contaminated soil into plant tissue is very limited (Hﬁlster and

Marschner 1993; Welsch-Pausch et al. 1995), other organisms are exposed through the

17

incidental ingestion of contaminated particulate matter and the consumption of prey
items. The lipophilic nature and resistance to biological degradation of PCDD/DPS make
these compounds likely to bioaccumulate up the food web. Consequently, species
located at the top of the food chain are the most likely to experience the greatest exposure
to PCDD/DPS through biomagniﬁcation. As an example, ﬁsh, ducks, and ﬁsh—eating
birds that were part of a food web associated with sediments contaminated with
PCDD/DFs had much greater concentrations of these compounds in their tissues than
aquatic vegetation and benthic invertebrates collected in the same area (Wu et al. 2001).
When selecting a receptor species to investigate bioaccumulative compounds, such as
PCDD/DPS, it should be a species that is situated near the top of the food chain,
representing the greatest dietary exposure potential on a body weight normalized basis.
Another important factor to consider during the selection of a receptor species is the
relative sensitivity of a species to a contaminant (USEPA 1994a). It is not feasible to
study all species present at a site, so only those which are sensitive to the COPECs should
be considered for selection as a receptor. Despite the fact that terrestrial and aquatic
invertebrates are in direct contact with and ingest relatively great quantities of soils and
sediments, they lack an AhR-mediated pathway, and are thus not sensitive to PCDFs and
PCDDS. Many reptiles and amphibians are also present on site, but studies indicate that
they are also not particularly sensitive to the effects of PCDD/DFs (Jung and Walker
1997). Fish have been shown to have the necessary receptor to elicit the toxic effects of
PCDD/DFS, but hepatic concentrations of the receptor appear to be lower in ﬁsh than in
mammals (Hahn 1998). In mammals, toxicity of PCDD/DFs seems to vary signiﬁcantly

among species, but in general, mammals have shown moderate sensitivity to dioxin-like

18

compounds. American mink (Mustela vison), was included as the only mammalian
receptor species in the TR ERA, as previous studies have shown it to be sensitive to the
toxic effects of dioxin—like compounds (Aulerich et al. 1988; Beckett er al. 2008; Heaton
et al. 1995; Hochstein et al. 1988; Tillitt er al. 1996). Laboratory studies have shown
birds, particularly during the embryonic stage, are particularly sensitive to the effects of
dioxin-like compounds (Barron et al. 1995). However, great differences in susceptibility
between species to these toxic effects has been observed, with the domestic chicken being
10— to 100- fold more sensitive to these effects (Brunstrom 1988; Hoffman et al. 1998;
Kennedy et al. 1996). The observed differences in sensitivity to dioxin-like compounds
among taxa may be attributable to varying concentrations of the AhR in certain tissues,
differences in degradation potential, or species differences in the AhR construct and
associated ligand binding afﬁnity (Hahn 1998). Further work has been done to
characterize differences in the AhR between avian species which exhibit different levels
of sensitivity to AhR-active compounds (Head er al. 2008; Karchner et al. 2006).
Brieﬂy, avian species may be broadly categorized to three levels of sensitivity to AhR-
active compounds based on key amino acid residues in the ligand binding domain of the
AhR (Head et al. 2008). An ideal receptor species would have relatively great sensitivity
to the COPECs so that it could be considered protective of other, potentially less

sensitive, species on-site.

Site-Specific Receptor Species
The great blue heron and belted kingﬁsher were selected to investigate the movement

of PCDD/DPS through an aquatic-based exposure pathway and the potential risk these

19

contaminants may pose to wildlife in the Tittabawassee River ﬂoodplain. In addition to
each of these species being piscivorous birds that nest in the Tittabawassee River
ﬂoodplain, each has other species-speciﬁc attributes which make it appropriate for use as
a receptor species in the TR ERA. I

GBH possess many of the characteristics that are desirable in a receptor species, and
as such, are often selected as an ecological receptor of concern in risk assessments. GBH
have a broad distribution across geographic regions and habitat types, residing in
freshwater, estuarine, and marine habitats throughout North America (Butler 1992).
GBH are a colonially-nesting species, with a rookery containing as many as 1300
breeding pairs recorded (Desgranges and Desrosiers 2006). With breeding pairs
concentrated in one area, the colonies are more conspicuous to researchers than single-
nesting species and allows for the assessment of population health rather than the
outcome of a few nesting pairs. There are many closely related species for. which GBH
could serve as a surrogate species or that could be studied utilizing the described
methods. Additionally, GBH are a charismatic species that is widely recognized by the
general public, which would have an interest in preserving this species.

As a long-lived territorial species at the top of the aquatic food web, GBH have the
potential to bioaccumulate local contaminants over a long period of time (Custer et al.
1991). Band recoveries have shown GBH may live to be at least 20 years old (Bayer
1981). GBH are year-round residents in areas of its range where foraging remains
available during winter months, particularly in coastal areas (Butler 1997). The
territorial nature of GBH leads to the active defense of distinct, identiﬁable foraging areas

local to the breeding colony (Marion 1989; Peifer 1979), thus GBH exposure may have

20

better deﬁned spatial boundaries as compared to other more opportunistically feeding
piscivorous birds. Studies tracking adult GBH from rookeries to foraging areas have
determined that adult GBH forage a mean distance between 3.1 km and 6.5 km from
breeding colonies, although a distance as great as 34.1 km has been recorded (Dowd and
Flake 1985; Peifer 1979; Thompson 1978). To directly characterize site-speciﬁc dietary
exposure, the habit of GBH nestlings to regurgitate their stomach contents when under
duress can be exploited by collecting the regurgitant to determine dietary composition
and contaminant concentrations. Previous studies have detected local organochlorine
contaminants in the tissues of GBH (Champoux et al. 2002; Custer et al. 1997; Elliott at
al. 2001; Harris et al. 2003; Straub et al. 2007; Thomas and Anthony 1999). The beneﬁts
and potential drawbacks of using GBH as a receptor species have previously been
outlined (Seston et al. 2009).

Another avian piscivore along the Tittabawassee River, the belted kingﬁsher, has
species-speciﬁc attributes which make it another attractive receptor species for inclusion
in the TR ERA. As a piscivorous species, the KF has a high trophic status and thus has
great exposure potential to bioaccumulative COPECs. Additionally, BKF have a high
rate of food intake, consuming nearly 50% of their body weight in food daily (Davis
1980). BKF forage primarily on the most available ﬁsh species present within its
foraging range, but will also take crayﬁsh, frogs, salamanders, lizards, small snakes, and
insects when water conditions decrease ﬁshing success (Davis 1982; Salyer and Lagler
1949; White 1938). During the breeding season, BKF defend a mean territory size of
1.03 km of river length (Davis 1982). Individuals may maintain their territories year-

round, unless the water freezes over, restricting food resources (Hamas 1994). Due to

21

their limited foraging range in combination with the aggressive territorial behavior of
BKF and the fact that nestlings are restricted to food brought to them by adults,
contamination in nestlings can be related to prey from a deﬁned reach of river proximal
to the nest site.

The BKF has a widespread distribution, inhabiting diverse aquatic habitats
throughout North America (Hamas 1994). Unlike the colonially-nesting GBH, BKF are
solitary nesters, aggressively defending their breeding territory from conspeciﬁcs (Davis
1980). Physical characteristics of an ideal nest site include steep earthen banks with
substrate comprised of approximately 75% sand, with clay and silt, and with little
vegetation or other potential impediments to the excavation of the nest burrow (Brooks
and Davis 1987). BKF construct a subterranean burrow, usually between 1—2 m in
length, that terminates is a nest chamber. Burrows are generally located near the top of
high banks as a means to deter predation and prevent ﬂooding of the nest chamber (Davis
1980). Locating and securing suitable nesting habitat is the most important factor in the
establishment of BKF breeding territories, and is often a limiting factor in overall
abundance (Brooks and Davis 1987; Davis 1982). Where there is a lack of natural
nesting sites but available food resources, BKF have been shown to utilize artiﬁcial nest
sites created through anthropogenic activities, such as gravel pits and road cuts (Hamas
1974). As with the GBH, the widespread distribution and charismatic nature of the BKF

draws the attention of the public to its preservation.

22

Research Objectives

Due to the presence of PCDD/DFS at elevated levels in the Tittabawassee River and
associated ﬂoodplains, there is concern that resident wildlife species may be exposed to
concentrations of these compounds which could potentially affect their health, at either
the individual or population level. To understand both exposure and associated effects,
ﬁeld studies were conducted over several years in support of a site-speciﬁc ecological
risk assessment. Many different receptor species were included in this site-speciﬁc
assessment to most accurately assess the complex Tittabawassee River ﬂoodplain
ecosystem. The focus of the research in the following dissertation is the movement of
PCDFs and PCDDS through an aquatic-based exposure pathway, monitoring population-
health parameters, and assessing overall risk posed by these contaminants to piscivorous
avian species nesting in the Tittabawassee River ﬂoodplain. As top-tier representatives
of an aquatic-based food web, the great blue heron and belted kingﬁsher were selected as
the receptor, species which could best meet this objective. Dietary exposure and tissue-
based exposure, combined with assessments of population health of each receptor
species, are the multiple lines of evidence to be employed in assessing the risk present to

resident species of the Tittabawassee River ﬂoodplain.

Permits, Approvals, and Funding

All aspects of the study that involved the use of animals were conducted in the most
humane means possible. To achieve that objective, all aspects of the study were
performed following standard operation procedures (GBH adult handling 05/07-069-00;

GBH nest monitoring 05/07-066-00; Protocol for belted kingﬁsher monitoring and tissue

23

collection 05/07-071-00; Field studies in support of TR ERA 03/04-042-00; Protocol for
ﬁsh sampling 05/07-059-00) approved by Michigan State University’s Institutional
Animal Care and Use Committee (IACUC). All of the necessary state and federal
approvals and permits (Michigan Department of Natural Resources Scientiﬁc Collection
Permit SC1254 for GBH and BKF/SC permit for ﬁsh (Zwiemik)/SC permit for
amphibians (Zwiemik); USFWS Migratory Bird Scientiﬁc Collection Permit
MBIOOOO62-0; and subperrnitted under US Department of the Interior Federal Banding
Permit 22926) are on ﬁle at Michigan State University—Wildlife Toxicology Laboratory
are on ﬁle at MSU-WTL.

Additionally, James Dastyek and Steven Kahl of the US Fish and Wildlife Service
Shiawassee National Wildlife Refuge granted approval to access to the refuge property,
the Saginaw County Park and Tittabawassee Township Park rangers granted access to
Tittabawassee Township Park and Freeland Festival Park, Tom Lenon and Dick Touvell
of the Chippewa Nature Center granted property access, and Michael Bishop of Alma
College provided oversight as the Master Bander. More than 50 cooperating landowners
throughout the research area granted access to their property, making this research
possible. Funding was provided through an unrestricted grant from The Dow Chemical
Company, Midland, Michigan to JP. Giesy and M.J. Zwiemik of Michigan State

University.

24

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33

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34

CHAPTER 2

Utilizing the great blue heron (Ardea herodias) in ecological risk assessments of

bioaccumulative contaminants

Rita M. Seston‘, Matthew J. Zwiemikz, Timothy B. Fredricksl, Sarah J. Coeﬁeld‘, Dustin

LL. Tazelaarz, David W. Hamman3, John D. Paulson4, and John P. Giesyl’5

lZoology Department, Center for Integrative Toxicology, Michigan State University,

East Lansing, MI 48824, USA

2Animal Science Department, Michigan State University, East Lansing, MI 48824, USA

3College of Veterinary Medicine, Center for Comparative Epidemiology, Michigan State

University, East Lansing, MI 48824, USA

4USDA-APHIS, Wildlife Services, Bismarck, ND 58501 USA

5 . . . . . .
Department of Veterinary Blomedlcal Selences and Tox1cology Centre, UnlverSlty of

Saskatchewan, Saskatoon, Saskatchewan, S7J 5B3, Canada

35

Abstract

Selection of an appropriate species is a key element of effective ecological risk
assessments (ERA), especially when site-speciﬁc ﬁeld studies are to be employed. Great
blue herons (GBH) possess several ideal characteristics of a receptor species for the
assessment of bioaccumulative compounds in the environment, such as case of study,
high potential for exposure, widespread distribution, and territorial foraging behavior.
Methodologies for assessing exposure and population health are described herein. As
outlined, the collection of GBH eggs, GBH nestling blood, and adult GBH blood allows
for the determination of contaminant concentrations in various GBH tissues, a top-down
assessment, which can be done in conjunction with predicted dietary exposure, a bottom-
up assessment, to support a multiple lines of evidence approach. Additionally,
population parameters, such as productivity and survival, can also be measured to
elucidate if the contaminant exposure may be causing population level effects. Over the
course of two years, three GBH rookeries were monitored for productivity and nestling
exposure. Nests were monitored from blinds and individually accessed at multiple time
points to obtain measures of nestling health, band nestlings, and collect eggs and nestling
plasma. Multiple nests could frequently be accessed by climbing one tree, resulting in
minimal effort to obtain the necessary sample size. Additionally, 51 adult GBH, captured
in their foraging areas, were banded, and provided a blood sample. With these samples, a
statistical difference in tissue baSed exposure was identiﬁed between the reference and
target area. Statistically signiﬁcant differences were also identiﬁed between the upper and
lower reaches of the target area, thereby identifying a range of doses geographically

which could be correlated to speciﬁc measurement endpoints. The ability to identify a

36

dose response greatly increases the ability of the dataset to determine causation, a key
goal of such studies. Overall, the use of the described methods allowed for the collection
of a statistically sufﬁcient and ecologically relevant dataset with reasonable effort and

minimal impact on GBH.

Introduction

Selection of appropriate species is a key element to allow effective ecological risk
assessments (ERA), especially when site-speciﬁc ﬁeld studies are to be employed.
Ideally, representative species used in assessments of bioaccumulative compounds should
have an elevated exposure potential, a widespread distribution, and be territorial. Data
collected using the selected species should ultimately provide insight into the health of
the entire ecosystem of the study site. Piscivorous birds are frequently selected as
receptors for evaluating aquatic systems because they can be sensitive to the effects of
contaminants and have the potential to accumulate persistent, lipophylic contaminants
through trophic-transfer. The great blue heron (Ardea herodias; GBH) possesses several
characteristics that make it an appropriate Species to use as a receptor in ERAS
concerning bioaccumulative contaminants in aquatic environments. Here we describe a
multiple lines of evidence approach to elucidating exposure of GBH to contaminants
through the diet and measured concentrations in speciﬁc GBH tissues. As a case study of
the methodology, we have investigated the exposure of GBH to polychlorinated
dibenzofurans (PCDFS) and dibenzo-p-dioxins (PCDDS) in the TittabawasSee River

basin, Michigan, USA.

37

The Tittabawassee River study area includes approximately 37 km of the
Tittabawassee River from the upstream boundary of the city limits of Midland, M1 to the
conﬂuence of the Tittabawassee and Saginaw Rivers downstream of Green Point Island
(Figure 2.1). Just above the upstream boundary of the study area is a low-head dam.
Throughout the study area, the river is free ﬂowing to the conﬂuence with the Saginaw
River and eventually the Saginaw Bay and Lake Huron. The study area was selected
because soils and sediments were found to contain elevated concentrations of PCDFS and

PCDDS. Soils and sediments collected from within the study area contained mean
PCDF/PCDD concentrations ranging from 1.0xlO2 to 5.4x104 pg/g dw, which were 10-

to 20- fold greater than those collected upstream in reference areas (Hilscherova et al.,
2003). The source of this contamination has been identiﬁed as The Dow Chemical
Company (USEPA, 1986)

PCDFS and PCDDS occur in the environment as mixtures and due to their
hydrophobic characteristics and resistance toward metabolism, they have great potential
to be accumulated through the food web. The toxicological response of primary concern
is mediated through the aryl hydrocarbon receptor (AhR) and effects include
carcinogenicity, immunotoxicity, and adverse effects on reproduction, development, and
endocrine functions (van den Berg et al., 1998). In particular, AhR-mediated compounds
have been shown to decrease hatching success and ﬂedging success in aquatic avian
species (Gilbertson, 1983; Hoffman et al., 1987; Ludwig et al., 1993; van den Berg et al.,
1994)

Desirable, species-Speciﬁc characteristics of the GBH led to its inclusion as a receptor

species in an ERA concerning PCDFS and PCDDS along the Tittabawassee River and its

38

   
 

Midland Area

 

\

Lake Huron

 

Reference
trapping reach Midland
‘vci

chivl’ﬂwm

as“

it“ * FR
Upper '
Target trapping
reaches Lower

0 2.5 5 10
E Kilometers

 

(Figure 2.1. Location of great blue heron rookeries (FRE, SNWR, and CAS) and reaches

within the Tittabawassee River study area, MI, USA where trapping occurred.

39

ﬂoodplain. The objective of this paper is to outline a series of methods and the
associated effort necessary to effectively employ the GBH as a receptor in an ERA

utilizing a multiple line of evidence approach.

Species Applicability

Species-speciﬁc attributes need to be considered when selecting a species for use as a
receptor. GBH possess many of the characteristics that are desirable in a receptor
species, and as such, are often selected as an ecological receptor of concern in risk
assessments. GBH have a broad distribution across geographic regions and habitat types,
residing in freshwater, estuarine, and marine habitats throughout North America (Butler,
1992). GBH are a colonially-nesting species, with a rookery containing as many as 1300
breeding pairs recorded (DesGranges and Desrosiers, 2006). With breeding pairs
concentrated in one area, the colonies are more conspicuous to researchers than single—
nesting species and allows for the aSsessment of population health rather than the
outcome of a few nesting pairs. There are many closely related species for which GBH
could serve as a surrogate species or that could be studied utilizing the described
methods. Additionally, GBH are a charismatic species that is widely recognized by the
general public, which would have an interest in preserving this species.

As a long-lived territorial species at the top of the aquatic food web, GBH have the
potential to bioaccumulate local contaminants over a long period of time (Custer et al.,
1991). Band recoveries have shown GBH may live to be at least 20 years old (Bayer,
1981). GBH are year-round residents in areas of its range where foraging remains

available during winter months, particularly in coastal areas (Butler, 1997). The

40

territorial nature of GBH leads to the active defense of distinct, identiﬁable foraging areas
local to the breeding colony (Peifer, 1979; Marion, 1989), thus GBH exposure may have
a greater spatial resolution as compared to other more opportunistically feeding
piscivorous birds. Previous studies have detected local organochlorine contaminants in
the tissues of GBH (Custer et al., 1997; Thomas and Anthony, 1999; Elliott et al., 2001;

Champoux et al., 2002; Harris et al., 2003; Straub et al., 2007).

Methods
Nest monitoring and fresh egg sampling

Great blue heron nests were monitored during the nesting season, which begins in
mid- to late March and runs through mid-July, of 2006 and 2007. Colonies were visited
several times over the breeding season to monitor reproductive success. Visits were
coincident with estimated mean nesting, hatching, chick rearing, and ﬂedgling periods
and separated by a minimum of lwk to .minimize disturbance to breeding pairs.
Calculation of events was based on a 2wk courtship/nesting period, 4wk incubation
period, and a minimum of 8 wk from hatching to ﬂedging (Harris et al., 2003). For the
second year of nest monitoring (2007), a helicopter with a stabilized zoom lens was
employed to determine the number of eggs in each nest. Surveys were conducted from
an altitude of 100 meters to minimize nest disturbance. By ﬂying at this altitude,
incubating GBH were not ﬂushed from nests and were not visually disturbed by the
helicopter. An entire rookery could be surveyed in approximately 15 min, capturing'both
video footage and still images. The contents of the nest could only be determined and

counted for nests which the adults were not incubating at the time of survey, but the

41

number of active nests in the rookery could be determined. Hatching date was estimated
by the presence of eggshells on the ground beneath nests and observing the act of herons
presenting sticks to their mates (Moul et al., 2001), along with hearing the nestlings
calling in nests. Estimates of hatching and ﬂedging success were then made when chicks
were estimated to be 4 and 8 wk of age, respectively. At 4 wk of age the nestlings could
be seen and counted in the nest and at 8 wk‘of age nestlings would perch on branches
proximal to the nest. Observations were conducted from semi-permanent blinds erected
in the nesting colony. A ﬁnal visit to the colony was made each year once the leaves
have fallen from the trees to make a count of the total number of nests.

Nests were located in Eastern cottonwood (Populus deltoides), silver maple (Acer
saccharinum), or white ash (Fraxinus americana) trees at heights ranging between 15-25
m. Viable and nonviable eggs were collected from accessible nests in each nesting
colony. Nest trees were selected based on the safety of access and the potential to reach
multiple nests. Tree climbers accessed nests using tree—climbing spikes. Eggs were
collected with a nylon stocking cup attached to the end of an extendable pole from the
nesting tree or a neighboring tree that was in near enough proximity (Hines and Custer,
1995). A maximum of one viable egg was collected at random from each accessed nest
that contained _>_ 2 eggs. In addition to viable eggs, all eggs which failed to hatch were
collected for analysis of developmental stage and contaminant content. Eggs were
weighed and measured, and then carefully transported to the laboratory in a crush-proof,

water-proof container and kept at 4 °C until processing.

42

Capture and handling of nestling GBH

Blood was collected from nestlings when they were approximately four to ﬁve wk old
based on methods previously described in (Henny and Meeker, 1981). At this age,
nestlings were still limited to movement within the nest yet had sufﬁcient mass to provide
an adequate volume of plasma for residue analyses (McAloney, 1973). Some of the
nestlings handled were older than the target age and proved to be more difﬁcult to
retrieve from and replace into the nest. An extendable pole with a retractable wire hoop
was used to reach nestlings (Ketch-All C0,, 4149 Santa Fe Rd. #2, San Luis Obispo, CA
93401). Individual nestlings were placed in a cloth bag and lowered to the ground from
the nest. A 7-10 cm piece of closed-cell polyethylene foam tubing (7 cm OD X 3 cm ID
Swim Noodle) was used to shield the potentially hazardous beak. The bill of the GBH
was inserted into the tubing and a sock was pulled over both the tubing and the bird’s
head to cover the eyes of the captured bird and to keep the tubing in place. This
combination reduced the chance of injury to personnel by covering the sharp beak,
minimized visual stimulation resulting in a calming of the bird, while allowing it to
breathe freely. Individuals were placed in the cloth bag to determine their body weight
by a spring scale (Model 42500, Pesola AG, Switzerland). Lengths of exposed culmen
and tarsus and masses were determined for each individual nestling. The age of each
nestling was estimated using anlequation relating age and culmen length from Quinney
(1982). Individuals were ﬁtted with USFWS bands on the tarsus and colored leg bands
on the tibia (Simpson and Kelsall, 1978). Color leg bands were made from 49mm high x
66mm wide pieces of 2-ply plastic (1/ 16” Gravoglas 2-plex, matte-ﬁnish; Gravograph-

New Hermes, Inc., Duluth, GA) wrapped around a wooden dowel to have a diameter

43

equal to that of the 7B USFWS leg bands (14mm), as described in Hayes and Barzen
(2006)

Salvage nestlings were collected opportunistically following weather events or as a
result of siblicide. Nestlings were examined for any gross external or internal
abnormalities including liver, kidney, spleen, intestine, and gonad histology. Nestling
stomach contents were analyzed to the lowest taxonomic identiﬁcation possible to aid in
the elucidation of a site-speciﬁc dietary composition. Contaminant concentrations were

determined for liver, adipose, and Skeletal muscle tissues of each individual nestling.

Capture and handling of adult GBH

Adult GBH were captured using modiﬁed foot hold traps set around feeding stations
in predetermined GBH foraging areas. Foot hold traps were modiﬁed in a manner similar
to that described by King et al. (1998). Brieﬂy, the factory coil springs of Victor #3
Softcatch traps (Oneida Victor, Inc., Ltd., Euclid, OH) were replaced with weaker Victor
#125 Softcatch coil springs. This modiﬁcation lessened the initial impact of the padded
jaws but still kept enough tension to hold the trapped bird’s leg in place. The chain
supplied with the trap was replaced with either a 15 cm or 30 cm length of elastic shock-
cord attached with swivels on both ends to allow freedom of movement and minimization
of injury to captured birds. Feeding stations were established in areas of the river with
substrate ranging from sandy-silt to small pebbles, water depth of approximately 15 cm to
46 cm, and where GBH were observed foraging or tracks were present. The stations
were placed in areas with little current to reduce stress on the bait ﬁsh, and free of debris

to reduce the chance of injury to captured GBH. The feeding stations were open-top 46

44

cm L x 30 cm W x 41 cm H cages constructed of 1.25 cm galvanized hardware cloth on a
frame of 0.60 cm hot-roll rod. Each station was ﬁtted with 1.25 cm urethane pipe
insulation along each of the long edges and anchored in the river by a 1.25 m piece of
smooth rebar passed through two hoops on one comer of the cage. This design allowed
the stations to ﬂoat while remaining anchored, accommodating the ﬂuctuating water
levels of the river, and preventing the loss of the bait ﬁsh. The top-edge of the cage was
ﬁtted with 0.95 cm Tygon® tubing to protect the trapped bird from any potentially sharp
edges. The feeding stations were stocked with forage ﬁsh collected from the immediate
area. Fish were collected by seine net or backpack electro-ﬁsher (Smith-Root LR-24,
Smith-Root Inc., Vancouver, WA). Once GBHs were regularly foraging from the
feeding station, approximately 40 modiﬁed traps were placed in a staggered conﬁguration
around the feeding station. The cord of each trap was outﬁtted with a clip, which
attached the traps to a galvanized steel cord secured with stakes around the station.
Loaded traps were set by placing ﬁrmly into the sediment to stabilize the trap, taking care
not to bury the springs, pan, or pin. Feeding stations were monitored by personnel in a
nearby blind (Doghouse blind, Ameristep Inc., Clio, MI) anytime the traps were set.
Optimal blind location Was on the bank opposite the feeding station, if the river could
easily be crossed. This allowed for the largest ﬁeld of view and minimized potential
disturbance of GBH approaching the feeding station. If this was not possible, blinds were
placed at least-30m away from the feeding station in as much cover that still allowed a
clear view of the feeding station and shoreline. Trapping along this river system was
limited to summer months, when the river’s water levels were lower and more stable.

Captured birds were approached with extreme caution as great blue herons are equipped

45

with strong, sharp beaks and are known to be aggressive (Butler, 1997). Personnel
handling the birds were. outﬁtted with appropriate protective gear including helmets ﬁtted
with face shields and thick woven clothing. Captured adult GBH were hooded in the
same manner as nestlings.

Once safely immobilized, individuals were color marked using numbered color leg
bands placed on the tibia. Color leg bands used on the adults were identical to those used
on nestlings with the addition of unique numerical codes of 14 mm numbers spaced 11
mm apart in 3 vertical rows to increase visibility on the banded bird. Color marking was
done to enable identiﬁcation of captured adults from a distance. Measurements of other
physical attributes such as the length of the exposed culmen, wing chord, and tarsus and
mass were also recorded. Each individual was also ﬁtted with a US. Fish & Wildlife

Service band on the tarsus. ,

Blood plasma sampling

Blood from nestling and adult GBH was drawn from the brachialis vein using needles
afﬁxed to sterile syringes pre-rinsed with sodium heparin solution. 22-gauge needles
(Becton Dickinson, Franklin Lakes, NJ) were used for nestlings while smaller 25-gauge
needles were used for adults due to smaller vein size. To determine the maximum
volume of blood that could be collected, the following set of equations were utilized; 7%
body weight = total blood volume, 10% total blood volume = acceptable sample volume.
Bloodcollection was most effectively performed with three people, one to hold'the head,
legs, and body of the GBH still, one keeping the wing outstretched and steady, and one to

perform the blood draw. The blood sample was then transferred to a heparinized

46

VacutainerTM (Becton Dickinson, Franklin Lakes, NJ) for transport back to the ﬁeld

laboratory. Each VacutainerTM was labeled with the band number, trapping station 1D,

GPS coordinates of trapping station, date, and collector’s initials. Whole blood samples
were centrifuged and the plasma (supernatant) was decanted. Both plasma and packed
cell volume were stored at -20 °C until analysis. Red. blood cells were saved for future

sexing of individuals.

Collection of prey items

Site-speciﬁc GBH dietary items, including forage ﬁsh, amphibians, and crayﬁsh,
were collected and analyzed for contaminant concentrations (Alexander, 1977).
Collection of the dietary items occurred at 6 sampling locations, 2 in the reference area
and 4 approximately equally spaced throughout the 27 km target area. The sampling
scheme maximized information on dietary exposure including geographically associated

contaminant variability and trends.

Sample processing and analytical techniques

Collected eggs were opened around the girth with a chemically cleaned scalpel blade
and assessed for stage of development and the presence of any abnormalities. Contents
were then homogenized in a chemically cleaned Omni-mixer, lyophilized, and stored in
clean jars until analysis (I-CHEM brand, Rockwood, TN). Tissues collected from
salvage nestlings were also homogenized using a chemically cleaned Omni-mixer. All
samples were analyzed for concentrations of the seventeen 2,3,7,8-substituted PCDF/D

congeners, in addition to a subset of egg and tissue samples also being analyzed for PCB

47

and DDXs. Analyses were conducted in accordance with EPA Method 8290 with minor
modiﬁcations (USEPA, 1998). In summary, biotic matrices were homogenized with
anhydrous sodium sulfate and Soxhlet extracted for 16 hr using 400 mL toluene. The
extraction solvent was transferred to hexane and the extract was concentrated to 10 mL.

Before extraction known amounts of I3C-labeled PCDF/Ds were added as internal

standards to the sample. Extracts were initially puriﬁed by treatment with concentrated
sulfuric acid. The extract was then passed through a multilayer silica gel column
containing silica gel and sulfuric acid silica gel and eluted with 150 mL of 10%
dichloromethane in hexane. The extract is then passed through a carbon column packed
with l g of activated carbon-impregnated Silica gel. The ﬁrst fraction, eluted with 100 ml
hexane, was kept for PCB analysis. The second fraction, eluted with 200 mL of toluene,
contained the 2,3,7,8-substituted PCDF/D5. PCDF/D5 were analyzed using HRGC-
HRMS, a Hewlett-Packard 6890 GC (Agilent Technologies, Wilmington, DE) connected
to a MicroMass high resolution mass spectrometer (Waters Corporation, Milford, MA).
PCDF and PCDD congeners were separated on a DB-S capillary column (Agilent
Technologies, Wilmington, DE) coated at 0.25 pm (60 m x 0.25 mm i.d.). Generally, the
mass spectrometer was operated at an El energy of 60 eV and an ion current of 600 uA.
PCDD/DP congeners were monitored by single ion monitoring (SIM) at the two most
intensive ions at the molecular ion cluster. Concentrations of certain PCDF/D congeners,
particularly TCDD and TCDF congeners were conﬁrmed by using a DB-17 (60 m x 0.25
mm id, 0.25 pm ﬁlm thickness) column (Agilent Technologies, Wilmington, DE).

Chemical analyses included pertinent quality assurance practices, including surrogate

48

Spikes, blanks, and duplicates. Soxhlet extractions and chemical analyses were

conducted at AsureQuality Limited, Lower Hutt, New Zealand.

Results
Rookery

Three active GBH rookeries were located within the study area. The Freeland
rookery (FRE) was established in 2001, and contained 44 nests in 2006 and 46 nests in
2007. The Shiawassee National Wildlife Refuge rookery (SNWR), established in 1999,
contained 161 nests in 2006, but nest occupancy has drastically decreased after the recent
establishment of predatory avian species, including bald eagles (Haliaeetus
leucocephalus), great horned owls (Bubo virginianus), and red-tailed hawks (Buteo
jamaicensis), within the rookery. A second rookery located on the Shiawassee National
Wildlife Refuge near the conﬂuence of the Cass and Shiawassee Rivers (CAS), contains
approximately 35 nests. CAS was established in 1989 but has only been occupied
periodically. No rookeries were located within the reference area. From each rookery, a
target sample size of eight fresh eggs was collected, each from separate nests in the
rookery. This collection of tissue from 24 different nests required climbing only 7
different trees, with up to 4 nests being accessible in one tree. Fresh egg sampling oﬁen
. involved incubation disturbance within the rookery so it was only done when
temperatures were greater than 15 °C. Time; required for egg sampling was on average
0.66 h/egg or 2.25 h/tree, with sampling efﬁciency increasing as climbers became more
experienced with the technique. Nestling banding and blood plasma collection occurred

at all rookeries. At FRE, 12 nests were accessed for nestling banding and blood plasma

49

collection, for a total of 20 nestling blood plasma samples. Four nests at SNWR
produced 8 nestling blood plasma samples. Fourteen nestling blood samples were
collected from 7 separate nests at CAS. Time required for nestling blood plasma
collection averaged 0.90 h/sample or 1.63 h/nest. Two salvage nestlings were collected
from 1 nest at FRE, 7 salvage nestlings were collected from 3 separate nests at SNWR,
and 2 salvage nestlings from 2 separate nests at CAS. A summary of collected tissues is

outlined in Table 2.1.

Adult trapping

Twelve GBH trapping stations were established, 3 located in the reference area and 9
in the target study area. By employing the described methods, there were 62 capture
events, which included the capture of 51 GBH, 9 recaptures, and 2 escapes. All GBH
recaptures occurred at their original trapping station, with one exception where the
trapping stations were less than 500 m apart. Once recaptures occurred at any given
feeding station, that station was moved to a new foraging territory to target new GBH.
One GBH was recaptured two consecutive years at the same feeding station. As many as
5 individuals were trapped at one feeding station over the course of a ﬁeld season. Of the
51 GBH captured, 15 were in the reference area and 36 in the target area (Table 2.1). On
average, a GBH adult plasma sample was obtained for every 15.26 h of active trapping,
which was conducted by a two-person team. Including recaptures and escapes, on
average one GBH was captured for every 12.56 h of active trapping. These ﬁgures do not
include time spent maintaining the stock of ﬁsh in the feeding stations. Normalized to

the number of trapping hours, the most successful time of day to conduct trapping was

50

Table 2.1. Description of sampling effort and summary of great blue heron
tissues collected through utilizing described methodologies. Note that there was
no rookery located in the reference area to sample. Necessary adult plasma
sample size in reference area obtained aﬁer 2006 ﬁeld season.

 

 

 

 

 

Year Reference Target
Adult Adult Egg Nestling Nestling
Plasma Plasma Collection * Plasma * Tissue *
2005 5 15
2006 10 9 6 (6) l3 (8) 9 (4)
2007 12 19 (18) 29 (15) 2 (2)
Total 15 36 25 (24) 42 (23) 11 (6)

 

* values in parentheses indicate the total number of nests samples collected from

(n)

51

from 0600 to 1000 (Figure 2.2), with a trapping success rate of approximately 0.12
GBH/h. Average time from capture to release of GBH was 60 min. Injuries associated
with adult trapping and handling were low and no injuries were sustained to adult GBH
that would be expected to impact survival. A damaged or torn leg scale was noted for 8

of the 62 birds trapped and a broken phalange (non-hallux) was noted for a single bird.

Sampling ejfectivity

Power analyses were conducted using total TEQS (pg/mL) in the blood plasma of
sampled adults from the 2005 ﬁeld season (Table 2.2). These analyses revealed that a
signiﬁcant difference could be discerned between the reference and target area with a
Type I (01) error rate of 0.05 and a Type II ([3) error rate of 0.20 with as few as 4 samples
from each area. Additionally, the target area could be divided into an upper and lower
reach with signiﬁcant differences discemable at the same errors rates with 14 samples
collected from each area. Analytical results from the other matrices were not available to

run site-speciﬁc power analyses.

52

.230: @3935 m0 598:: of 8 BERES: EmO mo 536:2 .voﬁoE waﬁmmb 22-88 28

228% :3 2: maﬁa boom 93 moom 52,309 8.8 madam 53M commmamnmﬁmh c5 5 vegan: macho: 0:3 38w mo 538:2 .N.N omawmm

 

 

25H
oonom coH a Gong oar: OOHN— cone conwo oonoo

_ . .l- p b . 1 O0.0
- 8o 9
8
H
. cod 1
.m
.w
. 23 .w
m
w.
. Nﬁo m

t 2.0

mﬁo

 

 

 

 

 

 

53

Table 2.2. Total TEQWHO-Avian (pg/mL) in adult GBH blood plasma from

Tittabawassee River study area collected during 2005 ﬁeld season.

