PHASE 2 REPORT- REVIEW COPY
FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
DECEMBER 1999
For
U.S. Environmental Protection Agency
Region II
and
U.S. Army Corps of Engineers
Kansas City District
Book 1 of 1
TAMS Consultants, Inc.
Menzie-Cura & Associates, Inc.
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.
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
1 REGION 2
' 29° BROADWAY
? NEW YORK, NY 10007-1866
December 29, 1999
To All Interested Parties:
The U.S. Environmental Protection Agency (USEPA) is pleased to release the baseline Ecological
Risk Assessment - Future Risks in the Lower Hudson River, which evaluates the future ecological
risks in the Lower Hudson River (Federal Dam to the Battery in New York City) posed by PCBs in
sediments at the Hudson River PCBs Superfund site, in the absence of remediation. This report,
called the Ecological Risk Assessment (ERA) Addendum, is a companion volume to USEPA's
August 1999 baseline Ecological Risk Assessment (ERA), which evaluated the current and future
ecological risks in the Upper Hudson River and the current ecological risks in the Lower Hudson
River. The ERA Addendum is posted on USEPA's website for the Hudson River PCBs
Reassessment Remedial Investigation/Feasibility Study (Reassessment RI/FS) at
www.epa.gov/hudson.
The ERA Addendum is part of Phase 2 of the Reassessment RI/FS for the Hudson River PCBs
Superfund site. The ERA Addendum, together with the August 1999 ERA, will help establish
acceptable exposure levels for use in developing remedial alternatives in the Feasibility Study, which
is Phase 3 of the Reassessment RI/FS.
USEPA will accept comments on the ERA Addendum until January 28, 2000. Comments should
be marked with the name of the report and should include the report section and page number for
each comment. Comments should be sent to:
Alison A. Hess, C.P.G.
USEPA Region 2
290 Broadway - 19th Floor
New York, NY 10007-1866
Attn: Hudson River ERA Addendum Comments
USEPA will hold a Joint Liaison Group meeting to discuss the findings of the ERA Addendum on
January 11, 2000, at 7:30 p.m. at the Sheraton Hotel, 40 Civic Center Plaza, Poughkeepsie, New
York. The meeting is open to the general public. Notification of the meeting was sent to Liaison
Group members, interested parties, and the press several weeks prior to the meeting.
During the public comment period, USEPA will hold an availability session to answer questions
from the public regarding the ERA Addendum. The availability session will be held from 6:30 to
8:30 p.m. on January 18, 2000 at Sheraton Hotel, 40 Civic Center, Poughkeepsie, New York.
Internet Address (URL) • http://www.epa.gov
Recycled/Recyclable • Printed with Vegetable Oil Based Inks on Recycled Paper (Minimum 25% Postconsumer)
-------
If you need additional information regarding the ERA Addendum or the Reassessment RI/FS in
general, please contact Ann Rychlenski, the Community Relations Coordinator for this site, at (2 1 2)
637-3672.
Sincerely yours,
/Richard L. Caspe, Director
"V Emergency and Remedial Response Division
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PHASE 2 REPORT- REVIEW COPY
FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
DECEMBER 1999
For
U.S. Environmental Protection Agency
Region II
and
U.S. Army Corps of Engineers
Kansas City District
Book 1 of 1
TAMS Consultants, Inc.
Menzie-Cura & Associates, Inc.
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Table of Contents
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PHASE 2 REPORT - REVIEW COPY
FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
Page
TABLE OF CONTENTS i
LIST OF TABLES xii
LIST OF FIGURES xix
1.0 INTRODUCTION ' 1
1.1 Purpose of Report 1
1.2 Report Organization 1
2.0 PROBLEM FORMULATION 3
2.1 Site Characterization 3
2.2 Contaminants of Concern 3
2.3 Conceptual Model 3
2.3.1 Exposure Pathways in the Lower Hudson River Ecosystem 4
2.3.2 Ecosystems of the Lower Hudson River 4
2.3.3 Exposure Pathways 6
2.3.3.1 Aquatic Exposure Pathways 6
2.3.3.2 Terrestrial Exposure Pathways 6
2.4 Assessment Endpoints 6
2.5 Measurement Endpoints (Measures of Effect) 7
2.6 Receptors of Concern 9
2.6.1 Fish Receptors 9
2.6.2 Avian Receptors 10
2.6.3 Mammalian Receptors 10
2.6.4 Threatened and Endangered Species 10
2.6.5 Significant Habitats 11
3.0 EXPOSURE ASSESSMENT 13
3.1 Quantification of PCB Fate and Transport: Modeling Exposure Concentrations 13
3.1.1. Modeling Approach 14
3.1.1.1 Use of the Farley Model 14
3.1.1.2 Use of FISHRAND 16
3.1.1.3 Comparison to the March 1999 Farley Model (1987-1997) 17
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3.1.1.4 Comparison Between Model Output and Sample Data 19
3.1.1.5 Comparison of White Perch Body Burden between the Farley
Model (Using Upper River Loads from HUDTOX) and
FISHRAND 21
3.1.1.6 Comparison Between FISHRAND Output and Sample Data ... 21
3.1.2. Model Results 23
3.1.2.1 Farley Model Forecast Water Column and Sediment
Concentrations 23
3.1.2.2 Farley Model Forecast Fish Body Burdens 23
3.1.2.3 FISHRAND Forecast Fish Body Burdens 24
3.1.3 Modeling Summary 24
3.2 Exposure Point Concentrations 24
3.2.1 Modeled Water Concentrations 25
3.2.2 Modeled Sediment Concentrations 25
3.2.3 Modeled Benthic Invertebrate Concentrations 25
3.2.4 Modeled Fish Concentrations 26
3.3 Identification of Exposure Pathways 27
3.3.1 Benthic Invertebrate Exposure Pathways 27
3.3.2 Fish Exposure Pathways 27
3.3.3 Avian Exposure Pathways, Parameters, Daily Doses, and Egg
Concentrations 27
3.3.3.1 Summary of ADDExpected, ADD95%UCL, and Egg Concentrations
for Avian Receptors 28
3.3.4 Mammalian Exposure Pathways, Parameters, and Daily Doses 29
3.3.4.1 Summary of ADDExpected and ADD95%UCL for Mammalian
Receptors 29
4.0 EFFECTS ASSESSMENT 31
4.1 Selection of Measures of Effects 31
4.1.1 Methodology Used to Derive TRVs 33
4.1.2 Selection of TRVs 36
5.0 RISK CHARACTERIZATION 37
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TABLE OF CONTENTS
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5.1 Evaluation of Assessment Endpoint: Benthic Community Structure as a Food
Source for Local Fish and Wildlife 38
5.1.1 Do Modeled PCB Sediment Concentrations Exceed Appropriate
Criteria and/or Guidelines for the Protection of Aquatic Life and
Wildlife? 38
5.1.1.1 Measurement Endpoint: Comparisons of Modeled Sediment
Concentrations to Guidelines 38
5.1.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria
and/or Guidelines for the Protection of Aquatic Life and Wildlife? .... 40
5.1.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations of PCBs to Criteria 40
5.2 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Local Fish Populations 40
5.2.1 Do Modeled Total and TEQ-Based PCB Body Burdens in Local Fish
Species Exceed Benchmarks for Adverse Effects on Forage Fish
Reproduction? 40
5.2.1.1 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for Forage
Fish 40
5.2.1.2 Measurement Endpoint: Comparison of Modeled PCB TEQ
Fish Body Burdens to Toxicity Reference Values for Forage
Fish 41
5.2.1.3 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for Brown
Bullhead 41
5.2.1.4 Measurement Endpoint: Comparison of Modeled TEQ Basis
Fish Body Burdens to Toxicity Reference Values for Brown
Bullhead 41
5.2.1.5 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for White
and Yellow Perch 42
iii TAMS/MCA
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5.2.1.6 Measurement Endpoint: Comparison of Modeled TEQ Basis
Body Burdens to Toxicity Reference Values for White and
Yellow Perch 42
5.2.1.7 Measurement Endpoint: Comparison of Modeled Tri+PCB
Fish Body Burdens to Toxicity Reference Values for Large-
mouth Bass 43
5.2.1.8 Measurement Endpoint: Comparison of Modeled TEQ Based
Fish Body Burdens to Toxicity Reference Values for Large-
mouth Bass 43
5.2.1.9 Measurement Endpoint: Comparison of Modeled Tri+ PCB
Fish Body Burdens to Toxicity Reference Values for Striped
Bass 43
5.2.1.10 Measurement Endpoint: Comparison of Modeled TEQ Based
Fish Body Burdens to Toxicity Reference Values for Striped
Bass 43
5.2.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria
and/or Guidelines for the Protection of Aquatic Life and Wildlife? .... 44
5.2.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations of PCBs to Criteria 44
5.2.3 Do Modeled PCB Sediment Concentrations Exceed Appropriate
Criteria and/or Guidelines for the Protection of Aquatic Life and
Wildlife? 44
5.2.3.1 Measurement Endpoint: Comparisons of Modeled Sediment
Concentrations to Guidelines 44
5.2.4 What Do the Available Field-Based Observations Suggest About the
Health of Local Fish Populations? 45
5.2.4.1 Measurement Endpoint: Evidence from Field Studies 45
5.3 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Lower Hudson River Insectivorous
Bird Populations (as Represented by the Tree Swallow) 46
5.3.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Insectivorous
Birds and Egg Concentrations Exceed Benchmarks for Adverse Effects
on Reproduction? 46
iv TAMS/MCA
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5.3.1.1 Measurement Endpoint: Modeled Dietary Doses on a Tri+
PCB Basis to Insectivorous Birds (Tree Swallow) 46
5.3.1.2 Measurement Endpoint: Predicted Egg Concentrations on a Tri+
PCB Basis to Insectivorous Birds (Tree Swallow) 46
5.3.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed on a TEQ Basis to Insectivorous Birds (Tree
Swallow) 47
5.3.1.4 Measurement Endpoint: Predicted Egg Concentrations
Expressed on a TEQ Basis to Insectivorous Birds (Tree
Swallow) 47
5.3.2 Do Modeled Water Concentrations Exceed Criteria for Protection
of Wildlife? 47
5.3.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations to Criteria for the Protection of
Wildlife 47
5.3.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Insectivorous Bird Populations? 47
5.3.3.1 Measurement Endpoint: Evidence from Field Studies 47
5.4 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth and Reproduction) of Lower Hudson River Waterfowl
Populations (as Represented by the Mallard) 48
5.4.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Waterfowl and
Egg Concentrations Exceed Benchmarks for Adverse Effects on
Reproduction? 48
5.4.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ PCBs
to Waterfowl (Mallard) 48
5.4.1.2 Measurement Endpoint: Predicted Egg Concentrations of
Tri+ PCBs to Waterfowl (Mallard) 48
5.4.1.3 Measurement Endpoint: Modeled Dietary Doses of TEQ-
Based PCBs to Waterfowl (Mallard) 49
5.4.1.4 Measurement Endpoint: Predicted Egg Concentrations of TEQ-
Based PCBs to Waterfowl (Mallard) 49
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TABLE OF CONTENTS
BOOK 1 of 1
5.4.2 Do Modeled PCB Water Concentrations Exceed Criteria for the
Protection of Wildlife? 49
5.4.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria 49
5.4.3 What Do the Available Field-Based Observations Suggest About the
Health of Lower Hudson River Waterfowl Populations? 50
5.4.3.1 Measurement Endpoint: Observational Studies 50
5.5 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Hudson River Piscivorous Bird
Populations (as Represented by the Belted Kingfisher, Great Blue Heron, and
Bald Eagle) 50
5.5.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Piscivorous
' Birds and Egg Concentrations Exceed Benchmarks for Adverse Effects
on Reproduction? 50
5.5.1.1 Measurement Endpoint: Modeled Dietary Doses of Total
PCBs for Piscivorous Birds (Belted Kingfisher, Great Blue
Heron, Bald Eagle) 50
5.5.1.2 Measurement Endpoint: Predicted Egg Concentrations
Expressed as Tri+ to Piscivorous Birds (Eagle, Great Blue
Heron, Kingfisher) 51
5.5.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed as TEQs to Piscivorous Birds (Belted Kingfisher,
Great Blue Heron, Bald Eagle) 52
5.5.1.4 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed as TEQs to Piscivorous Birds (Belted Kingfisher,
Great Blue Heron, Bald Eagle) 52
5.5.2 Do Modeled Water Concentrations Exceed Criteria for the Protection
of Wildlife? 53
5.5.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria 53
5.5.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Piscivorous Bird Populations? 53
5.5.3.1 Measurement Endpoint: Observational Studies 53
vi TAMS/MCA
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TABLE OF CONTENTS
BOOK 1 of 1
5.6 Evaluation of Assessment Endpoint: Protection (i.e., Survival and Repro-
duction) of Local Insectivorous Mammal Populations (as represented by
the Little Brown Bat) 54
5.6.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to
Insectivorous Mammalian Receptors Exceed Benchmarks for Adverse
Effects on Reproduction? 54
5.6.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Insectivorous Mammalian Receptors (Little Brown Bat) ... 54
5.6.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Insectivorous Mammalian Receptors (Little
Brown Bat) 54
5.6.2 Do Modeled Water Concentrations Exceed Criteria for Protection of
Wildlife? 55
5.6.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife ... 55
5.6.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Insectivorous Mammalian Populations? 55
5.6.3.1 Measurement Endpoint: Observational Studies 55
5.7 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Omnivorous Mammal Populations (as represented
by the Raccoon) 56
5.7.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to
Omnivorous Mammalian Receptors Exceed Benchmarks for Adverse
Effects on Reproduction? 56
5.7.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Omnivorous Mammalian Receptors (Raccoon) 56
5.7.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Omnivorous Mammalian Receptors (Raccoon) 56
5.7.2 Do Modeled Water Concentrations Exceed Criteria for Protection of
Wildlife? 56
5.7.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 56
vii TAMS/MCA
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BOOK 1 of 1
5.7.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Omnivorous Mammalian Populations? 57
5.7.3.1 Measurement Endpoint: Observational Studies 57
5.8 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Piscivorous Mammal Populations (as represented
by the Mink and River Otter) 57
5.8.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to
Piscivorous Mammalian Receptors Exceed Benchmarks for Adverse
Effects on Reproduction? 57
5.8.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Piscivorous Mammalian Receptors (Mink, River Otter) ... 57
5.8.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Piscivorous Mammalian Receptors (Mink, River
Otter) 58
5.8.2 Do Modeled Water Concentrations Exceed Criteria for the Protection
of Piscivorous Mammals? 59
5.8.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife ... 59
5.8.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Mammalian Populations? 59
5.8.3.1 Measurement Endpoint: Observational Studies 59
5.9 Evaluation of Assessment Endpoint: Protection of Threatened and
Endangered Species 60
5.9.1 Do Modeled Total and TEQ-Based PCB Body Burdens in Local
Threatened or Endangered Fish Species Exceed Benchmarks for
Adverse Effects on Fish Reproduction? 60
5.9.1.1 Measurement Endpoint: Inferences Regarding Shortnose
Sturgeon Population 60
5.9.2 Do Modeled Total and TEQ-Based PCB Body Burdens/Egg
Concentrations in Local Threatened or Endangered Species Exceed
Benchmarks for Adverse Effects on Avian Reproduction? 61
5.9.2.1 Measurement Endpoint: Inferences Regarding Bald Eagle and
Other Threatened or Endangered Species Populations 61
viii TAMS/MCA
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5.9.3 Do Modeled Water Concentrations Exceed Criteria for the Protection
of Wildlife? 61
5.9.3.1 Measurement Endpoint: Comparisons of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 61
5.9.4 Do Modeled Sediment Concentrations Exceed Guidelines for the
Protection of Aquatic Health? 61
5.9.4.1 Measurement Endpoint: Comparisons of Modeled
Sediment Concentrations to Guidelines 61
5.9.5 What Do the Available Field-Based Observations Suggest About the
Health of Local Threatened or Endangered Fish and Wildlife
Species Populations? 62
5.9.5.1 Measurement Endpoint: Observational Studies 62
5.10 Evaluation of Assessment Endpoint: Protection of Significant Habitats 62
5.10.1 Do Modeled Total and TEQ-Based PCB Body Burdens/Egg Concen-
trations in Receptors Found in Significant Habitats Exceed Bench-
marks for Adverse Effects on Reproduction? 63
5.10.1.1 Measurement Endpoint: Inferences Regarding Receptor
Populations 63
5.10.2 Do Modeled Water Column Concentrations Exceed Criteria for the
Protection of Aquatic Wildlife? 63
5.10.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife .... 63
5.10.3 Do Modeled Sediment Concentrations Exceed Guidelines for the
Protection of Aquatic Health? 64
5.10.3.1 Measurement Endpoint: Comparison of Modeled Sediment
Concentrations to Guidelines for the Protection of Aquatic
Health 64
5.10.4 What Do the Available Field-Based Observations Suggest About the
Health of Significant Habitat Populations? 64
5.10.4.1 Measurement Endpoint: Observational Studies 64
6.0 UNCERTAINTY ANALYSIS 67
6.1 Conceptual Model Uncertainties 67
ix TAMS/MCA
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6.2 Toxicological Uncertainties 67
6.3 Exposure and Modeling Uncertainties 70
6.3.1 Natural Variation and Parameter Error 70
6.3.2 Model Error 70
6.3.2.1 Uncertainty in the Farley Model 70
6.3.2.2 Uncertainty in FISHRAND Model Predictions 71
6.3.3 Sensitivity Analysis for Risk Models for Avian and Mammalian
Receptors 73
7.0 CONCLUSIONS 75
7.1 Assessment Endpoint: Benthic Community Structure as a Food Source for
Local Fish and Wildlife 75
7.2 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth,
and Reproduction) of Local Fish (Forage, Omnivorous, and Piscivorous)
Populations 75
7.3 Assessment Endpoint: Protection and Maintenance (i.e.,Survival, Growth,
and Reproduction) of Hudson River Insectivorous Bird Species (as Represented
by the Tree Swallow) 76
7.4 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth
and Reproduction) of Lower Hudson River Waterfowl (as Represented by
the Mallard) 76
7.5 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth,
and Reproduction) of Hudson River Piscivorous Bird Species (as Represented
by the Belted Kingfisher, Great Blue Heron, and Bald Eagle) 77
7.6 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Insectivorous Mammals (as represented by the Little Brown Bat) 78
7.7 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Local Omnivorous Mammals (as represented by the Raccoon) 78
7.8 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Local Piscivorous Mammals (as represented by the Mink and River Otter) .... 79
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BOOK 1 of 1
7.9 Assessment Endpoint: Protection of Threatened and Endangered Species .... 79
7.10 Assessment Endpoint: Protection of Significant Habitats 80
7.11 Summary 80
REFERENCES 81
APPENDICES
APPENDIX A - Conversion from Tri+ PCB Loads to Dichloro through Hexachloro
Homologue Loads at the Federal Dam
APPENDIX B - Effects Assessment
xi TAMS/MCA
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LIST OF TABLES:
2-1 Lower Hudson Assessment Endpoints, Receptors, And Measures
2-2 Lower Hudson River Endpoints and Risk Hypotheses
2-3 Lower Hudson River Significant Habitats
3-1 Summary of Conversion for the Di through Hexa Homologues
3-2 Ratio of Striped Bass to Largemouth Bass Concentrations
3-3 Sum of Monthly Average Loads Over the Troy Dam (kg)
3-4a Relative Percent Difference Between FISHRAND Results and Measured Fish Levels in the
Lower Hudson
3-4b Relative Percent Difference Between FISHRAND Results and Measured Spottail Shiner
Levels in the Lower Hudson
3-5 Summary of Tri+ Whole Water Concentrations from the Farley Model and TEQ-Based
Predictions for 1993-2018
3-6 Summary of Tri+ Sediment Concentrations from the Farley Model and TEQ-Based
Predictions for 1993-2018
3-7 Organic Carbon Normalized Sediment Concentrations Based on USEPA Phase 2 Dataset
3-8 Summary of Tri+ Benthic Invertebrate Concentrations from the FISHRAND Model and
TEQ-Based Predictions for 1993-2018
3-9 Spottail Shiner Predicted Tri+ Concentrations for 1993-2018
3-10 Pumpkinseed Predicted Tri+ Concentrations for 1993 - 2018
3-11 Yellow Perch Predicted Tri+ Concentrations for 1993 - 2018
3-12 White Perch Predicted Tri+ Concentrations for 1993 - 2018
3-13 Brown Bullhead Predicted Tri+ Concentrations for 1993-2018
3-14 Largemouth Bass Predicted Tri+ Concentrations for 1993 - 2018
3-15 Striped Bass Predicted Tri+ Concentrations for 1993 - 2018
3-16 Exposure Parameters for Tree Swallow (Tachycineta bicolor)
3-17 Exposure Parameters for Mallard (Anas platyrhynchos)
3-18 Exposure Parameters for Belted Kingfisher (Ceryle alcyori)
3-19 Exposure Parameters for Great Blue Heron (Ardea herodias)
xii TAMS/MCA
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3-20 Exposure Parameters for Bald Eagle (Haliaeetus leucocephalus)
3-21 Exposure Parameters for Little Brown Bat (Myotis lucifugus)
3-22 Exposure Parameters for Raccoon (Proycon lotor)
3-23 Exposure Parameters for Mink (Mustela visori)
3-24 Exposure Parameters for River Otter (Lutra canadensis) ,
3-25 Summary of ADDEXPECTED and Egg Concentrations for Female Swallow Based on Tri+
Congeners for Period 1993-2018
3-26 Summary of ADD95%UCL and Egg Concentrations for Female Swallow Based on Tri+
Congeners for Period 1993-2018
3-27 Summary of ADDEXPECTED and Egg Concentrations for Female Mallard Based on Tri+
Congeners for Period 1993-2018
3-28 Summary of ADD95%UCL and Egg Concentrations for Female Mallard Based on Tri+
Congeners for Period 1993-2018
3-29 Summary of ADDEXPECTED and Egg Concentrations for Female Belted Kingfisher Based on
Tri+ Congeners for Period 1993-2018
3-30 Summary of ADD95%UCLand Egg Concentrations for Female Belted Kingfisher Based on Tri+
Congeners for Period 1993-2018
3-31 Summary of ADDEXPECTED and Egg Concentrations for Female Great Blue Heron Based on
Tri+ Congeners for Period 1993 - 2018
3-32 Summary of ADD95%UCLand Egg Concentrations for Female Great Blue Heron Based on
Tri+ Congeners for Period 1993-2018
3-33 Summary of ADDEXPECTED and Egg Concentrations for Female Bald Eagle Based on Tri+
Congeners for Period 1993-2018
3-34 Summary of ADD95%UCL and Egg Concentrations for Female Bald Eagle Based on Tri+
Congeners for Period 1993-2018
3-35 Summary of ADDEXPECTED and Egg Concentrations for Female Tree Swallow for the Period
1993-2018 on TEQ Basis
3-36 Summary of ADD95%UCL and Egg Concentrations for Female Tree Swallow for the Period
1993-2018 on TEQ Basis
3-37 Summary of ADDEXPECTED and Egg Concentrations for Female Mallard for the Period 1993
-2018 on TEQ Basis
xiii TAMS/MCA
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FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
3-38 Summary of ADDEXPECrED and Egg Concentrations for Female Mallard for the Period 1993
-2018onTEQBasis
3-39 Summary of ADDEXPECTED and Egg Concentrations for Female Belted Kingfisher for the
Period 1993 - 2018 on TEQ Basis
3-40 Summary of ADD95%UCLand Egg Concentrations for Female Belted Kingfisher for the Period
1993-2018 on TEQ Basis
3-41 Summary of ADDEXPECTED and Egg Concentrations for Female Great Blue Heron for the
Period 1993 - 2018 on TEQ Basis
3-42 Summary of ADD95%UCLand Egg Concentrations for Female Great Blue Heron for the Period
1993-2018 on TEQ Basis
3-43 Summary of ADDEXPECTED and Egg Concentrations for Female Eagle for the Period 1993 -
2018 on TEQ Basis
3-44 Summary of ADD95%UCLand Egg Concentrations for Female Eagle for the Period 1993 -
2018 on TEQ Basis
3-45 Summary of ADDEXPECTED for Female Bat Based on Tri+ Predictions for the Period 1993 -
2018
3-46 Summary of ADD95%UCLfor Female Bat Based on Tri+ Predictions for the Period 1993 -
2018
3-47 Summary of ADDEXPECTEDfor Female Raccoon Based on Tri+ Predictions for the Period 1993
-2018
3-48 Summary of ADD95%UCLfor Female Raccoon Based on Tri+ Predictions for the Period 1993
-2018
3-49 Summary of ADDEXPECTED for Female Mink Based on Tri+ Predictions for the Period 1993
-2018
3-50 Summary of ADD95%UCLfor Female Mink Based on Tri+ Predictions for the Period 1993 -
2018
3-51 Summary of ADDEXPECTED for Female Otter Based on Tri+ Predictions for the Period 1993
-2018
3-52 Summary of ADD95%UCL for Female Otter Based on Tri+ Predictions for the Period 1993 -
2018
3-53 Summary of ADDEXPECTED for Female Bat on a TEQ Basis for the Period 1993-2018
xiv TAMS/MCA
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FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
3-54 Summary of ADD95%UCL for Female Bat on a TEQ Basis for the Period 1993 - 2018
3-55 Summary of ADDEXPECTED for Female Raccoon on a TEQ Basis for the Period 1993 - 2018
3-56 Summary of ADD95%UCL for Female Raccoon on a TEQ Basis for the Period 1993 - 2018
3-57 Summary of ADDEXPECTED for Female Mink on a TEQ Basis for the Period 1993 - 2018
3-58 Summary of ADD95%UCLfor Female Mink on a TEQ Basis for the Period 1993-2018
3-59 Summary of ADDEXPECrED for Female Otter on a TEQ Basis for the Period 1993 - 2018
3-60 Summary of ADD95%UCL for Female Otter on a TEQ Basis for the Period 1993 - 2018
4-1 Toxicity Reference Values for Fish - Dietary Doses and Egg Concentrations of Total PCBs
and Dioxin Toxic Equivalents (TEQs)
4-2 Toxicity Reference Values for Birds - Dietary Doses and Egg Concentrations of Total PCBs
and Dioxin Toxic Equivalents (TEQs)
4-3 Toxicity Reference Values for Mammals - Dietary Doses of Total PCBs and Dioxin Toxic
Equivalents (TEQs)
4-4 World Health Organization - Toxic Equivalency Factors (TEFs) for Humans, Mammals,
Fish, and Birds
5-1 Ratio of Predicted Sediment Concentrations to Sediment Guidelines.
5-2 Ratio of Predicted Whole Water Concentrations to Criteria and Benchmarks
5-3 Ratio of Predicted Pumpkinseed Concentrations to Field-Based NOAEL for Tri+ PCBs
5-4 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived NOAEL for Tri+
PCBs
5-5 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived LOAEL for Tri+
PCBs
5-6 Ratio of Predicted Pumpkinseed Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis
5-7 Ratio of Predicted Pumpkinseed Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis
5-8 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis
5-9 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis
xv TAMS/MCA
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FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
5-10 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived NOAEL For Tri+
PCBs
5-11 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived LOAEL For Tri+
PCBs
5-12 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis
5-13 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis
5-14 Ratio of Predicted White Perch Concentrations to Field-Based NOAEL for Tri+ PCBs
5-15 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived NOAEL for Tri+
PCBs
5-16 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived LOAEL for Tri+
PCBs
5-17 Ratio of Predicted White Perch Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis
5-18 Ratio of Predicted White Perch Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis
5-19 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis
5-20 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis
5-21 Ratio of Predicted Largemouth Bass Concentrations to Field-Based NOAEL For Tri+ PCBs
5-22 Ratio of Predicted Largemouth Bass Concentrations to Laboratory-Derived NOAEL on a
TEQ Basis
5-23 Ratio of Predicted Largemouth Bass Concentrations to Laboratory-Derived LOAEL on a
TEQ Basis
5-24 Ratio of Predicted Striped Bass Concentrations to Tri+ and TEQ PCB-Based TRVs
5-25 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Tree Swallow Based on
the Sum of Tri+ Congeners for the Period 1993 -2018
5-26 Ratio of Modeled Egg Concentrations to Benchmarks for Female Tree Swallow Based on
the Sum of Tri+ Congeners for the Period 1993-2018
xvi TAMS/MCA
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FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
5-27 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Tree Swallow Using TEQ
for the Period 1993-2018
5-28 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Tree Swallow Using
TEQ for the Period 1993 - 2018
5-29 Ratio of Modeled Dietary Dose for Female Mallard Based on FISHRAND Results for the
Tri+ Congeners
5-30 Ratio of Egg Concentrations for Female Mallard Based on FISHRAND Results for the Tri+
Congeners
5-31 Ratio of Modeled Dietary Dose to Benchmarks for Female Mallard for Period 1993 - 2018
on a TEQ Basis
5-32 Ratio of Modeled Egg Concentrations to Benchmarks for Female Mallard for Period 1993
-2018 on a TEQ Basis
5-33 Ratio of Modeled Dietary Dose to Benchmarks Based on FISHRAND for Female Kingfisher
Based on the Sum of Tri+ Congeners for the Period 1993-2018
5-34 Ratio of Modeled Dietary Dose to Benchmarks Based on FISHRAND for Female Blue
Heron Based on the Sum of Tri+ Congeners for the Period 1993-2018
5-35 Ratio of Modeled Dietary Dose to Benchmarks Based on FISHRAND for Female Bald Eagle
Based on the Sum of Tri+ Congeners for the Period 1993 - 2018
5-36 Ratio of Modeled Egg Concentrations to Benchmarks for Female Belted Kingfisher Based
on the Sum of Tri+ Congeners for the Period 1993-2018
5-37 Ratio of Modeled Egg Concentrations to Benchmarks for Female Great Blue Heron Based
on the Sum of Tri+ Congeners for the Period 1993-2018
5-38 Ratio of Modeled Egg Concentrations to Benchmarks for Female Bald Eagle Based on the
Sum of Tri+ Congeners for the Period 1993-2018
5-39 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Belted Kingfisher Using
TEQ for the Period 1993-2018
5-40 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Great Blue Heron Using
TEQ for the Period 1993-2018
5-41 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Bald Eagle Using TEQ
for the Period 1993-2018
xvii TAMS/MCA
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PHASE 2 REPORT - REVIEW COPY
FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
5-42 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Belted Kingfisher
Using TEQ for the Period 1993 - 2018
5-43 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Great Blue Heron
Using TEQ for the Period 1993-2018
5-44 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Bald Eagle Using
TEQ for the Period 1993 - 2018
5-45 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Bat for Tri+ Congeners
for the Period 1993-2018
5-46 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Bat on a TEQ Basis for
the Period 1993-2018
5-47 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Raccoon for Tri+
Congeners for the Period 1993-2018
5-48 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Raccoon on a TEQ
Basis for the Period 1993-2018
5-49 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Mink for Tri+
Congeners for the Period 1993-2018
5-50 Ratio of Modeled Dietary Dose to Toxicity Benchmarks for Female Otter for Tri+ Congeners
for the Period 1993-2018
5-51 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Mink on a TEQ Basis
for the Period 1993-2018
5-52 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Otter on a TEQ Basis
for the Period 1993-2018
xviii TAMS/MCA
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PHASE 2 REPORT - REVIEW COPY
FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
LIST OF FIGURES
1-1 Hudson River Drainage Basin and Site Location Map
1-2 Eight-Step Ecological Risk Assessment Process for Superfund - Hudson River PCB
Reassessment Ecological Risk Assessment
2-1 Phase 2 Ecological Sampling Locations - Lower Hudson River Stations
2-2 Hudson River PCB Reassessment Conceptual Model Diagram Including Floodplain Soils
3-1 Revised Segments and Regions of the Farley Model for PCBs in Hudson River Estuary and
Surround Area
3-2 Comparison of Cumulative PCB Loads at Waterford from Farley et al, 1999 and USEPA,
2000
3-3 Comparison Between the White Perch Body Burdens Using the March, 1999 Model and the
Farley Model Run with HUDTOX Upper River Loads (1987-1997)
3-4 Comparison Between the Striped Bass Body Burdens Using the March., 1999 Model and the
Farley Model Run with HUDTOX Upper River Loads
3-5 Comparison Between Field Data and Model Estimates for 1993 Dissolved PCB
Concentrations (Farley Model with HUDTOX Upper River Loads)
3-6 Comparison of Model and Measured Homologue Pattern for 1993 Dissolved Phase PCB
Concentrations
3-7 Comparison of Model and Measured PCB Surface Sediment Concentration for 1993
3-8 Comparison Between Model and Measured White Perch Body Burdens NYSDEC Fish
Samples vs. Farley Model with HUDTOX Upper River Loads
3-9 Comparison Between Model and Measured Striped Bass Body Burdens NYSDEC Fish
Samples vs. Farley Model with HUDTOX Upper River Loads
3-10 Comparison of Model Estimates for White Perch Body Burdens Farley Model with
HUDTOX Upper River Loads vs. FISHRAND in Food Web Regions 1 and 2
3-11 Comparison of White Perch Body Burdens (Farley Model vs. FISHRAND)
3-12a Comparison Between FISHRAND Results and Measurements at RM 152
3-12b Comparison Between FISHRAND Results and Measurements at RM 113
xix TAMS/MCA
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FURTHER SITE CHARACTERIZATION AND ANALYSIS
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
HUDSON RIVER PCBs REASSESSMENT RI/FS
TABLE OF CONTENTS
BOOK 1 of 1
3-12c Comparison Between FISHRAND Results and Measurements of Pumpkinseed
3-12d Comparison Between FISHRAND Results and Measurements of Spottail Shiner
3-13 Comparison Among the HUDTOX Upper River Load and Farley Model Estimates of
Dissolved Water Column Concentrations in Food Web Regions 1 and 2 (1987-2067)
3-14 Comparison Among the HUDTOX Upper River Load and Farley Model Estimates of
Particulate and Whole Water Column Concentrations in Food Web Region 1 (1987-2067)
3-15 Comparison Among the HUDTOX Upper River Load and Farley Model Estimates of Surface
Soil (0-2.5 cm) in Food Web Regions 1 and 2 (1987-2067)
3-16 Comparison Among the HUDTOX Upper River Load and Farley Model Estimates of White
Perch Body Burdens in Food Web Regions 1 and 2 (1987-2067)
3-17 Comparison Among the HUDTOX Upper River Load and Farley Model Estimates Striped
Bass Body Burdens in Food Web Regions 1 and 2 (1987-2067)
3-18 Forecasts of Large Mouth Bass Body Burdens from FISHRAND
3-19 Forecasts of White Perch Body Burdens from FISHRAND
3-20 Forecasts of Yellow Perch Body Burdens from FISHRAND
3-21 Forecasts of Brown Bullhead Body Burdens from FISHRAND
3-22 Forecasts of Pumpkinseed Body Burdens from FISHRAND
3-23 Forecasts of Spottail Shiner Body Burdens from FISHRAND
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ACRONYMS
ATSDR
GDI
CERCLA
CSF
EPC
GE
HI
HHRA
HHRASOW
HQ
NCP
NPL
NYSDEC
NYSDOH
PCB
RfD
RI
RI/FS
ROD
RM
RI/FS
SARA
TCDD
TEF
TSCA
UCL
USEPA
Agency for Toxic Substances and Desease Registry
Chronic Daily Intake
Comprehensive Environmental Response, Compensation, and Liability Act
Carcinogenic Slope Factor
Exposure Point Concentration
General Electric
Hazard Index
Human Health Risk Assessment
Human Helath Risk Assessment Scope of Work
Hazard Quotient
National Oil and Hazardous Substances Pollution Contingency Plan
National Priorities List
New York State Department of Environmental Conservation
New York State Department of Health
Polychlorinated Biphenyl
References Dose
Remedial Investigation
Remedial Investigation/Feasibility Study
Record of Decision
River Mile
Remedial Investigation/Feasibility Study
Superfund Amendments and Reauthorization Act of 1986
2,3,7,8-Tetrachlorodibenzo-p-dioxin
Toxicity Equivalency Factor
Toxic Substances Control Act
Upper Confidence Limit
United States Environmental Protection Agency
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Executive Summary
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Ecological Risk Assessment Addendum:
Future Risks in the Lower Hudson River
Executive Summary
December 1999
This document presents the baseline Ecological Risk Assessment for Future Risks in the
Lower Hudson River (ERA Addendum), which is a companion volume to the baseline Ecological
Risk Assessment (ERA) that was released by the U.S. Environmental Protection Agency (USEPA)
in August 1999. Together, the two risk assessments comprise the ecological risk assessment for
Phase 2 of the Reassessment Remedial Investigation/Feasibility Study (Reassessment RI/FS) for the
Hudson River PCBs site in New York.
The ERA Addendum quantitatively evaluates the future risks to the environment in the
Lower Hudson River (Federal Dam at Troy, New York to the Battery in New York City) posed by
polychlorinated biphenyls (PCBs) from the Upper Hudson River (Hudson Falls, New York to the
Federal Dam at Troy, New York), in the absence of remediation. This report uses current USEPA
policy and guidance as well as additional site data and analyses to update USEPA's 1991 risk
assessment.
USEPA uses ecological risk assessments to evaluate the likelihood that adverse ecological
effects are occurring or may occur as a result of exposure to one or more chemical or physical
stressors. The Superfund ecological risk assessment process includes the following: 1) identification
of contaminants of concern; 2) development of a conceptual model, which identifies complete
exposure pathways for the ecosystem; 3) identification of assessment endpoints, which are ecological
values to be protected; 4) development of measurement endpoints, which are the actual
measurements used to assess risk to the assessment endpoints; 5) selection of receptors of concern;
6) the exposure assessment, which describes concentrations or dietary doses of contaminants of
concern to which the selected receptors are or may be exposed; 7) the effects assessment, which
describes toxicological effects due to chemical exposure and the methods used to characterize those
effects to the receptors of concern; and 8) risk characterization, which compares the results of the
exposure assessment with the effects assessment to evaluate the likelihood of adverse ecological
effects associated with exposure to chemicals at a site.
The ERA Addendum indicates that, for some species, future concentrations of PCBs in the
Lower Hudson River generally exceed levels that have been shown to cause adverse ecological
effects through 2018 (the entire forecast period). The results of the ERA Addendum will help
establish acceptable exposure levels for use in developing remedial alternatives for PCB-
contaminated sediments in the Upper Hudson River, which is Phase 3 (Feasibility Study) of the
Reassessment RI/FS.
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Contaminants of Concern
The contaminants of concern identified for the site are PCBs. PCBs are a group of synthetic
organic compounds consisting of 209 individual chlorinated biphenyls called congeners. Some PCB
congeners are considered to be structurally similar to dioxin and are called dioxin-like PCBs. Toxic
equivalency (TEQ) factors, based on the toxicity of dioxin, have been developed for the dioxin-like
PCB congeners. PCBs have been shown to cause adverse reproductive and developmental effects
in animals. Ecological exposure to PCBs is primarily an issue of bioaccumulation rather than direct
toxicity. PCBs bioaccumulate in the environment by both bioconcentrating (being absorbed from
water and accumulated in tissue to levels greater than those found in surrounding water) and
biomagnifying (increasing in tissue concentrations as they go up the food chain through two or more
trophic levels).
Site Conceptual Model
The Hudson River PCBs site is the 200 miles (322 km) of river from Hudson Falls, New
York to the Battery in New York City. As defined in the ERA and ERA Addendum, the Lower
Hudson River extends approximately 160 miles (258 km) from the Federal Dam at Troy (River Mile
153) to the Battery.
The Hudson River is home to a wide variety of ecosystems. The Lower Hudson River is
tidal, does not have dams, and is freshwater in the vicinity of the Federal Dam, becoming brackish
and increasingly more saline towards the Battery. Spring runoffs and major storms can push the salt
front well below the Tappan Zee Bridge, and sometimes south to New York City. The Lower
Hudson has deep water environments, shallow nearshore areas (shallows, mudflats, and shore
communities), tidal marshes, and tidal swamps.
PCBs were released from two General Electric Company capacitor manufacturing facilities
located in the Upper Hudson River at Hudson Falls and Fort Edward, New York. Many of these
PCBs adhered to river sediments. As PCBs in the river sediments are released slowly into the river
water, these contaminated sediments serve as a continuing source of PCBs. During high flow events,
the sediments may be deposited on the floodplain and PCBs may thereby enter the terrestrial food
chain. High flow events may also increase the bioavailability of PCBs to organisms in the river
water.
Animals and plants living in or near the river, such as invertebrates, fish, amphibians, and
water-dependent reptiles, birds, and mammals, may be directly exposed to the PCBs from
contaminated sediments, river water, and air, and/or indirectly exposed through ingestion of food
(e.g., prey) containing PCBs.
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Assessment Endpoints
Assessment endpoints are explicit expressions of actual environmental values (i.e., ecological
resources) that are to be protected. They focus a risk assessment on particular' components of the
ecosystem that could be adversely affected due to contaminants at the site. These endpoints are
expressed in terms of individual organisms, populations, communities, ecosystems, or habitats with
some common characteristics (e.g., feeding preferences, reproductive requirements). The assessment
endpoints for the ERA Addendum were selected to include direct exposure to PCBs in Lower
Hudson River sediments and river water through ingestion and indirect exposure to PCBs via the
food chain. Because PCBs are known to bioaccumulate, an emphasis was placed on indirect
exposure at various levels of the food chain to address PCB-related risks at higher trophic levels.
The assessment endpoints that were selected for the Lower Hudson River are:
• Benthic community structure as a food source for local fish and wildlife
• Protection and maintenance (survival, growth, and reproduction) of local fish populations
(forage, omnivorous, and piscivorous)
• Protection and maintenance (survival, growth, and reproduction) of local insectivorous bird
populations
• Protection and maintenance (survival, growth, and reproduction) of local waterfowl
populations
• Protection and maintenance (survival, growth, and reproduction) of local piscivorous birds
populations
• Protection and maintenance (survival, growth, and reproduction) of local insectivorous
wildlife populations
• Protection and maintenance (survival, growth, and reproduction) of local omnivorous
wildlife populations
• Protection and maintenance (survival, growth, and reproduction) of local piscivorous wildlife
populations
• Protection of threatened and endangered species
• Protection of significant habitats
ES-3 TAMS/MCA
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Measurement Endpoints
Measurement endpoints provide the actual measurements used to evaluate ecological risk and
are selected to represent mechanisms of toxicity and exposure pathways. Measurement endpoints
for future risk generally include modeled concentrations of chemicals in water, sediment, fish, birds,
and/or mammals, laboratory toxicity studies, and field observations. The measurement endpoints
identified for the ERA Addendum are:
1) Modeled concentrations of PCBs in fish and invertebrates to evaluate food-chain exposure;
2) Modeled total PCB body burdens in receptors (including avian receptor eggs) to determine
exceedance of effect-level thresholds based on toxicity reference values (TRVs);
3) Modeled TEQ-based PCB body burdens in receptors (including avian receptor eggs) to
determine exceedance of effect-level thresholds based on TRVs;
4) Modeled concentration of PCBs in river water to determine exceedence of criteria for
concentrations of PCBs in river water that are protective of benthic invertebrates, fish and
wildlife;
5) Modeled concentrations of PCBs in sediment to determine exceedence of guidelines for
concentrations of PCBs in sediments that are protective of aquatic health; and
6) Field observations.
Receptors of Concern
Risks to the environment were evaluated for individual receptors of concern that were
selected to be representative of various feeding preferences, predatory levels, and habitats (aquatic,
wetland, shoreline). The ERA Addendum does not characterize injury to, impact on, or threat to
every species of plant or animal that lives in or adjacent to the Hudson River; such a characterization
is beyond the scope of the Superfund ecological risk assessment. The following receptors of concern
were selected for the ERA Addendum:
Aquatic Invertebrates /
V.
• Benthic macroinvertebrate community (e.g., aqu'atic worms, insect larvae, and isopods)
Fish Species
• Pumpkinseed (Lepomis gibbosus)
• Spottail shiner (Notropis hudsonius)
ES-4 TAMS/MCA
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• Brown bullhead (Ictalurus nebulosus)
• White perch (Morone americana)
• Yellow perch (Percaflavescens)
• Largemouth bass (Micropterus salmoides)
• Striped bass (Morone saxatilis)
• Shortnose sturgeon (Acipenser brevirostrum)
Birds
• Tree swallow (Tachycineta bicolor)
• Mallard (Anas platyrhychos)
• Belted kingfisher (Ceryle alcyon)
• Great blue heron (Ardea herodias)
• Bald eagle (Haliaeetus leucocephalus)
Mammals
• Little brown bat (Myotis lucifugus)
• Raccoon (Procyon lotof)
• Mink (Mustela vison)
• River otter (Lutra canadensis)
Exposure Assessment
The Exposure Assessment describes complete exposure pathways and exposure parameters
(e.g., body weight, prey ingestion rate, home range) used to calculate the concentrations or dietary
doses to which the receptors of concern may be exposed due to chemical exposure. USEPA
previously released reports on the nature and extent of contamination in the Hudson River as part
of the Reassessment RI/FS (e.g., February 1997 Data Evaluation and Interpretation Report, July 1998
Low Resolution Sediment Coring Report, August 1998 Database for the Hudson River PCBs
Reassessment RI/FS [Release 4.1], and May 1999 Baseline Modeling Report). The Reassessment
RI/FS documents form the basis of the site data collection and analyses that were used in conducting
the ERA Addendum. Future (i.e., modeled) concentrations of PCBs in fish, sediments and river
water are provided in the ERA Addendum, based on fate and bioaccumulation models by Farley et
al. (1999) and USEPA's Revised Baseline Modeling Report (USEPA, 2000). Exposure parameters
were obtained from USEPA references, the scientific literature, and directly from researchers as
reported in the ERA.
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Effects Assessment
The Effects Assessment describes the methods used to characterize particular toxicological
effects of PCBs on aquatic and terrestrial organisms due to chemical exposure. These measures of
toxicological effects, called TRVs, provide a basis for estimating whether the chemical exposure at
a site is likely to result in adverse ecological effects.
In conducting the ERA Addendum, USEPA used the TRVs selected in the ERA based on
Lowest Observed Adverse Effects Levels (LOAELs) and/or No Observed Adverse Effects Levels
(NOAELs) from laboratory and/or field-based studies reported in the scientific literature. These
TRVs examine the effects of PCBs and dioxin-like PCB congeners on the survival, growth, and
reproduction offish and wildlife species in the Lower Hudson River. Reproductive effects (e.g., egg
maturation, egg hatchability, and survival of juveniles) were generally the most sensitive endpoints
for animals exposed to PCBs.
Risk Characterization
Risk Characterization examines the likelihood of adverse ecological effects occurring as a
result of exposure to chemicals and discusses the qualitative and quantitative assessment of risks to
ecological receptors with regard to toxic effects. Risks are estimated by comparing the results of the
Exposure Assessment (e.g., modeled concentrations of chemicals in receptors of concern) to the
TRVs developed in the Effects Assessment. The ratio of these two numbers is called a Toxicity
Quotient, or TQ.
TQs equal to or greater than one (TQ > 1) are typically considered to indicate potential risk
to ecological receptors, for example reduced or impaired reproduction or recruitment of new
individuals. The TQs provide insight into the potential for adverse effects upon individual animals
in the local population resulting from chemical exposure. If a TQ suggests that effects are not
expected to occur for the average individual, then they are probably insignificant at the population
level. However, if a TQ indicates risks are present for the average individual, then risks may be
present for the local population.
At each step of the risk assessment process there are sources of uncertainty. Measures were
taken in the ERA to address and characterize the uncertainty. For example, in some cases
uncertainty factors were applied in developing TRVs. The purpose of these uncertainty factors is
to ensure that the calculated TRVs are protective of the receptor species of concern. Another source
of uncertainty is associated with the future PCB concentrations in fish. The PCB concentrations in
fish presented in the ERA Addendum (forecast from models in Farley et al. (1999) and the Revised
Baseline Modeling Report (USEPA, 2000) may be significantly underestimated, which may
underestimate risks to fish species. However, based on a comparison of measured concentrations
of PCBs in fish to modeled concentrations, the forecasts presented in the ERA Addendum are not
expected to overestimate future PCB concentration in fish, so that the risks to fish are not expected
to be overestimated.
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To integrate the various components of the ERA Addendum, the results of the risk
characterization and associated uncertainties were evaluated using a weight-of-evidence approach
to assess the risk of adverse effects in the receptors of concern as a result of exposure to PCBs in the
Lower Hudson River. The weight-of-evidence approach considers both the results of the TQ
analysis and field observations for each assessment endpoint. For the mammals and most birds, TQs
for the dioxin-like PCBs were greater than the TQs for total PCBs.
Benthic Community Structure
Risks to local benthic invertebrate communities were examined using two lines of evidence.
These lines of evidence are: 1) comparison of modeled water column concentrations of PCBs to
criteria and 2) comparisons of modeled sediment concentrations to guidelines. Both suggest an
adverse effect of PCBs on benthic invertebrate populations serving as a food source to local fish in
the Lower Hudson River. Uncertainty in this analysis is considered low.
Local Fish (Forage. Omnivorous. Piscivorous and Semi-piscivorous')
Risks to local fish populations were examined using five lines of evidence. These lines of
evidence are: 1) comparison of modeled total PCB fish body burdens to TRVs; 2) comparison of
modeled TEQ fish body burdens to TRVs; 3) comparison of modeled water column concentrations
of PCBs to criteria; 4) comparison of modeled sediment concentrations to guidelines; and 5) field-
based observations. Multiple receptors were evaluated for forage and semi-piscivorous/piscivorous
fish.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
fish species in the Lower Hudson River. However, based upon toxicity quotients, future exposure
to PCBs may reduce or impair the survival, growth, and reproductive capability of some forage
species (e.g., pumpkinseed) and semi-piscivorous/piscivorus fish (e.g., white perch, yellow perch,
largemouth bass, and striped bass), particularly in the upper reaches of the Lower Hudson River.
There is a moderate degree of uncertainty in the modeled body burdens used to evaluate
exposure, and at most an order of magnitude uncertainty in the TRVs (for the TEQ-based TRVs, no
uncertainty factors were needed).
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for protection of fish and wildlife through the duration of the
forecast period (1993 - 2018).
Insectivorous Birds
Risks to local insectivorous bird populations were examined using six lines of evidence.
These lines of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs; 2)
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comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled total PCB egg
concentrations to TRVs; 4) comparison of modeled TEQ egg concentrations to TRVs; 5) comparison
of modeled water column concentrations of PCBs to criteria; and 6) field-based observations. The
tree swallow was selected to represent insectivorous bird species.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
insectivorous bird species in the Lower Hudson River Valley. TQs are all below one for all locations
for the entire forecast period (1993 to 2018). However, given that U.S. Fish and Wildlife Service
field studies suggest PCBs may cause abnormal nest construction of Upper Hudson River tree
swallows, it is possible that future exposure to PCBs in the Lower Hudson River may reduce or
impair the reproductive capability of tree swallows, particularly in the upper reaches of the Lower
Hudson River.
There is a moderate degree of uncertainty in the calculated modeled concentrations of PCBs
in tree swallow diets and the concentrations of PCBs in eggs. There is a low degree of uncertainty
associated with tree swallow TRVs, which were derived from field studies of Hudson River tree
swallows.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993-2018).
Waterfowl
Risks to local waterfowl populations were examined using six lines of evidence. These lines
of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs; 2) comparison of
modeled TEQ dietary doses to TRVs; 3) comparison of modeled total PCB egg concentrations to
TRVs; 4) comparison of modeled TEQ egg concentrations to TRVs; 5) comparison of modeled
water column concentrations of PCBs to criteria; and 6) field-based observations. The mallard was
selected to represent waterfowl.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
waterfowl in the Lower Hudson River Valley. However, based upon toxicity quotients, future
exposure to PCBs may reduce or impair the survival, growth, and reproductive capability of some
waterfowl, particularly in the upper reaches of the lower river.
Calculated dietary doses of PCBs and concentrations of PCBs in eggs typically exceed their
respective TRVs throughout the modeling period. Toxicity quotients for the TEQ-based (i.e., dioxin-
like) PCBs consistently show greater exceedances than for total (Tri+) PCBs. There is a moderate
degree of uncertainty in the dietary dose and egg concentration estimates. Given the magnitude of
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the TEQ-based TQs, they would have to decrease by an order of magnitude or more to fall below one
for waterfowl in the Lower Hudson River.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993-2018).