 

 

Mean St. Dev. Min Max
Reference 1.8 0.75 1.2 3.0
(n=5)
Upper Target 8.4 5.5 2.2 17
(n=10)
Lower Target 12 5.5 8.5 20
(11:5)

 

54

Discussion

Selection of a receptor species is a key element to an effective ERA, especially when
site-speciﬁc ﬁeld studies are to be employed. As a long-lived species near the top of the
aquatic food web, GBH have the potential to be highly exposed to contaminants for many
years. Residing in freshwater, estuarine, and marine habitats throughout North America,
GBH have the potential to be utilized in ERAS in many different locales. The territorial
foraging behavior of GBH leads to the active defense of distinct and identiﬁable foraging
areas. Additionally, GBH have many closely related species which share some of these
desirable attributes and could be studied using these same methods. All of these
characteristics make the GBH a model receptor species for the assessment of
bioaccumulative compounds in an aquatic food web, and led to their inclusion in the
ERA conducted for the Tittabawassee River ﬂoodplain.

Although all of the aforementioned characteristics are important to have in a receptor
species, they become irrelevant if the species is too difﬁcult to study and acquire the
necessary samples. Since GBH are a colonial nesting species, the discovery of one
rookery results in the location of tens to hundreds of breeding pairs. This allows a
multitude of nests to be monitored simultaneously for reproductive success and nestling
dietary composition, and an assessment of population health rather than the health of a
few individuals. In the studied rookeries, multiple nests could oﬁen be sampled for eggs
or nestling handling by climbing one tree, reducing the effort needed to obtain the
necessary sample size. In addition to comparing population health parameters from study

areas to appropriate reference areas, comparisons may also be made with other studies

55

which report productivity that are available in the literature (Pratt, 1970; Thomas and
Anthony, 1999; Harris et al., 2003; Witt, 2006). Conversely, GBH characteristics of
nesting at great heights in small diameter and sometimes dead trees can make both
ground-based observations and physical nest access challenging. For the rookeries
monitored here, ground blind observations were supplemented with observations from
rotary wing aircraft to assess clutch size. However physical nest access was limited to
trees that were safely climbable. Date, time of ﬂight, and nonrandom limitations to
physical nest access may add bias to measurements and should be noted for within and
across study comparisons.

The collection of nestling blood plasma and eggs from the same nest allows for the
possible derivation of a plasma-to-egg ratio, which would eliminate the need for
destructive egg sampling. GBH are a migratory species, so it is possible that
contaminants transferred to the egg from the female were accumulated elsewhere (Henny,
1986); however, eggs collected from rookeries in the study area exhibited low variation
in total TEQS and congener proﬁles within and among rookeries, suggesting this was not
an important factor at this site. Nestlings are considered to be more representative of
local contamination as they are conﬁned to the nest and rely on the food brought to them
by adults (Olsson et’ al., 2000; Neigh et al., 2006). Studies tracking adult GBH from
rookeries to foraging areas have determined that adult GBH forage a mean distance
between 3.1 km and 6.5 km from breeding colonies, although a distance as great as 34.1
km has been recorded (Thompson, 1978; Peifer, 1979; Dowd and Flake, 1985). To
further characterize site-speciﬁc dietary exposure, the habit of GBH nestlings .to

regurgitate their stomach contents when under duress can be exploited by collecting the

56

regurgitant to determine dietary composition and contaminant concentrations.
Additionally, the described methods facilitate the collection of plasma samples from
multiple nestlings within a single nest to determine intra-brood variation. This dataset
combined with eggs from the same nest can generate plasma-to-egg ratios, which can be
a useful tool; however, they must only be used when both eggs and plasma are
representative of local contamination. These ratios are especially desirable when dealing
with endangered or threatened species when avoiding any destructive sampling is of great
importance (Strause, et al., 2007).

The territorial foraging of GBH facilitates the establishment of feeding stations in
multiple foraging territories in the area of interest to capture different individuals with a
low rate of recapture. Throughout the three ﬁeld seasons during which trapping of adult
GBH was performed, the feeding Station and foot-hold trap method proved to be very
effective, as demonstrated by the ﬁfty different individuals that were captured, banded,
and provided a blood plasma sample. As many as ﬁve individuals were trapped at one
feeding station over the course of one ﬁeld season. Along a river system in South
Dakota, Dowd and Flake (1985) determined that radio-tagged GBH would return to the
same general areas of the river, but other GBH were also observed using the same areas.
Additionally, ﬂedgling GBH did not seem to display aggressive territorialism over
foraging areas and were often seen foraging in ﬂocks. In the closely related grey heron
(Ardea cinerea), other foragers would visit actively defended territories in the absence of
the territory owner (Marion, 1989). One individual GBH was recaptured two consecutive
years at the same feeding station, suggesting territories may be maintained over multiple

years. No banded birds were recaptured at feeding stations other than where they were

57

initially trapped, except where the feeding stations were less than 500 m apart, again
reinforcing the territoriality 0f GBH foraging. Additionally, the recapture of GBH,
sometimes multiple times in the same day, suggests that either this trapping method was
not traumatic or injurious as compared to the desire for easy prey. The level of effort
involved in adult trapping could potentially be lessened by focusing trapping effort
during certain times of the day, during the nesting season, or in areas of group foraging.
When normalized to the number of trapping hours, the period from 0600 to 1000 had the
highest trapping success rate, at approximately 0.12 GBH/hr. In the present study,
trapping was focused in solitary feeding areas in an attempt to quantify site ﬁdelity and to
minimize the effects of foraging disturbance on additional birds.

The temporal consistency of both data access and exposure potential for GBH adds
potential ﬂexibility in study design and sampling efforts. For instance, we used power
analysis of ﬁrst year data to identify spatially explicit boundaries for which statistically
signiﬁcant differences could potentially be identiﬁed at a reasonable level of effort. This
adds value to the study by providing for a real-time cost beneﬁt analysis and the most
efﬁcient allocation of resources.

The methodologies employed in this study provided multiple ways of estimating
exposure, including dietary exposure and tissue-based exposure assessments. Although
an exact site-speciﬁc dietary composition was not calculated, a combination of literature-
based diets and observations of Site-speciﬁc GBH foraging led to the collection of forage
ﬁsh, crayﬁsh, and amphibians as the primary diet components. Analysis of these items

allowed for the calculation of estimated average daily intake and resulting HQS for

dietary exposure. Determination of concentrations of PCDDS/F5 in egg, nestling tissues,

58

and nestling and adult plasma allowed for the determination of HQS based on tissue
concentrations. Comparisons can then be made between the HQs derived from the
varying approaches to determine the accuracy of predicting exposure through the diet and

the importance of collecting receptor tissues.

Acknowledgements

The authors would like to thank all staff and students of the Michigan State
University -— Wildlife Toxicology Laboratory ﬁeld crew; namely Michael W. Nadeau,
Will R. Folland and Stephanie C. Plautz for their tree-climbing abilities, along with
Emily M. Koppel, Jeremy N. Moore, Bretton J. Joldersma, Megan L. Barker, Joost van
Dam, Lori E. Williams, and Casey L. Bartrem. We gratefully acknowledge Mikes Fales
for his design and fabrication of specialized equipment which was pivotal to the success
of this research. Additionally, we would like to recognize Patrick W. Bradley, Michael J.
Kramer, and Nozomi Ikeda for their assistance in the laboratory, James Dastyek and
Steven Kahl of the United States Fish and Wildlife Service —- Shiawassee National
Wildlife Refuge, for their assistance and access to the refuge, the Saginaw County Parks
and Recreation Commission for access to Imerma'n Park, and the Tittabawassee
Township Park Rangers for access to Tittabawassee Township Park and Freeland Festival
Park. We would also like to acknowledge the greater than 50 cooperating landowners
throughout the study area who have granted us access to their property, making our
research possible. Funding was provided through an unrestricted grant from The Dow
Chemical Company, Midland, Michigan to JP. Giesy and M. J. Zwiemik of Michigan

State University.

59

Animal Use

All aspects of the study that involved the use of animals were conducted in the most
humane means possible. To achieve that objective, all aspects of the study were
performed following standard operation procedures (GBH adult handling 03/05-036-00;
GBH nest monitoring 05/07-066-00; Field studies in support of TR ERA 03/04-042-00;
Protocol for ﬁsh sampling 03/04-043-00) approved by Michigan State University’s
Institutional Animal Care and Use Committee (IACUC). All of the necessary state and
federal approvals and permits (Michigan Department of Natural Resources Scientiﬁc
Collection Permit SC1254/SC permit for ﬁsh (Zwiemik)/SC permit for amphibians
(Zwiemik) and USFWS Migratory Bird Scientiﬁc Collection Permit MBIOOOO62-0) are

on ﬁle at MSU-WTL.

6O

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64

CHAPTER 3

Dietary exposure of great blue heron (Ardea herodias) to PCDD/DFs in the

Tittabawassee River ﬂoodplain, MI, USA

Rita M. Seston], Timothy B. Fredricksl, Dustin L. Tazelaarz, Sarah J. Coeﬁeldl, Patrick
W. Bradleyz, Shaun A. Roark3, John L. Newsted3, Denise P. Kay3, Matthew J.

Zwiemikz, and John P. Giesyl’4’5

lZoology Department, Center for Integrative Toxicology, Michigan State University,

East Lansing, MI 48824, USA

2Animal Science Department, Michigan State University, East Lansing, MI 48824, USA

3ENTRIX, Inc., Okemos, MI, USA

4 . . . . . . .
Department of Veterinary Biomedical Selences and Tox1cology Centre, Umversrty of

Saskatchewan, Saskatoon, Saskatchewan, S7J SB3, Canada

5 . ‘
Department of Biology and Chemistry, City Universrty of Hong Kong, Hong Kong,

SAR, China

65

Abstract
Concentrations of dioxin-like compounds, primarily polychlorinated dibenzofurans
(PCDFS), in soils and sediments of the Tittabawassee River (TR) and associated
ﬂoodplains downstream of Midland, Michigan (USA) were greater than upstream sites
and prompted a site-specific risk assessment of great blue herons (GBH). Dietary
exposure of GBH to PCDFS and polychlorinated dibenzo-p-dioxins (PCDDS) was
evaluated based on site-speciﬁc concentrations of residues in prey items. Concentrations

of ZPCDD/DFS and 2,3,7,8-tetrachlorodibenzo—p-dioxin equivalents (TEQWH0.AV,-an) in

prey items collected from the TR were consistently greater than those collected from
associated reference areas (RAs) and further downstream in the Saginaw River (SR). The
average daily dose (ADme) of ZPCDD/DFS to GBH was 45- to 54-fold greater along the
TR and 12-fold greater along the SR when compared to the RA. ZPCDD/DFS were
normalized to TEQWHOAW, and fold differences in the ADme increased, being 150- to
190-fold greater along the TR and 36-fold greater along the SR than they were in the RA.
Greater fold changes in the ADme based on TEQWH()_AV,,,,, between the RA and the TR
and SR was due to prey items from the latter reaches having a greater relative toxic
potency of EPCDD/DFs, primarily from greater amounts of 2,3,7,8-
tetrachlorodibenzofuran but also 2,3,4,7,8-pentachlorodibenzofuran. Potential for
adverse population-level effects from site-specific contaminant exposures were evaluated

via comparison to selected toxicity reference values. The prediction of minimal to no

risk of adverse population-level effects resultant from the assessment of - site-speciﬁc

dietary exposure of GBH to ZPCDD/DFs along the TR and SR is consistent with site-

speciﬂc assessments of tissue-based exposures as well as population condition.

66

Introduction
A screening-level risk assessment pertaining to dioxin-like compounds in the
Tittabawassee River (TR) ﬂoodplain predicted the great blue heron (Ardea herodias;
GBH) to have one of the greatest potential exposures to dioxin-like compounds of avian
species nesting along the TR (Galbraith Environmental Sciences LLC. 2003). That
preliminary assessment was based on minimal information and had to make a number of
assumptions. In the study described herein, additional data was collected that alloWed a
refined estimate of dietary exposure of GBH to polychlorinated dibenzo-p-dioxins and
dibenzofurans (PCDD/DFS) along the TR. Measured site-speciﬁc dietary exposure was
then compared to toxicity reference values (TRVs) to evaluate the potential for adverse
effects of PCDD/DPS on GBH that forage within the TR ﬂoodplain.
As a result of historical chemical production and management of associated
wastes, the Tittabawassee and Saginaw Rivers downstream of Midland, MI, USA contain
elevated concentrations of PCDD/DPS (USEPA 1986). Sediments and associated

ﬂoodplain soils collected from downstream study areas (SA) were found to contain total

concentrations of the 2,3,7,8-substituted PCDD/DFs (XPCDD/DFs) ranging from 1.0x102

to 5.4x104 ng/kg dw. Mean concentrations of XPCDD/PCDF in sediments and ﬂoodplain
soils in the reference area (RA) upstream of Midland were 10— to 20-fold less
(Hilscherova et al. 2003). The persistence of these compounds in the environment,
combined with their toxicity and potential to bioaccumulate, has led to concern over the

potential exposures of wildlife species foraging in the Tittabawassee and Saginaw River

ﬂoodplains.

67

PCDD/DFs often occur in the environment as complex mixtures, and may also be
in the presence of structurally related polychlorinated biphenyls (PCBs). Due to the
lipophilic characteristics of these compounds and their resistance to degradation in the
environment (Mandal 2005) they have the potential to be accumulated through the food

web. The critical responses of these compounds are mediated through the aryl
hydrocarbon receptor (AhR) and include enzyme induction, teratogenicity,
immunotoxicity, and adverse effects on reproduction, development, and endocrine
functions (Allred and Strange 1977; Brunstrom and Andersson 1988; Powell et al. 1996a;
Verrett 1976). In particular, AhR-mediated compounds have been shown to decrease
hatching success and ﬂedging success in aquatic avian species (Gilbertson 1983;
Hoffman et al. 1987; van den Berg et al. 1994a). The sensitivities of a number of species
to AhR-mediated effects have been determined in laboratory studies or inferred from
observations of populations exposed in the wild. Sensitivities have been shown to vary
among species (Brunstrom 1988). For example, the domestic chicken (Gallus gallus),
which is considered to be the most sensitive avian species, is over 1000-times more
sensitive to embryo-lethal effects than is the mallard (Anas platyrhynchos) (Head et al.
2008). Recent research has suggested the ligand binding domain (LBD) of the AhR in
avian species is highly conserved and can be classiﬁed by a few amino acid sequences.
The speciﬁc conﬁguration of the LED directly affects the binding afﬁnity of ligand
(dioxin-like compounds), thereby inﬂuencing the organisms response and sensitivity to
toxic effects. (Head et al. 2008', Karchner et al. 2006). This conﬁrms the importance of

species-speciﬁc exposure and responses among taxa.

68

The primary objective of this study was to describe the dietary exposure of GBH to
PCDD/DFs in the Tittabawassee River basin and predict the risk this exposure poses to
GBH breeding on-site. To characterize PCDD/DF exposure, concentrations of

ZPCDD/DF, EPCB, and 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) equivalents

(TEQWHO-AVian) based on World Health Organization 2,3,7,8-TCDD equivalency factors

for birds (TEFWH0-AVim,) (van den Berg et al. 1998) were measured in dietary items

collected from reference and study areas. Concentrations of PCDD/DPS and the patterns
of their relative congener concentrations in dietary items were evaluated for spatial

trends. To estimate the potential for adverse effects of the measured dietary exposure,
predicted average daily doses (ADDpot) were compared to TRVs. To facilitate a multiple

lines of evidence approach (F airbrother 2003) to determine the risk present to GBH along
the Tittabawassee River ﬂoodplain the results of the dietary exposure assessment were
compared to previously conducted assessments of GBH tissue exposure and population
condition (Seston et al. 2010a). Integration of multiple lines of evidence can help reduce

the uncertainty inherent in the risk assessment process (Leonards et al. 2008).

, Methods
Site description

The assessment was conducted in the vicinity of the city of Midland, located in the
east-central lower peninsula of Michigan (USA). The Tittabawassee River (TR) is a
tributary of the Saginaw River (SR), which eventually empties into Saginaw Bay of Lake
Huron. The TR runs through The Dow Chemical Company (DOW) property, which is

located on the southern edge of Midland and is the accepted source of the PCDD/DP

69

contamination (USEPA 1986). The area henceforth referred to as the study area (SA)
included approximately 37 km of the TR (sites T-3 to T—6) and associated wetlands from
DOW to the conﬂuence of the TR and SR. Additionally, the SA included 35 km of the
SR (sites S-7 to S-9) downstream of the conﬂuence to where it enters Saginaw Bay
(Figure 1). Sampling sites located in the SA were chosen to characterize maximal
exposure potential designated as potential “worst case scenario” locations based on a
previous study that measured soil and sediment concentrations (Hilscherova et al. 2003)
and availability of landowner access to sites. The reference area (RA) was composed of
the TR upstream of Midland, together with the Pine and Chippewa Rivers, both of which
are tributaries of the TR upstream of Midland. Sampling locations in the RA were on the
upstream TR (R-1) and on the Pine River, just upstream of its conﬂuence with the
Chippewa River (R—2). Distinct sampling areas were assessed individually as well as
grouped spatially based on river characteristics. Spatial groupings included reference
(RA) R-1 and R-2, upper Tittabawassee River (UTR) T-3 and T-4, lower Tittabawassee

River (LTR) T-5 to S-7, and Saginaw River (SR) S-8 and S-9.

Receptor species selection

Selection of an appropriate study species is a key element of effective ecological risk
assessments (ERAS), especially when site-speciﬁc 'ﬁeld studies are to be employed
(USEPA 1994). Piscivorous birds are frequently selected as receptors for evaluating
aquatic systems because they can be sensitive to the effects of contaminants and have the
potential to accumulate persistent, lipophilic contaminants through trophic-transfer. The

GBH possesses several characteristics that make it a suitable species to use as a receptor

70

Figure 3.1. Sampling locations for dietary components along the Chippewa,
Tittabawassee, and Saginaw river ﬂoodplains, Michigan, USA. Reference area (R-1 and
R-2; RA); Upper Tittabawassee River (T-3 and T-4; UTR); Lower Tittabawassee River

(T-S, T-6, and S-7; LTR); and Saginaw River (S—8 and S-9; SR).

        

 

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72

. in ERAS concerning bioaccumulative contaminants in aquatic environments (Seston et al.
2009). As a long-lived, territorial, apex predator in the aquatic food web, the GBH has
great potential to accumulate local contaminants over a long period of time, especially in
areas of its range where foraging is available over the winter months, allowing them to be
year-round residents (Butler 1997; Custer et al. 1991). GBH have a broad distribution
across geographic regions and habitat types, residing in freshwater, estuarine, and marine
habitats throughout North America (Butler 1992), which makes them a potential receptor
for sites of aquatic-based contamination at many different localities (Champoux et a1.
2002; Custer et al. 1997; Elliott et al. 2001; Harris et al. 2003; Straub et al. 2007;
Thomas and Anthony 1999). The territorial nature of GBH leads to the active defense of
distinct, identiﬁable foraging areas within 3.2 km to 6.5 km of the breeding colony
(Marion 1989; Peifer 1979). Therefore, the spatial boundaries of exposure of GBH to
persistent and bioaccumulative compounds can be better deﬁned than may be possible

with other more opportunistically feeding piscivorous birds.

Collection of prey items

Prey items of GBH, including forage ﬁsh, crayﬁsh, and amphibians, were collected
and concentrations of PCDD/DFs (n=188) and PCBs (n=20) determined. Dietary items
were collected from nine sampling locations (Figure 3.1). The sampling scheme
maximized information on dietary exposure including geographically associated
contaminant variability and trends, Forage ﬁsh of a size class consumed by GBH
(S25 cm in length) were collected by either electro-ﬁshing or seine netting. Three

species of amphibian common to the area, the wood frog (Rana sylvatica), leopard frog

73

(Rana pipiens), and green frog (Rana clamitans) were collected. Individual frogs were
captured by hand or dip net. Crayﬁsh were collected with a seine net or modiﬁed
minnow traps. Forage ﬁsh were analyzed as composite samples, whereas individual
amphibians and crayﬁsh were analyzed separately unless they needed to be combined to

obtain sufficient sample mass.

Dietary exposure calculations

Exposure of GBH to PCDD/DP via the diet was estimated by use of information
provided in the US. Environmental Protection Agency (U SEPA) Wildlife Exposure
Factors Handbook (WEF H; USEPA 1993). Major factors inﬂuencing dietary exposure
included body mass (BW), daily food intake rate [FIR; g wet wt food/g body weight
(BW)/d], dietary concentrations (C), and proportion of foraging time spent on-site (AUF).
A mean body mass of 2.3 kg was determined by (Henning et al. 1999) after an extensive
review of data available in the literature. Using the equation developed to determine FIR

for wading birds (Kushlan 1978) the USEPA WEFH reports a body-weight normalized
FIR of 0.18 kg food (wet wt) 'kg bw'I 'd'1 for GBH. The ADDpot, expressed as ng/kg
bw/d was calculated using equation 4-3 from the WEF H (USEPA 1993). Incidental
sediment ingestion was also included in the ADDpot using equation 4-23 (USEPA 1993).

GBH have a relatively large foraging range, which can result in site-nesting GBH
spending a portion of their time foraging outside of the SA (Dowd and Flake 1985; Peifer
1979; Thompson 1978). In addition, GBH are a migratory species in areas of its range

where foraging is not available during winter months due to ice cover (Butler 1992). To

examine the effects of fOraging range and site-use, ADme was calculated by use of three

74

different area use factors, including 25%, 75%, 100% on-site foraging. The off-site
portion of the diet was estimated based on contaminant concentrations in prey collected
from the RA. Contaminant concentrations in RA prey items were signiﬁcantly less than
those in prey items collected from the study area and were assumed to be representative
of non-point source exposures.

The relative proportion of each type of prey to the GBH diet was determined through
a combination of site-speciﬁc observations and data reported in the literature.
Observations of GBH foraging, in combination with prey remains and stomach contents
revealed the site-speciﬁc diet to contain primarily ﬁsh in addition to amphibians and
crayﬁsh. However, a small sample size precluded the determination of the relative
contribution of each prey item taxa to the dietary composition. A previous study of GBH
in Michigan reported the diet to be composed of 94 to 98% ﬁsh (Alexander 1977). Thus,
the diet selected here to investigate dietary exposure comprised 96% forage ﬁsh, 2%
amphibians, and 2% crayﬁsh. Reach-speciﬁc concentrations of ZPCDD/DFs in prey
items were multiplied by their relative contribution to the dietary composition. Exposure
was estimated using the geometric mean and associated 95% conﬁdence interval of
concentrations of residues in prey items from each reach (RA, UTR, LTR, and SR)

(Table 3.2). PCBs were not measured in all frogs or crayﬁsh, thus it was not possible to
directly calculate total TEQWHO-Avian exposure associated with PCDD/DFs and PCBs for
all samples. Where both PCDD/DF and PCB data was available, concentrations of total
TEQSWHO-Avian in frogs and crayﬁsh were less than proximally collected ﬁsh. To
estimate a conservative exposure of GBH to total TEQSwHo-Avian a dietary Composition

of 100% ﬁsh was utilized.

75

Sample processing and analytical techniques

Concentrations of seventeen 2,3,7,8-substituted PCDD/DF congeners were measured
in all samples while concentrations of the twelve dioxin-like PCBs were determined in a
subset of samples. Tissues were homogenized in a chemically cleaned Omni-mixer and
stored in clean jars until analysis (I-CHEM brand, Rockwood, TN). PCDD/DPS and
PCBs were quantiﬁed in accordance with EPA Method 8290/1668A with minor
modiﬁcations (USEPA 1998). In summary, biotic matrices were homogenized with
anhydrous sodium sulfate and Soxhlet extracted for 16 hr with toluene. The extraction

solvent was transferred to hexane and the extract was concentrated to 10 mL. Before
extraction known amounts of l3C-labeled analytes were added as internal standards to the

sample. Extracts were initially puriﬁed by treatment with concentrated sulfuric acid.
The extract was then passed through a multilayer silica gel column containing silica gel
and sulfuric acid silica gel and eluted with 10% dichloromethane in hexane. The extract
was then passed through a carbon column packed with activated carbon-impregnated
silica gel. The ﬁrst fraction, eluted with hexane, was kept for PCB analysis. The second
fraction, eluted with toluene, contained the 2,3,7,8-substituted PCDD/DFs. PCDD/DFs
were analyzed via HRGC-HRMS, using a Hewlett-Packard 6890 GC (Agilent
Technologies, Wilmington, DE) connected to a MicroMass® high resolution mass
spectrometer (Waters Corporation, Milford, MA). PCDF, PCDD, PCB, and DDX
congeners were separated on a DB—S capillary column (Agilent Technologies,

Wilmington, DE) coated at 0.25 pm (60 m x 0.25 mm i.d.). The mass spectrometer was

operated at an El energy of 60 eV and an ion current of 600 uA. PCDD/DF congeners

76

were monitored by single ion monitoring (SIM) at the two most intensive ions of the
molecular ion cluster. Concentrations of certain PCDD/DF congeners, particularly
TCDD and TCDF congeners were conﬁrmed by using a DB-l7 (60 m x 0.25 mm id,

0.25 pm ﬁlm thickness) column (Agilent Technologies, Wilmington, DE). Losses of
congeners during extraction were corrected based on recoveries of l3C-labeled as

outlined in EPA Method 8290/1668A. Quality control samples generated during
chemical analyses included laboratory method blanks, sample processing blanks
(equipment rinsate and atmospheric), matrix spike and matrix spike duplicate pairs,
unspiked sample replicates, and blind check samples. Results of method and ﬁeld blank
analyses indicated no systematic laboratory contamination issues. Evaluation of the
percent recovery and relative percent difference data for the matrix spike and matrix
spike duplicate samples and unspiked replicate samples were within i30% at a rate of
greater than 95% acceptability. Soxhlet extractions and instrumental analyses were

conducted at AsureQuality Ltd, Lower Hutt, New Zealand.

Statistical analyses

Residues in dietary items were reported several different ways. Total concentrations
of the seventeen 2,3,7,8-substituted PCDD/DF congeners are reported as the sum of all
congeners (ZPCDD/DFS) as well as TEQWHO-Avian (ng/kg wet weight (wet wt)).
Individual congeners that were less than the limit of quantiﬁcation were assigned a value
of half the sample method detection limit on a per sample basis. Total concentrations of
twelve mm and mono-ortho-substituted PCB congeners are reported as the sum of these

congeners (ng/kg wet wt) (ZPCBs) for a subset of samples. Concentrations of TEQWH0_

77

Avian (ng/kg wet wt) were calculated for both PCDD/DFs and PCBs by summing the
product of the concentration of each congener, multiplied by its avian TEFwHo-Avian (van
den Berg et al. 1998). The term DF-TEQWHO_AV;a,, refers to summation of the TEQS of
individual PCDD and PCDF congeners while the term PCB-TEQWH0-A,.;an refers to the
TEQ from PCBs. Total TEQSWHO-AVian refers to the summation of both DF-TEQSWHO-
Avian and PCB-TEQSWH0.Avian.

Statistical analyses were performed using SAS® software (Release 9.1; SAS Institute

Inc., Cary, NC, USA). Prior to the use of parametric statistical procedures, normality was
evaluated using the Shapiro-Wilks test and the assumption of homogeneity of variance
was evaluated using Levene’s test. Values that were not normally distributed were
transformed using the natural log (ln) before statistical analyses. PROC GLM was used
to make comparisons for three or more locations. When signiﬁcant differences among
locations were indicated, the Tukey-Kramer test was used to make comparisons between
individual locations. PROC TTEST was used to make comparisons between two groups.

Differences were considered to be statistically signiﬁcant at p<0.05.

Selection of toxicity reference values

Literature-based no observed adverse effect concentrations (NOAECs) and lowest
observed adverse effect concentrations (LOAECs) were used for calculation of hazard
quotients (HOS) and subsequent assessment of risk. In this study, dietary exposure-based
TRVs based on the same or similar compounds were identiﬁed from the literature and

compared to predicted site-speciﬁc dietary exposure of GBH. Resulting HQs are

78

presented as a range bounded by the LOAEC associated HQs at the low end and the
NOAEC associated HQs at the high end. It should be noted that the NOAEC and
LOAEC associated HQs are a function of the experimental design (dosing regime) and
the actual threshold concentration at which effects would be expected to occur lies
somewhere within the described range.

Laboratory studies of effects from dietary exposure to PCDD/DFs are limited for
avian species. The TRV selected for use in this assessment was derived from a study
which dosed adult hen ring-necked pheasants (Phasianus colchicus) with TCDD through
intraperitoneal (IP) injection (Nosek et al. 1992). The dietary-based TRVs were
determined by converting the weekly exposure at which adverse effects on fertility and
hatching success were determined (1000 ng TCDD/kg/wk) to a LOAEC for daily
exposure of 140 ng TCDD/kg/d. Adverse effects were not observed at the next lesser

dose, which was determined to be the NOAEC for dietary exposure (14 ng TCDD/kg/d).

Results
ZPCDD/DF and [PCB concentrations

Concentrations of EPCDD/DFs in each type of prey, including frogs, crayﬁsh, and
forage ﬁsh, exhibited consistent spatial trends that differed slightly from the trend
observed in sediment. In prey items, concentrations of ZPCDD/DFS were least in RA,
greater in UTR, greatest in LTR, and intermediate in the SR (Table 3.1). Mean
concentrations of ZPCDD/DFS were 9-, 18-, and 1- fold greater in frogs, 28-, 72-, and 10-
fold greater in crayﬁsh, and 47 -, 58-, and 13-fold greater in forage ﬁsh collected from the

UTR, LTR, and SR than those from the RA, respectively. Mean concentrations of

79

Table 3.1. TEQWH0-Aviana in prey items of great blue herons collected during 2004-2006
from the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.
Valuesb (ng/kg wet wt) are given as the geometric mean and sample size in parentheses
(11) over the 95% conﬁdence interval and range (min-max).

 

 

 

Reach
Reference Upper Lower Saginaw
Area Tittabawassee Tittabawassee River
Frog
EPCDD/DF 5.5 (29) A 49 (51) B 100 (55) C 6.7 (12) A
4.3-6.9 33-73 78-140 4.7-9.4
(1.8-21) (4.4-920) (17-3300) (3.3-26)
ZPCB N/A N/A 1300 (4) N/A
880-2000
(940-1700)
DF'TEQWHO-Avian 1.0 (29) A 20 (51) B 56 (55) C 2.9 (12) A
081—] .3 13—31 42—76 1.9—4.5
(029—3 .4) (1.1—460) (9.1—1900) (1 .5—14)
PCB'TEQWHO-Avian N/A C N/A 1.5 (4) N/A
' 086—25
(1 .1—2.0)
Crayﬁsh
ZPCDD/DF 5.0 (5) A 140 (7) B 360 (8) C 50 (8) D
1 .8—1 3 83—240 200—650 34—72
(2.3—12) (86—340) (140—1300) (28—110)
ZPCB N/A -6 (2) — (1) N/A
(2200-2200) (2100)
DF‘TEQWHO-Avian 0.91 (5) A 55 (7) B 160 (8) C 27 (8) B
0294.8 24—130 110—250 18-41
(034—29) (12—190) (75—420) (13—61)
PCB‘TEQWHO-Avian N/A “ (2) ‘ (1) N/A
(3.3-4.3) (5.3)
Forage Fish \
ZPCDD/DF —— (2) -— (2) 260 (5) A 59 (4) B
_. — 95—730 26—130
(4. 1—5. 1) (200—220) (83—610) (37-210)
ZPCB — (2) — (2) 18000 (5) A 25000 (4) A
_ — 4600—72000 10000—62000

80

Table 3.1 (cont’d)

 

(880-1000) (9300—15000) (7100—110000) (15000—47000)

DF-TEQWHO-Av.-an — (2) - (2) 180 (5) A 33 (4) B
-— - 66—480 20—56

(0.90—.091) (130—170) (74—440) (2553)

PCB-TEQWHO-Avian - (2) — (2) 31 (5) A 55 (4) A
-— —— 10—92 12—250

(1.3-1.8) (30—39) (13—100) (24-150)

 

a TEQwHo-Avian were calculated based on the 1998 avian WHO TEF values
b Values have been rounded and represent only two signiﬁcant ﬁgures
c N/A = no samples collected from this location

d Means identiﬁed with the same uppercase letter are not signiﬁcantly different among
reach at the p=0.05 level using Tukey-Kramer means separation test.

e - geometric mean and conﬁdence intervals not calculated for sites with fewer than 3
samples. These sites were not included in reach comparisons.

81

ZPCDD/DFS in sediments were least in the RA (31 ng/kg, dw), greatest in the UTR (1400
ng/kg, dw), and then decreased downstream in the LTR (960 ng/kg, dw) and the SR (350
ng/kg, dw). Consistent with the aforementioned spatial trends, concentrations of
ZPCDD/DFS were statistically signiﬁcantly greater in dietary components collected from
reaches of the Tittabawassee River compared to the RA. Similarly, concentrations of
EPCDD/DFs in dietary items from the SR were also greater that their RA collected
counterparts, but the differences varied in their signiﬁcance (Table 3.1). Concentrations
of ZPCDD/DFs were signiﬁcantly greater in sediment and crayﬁsh collected from the SR
compared to those from the RA (p=0.0066 and p=0.0101). In frogs, concentrations of
were not signiﬁcantly different between the RA and SR (p=0.9526).

Trends in concentrations of ZPCDD/DF were also observed among prey item types.
In general, concentrations of ZPCDDfDF were least in frogs and similar between crayﬁsh
and forage ﬁsh. Mean ZPCDD/DF concentrations in frogs were 3- to 4-fold less than in
crayﬁsh and forage ﬁsh collected from the UTR and LTR and 7- to 9-fold less than in
crayﬁsh and forage ﬁsh from the SR reach. Concentrations of ZPCDD/DF were not
different among prey items in the RA (Table 3.1).

Spatial trends also occurred in the relative contribution of PCDDS, PCDFS, and
individual congeners to ZPCDD/DF in dietary components. Sediment from the RA had a
greater percent contribution of dioxins to ZPCDD/DF than those collected from the UTR,
LTR, and SR reaches (80% compared to 56%, 43%, and 42%, respectively). A similar
pattern was observed in prey of GBH, with RA frogs, crayﬁsh, and forage ﬁsh having
67%, 61%, ad 74%, of their ZPCDD/DF contributed by dioxins, respectively. In the

RA, octachlorodibenzo-p-dioxin (OCDD) was consistently the greatest proportion of

82

ZPCDD/DF in all dietary components. F urans contributed between 60% to 85% of the
ZPCDD/DF in prey items collected from TR reaches, and 54% to 74% to those collected
from the SR. 2,3,7,8-tetrachlorodibenzofuran (TCDF) is the predominant congener
contributing to ZPCDD/DF in prey items (23% to 56%) collected from the UTR, LTR,
and SR. These patterns remained constant among the different dietary components
(Figure 3.2).

Concentrations of ZPCBs in prey items followed a different spatial trend than that of
EPCDD/DFs, being greatest in the SR. Inadequate sample size precluded statistical
comparisons among reaches for most prey item taxa. Concentrations of EPCBs in forage
ﬁsh collected from the SR reach were greater than those from the LTR, although the
difference was not signiﬁcant (p=0.48l2). PCB congeners 105 and 118 were the

predominant congeners present.

DF‘TEQli’HO-Avian and P CB‘TEQll-’HO-Avian
Concentrations of DF-TEQWH0_A,.,an in prey of GBH were least in the RA, greater in

UTR, greatest in LTR, and intermediate in the SR (Table 3.1). This was the same trend

observed for concentrations of ZPCDD/DFS from which they were calculated. Mean
concentrations of DF-TEQWHO_Avian were 20-, 56—, and 3-fold greater in frogs, 60-, 180-,

and 30-fold greater in crayﬁsh, and 170-, 200-, and 36-fold greater in forage ﬁsh

collected from the UTR, LTR, and SR when compared to the RA, respectively. In
sediment, mean concentrations of DF‘TEQWHQ-Avian were 130-, 240-, and 84-fold greater
at UTR, LTR, and SR reaches than the RA, respectively. Concentrations of DF-

TEQSWHO_ Avian in prey items from the UTR and LTR were statistically greater than those

83

 

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collected from RA, which was consistent with concentrations of EPCDD/DF. Statistical
comparisons of the concentrations of DF-TEQWHO.Avian in prey from SR to those from

RA were also similar to those based on ZPCDD/DFs, except for crayﬁsh from UTR
which were not statistically different from those collected from the SR (p=0.1322).