Piscivorous Birds
Risks to local semi-piscivorous/piscivorous bird populations were examined using six lines
of evidence. These lines of evidence are: 1) comparison of modeled total PCB dietary doses to
TRVs; 2) comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled total PCB
egg concentrations to TRVs; 4) comparison of modeled TEQ egg concentrations to TRVs; 5)
comparison of modeled water column concentrations of PCBs to criteria; and 6) field-based
observations. The belted kingfisher, great blue heron, and bald eagle were selected to represent
piscivorous birds.
Collectively, the evidence indicates that future PCB exposures (predicted i'rom 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of these
piscivorous species. However, based upon toxicity quotients, future exposure to PCBs may reduce
or impair the survival, growth, and reproductive capability of some piscivorous birds, particularly
in the upper reaches of the Lower Hudson Rver. Calculated dietary doses of PCBs and
concentrations of PCBs in eggs exceed all TRVs (i.e., NOAELs and LOAELs) for the belted
kingfisher and bald eagle throughout the modeling period, and exceed NOAELs for the great blue
heron. Toxicity quotients for egg concentrations are generally higher than body burden TQs.
There is a moderate degree of uncertainty in the dietary dose and egg concentration estimates.
Given the magnitude of the TQs, they would have to decrease by an order of magnitude or more to
fall below one for piscivorous birds in the Lower Hudson River. In particular, the bald eagle TQs
exceeded one by up to three orders of magnitude. Therefore, even if the factor of 2.5 to adjust from
largemouth bass fillets to whole body burden and the subchronic-to-chronic uncertainty factor of 10
used for the body burden TRV are removed, the TQs would remain well over one. These results
coupled with the lack of breeding success in Lower Hudson River bald eagles (USGS, 1999) indicate
that reproductive effects may be present.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993-2018).
Insectivorous Mammals
Risks to local insectivorous mammal populations were examined using four lines of
evidence. These lines of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs;
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2) comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled water column
concentrations of PCBs to criteria; and 4) field-based observations. The little brown bat was selected
to represent insectivorous mammals.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
insectivorous mammals in the Lower Hudson River Valley. However, exposure to PCBs may reduce
or impair the survival, growth, or reproductive capability of insectivorous mammals in the Lower
Hudson River. Modeled dietary doses for the little brown bat exceed TRVs by up to two orders of
magnitude at all locations modeled. There is a moderate degree of uncertainty in the calculated
dietary doses.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993-2018).
Omnivorous Mammals
Risks to local omnivorous mammal populations were examined using four lines of evidence.
These lines of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs; 2)
comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled water column
concentrations of PCBs to criteria; and 4) field-based observations. The raccoon was selected to
represent omnivorous mammals.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
omnivorous mammals in the Lower Hudson River Valley. However, exposure to PCBs may reduce
or impair the survival, growth, or reproductive capability of omnivorous mammals in the Lower
Hudson River. Modeled dietary doses for the raccoon exceed dietary dose NOAELs on a total PCB
(Tri+) basis and all TRVs on a TEQ-basis. There is a moderate degree of uncertainty in the
calculated dietary doses.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993-2018).
Piscivorous Mammals
Risks to local semi-piscivorous/piscivorous mammal populations were examined using four
lines of evidence. These lines of evidence are: 1) comparison of modeled total PCB dietary doses
to TRVs; 2) comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled water
column concentrations of PCBs to criteria; and 4) field-based observations. The mink and river otter
were selected to represent piscivorous mammals.
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Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of these
piscivorous species. However, based upon toxicity quotients, future exposure to PCBs may reduce
or impair the survival, growth, and reproductive capability of piscivorous mammals, particularly in
the upper reaches of the Lower Hudson River. Calculated dietary doses of PCBs exceed the NOAEL
on a total PCB basis for both the mink and river otter and exceed all TEQ-based TRVs by up to three
orders of magnitude.
There is a moderate degree of uncertainty in the dietary dose estimates. However, given the
magnitude of the TQs, they would have to decrease at least an order of magnitude to fall below one.
In particular, the river otter TQs exceeded one by up to three orders of magnitude. Therefore, even
if the factor of 2.5 to adjust from largemouth bass fillets to whole body burden is removed, the TQs
would remain well over one.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993-2018). In addition, preliminary results from a NYSDEC study indicate that PCBs may
have an adverse effect on the litter size and possibly kit survival of river otter in the Hudson River
(Mayack, 1999b).
Threatened and Endangered Species
Risks to threatened and endangered species were examined using five lines of evidence.
These lines of evidence are: 1) comparison of modeled total PCB dietary doses/egg concentrations
to TRVs; 2) comparison of modeled TEQ dietary doses/egg concentrations to TRVs; 3) comparison
of predicted modeled water column concentrations of PCBs to criteria; 4) comparison of modeled
sediment concentrations of PCBs to guidelines; and 5) field-based observations. The shortnose
sturgeon and bald eagle were selected to represent threatened and endangered species.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of threatened
or endangered species. However, using the TEQ-based toxicity quotients, potential for adverse
reproductive effects in shortnose sturgeon exists, particularly when considering the long life
expectancy of the sturgeon. Almost all TQs calculated for the bald eagle (across all locations)
exceeded one, in some instances by more than three orders of magnitude. Both the dietary dose and
egg-based results were consistent in this regard. Other threatened or endangered raptors, such as the
peregrine falcon, osprey, northern harrier, and red-shouldered hawk may experience similar
exposures.
There is a moderate degree of uncertainty in the dietary dose estimates. However, the bald
eagle TQs exceeded one by up to three orders of magnitude. Therefore, even if the factor of 2.5 to
adjust from largemouth bass fillets to whole body burden and the subchronic-to-chronic uncertainty
factor of 10 used for the body burden TRV are removed, the TQs would remain well over one.
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These results coupled with the lack of breeding success in Lower Hudson River bald eagles (USGS,
1999) indicate that reproductive effects may be present.
Modeled concentrations of PCBs in river water and sediment in the Lower Hudson River
show exceedances of the majority of their respective criteria and guidelines through the duration of
the forecast period (1993-2018).
Significant Habitats
Risks to significant habitats were examined using four lines of evidence. These lines of
evidence are: 1) toxicity quotients calculated for receptors in this assessment; 2) comparison of
modeled water column concentrations of PCBs to criteria; 3) comparison of modeled sediment
concentrations of PCBs to guidelines; and 4) field-based observations.
Based on the toxicity quotients for receptors of concern, future PCB concentrations modeled
for the Lower Hudson River exceed toxicity reference values for some fish, avian, and mammalian
receptors. These comparisons indicate that animals feeding on Hudson River-based prey may be
affected by the concentrations of PCBs found in the river on both a total PCB and TEQ basis. In
addition, based on the ratios obtained in this evaluation, other taxononic groups not directly
addressed in this evaluation (e.g., amphibians and reptiles) may also be affected by PCBs in the
Lower Hudson River. Many year-round and migrant species use the significant habitats along the
Lower Hudson River for breeding or rearing their young. Therefore, exposure to PCBs may occur
at a sensitive time in the life cycle (i.e., reproductive and development) and have a greater effect on
populations than at other times of the year.
Modeled concentrations of PCBs in river water and sediment in the Lower Hudson River
show exceedances of the majority of their respective criteria and guidelines through the duration of
the forecast period (1993-2018).
Major Findings of the ERA Addendum
The results of the risk assessment indicate that receptors in close contact with the Lower
Hudson River are at an increased ecological risk as a result of future exposure to PCBs in sediments,
water, and/or prey. This conclusion is based on a TQ approach, in which modeled body burdens,
dietary doses, and egg concentrations of PCBs were compared to TRVs, and on field observations.
On the basis of these comparisons, all receptors of concern except the tree swallow are at risk. In
summary, the major findings of the report are:
• Fish in the Lower Hudson River are at risk from future exposure to PCBs. Fish that eat other
fish (i.e., which are higher on the food chain), such as the largemouth bass and striped bass,
are especially at risk. PCBs may adversely affect fish survival, growth, and reproduction.
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Mammals that feed on insects with an aquatic stage spent in the Lower Hudson River, such
as the little brown bat, are at risk from future PCB exposure. PCBs may adversely affect the
survival, growth, and reproduction of these species.
Birds that feed on insects with an aquatic stage spent in the Lower Hudson, such as the tree
swallow, are not expected to be at risk from future exposure to PCBs.
Waterfowl feeding on animals and plants in the Lower Hudson River are at risk from PCB
exposure. Future concentrations of PCBs may adversely affect avian survival, growth, and
reproduction.
Birds and mammals that eat PCB-contaminated fish from the Lower Hudson River, such as
the bald eagle, belted kingfisher, great blue heron, mink, and river otter, are at risk. Future
concentrations of PCBs may adversely affect the survival, growth, and reproduction of these
species.
Omnivorous animals, such as the raccoon, that derive some of their food from the Lower
Hudson River are at risk from PCB exposure. Future concentrations of PCBs may adversely
affect the survival, growth, and reproduction of these species.
Fragile populations of threatened and endangered species in the Lower Hudson River,
represented by the bald eagle and shortnose sturgeon, are particularly susceptible to adverse
effects from future PCB exposure.
Modeled PCB concentrations in water and sediments in the Lower Hudson River generally
exceed standards, criteria and guidelines established to be protective of the environment.
Animals that use areas along the Lower Hudson designated as significant habitats may be
adversely affected by the PCBs.
The future risks to fish and wildlife are greatest in the upper reaches of the Lower Hudson
River and decrease in relation to decreasing PCB concentrations down river. Based on
modeled PCB concentrations, many species are expected to be at risk through 2018 (the
entire forecast period).
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Chapter 1
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1.0 INTRODUCTION
1.1 Purpose of Report
This document presents the baseline Ecological Risk Assessment for Future Risks in the
Lower Hudson River (ERA Addendum), which is a companion volume to the baseline Ecological
Risk Assessment (ERA) that was released by the U.S. Environmental Protection Agency (USEPA)
in August 1999. Together, the two risk assessments comprise the ecological risk assessment for
Phase 2 of the Reassessment Remedial Investigation/Feasibility Study (Reassessment RI/FS) for the
Hudson River PCBs site in New York.
The ERA Addendum quantitatively evaluates the future risks to the environment in the
Lower Hudson River (Federal Dam at Troy, New York to the Battery in New York City) posed by
polychlorinated biphenyls (PCBs) from the Upper Hudson River (Hudson Falls, New York to the
Federal Dam at Troy, New York), in the absence of remediation. This report uses current USEPA
policy and guidance as well as additional site data and analyses to update USEPA's 1991 risk
assessment.
Consistent with USEPA guidance (USEPA, 1997b), the ERA addendum calculates the risk
to individual receptor species of concern. The ERA addendum uses the same receptor species as the
baseline ERA (USEPA, 1999c). The species were selected to represent various trophic levels, a
variety of feeding types, and a diversity of habitats associated with the Hudson River. Receptor
species were selected as surrogates for the range of species potentially exposed to PCBs in the
Hudson River.
Because of the focused nature of the Reassessment RI/FS, a number of technical decisions
were made to structure and focus the ERA, as described in the baseline ERA (USEPA, 1999c). The
ERA and ERA Addendum focus on particular categories of PCBs that can be supported by the
available data and are amenable to modeling. Selection of PCBs categories to measure, model, and
assess was based on risk assessment considerations as well as on practical-considerations related to
modeling requirements. For the ecological risk assessment this led to a decision to evaluate total
PCBs as represented by "tri and higher" chlorinated compounds, as well as select congeners. The
"tri and higher" group includes the PCB compounds that are most toxic to fish and wildlife and
therefore captures most of the toxicity associated with these compounds. Tri and higher totals for
the Lower Hudson River that are compared to total PCBs (which include mono and dichlorinated
PCBs) may underestimate risks in some instances.
1.2 Report Organization
This ERA follows Ecological Risk Assessment Guidance for Superfund, Process for
Designing and Conducting Ecological Risk Assessments (ERAGS) (USEPA, 1997bj, as detailed in
the baseline ERA (USEPA, 1999c). The ERAGS guidance has of eight steps, as shown in Figure
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1-2. This ERA Addendum covers Steps 6 and 7 of the ERAGS process (analysis of ecological
exposures and effects and risk characterization) for the future risks in the Lower Hudson River.
Steps 1-5 were completed in previous reports (e.g., USEPA, 1999c). Step 8, Risk Management,
occurs after the completion of the ERA and is the responsibility of the USEPA site risk manager,
who balances risk reductions associated with cleanup of contaminants with potential impacts of the
remedial actions themselves.
Much of the information used in this addendum was originally presented in the baseline ERA
(USEPA, 1999c), where a detailed description of the assumptions and methodology that were used
can be found. In keeping with ERAGS, the format of this ERA Addendum is as follows:
• Chapter 1, the introduction, provides an overview of purpose of the report.
• Chapter 2, problem formulation, summarizes the conceptual model, assessment and
measurement endpoints, and the receptors of concern from the baseline ERA (USEPA,
1999c).
• Chapter 3, the exposure assessment, discusses modeled PCB concentrations forecast using
the Farley et al. (1999) and FISHRAND models, identifies exposure pathways for receptors,
and summarizes exposure parameters selected for avian and mammalian receptors in the
baseline ERA (USEPA, 1999c).
• Chapter 4, the effects assessment, summarizes toxicity reference values (TRVs) selected
for each receptor in the baseline ERA (USEPA, 1999c).
• Chapter 5, the risk characterization, uses the exposure and effects assessments to provide
a quantitative estimate of risk to receptors. The results of the measurement endpoints are
used to evaluate the assessment endpoints selected in the problem formulation phase of the
assessment.
• Chapter 6, the uncertainty analysis, summarizes uncertainties associated with the
assessment based on the baseline ERA (USEPA, 1999c).
• Chapter 7, conclusions, presents the conclusions of the risk assessment. This section
integrates the results of the risk characterization with the uncertainty analysis to provide
perspective on the overall confidence in the assessment.
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Chapter 2
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2.0 PROBLEM FORMULATION
Problem formulation establishes the goals, breadth, and focus of the assessment. It defines
the questions and issues based on identifiable complete exposure pathways and ecological effects.
A key aspect of problem formulation is the development of a conceptual model that illustrates the
relationships among sources, pathways, and receptors.
2.1 Site Characterization
The Hudson River PCBs Site includes the 200 miles (322 km) of river from Hudson Falls,
NY to the Battery in New York City, as described in the baseline ERA (USEPA, 1999c). The ERA
Addendum covers future risks to the Lower Hudson River, which stretches from the Federal Dam
to the Battery. Phase 2 ecological sampling locations are shown in Figure 2-1. The Lower Hudson
River is tidal and includes freshwater, brackish, and estuarine habitats, as described below.
2.2 Contaminants of Concern
Consistant with the scopr of the Reassessment RI/FS, the contaminants of concern (COCs)
are limited to PCBs. While there are other contaminants at various locations in the Hudson (e.g.,
metals, polycyclic aromatic hydrocarbons), PCBs are the chemicals that are the basis for the 1984
ROD and the Reassessment RI/FS. Consistent with that focus, the evaluation examines risks posed
by the presence of in-place PCBs in river sediments. PCBs can be described as individual congeners,
Aroclors, and total PCBs. Total PCBs in this assessment are represented by the trichlorinated and
higher congeners (designated Tri+) for the purposes of modeling (USEPA, 1999b), which
approximate total PCBs in biota.
2.3 Conceptual Model
A site conceptual model identifies the source, media, pathway, and route of exposure
evaluated in the ecological risk assessment, and the relationship of the measurement endpoints to
the assessment endpoints (USEPA, 1997b). An integrated site conceptual model was developed for
the Hudson River baseline ERA (Figure 2-2). In this model, the initial sources of PCBs are releases
from the two GE capacitor manufacturing facilities located in Hudson Falls and Fort Edward, NY.
PCBs enter the Hudson River and adhere to sediments or are redistributed into the water
column. Sediments may be deposited on the floodplain during high flow events and provide a
pathway for PCBs to enter the terrestrial food chain.
Animals and plants living in or near the Hudson River, such as invertebrates, fish,
amphibians, and water-dependent reptiles, birds, and mammals, are potentially exposed to the PCBs
from contaminated sediments, surface water, and/or prey. Species representing various trophic levels
living in or near the river were selected as receptor species for evaluating potential risks associated
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with PCBs. Exposure pathways by which these species could be exposed to PCBs were discussed
in the baseline ERA (USEPA, 1999c) and are summarized in the following section.
2.3.1 Exposure Pathways in the Lower Hudson River Ecosystem
Ecological receptors may be exposed to PCBs via various pathways. A complete exposure
pathway involves a potential for contact between the receptor and contaminant either through direct
exposure to the media or indirectly through food. Pathways are evaluated by considering
information on contaminant fate and transport, ecosystems at risk, and the magnitude and extent of
contamination (USEPA, 1997b).
Contaminant fate and transport and the magnitude and extent of contamination have been
discussed extensively in other Reassessment RI/FS reports, including the Baseline Modeling Report
(USEPA, 1999b), Data Evaluation and Interpretation Report (USEPA, 1997a), Low Resolution
Sediment Coring Report (USEPA, 1998a), and the baseline ERA (USEPA, 1999c). Exposure
pathways considered in this assessment are: ingestion of contaminated prey, ingestion of
contaminated sediments, and ingestion of contaminated surface water.
2.3.2 Ecosystems of the Lower Hudson River
The Lower Hudson River estuary is home to a wide variety of habitats. It is a valuable state
and local resource (NYSDEC, 1998a). Many commercially valuable fish and shellfish species
including striped bass, shad, Atlantic sturgeon, and blue crab use the estuary for spawning and as a
nursery ground. Over 16,500 acres in the estuary have been inventoried and designated significant
coastal fish and wildlife habitat. The NYS Natural Heritage Program has identified many areas
along the Hudson River estuary where rare plants, animals, or natural communities are found
(NYSDEC, 1999b). The estuary is also an important resting and feeding area for migratory birds,
such as eagles, osprey, songbirds, and waterfowl (NYSDEC, 1998a).
A number of distinct ecological communities including deepwater; shallows, mudflats, and
shore; tidal marsh; and tidal swamp communities are found in the Lower Hudson River. Brief
descriptions of these communities are provided below based on a publication of the New York State
Department of State and the Nature Conservancy (1990).
Deepwater- The deepwater community includes sections of the lower river with water depths
greater than six feet at low tide. Vegetation is limited to phytoplankton in the upper layers of the
water column, as light does not generally penetrate deep enough to support photosynthesis of rooted
plants. The deepwater community is composed of abundant animal life supported by organic material
originating in the watershed. Benthic invertebrates, fish, and fish eating predators (e.g., birds,
mammals) are found in this habitat. Fish found in the deepwater community include species such
as American shad, blueback herring, alewife, striped bass, Atlantic tomcod, and Atlantic and
shortnose sturgeon. Predators of deepwater fish can capture fish near the water's surface (e.g., bald
eagles, osprey) or below the surface of the water (e.g., cormorants, loons, and diving ducks).
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Shallows, Mudflats, and Shore- These communities include sections of the river found near
the low tide mark. Shallows are always below the low tide mark, mudflats are barely exposed at
low tide, and the shore is a zone largely exposed at low tide but inundated at high tide. The shallows
support a variety of vascular plants rooted in the bottom (e.g., waterweed, water celery, and various
pondweeds) and free floating plants (either in the water column or on the surface). Mudflats support
plants adapted to being submerged most of the day and then briefly exposed at low tide when they
are typically found encrusted in mud. In addition to vascular species, mudflats support significant
numbers of periphyton (attached algae) and bacteria that grow on mud or surfaces of vascular plants.
Shore areas are found along rocky or gravelly banks. Vegetation may be limited in areas subject to
waves, ice scour, and upland erosion.
Shallow waters support many zooplankton species and the animals that feed on them (e.g.,
fish larvae and fish). Many adult fish found in the shallow water are year-round Hudson River
residents including shiners, carp, white catfish, suckers, white and yellow perch, bass, sunfishes, and
darters in freshwater regions. Bay anchovies, killifish, silversides, winter flounder, and hog chokers
are found in more brackish sections of the river. Many anadromous (i.e., migrating) fish of the
deepwater community feed extensively in the shallows while preparing to return to the ocean. Many
fish also use the shallows as spawning and nursery grounds.
Numerous upper trophic level bird species (e.g., great blue heron, great egrets, least bittern)
feed in shallows and mudflats. Waterfowl feeding on aquatic plants and small fish and sandpipers
feeding on seeds, insects, and aquatic invertebrates are found in these communities.
Tidal Marsh- The tidal marsh community includes sections of the Hudson River where tidal
waters inundate plants specifically adapted to daily flooding. Lower marsh plants, adapted to daily
submersions, include broad-leaved plants such as spatterdock, pickerelweed, arrowhead, bulrushes,
and plantains. Upper marsh vegetation consists of plants adapted to partial flooding, which are
seldomly or never completely submerged. The upper marsh has a grassy appearance and is
dominated by narrow-leaved cattail and common reed.
Tidal marshes provide important feeding and breeding areas for many resident and transient
aquatic and terrestrial animals. Fish (e.g., killifish, darters, mummichogs, sunfish, and carp) come
into marshes at high tide to feed on invertebrates such as cladocerans, copepods, ostracods, and
chironomids. A variety of amphibians, reptiles, birds, and mammals feed on the fish and
invertebrates found in marshes. Hudson River tidal marshes support many bird species and large
populations of nesting birds, which includes a high density of breeding marsh birds.
Tidal Swamp- The tidal swamp community includes land adjacent to the Hudson River that
is regularly flooded by tidal waters. It is dominated by a closed canopy of trees (e.g., green and black
ash, red maple, and slippery elm). Below the canopy is a layer of shrubs and vines and at ground
level there is a layer of herbs. Tidal swamps occur exclusively in freshwater, either near freshwater
tributaries in brackish portions of the estuary or in upstream freshwater sections of the River.
The tidal swamp supports invertebrates and vertebrates feeding on plants, seeds, and organic
materials found in the swamp. Terrestrial herbivores and granivores include pheasants, rabbits,
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squirrels, muskrats, beaver, and deer. Predators of invertebrates and vertebrates found in the swamp
include salamanders, toads, snakes, turtles, shrews, foxes, weasels, and mink.
In addition to these communities, freshwater creek and upland forest communities are also
ecologically linked to the Hudson River. Exposure to PCBs originating in the River may occur via
the food chain or floodplain sediments.
Fish, amphibians, reptiles, birds, and mammals potentially found in or along the Hudson
River are listed in Tables 2-1 and 2-3 to 2-6 of the baseline ERA (USEPA, 1999c), respectively.
2.3.3 Exposure Pathways
The aquatic and terrestrial pathways for the Lower Hudson River are outlined below and
described in detail in Chapter 2 of the baseline ERA (USEPA, 1999c).
2.3.3.1 Aquatic Exposure Pathways
Aquatic and semi-aquatic organisms, such as fish, invertebrates, amphibians, and
reptiles (e.g., water snakes), are exposed to PCBs through:
• Direct uptake from water;
• Uptake from sediment; and
• Uptake via food.
2.3.3.2 Terrestrial Exposure Pathways
Terrestrial and semi-terrestrial animals, such as amphibians, reptiles, birds, and mammals,
can be exposed to PCBs via:
• Food uptake;
• Surface water ingestion;
• Incidental sediment ingestion;
• Contact with floodplain sediments/soils; and
• Inhalation of air.
Food uptake of contaminated prey is considered to be the primary PCB exposure pathway (USEPA,
1999c).
2.4 Assessment Endpoints
Assessment endpoints are explicit expressions of actual environmental values (e.g.,
ecological resources) that are to be protected (USEPA, 1992). They focus the risk assessment on
particular components of the ecosystem that could be adversely affected by contaminants from the
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site (USEPA, 1997b). These endpoints are expressed in terms of individual organisms, populations,
communities, ecosystems, or habitats with some common characteristics (e.g., feeding preferences,
reproductive requirements). In addition to protection of ecological values, assessment endpoints may
also encompass a function or quality that is to be maintained or protected.
The assessment endpoints selected for the ERA Addendum focus on the protection and
maintenance of local fish and wildlife populations exposed to PCBs in Hudson River sediments and
water through sediment and surface water ingestion, uptake from water, and indirect exposure to
PCBs via the food chain. Because PCBs are known to bioaccumulate, an emphasis was placed on
exposure at various levels of the food chain to address PCB-related risks at higher trophic levels.
The assessment endpoints selected to evaluate future risks in the Lower Hudson are:
• Benthic aquatic life as a food source for local fish and wildlife.
• Survival, growth, and reproduction of:
- local forage fish populations;
- local omnivorous fish populations; and
- local piscivorous fish populations.
• Protection (i.e., survival, growth, and reproduction) of local wildlife including:
- insectivorous bird populations;
- waterfowl populations;
- semi-piscivorous/piscivorous bird populations;
- insectivorous mammal populations;
- omnivorous mammal populations; and
- semi-piscivorous/piscivorous mammals populations.
• Protection of threatened and endangered species.
• Protection of significant habitats.
The selected assessment endpoints along with specific ecological receptors and measures of
effect are listed in Table 2-1. These endpoints reflect a combination of values that have been
identified by USEPA, New York State Department of Environmental Conservation (NYSDEC), US
Fish and Wildlife Service (USFWS), and National Oceanic and Atmospheric Administration
(NOAA) as being important, and/or habitats or species that have been identified as ecologically
valuable.
2.5 Measurement Endpoints (Measures of Effect)
Measures of effect provide the actual measurements used to estimate risk, as described in the
baseline ERA (USEPA, 1999c). Because of the complexity and inherent variability associated with
ecosystems, there is always a certain amount of uncertainty associated with estimating risks.
Measurement endpoints typically have specific strengths and weaknesses related to the data quality,
study design and execution, and strength of association between the measurement and assessment
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endpoint. Therefore, it is common practice to use more than one measurement endpoint to evaluate
an assessment endpoint, when possible.
Measures of effect used to evaluate each assessment endpoint in this addendum are the same
as those used in the baseline ERA (USEPA, 1999c) and include:
• Modeled total PCB (i.e., Tri+ congeners) body burdens in fish, birds, and mammals for 25
years (1993 to 2018) to determine exceedance of effect-level thresholds based on toxicity
reference values (TRVs) derived in the baseline ERA (USEPA, 1999c).
• Modeled TEQ-based PCB body burdens in fish, birds, and mammals for 25 years (1993 to
2018) to determine exceedance of effect-level thresholds based on TRVs derived in the
baseline ERA (USEPA, 1999c).
• Modeled total PCB egg concentrations in birds for 25 years (1993 to 2018) to determine
exceedance of effect-level thresholds based on TRVs derived in the baseline ERA (USEPA,
1999c).
• Modeled TEQ-based PCB egg concentrations in birds for 25 years (1993 to 2018) to
determine exceedance of effect-level thresholds based on TRVs derived in the baseline ERA
(USEPA, 1999c).
• Modeled PCB concentrations in fresh water for 25 years (1993 to 2018) compared to NYS
Ambient Water Quality Criteria (AWQC) for the protection of benthic aquatic life and
protection of wildlife from toxic effects of bioaccumulation (NYSDEC, 1998b).
• Modeled PCB concentrations in sediment for 25 years (1993 to 2018) compared to
applicable sediment benchmarks such as NOAA Sediment Effect Concentrations for PCBs
in the Hudson River (NOAA, 1999), NYSDEC Technical Guidance for Screening
Contaminated Sediments (1999a), Ontario sediment quality guideline (Persaud et al. 1993),
and Washington Department of Ecology guidelines for protection of aquatic life (1997).
• Available field observations on the presence and relative abundance of Lower Hudson
River fish and wildlife as an indication of the ability of the species to maintain populations.
• Available field observations on the presence and relative abundance of the wildlife species
using significant habitats within the Lower Hudson River as an indication of the ability of
the habitat to maintain populations.
Risk hypotheses posed as risk questions, along with specific measurement endpoints selected for
each assessment endpoint, are provided in Table 2-2.
Effect-level concentrations are measured by TRVs. TRVs are exceeded when the modeled
dose or concentration for the site is greater than the benchmark dose or concentration (i.e., toxicity
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quotient [TQ] exceeds 1). Equations for estimating avian and mammalian dietary doses, avian egg
concentrations, and fish body burdens are provided in Chapter 3 of the baseline ERA (USEPA,
1999c).
Population-level effects are determined for each receptor species by evaluating the species
life-history and the magnitude of the TQ over time. TQs equal to or greater than one across the
entire 25-year modeling period suggests sustained risk. If the life span of receptor covers only a
fraction of the modeling period, then population level effects are more likely given the time
trajectory. The results of all measurement endpoints, such as modeled total PCB dietary doses and/or
egg concentrations, modeled TEQ-based PCB dietary doses and/or egg concentrations, exceedances
of benchmarks and criteria, are used in a weight-of-evidence approach. For receptors with small
populations (e.g., threatened or endangered species), individual-level effects may place the
population at risk.
2.6 Receptors of Concern
Potential adverse effects are evaluated for selected receptor species that represent various
trophic levels living in or near the Lower Hudson River. These receptors are used to establish
assessment endpoints for evaluation of risk. Receptors were selected to represent different trophic
levels, a variety of feeding types, and a diversity of habitats (e.g., aquatic, wetland, shoreline).
Specific fish, avian, and mammalian species were selected for evaluation as suiTogate species for the
range of species likely to be exposed to PCBs in the Lower Hudson River. As described in the
baseline ERA (USEPA, 1999c), species were selected based on species sensitivity to PCBs, societal
relevance of selected species, discussions with agency representatives, and comments received on
the ERA Scope of Work (USEPA, 1998c; USEPA, 1999a).
2.6.1 Fish Receptors
The Hudson River is home to over 200 species of fish (Stanne et al. 1996). The following
eight fish species, representing a range of trophic levels were evaluated in the ERA and are also
evaluated in the ERA Addendum:
• Spottail shiner (Notropis hudsonius) - forage fish;
• Pumpkinseed (Lepomis gibbosus) - forage fish;
• Brown bullhead (Ictalurus nebulosus) - omnivore;
• White perch (Morone americand) - semi-piscivore;
• Yellow perch (Percaflavescens) - semi-piscivore;
• Largemouth bass (Micropterus salmoides) - piscivore;
• Striped bass (Morone saxatilis) - piscivore; and,
• Shortnose sturgeon (Acipenser brevirostrum) - omnivore (evaluated only in the
context of endangered and threatened species).
These forage fish, piscivorous/semi-piscivorous fish, and omnivorous fish provide a general estimate
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of PCB bioaccumulation potential according to trophic status and are designed to be protective of
potential PCB exposures to other, less common species. Detailed profiles of the fish species are
provided in Appendix D of the baseline ERA (USEPA, 1999c).
2.6.2 Avian Receptors
Five avian receptors were selected to represent various trophic levels and habitat use of the
numerous year-round residents and migratory bird species found along the Hudson River.
• Tree swallow (Tachycineta bicolor)- insectivore;
• Mallard (Anas platyrhychos) - aquatic plants and animals;
• Belted kingfisher (Ceryle alcyori) - piscivore;
• Great blue heron (Ardea herodias) - piscivore; and
• Bald eagle (Haliaeetus leucocephalus) - piscivore.
Detailed life history profiles of the avian species listed below are provided in Appendix E of the
baseline ERA (USEPA, 1999c).
2.6.3 Mammalian Receptors
The potential mammalian receptors found along the Hudson River also represent various
trophic levels and habitats. The four mammals selected to serve as representative receptors in
baseline ERA and the ERA Addendum are:
• Little brown bat (Myotis spp.) - insectivore;
• Raccoon (Procyon lotor) - omnivore;
• Mink (Mustela visori) - piscivore; and
• River Otter (Lutra canadensis) -piscivore.
Detailed profiles of these mammalian species are provided in Appendix F of the baseline ERA
(USEPA, 1999c).
2.6.4 Threatened and Endangered Species
Federal and State threatened and endangered species found in the Lower Hudson Valley are:
• Kamer blue butterfly (Lycaeides melissa samuelis) - federal- and State-listed endangered;
• Shortnose sturgeon (Acipenser brevirostrum) - federal- and State-listed endangered;
• Northern cricket frog (Acris crepitans)-Stale-\isled endangered;
• Bog turtle (Clemmys muhlenbergii) - State-listed endangered;
• Blanding's turtle (Emydoidea blandingii) - State-listed threatened;
• Timber rattlesnake (Crotalus horrldus)- State-listed threatened;
• Peregrine falcon (Falco peregrinus) - State-listed endangered;
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• Bald eagle (Haliaeetus leucocephalus) - State-listed endangered and federal-listed
threatened;
• Osprey (Pandion haliaetus) - State-listed threatened;
• Northern harrier (Circus cyaneus) - State-listed threatened;
• Red-shouldered hawk (Buteo lineatus) - State-listed threatened;
• Indiana bat (Myotis sodalis) - federal-listed endangered; and
• Eastern woodrat (Neotoma magister) - State-listed endangered.
Profiles of these threatened and endangered species are provided in Appendix G of the baseline ERA
(USEPA, 1999c).
New York State avian species of concern found in the vicinity of the Hudson River include
the least bittern (Ixobrychus exilis), Cooper's hawk (Accipiter cooperii), upland sandpiper (Bartramia
longicauda), shorteared owl (Asia flammeus), common nighthawk (Chordeiles minor), eastern
bluebird, (Sialia sialis), grasshopper sparrow (Ammodramus savannarum), and vesper sparrow
(Pooecetes gramineus).
Amphibians of special concern listed by NYS potentially found along the Lower Hudson
River include the Jefferson salamander (Ambystoma jeffersonianum), bluespotted salamander
(Ambystoma laterale, and spotted salamander (Ambystoma maculatum). Reptiles of special concern
include spotted turtle (Clemmys guttata), wood turtle (Clemmys insculpta), diamondback terrapin
(Malaclemys terrapin), and worm snake (Carphophis amoenus).
The Hudson's tidal habitats support a number of rare plant species. A list of these species
is provided in Appendix G of the baseline ERA (USEPA, 1999c).
This ERA Addendum evaluates risks to threatened and endangered species as represented
by the bald eagle and shortnose sturgeon, consistent with the baseline ERA.
2.6.5 Significant Habitats
All portions of the Hudson River have value for plants and animals. However, 34 specific
sites in the Lower Hudson River have been designated as Significant Coastal Fish and Wildlife
Habitats under NYS' Coastal Management Program. Five additional sites have been identified as
containing important plant and animal communities to bring the total number of sites to 39 (see
Table 2-11 of the baseline ERA [USEPA, 1999c]). Four of these areas comprise the Hudson River
National Estuarine Research Reserve (NERR), administered by NYS in partnership with NOAA.
Significant habitats contain areas that are unique, unusual, or necessary for continued
propagation of key or rare and endangered species. Rare ecological communities and areas of
concern often form part or all of the areas considered to be significant habitats. The community
types, rare species, and valuable species found at each of these sites are summarized in Table 2-3
based on information provided in New York State Department of State and The Nature Conservancy
(1990).
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Chapter 3
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3.0 EXPOSURE ASSESSMENT
The exposure assessment characterizes exposure concentrations or dietary doses for the
selected receptors. Exposure concentrations are estimates of the PCB concentrations modeled under
site-specific assumptions and are expressed as total PCBs (as Tri+) and dioxin-like toxic
equivalencies (TEQs) to which selected receptors are exposed.
Several exposure models were developed to evaluate the potential risk of PCB exposures
under baseline conditions. Sediment and water concentrations were estimated using the model
developed by Farley et al. (1999) for the Hudson River Foundation (i.e., independent of USEPA's
Reassessment RI/FS), as described later in this section. The FISHRAND model (USEPA, 1999c and
2000) was used to calculate all fish body burdens from the sediment and water column
concentrations forecast by the Farley model. The results of these models were used to estimate
dietary doses to the avian and mammalian receptors for the period 1993-2018. Modeled fish body
burdens were compared directly with the fish toxicity reference values to determine potential risk.
Egg concentrations in piscivorous receptors were estimated by applying a biomagnification
factor from the literature (Giesy et al, 1995) assumed to be 28 for total PCBs and 19 for TEQ-based
concentrations. These factors were applied to both the observed and modeled fish concentrations
to calculate egg concentrations in the bald eagle, great blue heron, and belted kingfisher. The
USFWS data were used to determine a tree swallow egg to emergent aquatic insect (assumed as
benthic invertebrate) biomagnification factor. The USFWS data were also used to establish a
mallard duck egg to emergent aquatic insect biomagnification factor.
PCB exposures are evaluated using total PCB concentrations expressed in terms of the
trichlorinated (Tri+) and higher PCB congeners in a series of body burden, dietary dose, and/or egg
concentration models and using dioxin-like TEQ exposure concentrations based on toxic equivalency
factors (TEFs) in a series of body burden, dietary dose and/or egg concentration models. As
discussed in Appendix K of the baseline ERA (USEPA, 1999b), the Tri+ sum is nearly identical to
the total PCB concentration in fish due to the lack of significant concentrations of monochloro or
dichloro congeners in fish tissue.
These approaches involve the construction of a series of models to first estimate PCB
concentrations in sediment, water and white perch via the Farley model (Farley et al., 1999) with
subsequent application of the FISHRAND model (USEPA, 1999c and 2000) to estimate
concentrations in fish tissue, and finally the construction of exposure models to estimate body
burdens, dietary doses, and/or egg concentrations in the various ecological receptors. These estimates
were then compared to the toxicity reference values (TRVs) discussed later in this report.
3.1 Quantification of PCB Fate and Transport: Modeling Exposure
Concentrations
The results of the sampling studies for the Reassessment RI/FS have been previously
described in several Phase 2 reports, in particular the DEIR (USEPA, 1997) and the ERA (USEPA,
1999c). In this report, a model of Lower Hudson PCB transport developed by Farley et al. (1999),
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supplemented by two USEPA models (HUDTOX and FISHRAND; USEPA, 1999b and 2000), is
applied to estimate current and future levels of PCB contamination in sediments, water and fish. The
ERA Addendum uses a forecast of 25 years, from 1993 to 2018) while the Mid-Hudson Human
Health Risk Assessment (USEPA, 1999d) uses up to a 41 year forecast (1999 to 2040). The forecast
data are identical for the overlapping period (i.e., 1999 to 2018).
The development and calibration of the model developed by Farley et al. is described in
Farley et al.(1999) and is not repeated here. The model's calibration used USEPA sampling data
from the Lower Hudson. The estimation of future PCB loads to the Lower Hudson from the Upper
Hudson was based on results from the USEPA's Upper Hudson model (HUDTOX) (USEPA, 1999c
and 2000). Estimation of fish body burdens was achieved through the use of the Farley et al. (1999)
model as well as USEPA's FISHRAND model which was also developed as part of the Upper
Hudson modeling effort (USEPA, 1999c and 2000).
This discussion of the modeling effort is comprised of three sections. The first, Section 3.1.1,
describes the modeling approach used and provides details on how the fate, transport and
bioaccumulation models were used. Because pre-existing models are used, no discussion of the
construction and calibration of the models is presented and the reader is referred to the original
modeling reports for additional information. Section 3.1.1 also provides a qualitative discussion on
model verification by comparing the model output to previous modeling efforts as well as to sample
data from the USEPA, NOAA and NYSDEC. Section 3.1.2 presents the model results which are
used in the ERA Addendum and the Mid-Hudson HHRA (USEPA, 1999d). Section 3.1.3 provides
a brief summary of the modeling analysis. Section 3.2 provides a summary of the exposure point
concentrations used in the ERA Addendum.
3.1.1 Modeling Approach
Four separate models are used to calculate the exposure point concentrations in the Lower
Hudson. The fate and transport model developed by USEPA for the Upper Hudson River
(HUDTOX) provides the flux of PCBs over the Federal Dam into the Lower Hudson River (USEPA,
1999b). These results represent an external input to the Lower Hudson River fate and transport
model (i.e., the Farley et al., 1999 model). The Farley et al. (1999) fate and transport model
developed specifically for the Lower Hudson River is used to generate the water and sediment
concentrations for the Lower Hudson River risk assessments. The water and sediment concentrations
from the Farley fate and transport model are used as input for the USEPA bioaccumulation model
(FISHRAND) to generate the PCB body burdens for all fish species examined in the Lower Hudson.
The Farley bioaccumulation model was applied to yield PCB concentrations in white perch and
striped bass for comparison purposes only.
3.1.1.1 Use of the Farley Models
The model segmentation for the Farley et al. (1999) fate and transport and bioaccumulation
models is shown in Figure 3-1. Water column segments 1 to 14 correspond to the Lower Hudson
between RM 153.5 and 14. There are 30 water column segments in all, which are combined into five
food web regions. Food web regions 1 and 2 cover the spatial extent of the Lower Hudson River risk
assessments. The sediment and dissolved water column concentrations of PCBs obtained for each
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of the segments of the fate and transport model are averaged by food web region utilized by the
bioaccumulation model. Detailed descriptions of the models are given in Farley et al. (1999). Few
changes were needed to make the models usable for the ERA Addendum and Mid-Hudson HHRA.
Unlike the HUDTOX model developed for the Upper Hudson, the Farley et al. (1999) model
is based on five separate homologue groups (dichloro to hexachloro hornologues) and requires
external load estimates for each group. For comparison, the HUDTOX model uses the sum of the
trichloro and higher hornologues (Tri+), total PCBs and 5 individual congeners. In the original
analysis by Farley et al. (1999), there were few bases on which to estimate future loads at the Federal
Dam and so the original model was only run through the year 2001 (i.e., to 2002).
For the ERA Addendum, the flux over the Federal Dam for each homologue is derived from
the flux of Tri+ PCBs given by the HUDTOX model (USEPA, 1999c and 2000). In order to use the
Tri+ flux given by the HUDTOX model, a basis for conversion of the Tri+ load to individual
homologue loads was required. This was accomplished through the use of Tri+ to homologue
conversion factor for each homologue group. These factors were determined by analyzing the
available USEPA and General Electric Company water column data. Table 3-1 gives the means of
conversion for each homologue during both the calibration and forecast periods. This conversion is
described in Appendix A.
The Farley etal. (1999) models were originally designed to run for a 15 year period, 1987-
2002. Because a 40 year forecast of concentrations is required for the Mid-Hudson HHRA, the
models are run in 15 year increments with the final conditions in each model segment and each
modeled species becoming the initial conditions for the next 15 years. The major external PCB load
to the Lower Hudson, i.e., the load from the Upper Hudson, was estimated using the 40-year forecast
from the HUDTOX model, assuming a constant concentration of 10 ng/L at the upstream boundary
of the HUDTOX model (USEPA, 2000). For the purposes of this ERA Addendum, only the model
output from the period 1993 to 2018 was used.
Prior to using the forecast from the Farley et al. (1999) models in the risk assessments, an
examination of the Farley model results was performed for the calibration period 1987 to 1997. In
this examination, the original calibration curve developed by Farley et al. (1.999) was compared with
model results produced using the HUDTOX model PCB loads to the Lower Hudson. In this fashion,
the effects of any differences in Upper Hudson load assumptions could be examined. The results of
this comparison are discussed later in Section 3.1.1.3.
The Farley et al. (1999) models have been updated since the report was finalized in March
1999. In the fate and transport model, the suspended solids loads to Newark Bay were found to be
too high and were corrected. This correction will have the greatest impact on food web region 3 and
water column segments 15 and higher. Because these areas are not considered in the ERA
Addendum and Mid-Hudson HHRA, the impact of these changes is minimal and this revision was
not included in this Lower Hudson modeling analysis. In ignoring this correction, the maximum
effect on food web region 2 (RM 14 to 60) would be slightly increased PCB concentrations,
potentially yielding a slight overestimate of the risks for RM 14 to 60. Because the resulting risk
estimate would still be protective of human health and the environment, no effort was made to
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update the Lower River fate and transport calculations to reflect the minor correction made to Farley
et al (1999).
The Farley et al. (1999) bioaccumulation model also underwent revisions after the original
report was finalized. These revisions relate to the absorption efficiencies for PCBs across the fish
digestive system and the estimation of lipid levels in fish. The July 1999 version of the Farley et al.
(1999) bioaccumulation model incorporating these revisions (Cooney, 1999) is used in this report.
3.1.1.2 Use of FISHRAND
The FISHRAND model was used to model PCB concentrations in all of the fish receptors
examined in the ERA Addendum except for striped bass. A full description of this model is given
in USEPA (2000). The differences from the application of the FISHRAND model to the Upper
Hudson River to the Lower Hudson River are:
• Water and sediment concentrations estimated from the Farley et al. (1999) fate and
transport model are used;
• The percent lipid distribution is significantly different for the Lower Hudson River
largemouth bass with an average lipid content of 2.5% in the Lower Hudson River versus
1.3% in the Upper Hudson River;
• The total organic carbon value for sediment segments used in the Farley et al. (1999) fate
and transport model is used; and
• the Kow values specified in USEPA (2000) for the Upper Hudson River below the
Thompson Island Dam are applied to the Lower Hudson River.
Estimation of Striped Bass Body Burdens in the Lower Hudson
The Farley bioaccumulation model was used to estimate PCB levels for striped bass which
migrate up to food web region 2 (i.e., fish which remain downstream of the salt front, approximately
RM 60). The model does not provide striped bass concentrations in food web region 1 (i.e., the
freshwater Lower Hudson). In order to estimate striped bass body burdens in food web region 1, the
largemouth bass body burdens estimated from the FISHRAND model were multiplied by the ratio
of striped bass to largemouth bass body burdens (MCA, 1999). Observed striped bass and
largemouth bass concentrations from NYSDEC data were used to construct the ratio at RMs 152 and
113. The averaged concentrations for each year and species are shown in Table 3-2. Ratios for
striped bass to white perch are also presented in the table for comparison.
Table 3-2a shows that the average ratio between measured striped bass and largemouth bass
at RM 152 is approximately 2.5 (standard deviation = 1.6). In all instances, the data were restricted
to fish larger than 25 cm to represent fish that would actually be caught and kept by an angler. This
criterion was met by all largemouth bass samples but resulted in the exclusion of several striped bass
samples. A similar ratio is obtained between striped bass and white perch, 3.43 (standard deviation
of 4.1). Notably, if the year 1990 is eliminated from the white perch comparison, then the ratio
becomes 1.62 (standard deviation of 0.4). However, elimination of an entire year of data given the
small sample size is unjustified and was not considered.
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The striped bass to largemouth bass ratio was also examined on a monthly basis at RM 152
as shown in Table 3-2b. All largemouth bass and white perch samples were collected in May and
June at this location. Striped bass were collected in June, July, August, arid October at RM 152.
Three separate ratios were calculated, comparing the May-June largemouth bass with the June-
August, June-July and June-only striped bass data. In all cases, the calculated ratios were essentially
the same, ranging between 2.5 and 2.6. Based on these results, the ratio of 2.5 was used to
approximate striped bass concentrations for 1998 to 2040 for RM 152. This is accomplished by
simply multiplying the modeled concentrations in largemouth bass at this location by 2.5 to estimate
the striped bass concentrations.
At RM 113, all of the largemouth bass and striped bass data were obtained in May and June
sampling events, so a similar comparison could not be made. At RM 113, the striped bass to
largemouth bass ratio is very different. The ratios in this region are much lower than at RM 152,
with an average ratio of 0.52 and also exhibit less variability (standard deviation = 0.2). the striped
bass concentrations are estimated in the same fashion as at RM 152, only with a multiplier of 0.52
instead of 2.5.
3.1.1.3 Comparison to the Farley et al (1999) Model for the Period 1987 to 1997
In order to assess the impact to the Farley et al. (1999) model made by changing the Upper
Hudson River PCB loads, the model inputs and outputs were compared. Specifically, the external
load estimates (i.e., an input to the Farley model) made by Farley et al.(1999) were first compared
with the external loads estimated via HUDTOX for the calibration period 1987-1997. Differences
in these load estimates should be evident in the model output because the Upper Hudson is such a
major source of PCBs to the Lower Hudson.
Secondly, the Farley et al. (1999) model output in the form of white perch and striped bass
body burdens were then compared between the March 1999 Farley et al. (1999) model results and
the Farley et al. (1999) models rerun with the HUDTOX estimates of PCB flux over the Federal
Dam.
The results of the Upper Hudson load comparison show the importance of the Upper Hudson
in smoothing loads originating above Thompson Island (TI) Dam. Overall, both the Farley et al.
(1999) and HUDTOX load estimates deliver approximately the same amount of PCBs to the Lower
Hudson over the ten year calibration period (1987 - 1997). The comparison of the fish body burdens
shows that the adjustments to the model made by Farley et al. (Cooney, 1999) are more important
than any differences in the sequence of PCB loads assumed by Farley et al. (1999) and HUDTOX.
Comparison of HUDTOX and Farley et al. (1999) PCB Load Estimates at the Federal Dam
The revision of the flux of PCBs over the Federal Dam at Troy is the only modification made
to the March 1999 Farley fate and transport model for the ERA Addendum and Mid-Hudson HHRA.
The difference in magnitude between Farley's original flux estimate and that derived from the
HUDTOX model can be seen in Table 3-3. This table shows the two estimates of the PCB
homologue loads. The cumulative tri-through-hexa-load estimates over the Federal Dam from the
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Farley model compare favorably with the estimates from HUDTOX for the period 1987-1997. The
largest difference is 101 kg for the tri homologue, representing a cumulative difference of about 4
percent relative to the estimate by Farley et al. (1999) (see Table 3-3). Conversely, the estimates for
the di homologue differ by a greater amount, 895 kg (76 percent relative to Farley et al. 1999). The
Farley et al. (1999) model used the General Electric Company water column samples at TI Dam to
estimate all homologue loads during the calibration period. As described in Appendix A and
presented in Table A-2, the di homologue fraction based on HUDTOX was calculated from the Tri+
PCBs by applying a ratio developed from the USEPA Phase 2 water column data. Notably, the
largest differences are for the homologue which matters least to Lower Hudson fish body burdens.
It is noteworthy as well that the cumulative HUDTOX loads are closer to the load estimates made
on a strictly statistical basis, as presented in the DEIR (USEPA, 1997).
The cumulative loads from both modeling estimates are plotted against time in Figure 3-2.
Evident in all diagrams is a distinct difference in the timing of the loads to the Lower Hudson.
Specifically, the loads estimated by Farley et a/.(1999) show a distinct rise in the 1991-1993 period
while those estimated from HUDTOX show a more gradual rise through the calibration period. This
is a result of the assumptions used in creating the two estimates. In the estimate by Farley et al.
(1999), the measured loads at TI Dam are directly translated to the Lower Hudson. In the HUDTOX-
based estimates, loads at TI Dam are affected by the intervening 35 miles of the Upper Hudson,
essentially buffering these loads and spreading them out over a longer time period. These
assumptions bear directly on the Lower Hudson fish body burdens because the external load
determines much of the fish exposure.
For tri through hexa homologues, the Farley et al. (1999) estimate is less than the HUDTOX
estimate from 1987-1991 and greater than the HUDTOX estimate for 1992-1997, yielding
cumulative loads which are quite similar. The Farley et al. (1999) estimate is always less than the
HUDTOX estimate for the di homologue. This is attributed in part to the lower sensitivity of the
General Electric Company data which was used by Farley et al. (1999) for this estimate, as discussed
above. In addition, the Farley et al. (1999) model estimates for the period 1987-1991 were based on
a total PCB load trajectory derived from an earlier modeling analysis prepared by Thomann (1989).
The homologue distribution was assumed to be the same as that measured in 1991 by the General
Electric Company. Conversely, the HUDTOX model is calibrated to the USGS data during this
period. Lastly, it is unclear whether the General Electric Company data used by Farley et al. (1999)
had been corrected for the BZ#4 bias as documented by QEA in O'Brien and Gere (1998). Overall,
it is apparent that the assumptions made by Farley et al. and the loads derived from HUDTOX will
yield different concentrations of PCBs on the Lower Hudson on a year-to-year basis. In the latter
period of record, 1994-1998, the results appear to converge as upstream loads become more regular
and predictable. (Note the parallel rates of increase in the cumulative curves.)
Comparison of White Perch and Striped Bass Body Burdens
Two changes in the Farley et al. (1999) bioaccumulation model are reflected in the
comparisons described below. First, the timing and magnitude of the Upper Hudson loads to the
Lower River have been changed as described above. Second, the bioaccumulation model itself has
been modified by Farley et al. (1999), changing the response between the exposures and the fish
18 TAMS/MCA
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body burdens. In this correction (Cooney, 1999), the lipid content of the modeled species was
decreased to match the lipid content of fish sampled by NYSDEC in the 1990s. This serves to
decrease the body burdens predicted by the application of the Farley et al. (1.999) model regardless
of the assumptions of the upstream loading.
The change in the body burden for white perch and striped bass resulting from these changes
can be seen by plotting the model results from the March 1999 report (Farley et al., 1999) and this
analysis on the x and y axes, respectively, for each time step (approximately a 2 week period) over
the entire calibration period (1987 to 1997). Tri+ PCBs (here defined as the sum of the tri through
hexa homologues) are plotted because this fraction is most prominent in the fish body burdens (there
is little contribution from the di fraction). This also minimizes the effect of the different bases used
to estimate the di homologue fraction.
The results are shown in Figure 3-3 for the white perch and Figure 3-4 for the striped bass.
The food web region 1 white perch values differ greatly, with the March 1999 values from Farley
et al. (1999) being distinctly higher. The scatter in the data is attributed to the sensitivity of the white
perch model in this food web region to the Upper Hudson River PCB loads. Nonetheless, the paired
results do form a linear trend (although not a line), indicating a similar kind of response in both
models. The displacement of the line away from the 1:1 line is largely attributed to the revisions to
the bioaccumulation model made since the modeling report was released (Cooney, 1999 and Farley
et al., 1999). The scatter about the line is attributed to the loading differences, with the points falling
above the line when the HUDTOX loading estimates are higher than those given by Farley et al.
(1999). The points fall below the line when the converse is true. The plot of white perch estimates
in food web region 2 is displaced from the 1:1 line by an amount similar to that for food web region
1 but the slope and the scatter in the data are much less as indicated by the difference in the R
values. The decreased scatter is attributed to a diminished sensitivity to the Upper Hudson loads in
this region of the Hudson, with food web region 1 of the Hudson serving to buffer the variations in
the Upper Hudson loads prior to their delivery to food web region 2.
The striped bass values (food web region 2 only) for both model runs is similar with slopes
and regression coefficients near 1, showing that the modeled striped bass is not sensitive to this
change in Upper Hudson River PCB loads.
3.1.1.4 Comparison Between Model Output and Sample Data
While the comparisons described in Section 3.1.1.3 are useful in examining the effects of
model assumptions relative to the original model, it is also important to examine the correlation of
the model output with the measurement results. Data from the Farley et at. (1999) model run with
the Upper Hudson River loads determined by HUDTOX were compared to the water, sediment and
fish samples taken from between 1987 and 1997 in order to test the accuracy of the Farley et al.
(1999) model with the revised upstream loads. USEPA Phase 2 water and sediment samples and
NYSDEC fish samples are available from the Lower Hudson River for this time period. Because the
water and sediment samples from this portion of the river are relatively few and limited to one or two
years, this comparison provides only a limited assessment of the fate and transport model approach.
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The NYSDEC fish data represent a more extensive data set and, therefore, provide a better basis for
assessing the overall modeling approach.
Dissolved Phase PCBs in Water Column
Modeled dissolved phase PGB concentrations are plotted by river mile for April and August
1993 against the USEPA Phase 2 water column samples in Figure 3-5. The dissolved phase data are
especially important because it is the data input from the Farley fate and transport model into the
bioaccumulation models. For April 1993, the model agrees reasonably well with the sampled data
at RMs 77 and 125, but is 0.02 ^ug/L lower than the sampled data at RM 152. For August 1993, the
modeled results are from 0.01 to 0.02 yUg/L (or a factor of 2 to 3) lower than the sampled data. These
results suggest that the Farley model may overestimate losses from the water column during the
summer period. Nonetheless, the model trend is similar to the measured trend, with a gradual
decline in concentration with RM, as would be expected in the absence of additional significant
external sources of PCBs.
The dissolved-phase homologue patterns for August and September 1993 are shown in Figure
3-6. The homologue pattern derived from the Farley et al. (1999) model with the HUDTOX loads
yields fairly good agreement with the sampled data based on the relative proportions of the
homologues. Again, the modeled concentrations are lower for this period than the sampled
concentrations, indicating that the possible overestimate of water column loss in the summer affects
the entire pool of congeners and not just a single homologue.
Sediment Concentrations
Modeled surface sediment concentrations from 0-2.5 cm and 2.5-5 cm are plotted against the
USEPA Phase 2 ecological samples (approximately 5 cm in depth). The modeled data fall within the
range of the sampled concentrations for all RMs except for RM 47. At this location, the modeled
values are about 0.1 ppm below the lowest sampled value. These results suggest that the model is
able to represent the general level of sediment contamination in the river as a function of distance
downstream.
Fish Body Burdens
The Farley bioaccumulation model yielded body burdens for white perch in regions 1 and 2
and striped bass in region 2 only. The modeled white perch and striped bass body burdens are plotted
against sample data from NYSDEC in Figures 3-8 and 3-9. For white perch, the modeled data fall
within the range of the sampled data for all years except 1990 in food web region 1. In addition, the
model values fall within + 50 percent of the mean value for all measurement years except 1990 (the
mean is represented by the horizontal bars). This includes five of the six sampling events in food
web region 1 and the one sampling event in food web region 2. In 1990, the modeled data are slightly
higher in concentration then the maximum sampled value.
For striped bass (shown in Figure 3-9), the modeled data nearly always fall within the range
of sampled values and are close to the mean sampled values, indicating a satisfactory level of
agreement.
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Although there is a relatively limited data set for PCBs in sediment, water and fish, the model
is able to replicate the measurements fairly well, particularly for the fish data. This indicates that the
use of the Farley etal. (1999) models with the HUDTOX Upper Hudson load estimate is consistent
with the available data and should provide a reasonable basis for estimating future concentrations
of PCBs in the Lower Hudson River.
3.1.1.5 Comparison of White Perch PCB Body Burden between the Farley Model (Using
Upper River Loads from HUDTOX) and FISHRAND
White perch is the only species that is common to both the Farley et al. (1999)
bioaccumulation model (as modified by Cooney, 1999) and the FISHRAND model, providing a
point of comparison between the models. Similar results for both models would suggest a consistent
basis on which to assess exposures and exposure-related risks to humans and the biota. As a basis
for comparison, the results of the 70-year forecast for each model are compared for several locations.
White perch body burdens of Tri+ PCBs are plotted against time for each location modeled
by FISHRAND in Figure 3-10. It is important to note that the Farley model predicts average fish
body burden for the entire food web region 1 while FISHRAND has been applied separately to
several locations within the region. In Region 1, the Farley model predicts lower concentrations than
the FISHRAND model at RM 152. At RMs 113 and 90 the FISHRAND and Farley models agree
fairly well, wherein FISHRAND results are only sometimes higher in concentration than the Farley
model. In food web region 2, the Farley model predicts higher PCB concentrations than the
FISHRAND model in the early portion of the forecast. Both models show a steady drop off in PCB
concentration with time and appear to approach a similar asymptote.
The Farley model estimates for white perch body burdens from each region of the river are
plotted against the corresponding FISHRAND estimates in Figure 3-11 for each time step in the
forecast. The linear fits to the data are reasonable with regression coefficients ranging from 0.825
to 0.916. The difference in the magnitude of the concentrations are evident in the slopes. At RM 152,
the slope is 1.27 where the FISHRAND concentrations are higher. At RM 50, the slope is 0.594
where the FISHRAND concentrations are lower. Overall, the agreement is considered good and
indicates that both models provide a consistent basis for estimating future fish body burdens. This
also indicates that it is reasonable to apply the FISHRAND outside its original calibration region
(i.e., the Upper Hudson River) and that the application of FISHRAND in the Lower Hudson will
produce reasonable future estimates of the various fish body burdens. This conclusion is further
supported by the comparisons to Lower Hudson data in the next subsection.
3.1.1.6 Comparison Between FISHRAND Output and Sample Data From NYSDEC and
USEPA
Fish body burdens modeled using FISHRAND were compared to the NYSDEC, NOAA and
USEPA sample data on both a wet weight basis and a lipid-normalized basis. This is shown in Figure
3-12a for the largemouth bass, white perch, brown bullhead and yellow perch at RM 152. Similarly,
results for largemouth bass, white perch and yellow perch at RM 113 are shown in Figure 3-12b.
These species plus striped bass represent the main human exposure routes. They are also important
21 TAMS/MCA
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for the larger ecological receptors. These species also have larger data sets than other species and
cover much of the Lower Hudson. In each diagram, the median fish body burden predicted by the
FISHRAND model is compared with measured median fish body burden as reported by the various
agencies. The error bars about each median represent the 95 percent confidence interval on the
median. The error bars were calculated assuming the underlying distribution to be lognormal using
the formulation given in Gilbert (1987). (Note that FISHRAND is a mechanistic model which also
incorporates probability distributions for the various parameters. The model result is a probability
distribution from which the mean, median or other statistical properties can be obtained.)
In general, the agreement between the modeled and sampled data is better on the wet weight
basis than on the lipid normalized basis. For the wet weight data, the model results fall close to the
median of the sampled data, in some cases mirroring the trend in the sample data. Nonetheless, the
data show substantive year-to-year variations which are not reflected in the model output.
Additionally, the model appears more accurate at RM 113 than at RM 152, falling within the
confidence limits for nearly all years of measurement for the three species shown at RM 113. At both
locations the model results reflect the general trend to lower PCB concentrations with time. On
average, the model values tend to fall below the mean value for each species, location and year.
The difference between the measured and predicted values can be expressed as a relative
percent difference (RPD). The RPD is calculated as follows:
RPD = (Model Median Estimate - Median Measurement)
Median Measurement
Table 3-4 summarizes the RPDs calculated from the FISHRAND results and the 1987 to
1996 NYSDEC, USEPA and NOAA data. The RPDs are calculated using the wet weight median
values from the model and the corresponding measurements. As was evident from the figures, the
FISHRAND results tend to fall below the measurement medians, yielding negative RPDs. However,
the measurements vary considerably so that both positive and negative deviations are obtained.
Averaging by species and river mile, the mean RPD + 2 standard errors rarely excludes zero,
indicating a lack of statistical significance for the calculated differences. The mean RPD for the
period 1986-1997 is -6 percent for all fish. For the potential game fish (largemouth bass, brown
bullhead, white perch and yellow perch), the mean RPD for the latter years (1993-1997) throughout
the Lower Hudson is -16 percent. Thus, while the model results tend to fall below the data (i.e.,
model concentrations are less than measured concentrations), the difference tends to be within the
uncertainty bounds of the measurements.
Figure 3-12c shows a comparison between model and measured fish body burdens for
pumpkinseed. Here again, the model differs from the measurements for individual years but is able
to reflect the overall trend. RPDs from these results are also included in Table 3-4. Pumpkinseed
represent an intermediate trophic level in the food web and indicate that the model is relatively
accurate at this level as well.
In 1993, USEPA in conjunction with NYSDEC and NOAA, collected and measured PCB
concentrations in the spottail shiner in the Lower Hudson. These data exist only for the one year and
are presented against the model results in Figure 3-12d. For this comparison, FISHRAND results
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were available for four locations and are summarized in the lower half of Table 3-4. These results
again indicate that the model estimates are low with a mean RPD of -27 percent. It is important to
note here, however, that the model appears to capture the spatial trend of the measurement values,
that is, a gradual trend to lower PCB concentrations in fish with decreasing river mile.
The agreement between the FISHRAND results and the measurements is considered
sufficiently good to support the use of FISHRAND in estimating fish body burdens in the Lower
Hudson using the model output from the Farley et al. (1999) model. Although the agreement is not
exact for each location examined with FISHRAND, the overall trends of food web region 1 appear
to be captured, just as they were in the original model by Farley et al. (1999). On average, the
FISHRAND model results tend to underpredict the measurements (by 16 percent in the most recent
period), but are probably within measurement error. Additionally, model agreement is better at some
locations than others but the differences appear to offset each other.
3.1.2 Model Results
The forecast results for the Farley fate and transport and bioaccumulation models and the
FISHRAND model are presented for parameters which are used in ERA Addendum. Relevant
examples of the model output are shown. This is appropriate because Section 3.1 serves as an
explanation of the use of the models and not a report on the models themselves. Complete
descriptions of the models are available in Farley et al. (1999) for the Farley model and USEPA
(1999b and 2000) for the FISHRAND model. The Federal Dam flux is presented on each figure to
show the effect of this parameter.
3.1.2.1 Farley Model Forecast Water Column and Sediment Concentrations
The averaged dissolved phase water column data for food web regions 1 and 2 are presented
in Figure 3-13 for Tri+ PCBs. Food web region 1 paniculate phase water column data for Tri+ PCBs
and whole water data for total PCBs are shown in Figure 3-14. Sediment data from 0-2.5 cm model
segments in the middle of the food web regions are plotted in Figure 3-15. Each of these diagrams
shows the gradual decline of PCB concentrations in the region and their correspondence to the
upstream loads. Additionally, the diagrams show that PCB levels appear to approach an asymptotic
value, suggesting a long-term residual level of contamination in the system, presumably resulting
from the continued upstream loads and the reworking of the existing sediment inventory.
3.1.2.2 Farley Model Forecast Fish Body Burdens
Modeled fish body burdens are plotted in Figures 3-16 and 3-17 for white perch and striped
bass. The flux of Tri+ PCBs over the Federal Dam is also presented in these figures to show the
correlation of this input with the fish body burden. Again, similar to the sediments and water, the fish
results suggest a long-term residual level of PCBs.
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3.1.2.3 FISHRAND Forecast Fish Body Burdens
The fish body burden forecasts for each receptor modeled using FISHRAND are shown in
Figures 3-18 through 3-23. Modeled receptors are the largemouth bass, white perch, yellow perch,
brown bullhead, pumpkinseed and spottail shiner. In these diagrams the mean PCB concentrations
at each RM are shown with the 95% upper confidence level on the mean. These mean values were
obtained based on the FISHRAND-predicted body burden distributions. The upper confidence level
is calculated from these distributions as well, assuming a lognormal distribution and applying the
calculation method given in Gilbert (1987). These confidence limits are based solely on the model
output distributions. It is likely that these are underestimates of the true confidence limits given that
the model is unable to capture the year-to-year variability evident in the data. Nonetheless, the model
is expected to accurately represent the long-term behavior of the mean, as shown by the agreement
between the model output and measurement medians presented previously.
3.1.3 Modeling Summary
This section describes the application of the model developed by Farley et al. (1999) to create
a 70-year forecast for the Lower Hudson. For use in the ERA Addendum and Mid-Hudson HHRA,
the Farley model was extensively supplemented by the USEPA models developed for the Upper
Hudson, namely HUDTOX and FISHRAND. HUDTOX provides a reasonable basis for estimating
future Upper Hudson loads to the lower river while FISHRAND provides estimates of PCB levels
in fish species based on Farley et al. (1999) model output. Supplementing the Farley model in this
manner provided acceptable agreement with the existing calibration data, particularly for fish and
sediments. In general, fish body burdens estimated by the models tended to fall below the
measurements by perhaps 16 percent. The model results were able to capture the general trend of
decreasing PCB concentration with time and distance down river, but not the year-to-year variability.
The agreement is considered sufficient for use in the ERA Addendum and Mid-Hudson HHRA.
3.2 Exposure Point Concentrations
Models have been developed to describe the fate, transport, and bioaccumulation potential
of PCBs in the Upper Hudson River. The Farley et al. (1999) model provides sediment and water
PCB concentrations and the FISHRAND model provides benthic invertebrate, water column
invertebrate, macrophyte, and fish PCB concentrations (USEPA, 1999b). FISHRAND predicts
probability distributions of expected concentrations of PCBs in fish based on mechanistic mass-
balance principles and an understanding of the underlying biology.
FISHRAND is a mechanistic, fully time-varying model based on the Gobas (1993) modeling
approach. The model relies on solutions of differential equations to describe the uptake of PCBs
over time, and incorporates both sediment and water sources to predict the uptake of PCBs based
on prey consumption and food web dynamics. The model provides expected fish species
concentrations of PCBs in the form of distributions. These distributions can be interpreted as
population-level concentrations; that is, at the 95th percentile, 95% of the population is expected to
experience the predicted concentration or less.
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Concentrations of PCBs in the Lower Hudson River ecosystem were estimated for the period
1993 to 2018 for the four reaches comprising the lower river. These reaches are:
• River Mile (RM) 152 - encompassing RM 153.5 - 123.5;
• RM 113 - encompassing RM 123.5 - 93.5;
• RM 90 - encompassing RM 93.5 - 63.5; and
• RM 50 - encompassing RM 63.5 - 33.5.
3.2.1 Modeled Water Concentrations
The Farley model (Farley et al. 1999) was used to predict whole water and dissolved water
concentrations of PCBs for four regions of the Lower Hudson River for the period of 1993 to 2018.
Table 3-4 provides the predicted average and 95% UCL whole water concentrations on a Tri+ total
PCB basis.
Table 3-5 also provides the predicted average and 95% UCL whole water concentrations
expressed on a TEQ basis. These values were obtained by multiplying the Tri+ predictions in Table
3-5 by the toxic equivalency weighting factors developed to describe the proportion of the Tri+ total
expressed as a TEQ (see USEPA, 1999c for details).
3.2.2 Modeled Sediment Concentrations
The Farley et al. (1999) model was also used to predict concentrations of PCBs in sediments
for the period 1993 to 2018. Table 3-6 provides the predicted average and 95% UCL sediment
concentrations on a Tri+ total PCB basis.
Table 3-7 provides total organic carbon (TOC) normalized predicted average and 95% UCL
sediment concentrations. To estimate the TOC-normalized sediment concentrations the predicted
dry weight was divided by the percent TOC, which was assumed to be 2.5% for the entire lower river
(Farley et al., 1999). TOC-normalized sediment concentrations are used for comparison to
guidelines based on organic carbon normalization (i.e., NYSDEC, 1999a and Persaud et al., 1993).
These tables also provide the predicted average and 95% UCL sediment concentrations
expressed on a TEQ basis. These values were obtained by multiplying the Tri+ predictions by the
toxic equivalency weighting factors developed to describe the proportion of the Tri+ total expressed
as a TEQ.
3.2.3 Modeled Benthic Invertebrate Concentrations
Benthic invertebrate concentrations of PCBs for the period 1993 to 2018 were predicted
using the biota sediment accumulation factor (BSAF) developed for the baseline ERA (USEPA,
1999c). Table 3-8 provides the predicted average and 95% UCL benthic invertebrate concentrations
expressed on a total PCB (Tri+) and a TEQ basis. The TEQ values were obtained by multiplying the
predicted benthic invertebrate concentration by the TEF for that receptor species based on the
analyses presented in subchapter 3.2 of the ERA (USEPA, 1999c).
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3.2.4 Modeled Fish Concentrations
Concentrations of PCBs in spottail shiner, pumpkinseed, yellow perch, white perch, brown
bullhead, and largemouth bass for the period 1993 to 2018 were predicted using the FISHRAND
model (USEPA, 1999b).
Striped bass PCB concentrations were predicted via a ratio to largemouth bass from
FISHRAND using the Farley model, as discussed in section 3.1.1.2. The average ratio between
measured striped bass and largemouth bass at RM 152 is 2.5 (standard deviation = 1.6) and 0.52
(standard deviation = 0.2) at RM 113. Striped bass concentrations were not calculated for the lower
regions because striped bass results for this region were already themselves averaged in the Farley
model, and would have to be re-averaged to generate results (i.e., taking the log of the already
averaged age classes is not the same as taking the log of the original values and then taking the
average). Using ratios to calculate the striped bass concentrations allows the population level risk,
rather than the average risk, to be estimated.
Tables 3-9 through 3-15 provide the 25th and 95th percentile values as well as the median of
the predicted distribution for the spottail shiner, pumpkinseed, yellow perch, white perch, brown
bullhead, largemouth bass, and striped bass, respectively, expressed on a wet weight basis for Tri+
total PCBs.
Forecasts are not provided for the shortnose sturgeon, because a specific bioaccumulation
model has not been developed for this species. For this analysis, brown bullhead results serve as
an order-of-magnitude surrogate fish species to assess potential risks to shortnose sturgeon.
The observed fish PCB concentrations for all species except pumpkinseed and spottail shiner
in both the USEPA Phase 2 and NYSDEC sampling programs are given as standard fillets. Because
ecological receptors do not distinguish between standard fillets and whole fish, and TRVs for fish
are typically based on whole body wet weight concentrations, the observed wet weight
concentrations require an adjustment to reflect the difference between the standard fillet and the
whole body. As PCBs are known to partition into lipid, the conversion was accomplished by
evaluating whole body versus standard fillet lipid content to obtain a multiplier for those species for
which data were available (USEPA, 1997c). For largemouth bass, this conversion factor is 2.5 and
for brown bullhead, the conversion factor is 1.5. These values were discussed with NYSDEC and
thought to be comparable to values for Hudson River fish (NYSDEC, 1999c). For those fish species
for which the ratio of lipid in the whole fish relative to the standard fillet could not be obtained (i.e.,
white perch and yellow perch), the observed and modeled body burdens expressed on a fillet basis
were used and the calculated concentrations are likely to be underpredicted. Note that this is likely
to underestimate wet weight concentrations in the whole body but has no effect on lipid-normalized
concentrations. No conversion factors were required for the pumpkinseed and spottail shiner
because they were modeled on a whole body basis.
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3.3 Identification of Exposure Pathways
Potential PCB exposure pathways for aquatic and terrestrial receptors were identified in
the baseline ERA (USEPA, 1999c), where the exposure equations can be found. The exposure
pathways included in the quantitative exposure calculations in this assessment are:
• Benthic invertebrate exposure pathways (as prey of fish and wildlife receptors);
• Fish exposure pathways;
• Avian exposure pathways; and
• Mammalian exposure pathways.
3.3.1 Benthic Invertebrate Exposure Pathways
Benthic invertebrates accumulate PCBs from water, including sediment porewater and the
overlying water, from ingestion of sediment particles, or from ingestion of particulate matter
(phytoplankton and detrital material) in the overlying water at the sediment/water interface.
Predicted benthic invertebrate concentrations for 1993 to 2018 were estimated by multiplying
the predicted sediment concentrations (from the Farley et a/., 1999 model) by a biota-sediment
concentration factor, as described in the baseline ERA (USEPA, 1999c). These benthic invertebrate
concentrations were used as prey concentrations for fish and wildlife receptors.
3.3.2 Fish Exposure Pathways
Fish are directly exposed to PCBs in water and sediments as well as indirectly through the
food chain. Fish exposure to PCBs is described by a wet weight PCB tissue concentration.
Concentrations of PCBs in spottail shiner, pumpkinseed, yellow perch, white perch, brown bullhead,
and largemouth bass were predicted using the FISHRAND model, while striped bass PCB
concentrations were predicted via a ratio to largemouth bass from FISHRAND using the Farley et
ai, 1999 model as updated (Cooney, 1999).
3.3.3 Avian Exposure Pathways, Parameters, Daily Doses, and Egg Concentrations
Avian receptors along the Hudson River are exposed to PCBs primarily through ingestion
of contaminated prey (i.e., diet), surface water ingestion, and incidental ingestion of sediments (see
USEPA, 1999c section 2.3.4). Intake is calculated as an average daily dosage (ADD) value,
expressed as mg PCB/kg/day. The ADD from each of these three calculated exposure pathways is
summed to develop the total ADD of PCBs from riverine sources. Exposure parameters for the tree
swallow, mallard, belted kingfisher, great blue heron, and bald eagle are provided in Tables 3-16 to
3-20. The equations used to calculate intakes for each of the average daily doses are provided in
Chapter 3 of the baseline ERA (USEPA, 1999c). All concentrations of PCBs in fish prey consumed
by avian receptors were calculated using the FISHRAND model (USEPA, 2000).
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3.3.3.1 Summary of ADDExpected, ADD95%UCL, and Egg Concentrations for Avian Receptors
Tree Swallow
Tables 3-25 and 3-26 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female tree swallow from water and dietary sources for the modeling period 1993 -
2018. Doses are based on the results from the Farley et al. (1999) model for water and FISHRAND
(USEPA, 2000) for benthic invertebrates. Tables 3-35 and 3-36 present the expected ADD and 95%
UCL daily dose on a TEQ PCB basis for the modeling period 1993 - 2018 using the same models.
All tables also show the predicted egg concentrations using biomagnification factors based on the
USFWS tree swallow data (2 for total PCBs and 7 on a TEQ basis).
Mallard Duck
Tables 3-27 and 3-28 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female mallard from water, sediment, and dietary sources for the modeling period 1993
- 2018. Doses are based on the results from the Farley et al. (1999) model for water and sediment
and FISHRAND (USEPA, 2000) for benthic invertebrates and macrophytes. Tables 3-37 and 3-38
present the expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling period
1993 - 2018 using the same models. All tables show the predicted egg concentrations using
biomagnification factors based on the USFWS mallard and wood duck data (3 for total PCBs and
28 on a TEQ basis).
Belted Kingfisher
Tables 3-29 and 3-30 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female belted kingfisher from water, sediment, and dietary sources for the modeling
period 1993 - 2018. Doses are based on the results from the Farley et al. (1999) model for water and
sediment and FISHRAND (USEPA, 2000) for benthic invertebrates and forage fish. Tables 3-39
and 3-40 present the expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling
period 1993 - 2018 using the same models. All tables also show the predicted egg concentrations
using biomagnification factors obtained from Giesy et al. (1995) for piscivorous birds (28 for total
PCBs and 19 on a TEQ basis).
Great Blue Heron
Tables 3-31 and 3-32 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female great blue heron from water, sediment, and dietary sources for the modeling
period 1993 - 2018. Doses are based on the results from the Farley et al. (1999) model for water and
sediment and FISHRAND for benthic invertebrates and forage fish. Tables 3-41 and 3-42 present
the expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling period 1993 -
2018 using the same models. All tables also show the predicted egg concentrations using
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biomagnification factors obtained from Giesy et al. (1995) for piscivorous birds (28 for total PCBs
and 19 on a TEQ basis).
Bald Eagle
Tables 3-33 and 3-34 present the expected ADD and 95% UCL dally dose on a total PCB
basis for the female bald eagle from water, sediment, and dietary sources for the modeling period
1993 - 2018. Doses are based on the results from the Farley et al. (1999) model for water and
sediment and FISHRAND (USEPA, 2000) for piscivorous fish. Tables 3-43 and 3-44 present the
expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling period 1993-2018
using the same models. All tables also show the predicted egg concentrations using
biomagnification factors obtained from Giesy et al. (1995) for piscivorous birds (28 for total PCBs
and 19 on a TEQ basis).
3.3.4 Mammalian Exposure Pathways, Parameters, and Daily Doses
Terrestrial mammals living along the Hudson River are exposed to PCBs primarily via
ingestion of contaminated prey (i.e., diet), surface water ingestion, and incidental ingestion of
sediments (see baseline ERA section 2.3.4). Intake is calculated as an ADD value expressed as mg
PCB/kg/day. The ADDs from each of the three calculated exposure pathways are summed to develop
the total ADD of PCBs from riverine sources. The equations and parameters used to calculate intakes
for each of the ADDs are provided in Chapter 3 of the baseline ERA (USEPA, 1999c). Exposure
parameters for the little brown bat, raccoon, mink, and river otter are provided in Tables 3-21 to 3-
24. The equations used to calculate intakes for each of the ADD are provided in the baseline ERA
(USEPA, 1999c). All concentrations of PCBs in fish prey consumed by mammalian receptors were
calculated using the FISHRAND model (USEPA, 2000).
3.3.4.1 Summary of ADDEjlpected and ADD95%UCL for Mammalian Receptors
Little Brown Bat
Tables 3-45 and 3-46 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female little brown bat from water and dietary sources for the modeling period 1993 -
2018. Doses are based on the results from the Farley et al. (1999) model for water and FISHRAND
(USEPA, 2000) for benthic invertebrates. Tables 3-53 and 3-54 present the expected ADD and 95%
UCL daily dose on a TEQ PCB basis for the modeling period 1993 - 2018 using the same models.
Raccoon
Tables 3-47 and 3-48 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female raccoon from water, sediment, and dietary sources for the modeling period 1993
- 2018. Doses are based on the results from the Farley et al. (1999) model for water and sediment
and FISHRAND (USEPA, 2000) for benthic invertebrates and forage fish. Tables 3-55 and 3-56
present the expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling period
1993-2018 using the same models.
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Mink
Tables 3-49 and 3-50 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female mink from water, sediment, and dietary sources for the modeling period 1993
- 2018. Doses are based on the results from the Farley et al. (1999) model for water and sediment
and FISHRAND (USEPA, 2000) for benthic invertebrates and forage fish. Tables 3-57 and 3-58
present the expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling period
1993 - 2018 using the same models.
River Otter
Tables 3-51 and 3-52 present the expected ADD and 95% UCL daily dose on a total PCB
basis for the female river otter from water, sediment, and dietary sources for the modeling period
1993 - 2018. Doses are based on the results from the Farley et al. (1999) model for water and
sediment and FISHRAND (USEPA, 2000) for forage fish and piscivorous fish. Tables 3-59 and 3-
60 present the expected ADD and 95% UCL daily dose on a TEQ PCB basis for the modeling period
1993 - 2018 using the same models.
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Chapter 4
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4.0 EFFECTS ASSESSMENT
This chapter provides a general overview of the toxicology of PCBs and provides a brief
overview of the methods used to characterize particular lexicological effects of PCBs on aquatic and
terrestrial organisms. Full details are provided in Appendix B. Toxicity reference values (TRVs)
selected to estimate the potential risk to receptor species resulting from exposure to PCBs are
presented following the background on PCB toxicology. TRVs are levels of exposure associated with
either Lowest Observed Adverse Effects Levels (LOAELs) or No Observed Adverse Effects Levels
(NOAELs). They provide a basis for judging the potential effects of measured or predicted
exposures that are above or below these levels.
Use of both LOAELs and NOAELS provides perspective on the potential for risk as a result
of exposure to PCBs originating from the site. LOAELs are values at which effects have been
observed (in either laboratory or field studies), while the NOAEL represents the lowest dose or body
burden at which an effect was not observed. Exceedance of a LOAEL indicates a greater potential
for risk.
4.1 Selection of Measures of Effects
Many studies examined the effects of PCBs on aquatic and terrestrial organisms, and results
of these studies are compiled and summarized in several reports and reviews (e.g., Eisler and Belisle,
1996; Niimi, 1996; Hoffman et ai, 1998; ATSDR, 1996; Eisler, 1986; and NOAA, 1999b). For the
present assessment, studies on the toxic effects of PCBs were identified by searching the National
Library of Medicine (NLM) MEDLINE and TOXLINE databases. Other studies were identified
from the reference section of papers that were identified by electronic search. Papers were reviewed
to determine whether the study was relevant to the topic.
Many different approaches and methodologies are used in these studies, some of which are
more relevant than others to the selection of TRVs for the ERA (USEPA, 1999c) and this ERA
Addendum. TRVs are levels of exposure associated with either LOAELs or NOAELs. They provide
a basis for judging the potential effects of measured or predicted exposures that are above or below
these levels. Some studies express exposures as concentrations or doses of total PCBs, whereas other
studies examine effects associated with individual congeners (e.g., PCB 126) or as total dioxin
equivalents (TEQs). This risk assessment develops separate TRVs for total PCBs and TEQs. This
chapter briefly describes the rationale that was used to select TRVs for various ecological receptors
of concern.
Some studies examine toxicity endpoints (such as lethality, growth, and reproduction) that
are thought to have greater potential for adverse effects on populations of organisms than other
studies. Other studies examine toxicity endpoints such as behavior, disease, cell structure,
immunological responses, or biochemical changes that affect individual organisms, but may not
result in adverse effects at the population level. For example, toxic effects such as enzyme induction
may or may not result in adverse effects to individual animals or populations. For the ERA and
ERA Addendum, TRVs were selected from studies that examine the effects of PCBs on lethality,
growth or reproduction. Studies that examined the effects of PCBs on other sublethal endpoints are
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not used to select TRVs, although effects may occur at these concentrations. Lethality, growth, and
reproductive-based endpoints typically present the greatest risk to the viability of the individual
organism and therefore survival of the population. Thus, these are considered to be the measurement
endpoints of greatest concern relative to the stated assessment endpoints.
When exposures are expected to be long-term, data from studies of chronic exposure are
preferable to data from medium-term (subchronic), short-term (acute), or single-exposure studies
(USEPA, 1997b). Because of the persistence of PCBs, exposure of ecological receptors to PCBs
from the Hudson River is expected to be long-term, and therefore studies of chronic exposure are
preferentially used to select the TRVs. Long-term studies are also preferred since reproductive
effects of PCBs are typically studied and evaluated following long-term exposure.
Dose-response studies compare the response of organisms exposed to a range of doses to that
of a control group. Ideally, doses that are below and above the threshold level that causes adverse
effects are examined. Toxicity endpoints determined in dose-response and other studies include:
• NOAEL (No-Observed-Adverse-Effect-Level) is the highest exposure level shown to
be without adverse effect in organisms exposed to a range of doses. NOAELs may be
expressed as dietary doses (e.g., mg PCBs consumed/kg body weight/day), as
concentrations in external media (e.g., mg PCBs/kg food), or as concentrations in tissue
of the affected organisms (e.g., mg chemical/kg egg).
• LOAEL (Lowest-Observed-Adverse-Effect-Level) is the lowest exposure level shown
to produce adverse effect in organisms exposed to a range of doses. LOAELs may also
be expressed as dietary doses (e.g., mg PCBs consumed/kg body weight/day), as
concentrations in external media (e.g., mg PCBs/kg food), or as concentrations in tissue
of the effected organisms (e.g., mg chemical/kg egg). The LOAEL represents a
concentration at which the particular effect has been observed and the occurrence of the
effect is statistically significantly different from the control organisms.
• LD50 is the Lethal Dose that results in death of 50% of the exposed organisms. The LD50
is expressed in units of dose (e.g., mg PCBs administered/kg body weight of test
organism/day).
• LC50 is the Lethal Concentration in some external media (e.g. food, water, or sediment)
that results in death of 50% of the exposed organisms. The LC50 is expressed in units
of concentration (e.g., mg PCBs/kg wet weight food).
• ED50 is the Effective Dose that results in a sublethal effect in 50% of the exposed
organisms (mg/kg/day).
• EC50 is the Effective Concentration in some external media that results in a sublethal
effect in 50% of the exposed organisms (mg/kg).
• CBR or Critical Body Residue is the concentration in the organism (e.g., whole body,
liver, or egg) that is associated with an adverse effect (mg PCBs/kg wet weight tissue).
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• EL-effect is the effect level that results in an adverse effect in organisms exposed to a
single dose, rather than a range of doses. Expressed in units of dose (mg/kg/day) or
concentration (mg/kg).
• EL-no effect is the effect level that does not result in an adverse effect in organisms
exposed to a single dose, rather than a range of doses. Expressed in units of dose
(mg/kg/day) or concentration (mg/kg).
Most USEPA risk assessments typically estimate risk by comparing the exposure of receptors
of concern to TRVs that are based on NOAELs. TRVs for the ERA (USEPA, 1999c) and ERA
Addendum were developed on the basis of both NOAELs and LOAELs to provide perspective on
the range of potential effects relative to measured or modeled PCB exposures. Because the LOAEL
represents a concentration at which effects were definitely observed, this is a stronger indicator of
the potential for risk. However, risk may occur at any concentration between the NOAEL and the
LOAEL, so exceedance of the NOAEL also indicates the potential for risk.
Differences in the feeding behavior of aquatic and terrestrial organisms determine the type
of toxicity endpoints that are most easily measured and most useful in assessing risk. For example,
the dose consumed in food is more easily measured for terrestrial animals than for aquatic organisms
because uneaten food can be difficult to collect and quantify in an aqueous environment. Therefore,
for aquatic organisms, toxicity endpoints are more often expressed as concentrations in external
media (e.g., water) or as accumulated concentrations in the tissue of the exposed organism (also
called a "body burden"). In some studies, doses are administered via gavage, intraperitoneal
injection into an adult, or injection into a fish or bird egg. If appropriate studies are available, TRVs
were selected on the basis of the most likely route of exposure, as described below:
• TRVs for fish are expressed as critical body residues (CBR) (e.g., mg/kg whole body
weight and mg/kg lipid in eggs).
• TRVs for terrestrial receptors (e.g., birds and mammals) are expressed as daily dietary
doses (e.g., mg/kg whole body weight/day).
• TRVs for birds are also expressed as concentrations in eggs (e.g. mg/kg wet weight
egg).
4.1.1 Methodology Used to Derive TRVs
The literature on toxic effects of PCBs to animals includes studies conducted solely in the
laboratory, as well as studies including a field component. Each type of study has advantages and
disadvantages for the purpose of deriving TRVs for a risk assessment. For example, a controlled
laboratory study can be designed to test the effect of a single formulation or congener (e.g. Aroclor
1254 or PCB 126) on the test species in the absence of the effects of other co-occurring
contaminants. This is an advantage because greater confidence can be placed in the conclusion that
observed effects are related to exposure to the test compound. However, laboratory studies are often
conducted on species that are easily maintained in the laboratory, rather than on wildlife species.
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Therefore, laboratory studies may have the disadvantage of being conducted on species that are less
closely related to a particular receptor of concern. Field studies have the advantage that organisms
are exposed to a more realistic mixture of PCB congeners (with differences in toxic potencies), than,
for example, laboratory tests that expose organisms to a commercial mixture, such as Aroclor 1254.
Field studies have the disadvantage that organisms are usually exposed to other contaminants and
observed effects may not be attributable solely to exposure to PCBs. Field studies can be used most
successfully, however, to establish concentrations of PCBs or TEQs at which adverse effects are not
observed (e.g., a NOAEL). Because of the potential contribution of other contaminants (e.g. metals,
pesticides, etc.) to observed effects in field studies, the ERA and ERA Addendum use field studies
to establish NOAEL TRVs, but not LOAEL TRVs.
If appropriate field studies are available for species in the same taxonomic family as the
receptor of concern, those field studies were used to derive NOAEL TRVs for receptors of concern.
Appropriateness of a field study was based on the following considerations:
• whether the study examines sensitive endpoints, such as reproductive effects, in a
species that is closely related (e.g. within the same taxonomic family) to the receptor
of concern;
• whether measured exposure concentrations of PCBs or dioxin-like compounds are
reported for dietary doses, whole organisms, or eggs;
• whether the study establishes a dose-response relationship between exposure
concentrations of PCBs or dioxin-like contaminants and observed effects; and
• whether contributions of co-occurring contaminants are reported and considered to be
negligible in comparison to contributions of PCBs or dioxin-like compounds.
If appropriate field studies are not available for a test species in the same taxonomic family
as the receptor species of concern, laboratory studies were used to establish TRVs for the receptor
species. The general methodology described in the following paragraphs was used to derive TRVs
for receptors of concern from appropriate studies.
When appropriate chronic-exposure toxicity studies on the effects of PCBs on lethality,
growth, or reproduction are not available for a species of concern, extrapolations from other studies
were made in order to estimate appropriate TRVs. For example, if toxicity data are unavailable for
a particular species of bird, toxicity data for a related species of bird were used if appropriate
information was available. Several methodologies have been developed for deriving TRVs for
wildlife species (e.g., Sample etal., 1996; California EPA, 1996; USEPA, 1996; and Menzie-Cura
& Associates, 1997). The general methodology used to develop LOAEL and NOAEL TRVs is
described below:
• If an appropriate NOAEL is unavailable for a phylogenetically similar species (e.g.
within the same taxonomic family), NOAEL values for other species (as closely related
as possible) were adjusted by dividing by an uncertainty factor of 10 to account for
extrapolations between species. The lowest appropriate NOAEL was used whenever
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several studies are available. However, if the surrogate test species is known to be the
most sensitive of all species tested in that taxonomic group (e.g. fish, birds, mammals),
then an interspecies uncertainty factor was not applied
• In the absence of an appropriate NOAEL, if a LOAEL is available for a phylogenetically
similar species, these may be divided by an uncertainty factor of 10 to account for a
LOAEL to NOAEL conversion. The LOAEL to NOAEL conversion is similar to
USEPA's derivation of human health RfD (Reference Dose) values, where LOAEL
studies are adjusted by a factor of 10 to estimate NOAEL values.
• When calculating chronic dietary dose-based TRVs (e.g. mg/kg/day) from data for sub-
chronic tests, the sub-chronic LOAEL or NOAEL values were divided by an additional
uncertainty factor of 10 to estimate chronic TRVs. The use of an uncertainty factor of
10 is consistent with the methodology used to derive human health RfDs. These factors
are applied to account for uncertainty in using an external dose (e.g., mg/kg/day in diet)
as a surrogate for the dose at the site of toxic action (e.g. mg/kg in tissue). Because
organisms may attain a toxic dose at the site of toxic action (e.g. in tissues or organs)
via a large dose administered over a short period, or via a smaller dose administered
over a longer period, uncertainty factors are used to estimate the smallest dose that, if
administered chronically, would result in a toxic dose at the site of action. USEPA has
not established a definitive line between sub-chronic and chronic exposures for
ecological receptors. The ERA and ERA Addendum follow recently developed guidance
(Sample et al., 1996) which considers 10 weeks to be the minimum time for chronic
exposure of birds and 1 year for chronic exposure of mammals.
• For studies that actually measure the internal toxic dose (e.g., mg PCBs/kg tissue), no
sub-chronic to chronic uncertainty factor was applied. This is appropriate because
effects are being compared to measured internal doses, rather than to external dietary
doses that are used as surrogates for the internal dose.
• In cases where NOAELs are available as a dietary concentration (e.g., mg contaminant
per kg food), a daily dose for birds or mammals was calculated on the basis of standard
estimates of food intake rates and body weights (e.g., USEPA, 1993b).
Professional judgment is used to determine relevant endpoints for selecting TRVs. For
example, hatching time in fish is considered less relevant than hatchability, which directly affects
the viability of offspring. The implication of hatching time on the viability of the population is less
clear than an effect such as hatchability. Specific endpoints relative to TRVs are provided in
Appendix B.
The sensitivity of the risk estimates to the use of uncertainty factors and the selected TRVs
will be examined in the uncertainty chapter (Chapter 6.0).
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4.1.2 Selection of TRVs
TRVs selected for Hudson River receptors are provided in Tables 4-1 to 4-3 for fish, birds,
and mammals, respectively. These tables provide both Total PCB (Tri+) TRVs and TEQ-based
TRVs (discussed below). A complete description of the selection process for each receptor can be
found in Appendix B.
As described in the baseline ERA (USEPA, 1999c), the Toxic Equivalency (TEQ)/Toxic
Equivalency Factors (TEF) methodology (TEQ/TEF), quantifies the toxicities of PCB congeners
relative to the toxicity of the potent dioxin 2,3,7,8-TCDD (see van den Berg et al., 1998 for review).
It is currently accepted that the carcinogenic potency of dioxin is affected by its ability to bind AhR
and dioxin is considered to be the most potent known AhR ligand. It is also generally accepted that
the dioxin-like toxicities of PCB congeners are directly correlated to their ability to bind the AhR.
Thus, the TEQ/TEF methodology provides a toxicity measurement for all AhR-binding compounds
based on their relative toxicity to dioxin. Since 2,3,7,8-TCDD has the greatest affinity for the AhR,
it is assigned a TCDD-Toxicity Equivalent Factor of 1.0. PCB congeners are then assigned a TCDD-
TEF relative to 2,3,7,8-TCDD, based on experimental evidence. For example, if the relative toxicity
of a particular congener is one-thousandth that of TCDD, it would have a TEF of 0.001. The potency
of a PCB congener is estimated by multiplying the tissue concentration of the congener in question
by the TEF for that congener to yield the toxic equivalent (TEQ) of dioxin. A TEQ for the total PCB
concentration can be determined from the sum of the calculated TEQs for each AhR-binding
congener. The World Health Organization (WHO) has derived TEFs for a number of PCB congeners
(van den Berg et al., 1998). These values, which are used in this assessment, are presented in Table
4-4.
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Chapter 5
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5.0 RISK CHARACTERIZATION
Risk characterization is made up of two steps, risk estimation and risk description (USEPA,
1992a and 19975). Risk estimation integrates stressor-response profiles (Chapter 4) with exposure
profiles (Chapter 3) to provide an estimate of risk (Chapter 5) and related uncertainties (Chapter 6).
The assessment endpoints and their associated measurement endpoints, selected during problem
formulation (Chapter 2), are evaluated in this section.
In the toxicity quotient (TQ) approach, potential risks to ecological receptors are assessed by
comparing measured or modeled concentrations (Chapter 3) to toxicity benchmarks developed in
(Chapter 4). Future PCB concentrations are predicted on total PCBs (Tri+) and TEQ bases.
The TQ is the direct numerical comparison of a measured or modeled exposure concentration
or dose to a benchmark dose or concentration. It is calculated as:
Toxicity Quotient = Modeled Dose or Concentration
Benchmark Dose or Concentration
TQs equal to or exceeding one are typically considered to indicate potential risk to ecological
receptors. The TQ method provides insight into the potential for general effects upon individual
animals in the local population resulting from exposure to PCBs. If effects are judged not to occur
at the aveiage individual level, they are probably insignificant at the population level. However, if
risks are present at the individual level they may or may not be important at the population level.
The risk characterization in the Hudson River is based on the following assessment endpoints:
• Benthic community structure as a food source for local fish and wildlife (Section 5.1)
• Health and maintenance of local fish populations (Section 5.2) by evaluating survival,
growth, and reproduction of:
- local forage fish populations;
- local omnivorous fish populations; and
- local piscivorous/semi-piscivorous fish populations.
• Protection (i.e., survival, growth, and reproduction) of local wildlife including:
- insectivorous birds (Section 5.3);
- waterfowl (Section 5.4);
- semi-piscivorous/piscivorous birds (Section 5.5);
- insectivorous mammals (Section 5.6);
- omnivorous mammals (Section 5.7); and
- semi-piscivorous/piscivorous mammals (Section 5.8)
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• Protection of threatened and endangered species (Section 5.9).
• Protection of significant habitats (Section 5.10).
5.1 Evaluation of Assessment Endpoint: Benthic Community Structure as a
Food Source for Local Fish and Wildlife
5.1.1 Do Modeled PCB Sediment Concentrations Exceed Appropriate Criteria and/or
Guidelines for the Protection of Aquatic Life and Wildlife?
5.1.1.1 Measurement Endpoint: Comparisons of Modeled Sediment Concentrations to
Guidelines For the Protection of Aquatic Life and Wildlife
Table 5-1 presents the ratios of forecast sediment concentrations to various sediment
guidelines. Comparisons are made on total PCB (Tri+) sediment concentrations (i.e., NOAA, 1999a;
Persaud et al., 1993; and Washington State, 1997) and TOC-normalized sediment concentrations
(/. e., N YSDEC, 1999a and Persaud et al., 1993). A summary of sediment concentrations is provided
in Table 3-2 and TOC-normalized sediment concentrations are shown in Table 3-3.
The NOAA (1999a) consensus-based sediment effect concentrations (SECs) for PCBs were
developed to support an assessment to sediment-dwelling organisms living in the Hudson River
Basin. They refer to all of the PCBs found in the Hudson River, plus the degradation products and
metabolites of these chemicals. The Hudson River SECs provide a threshold effect concentration
(TEC) of 0.04 mg/kg, a mid-range effect concentration (MEC) of 0.4 mg/kg, and an extreme effect
concentration (EEC) of 1.7 mg/kg. The TEC is intended to identify the concentration of total PCBs
below which adverse population-level effects (e.g., mortality, decreased growth, reproductive failure)
on sediment-dwelling organisms are unlikely to be observed (NOAA, 1999a). The MEC represents
the concentration of total PCBs above which adverse effects on sediment-dwelling organisms are
expected to be frequently observed. Adverse effects are expected to be usually or always observed
at PCB concentrations exceeding the EEC.
Forecast sediment concentrations based on the Farley et al. (1999) model exceed the NOAA
TEC at all four locations for both average and 95% UCL concentrations throughout the modeling
period (Table 5-1). MEC consensus values are exceeded using 95% UCL concentrations at RMs 152,
113, and 90 throughout the modeling period and at RM 50 until 2006. The average forecast
concentration at RM 152 exceeds the MEC throughout the modeling period and the average
concentrations lower down river exceed the MEC for portions of the modeling period. None of the
forecast concentrations exceed the EEC at any of the locations.
The NYSDEC has developed screening criteria concentrations that can be used to identify
areas of sediment contamination and evaluate the potential risk that the contaminated sediment may
pose to the environment (NYSDEC, 1999a). Criteria developed for the protection of aquatic life
from chronic toxicity and protection of wildlife from toxic effects of bioaccumulation are examined
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in this addendum. Forecast sediment concentrations exceed the NYSDEC benthic aquatic life
chronic toxicity criterion at RMs 152,113, and 90 for the duration of the modeling period based on
the 95% UCL. The benthic aquatic life criterion was exceeded until 2011 at RM 90 and until 1997
at RM 50 (Table 5-1). The average total PCB concentration exceeds the criterion for various portions
of the modeling period at RMs 152, 113, and 90. The freshwater criterion value of 19.3 mg/kg OC
was used, which based on the 2.5% OC assumption used in this assessment provides a dry weight
value of 0.48 mg/kg.
Forecast sediment concentrations exceed the NYSDEC wildlife bioaccumulation criterion
at all four locations for the duration of the modeling period using both average and 95th UCL results
(Table 5-1). The NYSDEC wildlife criterion is 1.4 mg/kg OC, which based on the 2.5% OC
assumption used in this assessment provides a dry weight value of 0.035 mg/kg.
The Ontario sediment quality guidelines for the protection and management of aquatic
sediment quality (Persaud et al., 1993) were developed to protect the aquatic environment by setting
safe levels for metals, nutrients, and organic compounds. The no effect level (NEL) is the level at
PCBs in the sediment that do not affect fish or the sediment-dwelling organism. The lowest effect
level (LEL) indicates a level of contamination that has no effect on the majority of sediment dwelling
organisms. At the severe effect level (SEL) sediments are likely to affect the health of sediment-
dwelling organisms. Forecast sediment concentrations exceeded the total PCB NEL of 0.01 mg/kg
at all locations for both the average and 95% UCL concentration for the duration of the sampling
period (1993-2018) by up to two orders of magnitude (Table 5-1). The total PCB LEL of 0.07 mg/kg
was also exceeded at all locations for both the average and 95% UCL concentration for the duration
of the sampling period. The total PCB SEL of 530 mg/kg OC (equal to a dry weight value of 1.3
mg/kg using 2.5% OC) was not exceeded at any location for the duration of the modeling period.
Washington State has also derived chemical criteria to predict possible biological effects in
sediments (Washington State, 1997). Bioassays for PCBs were conducted using both Microtox®
(endpoint = luminescence reduction) and Hyalella azteca (endpoint = mortality ). The Probable
Apparent Effects Thresholds (PAET) for Microtox® was 0.021 mg/kg (total PCBS), while the PAET
of Hyalella azteca was 0.45 mg/kg. The Microtox® PAET was exceeded at all locations for the
duration of the modeling period (1993-2018) using both average and 95% UCL concentrations
(Table 5-1). The PAET of Hyalella azteca was exceeded by predicted 95% UCL PCB concentrations
at RMs 152 and 113 for the duration of the modeling period and at RMs 90 and 50 for portions of
the modeling period. Using average PCB concentrations the Hyalella azteca PA.ET was exceeded
for a portion of the modeling period at all stations.
Many of the ratios of modeled sediment concentrations to appropriate guidelines exceed 10
or occasionally even 100. Forecast total PCB concentrations are Tri+ values, and do not include
mono or dichlorinated congeners that usually contribute a portion of the total PCB load. Thus, even
in the unlikely event that forecast sediment concentrations were to decrease by an order of magnitude
or more, comparisons to sediment guidelines would still show exceedances.
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5.1.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria and/or
Guidelines for the Protection of Aquatic Life and Wildlife?