The relative contribution of compound class and individual congeners to DF-
TEQWHO-Avian in dietary components also exhibited a spatial trend throughout the study

area. Reference area sediment had a greater contribution of dioxins to DF-TEQWH0-A,,,an

than those collected from the UTR, LTR, and SR reaches (38% compared to 3.7%, 2.6%,

and 3.4%, respectively). Similarly, prey items collected in the RA had a greater
contribution of dioxins to DF-TEQWHO-Avian, primarily from 2,3,7,8-TCDD (13% to
31%) and 1,2,3,7,8-PeCDD (13% to 18%), than those from UTR, LTR, and SR reaches.
F urans contributed between 86% to 99% of the DF-TEQWHO-AV,,m in prey itemslcollected

from the Tittabawassee River reaches, and 78% to 97% to those collected from the SR.

2,3,7,8-TCDF and 2,3,4,7,8—pentachlorodibenzofuran (PeCDF) are the predominant
congeners contributing to DF-TEQWHOAWM in prey items (51% to 82% and 12% to 25%,
respectively) collected from UTR, LTR, and SR.

Spatial trends are also seen in the relative contribution of DF-TEanomm and
PCB-TEQWHOAVian to total TEQwHO—Avian in prey items. In the RA, PCBs accounted for
a majority of the total TEQwHo-Avian (63%) in ﬁsh, compared to ﬁsh collected from the
Tittabawassee River which have a majority of their total TEQWHO-Avian attributed to DF-
TEanamm (81% and 85% for the UTR and LTR, respectively). Total concentrations

of TEQSWHO-Avian were dominated by PCB-TEst;.10_Avian in ﬁsh collected from the

85

 

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Saginaw River (61% PCB- TEQWH0_Avian). PCB-TEstmyAvian were dominated by

congeners PCB-126, PCB-77, andPCB-81 in all reaches.

Dietary exposure

ADDpot for GBH was consistently greatest within reaches of the Tittabawassee River
when compared to either the RA or SR, regardless of whether it was based on
ZPCDD/DF, DF-TEQWH0-A,,ian, or total TEQWHO_AV,an (Table 3.2). The ZPCDD/DF
ADDpot to resident GBH was 45- to 54-fold greater along the Tittabawassee River and

12—fold greater along the SR when compared to the RA. When based on DF-TEQWHO-

Avian fold differences in ADme were greater, being 150- to l90-fold greater along the
Tittabawassee River and 36-fold great along the SR when compared to the RA. The
ADDpot of total TEQWHO-Avian based on a 100% ﬁsh diet was 75— to 86-fold greater along

the Tittabawassee River and 39-fold greater along the SR when compared to the RA.

Dietary—exposure risk characterization

Estimates of dietary exposure based on 100% site use and geometric means of
measured concentrations of DF-TEQSWH0-A,.,an and total TEQSWHQ.Avian at

Tittabawassee River reaches were greater than the diet-based NOAEC TRV, which
resulted in HQs that were slightly greater than 1.0 (Figure 3.3). Maximum calculated

HQs along the Tittabawassee River, based on the most conservative parameters of 100%

site use and upper 95% upper conﬁdence limit (95% UCL) of measured DF-TEQSWHO-
Avian and total TEQSWHO-Aviana were between 1.0 and 10. When based on the more

86

Table 3.2. Predicted daily dietary dose of ZPCDD/DFs and TEstmmm a (ng/kg

body weight/d b’ c) for adult great blue herons breeding during 2004-2006 within the
Chippewa, Tittabawassee, and Saginaw river ﬂoodplains, Midland, Michigan, USA,
based on the geometric mean (95% conﬁdence interval) of site-speciﬁc dietary items.

 

 

 

Study Area ZPCDD/DFs DF—TEQW,-.()-A,,,,“ Total TEQWHO-AV.-ana‘ d
Reference C

100% On-site 0.97 (0.79—1.3) 0.17 (016—0. 18) 0.44 (0.40~0.49)
Upper Tittabawassee River f’ g’ h

100% On-site 44 (39—52) 26 (23~3 l) 33 (29—38)
75% On-site 33 (29—-39) 20 (17—23) 25 (22—28)
25% On-site 12 (ll-14) 6.7 (5.9—7.9) 8.5 (7.5-9.8)
Lower Tittabawassee River i

100% On-site 52 (19—140) 33 (12—87) 38 (14—100)
75% On-site 39 (l4~110) 24 (9.3—65) 29 (1 1—77)
25% On-site l4 (5.3—37) 8.3 (3.2—22) 9.8 (3.8—26)
Saginaw River

100% On-site 12 (4.8—45) 6.2 (3.5—13) 17 (5.4—51)
75% On-site 9.5 (3.935) 4.7 (2.7-9.7) 13 (4.l~38)
25% On-site 3.8 (1.8—12) 1.7 (1.0—3.4) 4.5 (1.6—13)

 

a TEQwHo-Avian were calculated based on the 1998 avian WHO TEF values

b . .
Values were rounded and represent only two Signiﬁcant ﬁgures

c Food ingestion rate was calculated from equations in The Wildlife Exposure Factors

Handbook (USEPA 1993)

d Total TEQWHGAWam based on a diet of 100% ﬁsh due to lack of PCB data in frogs and

crayﬁsh

6 Two ﬁsh composite samples collected from Reference reach. Range represents daily

dietary dose based on minimum-maximum of ﬁsh concentrations.

fOff-site diet assumed to be equal to reference area diet.
g Upper Tittabawassee River reach includes sites T-3 and T—4

h . . .
Two ﬁsh composne samples collected from Upper Tittabawassee River reach. Range
represents daily dietary dose based on minimum-maximum of ﬁsh concentrations.

i Lower Tittabawassee includes sites T-S, T-6, and S-7

87

Figure 3.3. Range of hazard quotients for dietary-based exposure of great blue herons to
DF-TEstmyMian and total TEQSWH()-A,.i,,n along the Chippewa, Tittabawassee, and

Saginaw river ﬂoodplains based on assumption of 100% (A) and 75% (B) site-use.

88

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realistic estimate of 75% site use, dietary exposure-based estimates of HQs based on the
geometric means of measured DF- TEQSWHamian and total TEQSWHOAVM“ at

Tittabawassee River reaches were still greater than the diet-based NOAEC TRV, which

resulted in HQs slightly greater than 1.0. Dietary exposure-based estimates based on the
95% UCL of measured concentrations of DF-TEstugAvian and total TEQSWH0-AV,an

compared to the diet-based LOAEC TRV never resulted in HQs greater than 1.0,

regardless of percent site use or reach.

Discussion
PCDD/DFs and PCBs in GBH prey

The observed trends in concentrations of PCDD/DFs in prey items of GBH were
consistent with those observed for dietary items and tissues of other receptor species
studied within the SA (Coeﬁeld et al. 2010; Fredricks et al. 2010b; Fredricks et al.
2010d; Seston et al. 2010b). The pattern observed is likely the result of the downstream
movement of this historical contamination from the upstream source. The presence of
this trend in predicted GBH dietary exposure is a result of the large proportion of the
GBH diet comprising small ﬁsh that tend to be young and are more intimately connected
to local sediment contamination. As such, GBH foraging in downstream reaches of the
TR would be consuming prey items exposed to sediment containing greater
concentrations of PCDD/DFs. The direct connection of concentrations of PCDD/DFs in
sediment and GBH exposure has previously been demonstrated. A rapid decrease in the

concentration of PCDD/DFs in GBH eggs was observed directly following changes in

91

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technologies employed at local pulp mills which decreased PCDD/DF releases (Elliott et
al. 2001).

The observation that concentrations of PCBs were greater in forage ﬁsh from the SR
compared to those from the TR is consistent with historical PCB contamination in areas
downstream of the TR, including the Saginaw River and Saginaw Bay (Kannan et al.
2008) as well as some lesser sources upstream of the RA. Anadromous ﬁshes from
Saginaw Bay can move up into the TR, but there is a dam which may impede the
movement of these ﬁsh into the RA. Furthermore, comparisons between ﬁsh and the
more spatially conﬁned frogs and crayﬁsh (Hazlett et al. 1974; Martof 1953), found that
ﬁsh generally contained greater concentrations of PCBs than frogs or crayﬁsh collected
from within the same reach. This also suggests that ﬁsh may be acquiring PCBs from
areas other than the TR. Additionally, there may be incidental PCB inputs from the

urbanized Midland, MI area.

Dietary exposure

Predictions of dietary exposure integrate measured contaminant body burdens of prey
randomly selected at the correct proportions based on designated dietary composition.
Concentrations of PCDD/DPS among species were generally spatially consistent while

the relative contribution of contaminant compound class and individual congeners were

not. The ADDpot of total TEQWHO- Avian was 13-21% greater than the ADDpot of
DF'TEQWHO.AVian along the reaches of the TR, but was 61% and 64% greater in the RA
and SR, respectively. The relative proportion of DF-TEanoAvm to ZPCDD/DF for
ADDpot of GBH was 18% in the RA, while in the SA they ranged from 59-62% and 52%

92

for reaches along the TR and SR, respectively. Greater percentages observed along the
TR and SR are due to prey items from those reaches having greater relative TCDD
potency of EPCDD/DF, primarily from 2,3,7,8-TCDF and secondarily from 2,3,4,7,8-
PeCDF.

Exposure of GBH through their diet was consistent with spatial trends of

concentration of contaminants in sediments and prey items. However, the predicted
ADDspot were greater than would be expected based on published congener-speciﬁc

biomagniﬁcation factors and site-speciﬁc concentrations in GBH tissues (Braune and
Norstrom 1989; Kubiak et al. 1989; Seston et al. 20103; Thomas and Anthony 1999).
One plausible explanation based on site—speciﬁc observations of GBH foraging and
published foraging characteristics is that GBH nesting along the TR forage off-site for a
portion of their diet. Adult GBH forage a mean distance between 3.2 km and 6.5 km
from rookeries during the breeding season, with a distance as great as 34 km recorded
(Dowd and Flake 1985; Peifer 1979; Thompson 1978). The large foraging range of GBH
leads to a potential for spending at least a portion of their time foraging off-site,
especially early in the breeding season when water levels and turbidity are less than
optimal for foraging in the river channel. Therefore, assuming 100% site-use is likely
overestimating percentage of time resident GBH are foraging within the TR ﬂoodplain.
As such, the dietary exposure estimates based on 75% site-use presented are likely more
realistic.

The relative contribution of individual congeners was similar between site-speciﬁc
prey items and GBH tissues. 2,3,7,8-TCDF was the predominant congener present in

prey items sampled from the SA, with 2,3,4,7,8-PeCDF also having a signiﬁcant

93

presence. In GBH tissues, the overall congener proﬁle was similar to that observed in
prey items, although there was a change in the predominant congener from 2,3,7,8-TCDF
to 2,3,4,7,8-PeCDF. This pattern was also observed in other species foraging in the SA,
including great horned owls, belted kingﬁshers, and several passerine species (Coeﬁeld et
al. 2010; Fredricks et al. 2010a; Fredricks et al. 2010c; Seston et al. 2010b). Field
studies have reported negligible bioaccumulation of 2,3,7,8-TCDF from prey by F orster’s
terns and herring gulls (Braune and Norstrom 1989; Kubiak et al. 1989). This may be
due to preferential metabolism of 2,3,7,8-TCDF. A number of studies have reported
similar clearance by various avian species exposed to mixtures of AhR-active compounds
(Elliott et al. 1996; Kubota et al. 2005; Senthil Kumar et al. 2002). Data on avian
toxicokinetics from controlled laboratory studies are limited, however, mammalian
studies have shown the rate of metabolism of 2,3,7,8-TCDF to be elevated with increased
concentrations of dioxin-like compounds and, the subsequent induction of cytochrome
P450 1A1 and/or 1A2, while 2,3,4,7,8-PeCDF is preferentially sequestered in the liver
(van den Berg et al. 1994b; Zwiemik et al. 2008). This difference in toxicokinetics is a
likely explanation for the observed difference in the relative contribution of the two furan

congeners from diet to tissue.

Dietary-exposure risk characterization
Dietary exposures of GBH along the TR are not expected to result in adverse
population-level effects. Maximum calculated HQs along the Tittabawassee River, based

on the most conservative parameters of 100% site use and 95% UCL of measured DF-

TEQSWHQ_Avian and total TEQSWHO-Aviana remained between 1.0 and 10. Estimates of

94

dietary exposure based on the 95% UCL of measured concentrations of DF-TEQSWHO-

Avian or total TEQSwHo-Avian in prey were never greater than the TRV based on the

LOAEC, regardless of site use factor or reach. The actual threshold concentration at
which effects would be expected to occur lies somewhere within the range delimited by
the NOAEC and LOAEC, which are a function of the experimental design of the study
from which they were derived. Since only a small portion of the range for GBH dietary
exposure exceeds an HQ of 1.0, it is possible that none of the dietary exposures predicted
here exceed the actual effect threshold concentration. However, forage ﬁsh, which were
the primary component of the GBH diet, were composites of all individuals captured per
location per sampling period. This provided an accurate estimate of the central tendency
of concentration estimates in forage ﬁsh but limited characterization of the variability of
concentrations among all forage ﬁsh at each sampling location. Thus, the upper end
estimates of dietary exposure presented here, and the risk associated with that exposure,

may be an underestimate of dietary exposure.

Uncertainty assessment

Concern associated with site-speciﬁc adverse effects can be assessed by using the HQ
approach or a more probabilistic approach when data permits; however both of these
processes are inﬂuenced by the selection of TRVs. The selection of TRVs often
introduces signiﬁcant uncertainty to hazard or risk assessments, especially in cases where
comprehensive site-speciﬁc data sets describe exposure with great certainty. HQs greater
than 1.0 are indicative of exposures that exceed the threshold for adverse effects and that

there is the potential for effects to occur. Such assessments are commonly conservative

95

 

 

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when extrapolating from individuals to populations, as they are based on assumptions of
maximal exposure and do not account for compensatory mechanisms associated with
wild populations (F airbrother 2001). Conversely, the limited number of endpoints
generally measured in laboratory studies may not capture an effect that could have
implications at the population-level. Thus, there is uncertainty inherent in the HQ
approach, meaning that HQ values greater than 1.0 do not necessarily translate into
population—level or ecologically relevant adverse effects (Blankenship et al. 2008).
Probabilistic exposure assessment might more accurately describe the exposure

distribution; however the assessment of risk is only as accurate as the, certainty associated
with applicability of the TRV to the population-level dose response curve.

Laboratory studies of effects from dietary exposure to PCDD/DPS are limited for
avian species. The TRV selected for use in this assessment was derived from a study
which dosed adult hen ring-necked pheasants (Phasianus colchicus) with TCDD through
intraperitoneal (IP) injection (Nosek et al. 1992). There are a few factors that make the
use of this TRV conservative in determining the potential hazard of these compounds to
GBH. Firstly, in Nosek et al. (1992) the hens were exposed to TCDD via IP injection
versus a true dietary-based exposure, potentially resulting in greater exposure efﬁciency
than that of gut transfer. Dietary exposure also allows for an increased potential for
metabolism and excretion prior to reaching target tissues. Additionally, this study was
not conducted on a piscivorous species, but on a member of Gallifomies, which are

generally considered to have greater sensitivity to dioxin-like compound exposures
(Brunstrom 1988; Brunstrtim and Reutergardh 1986; Powell et al. 1996b; Powell et al.

1997). Variability in sensitivity among species has been attributed to differences in the

96

 

 

hi

i (-

amino acid sequence of the ligand binding domain of the AhR (Head et al. 2008;
Karchner et al. 2006) with ring-necked pheasants having the construct which classiﬁes
them as moderately sensitive whereas the GBH construct classiﬁes them as insensitive.
Induction of 7-ethoxyresoruﬁn-O-deethylase (EROD) activity was 10-fold greater
following exposure to TCDD in the ring-necked pheasant than in the double-crested
cormorant, a piscivorous species with the same AhR construct as the GBH. Furthermore,
the dosing was done with only a single congener (TCDD), whereas the GBH being

assessed are exposed to a complex mixture of dioxin-like compounds. Although
previously assumed to be equitable in toxicity when normalized using TEFSWHO-Aviana

ﬁndings from recent studies investigating the relative potencies of individual PCDD/DP
congeners in avian species suggest that this may not hold true in every exposure scenario

(Herve’ et al. 20103; Herve’ et al. 2010b).

Multiple lines of evidence

To increase the certainty of the conclusions drawn from the risk assessment of GBH
along the TR ﬂoodplain, exposure of GBH to PCDD/DFs was assessed using both a
“bottom-up” and “top-down” approach (Leonards et al. 2008). The “bottom-up”
approach is exempliﬁed by the dietary exposure approach presented here. The “tOp-
down” approach was done by assessing the potential for effects based on site-speciﬁc
concentrations of PCDD/DFs in eggs and blood plasma of GBH nesting along the TR
(Seston et al., 2010a). Estimates of potential for effects based on these two measures of
exposure were compared to one another and also to a site-speciﬁc assessment of GBH

population condition. Trends in concentrations of PCDD/DFs and the relative

97

contribution of individual congeners were consistent between the “bottom-up” and “top-
down” assessments conducted for GBH nesting along the TR. The results of all three
approaches in this multiple lines of evidence approach were consistent. Based on
comparisons to TRVs from the literature, each line of evidence based on different
estimates of exposure suggests that GBH foraging and breeding along the TR ﬂoodplain
are at minimal to no risk of experiencing adverse population-level effects as a result of
their exposure to PCDD/DFs. This was supported by the site-speciﬁc assessment of

GBH population condition (Seston et al. 2010a).

Conclusions

The ﬁndings of this study predict that GBH foraging along the TR are more greatly
exposed to PCDD/DFs than those in associated reference areas. This prediction is
consistent with spatial trends observed in concentrations of PCDD/DFs in site-speciﬁc
GBH tissues. However, concentrations of PCDD/DFs in eggs of GBH collected from
rookeries within the TR ﬂoodplain were lesser than what would be expected given the
dietary exposure predicted here. This is likely a result of GBH foraging outside of the
assessment area. Direct foraging by GBH within the river corridor was established
through visual observations and trapping of GBH. However, GBH can ﬂy substantial
distances from their rookery and adverse water conditions associated with frequent spring
episodic rain events would be expected to preclude foraging within the river. Field
researchers noted an absence of GBH foraging associated with adverse foraging
conditions; however these observations were limited to the river corridor. Therefore, it is

likely that during these times GBH were foraging in areas unassociated with the

98

assessment area. Site-speciﬁc visual observations and published BMFs suggest that the
dietary exposure predictions presented here are likely an overestimate of their actual
exposure. Despite this potential overestimation, the predicted level of exposure of GBH
to PCDD/DFs has a low probability of causing adverse population level effects, based on
TRVs available in the literature. This conclusion is supported by ﬁndings of site-speciﬁc

assessments of GBH tissue exposure and population condition.

Acknowledgements

The authors would like to thank all staff and students of the Michigan State
University —- Wildlife Toxicology Laboratory ﬁeld crew; namely M. Nadeau, W. F olland
and D. Hamman, along with E. Koppel, J. Moore, L. Williams, and C. Bartrem. We
gratefully acknowledge M. Fales for his design and fabrication of specialized equipment
which was pivotal to the success of this research. Additionally, we would like to
recognize M. Kramer and N. Ikeda for their assistance in the laboratory. The following
people were key to the success of the project, J. Dastyek and S. Kahl of the USFWS —
Shiawassee National Wildlife Refuge for their assistance and access to the refuge;
Saginaw County Parks and Recreation Commission for access to Imennan Park;
Tittabawassee Township Park Rangers for access to Tittabawassee Township Park and
Freeland Festival Park; and Michael Bishop of Alma College for guidance as our Master
Bander. We would also like to acknowledge the greater than 50 cooperating landowners
throughout the study area who have granted us access to their property, making our

research possible. Funding was provided through an unrestricted grant from The Dow

99

 

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Pi

Chemical Company, Midland, Michigan to J. Giesy and M. Zwiemik of Michigan State

University.

Animal Use

All aspects of this study that involved the use of animals were conducted using the
most humane means possible. To achieve that objective, all aspects of the study were
performed following standard operation procedures (GBH adult handling 03/05-036-00;
GBH nest monitoring 05/07-066-00; Field studies in support of TR ERA 03/04-042-00;
Protocol for ﬁsh sampling 03/04-043-00) approved by Michigan State University’s
Institutional Animal Care and Use Committee (IACUC). All of the necessary state and
federal approvals and permits (Michigan Department of Natural Resources Scientiﬁc
Collection Permit SC1254 for GBH/SC permit for ﬁsh (Zwiemik)/SC permit for
amphibians (Zwiemik); USFWS Migratory Bird Scientiﬁc Collection Permit
MB]000062-0; and subpermitted under US Department of the Interior Federal Banding

Permit 22926) are on ﬁle at MSU-WTL.

100

 

 

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Supplemental Information

Table 3.3. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin
congeners in sediment collected during 2003-2006 from the Chippewa/Pine,

Tittabawassee, and Saginaw Rivers, Midland, MI, USA. Valuesa (ng/kg w) are
given as the arithmetic meanb :l: 1 SD over the range.

 

 

RA UTR LTR SR
Chemicalc n=1 6 n=23 n=1 9 n=8
2378-TCDF 0.80i1 .0 2603:490 63 0i960 2903:450
0.039—3.4 10-2100 1.8-3500 2.5—1300
4ND
23478-PeCDF 0.43i0.42 873:150 250i460 130i210
0.073—1.3 3.1—560 0.31—1900 0.57-610
3ND .
12378-PeCDF 05630.55 1103:170 3503750 180i290
0.083—1.8 3.4—570 0.43-3300 0.81-860
7ND
234678-HxCDF 0.51:t0.25 6.93:8.9 22140 233:18
0.14—0.82 0.36—28 0.69-160 6.9-48
l 1ND 2ND 4ND
123789-HXCDF 1.9123 5.2i9.4 213:24
021—73 024—37 2.2-48
16ND 5ND 4ND 5ND
123678-HxCDF 0.703:0.50 163:22 58i120 53365
0.14—l.5 1.7—130 1.6—520 0.34-170
9ND 2ND 2ND
123478-HXCDF 1.33:1.4 873:120 2603660 15032240
0.15—45 4.8—390 026-2400 0.54-710
5ND 1ND 2ND
1234789-HpCDF 0.622t0.25 173:28 193:33 38i21
0.23-40 1.7—130 1.2-140 18*67
IOND 2ND 4ND
1234678-HpCDF 4.3i6.6 210i410 1203110 370i450
0.16—22 ' 19—1900 031-480 0.3 8-1200
12346789-OCDF 9.23:13 7303:2100 2303:180 5203510
0.2240 25-10000 22-620 0.39-1200
1ND 2ND 1ND

101

 

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Table 3.3 (cont’d)

\m

2378-TCDD 0.263:0.16 2.0i1.9 1.5i0.9l 6.9:t7.2
0086—052 026—75 015—32 0.079—18
7ND 1ND 1ND
12378-PeCDD 0.28:t0.16 110i170 1.330.68 9.7:t4.2
0099—050 3.4—170 0.33—2.8 4.5—13
8ND 2ND 4ND
123789-HxCDD 06030.38 3.24:3.9 1931.2 1134.8
0.13—1.1 0.48-16 0.44—4.9 5.4-15
9ND 2ND 4ND
123678-HxCDD 08230.69 10420 4.43:3.0 263:10
016—18 081-96 0.95—13 15—37
8ND 2ND 4ND
123478-HxCDD 03540.13 1.33:1.4 0.85:1:053 3.24:1.3
0.13-0.43 019—52 0.21—1.9 1.2—4.1
llND 4ND 4ND
1234678-HpCDD 6.8310 200i500 64451 1403:160
034—30 1 1—2400 035—190 030—350
12346789-OCDD 643:97 190034800 630i500 130041400
2.0—280 92—23000 3.0—1800 1.7—3100
Sum PCDD/DP 883:130 360038200 260035400 3100i3700
2.9—390 280—40000 6.4—14000 9.7—9700

 

 

 

Mded and represent only two signiﬁcant ﬁgures
b Concentrations below limit of detection treated as zero in the calculation of
arithmetic means

C TCDF =tetrachlorodibenzoﬁ1ran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF =
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

102

 

 

 

ilﬁr
, .

 

 

Chemicalc ”:29 11:5 1 ”:5 5 ":1 2

RA

UTR

LTR

SR

2378-TCDF 04440.47 42468 864190 1.640.93
0.076—2.2 034—370 5.1—1400 070—35
2ND

23478-PeCDF 02040.10 1 1414 30464 1342.7
0.081—0.49 021—69 1.9—450 033—97
9ND

12378-PeCDF 01940.079 8.0413 264100 0.334016
0.10—0.28 0080-75 1.3—770 0.17—0.60
24ND

234678-HxCDF _ 0.654061 1.7459 0.57
0.13-0.18 015—28 011—37
27ND 24ND 16ND llND

123789-HxCDF 5.8
29ND 5an S4ND 12ND

123678-HxCDF 1.0415 2.5411
0.11—0.16 011—74 0087—76 014—1.]
27ND 14ND 4ND IOND

123478-HxCDF 02040.088 3.5467 10450 08941.2
0092—034 0.16—33 027—370 027—30
ZOND 8ND 7ND

1234789-HpCDF 1.24093 1.9438

0.20-2.6 023—14
29ND 42ND 43ND 12ND

1234678-HpCDF 0.494029 6.3410 5.6483 03740.29
0.16—0.92 023—42 029—50 019-13
24ND IOND

12346789-OCDF 0.724035 10418 9.9415 0.30
038—12 022—72 021—67
21ND IOND 2ND 1 1ND

2378-TCDD 0.434028 1341.1 09640.77 03540.13
0.12—1.1 027—74 0.20—4.1 0.20-0.58

103

 

 

Table 3.4 (cont’d)
I
4ND 1ND 1ND
123 78-PeCDD 0.254013 06640.44 0.574032 0.374021
0.12-0.55 0.20-2.7 0.18—1.6 0.18—091
9ND 1ND 3ND 1ND
l23789-HxCDD 0.18 0.304015 02940.081
0.13-0.70 0.17~04l
28ND 33ND 47ND 12ND
123678-HxCDD 0.304012 0.744061 0.724048 0.374033
0.16—0.43 014-26 023—22 013—1.]
24ND 5ND 13ND 4ND
123478-HxCDD 0.304012 02640.077 0.54
0079-059 0.19—0.42
29ND 23ND 43ND ~ llND
1234678-HpCDD 0684057 4.0458 4.0442 05340.39
022—28 029—28 045—22 0.20—1.7
1ND
12346789-OCDD 3.243.4 31456 31440 1.540.51
057-13 1.3—250 2.6—1220 067-23
Sum PCDD/DP 5.4448 1204180 2104450 7.246.]
096-19 3.5—920 17—3300 2.8—25

 

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of
arithmetic means

C TCDF = tetrachlorodibenzoﬁlran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF =
octachlorodibenzoﬁlran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

104

 

Table 3. 5.

w) are given as the arithmetic meanb 4 1 SD overt

 

Concentrations of seventeen 2 ,3 ,7 ,8
congeners in crayﬁsh collected during 2005- 2006

Tittabawassee, and Saginaw River ﬂoodplains, Mid

 

 

 

 

 

 

 

RA

Chemicalc
2378-TCDF 05540.44

019—13
23478-PeCDF 0.284023

0051—059

1ND
12378-PeCDF

0031-045

3ND
234678-HxCDF

0018—018

3ND
123789-HxCDF

5ND
123678-HxCDF

0019—041

3ND
123478-HxCDF 0.564045

0048—089

2ND
1234789-HpCDF

5ND
1234678-HpCDF 04740.32

0.11—0.68

2ND
12346789-OCDF 0.364020

0.17—0.57

1ND
2378-TCDD 02640.18

0056—037

UTR

60453
7.4—160

1249.1
2.6~26

16414
2.2-39
1ND
04340.26
0.20—0.77
3ND

0064—0. 12
5ND
09740.46
0.47—1.7

5242.6
1.9-8.7

07140.52
0.21—l.4
3ND
7.8493
2.3—28

13416
1.8—48

08940.19
054-1.1

'105

LTR

140474
5 9—3 00

35427
12—96

44429
16—110

1.741.6
0.44-4.9
1ND
04840.38
0.19-0.91
5ND
3943.8
1.2-13

19417
5.3—57

1.741.2
O.64-3.6
3ND
26425
3.3-64

31431
3.5—87

1.34034
O.84——1.8

-substituted furan and dioxin
from the within the Chippewa,

land, MI, USA. Valuesa (ng/kg
he range.

n=5 n=7 n=8 n=8
.

8ND

8ND
05040.24
0.25—0.72
5ND
1.34064
076-25

8ND
2.8423
077—7.]

2742.0 '
0.73—6.6

0.914065
035—21

 

Table 3.5 (cont’d)

 

 

 

 

m

2ND 2ND
12378-PeCDD 0.234020 0.494015 08640.38 04940.19
0030—044 0.30—0.65 0.44—1.6 0.25—0.71
2ND 4ND —
123789-HxCDD 0.254012
0.15—0.39 019—16
5ND 4ND 6ND 8ND
123678-HxCDD 0.594034 1.4413 0.36
024—11 022—37
5ND 3ND 2ND 7ND
123478-HxCDD 01440.049
0.1 1—0.20 0.14—0.90
- 5ND 4ND 6ND 8ND
1234678-HpCDD 04240.32 5,045.8 15419 1.44086
0.18—0.83 1.4—18 1.9-58 052—29
12346789-OCDD 2.8421 44454 1404170 1046.8
088—54 8.1—170 14—520 2.7-22
Sum PCDD/DF 5.4446 1604100 4504380 52427
1.5—11 83—340 140—1300 27—1 10

 

W rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of
arithmetic means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF =
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

106

 

 

 

~.l

I‘ll.

Table 3.6. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin
congeners in forage ﬁsh composites collected during 2004-2007 from the within the
Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA.

Valuesa (ng/kg w) are given as the arithmetic meanb 4 1 SD over the range.

 

 

RA UTR LTR SR
Chemicalc n=2 n=2 ":5 "=4
2378-TCDF 17041 10 2748.4
0.47—0.49 99—140 53—300 21-40
23478-PeCDF 35430 45426
0058—0077 19—22 59-78 2.1—8.0
12378-PeCDF 31430 4.0430
14—14 4.3—78 2.0—8.3
2ND
234678-HxCDF 2.9426 0.184021
0.82—0.91 035—63 00080—042
2ND 1ND 1ND
123789-HxCDF
0.38—0.65 -
2ND 2ND 3ND 4ND
123678-HxCDF 0.055 4545.2 0.554053
1.5—1.5 021—13 015—13
1ND
123478-HxCDF 19424 2.0420
3.1—5.2 084-58 054—49
2ND
1234789-HpCDF 1,341.1 01940.15
0.50—0.54 026—26 0063—013
2ND 1ND 1ND
1234678-HpCDF 8.2445 3.7426
0072-0092 51—55 2.8—13 l.3—7.1
12346789-OCDF 1248-1 4.5i3-9 ‘
0.29—0.35 6.5—11 3.0—25 1.3—9.8
2378-TCDD 4.6426 1.74096
0.20—0.23 4.7—6.0 1.4—10 094-30

107

 

Table 3.6 (cont’d)

12378-PeCDD 1.24065 05240.26
0095-013 1.2-1.6 O.33—l.9 028—084
123 789-HxCDD 0.33
0.12-0.49 0063-084
2ND 4ND 2ND
123678-HxCDD 1.04049 0.55403]
088*] .5 039-16 0.21~0.85
2ND 1ND
123478-HXCDD 1.24019 0.040
052-10 099-14
2ND 2ND 3ND
1234678-HpCDD 5 .0432 1941.2
0.33-0.34 3.6—4.1 1.4-10 080-35
12346789-OCDD 31414 1341 1
2.1-3.0 20—28 8.3—44 4.2-28
Sum PCDD/DF 3304220 64436
3.8-4.6 200—220 81-610 36-120

 

 

 

 

 

 

 

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of
arithmetic means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzoﬁiran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzoﬁiran; OCDF =
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

108

 

Table 3.7. Concentrations of twelve dioxin-like polychlorinated biphenyl congeners
in forage ﬁsh composites collected during 2004-2007 from the within the Chippewa,

Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesa (pg/kg

w) are given as the arithmetic meanb 4 1 SD over the range.

 

 

RA UTR LTR SR
Chemicalc n=2 n=2 n=5 n=4
PCB 77 0.284016 1.032095
0012—0016 036—047 012—051 022-22
PCB 81 01740.20 01840.12
00042-00059 0084—01 1 0043—052 0084—033
PCB 126 09540.13 006140032
00026—00033 0031—0042 0017—033 0.03 3~0099
PCB 169 0008940010
00026—0021
2ND 2ND 2ND 4ND
PCB 105 7.5410 6.043.]
0.17—0.22 1.8-3.3 1.6—26 3.3—9.8
PCB 114 06140.85 05540.34
0012-0016 014-027 0.12-2.1 0.26-0.95
PCB 118 20426 1849.7
0.52~062 5.7—9.0 4.3-66 9.0-29
PCB 123 0564073 0434026
0020—0022 018-029 0097-19 0.21-0.74
PCB 156 2.1430 1.34046
0069—0077 056-084 045-75 089—19
PCB 157 0454063 02740085
0015—0017 014—019 0098-16 0.19-0.37
PCB 167 09441.3 0.654021
0040-0040 0.34-0.41 023-32 0.48-0.91
PCB 189 0224033 01340055
00074—00087 059-088 0049-080 0079—019
Sum PCB 33444 28415
0.88—1.0 9.3-15 7.1-110 15-47

M

a . .
Values have been rounded and represent only two Signiﬁcant ﬁgures

b O u o a O 0
Concentrations below limit of detection treated as zero in the calculation of

arithmetic means

109

 

Table 3.8. Concentrations of twelve dioxin-like polychlorinated biphenyl congeners
in frogs and crayﬁsh collected during 2004-2007 from the within the Chippewa,

Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesa (pg/kg

w) are given as the arithmetic meanb 4 1 SD over the range.

 

 

Frog Crayﬁsh Crayﬁsh
LTR UTR LTR
Chemicalc 11:4 n=2 n=l
PCB 77 0011400055 0.086
0.004-0.016 0050-0063
PCB 81 00039400013 0.0035
00024—00054 0.002 7—0003 1
PCB 126 00057400014 0.0057
0.0037~0.0072 00040400073
PCB 169 00048400017 0.0040
00030-00070 00050—00080
PCB 105 01440080 0.43
00504024 027—058
PCB 114 4ND 2ND 1ND
PCB 118 0814038 0.92
0.40-1.1 1.0—1.4
PCB 123 4ND 2ND 1ND
PCB 156 0.2940089 0.86 0.44
0.20—0.41 1ND
PCB 157 4ND 2ND 1ND
PCB 167 01440.090 0.16 0.20
0.033——0.20 1ND
PCB 189 4ND 2ND IND
Sum PCB 1.44034 2.1
0.94-1.7 2.2-2.2

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b . . . . . .
Concentrations below limit of detection treated as zero in the calculation of

arithmetic means

110

 

 

Jiii

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Powell, D. C., Aulerich, R. J., Meadows, J. C., Tillitt, D. E., Stromborg, K. L., Kubiak, T.
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115

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116

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dibenzofurans (PCDFs) and their relevance for toxicity. Critical Reviews in
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dioxins and dibenzofurans in the developing chicken embryo. Memorandum
Report. 1976. Washington, D.C., US. Food and Drug Administration.

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Woodbum, K., Shotwell, M., Khim, J. S., Giesy, J. P., and Budinsky, R. A.
(2008). Toxicokinetics of 2,3,7,8-TCDF and 2,3,4,7,8-PeCDF in mink (Mustela
vison) at ecologically relevant exposures. Toxicological Sciences 105(1), 33-43.

117

CHAPTER 4

Tissue-based risk assessment'of great blue heron (Ardea herodias) exposed to

PCDD/DFs in the Tittabawassee River ﬂoodplain, MI, USA

Rita M. Sestonl, Timothy B. Fredricks], Dustin L. Tazelaarz, Sarah J. Coeﬁeldl, Patrick
W. Bradleyz, John L. Newsted3, Denise P. Kay3, Scott D. F itzgerald4, John P. Giesyl‘s,

and Matthew J. Zwiemik2

IZoology Department, Center for Integrative Toxicology, Michigan State University,

East Lansing, MI 48824, USA

2Animal Science Department, Michigan State University, East Lansing, MI 48824, USA

3ENTRIX, Inc., Okemos, MI 48864, USA

4Department of Pathobiology and Diagnostic Investigation, Diagnostic Center for

Population and Animal Health, Michigan State University, East Lansing, MI 48824,

USA

5 . . . .
Department of Veterinary Biomedical Sc1ences and Tox1cology Centre, UnlverSlty of

Saskatchewan, Saskatoon, Saskatchewan, 37] 5B3, Canada

118

Abstract

Concentrations of dioxin-like compounds, primarily polychlorinated dibenzofurans
(PCDFs), in soils and sediments of the Tittabawassee River (TR) and associated
ﬂoodplains downstream of Midland, Michigan (USA) were greater than upstream sites
and prompted a site-speciﬁc risk assessment of great blue herons (GBH). Tissue exposure
of PCDFs and polychlorinated dibenzo-p-dioxins (PCDDS) was assessed in multiple
GBH tissue types, including blood plasma of adults, eggs, as well as blood plasma,
adipose, liver, and muscle of nestlings. Adult GBH exposure was associated with
foraging area and age class, with concentrations of PCDD/DFs being greater in blood
plasma of adult GBH foraging in the TR compared to those foraging in upstream
reference areas and in older birds as compared to their younger cohorts. Concentrations
of PCDD/DFs and dioxin-like polychlorinated biphenyls (PCBs) in eggs and nestling
tissues of GBH collected from rookeries within the TR ﬂoodplain were generally similar
among rookeries. Mean concentrations of PCDD/DFs in eggs of GBH ranged from 45 to
67 ng/kg,wet wt for the rookeries studied, with a maximum concentration of 210 ng/kg, .
wet wt observed. Adipose consistently had the greatest concentration of PCDD/DFS of
all tissues collected from nestlings of GBH, ranging from 98 to 430 ng/kg, wet wt.
Potential for adverse population-level effects from site-speciﬁc contaminant exposures
were evaluated by comparison to selected toxicity reference values (TRVs). Minimal
risk of adverse population-level effects was predicted when exposures measured in
tissues of GBH collected from rookeries within the TR were compared to appropriate
TRVs. This prediction is consistent with site-speciﬁc measures of population condition,

which included clutch size and number of nestlings per successful nest.

119

Introduction

The Tittabawassee and Saginaw rivers downstream of Midland, MI, USA contain
elevated concentrations of polychlorinated dibenzofurans (PCDFs) and polychlorinated
dibenzo-p-dioxins (PCDDS). This contamination is the result of the historical treatment
and storage of wastes associated with the manufacture of chlorinated compounds at The
Dow Chemical Company, prior to the adoption of modern waste management practices
(USEPA 1986). Sediments and associated ﬂoodplain soils collected from downstream

study areas (SA) were found to contain total concentrations of 2,3,7,8-substituted

PCDD/DFs (ZPCDD/DFS) ranging from 1.0x102 to 5.4x104 ng/kg dry wt, respectively.

Mean ZPCDD/PCDF concentrations in sediments and ﬂoodplain soils in the reference
area (RA) upstream of Midland were 10- to 20-fold less (Hilscherova et al. 2003). The
persistence of these compounds in the environment, combined with their toxicity and
potential to bioaccumulate, has led to concern over the potential exposures of wildlife
species residing in the Tittabawassee and Saginaw River ﬂoodplains.

Polychlorinated dibenzo-p-dioxins and dibenzofurans often occur in the environment
as complex mixtures, commonly with structurally related polychlorinated biphenyls
(PCBs). Due to their lipophilic characteristics and their resistance toward metabolism
(Mandal 2005) these compounds have the potential to be accumulated through the food
web. Threshold toxicological responses of primary concern are mediated through the
aryl hydrocarbon receptor (AhR) and include enzyme induction, teratogenicity,
immunotoxicity, and adverse effects on reproduction, development, and. endocrine
functions (Hoffman et al. 1987; van den Berg et al. 1994a). In particular, AhR-mediated

compounds have been shown to decrease hatching success and ﬂedging success in

120

 

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aquatic avian species (Custer et al. 2005; Giesy et al. 1994; Gilbertson I983; Kubiak et
al. 1989).

The sensitivities of a number of species to AhR-mediated effects have been
determined in laboratory studies or inferred from observations of populations exposed in
the wild. Sensitivities have been shown to vary among species (Brunstrtim 1988). For
example, the domestic chicken (Gallus gallus), which is considered the most sensitive
avian species, is over 1000-times more sensitive to embryo-lethal effects than is the
mallard (Anas platyrhynchos) (Head et al. 2008). Recent research has suggested that the
ligand binding domain (LBD) of the AhR in avian species is highly conserved and can be
classiﬁed by only a few structural configurations. The speciﬁc conﬁguration of the LBD
directly affects the binding afﬁnity of ligand (dioxin-like compounds), thereby
inﬂuencing the organisms response and sensitivity to toxic effects (Head et al. 2008;
Karchner et al. 2006). This conﬁrms the importance of species-speciﬁc exposure and
response data and its applicability between taxa.

Selecting an appropriate study species is a key element of effective ecological risk
assessments (ERAS), especially when Site-speciﬁc ﬁeld studies are to be employed
(USEPA 1994). Piscivorous birds are frequently selected as receptors for use in
evaluating aquatic systems as they can be sensitive to the effects of contaminants and
have the potential to accumulate persistent, lipophilic contaminants through trophic-
transfer. The great blue heron (Ardea herodias; GBH) possesses several characteristics
that make it a suitable species to use as a receptor in ERAS concerning bioaccumulative
contaminants in aquatic environments. As a long-lived, territorial, apex predator in the

aquatic food web, the GBH has great potential to accumulate local contaminants over a

121

long period of time, especially in areas of its range where foraging is available over the
winter months, allowing them to be year-round residents (Butler 1997; Custer et al.
1991). Breeding colonies of GBH are more conspicuous to researchers than single-
nesting species and the condition of the population can more easily be assessed. The
territorial nature of GBH leads to active defense of distinct, identiﬁable foraging areas
proximal to the breeding colony (Marion 1989). Therefore, exposure of GBH to
persistent and bioaccumulative compounds potentially has better deﬁned Spatial
boundaries than do more opportunistically feeding piscivorous birds. GBH have a broad
distribution across geographic regions and habitat types, residing in freshwater, estuarine,
and marine habitats throughout North America, which makes them a potential receptor
for sites of aquatic-based contaminatiOn at many different localities (Butler 1992).
Previous studies have detected local organochlorine contaminants in the tissues of GBH,
which demonstrates their utility in the assessment of Site-speciﬁc contamination (Custer
ct al. 1997; Elliott et al. 2001; Harris et al. 2003; Straub et al. 2007; Thomas and
Anthony 1999).

The primary objective of the present study was to assess the risk that PCDD/DFS in
the Tittabawassee River (TR) basin pose to GBH breeding on-Site. To characterize
PCDD/DF exposure, contaminant concentrations were directly measured in GBH tissues

in terms of XPCDD/DF, EPCB, and 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)

equivalents (TEQWHO_AVian) based on World Health Organization (WHO) 2,3,7,8-TCDD

equivalency factors for birds (TEFWHO-A,,,,,,,) (van den Berg et al. 1998). Concentrations

of PCDD/DPS and the patterns of their relative congener concentrations in GBH tissues

were also evaluated for spatial trends. Additionally, tissue concentrations were compared

122

 

 

 

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to toxicity reference values (TRVs) to estimate the risk present. Parameters associated
with individual and population condition were then assessed and used to conﬁrm risk
estimates. Assessments of GBH tissue-based exposure and population condition
presented here each represent a separate line of evidence, facilitating a multiple lines of
evidence approach to determine the risk present to GBH along the TR (F airbrother 2003).
Integrating the data resulting from these multiple assessments can reduce uncertainty
inherent in the risk assessment process and provide sound data for use in decision making

regarding the ﬁiture of the Site (Leonards et al. 2008).

Methods
Site description

The assessment was conducted in the vicinity of the city of Midland, located in the
east-central lower peninsula of Michigan, USA (Figure 4.1). The TR is a tributary of the
Saginaw River (SR), which empties into Saginaw Bay and Lake Huron. The TR runs
through The Dow Chemical Company (DOW) property, which is located on the southern
edge of Midland and is the accepted source of the PCDD/DP contamination (USEPA
1986). The area henceforth referred to as the study area (SA) includes approximately 37
km of the TR and associated ﬂoodplain wetlands from DOW to the conﬂuence of the TR
and SR and 35 km of the SR until it enters Saginaw Bay. The reference area (RA) is
composed of the TR upstream of Midland, together with the Pine and Chippewa Rivers,
both of which are tributaries of the TR upstream of Midland. Blood plasma was collected

from adult GBH at established trapping stations throughout the reference and study areas.

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Over the course of the present study 12 adult GBH trapping stations were established,
with three in the RA and nine in the SA. Eggs and tissues of GBH nestlings were
collected from rookeries located in the SA. Within the SA there were three active GBH
rookeries located in the ﬂoodplain; the Freeland (FRE), Shiawassee (SNWR), and Cass
River (CAS) rookeries. They were located 11, 27, and 36 km downstream of Midland,
MI, respectively (Figure 4.1). Established in 2001, FRE was discovered mid-season 2006
and contained 44 nests in 2006 and 46 nests in 2007 (Personal communication,
landowner). On the Shiawassee National Wildlife Refuge, SNWR was established in
1999 and at one time contained as many as 161 nests. Nest occupancy of GBH
drastically decreased after predatory avian species, including bald eagles (Haliaeetus
leucocephalus), great horned owls (Bubo virginianus), and red-tailed hawks (Buteo
jamaicensis) established nests in the rookery. It has since been abandoned by GBH. The
CAS was located near the conﬂuence of the Cass and Shiawassee rivers in the
Shiawassee National Wildlife Refuge and contained approximately 35 nests throughout

the present study. No rookeries were discovered in the RA.

Capture and handling of adult GBH

A total of 51 blood samples were collected from adult GBH trapped while foraging
along the Tittabawassee, Pine, and Chippewa rivers during 2005-2007. Adult GBH were
captured using modiﬁed foot hold traps set around feeding stations in predetermined
GBH foraging areas. Trapping along this river system was limited to summer months,
when the river’s water levels were lower and more stable. Individual GBH were ﬁtted

with USFWS bands on the tarsus and colored leg bands on the tibia (Simpson and Kelsall

125

1978). Individuals were aged using plumage characteristics, and were classiﬁed as either
hatch year (HY) or after hatch year (AHY). In-depth descriptions of adult GBH capture

and handling are described in Seston et al. (2009).

Nest monitoring and fresh egg sampling

Great blue heron rookeries were visited several times during the 2006 and 2007
nesting season to monitor reproductive success. In mid—Michigan,,the GBH nesting
season begins in March and runs through mid-July. Parameters of reproductive success
monitored include clutch Size and the number of nestlings per successful nest. The
number of nestlings per successful neSt was counted during nestling plasma sampling,
when chicks were estimated to be 5 weeks of age. An abbreviated description of the
methods is presented here, as they have been described in detail in Seston et al. (2009).

Eggs were collected from 24 GBH nests, eight in each of the studied rookeries. Nests
were located in Eastern cottonwood (Populus deltoides), silver maple (Acer
saccharinum), or white ash (Fraxinus americana) trees at heights ranging from 15 to 25
m. Fresh egg sampling often involved incubation disturbance within the rookery so it
was only done when temperatures were greater than 15°C. Viable and nonviable eggs
were collected from accessible nests in each rookery. A maximum of one viable egg was
collected at random from each accessed nest that contained 2 two eggs. In addition to
viable eggs, all eggs that failed to hatch were collected for analysis of developmental
stage and contaminant content. Eggs were weighed, measured, and carefully transported

to the laboratory in a crush-proof, water-proof container and kept at 4°C until processing.

126

Capture and handling of nestling GBH

Blood plasma of nestling GBH was collected from 23 nests (n= 12, 4, and 7 nests at
FRE, SNWR, and CAS, respectively). When possible blood plasma of nestlings was
collected from the same nests from which eggs were collected. Blood was collected from
nestlings when they were approximately 4 to 5 wk old based on methods previously
described in Henny and Meeker (1987). At this age, nestlings were still limited to the
nest yet had sufﬁcient mass to provide an adequate volume of plasma for residue analyses
(McAloney I973). Individuals were ﬁtted with USFWS bands on the tarsus and colored
leg bands on the tibia (Seston et al. 2009).

Salvaged nestlings were collected opportunistically following weather events or
siblicide. Tissues of salvaged GBH nestlings, including adipose, liver, and muscle, were
collected and analyzed from a total of 11 individuals, collected from six nests. Sample
size of tissues precluded inter-rookery comparisons. Nestlings were examined for gross
external or internal abnormalities including liver, kidney, spleen, intestine, and gonad
histology. Nestling stomach contents were analyzed to the lowest taxonomic

identiﬁcation possible to determine the site-speciﬁc dietary composition.

Blood plasma sampling

Blood from nestling and adult GBH was drawn from the brachialis vein with needles
afﬁxed to sterile syringes, which had been pre-rinsed with sodium heparin solution. The
volume of the blood sample procured was determined by calculating 10% of 7% of the
bird’s body mass (Hoysak and Weatherhead 1991; Sturkie 1986). The volume of blood

collected per individual ranged from 5 mL up to 10 mL. The blood sample was then

127

 

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transferred to a heparinized VacutainerTM (Becton Dickinson, Franklin Lakes, NJ) for
transport back to the ﬁeld laboratory. Whole blood samples were centrifuged and the
plasma (supernatant) was decanted. Both plasma and packed blood cells were stored at -

20°C until analysis.

Sample processing and analytical techniques

Concentrations of seventeen 2,3,7,8-substituted PCDD/DF congeners were measured
in all samples while concentrations of PCBs and dichloro-diphenyl-trichloroethane
(DDT) and related metabolites (DDXs) were determined in a subset of samples.
Collected eggs were opened around the girth with a scalpel blade. Contents were then
homogenized in a chemically cleaned Omni-mixer, lyophilized, and stored in clean jars
until analysis (I-CHEM brand, Rockwood, TN). Concentrations of PCDD/DP in eggs
were reported on a fresh weight basis adjusted to account for any desiccation during
incubation and storage. Adjusted fresh weight was calculated based on egg volume
(Hoyt 1979). The mass of egg contents was determined by subtracting the eggshell mass
at the time of processing from the adjusted fresh weight. Tissues collected from salvaged
nestlings were homogenized using a chemically cleaned Omni-mixer.

Residues were quantiﬁed in accordance with US. Environmental Protection Agency
(U.S. EPA) Method 8290/1668A with minor modiﬁcations (USEPA 1998). Analytical
methods have been detailed elsewhere (Fredricks et al. 2010b; (Seston et al. 2010b).

Brieﬂy, biotic matrices were homogenized with anhydrous sodium sulfate,s_piked with
known amounts of l3C-labeled analytes (as internal standards), and Soxhlet extracted.

Ten percent of the extract was removed for lipid content determination. Sample

128

 

 

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puriﬁcation included the following: treatment with concentrated sulfuric acid, silica gel,
sulfuric acid silica gel, acidic alumina and carbon column chromatography. Components
were analyzed using high-resolution gas chromatography/high-resolution mass
spectroscopy, a Hewlett-Packard 6890 GC (Agilent Technologies, Wilmington, DE)
connected to a MicroMass® high-resolution mass spectrometer (Waters Corporation,

Milford, MA). Losses of congeners during extraction were corrected based on recoveries
of l3C-labeled as outlined in EPA Method 8290/1668A. Quality control samples

generated during chemical analyses included laboratory method blanks, sample
processing blanks (equipment rinsate and atmospheric), matrix spike and matrix spike
duplicate pairs, unspiked sample replicates, and blind check samples. Results of method
and ﬁeld blank analyses indicated no systematic laboratory contamination issues.
Evaluation of the percent recovery and relative percent difference data for the matrix
spike and matrix spike duplicate samples and unspiked replicate samples were within
i30% at a rate of greater than 95% acceptability. Soxhlet extractions and instrumental

analyses were conducted at AsureQuality Ltd, Lower Hutt, New Zealand.

Statistical analyses

Total concentrations of the seventeen 2,3,7,8-substituted PCDD/DF congeners
(ZPCDD/DFS) are reported as the sum of all congeners (ng/kg wet weight (wet wt)).
Individual congeners that were less than the limit of quantiﬁcation were assigned a value
of half the sample method detection limit on a per sample basis (ND=0.5DL). Total
concentrations of 12 non- and mono-ortho-substituted PCB congeners are reported as the

sum of these congeners (ng/kg wet wt) (ZPCBs) for a subset of samples. Concentrations

129

 

 

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of TEQwHo-Avian (ng/kg wet wt) were calculated for both PCDD/DPS and PCBs by
summing the product of the concentration of each congener, multiplied by its avian.
TEFWH0.AV,an (van den Berg et al. 1998). The term DF-TEQSWHQ-AVian refers to
summation of the TEQS of individual PCDD and PCDF congeners while the term PCB-

TEQSw}{0-Avian refers to the TEQ from PCBs. Total TEQS refers to the summation of

both DF'TEQSWHO-AVian and PCB-TEQSWH0-AVian. In addition, concentrations of
dichloro-diphenyl—trichloroethane (2’,4’ and 4’,4’ isomers) and dichloro-diphenyl-
dichloroethylene (4’,4’) were measured in the same subset of samples as PCBs and are.
reported as the sum of the o, p and p, p isomers (ZDDXs; ug/kg wet wt).

Statistical analyses were performed using SAS® software (Release 9.1; SAS
Institute Inc., Cary, NC, USA). Prior to the use of parametric statistical procedures,
normality was evaluated using the Shapiro-Wilks test and the assumption of homogeneity
of variance was evaluated using Levene’s test. Values that were not normally distributed
were transformed using the natural log (ln) before statistical analyses. PROC GLM was
used to make comparisons for three or more locations. When signiﬁcant differences
among locations were indicated, the Tukey-Kramer test was used to make comparisons
between individual locations. PROC TTEST was used to make comparisons between
two groups. Differences were considered to be statistically signiﬁcant at p <0.05.

The experimental unit for residue concentrations in eggs and all nestling tissues
was considered to be individual GBH nests. An arithmetic mean was calculated for nests
from which multiple samples of the same tissue were collected. The only exception to
this was for examining correlations of residue concentrations among various nestling

tissues, where the individual nestling was considered the experimental unit. Tissue

130

correlations were run using Spearman’s rank correlation test. Concentrations of
ZPCDD/DF and TEQwHo-Avian in GBH nestling plasma and eggs collected from the
same nest were compared using PROC REG. Regression analysis was used to develop

an equation from which concentrations of EPCDD/DF and TEQWHOAvian in eggs could

be calculated from concentrations of ZPCDD/DF and TEQW}-{O-Avian measured in nestling
plasma. To better understand the potential distributions of concentrations of
DF-TEQWHOpMan and total TEQWHO-A,,ian in eggs and blood plasma of GBH nestlings, a

probabilistic modeling approach was used to portray the distributions. Probabilistic

models were developed as cumulative frequency distributions based on concentrations of
DF-TEQWH0.Avian and total TEQWHO-Avian in eggs and blood plasma of nestlings. The

mean and standard deviation of transformed concentrations values of each sample type
were used to generate a sample of 10,000 random concentration values based on a
lognorrnal distribution.

To investigate patterns of relative concentrations of individual congeners in GBH
blood plasma, principle component analysis (PCA) was performed with PROC FACTOR
using relative orderings of PCDD/DF congener concentrations normalized to the
ZPCDD/DF concentration. Only congeners that were present in 260% of the samples 1
were included as factors [TCDD, 2,3,7,8-tetrachlorodibenzofuran (TCDF), 1,2,3,7,8-
pentachlorodibenzo-p-dioxin (1 ,2,3 ,7,8-PeCDD), 1,2,3,7,8-pentachlorodibenzofuran
(1 ,2,3,7,8-PeCDF), 2,3 ,4,7,8-pentachlorodibenzofuran (2,3,4,7,8-PeCDF), and
octachlorodibenzo-p-dioxin (OCDD)]. Samples for which concentrations were less than

the limit of quantiﬁcation had the greatest inﬂuence on concentrations of ZPCDD/DF in

131

 

H

in

til:

it;

Pltl

GBH nestling plasma, with a mean difference between using ND=O and ND=O.5DL for

XPCDD/DF quantiﬁcation of 18%.

Selection of toxicity reference values

Literature-based no observed adverse effect concentrations (NOAECs) and lowest
observed adverse effect concentrations (LOAECs) were used in the determination of
hazard quotients (HQs) and subsequent assessment of risk. In the present study, tissue-
speciﬁc, egg, plasma, and adipose exposure-based toxicity reference values [(TRVs)
based on the same or similar compounds were identiﬁed from literature and compared to
measured site-speciﬁc GBH tissue concentrations. Resulting HQs are presented as a
range bounded by the LOAEC-associated HQs at the low end and the NOAEC-associated
HQs at the high end. It should be noted that the NOAEC and LOAEC associated HQs
are a function of the experimental design (dosing regime) and the actual threshold
concentration at which effects would be expected to occur lies somewhere within the
described range. It has recently been suggested that TRV selection should involve the
combination of multiple suitable studies into a dose-response curve to determine the most
accurate value (Allard et al. 2010), however an inadequate number of suitable studies
precluded the use of that approach here.

Toxicity reference values derived from both laboratory- and ﬁeld-based studies
were utilized to assess GBH eggs. The US. EPA previously developed egg—based TRVs
by taking the geometric mean of the effect concentrations in three double-crested
cormorant (Phalacrocorax auritus) egg-injection studies (Powellet al. 1997a; Powell et

al. 1997b; Powell et al. 1998; USEPA 2003). In each study, dioxin—like compounds were

132

injected into cormorant eggs that were artiﬁcially incubated until hatch. Based on a

measurement endpoint of embryo mortality, the resulting NOAECEGGLAB and

LOAECEGGLAB was 3670 and 11090 ng total TEQWHO-Avian /kg wet wt, respectively. In

addition to the aforementioned laboratory-based TRVs, there have been several ﬁeld
studies of GBH from which threshold concentrations of dioxin-like compounds in eggs

can be deduced. A GBH rookery with a relatively large mean concentration of 220 ng
total TEQWHO-Avian /kg wet wt (NOAEC) in eggs was not associated with the number of
successful nests and number of ﬂedglings per nest (Elliott et al. 2001). Great blue heron
eggs with a mean concentration of 360 ng total TEQWHO-Avian /kg wet wt (LOAEC)
which were collected and artiﬁcially incubated produced hatchlings with observed health

deﬁciencies, including decreased body weight, yolk-free body weight, stomach weight,

intestine weight, tibia length, and tibia dry, wet, and ash weights (Sanderson et al. 1994).

Thus a ﬁeld study based NOAECEGG-FIELD and LOAECEGG-FIELD Of 220 ng total

TEQWHO-Avian /kg wet wt and 360 ng total TEQ\VHQ.Avian /kg wet wt, respectively, was

selected.

Studies reporting concentrations of dioxin-like compounds in avian nestling tissues
and associated effects are limited, eSpecially for species that would be appropriate
comparison to GBH. After applying the 1998 WHO-Avian TEFs to the PCDD/DF and
DL-PCB data reported relative to the blood plasma of bald eagle nestlings in British
Columbia, a NOAEC of 2.9 ng total TEQw}_10_Avian /kg was estimated, based on

productivity of eagle pairs located in areas of lesser and greater contamination over a 5 yr

period (Elliott and Norstrom 1998). This value was selected as the N'OAECNg-pLASMA

133

for assessing GBH nestling blood plasma. Beyond egg and blood plasma data, one
additional study was available for comparison of other tissue concentrations. A group of
grey heron nestlings in the UK that were found to have bone abnormalities had adipose
concentrations of 640 ng total TEQWHO-Avian /kg wet wt (LOAECNS.AmpOgE) compared
to the group with no observed deformities that had a mean adipose concentration of 300
ng total TEQWHO-AVian /kg wet wt (NOAECN3-ADIPOSE) (Thompson et al. 2006). It

should be noted that these TEQS were primarily from dioxin-like PCBs.

Results

Adult plasma
Concentrations of ZPCDD/DF and DF-TEQWHO.A,,ian in GBH blood plasma were

compared spatially and by age category (Table 4.1). In general, concentrations of
ZPCDD/DF were greater in blood plasma of GBH captured in the SA than that of those
captured in the RA and in older birds (after hatch year; AHY) as compared to their
younger cohorts (hatch year; HY). The greatest mean difference between SA and RA
mean XPCDD/DF concentrations was for AHY-GBH (p<0.0001) while the spatial
difference in HY—GBH was less pronounced (p=0.0700). As mentioned above, the age of
the trapped individual was also associated with plasma concentrations. Concentrations of
ZPCDD/DF were consistently greater in older AHY-GBH than their HY-GBH cohorts,

irrespective of study location. This age class trend was statistically signiﬁcant in the SA
(p=o.0003). Comparisons based on DF-TEQWHO-Avian were consistent with those of

ZPCDD/DFS (Table 4.1).

134

Table 4.1. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF)
and 113me Ma, 3 in blood plasma of great blue herons collected during 2005-2007
as a result of trapping along the Chippewa and Tittabawassee River ﬂoodplains b,

Midland, MI, USA. Values c (pg/mL wet wt) are given as the geometric mean with
sample size in parentheses (n) over the 95% conﬁdence interval and range (min-max).

 

 

 

Age of Great Blue Heron
HY GBH Plasma AHY GBH Plasma
EPCDD/DF
Reference 3.5 (8) Aa d’ e 4.1 (7) A3
2.5—5.0 3.1—5.5
(1.8—5.9) (1.3—6.6)
Target 9.9 (16) Aa 18 (20) Bb
7.8—13 15-23
(5.2—19) (6.3-4l)
DF’TEQWHO-Avian a
Reference 1.2 (8) Aa 1.4 (7) Aa
074—19 093-22
(047—2. 1) (074—25)
Target 4.6 (16) Aa 9.6 (20) Bb
3.4—6.2 7.7—12
(2.0—10) (3.8-20)

 

a TEQWH0-Av,an were calculated based on the 1998 avian WHO TEF values

b Trapping was done in the river channel of the Pine, Chippewa, and Tittabawassee
Rivers as previously described in Seston et al. 2009.

c Values have been rounded and represent only two signiﬁcant ﬁgures

d Means identiﬁed with the same uppercase letter are not signiﬁcantly different
between locations at the p=0.05 level using Tukey-Kramer means separation test

6 Means identiﬁed with the same lowercase letter are not signiﬁcantly different between
age classes at the p==0.05 level using Tukey-Kramer means separation test

135

 

 

1C

01

CUE

([1?

 

Patterns of relative concentrations of individual congeners to ZPCDD/DF in GBH
blood plasma were also associated with the different sampling areas and age classes. The
PCA model used to characterize patterns of relative concentrations of congeners in GBH
blood plasma explained 75% of the total variance, with PCI explaining 62% and PC3
explaining 13%. Great blue heron plasma samples were separated by sampling area
along PC], which had a loading score of 0.95 for TCDD, with both HY— and AHY-GBH
within the RA having the greatest eigenvectors for PC]. The pattern of congeners in
blood plasma from the SA were separated by age along PC3, which had a loading score
of 0.996 for 2,3,4,7,8-PeCDF, with AHY-GBH having the greatest eigenvectors for PC3
(Figure 4.2). 2,3,4,7,8-PeCDF contributes a greater percentage to ZPCDD/DF in blood

plasma of AHY—GBH in the SA (31%) than in HY—GBH (12%) (Figure 4.3).

Nest-associated tissues

In general, tissues collected from nests within SA rookeries were similar in analyte
concentrations and relative contributions of individual PCDD/DP and dioxin-like PCB
congeners. However, a few tissue— and analyte-speciﬁc subtle differences and trends
were noted. Eggs and nestling plasma, adipose, liver and muscle had similar
concentrations and relative congener contributions among the rookeries in the TR SA.
Mean concentrations of all residues measured in GBH eggs were not signiﬁcantly
different among rookeries (Table 4.2). A single egg from F RE with an anomalously great
concentration was examined as an outlier, however removal of this egg from analysis did

not affect the outcome of rookery comparisons (data not presented).

136

Figure 4.2. Principle component analysis of PCDD/DF concentration congener proﬁles in
blood plasma of great blue heron collected during 2005-2007 from the Chippewa,
Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Symbols are labeled
by area (RA==Reference Area; SA=Study Area) and age class (HY=hatch year;
AHY=aftcr~hatch year). Ellipses indicate 95% conﬁdence intervals of each group.
Individual PCDD/DF congener loading scores for each principle component is depicted
in the inset. TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran;

TCDD = tetrachlorodibenzo-p-dioxin; PeC DD = pentachlorodibenzo-p-dioxin; OCDD =

octachlorodibenzo-p-dioxin.

137

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I ocor
[j TCDD
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I HxCDD

 

I OCDD

 

 

 

Percent of ZPCDD/DF

 

 

 

 

 

 

AHY

STUDY AREA
Figure 4.3. Pattern of percent mean EPCDD/DF congeners in blood plasma collected
from hatch year (HY) and after hatch year (AHY) great blue herons trapped within the

Tittabawassee River study area, Midland, MI, USA.

139

Table 4.2. Total concentrations of 2,3,7,8-substituted furan and dioxin (XPCDD/DF),

TEQWHO-Avian a , and co—contaminants in great blue heron eggs collected during 2006-
2007 from the Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA.

Valuesb (ng/kg .wet wt) are given as the geometric mean (n) over the 95% conﬁdence

intervals and range in parentheses. DDX values are reported in jig/kg wet wt.

 

 

 

Rookery
FRE SNWR CAS
GBH Egg

DF-TEQWHO-A.,,, a 44 (8) Ac 28 (8) A 27 (8) A
25—78 17—46 21—35

(17—150) (1799) (19—44)
PCB-TEQWHO-AViana 150 (8)A 110(2)d , 130(8)A
73—330 - 34—190
(23—360) (100—1 10) (62—250)
Total TEQWHOMan a 210 (8) A 130 (2) 150(8) A
1 10—390 — 1 10—220
(46~430) (120—140) (83—290)

spoon/or 67 (8) A 50 (8) A 45 (8) A

39-1 10 34—73 36—55

(27—210) (31—130) (31—65)

ZPCB 3.3x105(8)A 1.1x105(2) 2.3><105(8)A

1.5><105-7.5><105 - l.5><105—3.2><105
(6.1x104—9.0x105) (9.9x104—12x105) (1.7x105—4.9x105)

ZDDX 520 (8) A 320 (2) 680 (8) A
270—1000 - 420-1 100
(180-1600) (180—560) (320-1500)

 

a TEQWHO-AV;an were calculated based on the 1998 avian WHO TEF values.
b Values have been rounded and represent only two signiﬁcant ﬁgures.

c Means identiﬁed with the same letter are not signiﬁcantly different among locations
(across) at the p=0.05 level using Tukey-Kramer means separation test

d . . - . .
Range reported for srtes wrth 2 sarnples and were not included m the between locatlon
comparisons.

140

Mean concentrations of most analytes quantiﬁed in GBH nestling plasma were not

signiﬁcantly different among rookeries (Table 4.3). However, when ZPCDD/DFs were
normalized using TEFSWHO-Avians the mean DF-TEQWHOAvian concentration in nestling
plasma from FRE was signiﬁcantly greater than that from CAS. The signiﬁcant rookery
differences of DF-TEQWH0-Avian concentrations in nestling plasma, and the lack thereof

for ZPCDD/DF concentrations is largely due to a shift in the relative contributions of
individual congeners and not a change in ZPCDD/DFS. 2,3,4,7,8-PeCDF made up a
greater proportion of the ZPCDD/DFS in the upstream FRE nestling plasma while the
relative contribution of the less potent OCDD increased in the furthest downstream CAS

nestling plasma.
Concentrations of ZPCDD/DF, DF‘TEQ\VHQ-Avian, ZPCB, and PCB- TEQWHQ-AVian

in adipose, liver, and muscle of nestlings were consistently greatest at SNWR and least at
CAS (Table 4.3). Even though the relatively small number of samples collected
precluded determining statistical signiﬁcance, the ranges of residue concentrations by
rookery did not overlap for a number of tissues. In contrast to nestling plasma, the
relative contribution of individual congeners to ZPCDD/DF was similar among rookeries
and tissue types. In GBH eggs, a mean of 40% of the ZPCDD/DFS was contributed by
furan congeners, with. 2,3,4,7,8-PeCDF (23%) contributing the greatest proportion
(Figure 4.4). 2,3,4,7,8-PeCDF also contributed a signiﬁcant proportion of the
XPCDD/DFS in adipose (20%), liver (19%), and muscle (15%) tissues and blood plasma
(11%) of nestling GBH. One notable difference was the greater relative proportion of

OCDD in GBH nestling blood plasma compared to other nestling tissues.

141

Table 4.3. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF)
and TEQWHOAQan in great blue heron nestling tissuesa collected during 2006-2007
from the Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesb
are given as the geometric mean with sample size given in parentheses (72) over range
(min-max). PCDD/DP and PCB values are reported as ng/kg, wet wt in tissues and
pg/mL, wet wt in plasma. DDX values are reported in ug/kg wet wt in tissues and

ng/mL in plasma.

 

GBH NS Adipose
DF-TEQWHO-Aviand

PCB'TEQWHo-Aviand
EPCDD/DF

EPCB

ZDDX

GBH NS Liver
’ d
DF'TEQWHO-Avian

PCB'TEQWHO-Aviand
XPCDD/DF

ZPCB

ZDDX

GBH NS Muscle
DF‘TEQWHO-Aviand

PCB'TEQWHO-Aviand

EPCDD/DF

 

 

Rookery
FRE SNWR CAS
110(1) 140(3) 74(2)
(65—320) (63—86)
560 (1) 800 (3) , 250 (2)
(380-1400) (210—290)
170(1) 200 (3) 120(2)
(98—430) (100—150)
9.4x105(1) 1.1x106(3) 4.5x105(2)
(4.7XI05—1.7><106) (3.5x105—57x105)
2000(1) 1800(3) 1700(2)
(1500—2000) (1500—1900)
3.6 (1) 4.9 (3) 2.6 (2)
(2.0—7.4) (2.5-2.8)
13(1) 31(3) 6.5 (2)
(20—53) (5.9—7.1)
6.9 (1) 8.5 (3) 7.2 (2)
(4.4—12) (5.1—10)
l.6><104(l) 2.3x104 (3) 1.3x104(2)
(9.5x103—5.8x104) (1.3x104-l.4x104)
33 (1) 35(3) 40(2)
(27—55) (31—53)
3.6 (1) 6.8 (3) 2.1 (2)
(2.7—14) (1.6-2.7)
13(1) 30(3) 5.9 (2)
(1 1—82) (4.7—7.3)
6.7 (1 ) 14(3) 4.1 (2)
(9.3—22) (3.3—5.1)

142

Table 4.3 (cont’d)

 

 

ZPCB 2.0810‘1 (1) 3.9x10“(3) 1.3><104 (2)
(1.4x104—1.3x105) (1.0x104—1.7x104)
zoox 51 (1) 87 (3) 46 (2)
(51—220) (45—48)
GBH NS Plasma
l)l?-'I1~:Q,,,HO_,,m,d 1.5 (12) A 1.1 (4) AB 0.65 (7)13
(0.69—3.5) (070—20) (037-17)
PCB-T'EQWHO-A.,,,,“ 2.1 (4) 3.6 (4) N/Af
(0.90—3 .2) (2.8—62)
EPCDD/DF 4.9 (12) A 2.5 (4) A 3.8 (7) A
(1.8—13) (1.7—3.6) (1 .9—7.2)
ZPCB 3.3x 103 (4) 4.2><103 (4) N/A
(9.6X102—9.1><103) (3.1x103—5.4x103)
zoox 6.3 (4) 8.0 (4) N/A
(1.8—19) (5.9—1 1)

 

a Nest is considered the experimental unit. A mean value was used for nests from
which greater than one sample of the same tissue was collected.

b Values have been rounded and represent only two signiﬁcant ﬁgures

C Means identiﬁed with the same letter are not signiﬁcantly different among locations
(across) at the p=0.05 level using Tukey—Kramer means separation test

d TEQWH0-Avian were calculated based on the 1998 avian WHO TEF values

6 Range reported for sites with 2 samples and were not included in the between
location comparisons

fN/A = no samples collected from this location

143

 

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l . [j PeCDF
I HxCDF

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I HpCDD

 

 

 

 

 

 

 

 

 

 

 

  
 

 

 

 

 

       
 

   
  

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Egg Adipose Liver Muscle Plasma

Figure 4.4. Patterns of percent mean ZPCDD/DF congeners in great blue heron tissues
collected from the FRE (a), SNWR (b), and CAS (c) rookeries, located within the

Tittabawassee and Saginaw river ﬂoodplains, MI, USA.