5.1.2.1 Measurement Endpoint: Comparison of Modeled Water Column Concentrations of
PCBs to Criteria
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria and guidelines. All forecast water concentrations (i.e., average
and 95% UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 /zg/L and the
USEPA wildlife criterion of 1.2 x 10"4 /^g/L at all four locations throughout the modeling period. The
whole water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity
criterion of 0.014 /j.g/L for a portion of the modeling period for both average and 95% UCL at all
modeling locations. These comparisons are likely to underestimate the true risk, as concentrations
are expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.2 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Local Fish Populations
5.2.1 Do Modeled Total PCB and TEQ-Based PCB Body Burdens in Local Fish Species
Exceed Benchmarks for Adverse Effects on Forage Fish Reproduction?
5.2.1.1 Measurement Endpoint: Comparison of Modeled Total PCB Fish Body Burdens to
Toxicity Reference Values for Forage Fish
Table 5-3 presents the results of the comparison between forecast PCB body burdens in
pumpkinseed and spottail shiner to selected toxicity reference values on a total PCB basis (expressed
as Tri+) under future conditions (1993 - 2018). The total PCB (Tri+) body burden in pumpkinseed
exceeds a TQ of one using a field-based NOAEL at all four modeling locations (i.e., RMs 152, 113,
90, and 50) for the 25th percentile, median, and 95th percentile. On a 95th percentile basis, the
pumpkinseed exceeds one at RM 152 until the end of the modeling period (2018), at RM 133 until
2016, at RM 90 until 2007, and at RM 50 until 2005. This is interpreted to mean that 95% of
individual pumpkinseed fish will experience the shown TQ or less for that year.
The spottail shiner did not exceed a TQ of one at any time or location using the laboratory-
derived NOAEL and LOAEL (Tables 5-4 and 5-5). The TRY derived for the spottail shiner differ
from the TRY derived for the pumpkinseed by more than an order of magnitude (0.5 mg/kg on a
NOAEL basis for the pumpkinseed versus 15 mg/kg on a NOAEL basis for the spottail shiner).
Consequently, spottail shiner TQs are much lower than pumpkinseed.
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5.2.1.2 Measurement Endpoint: Comparison of Modeled PCB TEQs Fish Body Burdens to
Toxicity Reference Values for Forage Fish
Tables 5-6 and 5-7 present the results of the comparison between forecast percentiles of
pumpkinseed to laboratory-derived NOAEL and LOAEL on a TEQ basis under future conditions.
The TRVs for TEQs in fish are mostly based on egg injection studies; however, Hudson River data
are for concentrations in adult fish. These two numbers were not considered to be directly
comparable since lipid concentrations in eggs and adults may differ substantially. The
lipid-normalized egg concentration TRY (e.g., ng TEQs/kg lipid) compared to the lipid-normalized
concentration in adult fish (e.g., ng TEQs/kg lipid) was considered to provide the most appropriate
comparison.
On a NOAEL basis, the TQs exceed one on a 95th percentile basis at RM 152 until
approximately 1999, at RM 113 until 1998, at RM 90 until 1995, and at RM 50 until 1994. On a
LOAEL basis, all TQs fell below one.
Tables 5-8 and 5-9 presents the results for the spottail shiner. TQs for spottail shiners do not
exceed one at any time or location during the modeling period on either a LOAEL or NOAEL basis.
5.2.1.3 Measurement Endpoint: Comparison of Modeled Total PCB Fish Body Burdens to
Toxicity Reference Values for Brown Bullhead
Tables 5-10 and 5-11 present the results of the comparison between predicted percentiles of
brown bullhead concentrations a total PCB basis to laboratory-derived NOAEL and LOAEL under
future conditions (1993-2018). TQs for the brown bullhead exceed one at all locations during the
entire modeling period on NOAEL basis. Using the laboratory-derived LOAEL, the 95th percentile
concentration exceeds one at RMs 152 and 133 throughout the modeling period, at RM 90 until
2017, and at RM 50 until 2007. Because the FISHRAND model predicts standard fillet
concentrations in fish, the wet weight model results were adjusted by a factor of 1.5 for the brown
bullhead, as wildlife feeding on fish consumes them whole. Even without this adjustment, most of
ratios would exceed one on a NOAEL basis.
5.2.1.4 Measurement Endpoint: Comparison of Modeled TEQ Basis Fish Body Burdens to
Toxicity Reference Values for Brown Bullhead
Tables 5-12 and 5-13 present the results of the comparison between forecast percentiles of
brown bullhead concentrations on a TEQ basis to a laboratory-derived NOAEL and LOAEL for
TEQs under future conditions. TQs for the brown bullhead do not exceed one at any time or location
during the modeling period on either a LOAEL or NOAEL basis.
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5.2.1.5 Measurement Endpoint: Comparison of Modeled Total PCB Fish Body Burdens to
Toxicity Reference Values for White and Yellow Perch
Table 5-14 presents the results of the comparison between forecast percentiles of white perch
a total PCB basis to a field-based NOAEL for the period 1993 - 2018. The white perch exceeds a TQ
of one at RM 152 in 1993. The remainder of the ratios fall below one at all locations.
The yellow perch exceeded a TQ of one at all locations during the entire modeling period
using the laboratory-derived NOAEL (Table 5-15). All concentrations (i.e., 25th, median, and 95th
) were exceeded at all locations with the exception of the 25th percentile at RM 50 for 2016-2108.
A TQ of one was not exceeded at any location using the laboratory-derived LOAEL (Table 5-16).
The laboratory-based NOAEL TRY derived for the yellow perch is more than an order of magnitude
lower than the field-based NOAEL TRY derived for the white perch (0.16 mg/kg on a NOAEL basis
for yellow perch versus 3.1 mg/kg on a NOAEL basis for white perch).
Modeled concentrations are based on a standard fillet lipid content. Although an adjustment
is required to estimate whole body tissue concentrations, there was not enough data available to
make this adjustment. Thus, because the presented results are based on forecast standard fillet
concentrations, true risks are likely underestimated for these two species.
5.2.1.6 Measurement Endpoint: Comparison of Modeled TEQ Basis Body Burdens to Toxicity
Reference Values for White and Yellow Perch
Tables 5-17 and 5-18 present the results of the comparison between forecast percentiles of
white perch TEQ-based PCB body burdens to laboratory-derived NOAEL and LOAEL under future
conditions (1993-2018). The white perch exceeds a TQ of one on a TEQ basis at RMs 152, 113, and
90 for the 25th percentile, median, and 95th percentile and at RM 50 for the 95th percentile for a
portion of the modeling period. On a 95th percentile basis, the white perch exceeds one at RMs 152
and RM 133 throughout the modeling period (2018), at RM 90 until 2014, and at RM 50 until 2005.
The median-based TQs exceed one at RM 152 until 2008, at RM 113 until 2003, at RM 90 until
1997, and at RM 50 until 1994. On a LOAEL basis, the 95th percentile exceeds one at RM 152 until
2004, at RM 113 until 1999, and at RM 90 until 1995. All median-based ratios were below one at
RM50.
Results for yellow perch are shown in Tables 5-19 and 5-20. These tables show similar
results to white perch, but yellow perch TQs fall below one a few years before white perch.
Because modeled TEQ concentrations are expressed on a lipid-normalized basis, an
adjustment for standard fillet to whole body is not required for this analysis.
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5.2.1.7 Measurement Endpoint: Comparison of Modeled Tri+ PCB Fish Body Burdens to
Toxicity Reference Values for Largemouth Bass
Table 5-21 presents the results of the comparison between forecast percentiles of largemouth
bass total PCB body burdens to a field-based NOAEL for the period 1993-2018. The largemouth
bass total PCB tissue concentrations exceed the field-based NOAEL for all concentrations (i.e., 25th
percentile, median, and 95th percentile) at all RM s (i.e., 152, 113,90, and 50) for the duration of the
modeling period (1993-2018) with the exceptions of the 25th percentile at RM 90 for 2017 and 2018
and at RM 50 for 2014-2108. As the FISHRAND model predicts standard fillet concentrations in
fish, the wet weight model results were adjusted by a factor of 2.5 for the largemouth bass, because
wildlife feeding on fish consumes them whole. The majority of the ratios would exceed one even
without this adjustment.
5.2.1.8 Measurement Endpoint: Comparison of Modeled TEQ Based Fish Body Burdens to
Toxicity Reference Values for Largemouth Bass
Tables 5-22 and 5-23 present the results of the comparison between modeled largemouth bass
body burdens and laboratory-based NOAEL and LOAEL on a TEQ basis under future conditions
(1993-2018). On a 95th percentile basis, concentrations on a TEQ basis exceed the NOAEL at RM
152 and RM 133 throughout the modeling period (2018), at RM 90 until 2014, and at RM 50 until
2009. Using the LOAEL, the 95th percentile exceed one at RM 152 until about 2005, at RM 133 until
2003, at RM 90 until 1999, and at RM 50 until 1998.
5.2.1.9 Measurement Endpoint: Comparison of Modeled Tri+ PCB Fish Body Burdens to
Toxicity Reference Values for Striped Bass
Table 5-24 presents the results of the comparison between forecast percent) les of striped bass
total PCB body burdens to a field-based NOAEL at RMs 152 and 113 for the period 1993- 2018. At
RM 152, the striped bass Tri+ PCB tissue concentrations exceed the field-based NOAEL on 95th
percentile, median, and 25th percentile bases throughout the entire modeling period (1993-2018). At
RM 113, a ratio of one is exceeded on a 95th percentile basis until 2005, on a median basis until
1999, and on a 25th percentile basis until 1996.
5.2.1.10 Measurement Endpoint: Comparison of Modeled TEQ Based Fish Body Burdens
to Toxicity Reference Values for Striped Bass
Table 5-24 presents the results of the comparisons between forecast percentiles of striped
bass PCB egg concentrations and a TEQ-based laboratory-based NOAEL and LOAEL at RMs 152
and 113. At RM 152, the striped bass TEQ-based egg concentrations exceed the NOAEL on 95lh
percentile, median, and 25th percentile bases throughout the entire modeling period (1993-2018) and
the LOAEL is exceeded on all three bases for almost the entire modeling period. At RM 113, a
NOAEL ratio of one is exceeded on a 95lh percentile basis until 2003, on a median basis until 1997,
43 TAMS/MCA
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and on a 25th percentile basis until 1994. Using the LOAEL, the 95th percentile was only exceeded
in 1993.
5.2.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria and/or
Guidelines for the Protection of Aquatic Life and Wildlife?
5.2.2.1 Measurement Endpoint: Comparison of Modeled Water Column Concentrations of
PCBs to Criteria
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria and guidelines. All forecast water concentrations (i.e., average
and 95% UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 //g/L and the
USEPA wildlife criterion of 1.2 x 10"4 //g/L at all four locations throughout the modeling period. The
whole water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity
criterion of 0.014 /^g/L for a portion of the modeling period for both average and 95% UCL at all
modeling locations. These comparisons are likely to underestimate the true risk, as concentrations
are expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.2.3 Do Modeled PCB Sediment Concentrations Exceed Appropriate Criteria and/or
Guidelines for the Protection of Aquatic Life and Wildlife?
5.2.3.1 Measurement Endpoint: Comparisons of Modeled Sediment Concentrations to
Guidelines
Table 5-1 presents the ratios of forecast sediment concentrations to various sediment
guidelines. Comparisons are made on total PCB (Tri+) sediment concentrations (i.e., NOAA, 1999a;
Persaud et al, 1993; and Washington State, 1997) and TOC-normalized sediment concentrations
(i.e., NYSDEC, 1999a and Persaud et al. 1993) to NOAA sediment effect concentrations (NOAA,
1999a), NYSDEC criteria (NYSDEC, 1999a), Ontario sediment quality guidelines (Persaud et al.,
1993), and Washington State sediment quality values (Washington State, 1997), as described in
subsection 5.1.1.1.
Forecast total PCB sediment concentrations exceeded the NOAA threshold effect
concentration, NOAA mid-range effect concentration, NYSDEC criteria for the protection of aquatic
life from chronic toxicity and wildlife from toxic effects of bioaccumulation, Ontario no effect and
lowest effect levels, and Washington State Microtox® and Hyalella azteca probable effect levels.
Many of the ratios of modeled sediment concentrations to appropriate guidelines exceed 10
or occasionally even 100. Forecast total PCB concentrations are Tri+ values, and do not include
mono or dichlorinated congeners that usually contribute a portion of the total PCB load. Thus, even
in the unlikely event that forecast sediment concentrations were to decrease by an order of magnitude
or more, comparisons to sediment guidelines would show exceedances.
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5.2.4 What Do the Available Field-Based Observations Suggest About the Health of Local
Fish Populations?
5.2.4.1 Measurement Endpoint: Evidence from Field Studies
Observational data for Hudson River fish are available for the Lower Hudson River (e.g., see
Klauda et al. 1988). The strengths and limitations of observational data have been previously
described. Based on the available data, the following observations provide insights into the potential
future risks associated with the presence of PCBs. Each insight is qualified to reflect the limitations
inherent in using observational data. In particular, there are no wildlife field studies currently
available that have directly addressed impacts associated with the presence of PCBs to Lower
Hudson River fish and wildlife.
Monitoring studies in the Lower Hudson River indicate that the fish community composition
is probably very similar to that which was present over the past few centuries. Beebe and Savidge
(1988) note that, "Except for a few species that entered the estuary through direct introductions or
through canals connecting other watersheds, the species composition of the Hudson River estuary
has probably remained similar to what it was at the time the area was settled by Europeans. All but
five species (barndoor skate, Atlantic salmon, cobia, nine-spine stickleback, and sharksucker) have
been collected within the last 20 years." No obvious losses of species that have occurred over the
past few decades during which PCB exposures have been greatest; however recommendation have
been made to limit the consumption of fish from the Lower Hudson River and the striped bass
fishery has been closed since February 1976. The qualitative data can not be used to provide insight
into the possibility that PCBs have reduced or impaired reproduction or rates of recruitment. Risks
to these endpoints could exist even if the fish species are able to maintain themselves in these areas.
For this reason, the analysis presented in subsection 5.2.1 comparing forecast body burdens to TRY
values is required to judge the possible magnitude of these risks.
The shortnose sturgeon has been on the federal endangered species list since 1967. Studies
of the abundance of shortnose sturgeon indicate that this species is reproducing in the Lower Hudson
River (below the Federal Dam) and that the population numbers are increasing (Bain, 1997).
Increases in populations in the absence of fishing pressures have not been well documented.
Ecological studies on the Hudson River during the 1970s suggest possible increases during that
period, but those increases are at least partly an artifact of improved sampling (e.g., Hoff et ai,
1988). The changing ratio of shortnose sturgeon: Atlantic sturgeon catches is also indicative of an
increasing shortnose sturgeon population in the Hudson River. While there is evidence that
populations of shortnose sturgeon are increasing following their demise at the turn of the century and
following improvements in overall water quality, the growth of the species's populations is likely to
be slow as a result of its biology. Measurable increases in shortnose sturgeon populations should not
be expected over short time periods (i.e., decades) as the species matures late (at about 7-10 years)
and spawns infrequently. While available data indicate that the population growth of shortnose
sturgeon in the Hudson is positive, it is not possible to quantify from these data the extent to which
PCB exposures might impair or reduce these population growth rates.
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Population data indicate that white perch, a semi-anadromous fish in the Lower Hudson
River, has exhibited positive population growth during the 1970s and 1980s, a period when PCB
exposures in the Lower Hudson River may have been highest. The data indicate that PCB exposures
to this fish species are not sufficiently high to significantly reduce reproduction and recruitment
rates. Wells et al. (1992) have reported on studies of the white perch during the 1970s and 1980s.
This species is a permanent resident in the Hudson and, together with the shortnose sturgeon, one
of two Hudson River species that are representative primarily of the Lower Hudson River. Wells et
al. (1992) studied several sources of Hudson River data for the period 1975 through 1987 and
concluded that the population of white perch has increased over this period. This positive population
growth has occurred during a period when PCB exposures have been occurring. This indicates that
PCB exposure to white perch has not been sufficient to prevent reproduction or recruitment. In fact,
populations have increased in size during this period. However, as noted above, there are many
factors that influence population size and it is possible that PCBs could influence rates of
reproduction and recruitment to a degree that is not manifested in recent population trends. The
analyses performed in this chapter provide insight into the degree to which PCB body burdens in
Hudson River fish might pose a risk to their reproductive and recruitment rates.
5.3 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Lower Hudson River
Insectivorous Bird Populations (as Represented by the Tree Swallow)
5.3.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Insectivorous Birds and Egg
Concentrations Exceed Benchmarks for Adverse Effects on Reproduction?
5.3.1.1 Measurement Endpoint: Modeled Dietary Doses on a Tri+ PCB Basis to Insectivorous
Birds (Tree Swallow)
Table 5-25 compares modeled dietary doses for the period 1993 - 2018 for the tree swallow
to the field-based TRY derived in the baseline ERA (USEPA, 1999c). This TRY was derived from
the USFWS data from the Hudson River. For the entire modeling period, the TQs for the tree
swallow are below one at all locations.
5.3.1.2 Measurement Endpoint: Predicted Egg Concentrations on a Tri+ PCB Basis to
Insectivorous Birds (Tree Swallow)
Table 5-26 compares predicted egg concentrations for the period 1993 - 2018 for the tree
swallow to the field-based TRY derived in the baseline ERA (USEPA, 1999c) under future
conditions. This TRY was derived from the USFWS data from the Hudson River, and the
biomagnification factor from aquatic insects to eggs was also obtained from these data. The
predicted egg concentrations used a biomagnification factor of 2 based on the USFWS tree swallow
data. For the entire modeling period, the TQs for the tree swallow are below one at all locations.
46 TAMS/MCA
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5.3.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs Expressed on a TEQ Basis
to Insectivorous Birds (Tree Swallow)
Table 5-27 compares the estimated TEQ-based dietary dose and predicted egg concentration
to the piscivorous birds to the field-based TRY for TEQs derived from the Phase 2 database
(USEPA, 1998b). For the entire modeling period (1993-2018), the TQs for. the tree swallow are
below one at all locations.
5.3.1.4 Measurement Endpoint: Predicted Egg Concentrations Expressed on a TEQ Basis to
Insectivorous Birds (Tree Swallow)
Table 5-28 compares the estimated TEQ-based predicted egg concentrations for insectivorous
birds to the field-based TRY for TEQs derived for egg concentrations. The predicted egg
concentrations used a biomagnification factor of 7 based on the USFWS tree swallow data. For the
entire modeling period, the TQs for the tree swallow are below one at all locations for the entire
modeling period.
5.3.2 Do Modeled Water Concentrations Exceed Criteria for Protection of Wildlife?
5.3.2.1 Measurement Endpoint: Comparison of Modeled Water Column Concentrations to
Criteria for the Protection of Wildlife
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 yug/L and the USEPA
wildlife criterion of 1.2 x 10"* £ig/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/N YSDEC benthic aquatic life chronic toxicity criterion
of 0.014 //g/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.3.3 What Do the Available Field-Based Observations Suggest About the Health of Local
Insectivorous Bird Populations?
5.3.3.1 Measurement Endpoint: Evidence from Field Studies
A natural history study of the wildlife species known to forage and reproduce within the
project site represents an important measurement endpoint. Whereas a species is not required to be
currently using a site for inclusion in the ecological risk assessment (i.e., the species may have been
severely impacted by site contamination/conditions), evidence of past use is important in validating
the endpoints and toxicity factors utilized in the analysis.
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The last ten annual Audubon Society Christmas bird counts for Albany, Rensselear,
Dutchess, Putnam, Southern and East Orange, Rockland, Catskill, Lower Hudson, and
Bronx/Westchester count circles (Cornell University, 1999) were examined to determine whether
any general inferences on insectivorous bird populations along the Hudson River could be made.
Because many insectivorous bird species are migratory (e.g., flycatchers, swallows, gnatcatchers),
the Christmas count alone does not provide a good population estimate for these species.
Despite their migratory nature, tree swallows were observed in Christmas count circles along
the Lower Hudson River. The Saw Mill Audubon Society provided year-round information on bird
sightings at Croton Point Park in Westchester since January 1994 (Bickford, 1999). Tree swallows
have been sighted from March to September, with the exception of during July. Lack of adequate
nesting holes may account for the low numbers of summer sightings.
The Lower Hudson Valley Bird Line transcripts (sponsored by the Sullivan County, Saw Mill
River, Rockland, Putnam Highlands, and Bedford Audubon Society chapters) from January 1998
to August 1999 (Audubon, 1999) were reviewed. Tree swallows were noted in the transcripts in the
spring months (March, April, and May) and again in the fall and winter (October to January).
5.4 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Lower Hudson River Waterfowl
Populations (as Represented by the Mallard)
5.4.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Waterfowl and Egg
Concentrations Exceed Benchmarks for Adverse Effects on Reproduction?
5.4.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ PCBs to Waterfowl (Mallard)
Table 5-29 provides the results of the comparison between predicted dietary doses of the
female mallard based on predictions for the modeling period 1993 to 2018 to the laboratory-based
NOAEL and LOAEL TRVs developed in the baseline ERA (USEPA, 1999c). On a NOAEL basis,
the predicted TQs exceed one on both an average and 95% UCL period for a portion of the modeling
period at all four locations. At RM152, the 95% UCL exceeds one until 2007, and the average until
2004. On a LOAEL basis, predicted TQs do not exceed one at any location.
5.4.1.2 Measurement Endpoint: Predicted Egg Concentrations of Tri+ PCBs to Waterfowl
(Mallard)
Table 5-30 provides the results of the comparison between predicted egg concentrations and
laboratory-based TRVs for the period 1993 to 2018. The predicted egg concentrations used a
biomagnification factor of 3 based on the USFWS mallard and wood duck data. The TQs for mallard
eggs exceed one for the duration of the modeling period on a NOAEL basis, for both the average and
95% UCL, at all four locations for the entire modeling period. LOAEL-based comparisons exceed
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one for both the average and 95% UCL at RM 152 for the entire modeling period and at RM 113 for
most of the modeling period (until 2016). The LOAEL also exceeds one on an average and 95%
UCL basis for a portion of the modeling period at RMs 90 and 50.
5.4.1.3 Measurement Endpoint: Modeled Dietary Doses of TEQ-Based PCBs to Waterfowl
(Mallard)
Table 5-31 provides the results of the comparison between predicted dietary doses and female
mallard PCB dietary doses on a TEQ basis to laboratory-based TRVs. The results presented in this
table show that the NOAEL and LOAEL-based comparisons exceed one at all four locations for the
duration of the modeling period (1993-2018), for both the average and the 95% UCL concentrations
by up to two orders of magnitude.
5.4.1.4 Measurement Endpoint: Predicted Egg Concentrations of TEQ-Based PCBs to
Waterfowl (Mallard)
Table 5-32 provides the results of the comparison between predicted concentrations of PCBs
in mallard egg and the field-based TRY for TEQs derived in the baseline ERA (USEPA, 1999c),
using a biomagnification factor of 28. These results show that predicted TQs exceed one for all
locations, years, and concentrations. Predicted TQs exceed 100 on a NOAEL and LOAEL basis at
RMs 152 and 113 locations for the duration of the modeling period and exceed 100 on a NOAEL
basis at RMs 90 and 50. This suggests the potential for adverse reproductive effects to waterfowl
species.
5.4.2 Do Modeled PCB Water Concentrations Exceed Criteria for the Protection of Wildlife?
5.4.2.1 Measurement Endpoint: Comparison of Modeled Water Concentrations to Criteria
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All predicted water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 pg/L and the USEPA
wildlife criterion of 1.2 x 10"4 //g/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity criterion
of 0.014 /zg/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
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5.4.3 What Do the Available Field-Based Observations Suggest About the Health of Lower
Hudson River Waterfowl Populations?
5.4.3.1 Measurement Endpoint: Observational Studies
The last ten annual Audubon Society Christmas bird counts for the Lower Hudson Valley
count circles (Cornell University, 1999) were examined to determine whether any inferences on local
waterfowl populations along the Hudson River could be made. Mallards were generally one of the
most abundant species sighted during the Christmas count. Other waterfowl, including Canada geese,
American black duck, ring-necked duck, ruddy duck, and common merganser are commonly seen
in the Hudson River area. Mallards, Canada geese, and mute swans were sighted throughout the year
in Croton Point Park (Bickford, 1999).
The Saw Mill Audubon Society provided information on bird sightings at Croton Point Park
in Westchester since January 1994 (Bickford, 1999). Mallards are numerous at Croton Point Park,
but nesting is probably limited due to lack of proper habitat. On the basis of breeding surveys, the
mallard population using the Hudson River estuary is stable to increasing (NYSDEC, 1997).
Not all waterfowl are likely to be adversely impacted by PCBs (particularly in the less
contaminated stretches), but PCB sensitive species may experience total reproductive failure nesting
in more contaminated areas.
5.5 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Hudson River Piscivorous Bird
Populations (as Represented by the Belted Kingfisher, Great Blue Heron,
and Bald Eagle)
5.5.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Piscivorous Birds and Egg
Concentrations Exceed Benchmarks for Adverse Effects on Reproduction?
5.5.1.1 Measurement Endpoint: Modeled Dietary Doses of Total PCBs for Piscivorous Birds
(Belted Kingfisher, Great Blue Heron, Bald Eagle)
Tables 5-33 through 5-35 compare the estimated total PCB (i.e., Tri+) dietary dose of the
female belted kingfisher, great blue heron, and bald eagle to the laboratory-based TRVs presented
in Table 4-2 and derived in the baseline ERA (USEPA, 1999c). The site-related doses are based on
modeled concentrations in forage fish, piscivorous fish, benthic invertebrates, whole water, and
sediment using the results from the FISHRAND (fish and invertebrates) and Farley et al. (1999)
(water and sediment) models.
The ratio of the female belted kingfisher dietary doses to the TRVs exceed one at all four
locations for the entire modeling period on both a NOAEL and LOAEL basis (Table 5-33).
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The ratio of the female great blue heron dietary doses to the TRVs exceed one at all four
locations for the entire modeling period on a NOAEL basis (Table 5-34). Estimated TQs exceed one
on a LOAEL basis at all locations for portions of the modeling period.
Table 5-35 presents the results for the bald eagle. Again, all comparisons exceed one for the
duration of the modeling period at all locations on both a NOAEL and LOAEL basis for both
average and 95% UCL doses.
Reproductive effects TQs for great blue heron, belted kingfisher, and bald eagle using
average and upper confidence limits all exceed one. This indicates that exposure to PCBs from the
Hudson River via prey and water present a risk of reproductive effects to these species on the basis
of modeled Tri+ PCB dietary doses as compared to appropriate toxicity reference values. These
results suggest the possibility of population-level impacts, as these TQs are based on reproductive
effects, and consistently exceed one over the course of the modeling period.
5.5.1.2 Measurement Endpoint: Predicted Egg Concentrations Expressed as Tri+ to
Piscivorous Birds (Eagle, Great Blue Heron, Kingfisher)
Tables 5-36 through 5-38 compare the estimated total PCB (i.e., Tri+) predicted egg
concentrations for the belted kingfisher, great blue heron, and bald eagle to the toxicity benchmarks
summarized in Table 4-2. Laboratory-based NOAELs and LOAELs were used for the belted
kingfisher and the great blue heron, whereas a field-based NOAEL was selected for the bald eagle.
Egg concentrations are estimated using a biomagnification factor of 28 from Giesy et al. (1995).
Table 5-36 presents the results for the modeled belted kingfisher egg concentrations. These
results are similar to those shown for the dietary dose. All comparisons at all locations exceed one
a NOAEL and LOAEL basis using both average and 95% UCL concentrations for the duration of
the modeling period.
Table 5-37 presents the results for the great blue heron. Again, all comparisons at all four
locations exceed one on both a NOAEL and LOAEL basis for the duration of the modeling period.
Table 5-38 presents the results for the bald eagle. These results are similar to those shown
for the dietary dose. All comparisons at all locations exceed one for the duration of the modeling
period.
All of the predicted TQs exceeded one on the basis of estimated egg concentrations. These
results suggest that exposure of piscivorous birds to PCBs from the Hudson River may result in
adverse reproductive effects. The elevated TQ over time for the modeling period 1993 to 2018
suggests that exposure to PCBs over the long term has the potential to impact piscivorous birds, as
represented by these species, on a population level.
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5.5.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs Expressed as TEQs to
Piscivorous Birds (Belted Kingfisher, Great Blue Heron, Bald Eagle)
Tables 5-39 through 5-41 present the results of the comparison between modeled dietary
doses expressed on a TEQ basis to piscivorous receptors over the modeling period (1993 - 2018).
Dietary doses were estimated using modeled concentrations in forage fish, piscivorous fish, benthic
invertebrates, whole water, and sediment using the results from the FTSHRAND (fish and
invertebrates) and Farley et al. (1999) (water and sediment) models. Model results were multiplied
by the weighted TEF factors derived in the baseline ERA (USEPA, 1999c). Laboratory-based TRVs
for TEQs were used for all species (Table 4-2).
The ratio of the female belted kingfisher PCB dietary doses on a TEQ-basis to the TRVs
exceed one at all four locations for the entire modeling period on both a NOAEL and LOAEL basis
(Table 5-39).
The ratio of the female great blue heron dietary doses to the TRVs exceed one at all four
locations for the entire modeling period on a NOAEL basis using both average and 95%UCL doses
(Table 5-40). Estimated TQs exceed one on a LOAEL basis at all locations for portions of the
modeling period.
Table 5-41 presents the TEQ-basis ratios for the bald eagle. All comparisons exceed one for
the duration of the modeling period at all locations on both a NOAEL and LOAEL basis, with the
exception of the LOAEL ratios at RM 50 for 2106-2018.
Reproductive effects TQs for great blue heron, belted kingfisher, and bald eagle using the
average and 95% upper confidence limit on a TEQ basis often exceed one, and in many cases exceed
100. This indicates that PCBs from the Hudson River in the diet and water are likely to result in
adverse reproductive effects to these species on the basis of modeled TEQ-based PCB dietary doses
as compared to appropriate toxicity reference values. These results suggest adverse population-level
effects may occur, given the consistent exceedance of a reproductive-based endpoint.
5.5.1.4 Measurement Endpoint: Modeled Dietary Doses of PCBs Expressed as TEQs to
Piscivorous Birds (Belted Kingfisher, Great Blue Heron, Bald Eagle)
Tables 5-42 through 5-45 present the results of the comparison between piscivorous bird egg
concentrations expressed on a TEQ-basis to TRVs (laboratory-based for the kingfisher and eagle,
field-based for the heron) for the period 1993-2018. Egg concentrations were estimated using
modeled concentrations in forage fish and piscivorous fish from the FISHRAND. Model results were
multiplied by the weighted TEF derived in the ERA (USEPA, 1999c) and then multiplied by a
biomagnification factor of 19 (Giesy et al., 1995).
The belted kingfisher ratios exceed one for at all four locations throughout the entire
modeling period (Table 5-42).
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The ratio of the female great blue heron egg concentration to the TEQ-based TRY egg
concentration exceed one at all four locations for the entire modeling period on a NOAEL basis
(Table 5-34). Estimated TQs also exceed one on a LOAEL basis at RMs 152 and 113 for all of the
modeling period and at RMs 90 and 50 for most of the modeling period (i.e., up to 2014 or later).
The bald eagle TQs exceed one for at all four locations throughout the entire modeling period
(Table 5-45). Ratios are as high as three orders of magnitude above one.
TQs based on reproductive effects for the great blue heron, belted kingfisher, and bald eagle
using average and upper confidence limits on a TEQ basis all exceed one, and in many cases exceed
100, and several of the bald eagle TQs exceed 1000. This indicates that PCBs from the Hudson River
in fish as they translate to egg concentrations are likely to result in adverse reproductive effects to
these species on the basis of modeled TEQ-based PCB egg concentrations as compared to
appropriate TRVs. These results suggest adverse population-level effects may occur, given the
consistent exceedance of a reproductive-based endpoint.
5.5.2 Do Modeled Water Concentrations Exceed Criteria for the Protection of Wildlife?
5.5.2.1 Measurement Endpoint: Comparison of Modeled Water Concentrations to Criteria
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 yug/L and the USEPA
wildlife criterion of 1.2 x 10"4 fj.g/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity criterion
of 0.014 yug/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.5.3 What Do the Available Field-Based Observations Suggest About the Health of Local
Piscivorous Bird Populations?
5.5.3.1 Measurement Endpoint: Observational Studies
Both the New York State Endangered Species Unit and The Atlas of Breeding Birds in New
York (Andrle and Carroll, 1988) provide general information regarding the bird, species using the
Hudson River. The belted kingfisher (Ceryle alcyori) appears to breed along the Hudson River north
of Westchester County in areas such as Oscawana and George's Island Parks. Belted kingfishers may
also be found in the area year-round, as evidenced by sightings of it in the Christmas bird count
(Cornell University, 1999).
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The great blue heron (Ardea herodias) is found along the Lower Hudson River throughout
the year. It has been observed in most count circles during the Christmas bird count (Cornell
University, 1999). There is a breeding colony of herons in the freshwater portion of the Lower
Hudson River (Rensselaer County).
Bald eagles are slowly returning to the Lower Hudson River Valley. Up to 40 eagles have
wintered in the 30 miles between Danskammer Point (Orange County) and Croton Point
(Westchester County) in the last few years (USGS, 1999). Releases of young eagles in the 1980's
have resulted in two nesting pairs along the Hudson River. However, these two breeding pairs have
been unsuccessful in producing offspring (USGS, 1999). Bald eagles have been sighted
intermittently during Christmas counts conducted in the last 10 years (Cornell University, 1999).
5.6 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Insectivorous Mammal Populations (as represented
by the Little Brown Bat)
5.6.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Insectivorous Mammalian
Receptors Exceed Benchmarks for Adverse Effects on Reproduction?
5.6.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ to Insectivorous Mammalian
Receptors (Little Brown Bat)
Modeled total PCB (Tri+) dietary dose comparisons to laboratory-based TRVs (Table 4-3)
are presented for the female little brown bat in Table 5-45 for the period 1993 - 2018. Dietary doses
are estimated by using forecast water concentrations from the Farley et al. (1999) model and
predicted invertebrate (aquatic insect) concentrations derived from the FISHRAND model. These
results show that all comparisons exceed one for at all four locations throughout the modeling period
on both a NOAEL and LOAEL basis for both average and 95%UCL doses.
These results suggest the potential for adverse reproductive effects to insectivorous
mammalian species at all locations in the Lower Hudson River based on using predicted future
concentrations in the exposure models.
5.6.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ Basis to Insectivorous
Mammalian Receptors (Little Brown Bat)
Modeled PCB dietary dose on a TEQ basis comparisons to laboratory-based TRVs for TEQs
(Table 4-3) are presented for the little brown bat in Table 5-46. These results show that all
comparisons exceed one (by one or two orders of magnitude) at all locations during the entire
modeling period on both a NOAEL and LOAEL basis.
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These results suggest the potential for adverse reproductive effects to insectivorous
mammalian species at all locations in the river based on using the results from the baseline modeling
in the exposure models. Given the consistency of the results, the magnitude of the exceedances, and
the duration of the exceedances, these results suggest the potential for population-level adverse
reproductive effects.
5.6.2 Do Modeled Water Concentrations Exceed Criteria for Protection of Wildlife?
5.6.2.1 Measurement Endpoint: Comparison of Modeled Water Concentrations to Criteria for
the Protection of Wildlife
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 /zg/L and the USEPA
wildlife criterion of 1.2 x 10"4 //g/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity criterion
of 0.014 //g/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.6.3 What Do the Available Field-Based Observations Suggest About the Health of Local
Insectivorous Mammalian Populations?
5.6.3.1 Measurement Endpoint: Observational Studies
A limited amount of data is available on little brown bat populations in the Lower Hudson
River, and only a small subset of that data is within a time frame relevant to this study. Therefore,
field-based observations do not provide sufficient information to evaluate this measurement
endpoint.
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5.7 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Omnivorous Mammal Populations (as represented
by the Raccoon)
5.7.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Omnivorous Mammalian
Receptors Exceed Benchmarks for Adverse Effects on Reproduction?
5.7.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ to Omnivorous Mammalian
Receptors (Raccoon)
Modeled total PCB (Tri+) dietary dose comparisons to laboratory based TRVs (Table 4-3)
are presented for the female raccoon in Table 5-47 for the period 1993 - 2018. Dietary doses are
estimated by using forecast water concentrations from the Farley et al. (1999) model and predicted
forage fish and benthic invertebrate concentrations from the FISHRAND model.
Predicted TQs for RMs 152, 113, and 90 exceed one on a NOAEL basis for both the average
and 95% UCL. At RM 50 TQs exceed one on using the 95% UCL concentration until 2011 and
using the average concentration until 2007. TQs were below one at all locations on a LOAEL basis.
5.7.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ Basis to Omnivorous
Mammalian Receptors (Raccoon)
Modeled PCB dietary dose on a TEQ basis comparisons to laboratory-based TRVs for TEQs
(Table 4-3) are presented for the female raccoon in Table 5-48 for the period 1993 - 2018. All
comparisons exceed one at all four locations for the duration of the modeling period on both a
NOAEL and LOAEL basis for both average and 95% UCL concentrations.
These results suggest the potential for adverse reproductive effects to omnivorous
mammalian species in the Lower Hudson River. Given the consistency of the results, the magnitude
of the exceedances, and the duration of the exceedances, these results suggest the potential for
population-level adverse reproductive effects in the Lower Hudson River.
5.7.2 Do Modeled Water Concentrations Exceed Criteria for Protection of Wildlife?
5.7.2.1 Measurement Endpoint: Comparison of Modeled Water Concentrations to Criteria for
the Protection of Wildlife
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 //g/L and the USEPA
wildlife criterion of 1.2 x 10"* /ug/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/N YSDEC benthic aquatic life chronic toxicity criterion
56 TAMS/MCA
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of 0.014 yug/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations, are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.7.3 What Do the Available Field-Based Observations Suggest About the Health of Local
Omnivorous Mammalian Populations?
5.7.3.1 Measurement Endpoint: Observational Studies
A limited amount of quantitative data is available on raccoon populations in the Lower
Hudson River. However, casual observations imply that raccoons are abundant along the Lower
Hudson River Valley. However, a large proportion of the raccoon population in the Lower Hudson
River Valley is likely to be obtaining food from sources other than the Hudson River, as the raccoon
is an opportunistic feeder. Therefore, only a small subset of the Lower Hudson River Valley raccoon
population is likely to be experience the daily doses calculated in the ERA Addendum.
5.8 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Piscivorous Mammal Populations (as represented
by the Mink and River Otter)
5.8.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Piscivorous Mammalian
Receptors Exceed Benchmarks for Adverse Effects on Reproduction?
5.8.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ to Piscivorous Mammalian
Receptors (Mink, River Otter)
Tables 5-49 and 5-50 present the results of the comparison between modeled dietary doses
to female mink and river otter under future conditions (1993-2018). Field-based TRVs derived in
the baseline ERA (Table 4-3) are used for both species. Modeled dietary doses are estimated by
using Farley et al. (1999) model results for water and sediment, and FISHRAND results for forage
fish and piscivorous fish concentrations.
On a dietary dose basis for total (Tri+) PCBs, predicted TQs for the female mink exceed one
on a NOAEL basis at all four locations for both the average and 95% UCL (Table 5-49). TQs were
below one at all locations on a LOAEL basis.
Table 5-50 shows the results for the female river otter. On a dietary dose basis for total (Tri+)
PCBs, predicted TQs exceed one on both a NOAEL and LOAEL basis at RMs 152 and 113 for
average and 95% UCL doses. At RMs 90 and 50, a ratio of one is exceeded for on a NOAEL basis
(average and 95%UCL). On a LOAEL basis, one is exceeded until 2004 at RM 90 and until 2002
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at RM 50. The river otter consumes a larger size range of fish than the mink and is likely to obtain
fish from deeper in the river. Thus, the exposure of the river otter is greater than that of the mink.
These results suggest the potential for adverse reproductive effects to piscivorous mammalian
species in the Hudson River based on using model results in the exposure models for dietary dose.
Reproductive effects TQs for the mink and otter using average and upper confidence limits exceed
one for the duration of the modeling period, often by more than two orders of magnitude. Given the
consistency of the results, the magnitude of the exceedances, and the duration of the exceedances,
these results suggest that PCBs from the Lower Hudson River in the diet and water are likely to
present a significant risk of reproductive effects to the mink and river otter.
5.8.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ Basis to Piscivorous
Mammalian Receptors (Mink, River Otter)
Tables 5-51 and 5-52 present the results of the comparison between modeled dietary doses
to mink and river otter under future conditions for the period 1993 - 2018 on a TEQ basis. Modeled
mink dietary doses on a TEQ basis exceed the field-based NOAEL and LOAEL for TEQs (Table 4-
3) at all four locations for the duration of the modeling period for both the average and 95% UCL
(Table 5-51).
Table 5-52 shows the results for the female river otter. Modeled otter dietary doses on a TEQ
basis exceed the field-based NOAEL and LOAEL for TEQs one at all four locations for the duration
of the modeling period for both the average and 95% UCL by up to three orders of magnitude. The
river otter, which consumes larger fish than the mink, demonstrates higher TQs than the mink, as
seen by comparing Tables 5-51 and 5-52.
These results suggest the potential for adverse reproductive effects to piscivorous mammalian
species in the Hudson River based on using Farley et al. (1999) and FISHRAND model results in
the exposure models for dietary dose. Given the consistency of the results, the magnitude of the
exceedances, and the duration of the exceedances, these results suggest the potential for population-
level adverse reproductive effects for mink and river otter consuming fish from the Hudson River.
Reproductive effects TQs for the mink and river otter using average and upper confidence
limits all exceed one on both a total PCB and TEQ basis, with generally higher TEQ based TQs. This
indicates that PCBs from the Lower Hudson River in the diet and water are likely to present a
significant risk of reproductive effects to the mink and river otter on the basis of modeled PCB
dietary doses as compared to appropriate toxicity reference values.
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5.8.2 Do Modeled Water Concentrations Exceed Criteria for the Protection of Piscivorous
Mammals?
5.8.2.1 Measurement Endpoint: Comparison of Modeled Water Concentrations to Criteria for
the Protection of Wildlife
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 yUg/L and the USEPA
wildlife criterion of 1.2 x 10"* //g/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity criterion
of 0.014 //g/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.8.3 What Do the Available Field-Based Observations Suggest About the Health of Local
Mammalian Populations?
5.8.3.1 Measurement Endpoint: Observational Studies
NYSDEC is currently performing a comprehensive study of three distinct aspects of injury
to Hudson River semi-aquatic mammals (Mayack, 1999a). This study consists of:
• Measuring the levels and nature of contamination in mink, muskrat, and otter from
within the Hudson River watershed.
• Measuring the population size and distribution of selected mammals throughout the
Hudson River ecosystem.
• Comparing mammalian reproductive success in the Upper Hudson River with that in
the Lower Hudson River.
A primary objective of the NYSDEC study is to evaluate the extent of PCB contamination
in mink, river otter, and muskrat populations downstream of a major point source at Fort Edward,
NY. Analysis of a small number of mink and otter collected from the Hudson River region (Foley
et al., 1988) suggests that concentrations of PCBs in mink may cause reproductive impairment and
a consequent decease in wild populations. Contaminant levels in populations upstream of Fort
Edward will be compared to levels in populations downstream. The study aims to establish a
downstream limit of potential contaminant impact on mammal populations in the Hudson River
ecosystem. A second objective is to determine if the abundance of mink can be related to the
distribution of PCB contamination within the Hudson River drainage.
Preliminary results from this study indicate that PCBs may have an adverse effect on the litter
size and possibly kit survival of river otter in the Hudson River (Mayack, 1999b). Mink appear to
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be accumulating PCBs to a lesser extent than river otter, possibly because their diet has a greater
proportion of uncontaminated prey. However, given the variability in diet and opportunistic nature
of mink foraging a portion of the population may be exposed to high dietary levels of PCBs if
aquatic prey are available. Levels of PCBs in river otter may represent a diet more highly
contaminated with PCBs than that of mink, because fish comprise the majority of the river otter diet.
Mink, river otter, and muskrats are found in several localized areas along the Lower Hudson
River. The herbivorous/omnivorous muskrat has had low pup abundances up and down the Hudson
River (Kiviat, 1999). The reason is unknown.
5.9 Evaluation of Assessment Endpoint: Protection of Threatened and
Endangered Species
Two threatened and/or endangered species, the shortnose sturgeon and bald eagle, were
selected as receptors in this assessment. The populations of other endangered, protected, and species
of concern found along the Hudson River (Chapter 2.6.5) may also be affected by PCBs. The bald
eagle is considered to be a representative surrogate for wildlife species, and the shortnose sturgeon
a representative surrogate for fish.
5.9.1 Do Modeled Total and TEQ-Based PCB Body Burdens in Local Threatened or
Endangered Fish Species Exceed Benchmarks for Adverse Effects on Fish
Reproduction?
5.9.1.1 Measurement Endpoint: Inferences Regarding Shortnose Sturgeon Population
There are no experimental data available to assess uptake of PCBs by shortnose sturgeon. To
evaluate the potential impact of PCBs on shortnose sturgeon, observed and modeled largemouth bass
total and TEQ based PCB concentrations were compared to toxicity reference values.
The derived toxicity reference values (Table 4-1) are considered protective of this species.
This analysis assumes that shortnose sturgeon are likely to experience patterns of uptake somewhere
between a largemouth bass and a brown bullhead. Shortnose sturgeon are primarily omnivorous, but
can live in excess of 30 years and thus might be expected to accumulate more PCBs than their diet
a/one would suggest.
For PCBs expressed as total PCBs, the comparison is no different from the results already
presented for the brown bullhead for Tri+ PCBs (Tables 5-10 and 5-11) and largemouth bass on a
TEQ basis (Tables 5-22 and 5-23), because the toxicity reference values are the same.
The analyses performed for both total (Tri+) and TEQ-based PCBs indicate the potential for
adverse effects as compared to the NOAEL and LOAEL TRY values. Therefore, the potential for
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adverse reproductive effects in shortnose sturgeon exists, particularly in the upper reaches of the
Lower Hudson River (i.e., RMs 152 and 113).
5.9.2 Do Modeled Total and TEQ-Based PCB Body Burdens/Egg Concentrations in Local
Threatened or Endangered Species Exceed Benchmarks for Adverse Effects on Avian
Reproduction?
5.9.2.1 Measurement Endpoint: Inferences Regarding Bald Eagle and Other Threatened or
Endangered Species Populations
The modeled results for the bald eagle were presented in Section 5.5. Almost all comparisons
across all locations and on a total PCB and TEQ-basis exceeded one, in some instances by more than
three orders of magnitude. Both the dietary dose and egg-based results were consistent in this regard.
Other threatened or endangered raptors, such as the peregrine falcon, osprey, northern harrier, and
red-shouldered hawk may experience similar exposures.
5.9.3 Do Modeled Water Concentrations Exceed Criteria for the Protection of Wildlife?
5.9.3.1 Measurement Endpoint: Comparisons of Modeled Water Concentrations to Criteria
for the Protection of Wildlife
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 /zg/L and the USEPA
wildlife criterion of 1.2 x 10"4 /zg/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity criterion
of 0.014 //g/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
5.9.4 Do Modeled Sediment Concentrations Exceed Guidelines for the Protection of Aquatic
Health?
5.9.4.1 Measurement Endpoint: Comparisons of Modeled Sediment Concentrations to
Guidelines
Table 5-1 presents the ratios of forecast sediment concentrations to various sediment
guidelines. Comparisons are made on total PCB (Tri+) sediment concentrations (i.e., NOAA, 1999a;
Persaud et al., 1993; and Washington State, 1997) and TOC-normalized sediment concentrations
(Le., NYSDEC, 1999a and Persaud et al. 1993) to NOAA sediment effect concentrations (NOAA,
1999a), NYSDEC criteria (NYSDEC, 1999a), Ontario sediment quality guidelines (Persaud et a/.,
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1993), and Washington State sediment quality values (Washington State, 1997), as described in
subchapterS.l.l.l.
Forecast total PCB sediment concentrations exceeded the NOAA threshold effect
concentration, NOAA mid-range effect concentration, NYSDEC criteria for the protection of aquatic
life from chronic toxicity and wildlife from toxic effects of bioaccumulation, Ontario no effect and
lowest effect levels, and Washington State Microtox® and Hyalella azteca probable effect levels.
Many of the ratios of modeled sediment concentrations to appropriate guidelines exceed 10
or occasionally even 100. Forecast total PCB concentrations are Tri+ values, and do not include
mono or dichlorinated congeners that usually contribute a portion of the total PCB load. Thus, even
in the unlikely event that forecast sediment concentrations were to decrease by an order of magnitude
or more, comparisons to sediment guidelines would show exceedances.
5.9.5 What Do the Available Field-Based Observations Suggest About the Health of Local
Threatened or Endangered Fish and Wildlife Species Populations?
5.9.5.1 Measurement Endpoint: Observational Studies
While available data indicate that the population growth of shortnose sturgeon in the Hudson
is positive, it is not possible to quantify from these data the extent to which PCB exposures might
impair or reduce these population growth rates. The kinds of effects expected in the field include
reduced fecundity, decreased hatching success, and similar kinds of reproductive impairment
indicators, which are often difficult to discern. These effects may be masked by populations increases
due to protection from fishing pressures.
The bald eagle was discussed in subsection 5.5.3.1. Bald eagles are slowly returning to the
Lower Hudson River Valley, however their long-term breeding success is unknown. Releases of
young eagles in the 1980's have resulted in two nesting pairs along the Hudson River. However,
these two breeding pairs have been unsuccessful in producing offspring (USGS, 1999). Part of the
difficulty of assessing populations is that there are no reference data to measure abundance against,
as bald eagles have not breed along the Hudson River for decades.
5.10 Evaluation of Assessment Endpoint: Protection of Significant Habitats
The significant habitats found along the Hudson River (Tables 2-3) are unique, unusual, or
necessary for the propagation of key species. Various measurement endpoints developed throughout
this risk assessment are used to determine the potential for adverse effects on significant habitats and
the animals and plants associated with them, rather than performing a quantitative evaluation of risks
to ecological communities.
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5.10.1 Do Modeled Total and TEQ-Based PCB Body Burdens/Egg Concentrations in
Receptors Found in Significant Habitats Exceed Benchmarks for Adverse Effects on
Reproduction?
5.10.1.1 Measurement Endpoint: Inferences Regarding Receptor Populations
Based on the comparisons of observed and modeled body burdens to toxicity reference values
presented in this chapter, current PCB concentrations found in the Lower Hudson River (i.e., RMs
152, 113, 90, and 50) exceed toxicity reference values for some fish, avian, and mammalian
receptors. These comparisons indicate that animals feeding on Lower Hudson River-based prey may
be affected by the concentrations of PCBs found in the river on both a total PCB and TEQ basis. In
addition, based on the ratios obtained in this evaluation, other taxononic groups not directly
addressed in this evaluation (e.g., amphibians and reptiles) may also be affected by exposure to PCBs
in the Lower Hudson River.
Many year-round and migrant species use the significant habitats along the Lower Hudson
River for breeding or rearing their young. Therefore, exposure to PCBs may occur at a sensitive time
in the life cycle (i.e., reproductive and development) and have a greater effect on populations than
at other times of the year.
5.10.2 Do Modeled Water Column Concentrations Exceed Criteria for the Protection of
Aquatic Wildlife?
5.10.2.1 Measurement Endpoint: Comparison of Modeled Water Concentrations to Criteria
for the Protection of Wildlife
Table 5-2 presents the results of the comparison between modeled whole water PCB
concentrations and appropriate criteria. All forecast water concentrations (i.e., average and 95%
UCL) exceed the NYSDEC wildlife bioaccumulation criterion of 0.001 jzg/L and the USEPA
wildlife criterion of 1.2 x 10"4 //g/L at all four locations throughout the modeling period. The whole
water concentrations also exceed the USEPA/NYSDEC benthic aquatic life chronic toxicity criterion
of 0.014 //g/L for a portion of the modeling period for both average and 95% UCL at all modeling
locations. These comparisons are likely to underestimate the true risk, as concentrations are
expressed as the sum of the Tri+ and higher congeners, while the criteria are based on total PCBs
(the sum of all congeners).
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5.10.3 Do Modeled Sediment Concentrations Exceed Guidelines for the Protection of Aquatic
Health?
5.10.3.1 Measurement Endpoint: Comparison of Modeled Sediment Concentrations to
Guidelines for the Protection of Aquatic Health
Table 5-1 presents the ratios of forecast sediment concentrations to various sediment
guidelines. Comparisons are made on total PCB (Tri+) sediment concentrations (i.e., NOAA, 1999;
Persaud et a/., 1993; and Washington State, 1997) and TOC-normalized sediment concentrations
(i.e., NYSDEC, 1999a and Persaud et al. 1993) to NOAA sediment effect concentrations (NOAA,
1999a), NYSDEC criteria (NYSDEC, 1999a), Ontario sediment quality guidelines (Persaud et al,
1993), and Washington State sediment quality values (Washington State, 1997), as described in
subchapterS.l.l.l.
Forecast total PCB sediment concentrations exceeded the NOAA threshold effect
concentration, NOAA mid-range effect concentration, NYSDEC criteria for the protection of aquatic
life from chronic toxicity and wildlife from toxic effects of bioaccumulation, Ontario no effect and
lowest effect levels, and Washington State Microtox® and Hyalella azteca probable effect levels.
Many of the ratios of modeled sediment concentrations to appropriate guidelines exceed 10
or occasionally even 100. Predicted total PCB concentrations are Tri+ values, and do not include
mono or dichlorinated congeners that usually contribute a portion of the total PCB load. Thus, even
in the unlikely event that forecast sediment concentrations were to decrease by an order of magnitude
or more, comparisons to sediment guidelines would show exceedances.
5.10.4 What Do the Available Field-Based Observations Suggest About the Health of
Significant Habitat Populations?
5.10.4.1 Measurement Endpoint: Observational Studies
The Waterfront Revitalization and Coastal Resources Act (WRCR) of 1981 declares it to be
the public policy of New York State to conserve, protect, and, where appropriate, promote
commercial and recreational use of fish and wildlife resources and to conserve fish and wildlife
habitats identified by NYSDEC as critical to the maintenance or re-establishment of species of fish
and wildlife (Executive Law of New York, Article 42, Sections 910-920). The implementation of
this policy required that significant coastal habitats be identified and designated for protection. It was
not feasible to designate very large ecosystem, such as the Hudson River, even though they support
significant fish and wildlife populations. This would diminish the ability of the area's fish and
wildlife values to compete with other land uses. Therefore, only smaller, discrete communities that
contribute to the overall significance of the large ecosystem were evaluated (NYSDEC, 1984).
Because the effort to designate significant habitats was undertaken in the early 1980s, it can
be assumed that these areas support important biological resources although they have been exposed
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to PCBs since the 1940s. Information on species observed using significant habitats in the Lower
Hudson River is of limited use because there are no data available for the comparison of biological
resources prior to exposure to PCBs. In addition, many areas experience other effects (e.g.,
development and habitat loss) at the same time as PCB exposure, so it would be difficult to segregate
out the cause for changes in communities, even if data were available. However, based on the
receptor analyses provided in the previous sections, some sensitive species may experience
reproductive effects when attempting to breed in Lower Hudson River significant habitats.
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Chapter 6
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6.0 UNCERTAINTY ANALYSIS
A qualitative or quantitative assessment of risk is inherently uncertain. At each step of the
risk assessment process there are sources of uncertainty. The sources of uncertainty in this ERA
Addendum include:
• Sampling error and representativeness;
• Analysis and quantitation uncertainties;
• Conceptual model uncertainties;
• Toxicological study uncertainties; and,
• Exposure and modeling uncertainties.
The first two sources of uncertainty are discussed in greater detail in the baseline ERA
(USEPA, 1999c). The remaining three sources of uncertainty are discussed in the following sections.
6.1 Conceptual Model Uncertainties
The conceptual model links PCB sources, likely exposure pathways, and potential ecological
receptors. It is intended to provide broad linkages of various receptor groups found along the
Hudson River to PCB contamination in Hudson River sediments and surface waters. However,
because it is a generalized model, it is not intended to mimic actual individuals or species currently
living in or around the Hudson River. The actual linkages between the biotic levels often depend
on seasonal availability of various prey and food items. Specific uncertainties in the exposure and
food web modeling are discussed in section 6.3.
The conceptual model used in the ERA Addendum is limited to animals exposed to Lower
Hudson River sediment and water, either directly or via the food chain. Many animals may be
exposed to PCBs from the Hudson River via floodplain soil pathways. These pathways are outside
of the scope of the ERA and ERA Addendum. Inclusion of these pathways would increase the risks
to the mink and raccoon, whose risks were calculated assuming 49.5% and 60% non-river related
diet sources, respectively (see Tables 3-21 and 3-22). In addition, risks for terrestrial species (e.g.,
shrews and moles) exposed to PCBs originating in the Hudson River are outside the scope of the
Reassessment RI/FS and therefore were not quantified, but may be above acceptable levels.
6.2 Toxicological Uncertainties
PCB lexicological studies cover a wide range of test species, doses, exposures, instruments,
and analytical methods. Toxicity can be measured in units of total PCBs, Aroclor mixtures, PCB-
congeners, or normalized toxic equivalency factors. The results of typical lexicological studies can
be reported based on doses by diet, doses per body weighl, and as body burdens, as a lolal PCB
concenlration, or lipid normalized concentralion. The TRVs lhal were selected in this assessment
were based on best-available information and professional judgment. There are other TRVs which
could have been selected which would result in higher or lower loxicily quolienls.
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Aquatic studies are further complicated by various exposure methods. The test species can
be exposed to PCBs via water, sediment, or direct dosing either by food or injection. Given the
insolubility of PCBs, they often partition/adhere to non-aqueous phase materials. Not all studies
consider the effect of sediment or some other matrices (e.g., glass, cotton) on the actual exposure
concentration and availability to test organisms.
Most TRVs are based upon laboratory exposures. Laboratory experiments offer the advantage
of being able to control exposure conditions, while field experiments may be are closer to actual
exposure conditions. Some of the possible reasons for differences between laboratory and field
studies include:
• Laboratory stress on the organisms;
• The lab does not create the actual environmental conditions experienced in the field;
• Contaminant concentration in the water at the study area may be below the instrument
detection limit and therefore will not be reproduced accurately in a laboratory;
• Increases in concentrations along the food chain are not always reflected in the laboratory;
and
• Confounding effects of other environmental contaminants associated with PCBs in the
environmental media.
Furthermore, differences in species sensitivity between laboratory test populations and
endemic populations are often unknown.
There are several uncertainties associated with the toxicological studies that were used to
develop the TRVs for this ERA Addendum. Uncertainty Factors (UFs) may be applied to toxicity
values to address interspecies uncertainty, intraspecies uncertainty, less-than-lifetime at steady state,
acute toxicity to chronic NOAELs, LOAELs to NOAELs, and modifying factors (Calabrese and
Baldwin, 1993).
When toxicological data are not available for specific receptor species, a species-to-species
extrapolation must be made. Generally, the closest taxonomic linked TRY (e.g., species >genus
>family >order >class) is preferred. Extrapolations can be made with a fair degree of certainty
between aquatic species within genera and genera within families (USEPA, 1996). In contrast,
uncertainties associated with extrapolating between orders, classes, and phyla tend to be very high
and are not preferred over more taxonomically similar comparisons (Suter, 1993). Species level
adjustments may be made to address specific developmental or reproductive endpoints or for
application to an endangered species. Under such circumstances, an uncertainty factor (UF) can be
used to account for species to species variation or for accounting for specific sensitive life stages.
A less-than-lifetime UF may be used if the test species is exposed to a contaminant for a
fraction of its lifespan. The purpose of this factor is to ensure that growth, maintenance, and
reproductive functions are accounted for within a protective range of uncertainty. Additional UF
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factors may be added for extrapolating acute toxicity to chronic studies and adapting a LOAEL to
a NOAEL. An additional modifying factor may be added if there are aspects of the TRY study that
are not covered by the other UFs.
Fish TRVs were expressed as a body burden. The pumpkinseed, largemouth bass, white
perch, and striped bass field-based NOAEL TRVs did not require any uncertainty factors. The
laboratory-based TRVs developed for yellow perch and brown bullhead required an interspecies
uncertainty factor of 10. The laboratory-based TRV developed for the spottail shiner required no
uncertainty factor.
For the avian receptors, the tree swallow and kingfisher dietary dose based TRVs required
no uncertainty factors. The dietary dose TRV for the mallard duck, great blue heron, and the bald
eagle all required a factor of 10 uncertainty to account for subchronic to chronic extrapolation. TRVs
developed for the concentration in avian eggs required no uncertainty factors for any avian receptor.
Mammalian receptors all required a factor of 10 uncertainty on a total PCB basis except for
the otter, which required no uncertainty factors. For the raccoon and bat, this value was for
interspecies comparisons. For mink, this value was for extrapolation from a subchronic study to a
chronic value.
There is also uncertainty in the manner in which TEQ concentrations are characterized in the
original studies upon which the TEQ-based TRV was based. Some toxicity studies used slightly
different TEFs when evaluating TEQ concentrations. Where available, a comparison of the
difference in the result between using the TEF reported in the paper as compared to the TEF used
in this analysis was conducted. This difference was no more than 30% and typically on the order of
13% - 20%.
For fish, the selected TRVs were based on egg concentrations in lake trout. Because lake
trout are among the most sensitive species tested, and the concentration was in the egg rather than
an estimated dose, the interspecies and subchronic-to-chronic uncertainty factors were not required.
For the avian receptors, the TEQ-based TRV for the tree swallow was based on Hudson River data
(USFWS), thus, no uncertainty factors were required. The egg-based TRVs for TEQ congeners for
the avian receptors was based on a study in gallinaceous birds, among the most sensitive of
receptors. For this reason, as with fish, no uncertainty factors were required. Dietary dose TRVs for
the avian receptors incorporated a factor of 10 subchronic-to-chronic uncertainty factor. For the
mammals, an uncertainty factor of 10 was applied in deriving the TEQ-based TRV to account for
potential interspecies differences. In conclusion, at most a factor of 10 was applied to the TEQ-based
TRV for mammals and for dietary-dose based TRVs for avian receptors. Fish and avian eggs did not
require any uncertainty factors.
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6.3 Exposure and Modeling Uncertainties
6.3.1 Natural Variation and Parameter Error
Parameter error includes both uncertainty in estimating specific parameters related to
exposure or the specific exposure point concentrations being applied in the exposure models (e.g.,
sediment and water concentrations) as well as variability (e.g., ingestion rate and body weight).
Some parameters can be both uncertain and variable. It is important to distinguish uncertainty from
variability. Variability represents known variations in parameters based on observed heterogeneity
in the characteristics of a particular endpoint species. Variability can be better understood by
collecting additional data, although never eliminated. Uncertainty can be reduced directly through
the confirmation of applied assumptions or inferences through direct measurement. Therefore, it is
theoretically possible to eliminate uncertainty but not variability.
A detailed description of sources of uncertainty and variability in the exposure model
parameters is presented in the baseline ERA (USEPA, 1999c).
6.3.2 Model Error
Model error is the uncertainty associated with how well a model approximates the true
relationships between environmental components (i.e., exposure sources and receptors). Model error
includes: inappropriate selection or aggregation of variables, incorrect functional forms, and
incorrect boundaries (Suter, 1993). This is the most difficult form of uncertainty to evaluate
quantitatively. In the ERA Addendum, model error is not expected to be a significant source of
uncertainty, for the reasons presented below. Relationships between trophic levels and food web
components in the Hudson River are well understood.
6.3.2.1 Uncertainty in the Farley Model
Uncertainty in the application of the Farley et al. (1999) model for the purposes of the ERA
Addendum and the Mid-Hudson HHRA arises from several sources. These sources of uncertainty
can be classified as one of two types: uncertainties which originate from the parameterization of the
model, and uncertainties concerning the assumptions of future conditions in the Hudson.
The uncertainties in model parameterization stem from the uncertainties in the individual
parameter estimates. Because the model is mechanistic, the various parameters are independently
obtained from the literature whenever possible. In this manner, the number of parameters which must
be determined in the calibration is minimized and model uncertainty is minimized. Nonetheless, the
data available for calibration are not sufficient to constrain the model completely and it is possible
that more than one model solution would satisfy all the available constraints. In particular, data on
sediment and water column PCB concentrations are very limited temporally. The more extensive fish
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data set provides an integrating constraint on model parameterization because it requires accuracy
of both the fate and transport and the bioaccumulation models. However, its constraints on the fate
and transport model are indirect and therefore limited. While the model uncertainty originating from
parameterization is not known quantitatively, it is likely to be less than that associated with
estimating future conditions. Indeed, the fact that the model is able to reproduce the general trends
of the existing sediment, water and fish data suggests that the model uncertainty from
parameterization is similar to the scale of the differences between the model calibration and the data
themselves.
The second and probably greater source of uncertainty in the model is inherent in the
assumption of future conditions. In order to estimate future PCB conditions, it is also necessary to
estimate future hydrology, sediment loads, external PCB sources and other concerns. To some
degree, hydrology and sediment loads can be estimated from historical records but the length of the
forecast required adds great uncertainty. In particular, changes in land use, population density and
other societal demands on the watershed are likely to change nature of water and sediment loads to
the Lower Hudson relative to those assumed for the forecast. Similarly, assumptions of future PCB
loads are also difficult to estimate and constrain. As demonstrated by the comparison of the
HUDTOX and original Farley et al. (1999) model loads at the Federal Dam, the loads from the
Upper Hudson have a significant effect on Lower Hudson fish body burdens. Thus, estimation of
external PCB loads such as that at the Federal Dam represent a potentially large source of
uncertainty. The use of HUDTOX model loads at Federal Dam is a direct attempt to minimize the
uncertainty of the Federal Dam load. By using the HUDTOX forecast, loads from the sediments of
the Upper Hudson, currently the most important external source to the Lower Hudson River, are
relatively well constrained. However, the loads originating from the General Electric facilities at
Hudson Falls and Fort Edward, NY remain an important source of long-term uncertainty to both
Upper and Lower Hudson models of PCB contamination.
It is important to note that uncertainties associated with the estimation of future conditions
affects any and all forecast models and is not unique to the models used by the USEPA. The reader
is referred to the original work by Farley et al. (1999) for additional discussion of uncertainty
associated with the Farley et al. (1999) fate and transport and bioaccumulation models.
6.3.2.2 Uncertainty in FISHRAND Model Predictions
A more detailed uncertainty and sensitivity analysis in the FISHRAND model is provided
in the Baseline Modeling Report (USEPA, 1999b). Those results are summarized here.
Two approaches were used to evaluate the impact of small changes in user-specified input
parameters (e.g., lipid content in the organisms, weight of the organisms, water temperature, total
organic carbon, sediment and water concentrations, and Kow) and model constants on predicted fish
body burdens.
In the first approach, a sensitivity analysis was conducted to evaluate the effect of varying
the input parameters using a Monte Carlo methodology. In this method, combinations of values for
the input parameters are generated randomly. Each parameter appears with the frequency suggested
by its probability distribution. For each combination of input parameters, the output of the model
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is recorded. Each individually recorded input parameter is then plotted against the predicted body
burden for that simulation. This is repeated many times to generate plots representing all possible
combinations of input parameters leading to predicted body burdens.
The partial rank and Spearman rank regression techniques (Morgan and Henrion, 1990) are
used as a formal method to find the most important parameters for the model performance. If the
Spearman or partial rank regression coefficient (PRRC or SRRC) is close to 1 or -1 for a specific
input model parameter, this parameter significantly influences model output. The percent lipid in
fish is strongly negatively correlated with PCB body burden expressed on a lipid-normalized basis.
This is because increases in lipid increase the PCB storage capacity of the fish, reducing the apparent
concentration. As expected, the percent lipid in fish is positively associated for the wet weight
results, but less so. This confirms that particularly on a lipid-normalized basis, the percent lipid
distribution is very important. K^ and benthic percent lipid are also important for some species on
a wet weight basis. Feeding preferences are only weakly correlated with body burdens in terms of
sensitivity to this parameter.
To evaluate changes in the model constants themselves, sensitivity to model constants was
evaluated by approximating an analytical solution and then taking partial derivatives of all the model
constants with respect to fish concentration. These partial derivatives were plotted to evaluate
changes in magnitude and sign over time. The assimilation efficiency and growth rate were
determined to be the most important parameters in terms of effect on predicted fish concentration.
The modeling results for this assessment show that the FISHRAND model tends to
underpredict at specific locations and for specific years. On a median basis, FISHRAND does not
overpredict. The FISHRAND calibration focused on optimizing wet weight concentrations, as
described in the Baseline Modeling Report (USEPA, 1999b). This was done for three reasons. First,
the model predicts a wet weight concentration in fish, and provides lipid normalized results by
dividing the predicted wet weight concentration by a percent lipid. Second, the lipid content of any
given fish is difficult to predict from first principles alone. Finally, potential target levels in fish are
typically described as wet weight concentrations.
Optimizing the model for wet weight concentrations provides a reasonable basis upon which
to make forecasts. In addition to forecasting fish responses to changes in sediment and water
concentrations, it is also necessary to predict lipid content. By simply relying on the observed lipid
for each year for which there are data, it is possible to obtain close to perfect agreement between
hindcast and observed body burdens. This approach makes forecasts tenuous, however. Instead, the
FISHRAND model forecasts wet weight concentrations by relying on a distribution of lipid values
in each fish species that is representative of the observed variability in lipid content. This provides
a more robust basis upon which to make predictions.
Focusing specifically on the wet weight results, largemouth bass hindcasts at RM 152 are
within between 60% and 17% less than the observed medians, and fall within the lower bound of the
error bars. This percentage represents 2 or 3 ppm on an absolute basis. At RM 113, hindcast
largemouth bass concentrations of PCBs are between 3% and 50% less than the observed medians.
For the period 1993 to 1996, the error between hindcast and observed is no more than 13%,
representing less than 0.5 ppm PCBs on an absolute basis.
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Brown bullhead concentrations of PCBs are typically within 6% and 30% less than the
observed medians at RM 152, except for 1991. This difference represents less than one ppm on an
absolute basis. White perch FISHRAND hindcasts at RM 152 are within 20% to 65% less than
observed values for 1992 - 1994, but exceed the observed median by 20% for 1996. Hindcast
concentrations of PCBs for 1993 and 1996 fall within the error bars of the observed median. These
values range from less than one ppm to slightly more than a one ppm on an absolute basis. At RM
113, the hindcast white perch concentration in 1994 exceeds the observed median by 100%.
However, for the remaining years, hindcast concentrations of PCBs fall below observed values by
40%, 6%, and 60% for 1993, 1995, and 1996, respectively. For 1996, this difference is 3 ppm PCBs
on an absolute basis. Hindcasts for yellow perch exceed in 1991, but fall below for 1992 and 1993
(50% and 21%, respectively), although for 1993 the hindcast concentration is within the error bounds
of the observed concentration. At RM 113, hindcast yellow perch concentrations of PCB s are 21 %
underpredicted for 1993 (but within the error bounds), and 36% overpredicted for 1994.
6.3.3 Sensitivity Analysis for Risk Models for Avian and Mammalian Receptors
Sensitivity analyses on the exposure and risk models were conducted by specifying
distributions for key parameters. This allows the generation of a distribution of toxicity quotients
to quantitatively evaluate the contribution of key parameters to the variance in the output based on
the inputs. Distributions were described as triangular and were based on the ranges for exposure
parameters presented in detail in Chapter 3 of the baseline ERA (USEPA, 1999c). Environmental
concentrations were described as lognormal by a geometric mean and geometric standard deviation.
Toxicity reference values were described as uniform and typically spanned an order of magnitude
(see discussion above). Results showed that toxicity quotients were most sensitive to changes in
concentrations in exposure media, followed by changes in the toxicity value, and finally by changes
in exposure parameters (e.g., ingestion rates and body weights). These results were consistent for
all avian and mammalian receptors.
The output distributions of toxicity quotients generated by this Monte Carlo analysis
represent population heterogeneity. Results are expressed as the ratio of selected percentiles to the
expected toxicity quotient (based on the average) and show that the 95th percentile of toxicity
quotients is typically 3.5 to 5 times the average, and the 99th percentile of toxicity quotients is
typically at 10 to 15 times the average. Ninety-nine percent of the population is expected to
experience the 99th percentile toxicity quotient or less, and which is estimated as between 10 and
15 times greater than the values shown in the tables for the average. These results were consistent
for both avian and mammalian receptors.
Ratios of the 25th percentile to the average typically range from 0.6 to 0.8 for the avian and
mammalian receptors. This result suggests that even at the 25th percentile, modeled dietary doses
and/or egg concentrations exceed toxicity reference values for most of the receptors (with the
exception of the tree swallow).
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Chapter 7
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7.0 CONCLUSION
This chapter summarizes the results of the ERA Addendum. A summary of the results for
each assessment endpoint is presented. The results of the risk characterization are evaluated in the
context of uncertainties in a weight-of-evidence approach to assess the potential for adverse
reproductive effects in the receptors of concern as a result of exposure to PCBs in the Lower Hudson
River originating in the Upper Hudson River.
7.1 Assessment Endpoint: Benthic Community Structure as a Food Source for
Local Fish and Wildlife
Risks to local benthic invertebrate communities were examined using two lines of evidence.
These lines of evidence are: 1) comparison of modeled water column concentrations of PCBs to
criteria and 2) comparisons of modeled sediment concentrations to guidelines.
Modeled concentrations of PCBs in river water and sediment in the Lower Hudson River
show exceedances of the majority of their respective criteria and guidelines through the duration of
the forecast period (1993 to 2018), indicating the potential for adverse effects on benthic invertebrate
communities.
The uncertainty associated with the application of the Farley et al. (1999) model to estimate
sediment and water concentrations is fairly low. The model is well constrained by the available
sediment, water and fish data. Far greater uncertainty is associated with estimating future forcing
conditions for the model (i.e., external PCB loads, sediment loads and river hydrology). This
uncertainty applies to all such forecasts and is not limited to the Farley et al. (1999) model. It is
likely that the uncertainty in the model forecasts of sediment and water is on the order of a factor of
two.
7.2 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth,
and Reproduction) of Local Fish (Forage, Omnivorous, and Piscivorous)
Populations
Risks to local fish populations were examined using five lines of evidence. These lines of
evidence are: 1) comparison of modeled total PCB fish body burdens to TRVs; 2) comparison of
modeled TEQ fish body burdens to TRVs; 3) comparison of modeled water column concentrations
of PCBs to criteria; 4) comparisons of modeled sediment concentrations to guidelines; and 5) field-
based observations. Multiple receptors were evaluated for forage and semi-piscivorous/piscivorous
fish.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
fish species in the Lower Hudson River. However, based upon toxicity quotients, future exposure
to PCBs may reduce or impair the survival, growth, and reproductive capability of some forage
species (e.g., pumpkinseed), omnivorous fish (e.g., brown bullhead) and semi-piscivorous/piscivorus
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fish (e.g., white perch, yellow perch, largemouth bass, and striped bass), particularly in the upper
reaches of the Lower Hudson River.
There is a moderate degree of uncertainty in the modeled body burdens used to evaluate
exposure, and at most an order of magnitude uncertainty in the TRVs (for the TEQ-based TRVs no
uncertainty factors were needed).
Modeled concentrations of PCBs in river water and sediment in the Lower Hudson River
show exceedances of the majority of their respective criteria and guidelines through the duration of
the forecast period (1993-2018).
7.3 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth,
and Reproduction) of Hudson River Insectivorous Bird Species (as
Represented by the Tree Swallow)
Risks to local insectivorous bird populations were examined using six lines of evidence.
These lines of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs; 2)
comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled total PCB egg
concentrations to TRVs; 4) comparison of modeled TEQ egg concentrations to TRVs; 5) comparison
of modeled water column concentrations of PCBs to criteria; and 6) field-based observations. The
tree swallow was selected to represent insectivorous bird species.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
insectivorous bird species in the Lower Hudson River Valley.
There is a moderate degree of uncertainty in the calculated modeled concentrations of PCBs
in tree swallow diets and the concentrations of PCBs in eggs. There is a low degree of uncertainty
associated with tree swallow TRVs, which were derived from field studies of Hudson River tree
swallows.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993 to 2018).
7.4 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth
and Reproduction) of Lower Hudson River Waterfowl (as Represented by
the Mallard)
Risks to local waterfowl populations were examined using six lines of evidence. These lines
of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs; 2) comparison of
modeled TEQ dietary doses to TRVs; 3) comparison of modeled total PCB egg concentrations to
TRVs; 4) comparison of modeled TEQ egg concentrations to TRVs; 5) comparison of modeled
water column concentrations of PCBs to criteria; and 6) field-based observations. The mallard was
selected to represent waterfowl.
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Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
waterfowl in the Lower Hudson River Valley. However, based upon toxicity quotients, future
exposure to PCBs may reduce or impair the survival, growth, and reproductive capability of some
waterfowl, particularly in the upper reaches of the Lower Hudson River.
Calculated dietary doses of PCBs and concentrations of PCBs in eggs typically exceed their
respective TRVs throughout the modeling period. Toxicity quotients for the TEQ-based (i.e., dioxin-
like) PCBs consistently show greater exceedances than for total (Tri+) PCBs. There is a moderate
degree of uncertainty in the dietary dose and egg concentrations estimates. Given the magnitude of
the TEQ-based TQs, they would have to decrease by an order of magnitude or more to fall below one
for waterfowl in the Lower Hudson River.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993 to 2018).
7.5 Assessment Endpoint: Protection and Maintenance (Le., Survival, Growth,
and Reproduction) of Hudson River Piscivorous Bird Species (as
Represented by the Belted Kingfisher, Great Blue Heron, and Bald Eagle)
Risks to local semi-piscivorous/piscivorous bird populations were examined using six lines
of evidence. These lines of evidence are: 1) comparison of modeled total PCB dietary doses to
TRVs; 2) comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled total PCB
egg concentrations to TRVs; 4) comparison of modeled TEQ egg concentrations to TRVs; 5)
comparison of modeled water column concentrations of PCBs to criteria; and 6) field-based
observations. The belted kingfisher, great blue heron, and bald eagle were selected to represent
piscivorous birds.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of these
piscivorous species. However, based upon toxicity quotients, future exposure to PCBs may reduce
or impair the survival, growth, and reproductive capability of some piscivorous birds, particularly
in the upper reaches of the Lower Hudson River. Calculated dietary doses of PCBs and
concentrations of PCBs in eggs exceed all TRVs (i.e., NOAELs and LOAELs) for the belted
kingfisher and bald eagle throughout the modeling period, and NOAELs for the great blue heron.
Toxicity quotients for egg concentrations are generally higher than body burden TQs.
There is a moderate degree of uncertainty in the dietary dose and egg concentrations
estimates. Given the magnitude of the TQs, they would have to decrease by an order of magnitude
or more to fall below one for piscivorous birds in the Lower Hudson River. In particular, the bald
eagle TQs exceeded one by up to three orders of magnitude. Therefore, even if the factor of 2.5 to
adjust from largemouth bass fillets to whole body burden and the subchronic-to-chronic uncertainty
factor of 10 used for the body burden TRV are removed, the TQs would remain well over one.
These results, coupled with the lack of breeding success in Lower Hudson River bald eagles (USGS,
1999), indicate that reproductive effects may be present.
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Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993 to 2018).
7.6 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Insectivorous Mammais (as represented by the Little Brown Bat)
Risks to local insectivorous mammal populations were examined using four lines of
evidence. These lines of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs;
2) comparison of modeled TEQ mammal dietary doses to TRVs; 3) comparison of modeled water
column concentrations of PCBs to criteria; and 4) field-based observations. The little brown bat was
selected to represent insectivorous mammals.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
insectivorous mammals in the Lower Hudson River Valley. However, exposure to PCBs may reduce
or impair the survival, growth, or reproduction capability of insectivorous mammals in the Lower
Hudson River. Modeled dietary doses for the little brown bat exceed TRVs by up to two orders of
magnitude at all locations modeled. There is a moderate degree of uncertainty in the calculated
dietary doses.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993 to 2018).
7.7 Assessment Endpoint: Protection (i. e., Survival and Reproduction) of Local
Omnivorous Mammals (as represented by the Raccoon)
Risks to local omnivorous mammal populations were examined using four lines of evidence.
These lines of evidence are: 1) comparison of modeled total PCB dietary doses to TRVs; 2)
comparison of modeled TEQ dietary doses to TRVs; 3) comparison of water column concentrations
of PCBs to criteria; and 4) field-based observations. The raccoon was selected to represent
omnivorous mammals.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of common
omnivorous mammals in the Lower Hudson River Valley. However, exposure to PCBs may reduce
or impair the survival, growth, or reproduction capability of omnivorous mammals in the Lower
Hudson River. Modeled dietary doses for the raccoon exceed dietary dose NOAELs on a total PCB
(Tri+) basis and all TRVs on a TEQ-basis. There is a moderate degree of uncertainty in the
calculated dietary doses.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993 to 2018).
78 TAMS/MCA
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7.8 Assessment Endpoint: Protection (i. e., Survival and Reproduction) of Local
Piscivorous Mammals (as represented by the Mink and River Otter)
Risks to local semi-piscivorous/piscivorous mammal populations were examined using four
lines of evidence. These lines of evidence are: 1) comparison of modeled total PCB dietary doses
to TRVs; 2) comparison of modeled TEQ dietary doses to TRVs; 3) comparison of modeled water
column concentrations of PCBs to criteria; and 4) field-based observations. The mink and river otter
were selected to represent piscivorous mammals.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of these
piscivorous species. However, based upon toxicity quotients, future exposure to PCBs may reduce
or impair the survival, growth, and reproductive capability of piscivorous mammals, particularly in
the upper reaches of the Lower Hudson River. Calculated dietary doses of F'CBs exceed the NOAEL
on a total PCB basis for both species and exceed all TEQ-based TRVs by up to three orders of
magnitude.
There is a moderate degree of uncertainty in the dietary dose estimates. However, given the
magnitude of the TQs, they would have to decrease at least an order of magnitude to fall below one.
In particular, the river otter TQs exceeded one by up to three orders of magnitude. Therefore, even
if the factor of 2.5 to adjust from largemouth bass fillets to whole body burden is removed, the TQs
would remain well over one. Preliminary results from a NYSDEC study indicate that PCBs may
have an adverse effect on the litter size and possibly kit survival of river otter in the Hudson River
(Mayack, 1999b), validating the TQ results.
Modeled concentrations of PCBs in river water in the Lower Hudson River show
exceedances of criteria developed for the protection of wildlife through the duration of the forecast
period (1993 to 2018).
7.9 Assessment Endpoint: Protection of Threatened and Endangered Species
Risks to threatened and endangered species were examined using five lines of evidence.
These lines of evidence are: 1) comparison of modeled total PCB dietary doses/egg concentrations
to TRVs; 2) comparison of modeled TEQ dietary doses/egg concentrations to TRVs; 3) comparison
of modeled water column concentrations of PCBs to criteria; 4) comparison of modeled sediment
concentrations of PCBs to guidelines; and 5) field-based observations. The shortnose sturgeon and
bald eagle were selected to represent threatened and endangered species.
Collectively, the evidence indicates that future PCB exposures (predicted from 1993 to 2018)
are not expected to be of a sufficient magnitude to prevent reproduction or recruitment of threatened
or endangered species. However, using the TEQ-based toxicity quotients, potential for adverse
reproductive effects in shortnose sturgeon exists, particularly when considering the long life
expectancy of the sturgeon (30 years, [Bain, 1997]). Almost all TQs calculated for the bald eagle
(across all locations) exceeded one, in some instances by more than three orders of magnitude. Both
the dietary dose and egg-based results were consistent in this regard. Other threatened or endangered
79 TAMS/MCA
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raptors, such as the peregrine falcon, osprey, northern harrier, and red-shouldered hawk may
experience similar exposures.
There is a moderate degree of uncertainty in the dietary dose estimates. However, the bald
eagle TQs exceeded one by up to three orders of magnitude. Therefore, even if the factor of 2.5 to
adjust from largemouth bass fillets to whole body burden and the subchronic-to-chronic uncertainty
factor of 10 used for the body burden TRV are removed, the TQs would remain well over one.
These results, coupled with the lack of breeding success in Lower Hudson River bald eagles (USGS,
1999), indicate that reproductive effects may be present.
Modeled concentrations of PCBs in river water and sediment in the Lower Hudson River
show exceedances of the majority of their respective criteria and guidelines through the duration of
the forecast period (1993 to 2018).
7.10 Assessment Endpoint: Protection of Significant Habitats
Risks to significant habitats were examined using four lines of evidence. These lines of
evidence are: 1) toxicity quotients calculated for receptors in this assessment; 2) comparison of
modeled water column concentrations of PCBs to criteria; 3) comparison of modeled sediment
concentrations of PCBs to guidelines; and 4) field-based observations.
Based on the toxicity quotients calculated in ERA Addendum, future PCB concentrations
(predicted from 1993 to 2018) in the Lower Hudson River exceed toxicity reference values for some
fish, avian, and mammalian receptors. These comparisons indicate that animals feeding on Lower
Hudson River-based prey may be affected by the concentrations of PCBs found in the river on both
a total PCB and TEQ basis. In addition, based on the TQs, other taxononic groups not directly
addressed in the ERA and ERA Addendum (e.g., amphibians and reptiles) may also be affected by
PCBs in the river. Many year-round and migrant species use the significant habitats along the
Hudson River for breeding or rearing their young. Therefore, exposure to PCBs may occur at a
sensitive time in the life cycle (i.e., reproductive and development) and have a greater effect on
populations than at other times of the year.
Modeled concentrations of PCBs in river water and sediment in the Lower Hudson River
show exceedances of the majority of their respective criteria and guidelines through the duration of
the forecast period (1993 to 2018).
7.11 Summary
The results of the ERA Addendum indicate that receptors in close contact with the Lower
Hudson River may experience adverse effects as a result of exposure to PCBs in prey, water, and
sediments. Higher trophic level receptors, such as the bald eagle and the river otter, are considered
to be particularly at risk. Risks are generally highest up river (i.e., closer to the PCB source) and
decrease in relation to PCB concentrations down river. Based on modeled PCB concentrations, many
species are expected to be at considerable risk through the entire forecast period (1993 to 2018).
80 TAMS/MCA
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Wiemeyer, S.N., T.G. Lament, C.M. Bunck, C.R. Sindelar, F.J. Gramlich, J.D. Fraser, and M.A.
Byrd. Organochlorine pesticide, polychlorobiphenyl, and mercury residues in bald eagle eggs-1969-
79-and their relations to shell thinning and reproduction. Archives of Environmental Contamination
and Toxicology. 13:529-549.
TAMS/MCA
-------
Tables
-------
TABLE 2-1
LOWER HUDSON ASSESSMENT ENDPOINTS, RECEPTORS, AND MEASURES
Assessment Endpoint
Benthic aquatic life as a food source for
local fish and wildlife.
Survival, growth, and reproduction of
local forage fish populations.
Survival, growth, and reproduction of
local piscivorous/semi-piscivorous fish .
populations.
Survival, growth, and reproduction of
local omnivorous fish populations.
Protection (i.e., survival and
reproduction) of insectivorous birds and
mammals.
Protection (i.e., survival and
reproduction) of waterfowl.
Protection of piscivorous/semi-
piscivorous birds and mammals.
Protection of omnivorous mammals.
Protection of endangered and threatened
species.
Protection of significant habitats.
Specific Ecological
Receptor
("Endpoint Species")
• Benthic aquatic community
• Spottail shiner
• Pumpkinseed
• Yellow perch
• White perch
• Largemouth bass
• Striped bass
• Shortnose sturgeon
• Brown bullhead
• Tree swallow
• Little brown bat
• Mallard
• Belted kingfisher
• Great blue heron
•Mink
• River Otter
• Raccoon
• Bald eagle
• Shortnose sturgeon
• Hudson River NERR
• NYSDOS significant
habitats
Measures
Exposure
• Modeled PCB concentrations in sediments
and water column
• Modeled PCB body burdens
• Modeled PCB concentrations in sediments
and water column
• Modeled PCB body burdens
• Modeled PCB concentrations in sediments
and water column
• Modeled PCB body burdens
• Modeled PCB concentrations in sediments
and water column
• Modeled PCB concentrations in prey
items (aquatic insects)
• Modeled PCB concentrations in the water
column
• Modeled PCB concentrations in prey
(invertebrates, macrophytes)
• Modeled PCB concentrations in the water
column
• Modeled PCB concentrations in prey
(forage fish, invertebrates)
• Modeled PCB concentrations in sediments
and water column
• Modeled PCB concentrations in prey
items (fish, invertebrates)
• Modeled PCB concentrations in the water
column
• Modeled PCB body burdens (sturgeon)
• Modeled PCB concentrations in prey
(fish)
• Modeled PCB concentrations in sediments
and water column
• Modeled PCB concentrations in sediments
and water column
Effect
• Exceedance of AWQC and sediment
guidelines
• Estimated exceedance of TRVs
• Exceedance of AWQC and sediment
guidelines
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC and sediment
guidelines
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC and sediment
guidelines
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC for the protection
of wildlife
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC for the protection
of wildlife
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC for the protection of
wildlife
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC for the protection
of wildlife
• Field observations
• Estimated exceedance of TRVs
• Exceedance of AWQC and sediment
guidelines for the protection of wildlife
• Field observations
• Exceedance of federal and state AWQC
and sediment guidelines
• Field observations
Notes: Individual-level effects are considered to occur when the TQ is greater to or equal to one.
Receptor species are surrogates chosen to represent a wide range of species likely to use the Hudson River as habitat or foraging source.
-------
TABLE 2-2
LOWER HUDSON RIVER ENDPODSTS AND RISK HYPOTHESES
Assessment Endpoint: Bentbic aquatic life as a food source for local fish and wildlife
Do modeled total PCB -water concentrations exceed
criteria and/or guidelines for protection of aquatic
health?
Do modeled total PCB sediment concentrations exceed
guidelines for protection of aquatic health?
Measurement Endpoint I: Modeled PCB concentrations in
water (freshwater) compared to NYS Ambient Water Quality
Criteria (AWQC) for die protection of benthic aquatic life
1NYSDEC, 1998b).
Measurement Endpoint 2: Modeled PCB concentrations in
sediment compared to applicable sediment benchmarks (e.g.,
NOAA Sediment Effect Concentrations for PCBs in the
Hudson River [NOAA, 1999a], NYSDEC Technical
Guidance for Screening Contaminated Sediments [1999a],
etc.)
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Lower Hudson River Fish
Populations (forage, omnivorous, piscivorous)
Do modeled total PCB body burdens in local fish exceed
benchmarks for adverse effects on fish reproduction?
Do modeled total PCB body burdens in local fish
expressed on a TEQ basis exceed benchmarks for
adverse effects on fish reproduction?
Do modeled total PCB water concentrations exceed
criteria and/or guidelines for protection of aquatic
health?
Do modeled total PCB sediment concentrations exceed
guidelines for protection of aquatic health?
What do available field-based observations suggest
about the health of local fish populations?
Measurement Endpoint 1: Modeled total PCB body burdens
in fish for each river segment over 25 years to determine
exceedance of effect-level thresholds based on toxicity
reference values (TRVs) derived in the baseline ERA
(USEPA, 1999c).
Measurement Endpoint 2: Modeled TEQ-based PCB body
burdens in fish for each river segment over 25 years to
determine exceedance of effect-level thresholds based on
TRVs.
Measurement Endpoint 3: Modeled PCB concentrations in
water (freshwater) compared to NYS Ambient Water Quality
Criteria (AWQC) for the protection of benthic aquatic life
(NYSDEC, 1998b).
Measurement Endpoint 4: Modeled PCB concentrations in
sediment compared to applicable sediment benchmarks (e.g.,
NOAA Sediment Effect Concentrations for PCBs in the
Hudson River [NOAA, 1999a], NYSDEC Technical
Guidance for Screening Contaminated Sediments [1999a],
etc.)
Measurement Endpoint 5: Available field observations on
the presence and relative abundance of fish species within
the Lower Hudson River as an indication of the ability of the
species to maintain populations.
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Lower Hudson River
Insectivorous Bird Populations (represented by the tree swallow)
Do modeled total PCB dietary doses to insectivorous
exceed benchmarks for adverse effects on reproduction?
Do modeled TEQ-based dietary doses of PCBs to
insectivorous birds exceed benchmarks for adverse
effects on reproduction?
Do modeled total PCB concentrations in insectivorous
bird eggs exceed benchmarks for adverse effects on
reproduction?
Do modeled TEQ-based PCB concentrations in
insectivorous bird eggs exceed benchmarks for adverse
effects on reproduction ?
Measurement Endpoint I: Modeled total PCB body burdens
in the tree swallow to determine exceedance of effect-level
thresholds based on TRVs.
Measurement Endpoint 2: Modeled TEQ-based PCB body
burdens in the tree swallow to determine exceedance of
effect-level thresholds based on TRVs.
Measurement Endpoint 3: Modeled total PCB egg
concentrations in the tree swallow to determine exceedance
of effect-level thresholds based on TRVs.
Measurement Endpoint 4: Modeled TEQ-based PCB egg
concentrations in the trees swallow to determine exceedance
of effect-level thresholds based on TRVs.
-------
TABLE 2-2
LOWER HUDSON RIVER ENDPOINTS AND RISK HYPOTHESES
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
What do the available field-based observations suggest
about the health of local insectivorous bird
populations?
Measurement Endpoint 5: Modeled PCB concentrations in
water (freshwater) compared to NYS AWQC for the
protection of wildlife (NYSDEC, 1998b).
Measurement Endpoint 6: Available field observations on
the presence and relative abundance of insectivorous bird
species within the Lower Hudson River as an indication of
the ability of the species to maintain populations.
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Lower Hudson River
Waterfowl Populations (represented by the mallard)
Do modeled total PCB dietary doses to waterfowl
exceed benchmarks for adverse effects on reproduction?
Do modeled TEQ-based dietary doses of PCBs to
waterfowl exceed benchmarks for adverse effects on
reproduction ?
Do modeled total PCB concentrations in insectivorous
bird eggs exceed benchmarks for adverse effects on
reproduction ?
Do modeled TEQ-based PCB concentrations in
waterfowl eggs exceed benchmarks for adverse effects
on reproduction?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
What do the available field-based observations suggest
about the health of local waterfowl populations?
Measurement Endpoint 1: Modeled total PCB body burdens
in the mallard to determine exceedance of effect-level
thresholds based on TRVs.
Measurement Endpoint 2: Modeled TEQ-based PCB body
burdens in the mallard to determine exceedance of
effect-level thresholds based on TRVs.
Measurement Endpoint 3: Modeled total PCB egg
concentrations in the tree swallow to determine exceedance
of effect-level thresholds based on TRVs.
Measurement Endpoint 4: Modeled TEQ-based PCB egg
concentrations in the mallard to determine exceedance of
effect-level thresholds based on TRVs.
Measurement Endpoint 5: Modeled PCB concentrations in
water (freshwater) compared to NYS AWQC for the
protection of wildlife (NYSDIiC, 1998b).
Measurement Endpoint 6: Available field observations on
the presence and relative abundance of waterfowl along the
Lower Hudson River as an indication of the ability of the
species to maintain populations.
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Hudson River Piscivorous Bird
Populations (represented by the bald eagle, great blue heron, and belted kingfisher)
Do modeled total PCB dietary doses to piscivorous
birds exceed benchmarks for adverse effects on
reproduction?
Do modeled TEQ-based dietary doses of PCBs to
piscivorous birds exceed benchmarks for adverse effects
on reproduction ?
Do modeled total PCB concentrations in piscivorous
bird eggs exceed benchmarks for adverse effects on
reproduction?
Do modeled TEQ-based PCB concentrations in
piscivorous bird eggs exceed benchmarks for adverse
effects on reproduction?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
What do the available field-based observations suggest
about the health of local piscivorous bird populations?
Measurement Endpoint 1: Modeled total PCB body burdens
in receptor species (i.e., bald eagle, great blue heron, and
belted kingfisher) over 25 years to determine exceedance of
effect-level thresholds based on TRVs.
Measurement Endpoint 2: Modeled TEQ-based PCB body
burdens in receptor species for each river segment over 25
years to determine exceedance of effect-level thresholds
based on TRVs.
Measurement Endpoint 3: Modeled total PCB egg
concentrations in receptor species to determine exceedance
of effect-level thresholds based on TRVs.
Measurement Endpoint 4: Modeled TEQ-based PCB egg
concentrations in receptor species to determine exceedance
of effect-level thresholds based on TRVs.
Measurement Endpoint 5: Modeled PCB concentrations in
water (freshwater and saline) compared to NYS AWQC for
the protection of wildlife (NYSDEC, 1998b).
Measurement Endpoint6: Available field observations on
the presence and relative abundance of piscivorous birds
along the Lower Hudson River as an indication of the ability
of the species to maintain populations.
-------
TABLE 2-2
LOWER HUDSON RIVER ENDPOINTS AND RISK HYPOTHESES
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Lower Hudson River
Insectivorous Mammals (as represented by the little brown bat)
Do modeled total PCB dietary doses to local wildlife
species exceed benchmarks for adverse effects on
reproduction ?
Do modeled TEQ-based PCB dietary doses to local
wildlife species exceed benchmarks for adverse effects
on reproduction?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
What do the available field-based observations suggest
about the health of local wildlife populations?
Measurement Endpoint 1: Modeled total PCB body burdens
in the wildlife species to determine exceedance of
effect-levels based on TRVs.
Measurement Endpoint 2: Measured and modeled
TEQ-based PCB body burdens in the little brown bat to
determine exceedance of effect-level thresholds based on
TRVs.
Measurement Endpoint 3: Modeled PCB concentrations in
water (freshwater and saline) compared to NYS AWQC for
the protection of wildlife (NYSDEC, 1999a).
Measurement Endpoint 4: Available field observations on
the presence and relative abundance of insectivorous species
along the Lower Hudson River as an indication of the ability
of the species to maintain populations.
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Hudson River Omnivorous
Mammals (as represented by the raccoon)
Do modeled total PCB dietary doses to local wildlife
species exceed benchmarks for adverse effects on
reproduction ?
Do modeled TEQ-based PCB dietary doses to local
wildlife species exceed benchmarks for adverse effects
on reproduction?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife ?
What do the available field-based observations suggest
about the health of local wildlife populations?
Measurement Endpoint 1: Modeled total PCB body burdens
in the raccoon to determine exceedance of effect-levels
based on TRVs.
Measurement Endpoint 2: Measured and modeled TEQ-
based PCB body burdens in the raccoon to determine
exceedance of effect-level thresholds based on TRVs.
Measurement Endpoint 3: Modeled PCB concentrations in
water (freshwater and saline) compared to NYS AWQC for
the protection of wildlife (NYSDEC, 1999a).
Measurement Endpoint 4: Available field observations on
the presence and relative abundance of omnivorous
mammals along the Lower Hudson River as an indication of
the ability of the species to maintain populations.
Assessment Endpoint: Sustainability (i.e., survival, growth, and reproduction) of Lower Hudson River
Piscivorous Wildlife (as represented by the mink and river otter)
Do modeled total PCB dietary doses to local wildlife
species exceed benchmarks for adverse effects on
reproduction ?
Do modeled TEQ-based PCB dietary doses to local
wildlife species exceed benchmarks for adverse effects
on reproduction ?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
What do the available field-based observations suggest
about the health of local wildlife populations?
Measurement Endpoint 1: Modeled total PCB body burdens
in the wildlife species to determine exceedance of
effect-levels based on TRVs.
Measurement Endpoint 2: Measured and modeled
TEQ-based PCB body burdens in the wildlife species for
each river segment over 25 years to determine exceedance
of effect-level thresholds based on TRVs.
Measurement Endpoint 3: Modeled PCB concentrations in
water (freshwater and saline) compared to NYS AWQC for
the protection of wildlife (NYSDEC, 1999a).
Measurement Endpoint 4: Available field observations on
the presence and relative abundance of the wildlife species
along the Hudson River as an indication of the ability of the
species to maintain populations.
Assessment Endpoint: Protection of Threatened and Endangered Species
Do modeled total PCB body burdens in local threatened
or endangered species exceed benchmarks for adverse
effects on reproduction ?
Measurement Endpoint 1 : Modeled total PCB body burdens
in shortnose sturgeon (using surrogate upper trophic level
fish species) and the bald eagle to determine exceedance of
effect-level thresholds based on TRVs.
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TABLE 2-2
LOWER HUDSON RIVER ENDPOINTS AND RISK HYPOTHESES
Do modeled TEQ-based PCB body burdens in local
threatened or endangered species exceed benchmarks
for adverse effects on reproduction?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
Do modeled sediment PCB concentrations exceed
guidelines for the protection of aquatic health?
What do the available field-based observations suggest
about the health of local wildlife populations?
Measurement Endpoint 2: Modeled TEQ-based PCB body
burdens in shortnose sturgeon (using surrogate upper trophic
level fish species) and the bald eagle to determine
exceedance of effect-level thresholds based on TRVs.
Measurement Endpoint 3: Modeled PCB concentrations in
water (freshwater and saline) compared to NYS AWQC for
the protection of wildlife (NYSDEC, 1998b).
Measurement Endpoint 4: Modeled PCB concentrations in
sediment compared to applicable sediment benchmarks (e.g.,
NOAA, 1999a, NYSDEC 1999, etc.)
Measurement Endpoint 5: Available field observations on
the presence and relative abundance of threatened and
endangered species along the Lower Hudson River as an
indication of the ability of the species to maintain
populations.
Assessment Endpoint: Protection of Significant Habitats
Do modeled toxicity quotients in local receptor species
exceed benchmarks for adverse effects on reproduction?
Do modeled whole water concentrations exceed criteria
and/or guidelines for the protection of wildlife?
Do modeled sediment PCB concentrations exceed
guidelines for the protection of aquatic health?
What do the available field-based observations suggest
about the health of local wildlife populations?
Measurement Endpoint 1: Modeled total PCB and
TEQ-based PCB body burdens in receptor species to
determine exceedance of effect-level thresholds based on
TRVs.
Measurement Endpoint 2: Modeled PCB concentrations in
water (freshwater and saline) compared to NYS AWQC for
the protection of benthic aquatic life (NYSDEC, 1998b) or
wildlife (NYSDEC, I998b).
Measurement Endpoint 3: Modeled PCB concentrations in
sediment compared to applicable sediment benchmarks (e.g.,
NOAA, 1999a, NYSDEC 1999a, etc.).
Measurement Endpoint 4: Available field observations on
the presence and relative abundance of the wildlife species
using significant habitats along the Hudson River as an
indication of the ability of the habitat to maintain
populations.
Note: Effect level-concentrations are measured by TRVs. Toxicity quotients are exceeded when the modeled dose or
concentration is greater than the benchmark dose or concentration (i.e., toxicity quotient [TQ] exceeds 1 ). Calculation
of the modeled dose and selection of the benchmark dose are covered in the baseline ERA (USEPA, 1999c).
-------
TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
County
Community Types
Rare Species
Valuable Species
Freshwater Habitats
Normans Kill
Albany
Freshwater creek with
shallows associated with
creek mouth.
None identified.
Spawning area for anadromous fish species
including alewife, white perch, and blueback
herring. Large resident smallmouth bass
populations.
Shad and Schermerhorn
Island
Albany
Largely comprised of
shallows and mudflats with
lesser amounts of lower
marsh, upper marsh and
freshwater creek.
Heart leaf plantain and
estuary beggar ticks.
Large feeding areas for herons and other wading
birds, furbearers, deer and other upland game,
limited waterfowl usage, important spawning and
nursery grounds for American shad, blueback
herring, alewife, white perch, striped bass, and
resident fish species.
Papascanee Marsh and
Creek
Renssalear
Mainly upper marsh with
lesser amounts of shallows,
mudflats, lower marsh, and
freshwater creek.
Least bittern nesting area;
map turtles.
Waterfowl use during migrations. Breeding birds
incl. green-backed heron, Virginia rail, several
duck species, marsh wren, swamp swallow, and
others. Spawning and nursery grounds for
American shad, blueback herring, alewife, white
catfish, black bass, white perch and other fish.
Schodack and
Houghtaling Islands and
Schodack Creek
Renssalear,
Columbia,
Greene
Predominantly shallows,
mudflats, and sandy beach
with lesser amounts of lower
marsh and upper marsh.
Osprey roosting and
feeding; possible use by
shortnose sturgeon;
heart leaf plantain.
Waterfowl use during migrations and limited
nesting activity, nesting by other bird species.
Furbearers present. Schodack Creek provides
important spawning and nursery grounds for
American Shad, white perch, alewife, and blueback
herring, black bass and other species. Northmost
concentration of shad spawning on the Hudson.