144

Similarities were also present in the mean relative contribution of DF-TEQWH0-AVian
and PCB-TEQWHGAV,an to the total concentration of TEQWHO-Avian in GBH tissues. In
eggs and tissues of nestlings, 71—86% of the total concentration of TEQWHOAvian was
attributable to PCB-TEanamm (Supplemental, Data). PCB-126 contributed the

greatest proportion to PCB'TEQWHO-AVian (43%) in GBH eggs, with PCB-77 (26%) and -

81 (24%) also contributing a signiﬁcant proportion. In tissues of GBH nestlings,

PCB-77, PCB-81, and PCB-126 contributed nearly equal proportions of the PCB-
TEQWHO-Avian (30%, 36%, and 28%, respectively) (data not presented).

Correlations in concentrations of several residues in tissues of GBH nestlings were
observed. In individual nestlings, there was a signiﬁcant positive correlation between
concentrations of XPCDD/DF, DF-TEQWHGAW, 2,3,4,7,8-PeCDF, 2,3,7,8-TCDF,
2,3,7,8-TCDD, and EPCB in adipose and liver (n=11) (Table 4.4). Positive correlations
were also observed between concentrations of 2,3,7,8-TCDF and 2,3,7,8-TCDD in
plasma and adipose tissue (n=4) and between concentrations of ZPCDDfDF, DF-
TEQwHo-Avian, and 2,3,4,7,8-PeCDF plasma and liver tissue (n=4) (Table 4.4).

In addition to correlations between concentrations among tissues from the same
individual, a relationship was observed between concentrations in egg and nestling blood

plasma collected from the same nest, during the same nesting attempt; Concentrations of
ZPCDD/DF (r2=0.4933 p=0.0160) and DF-TEQWH0_AV;,,, (r2=0.4699 p=0.0199) in egg

and nestling blood plasma were correlated (Figure 4.5) (Eqns. 1 and 2). Predicted mean

145

 

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concentrations of ZPCDD/DF and DF-TEQWHO-AVian and ranges, along with

corresponding measured egg concentrations from each rookery, are listed in Table 4.5.

ZPCDD/DF: Egg (ng/kg, wet wt) = 19.87836[plasma(pg/mL,wet wt)] — 22.20420 (1)
DF-TEQWHO_Avian: Egg (ng/kg, wet wt) = 50.46363[plasma(pg/mL,wet wt)] —5.89248 (2)

Risk assessment

Predicted probabilistic distributions of expected cumulative percent frequencies based
on concentrations of DF-TEQS00H().,1\v,-a,1 and total TEQSWHO-Avian in eggs and. blood
plasma of nestling GBH were compared to selected TRVs. Predicted distributions of

concentrations of both DF-TEQSWHO-Avian and total TEQSWHO-A,,,-an in GBH eggs were
not greater than the NOAECEGGLAB or LOAECEGGLAB (Figure 4.6). Based on
concentrations of DF-TEQSWHGAW and total TEQSWHOAWM in GBH eggs, less than
1.0% and 34% of the cumulative frequency was greater than the NOAECEGGFIELD,
ICSpectively, and less than 1.0% and 10% of the cumulative frequency was greater than
the LOAECEGGFIELD, respectively. Based on the predicted distributions of
concentrations of DF-TEQSWHQ-AV;an and total TEQSWHO-Avian in blood plasma of
nestling GBH, 4% and 70% of the cumulative frequency was greater than the

NOAECNS-PLASMA-

148

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149

Figure 4.6. Modeled probabilistic distribution of expected cumulative percent
frequencies for great blue heron egg TEQWH(,)_mm concentrations ng/kg wet wt in site-
specific eggs collected from the river ﬂoodplains near Midland, Michigan in 2005-2007.
Concentrations of DF- TEQ“»’H()-AVian and total TEQW10A“;m indicated by dashed and

solid lines, respectively. NOAEC and LOAEC indicated by vertical bars.

150

Cumulative Frequency (%)

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Cumulative Frequency (%)

 

 

 

100
80
60 NOAEC (220 ng total TEQ/kg ww)
(Elliott et al. 2001)
40 LOAEC (360 ng total TEQ/kg ww)
‘ (Elliott et al. 2001)
20
0 1 l I ﬂ r j
400 600 800 1000
100
3 (B)
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1g —- Total TEQ
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(Elliott et al.1998)
0 I l' r I j l I l ] l T' . V
0 ‘ 5 10 15 20 25

Concentration (ng/kg w)

151

Hazard quotients were calculated by comparing measured tissue concentrations to the

selected TRVs. Upper 95% conﬁdence level (UCL; of the geometric mean)
concentrations of DF-TEQSWHO-AVian in GBH eggs (n=24) collected from within the SA
were not greater than either of the egg-based NOAEC or LOAEC TRVs, resulting in HQs

less than 1.0. Conversely, 95% UCL concentrations of total TEQSwHo-Avian in GBH eggs

(n=l8) did marginally exceed the NOAECEGGMELD TRV, resulting in 3 HQ of
approximately 1.0 (Figure 4.7).

Similar trends in HQs were observed for GBH nestling blood plasma as for GBH

eggs. The 95% UCL DF—TEstnOmm concentrations in GBH nestling blood plasma
did not exceed the NOAECNSPLASMA for any rookery, thus resulting in less than 1.0.

When based on total TEQstt/m).m.ian and the NOAECN5-pLASMA, HQs calculated from
95% UCL concentrations were greater than 1.0 for blood plasma of nestlings from FRE
and SNWR. Concentrations of total TEQSWHO-Avian in blood plasma ’of nestlings were
not available from CAS for comparison (Figure 4.7).

Comparisons of the geometric mean and 95% UCL concentrations of DF-TEQSWHO-
Avian in adipose of nestling GBH to both the NOAECN3-AD|POSE and LOAECNS-AD1POSE
resulted in HQs less than 1.0 (Figure 4.7). Total TEQSWHO-Avian geometric mean and
95% UCL concentrations exceeded both the NOAECNSADIPOSE and LOAECNS-ADIPOSE,

resulting in HQs greater than 1.0. Although concentrations of total TEQSWHQ.AVian did

152

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153

exceed the associated TRVs, none of the bone abnormalities observed in the study upon

which the TRV was based were observed in any GBH nestlings in the present study.

Population condition

Measures of population condition were similar among all studied rookeries
(Table 4.6). Mean clutch sizes for GBH from FRE, SNWR, and CAS were not
signiﬁcantly different. Additionally, the mean number of nestlings per successful nest

was not signiﬁcantly different among rookeries. The number of nestlings per nest was
not significantly correlated with concentrations of total TEQSWH0-Avian in eggs or blood

plasma, adipose, liver, or muscle of GBH nestlings.

Discussion
Tissue-based exposure

Spatial trends in the concentrations of ZPCDD/DFS and the relative contribution of
PCDDS and PCDFs were consistent between adult GBH blood plasma and prey items
(Seston et al. 2010b). Concentrations of highly lipophilic compounds in blood plasma
are largely inﬂuenced by the absorption of contaminants in recently ingested food items
from the gastrointestinal tract. Differences in concentrations of residues in the blood
plasma of both HY- and AHY-GBH trapped in the RA as compared to those trapped in
the SA are in agreement with the conclusion that GBH were consistently foraging in the
area of the study site in which they were trapped. The greater mean concentrations of

ZPCDD/DFS in blood plasma from GBH trapped in the SA than that in blood plasma

154

Table 4.6. Reproductive parameters for great blue heron rookeries located in the
Tittabawassee and Shiawassee river ﬂoodplains from 2006-2007. Values are given as the
arithmetic mean 5: 1 SD over the sample size given in parentheses (n).

 

 

 

Rookery
FRE SNWR CAS
Clutch size a 3.8:t1.0 Ab 4.0i1.6 A 3.92t1.4 A
(22) (9) (23)
Nestlings / successful nest 0' 2.7i0.86 A 250.1 A 3.4012 A
(21) (13) (8)

 

a Number of eggs per nest
b Means identified with the same letter are not signiﬁcantly (p=0.05) among rookeries
(across) using the Tukey-Kramer means separation test.

C Nestlings / successful nest calculated as the number of chicks present at 4-5 weeks of
age.

155

from GBH trapped in the RA is spatially consistent with predicted average daily doses
(ADDpot) based on site-speciﬁc prey item concentrations. The ADDpot in the RA and

SA, based on an assumption of 100% site use, were 0.97 and 44 to 52 ng ZPCDD/DF/kg
bw/d, respectively (Seston et al. 2010b). The relative contributions of PCDDS and
PCDFs to EPCDD/DFS in GBH blood plasma and prey items also exhibited consistent
spatial trends. Great blue heron blood plasma from the SA had a greater relative
contribution of PCDFs to ZPCDD/DFS compared to those from the RA. This trend was
also seen in ﬁsh, mammals, and other birds in the RA and SA (Coeﬁeld et al. 2010b;
Coeﬁeld et al. 2010a; Fredricks et al. 2010a; Fredricks et al. 2010b; Seston et al. 2010b;
Seston et al. 2010a; Zwiemik et al. 2008b).

Concentrations of ZPCDD/DFS in GBH blood plasma were associated with the age of
individual. The association between concentrations of ZPCDD/DFS in GBH blood
plasma and age, in which nestling < HY-GBH < AHY-GBH is similar to the trend
between ZPCDD/DF concentrations in blood plasma of great horned owl nestlings and
adults along the TR (Coeﬁeld et al. 2010b). This phenomenon is consistent with
developmental dilution in the younger birds (Kunisue et al. 2006) and duration of
exposure. As a long-lived species that can defend its foraging territories over multiple
years, AHY-GBH are potentially exposed longer than nestling or HY-GBH.

Patterns of relative concentrations of individual congeners in blood plasma from the
more highly exposed GBH trapped in the SA were also associated with age class. In the
SA, both HY- and AHY-GBH blood plasma congener proﬁles were dominated by furans
but there was a shift in predominant congeners between age classes. 2,3,4,7,8-PeCDF

was the predominant congener in blood plasma from AHY-GBH whereas 2,3,7,8-TCDF

156

and 2,3,4,7,8-PeCDF were nearly equal in HY-GBH. The shift in predominance between
these two congeners by age class has previously been observed in common cormorants
and albatross (Kubota ét al. 2004; Kunisue et al. 2006). Field studies have reported
negligible bioaccumulation of TCDF from prey items in Forster’s terns and herring gulls
(Braune and Norstrom 1989; Kubiak et a1. 1989). This may be due to preferential
metabolism of 2,3,7,8-TCDF, as a number of studies have reported in various avian
speciesexposed to mixtures of AhR-active compounds (Elliott et al. 1996; Kubota et al.
2005; Senthil Kumar et al. 2002). Data on avian toxicokinetics from controlled
laboratory studies are limited, but mammalian studies have shown the rate of metabolism
of 2,3,7,8-TCDF to be elevated with increased concentrations of dioxin-like compounds
and the subsequent induction of cytochrome P450 1A1 and/or 1A2 while 2,3,4,7,8-
PeCDF is preferentially sequestered in the liver (van den Berg er al. 1994b; Zwiemik et
al. 2008a). Due to these differences in persistence, it is unsurprising the older GBH in
the SA would have a relatively greater proportion of PCDD/DF in their body due to the
greater bioaccumulation of 2,3,4,7,8-PeCDF compared to younger GBH.

Although spatial trends were observed in the relative congener contributions in blood
plasma of adult GBH, the congener patterns among other tissues were similar among
rookeries. As a migratory species, there is potential for GBH to accumulate contaminants
while on their wintering grounds and then later deposit these compounds into eggs (Yates
et al. 2010). The small variability among congener patterns in the various tissues within
and among the three SA rookeries is consistent with adult GBH from all the rookeries
foraging in areas with a common source. The similarity of congener proﬁles in tissues of

GBH and those in site-speciﬁc prey items strongly suggests that the majority of the

157

resident GBH PCDD/DP exposure is site-specific and not from wintering grounds.
Moreover, nestlings are limited to resources collected proximal to the ‘ nest by parent
birds. Thus the similarity of congener proﬁles in both eggs and nestling tissues from
rookeries within the TR ﬂoodplain again suggests the contaminants in each are site-
speciﬁc and not acquired from wintering grounds.

Albeit of site-speciﬁc origin, the concentrations of ZPCDD/DFs found in GBH
tissues were less than expected based on those measured in a second site-speciﬁc avian
piscivore. The belted kingﬁsher (Ceryle alcyon; BKF) was assessed within the TR
ﬂoodplain using similar exposure studies (Seston et al. 2010a). Mean concentrations of
ZPCDD/DFS were greater in both eggs and nestling tissues of belted kingﬁsher compared
to those of GBH. Despite the difference in concentrations, the relative contributions of
individual PCDD/DP congeners were similar between the two species, indicating they are
both being exposed to the same PCDD/DF mixture. The observed difference in tissue
concentrations is likely a result of variation in metabolic rate or the proportion of diet
collected from the study area due to differences in foraging range size. During the
breeding season, GBH may travel a mean distance of 3.2 to 6.4 km from their rookery to
foraging grounds (Dowd and Flake 1985; Marion 1989; Thompson 1978) whereas BKF

' forage 0.92 to 2.9 km proximal to their nest burrow (Davis 1982; Mazeika et al. 2006).

Risk characterization
The risk of site-speciﬁc adverse effects can be assessed by the use of probabilistic

modeling of site-speciﬁc concentrations and comparison of the frequency distributions to

158

appropriate TRVs. The most conservative estimate of risk for GBH eggs was the

comparison of concentrations of total TEQ to the NOAECEGGHELD and LOAEngg-
FIELD- Compared to thC NOAECEGG-FIELD and LOAECEGG-FIELDa 34% and 10% Of the

predicted frequency distribution of concentrations of total TEQSWH0_AVian in GBH eggs

exceeded these values, respectively. Importantly, the actual effect threshold for

individuals is likely between the established no-and lowest-effect TRVs. When

compared to the NOAECEGGLAB or LOAECEGGLAB, 0% of the frequency distribution

for concentrations of DF-TEQSWHO-A,m or total TEQSWHO_AVian in GBH eggs exceeded

either value. Based on conservatively selected egg-based TRVs (see Uncertainty

assessment), the potential for effects on individual GBH is minimal. Based on the
predicted distributions of concentrations of DF-TEQSWH0_AVian and total TEQSWHO-Avian
in blood plasma of nestling GBH, 4% and 70% of the cumulative frequency was greater
than the NOAECN3_pLASMA. However, there is no LOAECN3-pLASMA for comparison, so

it is not clear where the effect threshold occurs.
The degree of concern associated with the potential presence of site-speciﬁc adverse
effects can also be assessed by use of the traditional point estimate HQ approach.

Despite the conservative nature of the selected TRVs, egg—based HQs were less than or
equal to 1.0 when based on NOAECS and the 95% UCL concentrations of DF-TEQSWHO.
AVian and total TEQSwHo—Avian, which is consistent with a conclusion that adverse effects
would not be expected. For SA rookeries, HQs for blood plasma of nestling GBH were

less than 1.0 based on the 95% UCL concentrations of DF-TEQSWH0-Avian and marginally

159

greater than 1.0 when based on 95% UCL concentrations of total TEst}.10_Avian, again
indicating a minimal expectation for risk of adverse effects.

Hazard-quotients calculated from 95% UCL concentrations of total TEQSWHO-Avian in
GBH nestling adipose tissue did exceed 1.0 based on both the NOAECNSADIPOSE and

LOAECNSADIPOSE, but did not exceed 10. This would suggest there is the potential for

individual GBH nestlings to exhibit the bone abnormalities observed in the grey heron
nestlings (Thompson et al. 2006). This difference in apparent sensitivity may be due to
site-speciﬁc differences in the environmental mixture of PCDD/DPS and dioxin-like
PCBs present in each study area. Additionally, the potential role of additional unknown
stressors, such as co-contaminants which may be associated with the bone abnormality
observed in the grey heron nestlings, must also be taken into consideration. None of the
salvaged nestlings or other nestlings monitored along the TR exhibited this deformity.

In addition to dioxin-like compounds, the potential for adverse effects of site-speciﬁc
ZDDX concentrations was assessed. GBH appear to be relatively insensitive to the
effects of DDE contamination, as an egg with a concentration of 78 mg/kg successfully
pipped (Vermeer and Reynolds 1970). Other studies have found mean DDE
concentrations in eggs ranging from 0.086 to 16 mg/kg and reported no adverse impacts
on breeding success (Blus et al. 1980; Harris et al. 2003; Laporte 1982; Thomas and
Anthony 1999). These observations suggest that egg ZDDX concentrations reported in
the present study would not cause adverse effects on breeding success of GBH along the
TR ﬂoodplain.

The prediction of minimal potential for adverse effects from the aforementioned

tissue-based assessments is consistent with site-speciﬁc measures of population

160

 

 

it

Hi

condition. The clutch size of GBH is reported to range from 2 to 6 eggs per nest (Butler
1992), with rookery averages ranging from 2.8 to 4.4 eggs per nest (Elliott et al. 1989;
Straub et a1. 2007; Thomas and Anthony 1999). Clutch sizes observed in nests within the
TR ﬂoodplain are within these reported ranges. Literature values for the number of
nestlings per successful nest range from rookery averages of 3.09 to 3.22 nestlings per

nest (Straub et al. 2007), which is similar to values observed along the TR.

Uncertainty assessment

The greatest limiting factor in the assessment of risk to avian species exposed to
dioxin-compounds is the lack of studies reporting effect threshold concentrations,
especially for wild species. Thus, TRV selection oﬁen introduces signiﬁcant uncertainty
to the risk assessment, especially in cases where comprehensive site-speciﬁc data sets
deScribe exposure with great certainty. HQs greater than 1.0 are indicative of exposures
that exceed the threshold for adverse effects and that there is the potential for effects to
occur. However, such assessments are conservative when extrapolating from individuals
to populations, as they are based on assumptions of maximal exposure and do not account
for compensatory mechanisms associated with resource availability (F airbrother 2001).
Therefore, HQ values greater than 1.0 do not necessarily translate into population-level or
ecologically relevant adverse effects (Blankenship et al. 2008).

Each TRV selected to assess the risk to GBH based on egg concentrations present
along the TR ﬂoodplain has associated uncertainties. The values based on the double-
crested cormorant egg injection studies were considered to be appropriate for use in the

present ‘study because they were derived from a related avian piscivore which has a

161

similar sensitivity to dioxin-like compounds as GBH (Sanderson and Bellward 1995).
Furthermore, both the double-crested cormorant and GBH have similar AhR ligand
binding domain constructs, which suggests the two species will respond similarly to
dioxin-like compounds (Head et a1. 2008; Karchner et al. 2006). Although the species
appear to be similar in sensitivity to dioxin-like compounds, the route of exposure in the
aforementioned double-crested cormorant studies and the GBH being assessed along the
TR is different. Residues in the GBH eggs are maternally deposited compared to being
artiﬁcially introduced as in the study conducted with the double-crested cormorant eggs.
This difference in route of exposure adds some uncertainty when using this TRV in the
risk assessment (Heinz et al. 2009; Hoffman et al. 1996). Selected TRVs based on ﬁeld
studies of GBH reduce the uncertainty associated with interspecies comparisons and
differences routes of exposure, but add uncertainty associated with confounding ,
environmental factors such as weather events, cyclical population trends of both
predators and prey, ecological relevance of observed endpoints, and the potential for
unknown co-contaminants. In general, the uncertainties associated with ﬁeld studies will
tend to add conservative bias to the selected TRV if the same co-contaminant issues are
not present at the assessment site.

Studies conducted on various raptor species have measured concentrations of dioxin-

like compounds in blood plasma, but none have reported a concentration at which
reproduction was impaired. A NOAEC of 0.8 ng PCB-TEQSWHO-AV;an /kg nestling

plasma was reported based on reproductive parameters, for great horned owl (GHO)
nestlings exposed to PCBs in Michigan (Strause et al. 2007a). Within the same study site

as the GBH being assessed in the present study, GHO nestlings and adults had 95% UCL

162

plasma concentrations of 2.6 ng DF—TEQ/kg and 14 ng DF-TEQ/kg, respectively

(Coeﬁeld et al. 2010b). Since no reproductive impairments were observed, these values

are also considered to be plasma-based NOAECs. The bald eagle NOAECN3-pLASMA of

2.9 ng total TEQSWH0-Avian /kg selected for assessing GBH nestlings in the present study

was the greatest concentration among those selected as potentially appropriate for use in

the present assessment, thus theoretically being closest to the effect concentration.
There is additional uncertainty associated with the application of TEFSWHO-Avian to

concentrations of PCDD/DPS. Recent studies have found that TCDD, TCDF, and
2,3,4,7,8-PeCDF may not be equipotent in birds, as was concluded by the most recent
assessment by the WHO (Cohen-Bamhouse etal. 2010; Hervé et al. 2010b; Herve' et al.
2010a; van den Berg et al. 1998). Based on the results of egg injection studies in
Japanese quail (Cohen-Bamhouse et al. 2010), a species that falls into the same broad
category of sensitivity to dioxin-like compounds as GBH, TCDF and 2,3,4,7,8-PeCDF
were found to be 2- and 6-fold more potent, respectively, than TCDD at causing
embryolethality. By) applying these relative potencies to the concentrations observed in
eggs of GBH collected in the present study, concentrations expressed as DF-TEQs
increased approximately 3-fold. Despite the increase in DF-TEQs seen by using these
revised relative potencies for TCDF and 2,3,4,7,8-PeCDF, minimal potential for adverse

effects would be expected based on comparisons to selected TRVs.

Plasma to egg relationship
In addition to assessing the risk that site-speciﬁc concentrations of PCDD/DFs pose

to resident GBH along the TR, a relationship between concentrations of these compounds

163

in egg and nestling plasma was developed. Concentrations in eggs are often regarded as
the most important, as embryonic development and hatching success are considered some
of the most sensitive endpoints for PCDD/DPS, PCBs, and many other environmental
contaminants. The nest is more susceptible to parental abandonment as a result of
disturbance while it contains eggs than later in the nesting season. For GBH, incubating
adults from nearby nests or potentially the entire rookery will ﬂush from the nest during
egg collection, leaving the eggs susceptible to drops in temperature and predation. The
development of a relationship between concentrations of PCDD/DFs in nestling plasma
and eggs would allow researchers to limit time in the rookery to after hatch out when
rookery disturbance may have a lesser impact on nest abandonment and/or success.
Additionally, nestling plasma sampling is non-destructive, which is beneﬁcial when
working with endangered or threatened species. Not collecting an egg also eliminates the
need to account for sampling in hatch success calculations. Predicting concentrations of
PCDD/DPS in eggs from those in nestling plasma is an effective, non-destructive method.
Use of this method does presuppose that foraging habits of adults remain similar
throughout the nesting cycle, so that adults during egg formation are exposed similarly to
nestlings. This type of relationship has also previously been developed for great horned

owls and bald eagles (Strause et al. 2007b).

Conclusions
Great blue heron nesting within the TR ﬂoodplain are exposed to greater
concentrations of PCDD/DPS compared to those from associated reference areas.

Comparison of observed concentrations of PCDD/DFs and dioxin-like PCBs in tissues of

164

 

 

an

 

GBH nesting within the TR ﬂoodplain to threshold effect concentrations suggest there is
minimal to no risk of adverse effects present. Minimal to no risk of adverse effects was
also predicted from a dietary exposure assessment of GBH in this same system (Seston et
al. 2010b). Conclusions from each of these lines of evidence are in agreement with site-

speciﬁc measures of GBH population condition.

Acknowledgements

The authors would like to thank all staff and students of the Michigan State
University — Wildlife Toxicology Laboratory ﬁeld crew; namely M. Nadeau, W. Folland,
D. Hamman, and S. Plautz for their tree-climbing abilities, along with E. Koppel, J.
Moore, L. Williams, and C. Bartrem. We gratefully acknowledge M. Fales for his design
and fabrication of specialized equipment which was pivotal to the success of this
research. Additionally, we would like to recognize P. Bradley, M. Kramer, and N. Ikeda
for their assistance in the laboratory. The following people were key to the success of the
project, J. Dastyek and S. Kahl of the United States Fish and Wildlife Service —
Shiawassee National Wildlife Refuge for their assistance and access to the refuge;
Saginaw County Parks and Recreation Commission for access to Imerman Park;
Tittabawassee Township Park Rangers for access to Tittabawassee Township Park and
Freeland Festival Park; and Michael Bishop of Alma College for guidance as our Master
Bander. We would also like to acknowledge the greater than 50 cooperating landowners
throughout the study area who have granted us access to their property, making our

research possible. Thanks also to S. Roark for reviewing drafts of this manuscript.

165

Funding was provided through an unrestricted grant from The Dow Chemical Company,

Midland, Michigan to J .P. Giesy and M.J. Zwiemik of Michigan State University.

Animal Use

All aspects of this study that involved the use of animals were conducted using the
most humane means possible. To achieve that objective, all aspects of the study were
performed following standard operation procedures (GBH adult handling 03/05-036—00;
GBH nest monitoring 05/07-066—00; Field studies in support of TR ERA 03/04-042-00;
Protocol for ﬁsh sampling 03/04-043-00) approved by Michigan State University’s
Institutional Animal Care and Use Committee (IACUC). All of the necessary state and
federal approvals and permits (Michigan Department of Natural Resources Scientiﬁc
Collection Permit SC1254 for GBH/SC permit for ﬁsh (Zwiemik)/SC permit for
amphibians (Zwiemik); USFWS Migratory Bird Scientiﬁc Collection Permit
MB1000062-0; and subperrnitted under US Department of the Interior Federal Banding

Permit 22926) are on ﬁle at MSU-WTL.

166

Supplemental Information

Table 4.7. Concentrations of 2378-TCDD equivalents (TEQsa) in eggs and nestling
tissues of great blue herons collected during 2006-2007 from rookeries within the
Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg wet
wt) are given as geometric mean (sample size) over the 95% conﬁdence interval.

 

 

Eggs Plasma Adipose Liver Muscle

PCDD‘TEQWHO-AvianB

16 (24) 0.33 (23) 43 (6) 1.3 (6) 1.5 (6)

12-20 0.25—0.43 29~63 083—21 074—32
I)CDF’II‘EQWHO-Avianc

16(24) 0.72 (23) 66 (6) 2.4 (6) 2.4 (6)

13—21 0.52-0.99 31—140 1.2—4.6 0.98-61
non-ortho PCB-TEQWHQ-AViand

130 (18) 2.6 (8) 480(6) 16(6) 14(6)

90-180 1.7—4.0 220—1000 6.6-40 4.7—43
mono-ortho PC B-TEQWHO- Aviane

9.8 (18) 0.14 (8) 29(6) 0.71 (6) 0.93 (6)

6.9“14 (1083-024 15—57 0.38’13 035—25
TOW TEQWHO-Avian

170 (13) 4.2 (8) 620(6) 21 (6) 19(6)

130-230 2.9—6.1 300-1300 9.6-46 6.8-55

 

a TEQWH0-AVian were calculated based on the 1998 avian WHO TEF values

b PCDD-TEQWHO-AV;an'—= summation of the TEQS of individual PCDD congeners

C PCDF-TEQWH0-Avian= summation of the TEQS of individual PCDF congeners

d non—ortho PCB-TEQWH0_A,.ian= summation of the TEQS of individual non-ortho
substituted PCB congeners

e mono-ortho PCB-TEQWHO_A,.ia,,= summation of the TEQS of individual mono-ortho
substituted PCB congeners

167

Table 4.8. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners
in blood plasma of great blue herons collected during 2005-2007 as a result of
trapping within the Chippewa and Tittabawassee River ﬂoodplains, Midland, MI,

USA. Valuesa (ng/kg wet wt) are given as the arithmetic meanb i 1 SD over the

 

 

 

 

range.
Reference Areas Study Areas
HY AHY HY AHY
Chemicalc n=8 n=7 n=16 n=20
2378-TCDF 02700.14 03900.17 2701.9 1.6013
0.095—0.48 0.21—0.61 0.76-7.3 0.12—5.9
2ND 3ND
23478-PeCDF 02700.15 0.340018 1.6013 6.4041
0.13—0.48 0.15—0.56 040-48 075—13
2ND
12378-PeCDF 0.094 09600.79 0.710053
0.10-0.14 0.16—2.9 0.12—1.7
7ND 5ND 4ND
234678-HxCDF 0.38 0.270020
0084—053
8ND 7ND 15ND 14ND
123789-HxCDF 8ND 7ND 16ND 20ND
123678-HxCDF 0.16 0.260013 0.420017
0.10—0.44 0.20—0.91
8ND 6ND IOND 5ND
123478-HxCDF 0.29 06700.46 1.10056
0.12—1.5 021—24
8ND 6ND 4ND
1234789-HpCDF 0-15
8ND 7ND 16ND 19ND
1234678-HpCDF 03100.19 05600.41
0089—064 0.1 1—1.0
8ND 7ND 7ND 14ND
12346789-OCDF 0.60 1-0i0-43
0.46—1.5
8ND 7ND 15ND 16ND
2378-TCDD 0.500029 0.620032 0.520031 1.40058
022—1 .0 023—1.1 019—13 035—21
1ND 1ND

168

Table 4.8 (cont’d)

 

123 78-PeCDD

123 78 9-HxCDD

123678-HxCDD

123478-HXCDD

l234678-HpCDD

12346789-OCDD

0.18-0.53 0.14-0.52
1ND 3ND
8ND 7ND

0.3 500. l 1
0.25-0.59 0.20-0.44
6ND 3ND
8ND 7ND
0.5000085 0.27
0.45-0.60
5ND 6ND .
1.20047 1.10080
076-1 .8 046-25
4ND 1ND

0.13—0.76
3ND

0.12—0. 15
14ND
04300.23
0.18—0.75
7ND

0.18—0.33
14ND
05300.25
0.25—0.94
5ND
2301.6
0.62—7.0
1ND

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection

arithmetic means

C TCDF = tetrachlorodibenzoﬁrran; P

hexachlorodibenzofuran; HpCDF =
octachlorodibenzofuran; TCDD =
pentachlorodibenzo
heptachlorodibenzo

169

03600.14 0.370017 0.380022 1.20071

0.25-3.2
1ND
0.49

19ND
1.100.64
0.30-2.6
3ND
0.390021
0.16-0.64
15ND
0.670042

' 0.12—1.5

9ND
5.1062
0.32—26
1ND

treated as zero in the calculation of

eCDF = pentachlorodibenzofuran; HxCDF=

heptachlorodibenzofuran;
tetrachlorodibenzo-p—dioxin; PeCDD =
-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

OCDF =

Table 4.9. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners
in eggs of great blue herons collected during 2006-2007 from rookeries within the
Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesa

(ng/kg wet wt) are given as the arithmetic meanb 0 1 SD over the range.

 

 

FRE SNWR CAS
Chemicalc n=8 n=8 n=8
2378-TCDF 3203.5 5202.8 3.1018
0.30-10 0.44—9.2 1.1—5.8
23478-PeCDF 20015 14016 1105.0
4.2—48 3.7-51 5.6—20
12378-PeCDF 04900.47 1.000.62 0.420037
007-12 047—19 0.18—1.3
2ND 2ND
234678-HxCDF 08700.71 08300.57 05500.27
0.30—2.1 O.39—1.8 0.28—0.98
3ND
123780ch01? 8ND 8ND 8ND
123678-HxCDF 1,901.8 1.300.87 1.50082
043—44 0.62-3.3 0.75—3.1
123478-HxCDF 3.6065 1.7016 1.10058
031—20 0.55—5.'1 053—20
1ND
1234789-HpCDF 0.220020
0.10-0.46
5ND 8ND 8ND
1234678-HpCDF 0.460036 0.710041 03100.13
0.20—1.3 0.27—1.1 0.12—0.53
5ND
12346789-OCDF 0.1700098
0085—034 0085—045 0071—012
3ND 6ND 6ND
2378-TCDD 18019 8,707.8 7.0027
5.9—65 3.2-27 4.3—8.2
l2378-PeCDD 1207.6 5603.3 6201.6
4.2—28 3.1—13 4.2—8.2
123789-HxCDD 0.650029 0.780035 0.580021
0.27—1.1 041—13 0.30—0.93
4ND
123678-HxCDD 9,406.9 5.6029 6.5019

170

Table 4.9 (cont’d)

 

123478-HxCDD

1234678-HpCDD

12346789-OCDD

2.7—23
1.9010
0.69—3.9
2.101.8
0.88—67

5103.7
0.90—12

2.5-10
1.100.51
0.50—1.8
2401.4

1.3—5.3

1ND

6.805.]

2.1—17

3.9—9.1
1.300.80
0.55—2.9
1.400.82
0.67—3.3

4502.1
1.3-6.9

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of arithmetic

means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=

hexachlorodibenzofuran;
octachlorodibenzofuran;

pentachlorodibenzo-p-dioxin; HxCDD
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

HpCDF
TCDD

heptachlorodibenzofuran;
tetrachlorodibenzo-p-dioxin;
hexachlorodibenzo-p-dioxin; HpCDD

171

OCDF
PeCDD

Table 4.10. Concentrations of selected co
collected during 2006-2007 from rookerie
Saginaw River ﬂoodplains, Midland, MI,
the arithmetic mean :1: 1 SD over the range.

 

Chemicalb

 

 

 

PCB 77

PCB 81

PCB 126

PCB 169

PCB 105

PCB 114

PCB 118

PCB 123

PCB 156

PCB 157

PCB 167

PCB 189

zntDDTb

rnuDDEC

4, 4’-DDT

05900.45
0.055~l .4
05100.36
0.030—1 .1
1.00065
015—1.9
01000067
0026—022
86060
1 1—180
9.4003
1.0—19
3000230
37—610
6804.4
078—14
34022
5.7—69
7.605.]
LLJS
15010
3.3—31
4.0026
0.65—7.6
07701.6

0.050-4.1
2ND
6700480

180—1600
9.6019
023—55

0.78—0.96

0.30-0.33

0.26—0.30

0023—0029

23—27

1.9-2.7

56-74

1.8—2.1

6.8-7.6

1.4—1.9

3.3-4.3

0.79-0.81

0.090—0.30

180-560

0.77-1.4

172

-contaminants in eggs of great blue herons
3 within the Chippewa, Tittabawassee, and

USA. Valuesa (pg/kg wet wt) are given as

F RE SN WR CAS
n=8 n=2 n=8

0.640043
0093—1 .6
0300016
0086—054
0.690034
040—1.3
007400.048
0.038—0. 1 7
49014
32—67
5.6024
2.9—11
150080
87—340
4.7022
2.8—9.3
2207.9
14—37
5001.9
3.0—8.5
1002.1
7.5—14
2.701 .1
1.4—5.1
0.1100071

0.016-0.23

5000200

320-820
0.660068
0.15-2.0

Table 4.10 (cont’d)

 

a Values have been rounded and represent only two signiﬁcant ﬁgures
b DDT = dichloro-diphenyl-trichloroethane
c DDE = dichloro-diphenyl-dichloroethylene

173

Table 4.11. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin
congeners in blood plasma of great blue herons nestlings collected during 2006-2007
from rookeries within the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains,

Midland, MI, USA. Valuesa (ng/kg wet wt) are given as the arithmetic meanb 0 1 SD

over the range.

 

 

FRE SNWR CAS
Chemicalc "=1 2 n=4 n=7
2378-TCDF 06700.64 1.1015 02600.19

0080—22 0.18—3.4 0.11—0.74

2ND 1ND
23478-PeCDF 05700.45 08401.1 02700.22

4.2—48 021—25 0.11—0.74
12378-PeCDF 03600.30 0.360045 01500.12

0025—095 0072-10 0073—033

3ND 3ND
234678-HxCDF 0044000091 0.032

0037—0054

9ND 3ND 7ND
123789-HxCDF 12ND 4ND 7ND
123678-HxCDF 009700.030 01400.12 0.079

0064-013 0074—032

7ND 6ND
123470ch1313 02100.14 01800.21 0.1400085

0057—053 0053—042 0055—022

2ND 1ND 4ND
1234789-HpCDF 12ND 4ND 7ND
1234678-HpCDF 0.1200070 004800019 0.1 100.032

0033—025 0034-0069 0062—013

2ND 1ND 2ND
12346789-OCDF 0.180010 0091

0.1 1—026 .