1 of 9
TAMS/MCA
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TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
Coeymans Creek
Hannacroix Creek
Mill Creek Wetlands
Stuyvesant Marshes*
Coxsackie Creek
County
Albany
Albany,
Greene
Columbia
Columbia
Greene
Community Types
Predominantly shallows with
smaller amounts of mudflats,
lower marsh, and swamp
forest.
Predominantly freshwater
creek with shallows,
mudflats, lower marsh, upper
marsh and swamp forest.
Swamp forest with some
shallows, mudflats, sandy
beach, lower marsh, and
upper marsh.
Roughly equal amounts of
shallows, mudflats, sandy
mudflats, sandy beach, rocky
shore, lower marsh, and
upper marsh.
Principally freshwater creek
with some shallows,
mudflats, sandy beach, lower
marsh, upper marsh, and
freshwater creek.
Rare Species
None.
None identified.
Estuary beggar ticks.
Heart leaf plantain, kidney
leaf mud plantain.
Estuary beggar ticks.
Valuable Species
Important spawning area for anadromous fish
including alewife, blueback herring, white perch,
and American Shad. Limited waterfowl during
migrations.
Important spawning area for alewife, blueback
herring, white perch, American Shad, and other
fish. Resting and feeding area for migratory
waterfowl. Feeding area for herons, various birds,
and furbearers.
Limited waterfowl use during migrations.
Populations of breeding birds include green-backed
herons, various ducks, and many passerines.
Limited use by migrating waterfowl, probable
heavy use by various nesting bird species.
Spawning habitat for alewife, blueback herring,
white perch, and American shad. Feeding grounds
for herons and other wading birds. Small mammal
and furbearer foraging.
2 of 9
TAMS/MCA
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TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
County
Community Types
Rare Species
Valuable Species
Coxsackie Island
Backwater
Greene
Shallows with peripheral
mud and sand flats, rocky
shore, lower marsh, and
upper marsh.
Heart leaf plantain, kidney
leaf mud plantain.
Important spawning and nursery ground for
resident fish including brown bullhead, largemouth
bass, yellow perch, and redfin pickerel. Also
feeding grounds for anadromous fish and wintering
areas for largemouth bass.
Stockport Creek and
Flats
Columbia
Shallows and mudflats with
substantial areas of lower
marsh, upper marsh, and
woody swamp. Three miles
of tidal and freshwater creek.
Some deepwater and sandy
beach associated with
navigation channel and
islands.
Heart leaf plantain, estuary
beggar ticks, golden club;
map turtle.
Very important spawning/nursery grounds for
anadromous and freshwater fish including alewife,
blueback herring, smelt, American shad, striped
bass, and smallmouth bass. Very important feeding
and resting habitat for migrating and overwintering
waterfowl. Use by wading, shore, and passerine
birds for feeding and breeding. Bank swallows nest
in the vertical sand banks. Extensive stands of wild
rice.
Vosburgh Swamp and
Middle Ground Flats
Greene
Largely comprised of creek,
deepwater, shallows, and
mudflats with lesser amounts
of sandy beach, lower marsh,
upper marsh, and freshwater
swamp.
Possible least bittern and
mud turtle; heart leaf
plantain, sublate
arrowhead, estuary beggar
ticks.
Important feeding and resting grounds for
migrating waterfowl and wintering waterfowl
(when open water is available). Extensive nesting
area for ducks, green-backed herons, and other
birds. Colony of bank swallows. Heavy use of
shallows for American shad spawning and
extensive spawning, nursery and feeding areas for
striped bass, alewife, blueback herring and resident
fish species.
Roger's Island
Columbia
Comprised of roughly equal
amounts of shallows and
mudflats with some sandy
beach, lower marsh, upper
marsh, and swamp forest.
Estuary beggar-ticks,
goldenclub.
Extensive waterfowl use during migrations and
overwintering, nesting sites for many birds,
extensive spawning areas for anadromous fish
including the American shad.
3 of 9
TAMS/MCA
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TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
County
Community Types
Rare Species
Valuable Species
Catskill Creek
Greene
Predominantly creek with
small amounts of shallows,
mudflats, and lower marsh.
Wood turtle, probably in
association with buffer
area.
Important spawning and nursery grounds for
anadromous and resident fishes including American
shad, alewife, blueback herring, white perch,
smallmouth and largemouth bass.
Ramshorn Marsh
Greene
Largely shallows, mudflats,
lower marsh, upper marsh,
and swamp forest with lesser
amounts of sandy beach and
rocky shore.
Least bittern nesting;
estuary beggar-ticks, and
heart leaf plantain.
Waterfowl use during migrations and
overwintering, important heron feeding grounds,
furbearer habitat, spawning and nursery grounds for
American shad and black bass.
Inbocht Bay and Duck
Cove
Greene
Principally shallows and
mudflats with some lower
marsh.
Estuary beggar-ticks.
Very extensive waterfowl concentrations during
spring and fall migrations, some waterfowl
overwintering, large muskrat and snapping turtle
populations.
Roeliff-Jansen Kill
Columbia
Predominantly freshwater
creek with limited shallows,
mudflats, and lower marsh.
None identified.
Extensive use as a spawning/nursery ground for
anadromous fish including American shad,
blueback herring, white perch, and striped bass.
Resident brown trout in upper reaches.
Smith's Landing
Cementon*
Greene,
Ulster
Limited mudflats, lower
marsh, and upper marsh.
Heart leaf plantain, kidney
leaf mud-plantain.
None identified.
Germantown/Clermont
Flats
Columbia
Deepwater, shallows,
mudflats, and limited lower
marsh.
None identified.
Extremely important American shad spawning area,
nursery areas for shad, striped bass, white perch,
and resident fish. Extensive waterfowl feeding
grounds during spring and fall migration periods.
Some waterfowl overwintering.
4 of 9
TAMS/MCA
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TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
County
Community Types
Rare Species
Valuable Species
Esopus Estuary
Ulster,
Dutchess
Comprised of freshwater
creek, deepwater, shallows,
mudflats, lower marsh, upper
marsh, and a small amount of
tidal swamp.
Shortnose sturgeon
spawning and wintering
area in deepwater;
migrating osprey feeding
grounds; heart leaf
plantain, goldenclub.
Important spawning and nursery grounds for
striped bass, white perch, American shad, alewife,
blueback herring, rainbow smelt, and resident fish.
Feeding and resting grounds for migrating
waterfowl.
North and South Tivoli
Bays
Dutchess
Comprised of shallows,
lower marsh, and upper
marsh, followed by tidal
swamp forest, rocky shore
and creeks.
Migrating osprey feeding
and resting, least bittern
nesting, king rail; map
turtles; heart leaf plantain,
estuary beggar-ticks,
goldenclub and other rare
plants.
Feeding, spawning and/or nursery areas for striped
bass, alewife, blueback herring, largemouth and
smallmouth bass, and other fishes. Large snapping
turtle population. Extensive waterfowl use for
feeding and resting during migrations. Many
breeding birds. Furbearer habitat.
Mudder Kill*
Dutchess
Equal amounts of mudflats,
lower marsh, upper marsh,
and tidal swamp forest.
Goldenclub, hirsute sedge,
Davis sedge, heavy sedge,
kidney leaf mud-plantain,
and spongy arrowhead.
None known.
The Flats
Ulster,
Dutchess
Comprised entirely of
shallows.
Potential shortnose
sturgeon feeding and
resting area.
Primary spawning grounds for American shad and
spawning and nursery area for striped bass, white
perch, and resident fishes. Feeding area during
migration periods for diving ducks and resting
areas for all duck species.
5 of 9
TAMS/MCA
-------
TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
Roundout Creek
Kingston Deepwater
Habitat
Vanderburgh Cove and
Shallows
Esopus Meadows
Poughkeepsie
Deepwater Habitat
County
Ulster
Dutchess,
Ulster
Dutchess
Ulster
Dutchess,
1 llctpr
•^r
-------
TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
County
Community Types
Rare Species
Valuable Species
Crum Elbow Marsh*
Dutchess
Small amount of shallows,
lower marsh, upper marsh,
and tidal swamp forest.
Map turtle population.
Waterfowl migration, value limited by size of the
marsh.
Brackish Water Habitats
Wappinger Creek
Dutchess
Predominantly creek with
smaller amounts of shallows,
mudflats, lower marsh, and
upper marsh.
Osprey feeding during
spring migrations.
Grassleaf arrowhead,
subulate arrowhead,
kidney leaf mud plaintain
and Maryland bur-
marigold.
Important spawning areas for anadromous fish
including alewife, blueback herring, white perch,
tomcod, and striped bass. Resident fish include
largemouth bass, bluegill, brown bullhead, and red-
breasted sunfish. Productive area for herons,
waterfowl, and turtles.
Fishkill Creek
Dutchess
Mostly shallows and wooded
upland with smaller amounts
of mudflats, lower marsh,
and upper marsh.
Important feeding site for
migrating osprey and a
potential osprey nesting
site. Least bittern
breeding. Estuary beggar-
ticks, subulate arrowhead,
kidney leaf mud- plantain.
Important spawning areas for anadromous fish
including alewife, blueback herring, white perch,
tomcod, and striped bass. Resident fish include
largemouth bass, bluegill, brown bullhead, and red-
breasted sunfish. Also blue claw crabs, herons and
turtles.
Moodna Creek
Orange
Predominantly freshwater
creek with shallows,
mudflats, lower marsh, and
upper marsh associated with
the creek mouth.
Major feeding and resting
ground for bald eagles and
osprey. Limited summer
feeding ground for bald
eagles. Least bittern
breeding area.
Important spawning areas for anadromous fish
including alewife, blueback herring, smelt, white
perch, tomcod, and striped bass. Resident fish
include largemouth bass, bluegill, brown bullhead,
and pumpkinseed. Also many herons, snapping
turtles, raccoons, and muskrats.
7 of 9
TAMS/MCA
-------
TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
County
Community Types
Rare Species
Valuable Species
Hudson River Miles 44-
56
Orange,
Rockland,
Putnam,
Westchester
Deepwater, shallows, and
forested uplands.
Bald eagle winter feeding
grounds. Possible nursery
area for shortnose
sturgeon.
The major spawning area along the Hudson for
striped bass and white perch (about 50% of
northeast striped bass stocks come from the
Hudson). Narrow migration corridor for all
anadromous fish spawning upriver. Marine species
(e.g., bluefish, bay anchovy) live here during
periods of low freshwater flow (generally July
through February).
Constitution Marsh
Putnam
Approximately equal
amounts of shallows,
mudflats, lower marsh, and
upper marsh.
Least bittern nesting site.
Osprey use during
migrations.
Very important nesting habitat for a variety of bird
species including green-backed heron, various
waterfowl, and passerine birds. Important feeding
grounds for herons and other wetland and shore
birds. Significant spawning and feeding grounds
for anadromous and resident fish. Muskrat
population.
lona Island Marsh
Rockland
Mainly upper marsh,
followed by shallows and
flats, with lesser amounts of
woody tidal swamp and non-
tidal freshwater marsh.
Least bittern nesting,
adjacent bald eagle winter
roosting. Walking fern
and prickly pear cactus.
Extensive breeding for many birds. Muskrat and
possibly other furbearers, amphibians, snapping
turtle, and blue claw crab. Heron and shorebird
feeding. Spawning and/or nursery for anadromous
and resident fish.
Camp Smith Marsh and
Annsville Creek*
Westchester
Largely shallows and creek
with smaller amounts of
mudflats and upper marsh.
Spongy arrowhead.
None identified.
gait Water Habitats
8 of 9
TAMS/MCA
-------
TABLE 2-3
LOWER HUDSON RIVER SIGNIFICANT HABITATS
Site Name
Haverstraw Bay
Croton River and Bay
Piermont Marsh
County
Rockland,
Westchester
Westchester
Rockland
Community Types
Deepwater and shallows.
Mostly shallows with lesser
amounts of mudflats and
brackish upper marsh.
Predominantly shallows and
brackish upper marsh with a
broad transition area of
mudflats.
Rare Species
Shortnose sturgeon
wintering area.
Possible osprey feeding
grounds during spring and
fall migrations.
Least bittern and
sedgewren nesting.
Diamondback turtle use.
Osprey feeding during
migration.
Valuable Species
Extensive nursery for anadromous fish species.
Nursery and feeding ground for marine species.
Spawning and wintering grounds for Atlantic
sturgeon. Waterfowl feeding and resting during
migration.
Productive nursery, foraging and resting area for
anadromous and resident fish.
Extensive use of mudflats by herons and egrets.
Large numbers of resident and breeding birds, blue
claw crabs, resident fish, and lesser numbers of
furbearers. Waterfowl, wading bird, and shorebird
feeding during migration.
Notes: * Indicates areas recognized by the NYS Natural Heritage Program as containing rare/important species or communities, but not designated as
significant habitats.
Source: NYSDOS and the Nature Conservancy, 1990.
9 of 9
TAMS/MCA
-------
Table 3-1 Summary of Conversion for the Di through Hexa Homologues
Mean Mass
Percent of Mean +2
Homologue
Calibration
Di-Hexa
Di
Di
Di
Di
Tri-Hexa
Tri-Hexa
Tri-Hexa
Period
Period
1987-1990
High Flow 1991 -1995
Low Flow 1991 -1995
High Flow 1996-1 998
Low Flow 1996-1 998
Fall-winter 1991 -1998
Spring 1991 -1998
Summer 1991-1998
Tri+ Using
TID Data
Standard
Errors
Mean -2 Mean Mass
Standard Percent Ratio
Errors
Repeat the 1991
32.17
48.40
70.64
96.46
36.28
53.02
76.69
102.16
GE TID Data
GE TID Data
GE TID Data
28.07
43.78
64.60
90.76
Waterford/TID
Distribution
1.04
0.52
1.04
0.52
Same as below
by homologue.
ti
"
Corrected
TID Mass
Percent
33.37
25.41
73.27
50.64
Varies
Varies
Varies
Mass
Percent of
Tri+ at
Waterford
33.37
25.41
73.27
50.64
Varies
Varies
Varies
Forecast Period
Di
Di
Tri
Tri
Tri
Tetra
Tetra
Tetra
Penta
Penta
Penta
Hexa
Hexa
Hexa
Tri-Hexa
Tri-Hexa
Tri-Hexa
High Flow 1999+
Low Flow 1999+
Fall-winter 1999+
Spring 1999+
Summer 1999+
Fall-winter 1 999+
Spring 1999+
Summer 1999+
Fall-winter 1999+
Spring 1999+
Summer 1999+
Fall-winter 1999+
Spring 1999+
Summer 1999+
Fall-winter 1 999+
Spring 1999+
Summer 1999+
70.64
96.46
47.21
45.90
54.30
29.66
34.41
30.12
18.10
15.65
12.95
5.00
4.04
2.62
99.97
100.00
99.99
76.69
102.16
48.82
47.71
55.12
30.51
35.55
30.55
19.22
16.88
13.54
5.58
4.61
2.82
64.60
90.76
45.60
44.09
53.48
28.81
33.26
29.69
16.98
14.41
12.37
4.42
3.48
2.41
1.04
0.52
0.98
0.98
0.91
0.97
0.97
1.09
1.19
1.19
1.28
1.23
1.23
1.39
73.27
50.64
46.11
44.83
49.18
28.76
33.36
32.81
21.49
18.58
16.64
6.15
4.97
3.64
102.50
101.74
102.26
73.27
50.64
44.97
44.06
48.08
28.05
32.79
32.08
20.96
18.26
16.27
6.00
4.89
3.56
99.97
100.00
99.99
-------
Table 3-2
Ratio of Striped Bass to Largemouth Bass Concentrations
RM152
Year STB
1990
1991
1992
1993
1995
1994
1996
Tri + ppm
9.02
NA
15.32
10.92
NA
5.61
4.28
LMB Tri+ ppm
3.53
NA
3.24
9.34
NA
NA
2.51
WP Tri+ ppm
0.84
NA
8.64
5.45
NA
4.81
2.78
Average --->»
STB/LMB
2.56
4.73
1.17
1.71
2.54
STB/WP
10.68
1.77
2
1.16
1.54
3.43
RM 152 Monthly Averages
Year
1990
1992
1993
1996
RM113
Year LMB
1988
1989
1990
1991
1992
1993
1994
1995
1996
LMB
June
3.53
3.24
9.34
2.51
Tri+ ppm
7.71
NA
7.84
NA
8.28
4.45
6.26
3.27
3.73
Striped Bass
June-Aug
9.02
15.32
11.38
4.28
WP Tri+ ppm
NA
NA
NA
NA
NA
3.25
1.04
1.86
4.94
June-July
9.39
15.32
11.38
4.28
STB Tri+ ppm
6.31
NA
4.64
NA
2.94
3.27
2.3
1.11
1.66
Average — >»
June Only
4.95
15.32
11.37
2.78
Average
STB/LMB
0.82
0.59
0.35
0.74
0.37
0.34
0.45
0.52
STB/LMB
June-Aug June-July
3.55 3.70
6.03 6.03
4.48 4.48
1 .69 1 .69
2.55 2.58
STB/WP
1.01
2.21
0.6
0.34
1.04
June Only
1.95
6.03
4.47
1.09
2.58
Note:
STB : Striped Bass; WP: White Perch; LMB: Large Mouth Bass.
NA: Data is not available.
-------
Table 3-3
Sum of Monthly Average Loads Over the Troy Dam
(kg)
Homologue
HUDTOX
Converted
According
Thomann/Fa to Appendix
rley Model A
Difference
Di
Tri
Tetra
Penta
Hexa
Total 1987-199;
1182
2320
1664
715
270
6151
2077
2421
1599
742
251
7091
895
101
-65
27
-18
939
Homologue
Di
Tri
Tetra
Total 4/91-2/96
Thomann/Fa
rley Model
857
1645
1081
3583
HUDTOX
Converted
According
to Appendix
A
566
856
593
2015
DEIR
540
1180
860
2580
-------
Table 3-4a
Relative Percent Difference Between FISHRAND Results and Measured Fish Levels in the Lower Hudson
River Mile
Year
1987
1988
1989
1990
1991
1992
1993
1994
1995
spring
fall
1996
Mean
Std Deviation
Std Error
Mean + 2 std errors
Mean - 2 std errors
Species
Largemouth Bass
152
67%
-5%
-21%
-64%
-38%
-29%
-15%
45%
18%
21%
-51%
113
17%
-29%
-39%
39%
-10%
12%
-2%
-2%
27%
10%
19%
-22%
Brown Bullhead
152
-12%
-31%
-22%
-35%
-21%
-24%
9%
4%
-16%
-32%
White Perch
152
100%
-67%
-28%
-41%
-50%
20%
-11%
62%
25%
40%
-62%
152 (seasonal]
-52%
-48%
113
-13%
137%
14%
-46%
23%
80%
40%
103%
-57%
Yellow Perch
152
-62%
-21%
-46%
-43%
21%
12%
-19%
-67%
152 (seasonal)
-32%
-60%
113
-16%
43%
14%
42%
30%
73%
-46%
Pumpkinseed
142
-28%
-30%
100%
-38%
-55%
17%
-6%
57%
23%
41%
-52%
60
18%
25%
-2%
9%
-7%
77%
-40%
-10%
9%
34%
12%
33%
-15%
Average RPD -6%
Note:
RPD = (Predicted Median Concentration - Observed Median Concentration)/Observed Median Concentration
Concentrations are all wet weight concentrations.
-------
Table 3-4b
Relative Percent Difference Between FISHRAND Results and
Measured Spottail Shiner Levels in the Lower Hudson
Location (RM)
Model
60
90
113
152
Measurement
58.7
88.9
113.8
143.5
RPD
-22%
-27%
-65%
5%
Mean RPD
-27%
Note:
RPD = (Predicted Median Concentration - Observed Median Concentration)/Observed Median Concentration
Concentrations are all wet weight concentrations.
-------
TABLE 3-5:
3-2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
Tri+ Average PCB Results
152 113
Whole Whole 90 Whole 50 Whole
Water Water Water Water
Cone Cone Cone Cone
mg/l mg/l mg/l mg/l
4.4E-05
4.0E-05
1 .6E-05
4.7E-05
3.IE-05
I.8E-05
I.6E-05
2.6E-05
2.9E-05
I.7E-05
1 .9E-05
l.OE-05
1.4E-05
I.9E-05
1.9E-05
7.9E-06
8.5E-06
1.5E-05
1.5E-05
1 .OE-05
1 .4E-05
1.10-05
l.OE-05
5.4E-06
5.IE-06
7.6E-06
3. OE-05
2.6E-05
1 .6E-05
2.6E-05
2.IE-05
1.5E-05
I.3E-05
I.5E-05
1.7E-05
I.3E-05
1 .3E-05
8.6E-06
9.1E-06
I.1E-05
1.IE-05
7.0E-06
6.5E-06
8.8E-06
9.IE-06
7.7E-06
8.6E-06
7.5E-06
7.IE-06
5.0E-06
4.4E-06
5.4E-06
2.3E-05
2.0E-05
1 .6E-05
.8E-05
.6E-05
.3E-05
.IE-05
.1E-05
.2E-05
.OE-05
9.7E-06
7.8E-06
7.2E-06
7.5E-06
7.4E-06
6.IE-06
5.6E-06
6.1E-06
6.2E-06
5.9E-06
6.0E-06
5.7E-06
5.4E-06
4.6E-06
4.IE-06
4.3E-06
1 .8E-05
1.6E-05
1.4E-05
1 .3E-05
1.2E-05
1. IE-05
9.7E-06
9.0E-06
8.7E-06
8.0E-06
7.5E-06
6.5E-06
6.0E-06
5.8E-06
5.5E-06
5.0E-06
4.6E-06
4.6E-06
4.6E-06
4.5E-06
4.4E-06
4.3E-06
4.IE-06
3.8E-06
3.5E-06
3.4E-06
Tri+ 95% UCL Results
152 113
Whole Whole 90 Whole 50 Whole
Water Water Water Water
Cone Cone Cone Cone
mg/l mg/l mg/l mg/l
6. IE-05
4.9E-05
1.8E-05
6.9E-05
4.0E-05
2.0E-05
1.7E-05
3. IE-05
4.0E-05
2.0E-05
2.5E-05
1. IE-05
1.8E-05
2.6E-05
3.2E-05
8.7E-06
1 .OE-05
2.3E-05
2.5E-05
1.3E-05
2.0E-05
1.3E-05
1.2E-05
5.9E-06
5.7E-06
1. IE-05
3.8E-05
3. IE-05
I.9E-05
3.2E-05
2.5E-05
1.8E-05
1 .5E-05
1 .8E-05
2. IE-05
I.5E-05
1 .5E-05
9.8E-06
1. IE-05
1.3E-05
1.4E-05
8.0E-06
7.6E-06
1. IE-05
I.2E-05
9.2E-06
l.OE-05
8.6E-06
8.1E-06
5.7E-06
5.0E-06
6.8E-06
2.8E-05
2.4E-05
1 .9E-05
2. IE-05
1 .9E-05
1 .6E-05
1 .4E-05
1 .3E-05
1.4E-05
1 .2E-05
1 .2E-05
9.3E-06
8.5E-06
8.8E-06
8.7E-06
7.2E-06
6.6E-06
7.2E-06
7.3E-06
7.1E-06
7.1E-06
6.7E-06
6.4E-06
5.4E-06
4.8E-06
5.2E-06
2.2E-05
.9E-05
.6E-05
.6E-05
.5E-05
.3E-05
.IE-05
.IE-05
l.OE-05
9.6E-06
9.0E-06
7.8E-06
7.0E-06
6.8E-06
6.5E-06
5.9E-06
5.5E-06
5.5E-06
5.4E-06
5.4E-06
5.2E-06
5.1E-06
4.9E-06
4.5E-06
4.IE-06
4.IE-06
152
Whole
Water
Cone
mg/l
3.7E-08
3.4E-08
1 .4E-08
4.0E-08
2.6E-08
1.6E-08
1 .3E-08
2.2E-08
2.4E-08
1 .4E-08
1.6E-08
8.6E-09
I.2E-08
1 .6E-08
1 .6E-08
6.7E-09
7.2E-09
I.3E-08
1 .3E-08
8.9E-09
1 .2E-08
9.2E-09
8.8E-09
4.6E-09
4.4E-09
6.5E-09
Average Avian TEF
113
Whole 90 Whole
Water Water
Cone Cone
mg/l mg/l
2.6E-08
2.2E-08
1.4E-08
2.2E-08
.8E-08
.3E-08
.IE-08
.3E-08
.5E-08
.IE-08
.IE-08
7.3E-09
7.7E-09
9.1E-09
9.3E-09
5.9E-09
5.5E-09
7.5E-09
7.8E-09
6.5E-09
7.3E-09
6.4E-09
6.0E-09
4.2E-09
3.7E-09
4.6E-09
2.0E-08
1.7E-08
1 .4E-08
1.5E-08
1.4E-08
1. IE-08
9.8E-09
9.7E-09
9.8E-09
8.7E-09
8.3E-09
6.7E-09
6.2E-09
6.4E-09
6.3E-09
5.2E-09
4.7E-09
5.2E-09
5.2E-09
5.0E-09
5.IE-09
4.8E-09
4.6E-09
3.9E-09
3.5E-09
3.6E-09
50 Whole
Water
Cone
mffl
1.6E-08
I.4E-08
1.2E-08
1. IE-08
1. IE-08
9.3E-09
8.2E-09
7.7E-09
7.4E-09
6.8E-09
6.4E-09
5.6E-09
5.IE-09
4.9E-09
4.7E-09
4.2E-09
3.9E-09
3.9E-09
3.9E-09
3.8E-09
3.8E-09
3.6E-09
3.5E-09
3.2E-09
3.0E-09
2.9E-09
152
Whole
Water
Cone
mg/l
5.2E-08
4.2E-08
I.5E-08
5.9E-08
3.4E-08
I.7E-08
1 .5E-08
2.6E-08
3.4E-08
I.7E-08
2. IE-08
9.5E-09
I.6E-08
2.2E-08
2.7E-08
7.4E-09
8.6E-09
I.9E-08
2. IE-08
1. IE-08
1 .7E-08
1. IE-08
l.OE-08
5.0E-09
4.8E-09
9.0E-09
95% Avian TEF
113
Whole 90 Whole
Water Water
Cone Cone
mg/l mg/l
3.2E-08
2.6E-08
I.6E-08
2.7E-08
2.2E-08
I.5E-08
1 .2E-08
1.5E-08
I.8E-08
I.3E-08
1.3E-08
8.4E-09
8.9E-09
1. IE-08
I.2E-08
6.8E-09
6.5E-09
9.3E-09
9.8E-09
7.8E-09
8.9E-09
7.3E-09
6.9E-09
4.9E-09
4.3E-09
5.8E-09
2.4E-08
2.0E-08
1 .6E-08
.8E-08
.6E-08
.4E-08
.2E-08
.IE-08
.2E-08
.OE-08
9.9E-09
7.9E-09
7.2E-09
7.5E-09
7.4E-09
6.1E-09
5.6E-09
6.1E-09
6.2E-09
6.0E-09
6.0E-09
5.7E-09
5.4E-09
4.6E-09
4.IE-09
4.4E-09
50 Whole
Water
Cone
mg/l
.9E-08
.6E-08
.4E-08
.3E-08
.3E-08
.IE-08
9.8E-09
9.1E-09
8.8E-09
8.2E-09
7.6E-09
6.6E-09
6.0E-09
5.8E-09
5.6E-09
5.0E-09
4.7E-09
4.7E-09
4.6E-09
4.6E-09
4.5E-09
4.3E-09
4.2E-09
3.8E-09
3.5E-09
3.5E-09
Average Mammalian 1th
152 113
Whole Whole 90 Whole 50 Whole
Water Water Water Water
Cone Cone Cone Cone
mg/l mg/l mg/l mg/l
2.9E-08
2.6E-08
1. IE-08
3. IE-08
2.0E-08
I.2E-08
l.OE-08
1 .7E-08
1.9E-08
1. IE-08
I.2E-08
6.6E-09
9.4E-09
1 .2E-08
1.3E-08
5.2E-09
5.6E-09
9.9E-09
l.OE-08
6.8E-09
8.9E-09
7.1E-09
6.8E-09
3.5E-09
3.3E-09
5.0E-09
2.0E-08
.7E-08
.IE-08
.7E-08
.4E-08
.OE-08
8.3E-09
l.OE-08
1. IE-08
8.3E-09
8.3E-09
5.6E-09
5.9E-09
7.0E-09
7.2E-09
4.5E-09
4.2E-09
5.7E-09
6.0E-09
5.0E-09
5.6E-09
4.9E-09
4.6E-09
3.3E-09
2.9E-09
3.5E-09
1.5E-08
1 .3E-08
l.OE-08
I.2E-08
l.OE-08
8.8E-09
7.5E-09
7.4E-09
7.6E-09
6.6E-09
6.3E-09
5.IE-09
4.7E-09
4.9E-09
4.8E-09
4.0E-09
3.6E-09
4.0E-09
4.0E-09
3.8E-09
3.9E-09
3.7E-09
3.5E-09
3.0E-09
2.7E-09
2.8E-09
1.2E-08
l.OE-08
9.0E-09
8.7E-09
8.1E-09
7.1E-09
6.3E-09
5.9E-09
5.7E-09
5.2E-09
4.9E-09
4.3E-09
3.9E-09
3.8E-09
3.6E-09
3.3E-09
3.0E-09
3.0E-09
3.0E-09
2.9E-09
2.9E-09
2.8E-09
2.7E-09
2.5E-09
2.3E-09
2.2E-09
95% UCL Mammalian TEF
152 113
Whole Whole 90 Whole 50 Whole
Water Water Water Water
Cone Cone Cone Cone
mg/l mg/l mg/l mg/l
4.0E-08
3.2E-08
1.2E-08
4.5E-08
2.6E-08
1.3E-08
1. IE-08
2.0E-08
2.6E-08
I.3E-08
1.6E-08
7.3E-09
I.2E-08
1.7E-08
2.1E-08
5.7E-09
6.6E-09
1.5E-08
1.6E-08
8.5E-09
1.3E-08
8.3E-09
8.0E-09
3.9E-09
3.7E-09
6.9E-09
2.5E-08
2.0E-08
I.2E-08
2. IE-08
1.7E-08
1. IE-08
9.5E-09
I.2E-08
1.4E-08
9.9E-09
l.OE-08
6.4E-09
6.9E-09
8.4E-09
9.1E-09
5.2E-09
5.0E-09
7.2E-09
7.5E-09
6.0E-09
6.8E-09
5.6E-09
5.3E-09
3.7E-09
3.3E-09
4.4E-09
1.8E-08
1.6E-08
I.2E-08
1.4E-08
1.2E-08
I. OE-08
8.9E-09
8.7E-09
9.0E-09
8.0E-09
7.6E-09
6.IE-09
5.5E-09
5.8E-09
5.7E-09
4.7E-09
4.3E-09
4.7E-09
4.8E-09
4.6E-09
4.6E-09
4.4E-09
4.2E-09
3.5E-09
3.2E-09
3.4E-09
1.4E-08
1 .2E-08
1. IE-08
l.OE-08
9.7E-09
8.5E-09
7.5E-09
7.0E-09
6.8E-09
6.3E-09
5.9E-09
5.1E-09
4.6E-09
4.5E-09
4.3E-09
3.9E-09
3.6E-09
3.6E-09
3.5E-09
3.5E-09
3.4E-09
3.3E-09
3.2E-09
2.9E-09
2.7E-09
2.7E-09
TAMS/MCA
-------
TABLE 3-6: SUMMARY OF TRI+ SEDIMENT CONCENTRATIONS FROM THE FARLEY MODEL AND TEQ-BASED PREDICTIONS FOR 1993 - 2018
Tri+ Average PCB Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 11 3. Total 90 Total 50 Total
Year Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg ms/kg rng/kg
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
0.967
0.882
0.806
0.809
0.787
0.728
0.680
0.666
0.672
0.646
0.616
0.586
0.566
0.561
0.549
0.528
0.508
0.501
0.494
0.480
0.471
0.457
0.443
0.429
0.418
0.407
0.757
0.720
0.676
0.649
0.630
0.600
0.568
0.547
0.537
0.524
0.506
0.486
0.468
0.457
0.446
0.434
0.421
0.411
0.403
0.394
0.386
0.377
0.367
0.357
0.348
0.339
0.610
0.581
2.181
2.179
0.503
0.482
0.460
0.440
0.425
0.415
0.401
0.387
0.372
0.360
0.350
0.340
0.329
0.320
0.312
0.305
0.298
0.291
0.284
0.276
0.269
0.261
0.449
0.426
0.406
0.387
0.370
0.355
0.341
0.327
0.315
0.306
0.296
0.286
0.276
0.267
0.259
0.251
0.244
0.237
0.230
0.225
0.219
0.214
0.208
0.203
0.198
0.193
1.072
1 .023
0.999
0.977
0.954
0.942
0.938
0.910
0.870
0.866
0.848
0.872
0.875
0.811
0.789
0.809
0.839
0.770
0.714
0.699
0.679
0.668
0.659
0.706
0.714
0.679
0.860
0.838
0.817
0.795
0.777
0.766
0.761
0.745
0.726
0.709
0.695
0.700
0.693
0.675
0.658
0.646
0.656
0.639
0.617
0.586
0.571
0.558
0.560
0.557
0.556
0.561
0.677
0.656
0.652
0.634
0.606
0.590
0.574
0.566
0.552
0.540
0.528
0.524
0.513
0.503
0.500
0.489
0.480
0.469
0.457
0.445
0.433
0.421
0.411
0.403
0.395
0.388
0.505
0.490
0.474
0.460
0.450
0.438
0.431
0.421
0.411
0.401
0.398
0.389
0.380
0.372
0.371
0.363
0.355
0.348
0.340
0.332
0.323
0.315
0.307
0.300
0.293
0.287
152 Total
Sed Cone
mg/kg
S.2E-04
7.5E-04
6.9E-04
6.9E-04
6.7E-04
6.2E-04
5.8E-04
5.7E-04
5.7E-04
5.5E-04
5.2E-04
5.0E-04
4.8E-04
4.8E-04
4.7E-04
4.5E-04
4.3E-04
4.3E-04
4.2E-04
4.1E-04
4.0E-04
3.9E-04
3.8E-04
3.6E-04
3.6E-04
3.5E-04
Average Avian TEF
113 Total 90 Total
Sed Cone Sed Cone
mg/kg mg/kg
6.4E-04
6.IE-04
5.7E-04
5.5E-04
5.4E-04
5.IE-04
4.8E-04
4.6E-04
4.6E-04
4.5E-04
4.3E-04
4.1E-04
4.0E-04
3.9E-04
3.8E-04
3.7E-04
3.6E-04
3.5E-04
3.4E-04
3.4E-04
3.3E-04
3.2E-04
3.IE-04
3.0E-04
3.0E-04
2.9E-04
5.2E-04
4.9E-04
I.9E-03
1.9E-03
4.3E-04
4.1E-04
3.9E-04
3.7E-04
3.6E-04
3.5E-04
3.4E-04
3.3E-04
3.2E-04
3.1E-04
3.0E-04
2.9E-04
2.8E-04
2.7E-04
2.7E-04
2.6E-04
2.5E-04
2.5E-04
2.4E-04
2.3E-04
2.3E-04
2.2E-04
50 Total
Sed Cone
mg/kg
3.8E-04
3.6E-04
3.4E-04
3.3E-04
3.1E-04
3.0E-04
2.9E-04
2.8E-04
2.7E-04
2.6E-04
2.5E-04
2.4E-04
2.3E-04
2.3E-04
2.2E-04
2.IE-04
2.IE-04
2.0E-04
2.0E-04
1 .9E-04
I.9E-04
I.8E-04
I.8E-04
1 .7E-04
1.7E-04
1 .6E-04
152 Total
Sed Cone
mg/kg
9.1E-04
8.7E-04
8.5E-04
8.3E-04
8.IE-04
8.0E-04
8.0E-04
7.7E-04
7.4E-04
7.4E-04
7.2E-04
7.4E-04
7.4E-04
6.9E-04
6.7E-04
6.9E-04
7.1E-04
6.5E-04
6.IE-04
5.9E-04
5.8E-04
5.7E-04
5.6E-04
6.0E-04
6.1E-04
5.8E-04
95% Avian TEF
113 Total 90 Total
Sed Cone Sed Cone
mg/kg mg/kg
7.3E-04
7.1E-04
6.9E-04
6.8E-04
6.6E-04
6.5E-04
6.5E-04
6.3E-04
6.2E-04
6.0E-04
5.9E-04
5.9E-04
5.9E-04
5.7E-04
5.6E-04
5.5E-04
5.6E-04
5.4E-04
5.2E-04
5.0E-04
4.9E-04
4.7E-04
4.8E-04
4.7E-04
4.7E-04
4.8E-04
5.8E-04
5.6E-04
5.5E-04
5.4E-04
5.2E-04
5.0E-04
4.9E-04
4.8E-04
4.7E-04
4.6E-04
4.5E-04
4.5E-04
4.4E-04
4.3E-04
4.3E-04
4.2E-04
4.IE-04
4.0E-04
3.9E-04
3.8E-04
3.7E-04
3.6E-04
3.5E-04
3.4E-04
3.4E-04
3.3E-04
50 Total
Sed Cone
mg/kg
4.3E-04
4.2E-04
4.0E-04
3.9E-04
3.8E-04
3.7E-04
3.7E-04
3.6E-04
3.5E-04
3.4E-04
3.4E-04
3.3E-04
3.2E-04
3.2E-04
3.2E-04
3.IE-04
3.0E-04
3.0E-04
2.9E-04
2.8E-04
2.7E-04
2.7E-04
2.6E-04
2.6E-04
2.5E-04
2.4E-04
Average Mammalian I bh
152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone
mg/kg mg/kg mg/kg mg/kg
6.3E-04
5.8E-04
5.3E-04
5.3E-04
5.1E-04
4.8E-04
4.4E-04
4.3E-04
4.4E-04
4.2E-04
4.0E-04
3.8E-04
3.7E-04
3.7E-04
3.6E-04
3.4E-04
3.3E-04
3.3E-04
3.2E-04
3.1E-04
3.1E-04
3.0E-04
2.9E-04
2.8E-04
2.7E-04
2.7E-04
4.9E-04
4.7E-04
4.4E-04
4.2E-04
4.IE-04
3.9E-04
3.7E-04
3.6E-04
3.5E-04
3.4E-04
3.3E-04
3.2E-04
3.IE-04
3.0E-04
2.9E-04
2.8E-04
2.7E-04
2.7E-04
2.6E-04
2.6E-04
2.5E-04
2.5E-04
2.4E-04
2.3E-04
2.3E-04
2.2E-04
4.0E-04
3.8E-04
1.4E-03
1.4E-03
3.3E-04
3.1E-04
3.0E-04
2.9E-04
2.8E-04
2.7E-04
2.6E-04
2.5E-04
2.4E-04
2.4E-04
2.3E-04
2.2E-04
2.2E-04
2.1E-04
2.0E-04
2.0E-04
1.9E-04
1.9E-04
1.9E-04
1.8E-04
1.8E-04
1.7E-04
2.9E-04
2.8E-04
2.6E-04
2.5E-04
2.4E-04
2.3E-04
2.2E-04
2.1E-04
2.1E-04
2.0E-04
.9E-04
.9E-04
.8E-04
.7E-04
.7E-04
.6E-04
.6E-04
.5E-04
.5E-04
.5E-04
1.4E-04
1.4E-04
1.4E-04
1.3E-04
1.3E-04
1.3E-04
95% UCL Mammalian TEF
152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone
mg/kg mg/kg rag/kg mg/kg
7.0E-04
6.7E-04
6.5E-04
6.4E-04
6.2E-04
6.1E-04
6.1E-04
5.9E-04
5.7E-04
5.7E-04
5.5E-04
5.7E-04
5.7E-04
5.3E-04
5.1E-04
5.3E-04
5.5E-04
5.0E-04
4.7E-04
4.6E-04
4.4E-04
4.4E-04
4.3E-04
4.6E-04
4.7E-04
4.4E-04
5.6E-04
5.5E-04
5.3E-04
5.2E-04
5.1E-04
5.0E-04
5.0E-04
4.9E-04
4.7E-04
4.6E-04
4.5E-04
4.6E-04
4.5E-04
4.4E-04
4.3E-04
4.2E-04
4.3E-04
4.2E-04
4.0E-04
3.8E-04
3.7E-04
3.6E-04
3.7E-04
3.6E-04
3.6E-04
3.7E-04
4.4E-04
4.3E-04
4.3E-04
4.1E-04
4.0E-04
3.8E-04
3.7E-04
3.7E-04
3.6E-04
3.5E-04
3.4E-04
3.4E-04
3.4E-04
3.3E-04
3.3E-04
3.2E-04
3.1E-04
3.1E-04
3.0E-04
2.9E-04
2.8E-04
2.8E-04
2.7E-04
2.6E-04
2.6E-04
2.5E-04
3.3E-04
3.2E-04
3.1E-04
3.0E-04
2.9E-04
2.9E-04
2.8E-04
2.7E-04
2.7E-04
2.6E-04
2.6E-04
2.5E-04
2.5E-04
2.4E-04
2.4E-04
2.4E-04
2.3E-04
2.3E-04
2.2E-04
2.2E-04
2.1E-04
.2.1E-04
2.0E-04
2.0E-04
I.9E-04
I.9E-04
TAMS/MCA
-------
TABLE 3-7: ORGANIC CARBON NORMALIZED SEDIMENT CONCENTRATIONS
BASED ON USEPA PHASE 2 DATASET
1 Tri+ Average PCB Results
Year
1993
1994
1995
1996
11997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152 Total
Sed Cone
mg/kg
38.67
35.29
32.25
32.38
31.47
29.13
27.20
26.66
26.88
25.85
24.64
23.42
22.66
22.42
21.96
21.12
20.31
20.05
19.76
19.20
18.85
18.28
17.71
17.16
16.73
16.26
113 Total
Sed Cone
mg/kg
30.29
28.81
27.04
25.97
25.19
24.00
22.73
21.87
21.47
20.97
20.26
19.45
18.74
18.27
17.86
17.37
16.82
16.43
16.11
15.77
15.44
15.08
14.70
14.29
13.91
13.58
^H^^^^^^^VM
90 Total
Sed Cone
mg/kg
24.39
23.23
87.23
87.14
20.14
19.29
18.40
17.59
16.99
16.60
16.06
15.49
14.90
14.40
13.98
13.59
13.18
12.80
12.48
12.19
11.91
11.63
11.34
11.03
10.74
10.44
— ^-^— — •—
50 Total
Sed Cone
mg/kg
17.97
17.05
16.22
15.47
14.82
14.21
13.62
13.07
12.58
12.23
11.82
11.43
11.04
10.67
10.35
10.05
9.75
9.47
9.22
8.98
8.76
8.54
8.34
8.12
7.93
7.71
— i— — ^-~
152 Total
Sed Cone
mg/kg
42.90
40.94
39.96
39.06
38.17
37.68
37.53
36.39
34.79
34.66
33.94
34.89
35.00
32.42
31.55
32.35
33.55
30.80
28.57
27.98
27.16
26.74
26.38
28.25
28.54
27.16
•—•——•
Tri+ 95% UCL Results
113 Total
Sed Cone
mg/kg
34.40
33.51
32.67
31.78
31.06
30.64
30.42
29.78
29.04
28.37
27.80
27.99
27.70
26.98
26.30
25.85
26.25
25.58
24.67
23.45
22.84
22.33
22.42
22.30
22.23
22.43
.— — — —
90 Total
Sed Cone
mg/kg
27.09
26.22
26.08
25.34
24.23
23.58
22.95
22.62
22.08
21.61
21.11
20.95
20.54
20.10
20.00
19.56
19.18
18.77
18.29
17.79
17.31
16.86
16.45
16.11
15.80
15.53
.— — —
50 Total
Sed Cone
mg/kg
20.18
19.60
18.97
18.39
18.02
17.53
17.26
16.83
16.42
16.05
15.91
15.56
15.21
14.89
14.84
1452 1
14.22 1
13.92 1
13.60 1
13.27 I
12.94 1
12.61 1
12.29 I
12.00
11.71
11.48 |
average TOC from Farley model 2.5%
TAMS/MCA
-------
TABLE 3-8: SUMMARY OF TRI+ BENTHIC INVERTEBRATE CONCENTRATIONS FROM THE FISHRAND MODEL AND TEQ-BASED PREDICTIONS FOR 1993 - 2018
Tri+ Average PCB Results Tri+ 95% UCL Results Average Avian TEF 95% Avian TEF Average Mammalian TEF 95% UCL Mammalian TEF
!52Total IBTotal 90Total SOToial 152Total H3Total 90Total SOTotal 152Total IBTotal 90Total SOTotal !52Total USTotal 90Total SOTotal 152Toial H3Total 90Total SOTotal 152Tolal H3Total 90Total SOTotal
Benihic Benihic Benthic Benthic Benthic Benthie Benthic Benthic Benthic Benthic Benthic Benthic Benlhic Semitic Benthic Benthic Bentrric Benthic Benthic Benthic Bcnthic Benthic Benlhic Benthic
Year Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone
mg/kR mg/kg me/kg mR/kR mg/kg mg/kg ms/ks nig/kg mg/kg mg/kg mg/kR mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
.754 .393 1.131 0.831 1.885
.573 .304 1.073 0.780 1.686
.522 .252 1.006 0.741 1.632
.502 .202 0.958 0.713 1.610
.422 .153 0.928 0.690 1.524
.362 .121 0.884 0.652 1.460
.291 .087 0.852 0.633
.298 .042 0.829 0.614
.269 1.027 0.804 0.595
.213 0.991 0.784 0.585
.140 0.946 0.767 0.564
.122 0.912 0.727 0.539
.091 0.904 0.700 0.519
.049 0.877 0.669 0.496
.035 0.859 0.652 0.482
2008 0.999 0.827 0.633 0.469
2009 0.978 0.802 0.619 0.459
2010 0.962 0.786 0.608 0.450
.386
.393
.360
.303
.225
.495 1.215 0.893 2.4E-04
.398 1.151 0.837 2.2E-04
.341 1.079 0.794 2.1E-04
.289 1.026 0.764 2.1E-04
.235 0.994 0.739 2.0E-04
.200 0.947 0.699
.166 0.912 0.678
.119 0.887 0.658
.103 0.861 0.637
.065 0.840 0.628
.016 0.823 0.606
.208 0.981 0.781 0.579
.174 0.972 0.752 0.557
.127 0.943 0.720 0.533
.113 0.924 0.701 0.518
.077 0.890 0.680 0.504
.055 0.864 0.665 0.494
.034 0.846 0.653 0.484
2011 0.922 0.779 0.587 0.443 0.991 0.838 0.631 0.477
2012 0.899 0.762 0.573 0.433 0.966 0.820 0.616 0.466
2013 0.879 0.745 0.556 0.420 0.945 0.802 0.598 0.452
2014 0.870 0.727 0.543 0.410 0.935 0.782 0.583 0.441
2015 0.845 0.700 0.532 0.400 0.911 0.754 0.572 0.430
2016 0.853 0.681 0.521 0.392 0.923 0.734 0.560 0.422
2017 0.842 0.675 0.515 0.382 0.912 0.729 0.553 0.411
2018 0.822 0.673 0.505 0.373 0.890 0.728 0.543 0.402
.9E-04
.8E-04
.8E-04
.8E-04
.7E-04
.6E-04
.6E-04
.5E-04
.5E-04
.4E-04
.4E-04
.4E-04
.3E-04
.3E-04
.2E-04
.2E-04
.2E-04
.9E-04
.8E-04
.7E-04
.7E-04
.6E-04
.6E-04
.5E-04
.4E-04
.4E-04
.4E-04
.3E-04
.6E-04 I.2E-04 2.6E-04 2.IE-04 1.7E-04 1.2E-04 1.9E-04
.5E-04 I.1E-04 2.3E-04 I.9E-04 1.6E-04 1.2E-04 1.7E-04
.4E-04 1. OE-04 2.3E-04 I.9E-04 1.5E-04 1.1E-04 1.6E-04
.3E-04 9.9E-05 2.2E-04
.3E-04 9.6E-05 2.1E-04
.2E-04 9.0E-05 2.0E-04
.2E-04 8.8E-05
.1E-04 8.5E-05
.1E-04 8.2E-05
.IE-04 8.1E-05
.IE-04 7.8E-05
.3E-04 1 .OE-04 7.5E-05
.3E-04 9.7E-05 7.2E-05
.2E-04 9.3E-05 6.9E-05
.2E-04 9.0E-05 6.7E-05
.1E-04 8.8E-05 6.5E-05
.IE-04 8.6E-05 6.4E-05
.1E-04 8.4E-05 6.2E-05
.1E-04
i.lE-05 6.1E-05
.1E-04 7.9E-05 6.0E-05
.OE-04 7.7E-05 5.8E-05
.OE-04 7.5E-05 5.7E-05
.2E-04 9.7E-05 7.4E-05 5.5E-05
.2E-04 9.4E-05
.2E-04 9.4E-05
7.2E-05 5.4E-05
ME-05 5.3E-05
.1E-04 9.3E-05 7.0E-05 5.2E-05
.9E-04
.9E-04
.9E-04
.8E-04
.7E-04
.7E-04
.6E-04
.6E-04
.5E-04
.5E-04
.5E-04
.4E-04
.4E-04
.3E-04
.3E-04
.3E-04
.8E-04 I.4E-04 1. IE-04 1.6E-04
.7E-04 I.4E-04 l.OE-04 I.5E-04
.7E-04 I.3E-04 9.7E-05 I.5E-04
.6E-04 1.3E-04 9.4E-05 1.4E-04
.6E-04 1.2E-04 9.IE-05 1.4E-04
.5E-04 1.2E-04 8.8E-05 1.4E-04
.5E-04 I.2E-04 8.7E-05 .3E-04
.4E-04 1. IE-04 8.4E-05 .2E-04
.5E-04 I.2E-04 9.0E-05 2.0E-04 1.6E-04 1.3E-04 9.6E-05
.4E-04 1.2E-04 8.4E-05 1.8E-04 1.5E-04 1.2E-04 9.0E-05
.4E-04 1. IE-04 8.0E-05 1.8E-04 I.4E-04 I.2E-04 8.6E-05
.3E-04 l.OE-04 7.7E-05 1.7E-04 I.4E-04 1. IE-04 8.2E-05
.2E-04 l.OE-04 7.4E-05 1.6E-04 .3E-04 1. IE-04 8.0E-05
.2E-04 9.5E-05 7.0E-05 1.6E-04 .3E-04 l.OE-04 7.5E-05
.2E-04 9.2E-05 6.8E-05 1.5E-04 .3E-04 9.8E-05 7.3E-05
.IE-04 8.9E-05 6.6E-05 I.5E-04 .2E-04 9.6E-05 7.1E-05
.IE-04 8.7E-05 6.4E-05 1.5E-04 .2E-04 9.3E-05 6.9E-05
.IE-04 8.5E-05 6.3E-05 1.4E-04 .IE-04 9.1E-05 6.8E-05