IOND 4ND 6ND
2378-TCDD 04000.32 04200.46 02000.12

0,11-12 0.16—1.1 0.12—0.34

2ND 4ND
12378-PeCDD 02600.15 03700-44

0.16-059 0.11-0.88 0079—031

174

Table 4.11 (cont’d)

 

5ND 1ND 5ND
123 789-HXCDD 12ND 4ND 7ND
123678-HxCDD 0.170010 0.300021
0.068-0.31 0.11-0.77 0.060-0.26
6ND 5ND
123478-HXCDD 0.085 0.048
1 1ND 3ND 7ND
1234678-HpCDD 0.20i0.095 0.15001 1 0.1700036
0069-032 0081-027 0.13—0.21
2ND 1ND 3ND
12346789—OCDD 2302.0 0.910044 2.5020
0.30-7.0 0.39—1.3 0.61-6.1

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of
arithmetic means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF =
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p—dioxin; HpCDD =
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

175

Table 4.12. Concentrations of selected co-contaminants in blood plasma of great blue
herons nestlings collected during 2006-2007 from rookeries within the Chippewa,
Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Values (ng/kg wet
wt) are given as the arithmetic mean 0 1 SD over the range.

 

 

 

 

FRE SNWR W

Chemical n=4 n=4 n=

PCB 77 1309.0 25014
3.2—24 16—45

PCB 81 7003.1 1 1506.9
2.6—9.7 1 1—25

PCB 126 8.4077 8.7025
1.7—19 6.2—12

PCB 169 1,200.13 1.60039
099—13 1.2—2.1

PCB 105 9400740 8900240
230—2000 600—1200

PCB 114 69054 71019
16—150 50—94

PCB 118 270002100 26000590
580—5700 2000—3300

PCB 123 61044 67016
15—120 56—91

PCB 156 3000230 280063
58—610 220—360

PCB 157 66052 63012
13—140 52—79

PCB 167 1400100 150027
28—280 120-180

PCB 189 36025 3706.3
02—69 32—44

2,4213131"a 00094000045 002800.028
00038—0014 0012—0071

224201313 b 8,407.2 8.1021
1.8—1 9 5.6-1 1

4, 4’-DDT 0.08700062 0.08100083
0032—017 0020—020

 

\\
a DDT = dichloro-diphenyl-trichloroethane
b DDE = dichloro-diphenyl-dichloroethylene

176

Table 4.13. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin
congeners in adipose of great blue heron nestlings collected during 2006-2007 from
rookeries within the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains,

Midland, MI, USA. Valuesa (ng/kg wet wt) are given as the arithmetic meanb 0 1 SD

over the range.

 

 

SNWR CAS
Chemicalc n=3 n=2
2378-TCDF 30 20019
, 057—37 13—15
23478-PeCDF 33 21023
1.0—46 17-32
12378-PeCDF 6.7 2,201.7
0.26—3.3 3.4-4.3
234678-HxCDF 2.3 1,201.1
0029—23 1.8—3.4
123789-HxCDF 1ND 3ND 2ND
123678-HxCDF 6.7 2,902.6
0058—52 3.1—7.2
123478ch1»: 4.7 1801.6
' 0.16—3.3 3.0—5.0
1234789—HpCDF 1ND 3ND 2ND
1234678-HpCDF 2.8 1.4012
0056—22 2.4—5.1
12346789-OCDF 0.57 1.1 .
2ND 2ND
2378-TCDD 28 16016
0.22—32 15-17
12378-PeCDD 21 12012
0.1 1—24 14—21
123780ch131) 2.2 -
1.1—2.1 2.3-3.0
1ND '
123678-HxCDD 18 9208.8
0.078—18 13-24
123478-HxCDD 3.2
1.7—2.7 2.5-3-3

Table 4.13 (cont’d)

 

1ND
1234678-HpCDD 5.3 I 1.9016

0.074—2.9 4.2-5.4
12346789-OCDD 3.3 2.6020

0.55-4.6 2.6-4.7

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of arithmetic

means
c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD

pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

178

Table 4.14. Concentrations of selected co-contaminants in adipose of great blue heron
nestlings collected during 2006-2007 from rookeries within the Chippewa,
Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesa (pg/kg wet
wt) are given as the arithmetic mean :t 1 SD over the range.

 

 

FRE SNWR CAs

Chemical n=1 ”:3 n=2
PCB 77 3.3 5.4032

2.7—9.0 2.2—2.5
PCB 81 1.7 3001.9

1.2—5.0 0.18—0.33
PCB 126 2.0 2,901.6

1.1-3.9 0.69—1.2
PCB 169 0.15 01800094

0081—027 0086—011
PCB 105 190 2500140

90—360 73—110
PCB 1 14 14 19i10

7.2—26 7.0—11
PCB 118 610 8100440

300—1100 200—350
PCB 123 13 182t9.0

7.2—24 5.4—8.3
PCB 156 63 79041

32-110 32—45
PCB 157 13 1808.7

8.1—25 6.9—11
PCB 167 25 32015

17—47 19—30
PCB 189 7.4 8,804.3

4.2—13 3.8—5.4
2,4’-DDT b 0.19 0.770045

026-1.] 0038—0054
2’,4’-DDB c 2000 18000240

1500—2000 1500—1900
4, 4’-DDT 9.4 9.9010

3.4—22 5.5—6.6

 

a
Values have been rounded and repres

 

 

 

 

 

 

 

ent only two signiﬁcant ﬁgures

Table 4.14 (cont’d)

b . .
DDT = dichloro-dlphenyl-trichloroethane

c . .
DDE = drchloro-drphenyl-dichloroethylene

180

Table 4.15.

Concentrations of seventeen 2,3,7,8-substituted furan and dioxin

congeners in liver of great blue heron nestlings collected during 2006-2007 from
rookeries within the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains,

Midland, MI, USA. Valuesa (ng/kg wet wt) are given as the arithmetic meanb 0 1 SD

over the range.

 

 

FRE SNWR CAS
Chemicalc 17:1 n=3 n=2
2378-TCDF 0.74 47080
052—140 0.42—0.49
23478-PeCDF 1.3 36058
. 0.66—100 083—13
12378-PeCDF 0.20
0.33—31
1ND 2ND
234678-HxCDF 0.16—3.5
1ND 1ND 2ND
123789-HxCDF 1ND 3ND 2ND
123678-HxCDF 0.18 4.1068
0.14-12
123478-HxCDF 0.24 4.9081 2ND
0.13—14
1234789-HpCDF 1ND 3ND 2ND
1234678-HpCDF 0.16—5.4
1ND 1ND 2ND
12346789-OCDF 1ND 3ND 2ND
2378-TCDD 0.77 13022
0.43—39 0.48—0.48
12378-PeCDD 0.65 1 1017
034—31 0.43—0.60
123789-HxCDD 3.5
1ND 2ND 2ND
123678—HxCDD 0.66 9.0015
0.30—26 0.46—0.51
123478-HxCDD 0.21-4-3 '
1ND 1ND 2ND
1234678-H CDD 0.30 2603.8 -
p 024—70 0.29—0.38
12346789- 1.3 2,703.4
OCDD 070—66 074—41

 

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

181

Table 4.15 (cont’d)
Concentrations below 11m1t of detection treated as zero in the calculation of arithmetic

means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p—dioxin; OCDD = octachlorodibenzo-p-dioxin

182

and Saginaw River ﬂoodplains, Midland, MI, USA. Values (pg/kg wet wt) are given as the

arithmetic mean d: 1 SD over the range.
F RE SNWR CAS

 

 

 

 

 

 

 

 

 

 

 

Chemical n=l n=3 n=2
PCB 77 0.11 0.2200085
- 0.14—0.3 1 0057—0076
PCB 81 0.091 014000.053
0.097—020 00015-00058
PCB 126 0.040 007100.070
0024-0. 15 0.022—0.024
PCB 169 0.0038 00050000047
00013—0010 00023—00026
PCB 105 3.6 6504.7
2.1—ll 2.8—2.9
PCB 114 ‘ 0.26 0.490038
0.15-0.90 0.23—0.25
PCB 118 9.3 19017
5.5—37 7.5-8.6
PCB 123 0.23 0.410031
’ 0.16—0.76 0.13—0.14
PCB 156 1.1 1.9015
067—35 099—] .2
PCB 157 0.24 0.390031
0.16-0.75 0.21—0.26
PCB 167 0.49 0.880078
0.35-1.8 0.55—0.64
PCB 189 0.12 01900.14
0076-035 0.1 1—0-21
2,421313Ta 0.0049 00077000059 0.027
00031—0014 1ND
2’,4’-DDE b 33 37016
27—55 31-53
4, 4’-DDT 0.01 1 0013000086 0.022
00062—0023 1ND
\

 

Note: Values have been rounded and represent only two signiﬁcant ﬁgures
a DDT = dichloro-diphenyl-trichloroethane
b DDE = dichloro-diphenyl-dichloroethylene

183

Table 4.17. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners in
muscle of great blue heron nestlings collected during 2006-2007 from rookeries within the

Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesa
(ng/kg wet wt) are given as the arithmetic meanb 0 1 SD over the range.

 

 

SNWR CAS
Chemicalc n=3 n=2
2378-TCDF 1.1 2,101.1
0.95—3.0 0.37—0.42
23478-PeCDF 0.83 2902.1
068—48 0.42-0.92
12378-PeCDF 0.41
0.33—0.50
1ND 2ND
234678-HxCDF 1ND 3ND 2ND
123789-HxCDF 1ND 3ND 2ND
123678-HxCDF 0.22 05000.38
0.22—0.93
123478-HxCDF 0.40—0.44
1ND 1ND 2ND
1234789-HpCDF 1ND 3ND 2ND
1234678-HpCDF 0.31
1ND 2ND 2ND
12346789-OCDF 0.62
1ND 2ND 2ND
2378-TCDD 0.77 1501.4
0.57-3.1 0.33—0.70
12378-PeCDD 0.54 1.100.96
041-22 0.39—0.56
123789-HxCDD 1ND 3ND 2ND
l23678—HxCDD 0.39 09200.64
038-16 0.24—0.61
123478-HxCDD 1ND 3ND 2ND
04800.19

1234678-HpCDD

Table 4.17 (cont’d)

 

1ND 0.27-0.62 2ND
12346789—OCDD 0.66 2.201 .1
1.5-3.4 2ND

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of arithmetic
means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzofuran; OCDF
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

185

and Saginaw River ﬂoodplains, Midland,

arithmetic mean 0 1 SD over the range.

 

 

Chemical n=l n=3 n=2
PCB 77 0W

0.083-0.40 0.048—0.063
PCB 81 0.037 01200.11

0028-024 00034—00079
PCB 126 0.043 0.140016

0034-033 0016—0027
PCB 169 0.0051 001400.010

00067—0025 00018—00020
PCB 105 3.8 13013

2.7-28 2.0-3.3
PCB 114 0.35 1.1012

0 0.23-2.4 0.15—0.31

PCB 118 13 35:1:37

8.9—77 6.5—ll
PCB 123 0.30 0.930098

0.21—2.0 0.065—0.21
PCB 156 1.6 4.6052

1.1-ll 0.75—1.3
PCB 157 0.30 1.001.]

0.25-2.3 0.16:0.31
PCB 167 0.79 2.3026

0.63-5.3 0.43-0.86
PCB 189 0.18 0.570064

0.14-1.3 0088-018
2,4213ma 0.019 0.04300010

0031—0050 00077—0025
2343-131»; b 51 110095

51-220 45‘48
4, 4’-DDT 0.061 0.16001 1 ‘

0068—027 0084-0090

MI, USA.

ppewa, Tittabawassee,
Values (pg/kg wet wt) are given as the

FEM

 

\
Note: Values have been rounded

DDT = dichloro-diphenyl-trichloroethane

DDE = dichloro-diphenyl-dichloroethylene

a

b

and represent only two signiﬁcant ﬁgures

1.86

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194

CHAPTER 5

Multiple lines of evidence risk assessment of belted kingﬁsher exposed to PCDFs and

PCDDS in the Tittabawassee River ﬂoodplain, Midland, MI, USA

Rita M. Seston], John P. Giesyl’2’3, Timothy B. F redricks], Dustin L. Tazelaar4, Sarah J.
Coeﬁeldl, Patrick W. Bradley4, Shaun A. Roarks, John L. Newsteds, Denise P. Kays, and

Matthew J. Zwiemik4

‘Zoology Department, Center for Integrative Toxicology, Michigan State University, East

Lansing, MI 48824, USA

2Department of Veterinary Biomedical Sciences and Toxicology Centre, University of

Saskatchewan, Saskatoon, Saskatchewan, S7J 5B3, Canada

3Department of Biology and Chemistry, City University of Hong Kong, Hong Kong,

SAR, China

4 . o 0
Department of Animal Science, Michigan State Umvers1ty, East Lansrng, MI 48824,

USA

5
ENTRIX, Inc., Okemos, MI, USA

195

Abstract

Concentrations of dioxin-like compounds, primarily polychlorinated dibenzofurans
(PCDFs), in soils and sediments of the Tittabawassee River (TR) and associated
ﬂoodplains downstream of Midland, Michigan (USA) are greater than upstream sites. As
a result of these elevated concentrations, a site-speciﬁc risk assessment of belted
kingﬁsher (BKF) breeding in the assessment area was conducted. To reduce the
uncertainty associated with predicting exposure from abiotic matrices, concentrations of
residues were quantiﬁed in site-speciﬁc prey items and also in eggs and nestlings BKF.
Simultaneously, site-speciﬁc assessments of reproductive effort and success of BKF were
conducted. Dietary exposure, expressed as the potential average daily dose, based on

site-speciﬁc concentrations of PCDFs, polychlorinated dibenzo-p-dioxins (PCDDS), and
2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents (TEQWH0-A,.ian) in prey items was

consistently greater along the TR than in associated reference areas (RAs) and the further
downstream sites in the Saginaw River (SR). Concentrations of PCDD/DPS in eggs and
nestlings of BKF were associated with sampling area, being greater in both eggs and
nestlings of BKF nesting along the TR compared to those of BKF from upstream
reference areas. Geometric mean concentrations of PCDD/DPS were 130 and 200 ng/kg,
wet wt in eggs and nestlings of BKF, respectively, collected from nests along the TR.
Potential for adverse population-level effects associated with site—speciﬁc diet and egg
contaminant exposures were evaluated by comparison to dietary and egg-based toxicity
reference values (TRVs). Minimal risk of adverse population-level effects was predicted

based on either measured dietary- or tissue-based exposures. This conclusion was

196

consistent with site-speciﬁc measures of population condition, which included clutch

size, hatching success, and ﬂedging success.

Introduction

Concentrations of polychlorinated dibenzofurans (PCDFs) and polychlorinated
dibenzo-p—dioxins (PCDDS) in the Tittabawassee River (TR) and associated ﬂoodplains
downstream of Midland, MI, USA are greater than at upstream locations and regional
background concentrations. The presence of PCDD/DPS is the result of historical
chemical production and associated waste management practices (USEPA 1986);
Sediments and ﬂoodplain soils downstream of Midland, MI contain total concentrations

of the seventeen 2,3,7,8-substituted PCDD/DP congeners (EPCDD/DFS) ranging from
1.0x102 to 5.4x104 ng/kg dry wt, respectively. In contrast, concentrations of ZPCDD/DFs

in sediments and soils from upstream reference areas were 10- to 20-fold less
(Hilscherova et al. 2003). The elevated concentrations of PCDD/DPS within the TR
ﬂoodplain led to concerns about their potential effects on resident wildlife species in the
TR and Saginaw River (SR) ﬂoodplains.

Both PCDD/DPS and dioxin-like polychlorinated biphenyls (PCBs) are persistent in
the environment and due to their lipophilic nature tend to biomagnify (Mandal 2005).
Additionally, these compounds are known to cause an array of negative effects in
mammalian and avian species. The PCDD/DF and PCB congeners with the greatest toxic
potency act via a common mechanism, the aryl hydrocarbon receptor (AhR) and effects
include enzyme induction, immunotoxicity, and adverse effects on reproduction,

development, and endocrine functions (Hoffman et al. 1987; van den Berg et al. 1994a).

197

In particular, exposure to compounds that bind to the AhR can result in lesser hatching
and ﬂedging success of bird species (Custer et al. 2005; Giesy et al. 1994; Gilbertson
1983; Kubiak et al. 1989; Ludwig et al. 1993).

The sensitivities of a number of species to AhR-mediated effects have been
determined in laboratory studies or inferred from observations of populations exposed in
the wild. Sensitivities have been shown to vary among species (Brunstrtim 1988). For
example, the domestic chicken (Gallus gallus), which is considered the most sensitive
avian species, is more than 1000-times more sensitive to embryo-lethal effects than is the
mallard (Anas platyrhyncos) (Head et al. 2008). One theory that has been proposed to
account for these differences in avian sensitivity is that toxicity can be attributed to
variations in the afﬁnity of dioxin-like compounds to the ligand-binding domain (LBD)
of the AhR (Head et al. 2008). Speciﬁcally, differences in three different amino acid
sequences in the LBD of the AhR among avian species leads to differential binding of
ligands, which subsequently results in differential sensitivities (Karchner et al. 2006).
These differences can be used to classify birds into groups with different sensitivities to
AhR-mediated effects.

The primary objective of the present study was to characterize the exposure of belted
kingﬁshers (Ceryle alcyon; BKF) foraging and nesting iwithin the TR ﬂoodplain to
2,3,7,8-substituted PCDD/DPS congeners and predict the associated risk of effects.
Species-speciﬁc characteristics of the BKF give it a great potential for exposure and
make it well suited for study using a multiple-lines-of-evidence approach to ecological
risk assessment. The lines of evidence used here include (I) predicted exposure to

contaminants through the diet, (2) measured concentrations in egg and nestlings of BKF,

198

and (3) site-speciﬁc measures of BKF population condition, including clutch size,
hatching success, and ﬂedging success. Concentrations of ZPCDD/DF were measured in
dietary items and BKF tissues collected from reference and study areas. These

concentrations were also expressed as 2,3,7,8-tetrachlorodibenzo-p—dioxin (TCDD)

equivalents (TEQWH0-Avian) based on World Health Organization 2,3,7,8-TCDD

equivalency factors for birds (TEFWHO-Avian) (van den Berg et al. 1998). Spatial trends

in concentrations and relative congener contributions in dietary items and BKF tissues
were also evaluated. Measured exposures were compared to appropriate toxicity
reference values (TRVs) to estimate the risk present. Integrating the data resulting from.
these multiple assessments reduces the uncertainty inherent in the risk assessment process
(Fairbrother 2003; Leonards et al. 2008; USEPA 1998) and provides better information

for use in risk management decisions.

Methods
Site description

The assessment was conducted in the vicinity of the city of Midland, located in the
east-central lower peninsula of Michigan (USA) (Figure l). The TR is a tributary of the
Saginaw River (SR), which empties into Saginaw Bay and Lake Huron. The TR runs
through The Dow Chemical Company (DOW), which is the accepted source of the
PCDD/DF contamination (USEPA 1986). The area henceforth referred to as the study
area (SA) includes approximately 37 km of the TR (sites T-3 to 'T-6) stretching
downstream from DOW to the convergence of the TR and SR and 35 km of the SR (sites

S-7 to S-9) until it enters Saginaw Bay. Sampling sites selected in the SA were chosen to

199

Figure 5.1. Assessment area along the Chippewa, Tittabawassee, and Saginaw river
ﬂoodplains, Michigan, USA. Sampling locations for dietary components of belted
kingﬁsher were located in the Reference area (R-1 and R-2); Upper Tittabawassee River
(T-3 and T-4); Lower Tittabawassee River (T-S, T-6, and S-7); and Saginaw River (S-8
and S-9). Tissues of belted kingﬁsher were collected in the reference area (RA) and

along the Tittabawassee River (study area; SA).

               

 

0.0.0.005. m I.
o

E 0.0 05.

some”. 0000030003.... .0264

 

 

. .078.
, _ somom 0000030003; .000: a
. o
, 02¢ 02,009;
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201

characterize maximal exposure potential designated as “worst case scenario” locations
based on a previous study that measured concentrations of PCDD/DPS in soils and
sediments (Hilscherova et al. 2003). The willingness of landowners to provide access
was also a factor in site selection. The reference area (RA) includes the TR upstream of
Midland, together with the Pine and the Chippewa River, both of which are tributaries of
the TR upstream of Midland. Reference area sampling locations were on the upstream
TR (R-1) and on the Pine River, just upstream of its conﬂuence with the Chippewa River
(R—2). Sampling areas were assessed both individually and in spatial groups based on
characteristics of the river and associated ﬂoodplain. Spatial groupings included
reference (RA) R-1 and R-Z, upper Tittabawassee River (UTR) T—3 and T-4, lower
Tittabawassee River (LTR) T-S to S-7, and Saginaw River (SR) S-8 and S-9.
Components of the diet, including soil, sediment, and prey items were collected at each
sampling area whereas BKF tissues were collected from nests and were designated to

either RA or SA.

Receptor Species Selection

Selection of a receptor species is a key element in effective risk assessment,
particularly when site-speciﬁc ﬁeld studies are to be employed. The belted kingﬁsher
was selected as a receptor species because it possesses characteristics desirable to
investigate exposure and potential effects of PCDD/DFs via an aquatic exposure
pathway. As top aquatic food web predators with high food consumption rates, BKF
have a relatively great potential for exposure (Hamas 1994; Vessel 1978). Additionally,

BKF excavate subterranean burrows in which they nest, thus adult and nestling BKF have

202

intimate contact with river bank soils. Because BKF are territorial of distinct foraging
ranges proximal to the nest burrow, the spatial boundaries of the area from which nestling
BKF are exposed can be better deﬁned than other species (Davis 1980). Their
widespread distribution has led to BKF being included in other ecological assessments

(Baron eta]. 1997; Evers and Lane 2000; Moore et al. 1999).

Diet-based exposure assessment

Collection of prey items. Prey items of BKF, including forage ﬁsh, crayﬁsh, and
amphibians were collected and concentrations of PCDD/DFS (n=188) and PCBs (n=20)
determined. Prey items were collected from nine sampling locations (Figure 5.1). The
sampling scheme maximized information on dietary exposure including geographically
associated contaminant variability and trends. Detailed sample collection methods have

previously been described (Seston et al. 2009).

Dietary exposure calculations. Exposure of BKF to PCDD/DPS via the diet was
estimated by use of information provided in the US. Environmental Protection Agency
(USEPA) Wildlife Exposure Factors Handbook (WEFH) (USEPA 1993). Major factors
inﬂuencing dietary exposure included body mass (BW), daily food intake rate [FIR; g
wet weight (wet wt) food/g body weight (bw)/d], dietary concentrations (C), and

proportion of foraging time spent on—site (AUF). A mean body mass of 147 g and a

body-weight normalized FIR of 0.50 kg food(wet wt) 'kg bw'I °d'I for BKF is given in the

USEPA WEF H. The potential average daily dose (ADme; ng/kg bw/d) was calculated

203

using equation 4-3 of the WEFH (USEPA 1993). Incidental sediment ingestion was also
included in the ADDpot using equation 4-23 (USEPA 1993). ,

The relative proportion of each type of prey to the BKF diet was determined through
collection and identiﬁcation of prey remains from active nest chambers located in the
assessment area. Collected remains were sorted to prey item type by distinct elements,
such as ﬁsh otoliths, pharyngeal arches, and dentary bones, crayﬁsh chelipeds, and
amphibian femurs or pelvic girdles. Reach-speciﬁc concentrations of PCDD/DFs in prey
items were multiplied by their relative contribution to the dietary composition. Exposure
was estimated using the geometric mean and associated 95% conﬁdence interval of
concentrations of residues in prey items from each reach (RA, UTR, LTR, and SR).

PCBs were not measured in all frogs or crayﬁsh, thus the data set is incomplete for

calculating total TEQWHO-Avian exposure associated with PCDD/DPS and PCBs. Where

both PCDD/DP and. PCB data were available, concentrations of total TEQSWHO-Avian in
frogs and crayﬁsh were less than proximally collected ﬁsh. To estimate a conservative
exposure of BKF to total TEQSWHO-Avian a dietary composition of 100% ﬁsh was

utilized.

Tissue-based exposure assessment

Nest excavation and monitoring. Active BKF nest burrows were located via canoe
surveys of the rivers in the assessment area during 2005-2007 from mid-April to mid-
May and again in late June to search for any pairs that may have re-nested. Burrows
were considered active when there was evidence of fresh digging around and beneath the

burrow entrance, the presence of tracks left by the feet of BKF along the base of the

204

entrance, and/or defensive behavior of BKFs in the area. Further conﬁrmation of burrow
activity was gained by use of an infrared video camera that was inserted into the burrow
opening to visualize the nest chamber.

Once a burrow was deemed active, the location of the nest chamber was determined
by using a folding wooden ruler to approximate the length and angle of the burrow
tunnel. A hole was dug behind the approximate location of the nest chamber, and then
the leading edge of the excavation pit was slowly moved forward until a small opening
was made in the rear of the nest chamber (Mazeika et al. 2006). After the nest chamber
was located, a wooden panel with an access door and video-port was installed to allow
access for nest monitoring and sample collection. The entire excavation was then
covered with a sheet of 2.0 cm exterior grade plywood and a tarpaulin to prevent water
and predators entrance to the excavation or nesting chamber. It was optimal to perform
the excavation once the clutch was complete and incubation had begun to lessen the risk
of nest abandonment by adults (Davis 1982).

Nest abandonments attributed to inﬂuences other than contaminants were not
included in comparisons of nesting success between study locations. The excavation
process employed in the present study could have been disruptive to normal behavior of
nesting BKF. Nest abandonment was considered to be due to excavation if either of the
following two conditions were met (1) the nest contained an incomplete clutch at the time
of excavation and egg laying never resumed, or (2) the nest contained eggs at the time of
excavation and they were never warm after excavation. These conditions also required
an indication that the nest was active at the time of excavation, such as presence of

defensive adults and/or warm eggs. During early May 2006, the assessment area

205

received approximately 7.5 cm of rain within 24 h. Water clarity remained reduced for
several days afterward, which likely reduced foraging success of BKF. Previous
observations have shown that BKF will abandon their foraging areas when turbidity
increases as a result of heavy rains (Davis 1980; Salyer and Lagler 1949). Nests of BKF
that were active prior to the rain and then abandoned up to three days following the rain
event were considered to have done so in response to the decreased foraging success
proximal to the nest during incubation.

Nests were monitored from late-April to mid-July. Nests with complete clutches
were checked every other day to determine hatch date, nestling status, and ﬂedge date.
All nestlings were banded with US. Fish and Wildlife Service (USFWS) bands. Nest
success, clutch size, hatching success, and ﬂedging success were all used to assess the
reproductive effort and success of BKFs. Clutch size was not adjusted for egg sampling,
as collection was done during incubation when the clutch was already complete.
However, hatching and ﬂedging success were calculated based on an adjusted clutch size
since the fertility and hatchability of the collected egg was unknown at collection.
Adjusted clutch size was deﬁned as the clutch size excluding any eggs that were
collected. Hatching success (number of eggs that hatch per adjusted clutch size) and
ﬂedging success (number of nestlings that ﬂedge per number of eggs that hatched) were
adjusted to account for egg collection. Mortality of nestling BKF is low and generally
occurs early in the nestling period (Hamas 1975). Since nestlings were collected past
midway into their development, it was assumed that any nestlings collected would have
successfully ﬂedged provided the remaining portion of the nesting attempt was

successful. Thus, ﬂedging success was not adjusted for sampled nestlings. Nestlings

206

reaching 25 d of age were considered successfully ﬂedged. Reproductive parameters for
all clutches were included in comparisons up to the point that they were preyed upon or
abandoned due to human interference or the aforementioned rain event.

Various nest activities, including incubation, hatching, and feeding of nestlings, were
recorded using the camera at the rear of the nest chamber, and during routine handling,
nestling BKF were monitored for gross external abnormalities. Additionally, adult BKF
were trapped and banded with USFWS bands to allow determination of nest site ﬁdelity

and survival.

Collection of BKF tissues. Both eggs and nestlings of BKF were collected from each
available nest chamber located within the assessment area for quantiﬁcation of
PCDD/DPS. One egg was selected at random from each clutch. Mass, length, and three
width measurements were recorded for each egg. Addled or abandoned eggs were also.
collected for possible quantiﬁcation of PCDD/DF congeners. Addled eggs were deﬁned
as those which failed to hatch 2 d post hatch of other eggs in the clutch. Abandoned eggs
were deﬁned as those which were cold on three consecutive visits to the nest. Individual
eggs were wrapped in chemically-cleaned foil and placed inside a glass jar (I-CHEM
brand, Rockwood, TN) for storage and transport to the laboratory. One nestling from
each nest was randomly selected for collection at age 15 d and euthanized via cervical
dislocation. Individual collected nestlings were stored in similar jars for storage and
transport.

During the 2005-2007 breeding seasons, a total of 37 nest chambers were excavated.

Eggs were collected from 6 and 19 unique clutches in the RA and SA, respectively. A

207

total of 5 and 12 nestlings were collected from unique broods in the RA and SA,
respectively. Of the collected nestlings, 9 (3 RA and 6 SA) were collected from nests

from which an egg was also analyzed.

Sample processing and analytical techniques

Concentrations of seventeen 2,3,7,8-substituted PCDD/DP congeners were measured
in all samples while concentrations of PCBs and dichloro-diphenyl-trichloroethane
(DDT) and related metabolites (DDXs) were determined in a subset of samples.
Collected eggs were opened around the girth with a scalpel blade. Contents were then
homogenized in a chemically cleaned Omni-mixer, lyophilized, and stored in chemically
cleaned jars until analysis (I-CHEM brand, Rockwood, TN). Concentrations of
PCDD/DF in eggs were reported on a fresh weight basis adjusted to account for any
desiccation during incubation and storage. Adjusted fresh weight was calculated based
on egg volume (Hoyt 1979). The mass of egg contents was determined by subtracting the
eggshell mass at the time of processing from the adjusted fresh weight. Nestlings were
homogenized in a chemically-cleaned Omni-mixer, without stomach contents, feathers,
legs below the tibiotarsus, or the beak.

Residues were quantiﬁed in accordance with USEPA Method 8290/1668A with
minor modiﬁcations (USEPA 1998). Analytical methods have been detailed elsewhere

(Fredricks et al. 2010b; Seston et al. 2010b). Brieﬂy, biotic matrices were homogenized
with anhydrous sodium sulfate, spiked with known amounts of l3C-labeled analytes (as

internal standards), and Soxhlet extracted. Ten percent of the extract was removed for

lipid content determination. Sarnple puriﬁcation included the following: treatment with

208

concentrated sulfuric acid, silica gel, sulfuric acid silica gel, acidic alumina and carbon
column chromatography. Components were analyzed using high-resolution gas
chromatography/high-resolution mass spectroscopy, a Hewlett-Packard 6890 GC
(Agilent Technologies, Wilmington, DE). connected to a MicroMass® high-resolution
mass spectrometer (Waters Corporation, Milford, MA). .Losses of congeners during

extraction were corrected based on recoveries of l3C-labeled as outlined in USEPA

Method 8290/ 1668A. Quality control samples generated during chemical analyses
included laboratory method blanks, sample processing blanks (equipment rinsate and
atmospheric), matrix spike and matrix spike duplicate pairs, unspiked sample replicates,
and blind check samples. Results of method and ﬁeld blank analyses indicated no
systematic laboratory contamination issues. Evaluation of the percent recovery and
relative percent difference data for the matrix spike and matrix spike duplicate samples
and unspiked replicate samples were within 130% at a rate of greater than 95%
acceptability. Soxhlet extractions and instrumental analyses were conducted at

AsureQuality Ltd, Lower Hutt, New Zealand.

Statistical analyses

Total concentrations of the seventeen 2,3,7,8-substituted PCDD/DF congeners
(ZPCDD/DFS) are reported as the sum of all congeners (ng/kg wet weight (wet wt)).
Individual congeners for which concentrations were less than the limit of quantiﬁcation
were assigned a value of half the sample method detection limit on a'per sample basis.
Total concentrations of the twelve dioxin-like non- and mono-ortho—substituted PCB

congeners are reported as the sum of these congeners (ng/kg wet wt) (ZPCBS) for a

209

subset of samples. Concentrations of TEQ\VHO_AVjan (ng/kg wet wt) were calculated for
both PCDD/DFs and dioxin-like PCBs by summing the product of the concentration of
each congener, multiplied by its avian TEFWHOA“an (van den Berg et al. 1998). Total
TEQS throughout this manuscript refers to the summation of TEQS from ZPCDD/DFS
(DF—TEQSWHO-AVian ) and XPCBS (PCB-TEQSWHO_AVjan ). Additionally, dichloro-
diphenyl-trichloroethane (2’,4’ and 4’,4’ isomers) and dichloro-diphenyl-
dichloroethylene (4’,4’) are reported as the sum of the o,p and p,p isomers (ZDDXS;
ug/kg wet wt) for the same subset of samples as for PCBs.

Statistical analyses were performed using SAS® software (Release 9.1; SAS Institute
Inc., Cary, NC, USA). Prior to the use of parametric statistical procedures, normality was
evaluated using the Shapiro—Wilks test and the assumption of homogeneity of variance
was evaluated using Levene’s test. Values that were not normally distributed were
transformed using the natural log (1n) before statistical analyses. PROC TTEST was used
to make comparisons between the RA and SA. PROC GLM was used to make
comparisons for three or more locations. When signiﬁcant differences among locations
were indicated, the Tukey—Kramer test was used to make comparisons between individual

locations. The association between concentrations of ZPCDD/DF or DF—TEstHO-AVian

and hatching success was evaluated with Spearman’s correlation coefﬁcients for nesting
attempts in which both data were collected. Statistical signiﬁcance was inferred at

p<0.05.

To better understand the potential distributions of concentrations of DF-TEQWHQ-

Avian and total TEQ\,VH()_,oWian in eggs of BKF, a probabilistic modeling approach was

210

used to portray the distributions. Probabilistic models were developed as cumulative

frequency distributions based on concentrations of DF-TEQWHO-Av;an and total TEQWHO.

Avian in eggs. The mean and standard deviation of transformed concentrations of each

sample type were used to generate 10,000 iterations of random concentration values

based on a lognormal distribution.

Selection of toxicity reference values

Literature-based no observed adverse effect concentrations (NOAECs) and lowest
observed adverse effect concentrations (LOAECs) were used in the determination of
hazard quotients (HQs) and subsequent assessment of risk. In the present study, matrix—
speciﬁc toxicity reference values (TRVs) based on the same or similar compounds were
identiﬁed from literature and compared to measured site—speciﬁc exposures of BKF.
. Resulting HQs are presented as a range bounded by the LOAEC-associated HQs at the
low end and the NOAEC-associated HQs at the high end. It should be noted that the
NOAEC and LOAEC associated HQs are a function of the experimental design (dosing
regimen) and the actual threshold concentration at which effects would be expected to
occur somewhere within the described range. It has recently been suggested that TRV
selection should involve the combination of multiple suitable studies into a dose-response
curve to determine the most accurate value (Allard et al. 2010). However an inadequate
number of suitable studies precluded the use of that approach in the present study.

Laboratory studies of effects from dietary exposure to PCDD/DPS are limited for
avian species. The TRV selected for use in this assessment was derived from a study that

dosed adult hen ring-necked pheasants (Phasianus colchicus) with TCDD through

211

intraperitoneal (IP) injection (Nosek et al. 1992). The dietary-based TRVs were

determined by converting the weekly exposure at which adverse effects on fertility and
hatching success were determined (1000 ng TCDD/kg/wk) to a LOAECDIET for daily
exposure of 140 ng TCDD/kg/d. Adverse effects were not present at the next smaller
dose, which was determined to be the NOAECDIET for dietary exposure (14 ng
TCDD/kg/d).

The number of laboratory studies that report the effects of PCDD/DFs from egg-
based exposures is limited. The USEPA previously developed egg-based TRVs by
taking the geometric mean of the effect concentrations in three double-crested cormorant
(Phalacrocorax auritus) egg-injection studies (P0well et al. 1997a; Powell et al. 1997b;

Powell et al. 1998; USEPA 2003). Based on a measurement endpoint of embryo

mortality, the resulting NOAECEGG and LOAECEGG was 3670 and 11090 ng total

TEQWHO-Avian /kg wet wt, respectively. No studies reporting effects of dioxin-like

compounds in nestlings were available for comparison.

Results
PCDD/DFS and PCBs in BKF prey

Concentrations of the contaminants of concern varied among sampling reach and prey

type, ranging from 1.8 ng EPCDD/DFs/kg and 0.29 ng DF-TEQWH0_Avian/1<g in the RA
to 3300 ng XPCDD/DFs/kg and 1900 ng DF-TEQWH0-Avia,,/kg in the SA (Table 5.1).

Consistent spatial trends in concentrations of ZPCDD/DFs and DF—TEanamm were

212

Table 5.1. TEQWHO-AV;ana in prey items of belted kingﬁsher collected during 2004-
2006 from the Chippewa, Tittabawassee, and Saginaw River ﬂoodplains, Midland,
MI, USA. Valuesb (ng/kg w) are given as the geometric mean TEQqumm and
sample size in parentheses (n) over the 95% conﬁdence interval and range (min-max).

 

Frog
ZPCDD/DF

XPCB

DF-TEQWHo-Avian

PCB-TEQWHO-

Avian

Crayﬁsh
EPCDD/DF

ZPCB

DF’TEQWHO-Avian

PCB-TEQWHQ-

Avian

Forage Fish
EPCDD/DF

 

 

Reach
Reference Upper Lower Saginaw
Area Tittabawassee Tittabawassee River
5.5 (29) A 49 (51) B 100 (55) C 6.7 (12) A
4.3-6.9 33-73 78-140 4.7-9.4
(1 .8-21) (4.4-920) (17-3300) (3 .3-26)
N/A N/A 1300 (4) N/A
880-2000
(940-1700)

1.0 (29) A 20 (51) B 56 (55) C 2.9 (12) A
O.81—1.3 13—31 42——76 1.9-4.5
(0.29—3.4) (1.1—460) (9.1—1900) (1.5—~14)
N/A C N/A 3 1.5 (4) N/A

0.86—2.5
(1.1—2.0)

5.0 (5) A 140 (7) B 360(8) C 50 (8) D
1.8—1 3 83-240 200—650 34—72
(2.3—12) (86—340) (140—1300) (28—110)
N/A _ c (2) —- (1) N/A

(2200—2200) (2100)

0.91 (5) A 55 (7) B 160 (8) C 27 (8) B
0.29-2.8 24—130 110—250 18—41
(034—29) (12—190) (75—420) (13-61)

N/A — (2) —— ( 1) N/A
(3.3—4.3) (5.3)
—- (2) ~ (2) 260 (5) A 59 (4) B
_ - 95~730 26—130

213

Table 5.1 (cont’d)

 

(4. 1-5. 1) (200.220) (83—610) (374210)
ZPCB — (2) ~ (2) 18000 (5) A 25000 (4) A
I —- —~ 4600—72000 10000—
62000
(880—1000) (9300—15000) (7100—1 10000) (15000—
47000)
DF-TEQWHQ-AVian ” (2) “ (2) 180 (5) A 33 (4) B
—— —— 66—480 20—56
(0.90—091) (130—170) (74440) (25—53)
PCB-TEQWHO- - (2) - (2) 31 (5) A 55 (4) A
Avian
—— —— 1. 0—92 1 2--250
(1.3—1.8) _ (30—39) (13—100) (24—150)

 

a TEQWHomian were calculated based on the 1998 avian WHO TEF values
b Values have been rounded and represent only two signiﬁcant ﬁgures
c N/A = no samples collected from this location

d Means identiﬁed with the same uppercase letter are not signiﬁcantly different among
reach at the p=0.05 level using Tukey-Kramer means separation test.

6 Geometric mean and conﬁdence intervals not calculated for sites with fewer than 3
samples. These sites were not included in reach comparisons.

214

observed for each type of prey. Concentrations were least in the RA, greater in UTR,
greatest in LTR. Concentrations in preyitems from the SR were intermediate to those
from the RA and UTR. Mean concentrations of ZPCDD/DFs were 9-, 18, and 1- fold
greater in frogs, 28-, 72-, and 10—fold greater in crayﬁsh, and 47-, 58-, and 13-fold greater

in forage ﬁsh collected from the UTR, LTR, and SR than those from the RA,
respectively. Mean concentrations of DF-TEQWHO_A,.ian were 20-, 56-, and 3-fold greater

in frogs, 60-, 180-, and 30-fold greater in crayﬁsh, and 170-, 200-, and 36-fold greater in
forage ﬁsh collected from the UTR, LTR, and SR when compared to the RA,
respectively. Concentrations of ZPCBs in prey items followed a different spatial trend
than that of ZPCDD/DFs, being greatest in the SR. A more detailed description of the
spatial trends in concentrations of PCDD/DPS and PCBs in the prey items, along with

discussion of relative congener concentrations, is available is Seston et al. (2010b).

Spatial trends were also seen in the relative contribution of DF'TEQWHO-Avian and
PCB-TEQ\)VH().AVian to total TEQWHO-Avian in prey items. A majority of the total

TEQWHO_AVian in ﬁsh from the RA and SR were attributed to PCB-TEQSWHO-A,.ian (63%
and 61%, respectively), in contrast to ﬁsh collected from the Tittabawassee River that

had a majority of their total TEQWHO-A,,.ianattributed to DF- TEQWIlO-Avian (81% and

85% for the UTR and LTR, respectively). PCB-TEQSWHGAW were dominated by

PCB-126, PCB-77, and PCB-81 in all reaches.

Dietary exposure

215

Prey remains collected from BKF nest chambers revealed the site-speciﬁc diet to be
90.2% ﬁsh, 5.4% crayﬁsh, and 4.4% frogs. Using this site-speciﬁc diet, the ADDpot was
consistently greatest within reaches of the Tittabawassee River when compared to either

the RA or SR, regardless of whether it was based on ZPCDD/DF, DF-TEQWH0_Avian, or

total TEQWH0-AVian (Table 5.2). ZPCDD/DF— ADDpot to BKF was 44- to 54-fold greater
along the Tittabawassee River and 12-fold greater along the SR when compared to the
RA. When normalized to DF-TEQWHOAvian, fold differences in ADDpot increased, being

150- to 190-fold greater along the Tittabawassee River and 35-fold greater along the SR
when compared to the RA. Site-speciﬁc prey items contained contaminant burdens that

were quite similar (wet wt), with forage ﬁsh composites being slightly greater than
crayﬁsh and frogs. Based on a diet of 100% ﬁsh, the ADDPO, of DF-TEQWH0.A,.ian
ranged from 0% to 6.6% greater, resulting in a slightly more conservative estimate. The
ADDpot expressed as total TEQ\VHO_AVian based on a 100% ﬁsh diet was 76- to 92-fold

greater along the Tittabawassee River and 38-fold greater along the SR when compared

to the RA.

PCDD/DFS and PCBs in BKF tissues
Concentrations of measured residues were consistently greater in tissues of BKF
collected within the SA compared to those of the RA. However, the differences varied in

their statistical signiﬁcance (Table 5.3). Mean concentrations of ZPCDD/DFS and DF-
TEQWHO_AV;an were signiﬁcantly greater in eggs of BKF collected from the SA compared

to those from the RA (4- and 6-fold greater, respectively; p<0.0001). Although mean

216

Table 5.2. Predicted daily dietary dose of XPCDD/DFs and TEQSWHO-A.,,,, a (ng/kg

body weight/d b’ c) for adult belted kingﬁsher breeding during 2004-2006 within the
Chippewa, Tittabawassee, and Saginaw river ﬂoodplains, Midland, Michigan, USA,
based on the geometric mean (95% conﬁdence interval) of site-speciﬁc dietary items.

 

 

 

Study Area spoon/DPS DF-TEQWHO-AV,ana Tom] TEQJVHO'AV‘T
Reference C

2.6 (2.1—3.5) 0.46 (0.44—0.53) 1.2 (1.1—1.4)
Upper Tittabawassee River f’ g‘ h

114 (100—130) 70 (60—83) 91 (80—110)
Lower Tittabawassee River i

140 (54—370) 88 (34—230) 110 (39—290)
Saginaw River

31 (13—110) 16 (9.5—32) 46 (15—140)

 

a TEQquMian were calculated based on the 1998 avian WHO TEF values

b Values were rounded and represent only two signiﬁcant ﬁgures

C Food ingestion rate was calculated from equations in The Wildlife Exposure Factors

Handbook (us EPA 1993)

d Total TEQWHO-AVian based on a diet of 100% ﬁsh due to lack of PCB data in frogs

and crayﬁsh

6 Two ﬁsh composite samples collected from Reference reach. Range represents daily
dietary dose based on minimum-maximum of ﬁsh concentrations.

1' . . .
Off-sue diet assumed to be equal to reference area d1et.

g Upper Tittabawassee River reach includes sites T-3 and T-4

h . . .
Two ﬁsh composrte samples collected from Upper Tittabawassee River reach. Range
represents daily dietary dose based on minimum-maximum of ﬁsh concentrations.

i Lower Tittabawassee includes sites T-S, T-6, and S-7

217

Table 5.3. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF) and
TEanGMm a belted kingﬁsher eggs and nestlings collected during 2005-2007 from the
Chippewa, Tittabawassee and Saginaw River ﬂoodplains, Midland, MI, USA. Valuesb
(ng/kg w) are given as the geometric mean (n) over the 95% conﬁdence intervals and
range in parentheses. ZPCB and ZDDX values are reported in jig/kg ww.

 

 

 

BKF Eggs BKF Nestlings
RA SA RA SA
DF-TEQs 15 (6) A 84 (19) B 5.0 (5) a 95 ( 12) b
' 6.3-35 62—1 10 2.9—8.6 73-120
(6.5-53) (36—260) (2.6-8.5) (49-180)
PCB-TEQS 29 (3) 64 (1 l) 6.0 (4) a 35 (9) b
12-74 41~99 3.8-9.6 14-87
(19~39) (23—270) (4.1-8.5) (9.4—590)
Total TEQS 47 (3) A 170 (11) B 9.7 (4) a 160 (9) b
10—210 120-240 8.0—12 100-250
(26-87) (79-520) (8.6-1 1) (110—660)
ZPCDD/DF 30 (6) A 130 (19) B 9.6 (5) a 200 (12) b
14—62 100—170 5.0-1 8 120-340
(13—73) (63—370) (4.7-20) (82-1300)
EPCB 120(3) 130 (11) 11 (4) a 67 (9) b
20-660 81-220 73-15 35—130
(52-—190) (42—650) (8.3-14) (23-430)
ZDDX 440 (3) 540 (11) 140 (4) 210 (9)
43-4600 310-940 43—430 160-280
_ (210—1300) (120-1500) (82-390) (120-350)

 

a TEQWHO_Avian were calculated based on the 1998 avian WHO TEF values.

b . .
Values have been rounded and represent only two Signiﬁcant ﬁgures.

c Means identiﬁed with a unique capitalized letter are signiﬁcantly different between

locations (across) at the p=0.05 level, for BK‘F eggs.

d Means identified with a unique lowercase letter are signiﬁcantly different between
locations (across) at the p=0.05 level, for BKF nestlings.

218

concentrations of PCB-TEmeyAvian were 2-fold greater in eggs of BKF collected from

the SA than those from the RA, the difference was not signiﬁcant (p=0.0767). Mean
concentrations of ZPCB and EDDX in eggs of BKF were similar between sampling
locations (p=0.7686 and p=0.73 72, respectively).

The trend in nestling contaminant concentrations was similar to that observed for
eggs of BKF. Mean concentrations of ZPCDD/DFS and DF-TEQWHOAWan were
Signiﬁcantly greater in nestlings of BKF collected from the SA compared to those
fromthe RA (21- and 19-fold greater, respectively; p<0.0001). In contrast to eggs, mean
concentrations of ZPCB and PCB-TEQWH0.A,.ian were signiﬁcantly greater in nestlings
of BKF collected from the SA compared to those from the RA (6— and 6-fold,
respectively; p=0.0014 and p==0.0167, respectively). Mean concentrations of EDDX in
nestlings of BKF were similar between sampling locations (p=0.1492).

Spatial trends in relative contribution of PCDDS, PCDFS, and individual congeners to
EPCDD/DFS and DF- TEQWHO-AVian in eggs of BKF were observed. Eggs of BKF
collected from the SA had a greater percent contribution of furans to both ZPCDD/DFS
and DF- TEQWHO_Avian (66% and 73%, respectively) compared to those collected from

the RA (27% and 29%, respectively). Predominant congeners of ZPCDD/DFS in eggs of
BKF from the RA included TCDD (20%) and 1,2,3,7,8-pentachlorodibenzo-p-dioxin
(PeCDD) (17%), in contrast to 2,3,4,7,8—pentachlorodibenzofuran (PeCDF) (39%) which

was the predominant congener in eggs collected from the SA (Figure 5.2). When

normalized to DF-TEQWHO_Avian, the predominant congeners remained the same in each

219

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area, with TCDD (36%) and 1,2,3,7,8-‘PeCDD (34%) in eggs of BKF collected from the
RA and 2,3,4,7,8-PeCDF (59%) in eggs collected from the SA.

Similar spatial trends in relative contribution of PCDDS, PCDFS, and individual
congeners to ZPCDD/DFS and DF- TEQwHo-Avian were also observed in nestlings of
BKF. Nestlings of BKF collected from the SA had a greater percent contribution of
furans to both ZPCDD/DFS and DF- TEQWHO-Avian (75% and 90%, respectively)

compared to those collected from the RA (37% and 38%, respectively). Predominant
congeners of ZPCDD/DFS in nestlings of BKF from the RA included TCDD (19%) and
1,2,3,7,8-PeCDD (13%), in contrast to 2,3,4,7,8-PeCDF (30%) which was the
predominant congener in nestlings collected from the SA (Figure 5.2). When normalized

to DF- TEQWHO-Avians the predominant congeners remained the same in each area, with

TCDD (36%) and 1,2,3,7,8-PeCDD (25%) in nestlings of BKF collected from the RA

and 2,3,4,7,8-PeCDF (57%) in nestlings collected from the SA.

There was also a spatial trend in the relative contribution of DF-TEQWHO-AVian and
PCB'TEQWHO-Avian to total TEQWHO-Avian in the tissues of BKF. In the RA, PCBS

accounted for the greatest proportion of total TEQWHO-Avian in both eggs and nestlings of
BKF (64% and 62%, respectively). Conversely, PCDD/DFS accounted for a majority of
the total TEQW'HO-A,,.,., in both eggs and nestlings of BKF collected from the SA (59%
and 72%, respectively). However, PCB-TEQSW”on,“an were dominated by PCB-126,

PCB-77, and PCB-81 in each sampling location.

221

Risk assessment
Estimates of dietary exposure based on the geometric means of measured
concentrations of DF-TEQSWHO-A,.ian and total TEQSWHO-Avian from Tittabawassee River

reaches were greater than the diet-based NOAEC TRV, which resulted in HQs that were
greater than 1.0 (Figure 5.3). Along the Tittabawassee River, maximum calculated HQs,

based on the most conservative parameters of 100% site use and 95% upper conﬁdence
limit (95% UCL) of measured DF-TEQS\)VHO./man and total TEQSW’HQ-AVian, ranged
between 5.0 and 20. Dietary exposure-based estimates based on the 95% UCL of
measured concentrations of DF-TEQSWH0_AVian and total TEQSWHO-AVian compared to the

diet-based LOAEC TRV resulted in HQS less than 1.0, regardless of reach.

Hazard quotients based on the 95% UCL of the geometric mean concentrations of

DF-TEQSWH()-AVian (n=19) and total TEQSWH0-A,,ian (n=18) in BKF eggs collected from
the RA or SA were not greater than 1.0 when compared to either the NOAECEGG or
LOAECEGG TRV (Figure 5.4). Based on predicted probabilistic distributions of
concentrations of either DF-TEQSWHO_A,.ian or total TEQSWHO-Aviam 100% of the

expected cumulative percent frequencies were below both the NOAECEGG and

LOAECEGG TRVS (Figure 5.5).

Population condition
During the 2005-2007 breeding seasons, a total of 37 nest chambers were excavated
and monitored for reproductive effort and success. Of the studied nests, 27 contained

eggs upon excavation while the remaining 10 were found with nestlings. The number of

222

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eggs per clutch ranged from six to seven for nests in both the RA (n=6) and SA (n=16).
The most frequently occurring number of eggs per clutch was six in nests in the RA and
seven in nests in the SA. A statistically signiﬁcant difference in mean (iSD) clutch size
of BKF was noted between the RA (6.3:t0.52) and SA (6.9i0.34; p=0.0477).

Hatching success was not correlated with concentrations of ZPCDD/DFS in eggs of
BKF for clutches with both data points measured. Eggs from the RA had lesser
concentrations of ZPCDD/DFS but similar hatching success compared to the downstream

SA, which resulted in a slightly positive correlation coefﬁcient (R=0.29, p=0.3576).

Similarly, no signiﬁcant correlation existed between concentrations of DF—TEQSWHO-
Avian and hatching success (R=0.24, p=0.4592; Figure 5.6). A limited number of clutches

(n=3) had both hatching success and concentration of total TEQSWHO-Avians precluding a
correlation assessment.

Hatching success was similar between the RA and SA, regardless of whether hatching
success calculations included extrinsic inﬂuences or were done on a per egg or per nest
basis (Table 5.4). When only examining eggs that were incubated until hatch, hatching
success was 100% (16/ 16) and 95% (40/42) for eggs from the RA and SA, respectively.
When clutches which were partially incubated and abandoned were included, hatching
success was reduced to 73% (16/22) and 73% (40/55) for eggs from the RA and SA,
respectively. If all eggs were included, hatching success was further reduced yet
remained similar between areas, at 38% (16/42) and 31% (40/128) for eggs from the RA
and SA, respectively. The percentage of nests that successfully hatched at least one egg

was also similar between study areas, at 75% (3/4) and 78% (7/9) for the RA and SA,

226

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227

Table 5.4. Parameters of reproductive effort and success of belted kingﬁshers breeding

along the Chippewa and Tittabawassee river ﬂoodplains during 2005-2007.

 

 

Reproductive Parameter RA SA
Incubation Period
No. of nests 7 20
No. successfulal 3 7
No. failedb 4 13
No. abandoned (unknown) 1 2
No. abandoned (disturbance)C 2 5
No. abandoned (2006 rain) 1 6
No. of eggs 42 128
No. incubated 22 55
No. successful 16 40
No. failed 0 2
No. abandoned (unknown) 6 13
Excluded from hatching success calculations
No. abandoned (disturbance) 11 27
No. abandoned (2006 rain) 7 39
No. missing 6 7
No. of eggs hatched
No. of eggs hatched/eggs incubated 16/22 (73%) 40/55 (73%)
No. of eggs hatched/egg fully incubated 16/ 16 (100%) 40/42 (95%)
Nestling Period
No. of nests 5 15
No. successful61 3 12
No. where all nestlings disappeared I 2
No. abandoned 1 1
No. of nestlings 28 83
No. abandoned/missing/depredated 12 29
No. dead in nest 0 1
No. reaching 25 d age 15 52
No. of nestlings ﬂedged
No. of nestlings ﬂedged (excluding 15/15 (100%) 52/53 (98%)
abandoned/missing/depredated) _
No. of nestling ﬂedged/all nestlings 15/28 (54%) 52/83 (63%)

 

 

a successful == at least one egg hatched or at least one nestling ﬂedged

b .
failed = no eggs hatched
C See Methods for deﬁnition of disturbance and 2006 rain event

228

respectively, excluding those nests which were abandoned as result of human disturbance
or the 2006 rain event.

Measures of ﬂedging success were also similar between the RA and SA (Table 5.4).
Mortality of nestlings that was likely unrelated to exposure to contaminants, such as those
that were abandoned, depredated, or were otherwise missing from the nest, was excluded
from the calculation of ﬂedging success. By excluding these nestlings, ﬂedging success
was 100% (15/ 15) and 98% (52/53) for nestlings from the RA and SA, respectively.
When abandoned, depredated, and missing nestlings Were included, ﬂedging success was '
reduced yet remained similar between the study location, at 54% (15/28) and 63%
(52/83) for nestlings from the RA and SA, respectively.

A number of band recoveries were observed during the present study. One adult BKF
breeding in the RA and fourteen adults breeding in the SA were banded. Band return
monitoring occurred for one and two breeding seasons in the RA and SA, respectively. A
breeding pair banded during the 2006 breeding season was recaptured during the 2007
breeding season. The pair was nesting in a new burrow located within 500m of the
previous year’s burrow. Another adult female banded during the 2006 breeding season
was recaptured in 2008 at a new burrow with a new mate. All band recoveries were

made in the SA of adults initially banded in the SA.

Discussion
BKF as receptor species
Belted kingﬁshers possess desirable characteristics for use in assessing ecological

effects of contaminants in the environment. As inhabitants of varied riverine and

229

lacustrine habitats with a widespread distribution across North America, BKF are a
species likely to be present in areas of aquatic-based contamination. They are also a
highly charismatic species that is easily recognized by the public. Additional attributes
that make BKF ideal for assessing bioaccumulative compounds include their high trophic
status, deﬁned foraging territories, and great food consumption rate. However, BKF can
be a difﬁcult species to study in the ﬁeld. In contrast to box-nesting passerine species
that are often used as receptor species, BKF are a subterranean nesting species and have
not been shown to nest in artificial structures. As such, time and effort must be expended
in locating and accessing nests for monitoring and sample collection, which is difﬁcult
without causing undue disturbance to the nest. Furthermore, BKF often appear to be
nest-site limited (Cornwell 1963; Davis 1982), which may make it difﬁcult to obtain the
necessary sample size in a short-term study. In spite of these potential disadvantages,
BKF should still be considered an optimal piscivorous avian species to study because
their deﬁned foraging range allows for greater spatial resolution of their dietary exposure
compared to other species commonly used, such as the great blue heron (Ardea herodias;

GBH) or black-crown night heron (Nycticorax nycticorax).

Dietary exposure

The observed trends in concentrations and relative contribution of individual
congeners of PCDD/DPS in prey items of BKF were consistent with those observed for
dietary items and tissues of other receptor species studied within the SA (Coeﬁeld et al.
2010; Fredricks et al. 2010a; Fredricks et al. 2010b; Seston et al. 2010b; Seston et al.

2010a; Zwiemik et al. 2008b). The pattern observed is likely the result of the

230

downstream movement of this historical contamination from the upstream source. When

expressed as dietary exposure, the relative proportion of DF-TEQWHOAvian to .

ZPCDD/DF for ADme of BKF was 18% in the RA, while in the SA it ranged from 61%

to 63% and 52% for reaches along the TR and SR, respectively. The greater percentages
observed along the TR and SR were due to prey items from those reaches having greater
relative TCDD potency of ZPCDD/DF, from primarily TCDF and secondarily from
2,3,4,7,8-PeCDF.

Greater concentrations of PCBs in forage fish from the SR compared to those from
the TR are most likely due to historical PCB contamination in areas downstream of the
TR, including the Saginaw River and Saginaw Bay (Kannan et al. 2008) as well as some
lesser sources upstream of the RA. Anadromous ﬁshes from Saginaw Bay can move up

into the TR, but a dam located in Midland, MI may impede the movement of these ﬁsh

into the RA. As such, the total TEQWHOAvian ADme was only 13-20% greater than the

DF-TEQqumm ADDpot along the reaches of the TR, but was 62% and 65% greater in

the RA and SR, respectively.

Tissue exposure

Consistency in the relative contribution of individual congeners to PCDD/DF
observed between prey items of BKF and both eggs and nestlings of BKF is likely
resultant of site-speciﬁc dietary exposure. TCDF was the predominant congener present
in prey items sampled from the SA, with 2,3,4,7,8-PeCDF also having a signiﬁcant
presence. In eggs and nestlings of BKF, the overall congener proﬁle was similar to that

observed in prey items, although there was a change in the predominant congener from

231

TCDF to 2,3,4,7,8-PeCDF. This pattern was also observed in other species foraging in
the SA, including GBH, great horned owls, and several passerine species (Coeﬁeld et al.
2010; Fredricks et al. 2010a; Fredricks et al. 2010b; Seston et al. 2010b; Seston et al.
20103; Zwiemik et al. 2008b). Bioaccumulation of TCDF from prey items in Forster’s
terns and herring gulls has been reported to be negligible (Braune and Norstrom 1989;
Kubiak et al. 1989). This may be due to preferential metabolism of TCDF as reported in
various avian species that have been exposed to mixtures of AhR-active compounds
(Elliott et a1. 1996; Kubota et al. 2005; Senthil Kumar et al. 2002). While data on avian
toxicokinetics from controlled laboratory studies are limited, marmnalian studies have
shown the rate of metabolism of TCDF to be elevated with increased concentrations of
dioxin—like compounds and the subsequent induction of cytochrome P450 1A1 and/or
1A2 while 2,3,4,7,8—PeCDF is preferentially sequestered in the liver (van den Berg et al.
1994b; Zwiemik et al. 2008a). This difference in toxicokinetics is a likely explanation
for the observed difference in the relative contribution of the two furan congeners from
diet to tissue observed for BKF in the present study.

The relative contribution of individual PCDD/DF congeners was consistent between
eggs and nestlings of BKF, which supports the conclusion that tissue exposure was the
result of site-speciﬁc dietary exposure. Nestlings concentrations are generally considered
more representative of local contamination as they are conﬁned to the nest and rely on
food brought to them by adults (Neigh et al. 2006; Olsson et al. 2000). During breeding
and nestling rearing, adult BKF have a foraging range that is typically limited to 0.92 to
2.9 km proximal to the nest (Davis 1982; Mazeika et al. 2006). Although the BKF is

generally a seasonally migratory species in this area of its range (Hamas 1994), because

232

PCDD/DF congener patterns of contamination in sediments and prey are similar to that
observed in tissues of BKF, it can be concluded that the contamination measured in eggs

and nestlings of BKF are site-speciﬁc.

Risk characterization
Based on the selected TRVS and calculated exposures for BKF foraging along the TR,
there is the potential for adverse effects on hatching success, but the likelihood is small

given the conservative nature of the assessment. For instance, based on the 95% UCL of

measured concentrations of DF-TEQSWHO_A,.;an and total TEQSWH0-AV;an compared to the
NOAECDIET TRV, maximum calculated HQs along the Tittabawassee River ranged
between 1.0 and 20 and were less than 1.0 when based on LOAECDlET. Since the HQs

derived from comparisons of the NOAECDIET and LOAECDIET to the range of BKF

dietary exposure bracket an HQ of 1.0, it is possible that none of the dietary exposures
predicted here exceed the actual effect threshold concentration, given that the actual
effects threshold lies somewhere between these values. However, it is important to note
that forage ﬁsh, which were the primary component of the BKF diet, were composite
samples of all individual ﬁsh captured at each location per sampling period. While this
provided an accurate estimate of the central tendency of concentration estimates in forage
ﬁsh, it limited the characterization of the variability of concentrations among all forage
ﬁsh at each sampling location. Thus, the upper end estimates of dietary exposure
- presented here, and the risk associated with that exposure, may be an underestimate.

Despite the conservative nature of the selected TRVS, egg-based HQs were less than
1.0 when based on NOAECs and the 95% UCL concentrations of DF-TEQth/H().,a,vian and

233

total TEQSWH0-AVian, which is consistent with the other lines of evidence that concluded

that adverse effects would not be expected. The few other studies which have monitored
contaminants in tissues of BKF have generally reported concentrations of total PCBs and
methylmercury (Baron et al. 1997; Evers and Lane 2000; Fecteau 2008; Heinz et al.

1984; Hudson River Natural Resource Trustees 2004). A study of BKF along the PCB-
contarninated Hudson River reported concentrations of total TEQSWHOAWan in eggs of

BKF (Custer et al. 2010) that ranged from 96 to 5200 ng/kg with a geometric mean of
620 ng/kg. In comparison, eggs of BKF collected along the Hoosic, Mohawk, and

Connecticut Rivers, considered to be the reference area in that study, contained

concentrations of total TEQSWHO_A,.,an ranging from 50 to 870 ng/kg with a geometric

mean of 120 ng/kg. PCBs accounted for 98% and 92% of the total TEQSWHO-Avian

concentration in eggs of BKF collected along the Hudson River and associated reference
areas, respectively. No association was found between these concentrations of total

TEQs and reproductive impairment. When the predicted frequency distribution for
concentrations of DF—TEst}.;O_A,.;an or total TEQSWHQ.AVian in BKF eggs from the RA

and SA of the present study were compared to the Hudson River geometric mean of 620
ng/kg, 0% and <1% of eggs of BKF exceeded either value, respectively (Figure 5). This
comparison to a ﬁeld-derived NOAEC also suggests that adverse effects on reproductive
success would not be expected along the TR. There is some uncertainty in comparing
these two studies given that TEQs along the Hudson River were primarily from PCBs
whereas those along the TR were primarily from PCDFs and secondarily from PCDDS.

In addition to dioxin-like compounds, the potential for adverse effects of site-speciﬁc

EDDX concentrations was assessed. Previous studies have found mean DDE

234

concentrations in eggs of great blue heron ranging from 0.086 to 16 mg/kg had no
adverse impacts on breeding success (Blus et al. 1980; Harris et al. 2003; Laporte 1982;
Thomas and Anthony 1999). However, concentrations of DDE greater than 8 mg/kg
were associated with decreases in clutch size and productivity in black—crowned night
herons (Henny et al. 1984). For reproductive effects in bald eagles, a threshold
concentration of 6 mg DDE/kg has been suggested (Elliott and Harris 2001). Eggs of
BKF collected from the TR assessment area were well below these thresholds, with mean
concentrations of ZDDX <l.0 mg/kg. A lack of adverse reproductive effects was
observed with similar concentrations of EDDX in BKF collected from the Hudson River
assessment area (Custer et al. 2010). These observations suggest that egg ZDDX
concentrations reported in the present study would not cause adverse effects on
reproductive success of BKF along the TR ﬂoodplain.

Despite elevated exposure of BKF to PCDD/DPS in the SA compared to the RA,
there were no discernable differences in parameters of reproductive success between the
two locations. The average clutch size observed here is similar to that reported in other
studies, which is reported to vary between ﬁve to eight eggs, with six or seven eggs
occurring most often (Bent 1940; Hamas 1975). A similar percentage of nests that
successfully hatched at least one egg has been reported in a previous study (Custer et al.
2010). When inﬂuences of human disturbance and depredation were excluded, rates of
hatching and .ﬂedging success observed here were comparable to those reported in
previous studies (Custer et al. 2010; Davis 1980; Hamas 1975). Both the RA and SA had
a rObust population of American mink (Mustela vison) (Zwiemik et al. 2008b), which

was the primary predator of nestlings of BKF in the assessment area.

235

Uncertainty assessment

For the study described herein, exposure and effects assessments were conducted
based on both dietary and tissue exposure, facilitating a multiple lines of evidence
approach. 1 Discrepancies in the outcomes of the two methodologies were examined in
terms of the uncertainties and associated conservative bias in order to weight each line
apprOpriately. In the assessment of dietary exposure, uncertainty is often introduced by
making assumptions in regards to dietary composition and the proportion of the diet that
is collected from the area of concern. For tissue-based assessments, uncertainty is often
associated with site-use characteristics as well as spatial, temporal, and individual
variations in exposure. Many of the uncertainties associated with both exposure
assessments were greatly reduced in the present study through the collection of extensive
site-speciﬁc data. Uncertainties may also be present in the effects assessment as a result
of comparing across species and exposure scenarios due to limited toxicological data
available for comparison. For the two exposure and effects assessments of BKF
presented here the least uncertainty is associated with the tissue-based approach, which is
derived from a more robust and directly linked data set.

Uncertainties in the quantiﬁcation of the dietary exposure of BKF were minimized
through the identiﬁcation of a site-speciﬁc dietary composition and the subsequent
collection of the appropriate prey items. However, unlike the sampling of BKF tissues,
prey items were collected at fewer time points, thus it is possible that there were temporal
variations in the contaminant burdens of prey items that were not captured. The

techniques used to determine dietary composition did integrate over the duration of

236

incubation of eggs through nestling rearing but did not directly identify prey items
consumed during egg laying. To investigate the potential inﬂuence of dietary
composition, the predicted dietary dose was calculated with the determined site-speciﬁc
diet (90.2% ﬁsh, 5.4% crayﬁsh, 4.4% frog), 3 literature-based diet (58% ﬁsh, 41%
crayﬁsh; (Salyer and Lagler 1949)) and an equally opportunistic diet (33% of each ﬁsh,
crayﬁsh, and frog). Use of the site-speciﬁc diet consistently resulted in the greatest
predicted dietary dose, thus making this the most conservative estimate. Additional
conservative bias was introduced by the assumption of 100% site use as the BKF was
very likely forced to forage off-site in non-ﬂowing water during times of extreme water
turbidity associated with heavy rains (Davis 1980; Hamas 1975). Therefore, while
dietary exposure is thoroughly characterized with an extensive amount of site-speciﬁc
data, there remains a considerable potential to overestimate exposure due to the
aforementioned factors.

Laboratory studies of effects from dietary exposure to PCDD/DFs are limited for
avian species. The TRV selected for use in this assessment was derived from a study in
which adult hen ring-necked pheasants (Phasianus colchicus) were dosed with TCDD
through intraperitoneal (IP) injection (Nosek et al. 1992). There are a few factors that
make the use of this TRV conservative in determining the potential hazard of these
compounds to BKF. First, in Nosek et al. (1992) the hens were exposed to TCDD via IP
injection rather than a true dietary-based exposure, potentially resulting in greater
exposure efﬁciency than that of a true dietary exposure. Dietary exposure also allows
greater potential for metabolism and excretion prior to reaching target tissues.

Additionally, the ring-necked pheasant is not a piscivorous species, but a galliforrnes,

237

which are generally considered to have greater sensitivity to dioxin-like compounds
(Brunstrdm 1988; Brunstrom and Reutergardh 1986; Powell et a1. 1996; Powell et al.
1997a). Variability in sensitivity among species has been attributed to differences in the
amino acid sequence of the ligand binding domain of the AhR (Head et al. 2008;
Karchner et al. 2006) with ring-necked pheasants having the construct which classiﬁes
them as moderately sensitive whereas the BKF construct classiﬁes them as insensitive.
Furthermore, dosing in the Nosek study was done with only a single congener (TCDD),
whereas the BKF being assessed are exposed to a complex mixture of dioxin-like

compounds. Although previously assumed to be equitable in toxicity when normalized
using TEFSWHQ-AVian, ﬁndings from recent studies investigating the relative potencies of

individual PCDD/DP congeners suggest that this may not hold true in every exposure
scenario (Herve et al. 2010b; Herve’ et al. 2010a).

Tissue-based exposure assessments integrate exposure characteristics including site
use, dietary composition, and contaminant metabolic potentials that may vary over time
and space. Directly measuring residue levels in tissues requires the collection of fewer
data and eliminates any uncertainties associated with modeling the exposure (Leonards et 1
al. 2008). To illustrate this, a screening level ERA of ﬁsh-eating birds exposed to
PCDD/DPS along the TR predicted there to be much greater concentrations in eggs than

what were directly measured in site-speciﬁc samples (Galbraith Environmental Sciences
LLC. 2003). In the screening-level assessment, concentrations of DF-TEQSWHO_A,.ian in

eggs of ﬁsh-eating birds were predicted through the application cf biomagniﬁcation

factors (BMFs) to concentrations measured in ﬁsh collected from the TR. Concentrations

in eggs 'were predicted to be 1031 ng‘DF-TEQSWHogvian/kg, wet wt, based on the mean

238

concentration of DF'TEQSWHQ-AViaD in collected ﬁsh. Eggs of BKF collected from the
SA had a geometric mean concentration of 84 ng DF-TEQSWHO-Avian/kg, wet wt, and a

maximum of 260 ng DF-TEQSWH0_AV,an/kg, wet wt. Factors which may have resulted in

the overestimation of exposure in the screening level assessment include; collection of
unrepresentative prey items, inaccurate dietary composition, assumptions regarding
foraging range and site use, and inaccurate BMFs. The direct quantiﬁcation of residues
in both eggs and nestlings of BKF collected from within the TR ﬂoodplain eliminated
these uncertainties from the assessment, allowing for a more accurate characterization of
tissue-based exposure.