.OE-04 8.3E-05 6.1E-05 1.3E-04 .IE-04 8.9E-05 6.5E-05
.4E-04 1. IE-04 8.0E-05 .2E-04 9.8E-05 7.8E-05 5.8E-05 I.3E-04 1. IE-04 8.4E-05 6.2E-05
.3E-04 l.OE-04 7.7E-05 .2E-04 9.8E-05 7.6E-05 5.6E-05 .3E-04 l.OE-04 8.1E-05 6.0E-05
.3E-04 l.OE-04 7.4E-05 .IE-04 9.5E-05 7.2E-05 5.3E-05 .2E-04 l.OE-04 7.8E-05 5.7E-05
.3E-04 9.7E-05 7.2E-05 .IE-04 9.3E-05 7.0E-05 5.2E-05 .2E-04 l.OE-04 7.6E-05 5.6E-05
.2E-04 9.4E-05 7.0E-05 .IE-04 8.9E-05 6.8E-05 5.1E-05 .2E-04 9.6E-05 7.3E-05 5.4E-05
.2E-04 9.2E-05 6.8E-05 .IE-04 8.7E-05 6.7E-05 5.0E-05 .IE-04 9.3E-05 7.2E-05 5.3E-05
.2E-04 9.IE-05 6.7E-05 l.OE-04 8.5E-05 6.6E-05 4.9E-05 .IE-04 9.1E-05 7.0E-05 5.2E-05
.2E-04 8.7E-05 6.6E-05 9.9E-05 8.4E-05 6.3E-05 4.8E-05 .IE-04 9.0E-05 6.8E-05 5.1E-05
.IE-04 8.5E-05 6.5E-05 9.7E-05 8.2E-05 6.2E-05 4.7E-05 l.OE-04 8.8E-05 6.7E-05 5.0E-05
.IE-04 8.3E-05 6.3E-05 9.5E-05 8.0E-05 6.0E-05 4.5E-05 l.OE-04 8.6E-05 6.4E-05 4.9E-05
.IE-04 8.IE-05 6.IE-05 9.4E-05 7.8E-05 5.9E-05 4.4E-05 l.OE-04 8.4E-05 6.3E-05 4.8E-05
.3E-04 1. OE-04 7.9E-05 6.0E-05 9.1E-05 7.6E-05 5.7E-05 4.3E-05 9.8E-05 8.1E-05 6.2E-05 4.6E-05
.3E-04 l.OE-04 7.8E-05 5.8E-05 9.2E-05 7.3E-05 5.6E-05 4.2E-05 l.OE-04 7.9E-05 6.0E-05 4.5E-05
.3E-04 l.OE-04 7.7E-05 5.7E-05 9.1E-05 7.3E-05 5.6E-05 4.IE-05 9.8E-05 7.9E-05 6.0E-05 4.4E-05
.2E-04 l.OE-04 7.5E-05 5.6E-05 8.9E-05 7.3E-05 5.4E-05 4.0E-05 9.6E-05 7.9E-05 5.9E-05 4.3E-05
TAMS/MCA
-------
TABLE 3-9: SPOTTAIL SHINER PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.36
0.28
0.22
0.29
0.25
0.18
0.15
0.16
0.19
0.15
0.13
0.10
0.11
0.12
0.10
0.09
0.07
0.10
0.09
0.09
0.10
0.09
0.08
0.06
0.06
0.07
0.46
0.41
0.29
0.40
0.32
0.22
0.20
0.22
0.24
0.19
0.18
0.14
0.15
0.17
0.14
0.11
0.11
0.14
0.13
0.13
0.13
0.12
0.11
0.09
0.08
0.09
0.76
0.63
0.51
0.66
0.51
0.34
0.31
0.35
0.39
0.30
0.30
0.22
0.23
0.29
0.22
0.19
0.18
0.21
0.21
0.20
0.22
0.20
0.19
0.14
0.13
0.14
River Mile 1
25th Median
(mg/kg (mg/kg
wet wet
weight) weight)
0.25
0.23
0.18
0.20
0.17
0.14
0.12
0.12
0.12
0.12
0.11
0.09
0.08
0.08
0.08
0.07
0.06
0.07
0.07
0.07
0.07
0.07
0.06
0.05
0.05
0.05
0.33
0.31
0.23
0.27
0.23
0.19
0.16
0.17
0.17
0.15
0.14
0.12
0.11
0.12
0.11
0.10
0.09
0.10
0.10
0.10
0.10
0.10
0.09
0.08
0.07
0.07
13
95th
Percentile
(mg/kg
wet
weight)
0.49
0.45
0.35
0.40
0.34
0.28
0.25
0.25
0.26
0.23
0.22
0.18
0.18
0.18
0.18
0.16
0.15
0.15
0.16
0.16
0.15
0.15
0.14
0.12
0.11
0.12
River Mile 90
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.21
0.19
0.16
0.15
0.14
0.12
0.11
0.10
0.09
0.09
0.09
0.08
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.27
0.24
0.21
0.20
0.18
0.16
0.14
0.13
0.13
0.12
0.12
0.10
0.10
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.06
0.06
0.39
0.35
0.31
0.29
0.27
0.24
0.21
0.20
0.19
0.19
0.18
0.16
0.15
0.14
0.13
0.13
0.12
0.12
0.12
0.12
0.11
0.11
0.11
0.10
0.09
0.10
River Mile 50
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.20
0.18
0.16
0.15
0.13
0.12
0.11
0.10
0.09
0.09
0.08
0.08
0.07
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.26
0.23
0.20
0.18
0.17
0.15
0.14
0.13
0.12
0.11
0.11
0.10
0.09
0.09
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.06
0.06
. 0.06
0.06
0.06
0.38
0.33
0.29
0.27
0.25
0.22
0.20
0.19
0.18
0.17
0.16
0.15
0.14
0.13
0.12
0.12
0.11
0.11
0.10
0.10
0.10
0.10
0.09
0.09
0.09
0.09
TAMS/MCA
-------
TABLE 3-10: PUMPKINSEED PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
1.16
0.86
0.74
0.92
0.78
0.53
0.47
0.49
0.55
0.45
0.43
0.32
0.33
0.40
0.32
0.28
0.26
0.29
0.32
0.29
0.32
0.29
0.26
0.20
0.19
0.20
1.57
1.17
1.03
1.26
1.06
0.77
0.68
0.67
0.75
0.65
0.60
0.46
0.46
0.55
0.45
0.41
0.37
0.41
0.45
0.42
0.45
0.42
0.37
0.30
0.29
0.29
2.54
1.87
1.71
2.03
1.72
1.28
1.13
1.10
1.22
1.10
1. 00
0.78
0.77
0.91
0.75
0.70
0.64
0.70
0.75
0.71
0.76
0.70
0.62
0.52
0.50
0.51
River Mile 113
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.76
0.67
0.53
0.59
0.51
0.42
0.37
0.36
0.37
0.34
0.32
0.27
0.26
0.26
0.26
0.23
0.21
0.21
0.23
0.22
0.22
0.21
0.20
0.18
0.16
0.16
1.05
0.95
0.77
0.81
0.74
0.61
0.54
0.50
0.52
0.50
0.46
0.39
0.36
0.37
0.36
0.34
0.30
0.30
0.32
0.31
0.32
0.30
0.29
0.26
0.24
0.23
1.73
1.54
1.28
1.33
1.24
1.02
0.90
0.84
0.87
0.85
0.77
0.67
0.62
0.63
0.61
0.57
0.52
0.52
0.54
0.53
0.54
0.52
0.48
0.44
0.41
0.40
River Mile 90
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.58
0.53
0.46
0.43
0.39
0.36
0.32
0.29
0.28
0.27
0.25
0.23
0.21
0.20
0.20
0.18
0.17
0.17
0.17
0.17
0.17
0.16
0.15
0.14
0.14
0.13
0.84
0.75
0.66
0.62
0.57
0.53
0.46
0.42
0.40
0.39
0.36
0.33
0.30
0.29
0.28
0.27
0.25
0.24
0.24
0.24
0.24
0.23
0.22
0.21
0.20
0.19
1.37
1.25
1.09
1.02
0.95
0.87
0.77
0.70
0.66
0.65
0.61
0.56
0.52
0.49
0.47
0.45
0.43
0.40
0.40
0.41
0.40
0.39
0.38
0.36
0.34
0.33
River Mile 50
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.57
0.50
0.45
0.40
0.37
0.34
0.30
0.28
0.26
0.25
0.23
0.21
0.20
0.19
0.18
0.17
0.16
0.15
0.15
0.15
0.15
0.14
0.14
0.13
0.13
0.12
0.79
0.71
0.63
0.58
0.53
0.49
0.44
0.40
0.37
0.35
0.33
0.31
0.28
0.27
0.25
0.24
0.23
0.22
0.21
0.21
0.21
0.20
0.20
0.19
0.18
0.17
1.31
1.17
1.04
0.94
0.86
0.79
0.72
0.65
0.60
0.58
0.55
0.51
0.47
0.44
0.42
0.40
0.38
0.36
0.35
0.35
0.35
0.33
0.32
0.32
0.30
0.29
TAMS/MCA
-------
TABLE 3-11: YELLOW PERCH PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.85
0.71
0.67
0.70
0.66
0.58
0.52
0.50
0.51
0.50
0.46
0.42
0.40
0.42
0.40
0.38
0.35
0.35
0.36
0.35
0.35
0.33
0.31
0.30
0.29
0.28
0.99
0.85
0.80
0.83
0.78
0.71
0.63
0.60
0.62
0.60
0.55
0.50
0.48
0.50
0.47
0.46
0.42
0.42
0.43
0.42
0.42
0.40
0.38
0.36
0.35
0.34
1.28
1.11
1.04
1.06
1.01
0.92
0.83
0.79
0.81
0.78
0.73
0.67
0.64
0.66
0.63
0.60
0.57
0.56
0.57
0.55
0.55
0.53
0.51
0.48
0.47
0.45
River Mile 113
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.64
0.58
0.54
0.52
0.50
0.47
0.43
0.40
0.40
0.39
0.37
0.35
0.33
0.32
0.31
0.31
0.29
0.28
0.28
0.28
0.27
0.27
0.26
0.25
0.24
0.23
0.75
0.69
0.64
0.61
0.59
0.56
0.51
0.49
0.48
0.47
0.45
0.42
0.40
0.39
0.38
0.37
0.35
0.34
0.34
0.34
0.33
0.32
0.31
0.30
0.29
0.28
0.98
0.90
0.84
0.81
0.78
0.73
0.68
0.64
0.63
0.62
0.59
0.56
0.53
0.52
0.51
0.49
0.47
0.46
0.46
0.45
0.44
0.43
0.42
0.40
0.39
0.37
River Mile 90
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.51
0.47
0.44
0.42
0.39
0.37
0.35
0.33
0.32
0.31
0.30
0.28
0.27
0.26
0.25
0.24
0.23
0.22
0.22
0.21
0.21
0.21
0.20
0.19
0.19
0.18
0.60
0.56
0.53
0.49
0.47
0.45
0.42
0.40
0.38
0.37
0.36
0.34
0.32
0.31
0.30
0.29
0.28
0.27
0.27
0.26
0.26
0.25
0.24
0.24
0.23
0.22
0.78
0.73
0.69
0.65
0.62
0.59
0.55
0.52
0.50
0.49
0.47
0.45
0.43
0.41
0.40
0.39
0.38
0.36
0.35
0.35
0.34
0.33
0.32
0.31
0.30
0.29
River Mile 50
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
0.41
0.38
0.35
0.33
0.31
0.29
0.27
0.26
0.25
0.24
0.23
0.22
0.21
0.20
0.19
0.19
0.18
0.17
0.17
0.17
0.16
0.16
0.15
0.15
0.15
0.14
0.47
0.44
0.41
0.39
0.36
0.35
0.33
0.31
0.29
0.28
0.27
0.26
0.25
0.24
0.23
0.22
0.22
0.21
0.20
0.20
0.20
0.19
0.19
0.18
0.18
0.17
0.61
0.57
0.53
0.50
0.47
0.45
0.43
0.40
. 0.39
0.37
0.36
0.34
0.33
0.32
0.31
0.30
0.29
0.28
0.27
0.26
0.26
0.25
0.25
0.24
0.23
0.23
TAMS/MCA
-------
TABLE 3-12: WHITE PERCH PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
RiverMilel52 RiverMilellS RiverMile90 RiverMileSO
95th 95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg
wet wet wet wet wet wet wet wet wet wet wet wet
Year weight) weight) weight) weight) weight) weight) weight) weight) weight) weight) weight) weight)
1993 2.69 2.86 3.30 2.08 2.21 2.55
1994 2.32 2.47 2.88
1995 2.16 2.32 2.70
1996 2.32 2.45 2.77
1997 2.10 2.24 2.61
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
.86 2.01 2.40
.72 1.84 2.17
.66 1.77 2.11
.72 1.82 2.12
.65 1.76 2.06
.51 1.62 1.92
.36 1.47 1.78
.31 1.42 1.72
.36 1.45 1.73
.91 2.03 2.37
.76 1.88 2.21
.70 1.80 2.10
.62 1.73 2.04
.54
.39
.31
.29
.27
.21
.13
.07
.05
.30 1.40 1.66 1.02
.23 1.33 1.61 1.00
.15 1.24 1.51 0.95
.17 1.26 1.52 0.92
.19 1.28 1.52 0.92
.63 1.91
.49 1.78
.41 1.69
.39 1.66
.37 1.63
.65
.54
.43
.35
.28
.21
.13
.07
.02
.00
.30 1.56 0.96
.75 2.03 1.32 1.39 .58
.64
.53
.44
.37
.30
.22
.15
.10
.07
.03
.23 1.48 0.91 0.99
.16 1.41 0.87 0.94
.14 1.38 0.83 0.90
.11 1.34 0.80 0.87
.08 1.31 0.78 0.85
.92 1.23 1.29 .47
.81 1.14 1.20 .38
.69 1.07 1.13 .29
.62 1.01 1.07 .23
.55 0.95 1.02 .18
.45 0.89 0.96 .11
.37 0.85 0.90 .05
.32 0.81 0.86 .01
.28 0.78 0.83 0.97
.24 0.75 0.80 0.94
.19 0.71 0.76 0.90
.13 0.68 0.73 0.86
.09 0.65 0.70 0.83
.06 0.63 0.67 0.80
.03 0.61 0.65 0.78
.03 1.25 0.75 0.82 0.99 0.58 0.63 0.75
.01 1.23 0.72 0.79 0.96 0.56 0.61 0.73
.00 1.21 0.71 0.77 0.94 0.55 0.59 0.71
.14 1.23 1.48 0.91 0.99 1.20 0.71 0.76 0.93 0.54 0.58 0.69
.15 1.24 1.47 0.90 0.97 1.17 0.69 0.75 0.90 0.53 0.57 0.67
.09 1.17 1.40 0.87 0.94 1.14 0.67 0.72 0.88 0.51 0.55 0.66
.03 1.11 1.34 0.84 0.91 1.10 0.65 0.70 0.86 0.50 0.53 0.64
2016 0.98 1.06 1.29 0.81 0.88 1.07 0.63 0.68 0.83 0.48 0.52 0.62
2017 0.94 1.02 1.25 0.77 0.84 1.03 0.61 0.66 0.81 0.47 0.51 0.61
2018 0.92 1.01 1.23 0.76 0.83 1.02 0.59 0.65 0.80 0.46 0.50 0.60
TAMS/MCA
-------
TABLE 3-13: BROWN BULLHEAD PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
RiverMilel52 RiverMilellS RiverMileQO RiverMileSO
95th 95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg (mg/kg
wet wet wet wet wet wet wet wet wet wet wet wet
Year weight) weight) weight) weight) weight) weight) weight) weight) weight) weight) weight) weight)
1993 2.34 3.32 5.48
1994 2.04 2.94 4.90
1995 1.90 2.74 4.56
1996 1.93 2.77 4.61
1997 1.83 2.63 4.38
1998 1.69 2.43 4.06
1999 1.52 2.20 3.70
2000 1.48 2.16 3.63
2001 1.50 2.17 3.62
2002 1.44 2.09 3.49
2003 1.35
2004 1.26
2005 1.21
2006 1.23
2007 1.17
2008 1.13
2009 1.08
2010 1.06
2011 1.07
2012 1.04
2013 1.02
2014 0.99
2015 0.95
2016 0.90
2017 0.88
2018 0.85
.96 3.29
.83 3.08
.78 2.99
.78 2.55 4.28 1.43 2.05 3.44 1.10 1.57 2.59
.66 2.39 4.00 1.35
.54 2.23 3.75 1.26
.49 2.14 3.60 1.19
.43 2.07 3.45 1.14
.34
.25
.20
.18
.15
.09
.04
.00
.78 2.98 0.98
.71 2.88 0.95
.64 2.77 0.93
.57 2.65 0.89
.57 2.64 0.87
.55 2.62 0.86
.52 2.55 0.84
.49 2.51 0.83
.44 2.42 0.81
.38 2.33 0.78
.32 2.24 0.76
.28 2.16 0.73
.25 2.12 0.71
.95 3.28 1.09
.81 3.05 1.02
.75 2.93 0.97
.72 2.87 0.93
.67 2.80 0.91
.60 2.69 0.87
.52 2.57 0.83
.46 2.46 0.80
.43 2.40 0.77
.39 2.34 0.75
.35 2.27 0.73
.30 2.19 0.70
.27 2.14 0.68
.93 3.25 1.03 1.47 2.44
.82 3.06 0.97 .39 2.31
.72 2.90 0.91 .31 2.18
.64 2.77 .0.87 .24 2.08
.57 2.64 0.83 .18
.48 2.50 0.78 .13
.41 2.36 0.74 .07
.36 2.28 0.71 .03
.32 2.21 0.69 0.99
.27 2.14 0.66 0.96
.22 2.06 0.63 0.92
.17
.13
.10
.06
.03
.00
.26 2.11 0.66 0.97
.24 2.07 0.65 0.96
.21 2.03 0.64 0.93
.18 1.98 0.62 0.91
.14 1.92 0.61 0.89
.10 1.86 0.59 0.86
.07 1.80 0.57 0.83
.04 1.77 0.55 0.81
.97 0.61 0.89
.90 0.59 0.85
.84 0.57 0.82
.78 0.55 0.80
.72 0.53 0.77
.67 0.52 0.75
.64 0.50 0.73
.61 0.49 0.72
.57 0.48 0.70
.53 0.47 0.68
.49 0.46 0.66
.44 0.44 0.64
.40 0.43 0.63
.37 0.42 0.61
.97
.88
.79
.71
.66
.60
.54
.48
.43
.38
.34
.29
.25
.22
.20
.16
.13
.11
.08
.05
.03
TAMS/MCA
-------
TABLE 3-14: LARGEMOUTH BASS PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
11.28
8.05
7.10
8.25
7.62
6.05
5.06
4.78
5.34
5.07
4.34
3.59
3.35
3.83
3.48
3.32
2.81
2.99
3.28
2.99
3.19
2.94
2.70
2.56
2.27
2,16
14.33
10.38
8.92
10.58
9.63
7.56
6.53
6.12
6.96
6.37
5.66
4.57
4.35
4.90
4.52
4.21
3.64
3.84
4.29
3.84
4.18
3.80
3.51
3.22
2.90
2.82
21.56
15.44
13.51
15.79
14.45
11.61
9.76
9.25
10.34
9.66
8.54
7.01
6.61
7.49
6.79
6.41
5.57
5.80
6.49
5.81
6.30
5.80
5.36
4.97
4.44
4.30
River Mile 1
25th Median
(mg/kg (mg/kg
wet wet
weight) weight)
7.50
6.55
5.89
5.39
5.26
4.73
3.96
3.57
3.64
3.62
3.31
2.96
2.68
2.65
2.60
2.53
2.29
2.18
2.31
2.27
2.33
2.22
2.11
1.99
1.82
1.71
9.58
8.37
7.45
6.94
6.71
6.10
5.10
4.64
4.70
4.65
4.27
3.79
3.48
3.44
3.37
3.24
2.96
2.83
3.01
2.94
3.03
2.87
2.74
2.55
2.35
2.23
13
95th
Percentile
(mg/kg
wet
weight)
14.39
12.63
11.24
10.40
10.08
9.19
7.73
7.04
7.11
7.07
6.52
5.81
5.31
5.23
5.10
4.96
4.54
4.31
4.57
4.49
4.62
4.38
4.17
3.91
3.59
3.42
River Mile 90
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
1.84
1.69
1.52
1.37
1.29
1.20
1.07
0.96
0.90
0.88
0.84
0.78
0.72
0.67
0.65
0.63
0.59
0.56
0.56
0.56
0.57
0.53
0.52
0.50
0.47
0.44
2.23
2.03
1.83
1.67
1.56
1.44
1.29
1.17
1.11
1.08
1.03
0.95
0.88
0.83
0.80
0.77
0.73
0.69
0.69
0.68
0.70
0.66
0.64
0.61
0.58
0.55
3.05
2.78
2.53
2.30
2.14
1.98
1.78
1.63
1.55
1.50
1.43
1.33
1.23
1.16
1.13
1.09
1.03
0.98
0.97
0.96
0.98
0.93
0.90
0.86
0.81
0.78
River Mile 50
95th
25th Median Percentile
(mg/kg (mg/kg (mg/kg
wet wet wet
weight) weight) weight)
.75
.57
.41
.28
.17
.09
0.98
0.89
0.83
0.79
0.75
0.70
0.65
0.61
0.58
0.55
0.53
0.50
0.48
0.48
0.49
0.46
0.45
0.43
0.42
0.40
2.11
.89
.70
.53
.42
.30
.18
.08
1.01
0.97
0.92
0.86
0.80
0.75
0.71
0.68
0.65
0.62
0.60
0.58
0.60
0.56
0.55
0.53
0.52
0.49
2.86
2.57
2.30
2.08
1.92
.78
.62
.48
.39
.32
.26
.18
.10
.03
0.98
0.94
0.90
0.86
0.83
0.82
0.82
0.79
0.77
0.74
0.72
0.68
TAMS/MCA
-------
TABLE 3-15: STRIPED BASS PREDICTED TRI+ CONCENTRATIONS
FOR 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
28.66
20.43
18.03
20.95
19.34
15.36
12.85
12.15
13.57
12.87
11.02
9.12
8.50
9.72
8.85
8.43
7.14
7.59
8.33 .
7.58
8.11
7.47
6.87
6.51
5.77
5.50
36.41
26.37
22.65 .
26.88
24.47
19.19
16.58
15.55
17.67
16.19
14.37
11.61
11.04
12.45
11.49
10.69
9.25
9.74
10.89
9.75
10.62
9.66
8.92
8.17
7.36
7.16
54.77
39.23
34.33
40.12
36.70
29.49
24.80 .
23.50
26.26
24.54
21.69
17.80
16.80
19.03
17.26
16.27
14.16
14.73
16.50
14.75
15.99
14.72
13.60
12.62
11.27
10.92
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
3.90
3.40
3.06
2.81
2.73
2.46
2.06
.86
.89
.88
.72
.54
.39
.38
.35
.32
.19
.14
.20
.18
.21
.15
.09
.03
0.95
0.89
4.98
4.35
3.88
3.61
3.49
3.17
2.65
2.41
2.44
2.42
2.22
.97
.81
.79
.75
.69
.54
.47
.56
.53
.58
.49
.42
.33
1.22
1.16
7.48
6.57
5.85
5.41
5.24
4.78
4.02
3.66
3.69
3.68
3.39
3.02
2.76
2.72
2.65
2.58
2.36
2.24
2.38
2.33
2.40
2.28
2.17
2.03
1.87
1.78
-------
TABLE 3-16
EXPOSURE PARAMETERS FOR THE TREE SWALLOW (Tachycineta bicolor)
Exposure Parameters
Range Reported
for Species
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg)'
Total Daily Dietary Ingestion (kg/day wet wt.)"
Total Daily Dietary Ingestion (kg/day dry wt.)3
General Dietary Characterization
Percent Diet Composition (% wet wt.)
Fish (Total Component)
Aquatic Invertebrates (Total Component)5
Non-river Related Diet Sources
Water Consumption Rate (L/day) ft
Percent Incidental Sediment Ingestion in Diet7
Foraging Territory (km)x
Behavioral Modification Factors in the Exposure Assessment9
Temporal Migration CorrectionFactor (l-%Annual Temporal Displaceme
Temporal Hibernation/Asetivation Correction Factor (1 -%Temporal Hib/,
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Birth)
ID. n
Tree Swallow
Tachycineta
bicolor
Female Male
Adult, Breeding
0.0210 0.0206
0.018 0.018
0.005
Insectivore
0%
100%
0%
0.0044
0.00%
O.I
April - June
0.017-0.0255 (M and F)
0.016-0.020
No Contact with Sediments
0%
95.0% - 100.0%
0%
0.0038-0.0050
No Contact with Sediments
0.1-0.2
Feeds over open water habitats
April - June
Notes:' Secord and McCarly (1997). Robertson el al. (1992); 2 Estimated from Nagy (1987) and USEPA (December, 1993);' No contact with sediments;
4Secordand McCarty (1997), McCarty and Winkler(In Press)'5 Emergent forms of insects with partial aquatic life histories; 6Calderand Braun(1983 In USE
December 1993), Davis (1982); 7 Robertson et al. (1992); * McCarty and Winkler (In Press);v Robertson et al. (1992), see text for rationale;10 Bull (1998), And
(1988). I
TAMS/MCA
-------
TABLE 3-17
EXPOSURE PARAMETERS FOR THE MALLARD (Anas platyrhynchos)
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) "
Total Daily Dietary Ingestion (kg/day dry wt.)
General Dietary Characterization
Percent Diet Composition (% wet wt.) 4
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Aquatic Vegetation/Seeds
Water Consumption Rate (L/day) s
Percent Incidental Sediment Ingestion in Diet ft
Foraging Territory ( km)
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFactor ( 1 -% Annual Temporal Displaceme
Temporal Hibernation/Asetivation Correction Factor ( 1 -%Temporal Hib/
Habitat Use Factor (Temporal use factor %)
U ]M
Temporal Reproductive Period (Mating/Gestation/Birth) '
Exposure Parameters
Mallard
Anas
platyrhychos
Female Male
Adult, Breeding
1.06 1.24
0.292 0.322
0.061 0.067
Opportunistic Omnivore
0%
50%
48%
0.061 0.068
2.00%
540.0 620.0
1
1
1
February -May
Range Reported
for Species
-
-
-
-
1.01 - 1.11F/M 1.21 - 1.27
0.270-0.279 F/0.3 17-0.326 M
0.058-0.063 F/ 0.066-0.068 M
-
0%
10- 100%
8 - 90 %
0.059-0.063 F/ 0.067 - 0.069 M
2.00%
40.0- 1440.0 Ha
Resident
Active Year Round
Riparian habitats preferred
February -May
1 Dunning (1993), USEPA (December 1993); 2 Estimated from Nagy (1987) and USEPA (December 1 993); ' Estimated from USEPA (December 1993);
4 Average of diet study summaries presented in USEPA (December 1993); 5 Calder and Braun (1983 In USEPA, December 1993); * Beyer et al. (1994);
7 Kirby et al. (1985 In USEPA, December 1993);" Bull (1998), USEPA (December 1993); ""Bull (1998), Andrleand Carroll (1988).
TAMS/MCA
-------
TABLE 3-18
EXPOSURE PARAMETERS FOR BELTED KINGFISHER (Ceryle alcyon)
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) 2
Total Daily Dietary Ingestion (kg/day dry wt.) 3
General Dietary Characterization
Percent Diet Composition (% wet wt.) 4
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day) '
Percent Incidental Sediment Ingestion in Diet
Foraging Territory ( km)
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFactor (1-% Annual TemporalDisplacement)
Temporal Hibernation/Asetivation Correction Factor ( 1 -%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Hatching)9'1"
Exposure Parameters
Belted Kingfisher
Ceryle
alcyon
Female Male
Adult, Breeding
0.147 0.147
0.058 0.058
0.017 0.017
Opportunistic Piscivore
78%
22%
0%
0.016
1.00%
0.70
1
1
1
April - June
Range Reported
for Species
-
-
-
-
0. 1 36-0.1 58 M and F
0.055-0.060 M and F
-
-
46%- 100%
5% -41%
0-4.3%
0.015-0.017
nests in banks, grooming
0.389-1.03
Resident
Active Year Round
Riparian habitats preferred
April - June
1 Brooks and Davis (1987), Poole(1932); Estimated from Nagy (1987) and USEPA (December 1 993); 3 No contact with sediments;
4 Gould unpublished data (ln USEPA, December 1993), Davis (1982); ' Calderand Braun(1983 In USEPA December 1993); 6 Best Professional
Judgment based on Davis (1982); 7 Davis (1982); * Bull (1998), USEPA (December 1993); "' '" Bull (1998), Andrle and Carroll (1988).
TAMS/MCA
-------
TABLE 3-19
EXPOSURE PARAMETERS FOR GREAT BLUE HERON (Ardea herodias)
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juvenile)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) '
Total Daily Dietary Ingestion (kg/day dry wt.) '
General Dietary Characterization
Percent Diet Composition (% wet wt.)
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day) s
Percent Incidental Sediment Ingestion in Diet 6
Foraging Territory ( km)
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFactor ( 1 -%Annual Temporal Displacement)
Temporal Hibernation/ Asetivation Correction Factor ( l-%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Birth)lJ'
Exposure Parameters
Great Blue Heron
Ardea
herodias
Female Male
Adult, Breeding
2.20 2.58
0.352 0.390
0.097 0.108
Opportunistic Piscivore
98%
1%
1%
0.100 0.111
2.00%
0.98
1
1
1
March - June
Range Reported
for Species
-
-
-
-
1. 87-2.54 F/ 2.28-2.88 M
0.284-0.431 F/ 0.33 1-0.455 M
-
72-98%
1-18%
0-4.3%
0.089-0. 1 1 0 F/ 0. 1 02-0. 1 1 9 M
-
0.6-1.37
Resident
Active Year Round
Riparian habitats preferred
March -June
Notes: ' Dunning (1993) ; 2 Estimated from Nagy (1987) and USEPA (December 1993); 4 Alexander (1977 In USEPA, December 1993), Cotaam and Uhler (1945);
5 Calder and Braun (1983 In USEPA, December 1993); " Best Professional Judgement based on Eckert and Karalus (1988); 7 Peifer (1979 In USEPA (December, 1993);
* USEPA (December, 1993); 9' '" Bull (1998) and Andrle and Carroll (1988).
TAMS/MCA
-------
TABLE 3-20
EXPOSURE PARAMETERS FOR BALD EAGLE (Haliaeetus leucocephalus)
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juvenile)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) ~
Total Daily Dietary Ingestion (kg/day dry wt.) 3
General Dietary Characterization 4
Percent Diet Composition (% wet wt.) 4
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day) s
Percent Incidental Sediment Ingestion in Diet 6
Foraging Territory (km) 7
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFactor (l-%Annual Temporal Displacement)
Temporal Hibernation/ Asetivation Correction Factor ( 1 -%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Geslation/Birth)9'"1
Exposure Parameters
Bald Eagle
Haliaeetus
leucocephalus
Female Male
Adult, Breeding
5.10 3.20
0.65 0.46
-
Opportunistic Piscivore
100%
0%
0%
0.175 0.129
0.00%
5.0
1
1
I
February - May
Range Reported
for Species
-
-
-
-
4.5-5.6 F/M 3.0-3.4
0.60-0.69 F/0.46-0.49 M
-
-
70-100%
0-18%
0-4.3%
0. 1 62-0. 1 87 F/0. 123-0.1 34 M
0.00%
3.0-7.0 Km
Resident
Active Year Round
Riparian habitats preferred
February - May
' Bopp (1999), USEPA (December 1993), Dunning (1993); ', } Estimated from Nagy (1987) and USEPA (December 1993);
4Nye (1999), Bull (1998), USEPA (December 1993), Nye and Suring (1978); 5 Caluderand Braun (1983 In USEPA December 1993);
" Best Professional Judgement - USEPA (December 1993);
7 Craig et al. (1988 In USEPA, December 1993); * Nye (1999), USEPA (December 1 993); 9I" Nye (1999), Andrle and Carroll (1988).
TAMS/MCA
-------
TABLE 3-21
EXPOSURE PARAMETERS FOR LITTLE BROWN BAT (Myotis lucifugus)
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) 2
Total Daily Dietary Ingestion (kg/day dry wt.)
General Dietary Characterization
Percent Diet Composition (% wet wt.)
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day) 5
Percent Incidental Sediment Ingestion in Diet
Home Range (km) 7
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFactor (l-%Annual Temporal Displacement)
Temporal Hibernation/Asetivation Correction Factor ( 1 -%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Birth)v'
Exposure Parameters
Little Brown Bat
Myotis
lucifugus
Female Male
Adult, Breeding
0.0071 0.0069
0.0025 0.0025
-
Insectivore
0.0%
100.0%
0.0%
0.0011 0.0011
0.00% 0.00%
0. 1 >0. 1
1
1
1
April to July
Proximal Range Reported
for Species
,
-
-
-
-
0.0042-0.0094 /0.0055-0.0077
0.0025-0.0037 F/ No Male Data
-
-
0%
87.0 % - 100.0%
0%-13.0%
Based upon 0.007 Kg
0.00%
0.1 ->0.1
Resident
See text
Feeds over waterbody
April to July
1 Bopp (1999); 2 Fenton and Barclay (1980); ' Dry weight basis of ingestion not required;
4 Anthony and Kunz(l977), Belwood and Fenton (1976), Buehler (1976); ' Parrel 1 and Wood (1968c In USEPA, December 1993); " No contact
with sediments; 7 Bulcher (1976); * Davis and Hitchcock (1965); *' '" Belwood and Fenton (1976), Wimbatt (1945).
TAMS/MCA
-------
TABLE 3-22
EXPOSURE PARAMETERS FOR RACCOON (Proycon lotor)
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) "
Total Daily Dietary Ingestion (kg/day dry wt.)
General Dietary Characterization
Percent Diet Composition (% wet wt.) 4
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day) '
Percent Incidental Sediment Ingestion in Diet 6
Home Range (hectare)
Behavioral Modification Factors in the Exposure Assessment1*
Temporal Migration CorrectionFactor (l-% Annual TemporalDisplacement)
Temporal Hibernation/Asetivation Correction Factor (l-%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Birth)''' '"
Exposure Parameters
Raccoon
Procyon
lotor
Female Male
Adult, Breeding
6.400 7.600
0.99 1.20
0.316 0.364
Opportunistic Omnivore
3.0%
37.0%
60.0%
0.526 0.614
9.4% 9.4%
48.0 48.0
1
1
i
January to May
Proximal Range Reported
for Species
_
,<•
-
-
-
5.6-7.1 F/7.0-8.3M
0.866-1.1 F/l. 1-1.30 M
0.283-0.344 F/0.340-0.391 M
-
0-3%
1.4-37.0%
0-1.5%
0.467-0.578 F/0.57 1-0.665 M
9.40%
5.3-376 F/18.2-814M
Resident
Active Year Round
Riparian habitats preferred
January to May
1 Bopp (1999), Sanderson (1984), USEPA (December 1993); 2, ' Estimated from NFMR and ME in USEPA (December 1993) and Nagy (1987);
4 Tabatabai and Kennedy ) 1988), Newell el al. (1987), Llewellyn and L'hler( 1952), Hamilton (1951); 5 Farrell and Wood (1968c In USEPA, 1993a);
"Beyer el al. (1994); 7 Urban (1970), Stuewer (1943); * USEPA (December, 1993), Hamilton (1951); 9 '" USEPA (December, 1993), Stuewer (1943).
TAMS/MCA
-------
TABLE 3-23
EXPOSURE PARAMETERS FOR MINK (Mustela vison )
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg) '
Total Daily Dietary Ingestion (kg/day wet wt.) "
Total Daily Dietary Ingestion (kg/day dry wt.) *
General Dietary Characterization 4
Percent Diet Composition (% wet wt.)
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day) "
Percent Incidental Sediment Ingestion in Diet 6
Home Range (km) 7
u
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFactor (l-%Annual TemporalDisplacement)
Temporal Hibernation/Asetivation Correction Factor ( 1 -%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Birth)
Exposure Parameters
Mink
Mustela
vision
Female Male
Adult, Breeding
0.83 1.02
0.132 0.132
0.059 0.069
Opportunistic Piscivore/Carnivore
34.0%
16.5%
49.5%
0.084 0.101
1.0%
1.9 3.4
1
1
1
March to June
Proximal Range Reported
for Species
-
-
-
0.550-1.101 F/0.681-1.362M
0.145 F/ 0.1 19 M
0.042-1.013 F/0.050-0.089 M
-
18.8-34.0%
13.9-16.5%
49.5 % - 67.0 %
0.052-0.107 F/0.070-0.131 M
1.0%
1.0-2.8 km F/l. 8-5.0 km M
Resident
Active Year Round
Riparian habitats preferred
March to June
1 Mitchell (1961); J. Bopp (1999), 2 Bleavins and Aulerich (1981); ' Estimated from Nagy (1987) and USEPA (December, I993);4 Hamilton (1951),
Hamilton (1940), Hamilton (1936); 5 Farrell and Wood (1968c In USEPA, December 1993);6 Best Professional Judgement - based upon observations
in Hamilton (1940); 7Gerell (1970), Mitchell (1961); * Allen (1986).
TAMS/MCA
-------
TABLE 3-24
EXPOSURE PARAMETERS FOR RIVER OTTER (Lutra canademis)
Exposure Parameters
Proximal Range Reported
for Species
Common Name
Genus
Species
Sex (M/F)
Age (Adult/Juv.)
Male/Female Body Weight (kg)'
Total Daily Dietary Ingestion (kg/day wet wt.)2
Total Daily Dietary Ingestion (kg/day dry wt.)'
General Dietary Characterization 4
Percent Diet Composition (% wet wt.)
Fish (Total Component)
Aquatic Invertebrates (Total Component)
Non-river Related Diet Sources
Water Consumption Rate (L/day)
Percent Incidental Sediment Ingestion in Diet
Home Range (km)
Behavioral Modification Factors in the Exposure Assessment
Temporal Migration CorrectionFaclor (l-%Annual TemporalDisplacement)
Temporal Hibernation/Asetivation Correction Factor (l-%Temporal Hib/Aset.)
Habitat Use Factor (Temporal use factor %)
Temporal Reproductive Period (Mating/Gestation/Birth)9
River Otter
Lutra
canadensis
Female Male
Adult, Breeding
7.32 10.9
0.900 0.900
0.353 0.491
Opportunistic Piscivore
700%
0.0%
0.0%
0.594 0.853
1.0%
10.0
1
March to March
6.73-7.90 F/9.20-12.7M
0.7-1.1
0.329-0.376 F/0.425-0.555 M
70-100%
5-15%
0-25%
0.551-0.636 F/0.730-0.975 M
1.0%
1.5-22.3 Km
Resident
Active Year Round
Riparian habitats preferred
March to March
Spinola et al., (undated), Bopp (1999), USEPA (December 1993); ,' Harris (1968 In USEPA, December 1993), Penrod (1999);
4Spinola(1999), Newell et al. (1987), Hamilton (1961): 5 Farrell and Wood (1968c In USEPA, December 1993); 6 Best Professional Judgement -
based upon Liers(195l) In USEPA, 1993); 7 Spinola et al. (undated); * USEPA (December 1993a);9 Hamilton and Eadie (1964); "' Period between
mating and birth extends for one full year due to delayed implantation of zygote.
TAMS/MCA
-------
TABLE 3-25: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE SWALLOW BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
Sex(M/F)
1997
1998
1999
2000
2001
2002
152
.50E+00
.35E+00
.30E+00
.29E+00
.22E+00
.17E+00
.11E+00
.I1E+00
.09E+00
.04E+00
2003 9.77E-01
2004 9.62E-01
2005 9.36E-01
2006 i
2007
2008 i
2009
2010
3.99E-01
S.87E-01
S.56E-01
3.38E-01
5.25E-01
2011 7.90E-01
2012 7.70E-01
2013 7.54E-01
2014 7.46E-01
2015 7.24E-01
2016 7.31E-OI
2017 7.22E-01
2018 7.05E-01
Average Dietary Dose
(mg/Kg/day)
113 90
1.19E+00
1.12E+00
1.07E+00
1.03E+00
9.88E-01
9.61E-01
9.31E-01
8.93E-01
8.81E-01
8.50E-01
8.11E-01
7.82E-01
7.75E-01
7.52E-01
7.36E-01
7.09E-01
6.87E-01
6.74E-01
6.68E-01
6.53E-01
6.39E-01
6.23E-01
6.00E-01
5.84E-01
5.79E-01
5.76E-01
9.69E-01
9.20E-01
8.62E-01
8.21E-01
7.95E-01
7.58E-01
7.30E-01
7.10E-01
6.89E-01
6.72E-01
6.57E-01
6.23E-01
6.00E-01
5.74E-01
5.59E-01
5.42E-01
5.30E-01
5.21E-01
5.03E-01
4.92E-01
4.77E-01
4.65E-01
4.56E-01
4.46E-01
4.41E-01
4.33E-01
50
7.13E-01
6.68E-01
6.35E-01
6.11E-01
5.91E-01
5.59E-01
5.43E-01
5.27E-01
5.10E-01
5.01 E-01
4.84E-01
4.62E-01
4.44E-01
4.25E-01
4.13E-01
4.02E-01
3.94E-01
3.86E-01
3.80E-01
3.71 E-01
3.60E-01
3.51 E-01
3.42E-01
3.36E-01
3.27E-01
3.20E-01
Average Egg Concentration
(mg/Kg)
152 113 90
3.51E+00 2.79E+00 2.26E+00
3.15E+00 2.61E+00 2.15E+00
3.04E+00 2.50E+00 2.01E+00
3.00E+00 2.40E+00 1.92E+00
2.84E+00 2.31E+00 1.86E+00
2.72E+00 2.24E+00 .77E+00
2.58E+00 2.17E+00 .70E+00
2.60E+00 2.08E+00 .66E+00
2.54E+00 2.05E+00 .61E+00
2.43E+00
2.28E+00
2.24E+00
2.18E+00
2.10E+00
2.07E+00
2.00E+00
1 .96E+00
1.92E+00
1.84E+00
1.80E+00
1.76E+00
.74E+00
.69E+00
.71 E+00
.68E+00
.64E+00
.98E+00 .57E+00
.89E+00 1 .53E+00
.82E+00 1.45E+00
.81E+00
.75E+00
.72E+00
.65E+00
.60E+00
.57E+00
.56E+00
.52E+00
.49E+00
.45E+00
.40E+00
.36E+00
.35E+00
.35E+00
.40E+00
.34E+00
.30E+00
.27E+00
.24E+00
.22E+00
.17E+00
.15E+00
. 1 1 E+00
.09E+00
.06E+00
.04E+00
.03E+00
.01 E+00
50
1.66E+00
1.56E+00
1.48E+00
1.43E+00
1.38E+00
1.30E+00
1.27E+00
1 .23E+00
1.19E+00
1.17E+00
1.13E+00
1 .08E+00
1 .04E+00
9.91 E-01
9.64E-01
9.38E-01
9.18E-OI
9.01 E-01
8.86E-01
8.66E-01
8.40E-01
8.19E-01
7.99E-01
7.84E-01
7.63E-01
7.46E-01
TAMS/MCA
-------
TABLE 3-26: SUMMARY OF ADD95%UCL AND EGG CONCENTRATIONS FOR
FEMALE SWALLOW BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year 95% UCL Dietary Dose 95% UCL Egg Concentration
(mg/Kg/day) (mg/Kg)
152 113 90 50 152 113 90 50
1993 1.62E+00 .28E+00 1.04E+00 7.65E-01 3.77E+00 2.99E+00 2.43E+00
1994 1.45E+00 .20E+00 9.87E-01 7.17E-01 3.37E+00 2.80E+00 2.30E+00
1995 1.40E+00 .15E+00 9.25E-01 6.81E-01 3.26E+00 2.68E+00 2.16E+00
1996 1.38E+00 .10E+00 8.80E-01 6.55E-01 3.22E+00 2.58E+00 2.05E+00
1997 1.31 E+00 .06E+00 8.52E-01 6.33E-01 3.05E+00 2.47E+00
1998 1.25E+00 .03E+00 8.12E-01 5.99E-01 2.92E+00 2.40E+00
1999 1.19E+00 9.99E-01 7.82E-01 5.81E-01 2.77E+00 2.33E+00
2000 1.19E+00 9.59E-01 7.60E-OI 5.64E-01 2.79E+00 2.24E+00
2001 1.17E+00 9.46E-01 7.38E-OI 5.46E-01 2.72E+00 2.21E+00
2002 1.12E+00 9.13E-01 7.20E-01 5.38E-01 2.61E+00 2.13E+00
2003 1.05E+00 8.71E-01 7.05E-01 5.19E-01 2.45E+00 2.03E+00
2004 1.04E+00 8.41E-01 6.69E-01 4.96E-01 2.42E+00 1.96E+00
2005 1.01 E+00 8.33E-01 6.44E-01 4.77E-01 2.35E+00 1.94E+00
2006 9.66E-01 8.09E-01 6.17E-01 4.57E-01 2.25E+00
2007 9.54E-01 7.92E-01 6.01E-01 4.44E-01 2.23E+00
2008 9.23E-01 7.63E-01 5.83E-01 4.32E-01 2.15E+00
2009 9.04E-01 7.40E-01 5.70E-01 4.23E-01 2.11E+00
2010 8.87E-01 7.25E-01 5.60E-01 4.15E-01 2.07E+00
2011 8.49E-01 7.18E-01 5.41E-01 4.09E-01 1.98E+00
2012 8.28E-01 7.02E-01 5.28E-01 3.99E-01 1.93E+00
2013 8.10E-01 6.87E-01 5.12E-01 3.87E-01 1.89E+00
2014 8.02E-01 6.70E-01 5.00E-01 3.78E-01 1.87E+00
2015 7.81E-01 6.46E-01 4.90E-01 3.68E-01 1.82E+00
2016 7.91E-01 6.29E-01 4.80E-01 3.61E-01 1.85E+00
2017 7.82E-01 6.25E-01 4.74E-01 3.52E-01 1.82E+00
2018 7.63E-01 6.24E-01 4.66E-01 3.44E-01 1.78E+00
.89E+00
.85E+00
.78E+00
.73E+00
.69E+00
.68E+00
.64E+00
.60E+00
.56E+00
.51 E+00
.47E+00
.46E+00
.46E+00
.99E+00
.89E+00
.82E+00
.77E+00
.72E+00
.68E+00
.65E+00
.56E+00
.50E+00
.44E+00
.40E+00
.36E+00
.79E+00
.67E+00
.59E+00
.53E+00
.48E+00
.40E+00
.36E+00
.32E+00
.27E+00
.26E+00
.21 E+00
.16E+00
.11E+00
.07E+00
.04E+00
.01 E+00
.33E+00 9.87E-01
.31 E+00 9.68E-01
.26E+00 9.53E-01
.23E+00 9.32E-01
.20E+00 9.04E-01
.17E+00 8.82E-01
.14E+00 8.60E-01
.12E+00 8.43E-0!
.11E+00 8.21E-01
.09E+00 8.03E-01
TAMS/MCA
-------
TABLE 3-27: SUMMARY OF ADDExpec,ed AND EGG CONCENTRATIONS FOR
FEMALE MALLARD BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
Average Dietary Dose
(mg/Kg/day)
152 113 90
5.69E-01 4.55E-01
4.99E-01 4.12E-01
4.21E-01 3.47E-01
5.19E-01 3.58E-01
4.37E-01 3.33E-01
3.54E-01 2.87E-01
3.13E-01 2.58E-01
3.37E-01 2.52E^01
3.56E-01 2.56E-01
3.10E-01 2.40E-01
2.75E-01 2.26E-01
2.43E-01
2.39E-01
2.42E-01
2.25E-01
2.13E-01
1.90E-01
2.14E-01
1.96E-01
2.00E-01
2.18E-01
1.95E-01
1.88E-01
1.69E-01
1.63E-01
.98E-01
.92E-01
.91E-01
.86E-01
.73E-01
.61E-01
.66E-01
.66E-01
.65E-01
.66E-01
.58E-01
.51E-OI
.36E-01
.30E-01
1.71E-01 1.33E-01
3.67E-01
3.36E-01
2.95E-01
2.82E-01
2.63E-01
2.36E-01
2.15E-01
2.04E-01
.97E-01
.90E-01
.83E-01
.65E-01
.56E-01
.50E-01
.45E-01
.38E-01
1.31E-01
1.29E-01
1.27E-01
1.25E-01
.23E-01
.20E-01
.16E-01
.10E-01
.06E-01
.04E-01
50
3.30E-01
2.96E-01
2.68E-01
2.46E-01
2.28E-01
2.08E-01
1.94E-01
1.79E-01
1.69E-01
1.63E-01
1.55E-01
1.45E-01
1.35E-01
1.28E-01
1.22E-01
1.17E-01
1.13E-01
1.09E-01
1.07E-01
1.04E-01
1.02E-01
9.98E-02
9.73E-02
9.44E-02
9.10E-02
8.77E-02
152
5.26E+00
4.72E+00
4.57E+00
4.51E+00
4.27E+00
4.09E+00
3.87E+00
3.89E+00
3.81E+00
3.64E+00
3.42E+00
3.37E+00
3.27E+00
3.15E+00
3.10E+00
3.00E+00
2.93E+00
2.89E+00
2.77E+00
2.70E+00
2.64E+00
2.61E+00
2.53E+00
2.56E+00
2.53E+00
2.47E+00
Average Egg Concentration
(mg/Kg)
113 90 50
4.18E+00
3.91E+00
3.75E+00
3.61E+00
3.46E+00
3.36E+00
3.26E+00
3.13E+00
3.08E+00
2.97E+00
2.84E+00
2.74E+00
2.71E+00
2.63E+00
2.58E+00
2.48E+00
2.41E+00
2.36E+00
2.34E+00
2.29E+00
2.24E+00
2.18E+00
2.10E+00
2.04E+00
2.02E+00
2.02E+00
3.39E+00 2.49E+00
3.22E+00 2.34E+00
3.02E+00 2.22E+00
2.87E+00 2.14E+00
2.78E+00 2.07E+00
2.65E+00
2.56E+00
2.49E+00
2.41E+00
2.35E+00
2.30E+00
2.18E+00
2.10E+00
2.01E+00
1 .96E+00
1.90E+00
.86E+00
.82E+00
.76E+00
.72E+00
.96E+00
.90E+00
.84E+00
.79E+00
.75E+00
.69E+00
.62E+00
.56E+00
.49E+00
.45E+00
..41E+00
.38E+00
.35E+00
.33E+00
.30E+00
.67E+00 1.26E+00
.63E+00 1 .23E+00
.60E+00 1 .20E+00
.56E+00 1.18E+00
1.54E+00 I.15E+00
1.51E+00 1.12E+00
TAMS/MCA
-------
TABLE 3-28: SUMMARY OF ADD95%UCL AND EGG CONCENTRATIONS FOR
FEMALE MALLARD BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
6.10E-01
5.34E-01
4.51E-01
5.57E-01
4.68E-01'
3.80E-01
3.36E-01
3.62E-01
3.82E-01
3.33E-01
2.95E-01
2.62E-01
2.57E-01
2.60E-01
2.42E-01
2.30E-01
2.04E-01
2.30E-01
2.11E-01
2.15E-01
2.34E-01
2.09E-01
2.02E-01
1.82E-01
1.77E-OI
1.84E-OI
95% UCL Dietary Dose
(mg/Kg/day)
113 90
4.88E-01
4.42E-01
3.72E-OI
3.84E-01
3.57E-01
3.07E-01
2.76E-01
2.70E-01
2.75E-01
2.58E-01
2.42E-01
2.12E-01
2.07E-01
2.06E-01
2.00E-01
1.87E-01
1.74E-01
1.79E-01
1.78E-01
1.77E-01
1.78E-01
1.70E-01
1.62E-01
1.47E-01
1.40E-0!