The TRVS selected to assess the risk to BKF based on egg concentrations present
along the TR ﬂoodplain also have assOciated uncertainties. The values based on the
double-crested cormorant egg injection studies were considered to be appropriate for use
in the present study because they were derived from another avian piscivore.
Furthermore, both the double-crested cormorant and BKF have similar AhR ligand
binding domain constructs, which suggests the two species will respond similarly to
dioxin-like compounds (Head et al. 2008; Karchner et al. 2006). Although the species
appear to be similar in sensitivity to dioxin-like compounds, the route of exposure in the
aforementioned double-crested cormorant studies and the BKF being assessed along the
TR is different. Residues in the BKF eggs are maternally deposited rather than being
artiﬁcially introduced as in the study conducted with the double-crested cormorant eggs.
This difference in route of exposure adds uncertainty when using this TRV in the risk

assessment (Heinz et al. 2009; Hoffman et al. 1996).

239

Hazard quotients are generally used in screening for the potential for risk and not for
direct risk characterization. That HQs based on the predicted dietary exposure exceeded
1.0 in some instances is not incompatible with HQs less than 1.0 for the tissue-based
assessment or for the inability of the study to identify adverse effects on individual and
population condition, despite the extensive dataset. HQs greater than 1.0 indicate
exposures that exceed an effect concentration given the present dataset and conservative
biases associated with uncertainties. Additionally, the HQ approach may be conservative
when extrapolating from individuals to populations, as they are usually based on
laboratory exposures and do not account for compensatory mechanisms associated with
wild populations (Fairbrother 2001). Conversely, the limited number of endpoints
generally measured in laboratory studies may not capture an effect that could have
implications at the population level. Thus, there is uncertainty inherent in the HQ
approach, meaning that HQ values greater than 1.0 do not necessarily translate into
population-level or ecologically relevant adverse effects (Blankenship et al. 2008).

The hazard assessment based on tissue exposure resulted in HQs less than 1.0. This
held true even when the assessment was made more conservative by including an
interspecies uncertainty factor of three. The robust dataset for contaminant
concentrations in eggs of BKF also allowedfor the use of a probabilistic assessment,
which gives a more complete picture of the exposure distribution and associated risk.
The ﬁndings of the probabilistic assessment were consistent with the prediction of
minimal risk and with ﬁeld measured parameters associated with individual and

population condition.

240

The accuracy and certainty of the assessment hazard in the HQ approach and the
assessment of risk in the probabilistic approach are each dependent on the quality
applicability of the dose response curves used to establish effect concentrations. The
greater certainty in both of exposure and effects assessment based on contaminant
concentrations in eggs of BKF within the TR ﬂoodplain gives greater weight to the

conclusions drawn from that line of evidence.

Conclusions

The BKF proved to be a useful receptor Species in the assessment of PCDD/DFs
along the TR. Given its relatively small, distinct foraging range, the exposure of BKF
has a greater spatial resolution than most other ﬁsh-eating birds. Belted kingﬁsher
foraging along the TR were predicted to be more greatly exposed to PCDD/DFS than
those in associated RAs. This spatial trend was consistent with what was noted in
concentrations of PCDD/DFs in both eggs and nestlings of BKF. Overall, the weighted
exposure and effects assessment suggests that BKF along the TR are exposed to
signiﬁcant concentrations of PCDD/DFs, however, presently there is minimal potential
for adverse effects. This conclusion was consistent with direct ﬁeld measures of
individual and. population condition. The greatest remaining uncertainty for the
assessment of risk to BKF along the TR include temporal and species variability in forage
ﬁsh contaminant burdens and uncertainties associated with threshold effects

concentrations for both dietary exposure and internal doses (tissue concentrations).

241

Acknowledgements

The authors would like to thank all staff and students of the Michigan State
University ~ Wildlife Toxicology Laboratory ﬁeld crew; namely M. Nadeau, D.
Hamman, W. Folland, E. Koppel, and L. Williams. We gratefully acknowledge M. Fales
for his design and fabrication of specialized equipment which was pivotal to the success
of this research. Additionally, we would like to recognize M. Kramer and N. Ikeda for
their assistance in the laboratory. The following people were key to the success of the
project, J. Dastyck and S. Kahl of the USFWS -— Shiawassee National Wildlife Refuge;
Saginaw County Parks and Recreation Commission for access to lmennan Park;
Tittabawassee Township Park Rangers for access to Tittabawassee Township Park and
Freeland Festival Park; Tom Lenon for access to Chippewa Nature Center; and Michael
Bishop of Alma College for guidance as our Master Bander. We also acknowledge the
greater than 50 cooperating landowners throughout the study area who made our research
possible through granting us access to their property. Funding was provided through an
unrestricted grant from The Dow Chemical Company, Midland, Michigan to J. Giesy and

M. Zwiemik of Michigan State University.

Animal Use

All aspects of the present study that involved the use of animals were conducted using
the most humane means possible. To achieve that objective, all aspects .of the present
study were performed following standard operation procedures (Protocol for belted
kingﬁsher monitoring and tissue collection 05/07-071-00; Field studies in support of TR

ERA 03/04-042-00; Protocol for ﬁsh sampling 03/04-043-00) approved by Michigan

State University’s Institutional Animal Care and Use Committee (IACUC).. All of the
necessary state and federal approvals and permits (Michigan Department of Natural
Resources Scientiﬁc Collection Permit SC1254 for BKF/SC permit for ﬁsh
(Zwiemik)/SC permit for amphibians (Zwiemik); USFWS Migratory Bird Scientiﬁc
Collection Permit MBlOOOO62-0; and subperrnitted under US Department of the Interior

Federal Banding Permit 22926) are on ﬁle at MSU-WTL.

243

Supplemental Information

Table 5.5. Concentrations of 2378-TCDD equivalents (TEQsa) in eggs and
nestlings of belted kingﬁsher collected during 2005~2007 ﬁ'om within the
Chippewa and Tittabawassee River ﬂoodplains, Midland, MI, USA. Valuesb
(ng/kg w) are given as the geometric mean (sample size) over the 95%

conﬁdence interval.
BKF Eggs BKF Nestlings
\\ \
_ RA SA RA SA
PCDD-TEQWHO-Aviant W

 

 

10(6) 19 (19) 3.1 (5) 8.5 (12)

3.9—28 15—26 l.6-—6.0 0.74—3.2
PCDF'TEQWHO-Aviand

4.1 (6) 59 (19) 1.8 (5) 86 (12)

2.4—7.0 40—87 1.2—2.9 65—1 10
non-ortho PCB-TEQWHO-A,.~,,f

26(3) 58 (11) 5.6 (4) 32(9)

1 1—63 37—89 3.5—9.0 12—81
mono-ortho PCB-TEQ\;\/H().,.(vianf

2.5 (3) 5.5 (11) 040(4) 2.6 (9)

6.9—14 3.2—9.3 0.26—0.63 1.3—5.1
T0181 TEQWHo-Avian

47(3) 170(11) 9.7(4) 170(9)

10-210 120-240 8.0—12 110—250

 

X
a TEQwH0-Avian were calculated based on the 1998 avian WHO TEF values
b

Values have been rounded and represent only two signiﬁcant ﬁgures
b PCDD-TEQWHO-A,,,,,= summation ofthe TEQS of individual PCDD
congeners
c PCDF-TEQWHO-A,.ian= summation of the TEQS of individual PCDF
congeners
° non-ortho PCB-TEQWHO-Avian= summation ofthe TEQS ofindividual non-
ortho substituted PCB congeners
fmono-ortho PCB-TEQWH0-AV,-an= summation of the TEQS of individual
mono-ortho substituted PCB congeners

244

Table 5.6. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin congeners
in eggs and nestlings of belted kingﬁsher collected during 2005-2007 within the
Chippewa and Tittabawassee River ﬂoodplains, Midland, MI, USA. Valuesa (ng/kg
w) are given as the arithmetic meanb i 1 SD over the range.

 

 

 

 

BKF Eggs BKF Nestlings
RA SA RA SA
Chemicalc "=6 n=1 9 n= 5 n=1 2
2378-TCDF 0.82:0.46 7.4:t3.8 0.91:1:0.42 26:1:10
0.22—1.3 1.6—14 0.52—1.6 12—49
23478-PeCDF 3.4m2.1 67m57 0.87i0.4l 63138
1.3—7.2 6.7—180 O.61—1.6 12—140
12378-PeCDF 0.30m0.14 5.1:t2.9 0.27m0.026 15:1:7.9
0.21—0.66 0.88—10 0.24—0.29 6.3—34
3ND 1ND 2ND
234678-HxCDF 0.46m0.l7 1.5il.l 2.33:2.2
0.21-0.66 0.47—4.6 0.15—0.16 0.69—7.0
1ND 3ND
123789-HxCDF 0.54—0.99
6ND 19ND 5ND IOND
123678-HxCDF 1.0i0.48 4.1:l:3.1 5.65:5.6
0.49-1.6 1.1—12 0.25—2.7 1.0—18
_ 1ND 3ND
123478-HxCDF 1.4:t0.95 16:1:15 0233:0096 26127
O.62—2.7 2.1—57 0.17-0.34 3.0—77
1ND 2ND
1234789-HpCDF ' 0.52:t0.18 3,713.7
0.28—0.77 0.40—9-9
6ND -11ND 5ND 7ND
1234678—HpCDF 0.31i0.077 2.8:t1.3 0.16 182t31
O.22—O.39 1.2-5.8 1.2-99
2ND 1ND 4ND
12346789-OCDF ‘ 0.84:1:0.60 0.16 25i44
O.36—2.5 , 0.34—130
6ND 5ND 4ND 2ND
2378-TCDD 85:97 13:1:9.l 2.0:1:0.97 5.6:t1.6
1.8—27 3.8—46 0.76—3.2 2.5—8.2

245

 

Table 5.6 (cont’d)

12378-PeCDD 66:57 10173 l.4i0.85 3.1:l:0.94
1.7—~17 2.5-29 0.69—29 1.3—4.6
123789-HxCDD 0.78:t0.29 1.1i0.68 0.86i0.80
0.43—1.1 0.21—2.23 0.23—1.9
2ND 5ND 5ND 8ND
123678-HxCDD 5.5i3.8 7.0i5.3 1511.3 3.3136
1.1—11 2.1—21 0.33—3.6 1.4—14
123478-HxCDD 1.6:t0.50 1981.4 0.531049 072350.45
0.70—1.9 0.60—5.2 0.23—1.1 O.34-l.6
1ND 3ND 2ND 5ND
1234678-HpCDD 1.610.98 3711.4 0.60i0.43 15329
~ 077—35 2.0—6.2 032—12 096—97
1ND
12346789-OCDD 3.4:I:2.6 9,015.2 0.67:t0.088 85:190
1.4—8.3 1.8—20 052—075 1.1—650

 

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of arithmetic
means

C TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzofuran; HxCDF=
hexachlorodibenzofuran; HpCDF = heptachlorodibenzoﬁiran; OCDF =
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p-dioxin; OCDD = octachlorodibenzo-p-dioxin

246

of belted kingﬁsher collected during 2005-2007 within the Chippewa and
Tittabawassee River ﬂoodplains, Midland, MI, USA. Valuesa (pg/kg w) are
given as the arithmetic meanb :t 1 SD over the range.

 

 

 

 

 

 

 

 

 

 

 

 

BKF Eggs BKF Nestlings
RA SA RA SA
ChemicalC n=3 n=1 1 n=4 "=9
—‘ \\I
PCB 77 009980.035 0.348025 004380.016 0.498098
0074—014 0.1 1—0.96 0028—0061 0039—31
PCB 81 004780014 0.1180073 0010800077 04481.1
0036—0063 0025—023 00037—0021 0.01 1—3.5
PCB 126 0.18:I:0.066 04580.54 0.026800067 0.178018
0.10—0.23 0.085—2.0 0020—0034 _ 0053—064
, 0025800006 0011800038
PCB 169 4 004180.035
00018—0030 0014—013 00030—00040 00063—0015
1ND 2ND 5ND
PCB 105 1681 1 43849 2.08065 26843
4.6—27 7.6—180 1.3—2.7 5.0—140
PCB 1 14 1.28066 4.0845 01680.045 2083.1
O.52-—1.8 0.79—17 0.10—0.20 0.44—10
PCB118 56845 100891 6681.4 58866
19—1 10 26—350 5.2—8.2 14—230
PCB 123 1.280075 2.7:l:3.2 0.1480072 1.6826
053-20 032—1 1 0067—023 0.23—8.5
PCB156 6.7842 14813 1.08028 5.78.5.2
2.4—11 3.6—47 O.82—l.4 1.7—19
PCB 157 1.7:l:l.2 3.3831 0.2180071 1.2812
054—29 076—1 1 0.17—0.32 0.36—4.3
PCB167 4.0832 6.8863 0.4980058 2781.9

247

Table 5.7 (cont’d)

 

1.2—7.5 2.0—23 043—057 O.89—7.4
PCB 189 06980.51 1,481.2 0.09880020 0.608049
0.24—1.3 045—45 0.081—O.13 O.19—1.8
2,4’-DDT° 0.738099 0.368018 007580045
0035—19 0084—062 0012—0031 0016—015
2ND 1ND
29,484)ng 60%530 6908500 1708150 220878
200—1300 120-1500 82—390 120—350
4, 420131 6481.7 20816 0268032 1.3812
5.0—8.3 4.0—55 00055—069 0.028-3.1

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of

arithmetic means
° DDT = dichloro-diphenyl-trichloroethane
d DDE = dichloro-diphenyl-dichloroethylene

248

Table 5.8. Concentrations of seventeen 2,3,7,8-substituted furan and dioxin
congeners in soils collected during 2006 from nest chambers of belted kingﬁshers
within the Chippewa and Tittabawassee River ﬂoodplains, Midland, MI, USA.

Valuesa (ng/kg w) are given as the arithmetic meanb 8 1 SD over the range.

 

 

Reference Area Study Area
Chemicalc n=2 "=12
2378-TCDF 320084300
7.5—8.1 12—15000
23478-PeCDF 1000i1500
0.64—0.73 1.7-5100
12378-PeCDF 130081900
1.1—1.3 2.4—6800
234678-HxCDF 0.22 1108130
0.21—430
1ND
123789-HxCDF 22827
1.2—90
2ND 2ND
12367 8-HxCDF 0.19 1808240
0.29—850
1ND
123478-HxCDF 99081300
0.22—0.47 1.1—4500
12347 89-HpCDF 0.10 82896
0.13—320
1ND
1234678-HpCDF 8003:1200
0.20—0.94 2.4—4300
12346789-OCDF 1500962400
0.35—1.1 3.6—8600
2378-TCDD 0.13 7.4811 .
0.26—40
1ND
12378-PeCDD 0.21 96812
0.22—36
1ND 2ND

249

Table 5.8 (cont’d)

 

 

 

 

 

 

123789-HXCDD 0.25 15820
0.27-55
1ND 2ND
l23678-HxCDD 0.30 28843
0.34-130
1ND
l23478-HXCDD 0.12 7.5892
0.13—23
1ND 4ND
1234678-HpCDD 5208810
0.73—3.0 4.3—2300
12346789-OCDD 6300810000
5.0—1 9 32-29000 ‘
SUM PCDD/DPS 16000816000

1 7—35 62-50000

 

 

 

a Values have been rounded and represent only two signiﬁcant ﬁgures

b Concentrations below limit of detection treated as zero in the calculation of
arithmetic means

c TCDF = tetrachlorodibenzofuran; PeCDF = pentachlorodibenzoﬁiran; HxCDF=
hexachlorodibenzoﬁiran; HpCDF = heptachlorodibenzofuran; OCDF =
octachlorodibenzofuran; TCDD = tetrachlorodibenzo-p-dioxin; PeCDD =
pentachlorodibenzo-p-dioxin; HxCDD = hexachlorodibenzo-p-dioxin; HpCDD =
heptachlorodibenzo-p-dioxin;' OCDD = octachlorodibenzo-p-dioxin

250

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CHAPTER 6

Conclusions

Rita Marie Seston

259

Comparison of Receptor Species

Selection of receptor species is a critical step of the problem formulation process of
ecological risk assessments. The great blue heron (Ardea herodias; GBH) and belted
kingﬁsher (Ceryle alcyon; BKF) are often selected as receptor species for ecological risk
assessments to determine the potential for effects of persistent bioaccumulative
contaminants in aquatic environments. Both species have similar diets, however their
exposure characteristics including metabolic rates, foraging range, and time on-site may
differ. Furthermore, issues of maximal obtainable sample size and level of effort
required to reach sample size requirements will be different between the two species
depending on site characteristics such as size and available habitat types. Matching
receptor characteristics to site characteristics and the risk management goals ultimately
determines the utility of any ecological risk assessment. Herein, the exposures of BKF
and GBH breeding within a river system contaminated with dioxin-like compounds are
compared. To provide risk assessors with information on species characteristics and
suitability, a number of exposure characteristics were compared between species to assess
spatial trends, spatial resolution, data interpretability, and required level of effort.

As a result of historical chemical production and management of associated wastes,
the Tittabawassee (TR) and Saginaw Rivers (SR) downstream of Midland, MI, USA
contain elevated concentrations of polychlorinated dibenzofurans (PCDFs) and dibenzo-
p-dioxins (PCDDS). The persistence of these compounds in the environment, combined
with their toxicity and potential to bioaccumulate, has led to concern over the potential

exposures of wildlife species foraging in the Tittabawassee and Saginaw River

260

ﬂoodplains. To understand both exposure and associated effects, ﬁeld studies were
conducted over several years in support of a large scale, site-speciﬁc ecological risk
assessment (ERA) utilizing multiple lines of evidence. To most accurately assess an
ecosystem as complex as the Tittabawassee River ﬂoodplain, the site-speciﬁc ERA
employed many different receptor species representative of various feeding guilds and
exposure pathways.

This dissertation details the exposure of GBH and BKF foraging and breeding within
the TR ﬂoodplain to the seventeen 2,3,7,8-substituted PCDD/DFs (ZPCDD/DFS). Due to
the tendency of PCDD/DPS to bioaccumulate through trophic transfer, species located
near the top of food webs have the greatest potential for exposure. Additionally, it has
been determined that avian species are more sensitive to the toxic effects of PCDD/DPS
than are mammals, ﬁsh, amphibians, or reptiles. As maximally-exposed, potentially
highly-sensitive species, the GBH and BKF were selected as top-tier avian receptors of
an aquatic-based exposure pathway.

Although both the BKF and GBH share an exposure pathway and several
characteristics which make both of them desirable species to investigate exposure and
potential effects of PCDD/DFs, there are a few key differences. As top aquatic food web
predators, both the BKF and GBH have a relatively great potential for exposure as a
result of biomagniﬁcation. Each species has a widespread distribution throughout North
America, making them potential receptors for ecological assessments of aquatic
environments in various localities. Many of the differences between the two species are
related to their difference in body size. Although the dietary composition is similar

between the two species, there may be differences in the size-class of prey consumed. As

261

the smaller species, BKF have a greater food consumption rate normalized to body
weight. Additionally, BKF defend deﬁned territories and foraging areas proximal to its
nest burrow (0.92 to 2.9 km (Davis 1982; Mazeika et al. 2006)) in contrast to the GBH
which defends territories within 3.2 to 6.5 km of the breeding colony (Dowd and Flake
1985; Marion 1989; Peifer 1979; Thompson 1978).

Concentrations of PCDD/DPS and PCBs were quantiﬁed in the dietary components of
GBH and BKF. The diet of both the GBH and BKF is primarily composed of ﬁsh, but
also includes crayﬁsh and amphibians. Consistent spatial trends in concentrations of
ZPCDD/DFs and DF-TEQ07H0-g,,ian were observed for each type of prey.
Concentrations were least in the RA, greater in the UTR, and greatest in the LTR.
Concentrations of PCDD/DFs in prey items from the SR were intermediate to those from
the RA and UTR (Table 6.1). Mean concentrations of ZPCDD/DFs‘ were 8-, 18-, and 1-
fold greater in frogs, 21-, 49-, and 7-fold greater in crayﬁsh, and 36—, 71-, and 23-fold
greater in forage ﬁsh collected from the UTR, LTR, and SR than those from the RA,

respectively. Mean concentrations of DF-TEQWHO-A,.,an were 17-, 52-, and 3-fold greater

in frogs, 43-, 120-, and 20-fold greater in crayﬁsh, and 130-, 240-, and 64-fold greater in
forage ﬁsh collected from the UTR, LTR, and SR when compared to the RA,
respectively. Greater fold differences observed in TEQWH0-AV;,m are a result of prey
items from the TR and SR having greater relative TCDD potency of ZPCDD/DF,
primarily from 2,3,7,8—TCDF and secondarily from 2,3,4,7,8-PeCDF. Concentrations of

EPCDD/DFS in estimated diets of BKF were 29-, 59—, and l6-fold greater and 32-, 65-,

and 20-fold greater for GBH compared to the RA for UTR, LTR, and SR, respectively

262

Table 6.1. TEQ\;VH().AV,-ana in prey items of belted kingﬁsher and great blue heron
collected during 2004-2006 from the Chippewa, Tittabawassee, and Saginaw River
ﬂoodplains, Midland, MI, USA. Valuesb (ng/kg lw) are given as the geometric mean

TEQWHO—Avian and sample size in parentheses (11) over the 95% conﬁdence interval and

range (min-max).

 

Egg
TCDD

TCDF

4-PeCDF
ZPCDD/DF
PCB426

DF'TEQWl-IO-Avian

T0131 TEQVVHO-

Avian
Smash
TCDD
TCDF
4-PeCDF

 

Reference Upper Lower Saginaw
Area Tittabawassee Tittabawassee River
22 65 52 24
16—31 53—79 42—63 20—28
(3.4-82) (13—800) (1 1——3 10) (14—37)
22 790 2800 110

15—32 460—1300 2200—3 700 98—130
(4.3—230) (21—19000) (310—68000) (84-150)
9.9 270 970 51
7.3—1 3 170~430 720—1300 33~—79
(2.3—3 8) ( 1 5—3 800) (1 90—23 000) (30—410)
430 (29) 3300 (51) 7700 (55) 520 (12)
330—540 2200—4900 6000—9900 440—620
(98—1500) (260—68000) (1000—160000) (410—1 100)
N/A N/A 240 (4) N/A
130-440
(140-310)

80 (29) 1300 (51) 4200 (55) 230 (12)
60—1 10 840—2100 3200—5500 180—280
(19——330) (79—-23000) (600—99000) (160—600)

N/ A C WA 24000 (4) N/A
5200—110000
(9800—97000)
1 1 61 86 30
2.3~55 45—83 55-1 30 12~77
(3.9—75) (36—93) (35—180) 03—120)
41 2900 8300 1500
11-160 1200—6800 6400—11000 910—2400
(1 5—260) (490—9400) (5700—15000) (680—3 300)
15 620 2000 220

263

Table 6.1 (cont’d)

 

XPCDD/DF

PCB-126

DF‘TEQWl-lO-Avian

Total TEano-

Avian

Forage Fish
TC DD

 

TCDF

4-PeCDF

ZPCDD/DF

PCB-126

DF-TEQWHO-Avian

Total TEQWHQ-

Avian

:29—81
619—120)
480(5)
130—1800

(180—2500)

N/A

88 (5)
20-390
(27——590)
N/A

—(2)
410-553
—(2)

186—11
~(2)

_(11—4.8)

—-(2)

(80—120)
-(2)

(60—64)
—(2)

(18—21)
-(2)

(52-52)

310—1300 1600~2400 140-340
(170—1500) (1300—2600) (1 10—570)
9900(7) 24000(8) 3200(8)
5800~17000 19000—30000 2000—5300
(5800—28000) (17000-35000) (1600-7700)
_. d (2) — (1) N/A
(330~430) (360)
3800(7) 11000(8) 1800(8)
1700—8400 8600-14000 1100—2900
(780—11000) (8000—18000) (84084200)
-(2) -—(1) 1464
(4900—8200) (14000)
—(2) 100 57
— 46—220 20-170
86—89 (37—200) (33—140)
._(2) 3800 970
~ 1700—8500 450_2100
1800—2000 (1400—7500) (690—1932)
—(2) 620 140
— 1802mm 4844)
(280—410) (160~2000) (85—390)
—(2) 6900(5) 2200(4)
— 2700-17000 750—6300
(3300—3600) (2200—15000) (1200~5700)
—(2) 1300(5) 2000(4)
8 320—5200 710—5700
(460—760) (500—8100) (1100—4800)
—(2) 4600(5) 1200(4)
- 1900—11000 540—2800
(2400—2500) (1700—10000) (840—2600)
—(2) 5500(5) 3400(4)
— 2300—13000 930—12000
(2900—3100) (2000—11000) (1600-10000)

264

Table 6.1 (cont’d) ‘
a TEQWH0-Avian were calculated based on the 1998 avian WHO TEF values

b Values have been rounded and represent only two signiﬁcant ﬁgures
c N/A = no samples collected from this location

d Geometric mean and conﬁdence intervals not calculated for sites with fewer than 3
samples. These sites were not included in reach comparisons.

265

(Table 6.2). To facilitate direct comparisons between prey and receptor tissues, all
concentrations are reported on a lipid-normalized basis.

A spatial trend similar to that observed for prey items was observed in concentrations
of measured residues in eggs and nestlings of BKF but not in those of GBH.
Concentrations of ZPCDD/DFS were 4- and 5-fold greater in BKF eggs collected from
the UTR and LTR compared to those collected in the RA, respectively (Table 6.3). This
trend was most pronounced for concentrations of 2,3,4,7,8-PeCDF in eggs of BKF
collected from the UTR and LTR, which were 10- and 23-fold greater, respectively, than
in eggs from the RA. Concentrations of ZPCDD/DFs were greater in BKF nestlings from
downstream study areas compared to the RA, however the spatial trend between the UTR
and LTR was not observed due to elevated concentrations of octachlorodibenzo-p-dioxin
(OCDD) in BKF nestlings collected from the UTR (Table 6.4). Similar to eggs, the
concentrations of the site-speciﬁc 2,3,4,7,8-PeCDF in BKF nestlings exhibited the same

spatial gradient, being 33- and 52-fold greater in the UTR and LTR, respectively, than
those collected from the RA. Concentrations of DF-TEQsWHOAVian were 5- and 7—fold

greater in BKF eggs and 10- and 15-fold greater in BKF nestlings from the UTR and

LTR, respectively, compared to those collected in the RA. Concentrations of
ZPCDD/DFS and DF-TEQSWHOAvian in both eggs and livers of nestling GBH were
similar among all studied rookeries (Table 6.3, 6.4).

Consistent spatial trends in relative contributions of DF-TEQSWHO-Avian and PCB-

TEQSWHO-AVian to total TEstt/HO_A,,,an between prey items and receptor tissues were
observed. In ﬁsh collected from the RA, PCBs accounted for the greatest proportion of

the total TEQSWHQ_AVian (63%), compared to those collected from the Tittabawassee

266

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Table 6.3. Total concentrations of 2,3,7,8-substituted furan and dioxin (ZPCDD/DF)

and TEQWHO-Avian a in eggs of belted kingﬁsher and great blue heron collected during
2005- 2007 from the Chippewa, Tittabawassee and Saginaw River ﬂoodplains,

Midland, MI, USA. Valuesb (ng/kg 1w) are given as the geometric mean (n) over the

95% conﬁdence intervals and range in parentheses.

 

TCDD

TCDF

4-PeCDF

ZPCDD/DF

PCB-126

DF-TEQS

Total TEQS

 

 

 

BKF Eggs GBH Eggs
RA. UTR LTR SA
72(6) 200(8) 130(11) 180(24)
22—240 110—370 94—190 130—230
(22—470) (64—680) (53—270) (so—1200)
586(6) 77(8) 100(11) 57(24)
418—19 49—120 63—180 36—88
(3.7—18) (24—130) (20—220) (5.0—320)
41(6) 410(8) 940(11) 240(24)
19—90 160-1100 610—1500 180—320
(14—120) (88—2600) (420—2200) (82-920)
410(6) 1600(8) 2100(11) 1100(24)
170-990 900—2900 1500—2900 920-1300
(140—1300) (820-5500) (1200—3900) (690-4000)
3700(3) 8900(4) 3200(7) 14000(18)
1700—7800 2000—40000 2100—5100 9400—20000

(2900—5200)

210(6)
77-550
(67—920)

750(3)
130—4400
(360—1500)

(3400—30000)

960(8)
490—1900
(380—3800)

3500(4)
1200—9700
(1600—7800)

(1200—5900)

1400(11)
980—2000
(750—2800)

2000(7)
1400—3000.
(1100—3900)

(2700-39000)

670(24)
530—850
(320—2800)

3700(18)
2700—5000
(850-8000)

 

a TEQwHo-Avian were calculated based on the 1998 avian WHO TEF values.

b . .
Values have been rounded and represent only two srgmﬁcant ﬁgures.

268

 

 

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270

River where the greatest proportion of total TEQSWHQ-AVian were attributed to DF-
TEQSWHO-Avian (81% and 85% for the UTR and LTR, respectively). Concentrations of
total TEQSWHO-Avian were dominated by PCB-TEQSWH0-AV;M in ﬁsh collected from the
Saginaw River (61% PCB- TEQWHO-Avian). In both eggs and nestlings of BKF from the
RA, PCB-TEQSWHO-AVian accounted for the greatest proportion of total TEQSWHQ-AVian
(69% and 61%, respectively). Conversely, DF-TEQSWHO-Av;an accounted for the greatest

proportion of total TEQSWHOAVian in both eggs and nestlings of BKF collected from the

UTR (51% and 73%, respectively) and LTR (64% and 66%, respectively). This

observation was in contrast to eggs and livers of nestling GBH collected from along the

TR, in which PCB-TEQSWHOAVian (79% and 81%, respectively) accounted for the

greatest proportion of total TEQSWHOAW. PCB'TEQSWHQ-AVian were dominated by

congeners PCB-126, PCB-77, and PCB-81 in all reaches.

The observed spatial trends in concentrations of PCDD/DPS and PCBs in prey of
BKF and GBH were consistent with those observed for prey and tissues of other receptor
species studied within the SA (Coeﬁeld et a1. 2010; Fredricks et al. 2010a; Fredricks et
al. 2010b; Seston et al. 2010). The occurrence of greater concentrations of PCDD/DFs
further downstream in the TR is likely the result of the downstream movement of this
historical contamination from the source near Midland, MI. The observation that
concentrations of PCBs were greater. in forage ﬁsh from the SR compared to those from
the TR is consistent with historical PCB contamination in areas downstream of the TR,
including the Saginaw River and Saginaw Bay (Kannan et al. 2008) as well as some

lesser sources upstream of the RA. Anadromous ﬁshes from Saginaw Bay can move up

271

into the TR, but there is a dam which may impede the movement of these ﬁsh into the
RA. Furthermore, comparisons between ﬁsh and the more spatially conﬁned frogs and
crayﬁsh (Hazlett er a1. 1974; Martof 1953), found that ﬁsh generally contained greater
concentrations of PCBs than frogs or crayﬁsh collected from within the same reach. This
also suggests that ﬁsh may be acquiring PCBs from areas other than the TR.
Additionally, there may be incidental PCB inputs from the urbanized Midland, MI area.
The observed difference in the relative concentrations of PCBs between BKF and
GBH may also be due to differences in contribution from the SR or offsite sources.
Belted kingﬁsher nesting along the TR have a lesser proportion of the total TEQS in their
tissues attributable to PCBs than GBH nesting in similar areas. This could be from a
difference in size of foraging ranges between the two species. To estimate how closely
concentrations in receptor tissues match those in site-speciﬁc prey, spatial trends in the
ratio of the ubiquitious PCB-126 to the site-speciﬁc 2,3,4,7,8-PeCDF in diet and receptor
tissues were compared. In prey, listed in descending order, this ratio is greatest in the
RA, then the SR, and nearly equal in the UTR and LTR. Both egg and nestlings of BKF
exhibited a similar spatial trend of this ratio. In contrast, the ratio in GBH egg and
nestling liver collected from rookeries along the TR most closely resembles that of prey
and BKF tissues from the RA. This dissimilarity could be a result the two species
utilizing different foraging grounds, with BKF exhibiting a greater spatial resolution in
their exposure. Additionally, as a larger species, it is possible that GBH target larger ﬁsh
that are more likely to be moving longer distances and integrating contaminants over a

larger spatial scale.

272

Thus, it can be seen that either BKF or GBH may be better suited as a receptor
species based on site characteristics and assessments goals. Sites with areas of
widespread contamination may be better represented by GBH, which integrate exposure
over a large area. Additionally, many GBH nests may be monitored simultaneously
within a single breeding colony to assess any potential impacts on reproductive and
population health. By contrast, assessments of sites with a point-source may be better
understood by employing BKF, whose exposure has a greater spatial resolution and may
be used to evaluate concentration gradients. Each Species has qualities. that make it a
good receptor as long as consideration is given to the different exposure characteristics of

each species and which would be most appropriate for a given assessment.

Overall Conclusions

This dissertation details the exposure of GBH and BKF foraging and breeding within
the TR ﬂoodplain to the seventeen 2,3,7,8-substituted PCDD/DPS and the associated risk
of adverse effects. EXposure was determined in both modeled dietary exposure and
measured concentrations in receptor speciﬁc tissues. Simultaneously, site-speciﬁc
population health of both BKF and GBH was monitored. Each measure was then
combined in the framework of a multiple lines of evidence assessment to minimize
uncertainty in conclusions. Over the course of this assessment, BKF and GBH breeding
along the TR successfully reproduced despite elevated exposure to PCDD/DFs.

Concentrations of PCDD/DFs in tissues of both BKF and GBH were greater at
downstream SAs compared to upstream RAs. For BKF, this spatial trend was seen in

eggs and nestlings. In GBH, this comparison was only available for blood plasma of

273

adults foraging in the river channel, as no rookeries were found in the RA. Elevated
exposures in the SA was composed primarily of 2,3,4,7,8-pentachlorodibenzofuran
(PeCDF) and secondarily 2,3,7,8-tetrachlorodibenzofuran (TCDF), which are speciﬁc to
the historical releases from DOW. Although GBH eggs and nestlings from within the SA
also contained these site-speciﬁc congeners, their concentrations were lesser than
expected based on predicted dietary exposures. Additionally, eggs and nestlings of GBH
had greater relative concentrations of PCBs compared to BKF, suggesting they may be
foraging partially off-site.

Site-specific dietary- and tissue-based exposures for both the BKF and GBH were

compared to toxicity reference values (TRVs) to estimate the potential for adverse
effects. Egg-based exposures based on DF-TEQSWHO,A,.,an and total TEQSWHOAWan for

both the BKF and GBH within the SA were at or below the no observed adverse effect
concentration (N OAEC). Dietary-based exposures did exceed associated NOAECs for
both species but were below the lowest observed adverse effect concentrations
(LOAECs). It is important to note that there is more uncertainty associated with the
prediction of dietary exposures compared to measured tissue concentrations due to the
greater number of assumptions that are necessary. Additionally, the best available TRVS
for dietary exposure were based on intraperitoneal injections which likely overestimate
exposures compared to actual ingestions. The prediction of minimal risk of adverse
effects from the dietary- and tissue-based assessments is in agreement with observations
of population health of BKF and GBH along the TR. Therefore, the overall conclusion of
the research presented here is that the populations of BKF and GBH breeding along the

TR are not at risk despite elevated concentrations of PCDD/DFS in the diet and tissues.

274

From this research, several areas were identiﬁed which should be addressed in future
work. Firstly, a better understanding of the foraging habits of GBH would greatly '
reduce the uncertainty currently associated with their predicted dietary exposure and may
explain the presence of greater proportion of PCBs in their tissues. The location and
study of a breeding rookery(s) in the RA would also add great value to the current study,
allowing direct comparison to regional background concentrations in tissues and
reproductive success. Furthermore, although there is great certainty in the site-speciﬁc
exposure assessments conducted here, further research needs to be done to expand upon
the TRVS available to assess risk in ecological studies. There is currently a large gap in
toxicological data based on ecologically relevant endpoints, particularly those based on
dietary exposure. Lastly, because the research presented here failed to ﬁnd any
signiﬁcant population-level or individual-based effects as a result of contaminant
exposure, any remediation actions suggested or taken along the TR should primarily
focus on the habit-based effects of those actions. This is of particular concern for BKF
which is often limited by the presence of suitable banks for nest sites. The alteration or
removal of that habitat would likely have immediate and long-lasting adverse effects on

the breeding population of BKF along the TR.

275

 

 

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MICHIGAN STATE UNIVERSITY LIBRARIES

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