I.44E-01
3.94E-01
3.60E-01
3.14E-01
3.01E-01
2.82E-01
2.53E-01
2.30E-01
2.19E-01
2.11E-01
2.04E-01
1.96E-01
1.78E-01
1.67E-01
1.61E-01
1.56E-01
1.48E-01
1.41E-01
1.39E-01
I.36E-01
I.35E-OI
1.32E-01
1.29E-01
1.25E-01
1.18E-01
1.14E-01
I.11E-01
50
3.54E-01
3.17E-01
2.87E-01
2.64E-01
2.45E-01
2.23E-01
2.08E-01
1.92E-01
1.81E-01
1.75E-01
1.66E-01
1.55E-OI
1.45E-01
1.37E-01
1.31E-01
1.26E-01
I.21E-01
1.17E-01
1.14E-01
1.12E-01
1.09E-01
1.07E-01
1.05E-01
1.01E-01
9.78E-02
9.43E-02
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
5.65E+00
5.06E+00
4.89E+00
4.83E+00
4.57E+00
4.38E+00
4.16E+00
4.18E+00
4.08E+00
3.91E+00
3.68E+00
3.62E+00
3.52E+00
3.38E+00
3.34E+00
3.23E+00
3.16E+00
3.10E+00
2.97E+00
2.90E+00
2.83E+00
2.81E+00
2.73E+00
2.77E+00
2.74E+00
2.67E+00
4.48E+00
4.19E+00
4.02E+00
3.87E+00
3.71E+00
3.60E+00
3.50E+00
3.36E+00
3.31E+00
3.19E+00
3.05E+00
2.94E+00
2.92E+00
2.83E+00
2.77E+00
2.67E+00
2.59E+00
2.54E+00
2.51E+00
2.46E+00
2.40E+00
2.35E+00
2.26E+00
2.20E+00
2.19E+00
2.18E+00
3.65E+00
3.45E+00
3.24E+00
3.08E+00
2.98E+00
2.84E+00
2.74E+00
2.66E+00
2.58E+00
2.52E+00
2.47E+00
2.34E+00
2.25E+00
2.16E+00
2.10E+00
2.04E+00
2.00E+00
1.96E+00
1.89E+00
1.85E+00
1.79E+00
1.75E+00
1 .72E+00
1 .68E+00
I.66E+00
1 .63E+00
2.68E+00
2.51E+00
2.38E+00
2.29E+00
2.22E+00
2.IOE+00
2.03E+00
1.97E+00
1.91E+00
1.88E+00
1.82E+00
1.74E+00
1.67E+00
1.60E+00
1.55E+00
1.51E+00
1.48E+00
1.45E+00
1.43E+00
1.40E+00
1.36E+00
1.32E+00
1.29E+00
1 .26E+00
I.23E+00
1.20E+00
TAMS/MCA
-------
TABLE 3-29: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE BELTED KINGFISHER BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
6.67E-01
5.22E-01
4.74E-01
5.44E-01
4.73E-01
3.77E-01
3.41E-01
3.35E-01
3.58E-01
3.24E-01
3.01E-01
2.55E-01
2.49E-01
2.75E-01
2.40E-01
2.26E-01
2.12E-01
2.23E-01
2.32E-01
2.19E-01
2.28E-01
2.17E-01
1.97E-01
1.78E-01
1.73E-01
1.72E-01
Average Dietary Dose
(mg/Kg/day)
113 90
4.68E-01
4.28E-01
3.66E-01
3.72E-01
3.49E-01
3.03E-01
2.76E-01
2.58E-01
2.64E-01
2.56E-01
2.37E-01
2.12E-01
2.02E-01
2.03E-01
1.96E-01
1.86E-01
1.74E-01
1.71E-01
1.76E-01
1.73E-01
1.73E-01
1.67E-01
1.58E-01
1.47E-01
1.40E-01
1.38E-01
3.76E-01
3.45E-01
3.15E-01
2.96E-01
2.72E-01
2.53E-01
2.27E-01
2.13E-01
2.03E-01
1.98E-01
1.89E-01
1.75E-01
1.64E-01
1.56E-01
1.51E-01
1.46E-01
1.39E-01
1.33E-01
1.31E-01
1.31E-01
1.29E-01
1.25E-01
1.21E-01
1.17E-01
1.12E-01
1.08E-01
50
3.34E-01
3.02E-01
2.73E-01
2.53E-01
2.35E-01
2.19E-01
2.00E-01
1.86E-01
1.74E-01
1.68E-01
1.60E-01
1.50E-01
1.40E-01
1.32E-01
1.27E-01
1.22E-01
1.17E-01
1.12E-01
1.10E-01
1.08E-01
1.06E-01
1.03E-01
l.OOE-01
9.74E-02
9.37E-02
9.10E-02
152
5.05E+01
3.94E+01
3.58E+01
4.11E+01
3.57E+01
2.84E+01
2.57E+01
2.53E+01
2.71E+01
2.45E+01
2.27E+01
1.92E+01
1.87E+01
2.07E+01
1.81E+01
1.70E+01
1.60E+01
1.68E+01
1.75E+01
1.65E+01
1.72E+01
1.64E+01
1.49E+01
1.34E+01
1.30E+01
1.30E+01
Average Egg Concentration
(mg/Kg)
113 90
3.54E+01
3.23E+01
2.76E+01
2.81E+01
2.63E+01
2.28E+OI
2.08E+01
1.95E+01
I.99E+01
1.93E+01
1.78E+01
1.60E+01
1.52E+01
1.53E+01
1.48E+01
1.40E+01
1.31E+01
1.29E+01
1.33E+01
1.30E+01
1.30E+01
1.26E+01
1.19E+01
1.11E+01
1.05E+01
1.04E+01
2.84E+01
2.60E+01
2.33E+01
2.18E+01
2.05E+01
1.91E+01
1.7IE+01
I.60E+01
1.53E+01
1.49E+01
1.43E+01
1.32E+01
1.24E+01
1.18E+01
1.13E+01
1.10E+01
1.05E+01
9.99E+00
9.90E+00
9.85E+00
9.73E+00
9.41E+00
9.15E+00
8.79E+00
8.41E+00
8.16E+00
50
2.53E+01
2.28E+01
2.06E+01
1.91E+01
1.77E+01
I.66E+01
1.51E+01
1.40E+01
1.32E+01
1.26E+01
1.21E+01
1.13E+01
1.06E+01
9.97E+00
9.55E+00
9.19E+00
8.81E+00
8.46E+00
8.26E+00
8.17E+00
8.02E+00
7.76E+00
7.57E+00
7.34E+00
7.06E+00
6.86E+00
TAMS/MCA
-------
TABLE 3-30: SUMMARY OF ADD9S%UCL AND EGG CONCENTRATIONS FOR
FEMALE BELTED KINGFISHER BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
6.93E-01
5.42E-01
4.95E-01
5.65E-01
4.92E-01
3.94E-01
3.57E-01
3.51E-01
3.74E-01
3.40E-01
3.15E-01
2.68E-01
2.62E-01
2.88E-01
2.53E-01
2.38E-01
2.24E-01
2.35E-01
2.43E-01
2.30E-01
2.39E-01
2.28E-01
2.08E-01
1.89E-01
1.83E-01
1.83E-OI
95% UCL Dietary Dose
(mg/Kg/day)
113 90
4.87E-01
4.45E-01
3.82E-01
3.87E-01
3.64E-01
3.17E-01
2.89E-01
2.70E-01
2.76E-01
2.68E-01
2.48E-01
2.23E-01
2.13E-01
2.13E-01
2.06E-01
1.96E-01
1.84E-01
1.80E-01
1.85E-01
1.82E-01
1.82E-01
1.75E-01
1.66E-01
1.55E-01
1.48E-01
1.46E-01
3.92E-01
3.59E-01
3.21E-01
3.02E-01
2.84E-01
2.65E-01
2.38E-01
2.23E-01
2.13E-01
2.07E-01
1.98E-01
1.84E-01
1.72E-01
1.64E-01
1.58E-01
1.53E-01
1.46E-01
1.40E-01
1.38E-01
1.38E-01
1.36E-01
1.31E-01
1.28E-01
1.23E-OI
1.18E-OI
1.14E-01
50
3.47E-01
3.14E-01
2.84E-01
2.63E-01
2.44E-01
2.28E-01
2.08E-01
I.94E-01
1.82E-01
1.75E-01
1.67E-01
1.57E-01
1.47E-01
1.38E-OI
1.33E-01
1.28E-01
1.23E-01
1.18E-01
1.15E-01
1.14E-01
1.12E-01
1.08E-01
1.05E-01
I.02E-01
9.84E-02
9.56E-02
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
5.24E+01
4.09E+01
3.73E+01
4.27E+01
3.71E+01
2.97E+01
2.69E+01
2.64E+01
2.82E+01
2.56E+01
2.37E+01
2.01E+01
1.96E+01
2.17E+01
1.90E+01
1.79E+01
1.68E+01
1.76E+01
1.83E+01
1.73E+01
1.80E+OI
1.71E+01
1.56E+01
1.42E+01
I.37E+01
1.37E+01
3.68E+01
3.36E+01
2.88E+01
2.92E+01
2.75E+01
2.38E+01
2.18E+01
2.03E+01
2.08E+01
2.02E+01
1.87E+01
1.67E+01
1.60E+01
1.60E+01
1.55E+01
1.47E+01
1.38E+01
1.35E+01
1.39E+01
1.36E+01
1.37E+01
1.32E+01
1.25E+01
1.16E+01
1.11E+01
1.09E+01
2.96E+01
2.71E+01
2.42E+01
2.28E+01
2.14E+01
1.99E+01
1.79E+01
1.68E+01
1.60E+01
1.56E+01
1.49E+01
1.38E+01
1.30E+01
1.23E+01
1.19E+01
1.15E+OI
1.10E+01
1.05E+01
1.04E+01
1.03E+01
1.02E+01
9.86E+00
9.59E+00
9.22E+00
8.83E+00
8.56E+00
2.63E+01
2.37E+01
2.14E+01
1.99E+01
1.84E+01
1.72E+01
1.57E+01
1.46E+01
1.37E+01
1.32E+01
1.26E+01
1.18E+01
I.11E+01
1.04E+01
9.98E+00
9.60E+00
9.21E+00
8.84E+00
8.64E+00
8.55E+00
8.39E+00
8.11E+00
7.92E+00
7.68E+00
7.39E+00
7.18E+00
TAMS/MCA
-------
TABLE 3-31: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE GREAT BLUE HERON BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
2.61E-01
.94E-01
.73E-01
2.09E-01
.76E-01
.30E-01
.15E-01
.12E-01
.25E-01
.10E-01
.01E-01
7.85E-02
7.67E-02
9.23E-02
7.51E-02
6.94E-02
6.31E-02
6.97E-02
7.59E-02
7.07E-02
7.62E-02
7.10E-02
6.19E-02
. 5.19E-02
4.94E-02
5.01E-02
Average Dietary Dose
(mg/Kg/day)
113 90
.75E-01
.59E-01
.29E-01
.35E-01
.25E-01
.03E-01
9.11E-02
8.39E-02
8.75E-02
8.52E-02
7.74E-02
6.62E-02
6.16E-02
6.32E-02
6.07E-02
5.70E-02
5.19E-02
5.12E-02
5.41E-02
5.31E-02
5.41E-02
5.17E-02
4.84E-02
4.36E-02
4.04E-02
3.93E-02
1.40E-01
1.27E-01
1.13E-01
1.05E-01
9.65E-02
8.88E-02
7.69E-02
7.06E-02
6.68E-02
6.51E-02
6.13E-02
5.62E-02
5.16E-02
4.88E-02
4.69E-02
4.53E-02
4.24E-02
3.97E-02
4.01E-02
4.04E-02
4.04E-02
3.88E-02
3.75E-02
3.56E-02
3.33E-02
3.21E-02
50
1.33E-01
1.19E-01
1.06E-01
9.66E-02
8.84E-02
8.22E-02
7.31E-02
6.67E-02
6.17E-02
5.88E-02
5.57E-02
5.17E-02
4.79E-02
4.47E-02
4.25E-02
4.07E-02
3.86E-02
3.66E-02
3.56E-02
3.54E-02
3.50E-02
3.37E-02
3.29E-02
3.17E-02
3.03E-02
2.93E-02
152
4.98E+01
3.71E+01
3.29E+01
4.00E+01
3.37E+01
2.48E+01
2.19E+01
2.13E+01
2.38E+01
2.10E+01
1.93E+01
1.49E+01
1.46E+01
1.76E+01
1.43E+01
1.32E+01
1.20E+01
1.33E+01
1.45E+01
1.35E+01
1.45E+01
1.35E+01
1.18E+01
9.86E+00
9.39E+00
9.54E+00
Average Egg Concentration
(mg/Kg)
113 90
3.35E+01
3.03E+01
2.47E+01
2.57E+01
2.39E+01
.97E+01
.74E+01
.60E+01
.67E+01
.62E+01
.47E+01
.26E+01
.17E+01
1.20E+01
1.15E+01
1.08E+01
9.87E+00
9.73E+00
1.03E+01
1.01E+01
1.03E+01
9.84E+00
9.20E+00
8.29E+00
7.68E+00
7.47E+00
2.68E+01
2.42E+01
2.12E+01
.98E+01
.84E+01
.69E+01
.47E+OI
.34E+01
.27E+01
.24E+01
.17E+01
.07E+01
9.82E+00
9.29E+00
8.93E+00
8.62E+00
8.06E+00
7.56E+00
7.63E+00
7.69E+00
7.69E+00
7.39E+00
7.14E+00
6.77E+00
6.34E+00
6.10E+00
50
2.54E+01
2.27E+01
2.01E+01
1.84E+01
1.69E+01
1.57E+01
1.39E+01
1.27E+01
1.18E+01
1.12E+01
1.06E+01
9.86E+00
9.13E+00
8.52E+00
8.11E+00
7.75E+00
7.35E+00
6.96E+00
6.78E+00
6.75E+00
6.67E+00
6.42E+00
6.27E+00
6.04E+00
5.77E+00
5.59E+00
TAMS/MCA
-------
TABLE 3-32: SUMMARY OF ADD95%UCL AND EGG CONCENTRATIONS FOR
FEMALE GREAT BLUE HERON BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
2.68E-01
2.00E-01
1.78E-01
2.15E-01
1.82E-01
1.35E-01
1.19E-01
1.16E-01
1.29E-01
1.14E-01
1.05E-01
8.22E-02
8.03E-02
9.59E-02
7.85E-02
7.28E-02
6.63E-02
7.29E-02
7.90E-02
7.38E-02
7.92E-02
7.40E-02
6.47E-02
5.47E-02
5.21E-02
5.28E-02
95% UCL Dietary Dose
(mg/Kg/day)
113 90
.81E-01
.64E-01
.34E-01
.39E-01
.30E-01
.07E-01
9.49E-02
8.74E-02
9.10E-02
8.87E-02
8.06E-02
6.93E-02
6.45E-02
6.61E-02
6.35E-02
5.97E-02
5.46E-02
5.37E-02
5.67E-02
5.56E-02
5.66E-02
5.41E-02
5.06E-02
4.58E-02
4.26E-02
4.14E-02
1.45E-01
1.31E-01
1.15E-01
1.08E-01
l.OOE-01
9.21E-02
8.00E-02
7.34E-02
6.95E-02
6.78E-02
6.39E-02
5.86E-02
5.40E-02
5.10E-02
4.90E-02
4.74E-02
4.44E-02
4.17E-02
4.20E-02
4.23E-02
4.22E-02
4.06E-02
3.93E-02
3.73E-02
3.50E-02
3.37E-02
95% UCL Egg Concentration
(mg/Kg)
50 152 113 90
1.37E-01 5.11E+01 3.44E+01
1.22E-01 3.81E+01 3.12E+01
1.09E-01 3.39E+01 2.54E+01
9.96E-02 4.10E+01 2.64E+01
9.12E-02 3.46E+01 2.46E+01
8.48E-02 2.55E+01 2.02E+01
7.56E-02 2.25E+01 1.79E+01
6.90E-02 2.19E+01 1.64E+01
6.39E-02 2.45E+01
6.09E-02 2.16E+01
5.77E-02
5.37E-02
4.98E-02
4.65E-02
4.42E-02
4.23E-02
4.02E-02
3.81E-02
3.71E-02
3.69E-02
3.65E-02
3.51E-02
3.43E-02
.98E+01
.54E+01
.50E+01
.81E+01
.47E+01
.36E+OI
.24E+01
.37E+01
.49E+01
.39E+01
.49E+01
.39E+01
.71E+01
.67E+01
.52E+01
.30E+01
.21E+01
.24E+OI
.19E+01
.12E+01
.02E+01
.OOE+01
.06E+OI
.04E+01
.06E+01
.01E+01
.21E+01 9.48E+00
3.31E-02 1.02E+01 8.55E+00
3.16E-02 9.69E+00 7.92E+00
3.06E-02 9.84E+00 7.70E+00
2.76E+01
2.49E+01
2.19E+01
2.04E+01
1.89E+01
1.74E+01
.51E+OI
.38E+01
.31E+01
.28E+01
.20E+01
.10E+01
.01E+01
9.57E+00
9.19E+00
8.89E+00
8.31E+00
7.79E+00
7.85E+00
7.92E+00
7.92E+00
7.61E+00
7.35E+00
6.98E+00
6.54E+00
6.28E+00
50
2.61E+01
2.33E+01
2.07E+01
.90E+01
.74E+01
.61E+01
.44E+01
.31E+01
.21E+01
1.15E+01
1.09E+01
.02E+01
9.40E+00
8.77E+00
8.34E+00
7.98E+00
7.56E+00
7.17E+00
6.98E+00
6.94E+00
6.87E+00
6.61E+00
6.45E+00
6.22E+00
5.94E+00
5.75E+00
TAMS/MCA
-------
TABLE 3-33: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE EAGLE BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.90E+00
1.38E+00
1.18E+00
1.40E+00
1.27E+00
l.OOE+00
8.65E-01
8.12E-01
9.22E-01
8.43E-01
7.52E-01
6.06E-01
5.78E-01
6.51E-01
6.00E-01
5.58E-01
4.84E-01
5.09E-01
5.70E-01
5.09E-01
5.56E-01
5.05E-01
4.67E-01
4.27E-01
3.84E-01
3.75E-01
Average Dietary Dose
(mg/Kg/day)
113 90
1.27E+00 2.91E-01
1.11E+00 2.65E-01
9.86E-01 2.39E-01
9.20E-01 2.18E-01
8.89E-01 2.03E-01
8.09E-01
6.77E-01
6.16E-01
6.24E-01
6.17E-01
5.67E-01
5.03E-01
4.63E-01
4.57E-01
4.47E-01
4.30E-01
.87E-01
.69E-01
.53E-01
.45E-01
.41E-01
.34E-01
.25E-01
.15E-01
.09E-01
.05E-01
.01E-01
3.93E-01 9.58E-02
3.76E-01 9.09E-02
3.99E-01 9.01E-02
3.90E-01 8.93E-02
4.03E-01 9.17E-02
3.82E-01 8.67E-02
3.64E-01 8.41E-02
3.39E-01 8.05E-02
3.13E-01 7.63E-02
2.97E-01 7.23E-02
50
2.74E-01
2.46E-01
2.21E-01
2.00E-01
.85E-01
.69E-01
.54E-01
.42E-01
.32E-01
.26E-01
.20E-01
.12E-01
.04E-01
9.77E-02
9.30E-02
8.91E-02
8.47E-02
8.06E-02
7.82E-02
7.65E-02
7.82E-02
7.39E-02
7.20E-02
6.98E-02
6.77E-02
6.39E-02
Average Egg Concentration
(rag/Kg)
152 113 90
4.17E+02
3.02E+02
2.59E+02
3.08E+02
2.80E+02
2.20E+02
1.90E+02
1.78E+02
2.03E+02
1.85E+02
1 .65E+02
1.33E+02
1 .27E+02
1.43E+02
1.32E+02
1.23E+02
1.06E+02
1.12E+02
1.25E+02
1.12E+02
1.22E+02
1.11E+02
1.03E+02
9.38E+01
8.44E+01
8.24E+01
2.79E+02
2.44E+02
2.17E+02
2.02E+02
.95E+02
.78E+02
.49E+02
.35E+02
.37E+02
.36E+02
.25E+02
.11E+02
.02E+02
l.OOE+02
9.83E+01
9.45E+01
8.64E+01
8.25E+01
8.78E+01
8.57E+01
8.85E+01
8.39E+01
8.00E+01
7.45E+OI
6.87E+01
6.52E+01
6.40E+01
5.82E+01
5.25E+01
4.78E+01
4.47E+01
4.12E+01
3.71E+01
3.36E+01
3.19E+01
3.10E+01
2.95E+01
2.74E+01
2.53E+01
2.39E+01
2.31E+01
2.23E+01
2.10E+01
2.00E+01
.98E+01
.96E+01
2.01E+01
.90E+01
.85E+01
.77E+01
1.68E+01
1.59E+01
50
6.03E+01
5.40E+01
4.86E+01
4.39E+01
4.06E+01
3.72E+01
3.39E+01
3.11E+01
2.90E+01
2.77E+01
2.63E+01
2.46E+01
2.28E+01
2.15E+OI
2.04E+01
.96E+01
1.86E+01
1.77E+01
1.72E+01
.68E+01
.72E+01
.62E+01
.58E+OI
.53E+01
.49E+01
1.40E+01
TAMS/MCA
-------
TABLE 3-34: SUMMARY OF ADD95%UCL AND EGG CONCENTRATIONS FOR
FEMALE EAGLE BASED ON TRI+ CONGENERS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
' 2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.94E+00
1.40E+00
1.21E+00
1.43E+00
1.30E+00
1.02E+00
8.84E-01
8.29E-01
9.42E-01
8.62E-01
7.68E-01
6.20E-01
5.91E-01
6.65E-01
6.14E-01
5.70E-01
4.95E-01
5.20E-01
5.83E-01
5.21E-01
5.68E-01
5.17E-01
4.78E-01
4.36E-01
3.93E-01
3.84E-01
95% UCL Dietary Dose
(mg/Kg/day)
113 90
1.30E+00
1.13E+00
1.01E+00
9.40E-01
9.08E-01
8.27E-01
6.92E-01
6.30E-01
6.38E-01
6.31E-01
5.80E-01
5.14E-01
4.73E-01
4.67E-01
4.57E-01
4.40E-01
4.02E-01
3.84E-01
4.09E-01
3.99E-01
4.12E-01
3.91E-01
3.72E-01
3.47E-01
3.20E-01
3.04E-OI
2.96E-01
2.69E-01
2.43E-01
2.21E-01
2.07E-01
1.90E-01
1.72E-01
1.56E-01
1.48E-01
1.44E-01
1.37E-01
1.27E-01
1.17E-01
1.11E-01
1.07E-01
1.03E-OI
9.75E-02
9.25E-02
9.18E-02
9.09E-02
9.33E-02
8.82E-02
8.56E-02
8.19E-02
7.76E-02
7.36E-02
50
2.79E-01
2.50E-01
2.25E-01
2.03E-01
1.88E-01
1.72E-01
1.57E-01
1.44E-01
1.34E-01
1.28E-01
1.22E-01
1.14E-01
1.06E-01
9.94E-02
9.46E-02
9.06E-02
8.61E-02
8.21E-02
7.96E-02
7.78E-02
7.95E-02
7.53E-02
7.32E-02
7.11E-02
6.89E-02
6.50E-02
95% UCL Egg Concentration
(mg/Kg)
152 113 90
4.26E+02
3.09E+02
2.65E+02
3.15E+02
2.86E+02
2.25E+02
1.94E+02
1.82E+02
2.07E+02
.89E+02
.69E+02
.36E+02
.30E+02
.46E+02
.35E+02
.25E+02
.09E+02
.14E+02
.28E+02
.14E+02
1.25E+02
1.14E+02
1.05E+02
9.58E+01
8.63E+01
8.43E+01
2.85E+02
2.49E+02
2.21E+02
2.06E+02
1.99E+02
1.82E+02
1.52E+02
1.38E+02
.40E+02
.39E+02
.27E+02
.13E+02
.04E+02
.03E+02
.OOE+02
9.66E+01
8.83E+01
8.44E+01
8.97E+01
8.77E+01
9.05E+01
8.58E+01
8.18E+01
7.62E+01
7.02E+01
6.67E+01
6.50E+01
5.92E+01
5.34E+01
4.86E+01
4.54E+01
4.18E+01
3.77E+01
3.42E+01
3.25E+01
3.16E+01
3.00E+01
2.79E+01
2.57E+01
2.44E+01
2.35E+01
2.27E+01
2.14E+OI
2.03E+01
2.02E+01
2.00E+01
2.05E+01
1.94E+01
1.88E+01
1.80E+01
1.71E+01
1.62E+OI
50
6.12E+01
5.48E+01
4.93E+01
4.46E+01
4.12E+01
3.78E+01
3.45E+01
3.16E+01
2.95E+01
2.82E+01
2.67E+01
2.50E+01
2.32E+01
2.18E+01
2.08E+01
.99E+OI
.89E+01
.80E+01
.75E+01
.71E+01
.75E+01
.65E+OI
.61E+01
.56E+01
1.51E+OI
1.43E+01
TAMS/MCA
-------
TABLE 3-35: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE TREE SWALLOW FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Total Average Dietary Dose
Year (mg/Kg/day)
152 113 90
1993 2.08E-04
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
.87E-04
.81E-04
.78E-04
.69E-04
.62E-04
.53E-04
.54E-04
.51E-04
.44E-04
.35E-04
.33E-04
.30E-04
.25E-04
2007 1.23E-04
.65E-04
.55E-04
.49E-04
.43E-04
.37E-04
.33E-04
.29E-04
.24E-04
.22E-04
.18E-04
.12E-04
.08E-04
.07E-04
.04E-04
.02E-04
2008 1.19E-04 9.82E-05
2009 1.16E-04 9.53E-05
2010 1.14E-04 9.34E-05
2011 1.10E-04 9.26E-05
2012 1.07E-04 9.05E-05
2013 1.04E-04 8.85E-05
2014 1.03E-04 8.63E-05
2015 l.OOE-04 8.32E-05
2016 1.01E-04 8.09E-05
2017 l.OOE-04 8.02E-05
2018 9.77E-05 7.99E-05
.34E-04
.27E-04
.19E-04
.14E-04
.10E-04
.05E-04
.01E-04
9.84E-05
9.55E-05
9.32E-05
9.11E-05
8.63E-05
8.31E-05
7.95E-05
7.74E-05
7.51E-05
7.35E-05
7.22E-05
6.98E-05
6.81E-05
6.61E-05
6.45E-05
6.32E-05
6.19E-05
6.11E-05
6.00E-05
50
9.88E-05
9.26E-05
8.80E-05
8.47E-05
8.19E-05
7.74E-05
7.52E-05
7.30E-05
7.07E-05
6.95E-05
6.70E-05
6.40E-05
6.16E-05
5.89E-05
5.72E-05
5.57E-05
5.46E-05
5.35E-05
5.26E-05
5.15E-05
4.99E-05
4.87E-05
4.75E-05
4.65E-05
4.53E-05
4.43E-05
152
1.70E-03
1.53E-03
1.48E-03
1.46E-03
1.38E-03
1.32E-03
1.25E-03
1.26E-03
1.23E-03
1.18E-03
1.11E-03
1.09E-03
1.06E-03
1 .02E-03
l.OOE-03
9.69E-04
9.49E-04
9.33E-04
8.94E-04
8.72E-04
8.53E-04
8.44E-04
8.19E-04
8.27E-04
8.17E-04
7.98E-04
Average Egg Concentration
(mg/Kg)
113 90
.35E-03
.27E-03
.21E-03
.17E-03
.12E-03
.09E-03
.05E-03
.01E-03
9.97E-04
9.62E-04
9.17E-04
8.85E-04
8.77E-04
8.51E-04
8.33E-04
8.02E-04
7.78E-04
7.63E-04
7.56E-04
7.39E-04
7.23E-04
7.05E-04
6.79E-04
6.60E-04
6.55E-04
6.52E-04
1.10E-03
1.04E-03
9.75E-04
9.29E-04
9.00E-04
8.57E-04
8.26E-04
8.04E-04
7.80E-04
7.61E-04
7.44E-04
7.05E-04
6.79E-04
6.49E-04
6.32E-04
6.14E-04
6.00E-04
5.90E-04
5.70E-04
5.56E-04
5.39E-04
5.26E-04
5.16E-04
5.05E-04
4.99E-04
4.90E-04
50
8.07E-04
7.56E-04
7.19E-04
6.91E-04
6.69E-04
6.32E-04
6.14E-04
5.96E-04
5.77E-04
5.67E-04
5.47E-04
5.23E-04
5.03E-04
4.81E-04
4.67E-04
4.55E-04
4.45E-04
4.37E-04
4.30E-04
4.20E-04
4.07E-04
3.97E-04
3.88E-04
3.80E-04
3.70E-04
3.62E-04
TAMS/MCA
-------
TABLE 3-36: SUMMARY OF ADD9S%UCL AND EGG CONCENTRATIONS FOR
FEMALE TREE SWALLOW FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
Total 95% UCL Dietary Dose
(mg/Kg/day)
152 113 90 50
1993 2.24E-04 1.78E-04
1994 2.00E-04
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
.94E-04
.91E-04
.81E-04
.73E-04
.65E-04
.65E-04
.62E-04
.55E-04
.46E-04
.43E-04
.40E-04
.34E-04
.32E-04
.28E-04
.25E-04
.23E-04
.66E-04
.59E-04
.53E-04
.47E-04
.43E-04
.38E-04
.33E-04
.31E-04
.26E-04
.21E-04
.16E-04
.15E-04
.12E-04
.10E-04
.06E-04
.03E-04
.01E-04
.18E-04 9.96E-05
.I5E-04 9.74E-05
.12E-04 9.52E-05
.11E-04 9.29E-05
.08E-04 8.96E-05
.10E-04 8.72E-05
.08E-04 8.66E-05
.06E-04 8.65E-05
.44E-04
.37E-04
.28E-04
.22E-04
.18E-04
.13E-04
.08E-04
.05E-04
.02E-04
9.98E-05
9.77E-05
9.27E-05
8.93E-05
8.55E-05
8.32E-05
8.08E-05
7.90E-05
7.76E-05
7.50E-05
7.32E-05
7.10E-05
6.93E-05
6.79E-05
6.65E-05
6.57E-05
6.45E-05
1.06E-04
9.94E-05
9.43E-05
9.07E-05
8.78E-05
8.30E-05
8.06E-05
7.81E-05
7.57E-05
7.45E-05
7.20E-05
6.87E-05
6.61E-05
6.33E-05
6.15E-05
5.99E-05
5.86E-05
5.75E-05
5.66E-05
5.54E-05
5.37E-05
5.24E-05
5.11E-05
5.01E-05
4.88E-05
4.77E-05
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
.83E-03
.64E-03
.58E-03
.56E-03
.48E-03
.42E-03
.34E-03
.35E-03
1.32E-03
1.26E-03
.19E-03
.17E-03
.14E-03
.09E-03
.08E-03
.05E-03
.02E-03
.OOE-03
9.61E-04
9.37E-04
9.17E-04
9.07E-04
8.84E-04
8.96E-04
8.85E-04
8.64E-04
1.45E-03
1.36E-03
1.30E-03
1.25E-03
1.20E-03
1.16E-03
1.13E-03
1.09E-03
1.07E-03
1.03E-03
9.85E-04
9.51E-04
9.43E-04
9.15E-04
8.96E-04
8.64E-04
8.38E-04
8.21E-04
8.13E-04
7.95E-04
7.78E-04
7.58E-04
7.31E-04
7.12E-04
7.07E-04
7.06E-04
1.18E-03
1.12E-03
1.05E-03
9.96E-04
9.64E-04
9.19E-04
8.85E-04
8.60E-04
8.35E-04
8.15E-04
7.98E-04
7.57E-04
7.29E-04
6.98E-04
6.80E-04
6.60E-04
6.45E-04
6.34E-04
6.12E-04
5.98E-04
5.80E-04
5.66E-04
5.55E-04
5.43E-04
5.37E-04
5.27E-04
8.66E-04
8.12E-04
7.70E-04
7.41E-04
7.17E-04
6.78E-04
6.58E-04
6.38E-04
6.18E-04
6.09E-04
5.88E-04
5.61E-04
5.40E-04
5.17E-04
5.03E-04
4.89E-04
4.79E-04
4.70E-04
4.62E-04
4.52E-04
4.38E-04
4.28E-04
4.17E-04
4.09E-04
3.98E-04
3.90E-04
TAMS/MCA
-------
TABLE 3-37: SUMMARY OF ADDExpecled AND EGG CONCENTRATIONS FOR
FEMALE MALLARD ON A TEQ BASIS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
2.27E-04
.97E-04
.54E-04
2.13E-04
.70E-04
.24E-04
.05E-04
.19E-04
.31E-04
.07E-04
9.14E-05
7.39E-05
7.32E-05
7.77E-05
6.87E-05
6.40E-05
5.13E-05
6.63E-05
5.85E-05
6.25E-05
7.39E-05
6.09E-05
5.85E-05
4.66E-05
4.42E-05
4.97E-05
Average Dietary Dose
(mg/Kg/day)
113 90
1.82E-04
1.62E-04
1.27E-04
1.37E-04
1.25E-04
9.96E-05
8.46E-05
8.38E-05
8.76E-05
8.01E-05
7.45E-05
6.00E-05
5.74E-05
5.85E-05
5.65E-05
5.10E-05
4.55E-05
4.92E-05
4.94E-05
4.99E-05
5.15E-05
4.80E-05
4.55E-05
3.83E-05
3.49E-05
3.66E-05
1.49E-04
1.32E-04
1.19E-04
1.07E-04
9.66E-05
8.70E-05
7.95E-05
7.20E-05
6.70E-05
6.40E-05
6.02E-05
5.58E-05
5.13E-05
4.80E-05
4.56E-05
4.34E-05
4.11E-05
3.92E-05
3.82E-05
3.74E-05
3.67E-05
3.61E-05
3.52E-05
3.39E-05
3.26E-05
3.12E-05
50
2.23E-04
1.93E-04
1.48E-04
2.11E-04
1.65E-04
1.18E-04
9.75E-05
1.12E-04
1 .26E-04
1.01E-04
8.50E-05
6.61E-05
6.57E-05
7.10E-05
6. 1 4E-05
5.68E-05
4.34E-05
5.97E-05
5.21E-05
5.66E-05
6.90E-05
5.51E-05
5.28E-05
3.97E-05
3.72E-05
4.33E-05
Average Egg Concentration
(mg/Kg)
152 113 90 50
6.81E-03
6.11E-03
5.91E-03
5.83E-03
5.52E-03
5.28E-03
5.01E-03
5.04E-03
4.93E-03
4.71E-03
4.42E-03
4.35E-03
4.24E-03
4.07E-03
4.02E-03
3.88E-03
3.80E-03
3.73E-03
3.58E-03
3.49E-03
3.41E-03
3.38E-03
3.28E-03
3.31E-03
3.27E-03
3.19E-03
5.40E-03
5.06E-03
4.86E-03
4.66E-03
4.47E-03
4.35E-03
4.22E-03
4.04E-03
3.99E-03
3.85E-03
3.67E-03
3.54E-03
3.51E-03
3.40E-03
3.33E-03
3.21E-03
3.11E-03
3.05E-03
3.02E-03
2.96E-03
2.89E-03
2.82E-03
2.72E-03
2.64E-03
2.62E-03
2.61E-03
4.39E-03
4.16E-03
3.90E-03
3.72E-03
3.60E-03
3.43E-03
3.31E-03
3.22E-03
3.I2E-03
3.04E-03
2.98E-03
2.82E-03
2.72E-03
2.60E-03
2.53E-03
2.45E-03
2.40E-03
2.36E-03
2.28E-03
2.23E-03
2.16E-03
2.11E-03
2.06E-03
2.02E-03
2.00E-03
1 .96E-03
3.23E-03
3.03E-03
2.87E-03
2.77E-03
2.68E-03
2.53E-03
2.46E-03
2.38E-03
2.31E-03
2.27E-03
2.19E-03
2.09E-03
2.01E-03
1.92E-03
1.87E-03
1.82E-03
.78E-03
.75E-03
.72E-03
.68E-03
.63E-03
.59E-03
.55E-03
.52E-03
1 .48E-03
1 .45E-03
TAMS/MCA
-------
TABLE 3-38: SUMMARY OF ADD95%IJCL AND EGG CONCENTRATIONS FOR
FEMALE MALLARD ON A TEQ BASIS FOR PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
2.43E-04
2.11E-04
1.65E-04
2.29E-04
1.82E-04
1.34E-04
I.12E-04
1.27E-04
1.41E-04
1.15E-04
9.82E-05
7.95E-05
7.88E-05
8.36E-05
7.39E-05
6.89E-05
5.53E-05
7.14E-05
6.30E-05
6.73E-05
7.95E-05
6.55E-05
6.30E-05
5.04E-05
4.78E-05
5.37E-05
95% UCL Dietary Dose
(mg/Kg/day)
113 90
1.95E-04
1.74E-04
1.36E-04
1.47E-04
1.34E-04
1.07E-04
9.08E-05
9.02E-05
9.42E-05
8.61E-05
8.01E-05
6.45E-05
6.18E-05
6.31E-05
6.08E-05
5.50E-05
4.90E-05
5.30E-05
5.32E-05
5.37E-05
5.55E-05
5.17E-05
4.91E-05
4.12E-05
3.77E-05
3.96E-05
1.57E-04
1.41E-04
1.19E-04
1.14E-04
1.05E-04
9.06E-05
7.92E-05
7.40E-05
7.13E-05
6.82E-05
6.49E-05
5.64E-05
5.22E-05
5.05E-05
4.84E-05
4.54E-05
4.I8E-05
4.14E-05
4.12E-05
4.13E-05
4.09E-05
3.98E-05
3.84E-05
3.52E-05
3.27E-05
3.20E-05
50
1.54E-04
1.35E-04
1.20E-04
1.08E-04
9.85E-05
8.84E-05
8.06E-05
7.27E-05
6.75E-05
6.44E-05
6.04E-05
5.59E-05
5.13E-05
4.79E-05
4.54E-05
4.32E-05
4.08E-05
3.88E-05
3.80E-05
3.73E-05
3.66E-05
3.60E-05
3.51E-05
3.37E-05
3.23E-05
3.08E-05
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
7.31E-03
6.54E-03
6.33E-03
6.25E-03
5.92E-03
5.67E-03
5.38E-03
5.40E-03
5.28E-03
5.06E-03
4.76E-03
4.69E-03
4.56E-03
4.37E-03
4.32E-03
4.18E-03
4.09E-03
4.01E-03
3.85E-03
3.75E-03
3.67E-03
3.63E-03
3.53E-03
3.58E-03
3.54E-03
3.45E-03
5.80E-03
5.43E-03
5.20E-03
5.00E-03
4.79E-03
4.66E-03
4.52E-03
4.34E-03
4.28E-03
4.13E-03
3.94 E-03
3.81E-03
3.77E-03
3.66E-03
3.59E-03
3.45E-03
3.35E-03
3.28E-03
3.25E-03
3.18E-03
3. 11 E-03
3.03E-03
2.93E-03
2.85E-03
2.83E-03
2.83E-03
4.72E-03
4.47E-03
4.19E-03
3.98E-03
3.86E-03
3.68E-03
3.54E-03
3.44E-03
3.34E-03
3.26E-03
3.19E-03
3.03E-03
2.92E-03
2.79E-03
2.72E-03
2.64E-03
2.58E-03
2.53E-03
2.45E-03
2.39E-03
2.32E-03
2.26E-03
2.22E-03
2.I7E-03
2.15E-03
2. 11 E-03
3.46E-03
3.25E-03
3.08E-03
2.96E-03
2.87E-03
2.71 E-03
2.63E-03
2.55E-03
2.47E-03
2.44E-03
2.35E-03
2.25E-03
2.16E-03
2.07E-03
2.01 E-03
1.96E-03
1 .92E-03
1 .88E-03
1.85E-03
1 .8 1 E-03
1.75E-03
1 .7 1 E-03
1.67E-03
1 .64E-03
1.59E-03
1 .56E-03
TAMS/MCA
-------
TABLE 3-39: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE BELTED KINGFISHER FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
Average Dietary Dose
(mg/Kg/day)
152 113 90
1.20E-04
9.32E-05
8.43E-05
9.76E-05
8.44E-05
6.64E-05
5.98E-05
5.87E-05
6.32E-05
5.70E-05
5.28E-05
4.41E-05
4.30E-05
4.83E-05
4.17E-05
3.91E-05
3.65E-05
3.87E-05
4.05E-05
3.83E-05
4.00E-05
3.79E-05
3.43E-05
3.06E-05
2.95E-05
2.95E-05
8.37E-05
7.64E-05
6.47E-05
6.61E-05
6.19E-05
5.32E-05
4.83E-05
4.50E-05
4.62E-05
4.48E-05
4.14E-05
3.68E-05
3.49E-05
3.51E-05
3.40E-05
3.22E-05
3.00E-05
2.95E-05
3.05E-05
2.99E-05
3.00E-05
2.89E-05
2.73E-05
2.53E-05
2.40E-05
2.35E-05
6.71E-05
6.14E-05
6.02E-05
5.69E-05
4.82E-05
4.48E-05
3.99E-05
3.72E-05
3.55E-05
3.46E-05
3.30E-05
3.06E-05
2.85E-05
2.71E-05
2.61E-05
2.53E-05
2.40E-05
2.29E-05
2.27E-05
2.27E-05
2.24E-05
2.I7E-05
2.10E-05
2.02E-05
1.92E-05
1 .86E-05
50
6.03E-05
5.44E-05
4.90E-05
4.53E-05
4.20E-05
3.92E-05
3.56E-05
3.29E-05
3.08E-05
2.96E-05
2.82E-05
2.64E-05
2.47E-05
2.32E-05
2.22E-05
2.14E-05
2.04E-05
1 .96E-05
1.91E-05
1 .89E-05
1.86E-05
1.80E-05
1.75E-05
1.70E-05
1 .63E-05
1 .59E-05
Average Egg Concentration
(rag/Kg)
152 113 90 50
5.63E-03
4.34E-03
3.93E-03
4.57E-03
3.94E-03
3.08E-03
2.77E-03
2.72E-03
2.94E-03
2.64E-03
2.44E-03
2.03E-03
1.98E-03
2.23E-03
1.92E-03
1.80E-03
1.67E-03
1 .78E-03
1.87E-03
1.76E-03
1.85E-03
1.75E-03
1 .58E-03
1 .40E-03
1 .35E-03
1 .36E-03
3.91E-03
3.56E-03
3.01E-03
3.08E-03
2.88E-03
2.47E-03
2.23E-03
2.08E-03
2.14E-03
2.07E-03
1.91E-03
1.69E-03
1.61E-03
1.62E-03
1.56E-03
1 .48E-03
1 .38E-03
1.35E-03
1 .40E-03
1.37E-03
1.38E-03
1.33E-03
1 .26E-03
1.16E-03
1.IOE-03
1.08E-03
3.13E-03
2.86E-03
2.55E-03
2.39E-03
2.24E-03
2.08E-03
1.85E-03
1 .72E-03
1.64E-03
1.60E-03
1.52E-03
1.41E-03
1.31E-03
1.25E-03
1.20E-03
1.16E-03
1.10E-03
1 .05E-03
1 .04E-03
1.04E-03
M.03E-03
9.97E-04
9.67E-04
9.27E-04
8.83E-04
8.55E-04
2.83E-03
2.55E-03
2.29E-03
2.12E-03
1 .96E-03
1 .83E-03
1.66E-03
1.53E-03
1 .43E-03
1 .38E-03
1.31E-03
1 .22E-03
1.14E-03
1 .07E-03
1 .03E-03
9.88E-04
9.45E-04
9.04E-04
8.83E-04
8.75E-04
8.60E-04
8.31E-04
8.11E-04
7.85E-04
7.54E-04
7.32E-04
TAMS/MCA
-------
TABLE 3-40: SUMMARY OF ADD95%IJCL AND EGG CONCENTRATIONS FOR
FEMALE BELTED KINGFISHER FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
95% UCL Dietary Dose
(mg/Kg/day)
152 113 90
1.25E-04
9.68E-05
8.80E-05
1.02E-04
8.80E-05
6.97E-05
6.31E-05
6.18E-05
6.62E-05
6.01E-05
5.56E-05
4.70E-05
4.59E-05
5.10E-05
4.43E-05
4.18E-05
3.92E-05
4.12E-05
4.29E-05
4.06E-05
4.23E-05
4.02E-05
3.64E-05
3.30E-05
3.19E-05
3. 1 8E-05
8.70E-05
7.95E-05
6.77E-05
6.90E-05
6.48E-05
5.59E-05
5.09E-05
4.75E-05
4.86E-05
4.73E-05
4.37E-05
3.91E-05
3.71E-05
3.74E-05
3.61E-05
3.43E-05
3.21E-05
3.15E-05
3.25E-05
3.18E-05
3.19E-05
3.07E-05
2.91E-05
2.71E-05
2.58E-05
2.54E-05
1.69E-04
1.59E-04
1.50E-04
1.44E-04
1.39E-04
1 .34E-04
1.29E-04
1.23E-04
1.18E-04
1.17E-04
1.13E-04
1.13E-04
1.11E-04
1 .04E-04
1.01E-04
1.02E-04
1 .03E-04
9.56E-05
9.03E-05
8.88E-05
8.67E-05
8.49E-05
8.34E-05
8.68E-05
8.65E-05
8.27E-05
50
1.42E-04
1.34E-04
1.27E-04
1.21E-04
1.16E-04
1.12E-04
1.08E-04
1.03E-04
9.95E-05
9.66E-05
9.39E-05
9.25E-05
9.01E-05
8.69E-05
8.43E-05
8.24E-05
8.24E-05
7.99E-05
7.73E-05
7.43E-05
7.25E-05
7.07E-05
7.04E-05
6.96E-05
6.87E-05
6.87E-05
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
5.82E-03
4.50E-03
4.08E-03
4.73E-03
4.08E-03
3.21E-03
2.89E-03
2.83E-03
3.05E-03
2.75E-03
2.54E-03
2.12E-03
2.07E-03
2.32E-03
2.00E-03
1.88E-03
1.75E-03
1.86E-03
1 .95E-03
1.84E-03
1.93E-03
1.83E-03
1 .65E-03
1.47E-03
1 .42E-03
1.42E-03
4.05E-03
3.69E-03
3.13E-03
3.19E-03
2.99E-03
2.57E-03
2.33E-03
2.17E-03
2.22E-03
2.16E-03
1.99E-03
1 .77E-03
1.68E-03
1.69E-03
1.63E-03
1.55E-03
1 .44E-03
1.42E-03
1 .46E-03
1.44E-03
1 .45E-03
1.39E-03
1.31E-03
1.21E-03
I.15E-03
1.13E-03
3.25E-03
2.97E-03
2.65E-03
2.48E-03
2.33E-03
2.16E-03
1.92E-03
1.79E-03
1.71E-03
1.67E-03
1.59E-03
1 .47E-03
I.37E-03
1.30E-03
1.25E-03
1.21E-03
1.15E-03
1.10E-03
1 .09E-03
1.09E-03
1.08E-03
1.04E-03
1.01E-03
9.69E-04
9.24E-04
8.94E-04
2.93E-03
2.64E-03
2.38E-03
2.20E-03
2.03E-03
1.90E-03
1.72E-03
1 .59E-03
1 .49E-03
1.43E-03
1.36E-03
1 .27E-03
1.19E-03
1.12E-03
1.07E-03
1 .03E-03
9.85E-04
9.43E-04
9.21E-04
9.12E-04
8.97E-04
8.66E-04
8.45E-04
8.19E-04
7.86E-04
7.64E-04
TAMS/MCA
-------
TABLE 3-41: SUMMARY OF ADDExpecled AND EGG CONCENTRATIONS FOR
FEMALE GREAT BLUE HERON FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
Average Dietary Dose
(mg/Kg/day)
152 113 90
3.09E-05
2.32E-05
2.06E-05
2.48E-05
2.10E-05
1.56E-05
1.38E-05
1.34E-05
1.49E-05
1.32E-05
1.21E-05
9.50E-06
9.29E-06
1.11E-05
9.08E-06
8.41E-06
7.66E-06
8.41E-06
9.13E-06
8.52E-06
9.14E-06
8.53E-06
7.48E-06
6.30E-06
6.01E-06
6.09E-06
2.11E-05
1.92E-05
1.57E-05
1.63E-05
1.52E-05
1.26E-05
1.12E-05
1.03E-05
1.07E-05
1.04E-05
9.51E-06
8.21E-06
7.65E-06
7.82E-06
7.52E-06
7.08E-06
6.48E-06
6.39E-06
6.71E-06
6.59E-06
6.69E-06
6.40E-06
6.01E-06
5.45E-06
5.08E-06
4.94E-06
1.69E-05
1.53E-05
1.47E-05
1.38E-05
1.17E-05
1.08E-05
9.42E-06
8.66E-06
8.22E-06
8.00E-06
7.55E-06
6.94E-06
6.40E-06
6.06E-06
5.83E-06
5.63E-06
5.28E-06
4.97E-06
5.00E-06
5.03E-06
5.02E-06
4.83E-06
4.67E-06
4.44E-06
4.17E-06
4.02E-06
50
1.59E-05
1.42E-05
1.27E-05
1.16E-05
1.07E-05
9.91E-06
8.85E-06
8.10E-06
7.51E-06
7.16E-06
6.79E-06
6.32E-06
5.86E-06
5.48E-06
5.22E-06
5.00E-06
4.75E-06
4.51E-06
4.39E-06
4.36E-06
4.31E-06
4.15E-06
4.05E-06
3.91E-06
3.74E-06
3.62E-06
152
3.68E-03
2.74E-03
2.43E-03
2.95E-03
2.49E-03
1.83E-03
1.62E-03
1.57E-03
1.76E-03
1.55E-03
1 .42E-03
1.10E-03
1.08E-03
1 .30E-03
1.06E-03
9.76E-04
8.87E-04
9.80E-04
1 .07E-03
9.95E-04
1 .07E-03
9.99E-04
8.71E-04
7.29E-04
6.94E-04
7.05E-04
Average Egg Concentration
(mg/Kg)
113 90
2.48E-03
2.24E-03
1.82E-03
1.90E-03
1 .77E-03
1 .45E-03
1 .28E-03
1.18E-03
1.23E-03
1.20E-03
1 .09E-03
9.31E-04
8.66E-04
8.89E-04
8.54E-04
8.01E-04
7.30E-04
7.19E-04
7.61E-04
7.47E-04
7.62E-04
7.27E-04
6.80E-04
6.13E-04
5.68E-04
5.52E-04
1.98E-03
1.79E-03
1.57E-03
1.46E-03
1 .36E-03
1.25E-03
1 .08E-03
9.94E-04
9.41E-04
9.17E-04
8.63E-04
7.90E-04
7.26E-04
6.87E-04
6.60E-04
6.37E-04
5.96E-04
5.59E-04
5.64E-04
5.68E-04
5.68E-04
5.46E-04
5.27E-04
5.00E-04
4.68E-04
4.51E-04
50
1.88E-03
1.68E-03
1.49E-03
1.36E-03
1.25E-03
1.16E-03
1.03E-03
9.40E-04
8.70E-04
8.28E-04
7.85E-04
7.29E-04
6.75E-04
6.30E-04
5.99E-04
5.73E-04
5.43E-04
5.15E-04
5.01E-04
4.99E-04
4.93E-04
4.74E-04
4.63E-04
4.46E-04
4.26E-04
4.13E-04
TAMS/MCA
-------
TABLE 3-42: SUMMARY OF ADD95%l,cL AND EGG CONCENTRATIONS FOR
FEMALE GREAT BLUE HERON FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
95%
152
3.20E-05
2.41E-05
2.15E-05
2.58E-05
2.19E-05
1.64E-05
1.46E-05
1.42E-05
1.57E-05
1.40E-05
1.29E-05
1.02E-05
9.97E-06
1.18E-05
9.71E-06
9.06E-06
8.32E-06
9.04E-06
9.72E-06
9.10E-06
9.72E-06
9.10E-06
8.01E-06
6.87E-06
6.58E-06
6.63E-06
UCL Dietary
(mg/Kg/day)
113
2.17E-05
1.97E-05
1.62E-05
1.68E-05
1.57E-05
1.30E-05
1.16E-05
1.07E-05
1.11E-05
1.08E-05
9.90E-06
8.57E-06
8.01E-06
8.18E-06
7.86E-06
7.42E-06
6.82E-06
6.71E-06
7.04E-06
6.89E-06
7.01E-06
6.71E-06
6.30E-06
5.73E-06
5.35E-06
5.22E-06
Dose
90
1.74E-05
1.58E-05
1.39E-05
1.30E-05
1.21E-05
1.12E-05
9.73E-06
8.96E-06
8.50E-06
8.29E-06
7.83E-06
7.21E-06
6.66E-06
6.31E-06
6.07E-06
5.88E-06
5.52E-06
5.19E-06
5.22E-06
5.24E-06
5.23E-06
5.03E-06
4.87E-06
4.63E-06
4.36E-06
4.20E-06
50
1.63E-05
1.46E-05
1.30E-05
1.19E-05
1.10E-05
1.02E-05
9.12E-06
8.35E-06
7.75E-06
7.39E-06
7.02E-06
6.54E-06
6.07E-06
5.68E-06
5.42E-06
5.19E-06
4.93E-06
4.69E-06
4.57E-06
4.54E-06
4.48E-06
4.32E-06
4.21E-06
4.07E-06
3.89E-06
3.77E-06
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
3.78E-03
2.81E-03
2.50E-03
3.03E-03
2.55E-03
1.89E-03
1.67E-03
1.62E-03
1.81E-03
1.60E-03
1.47E-03
1.14E-03
1.11E-03
1 .34E-03
1.09E-03
1.01E-03
9.14E-04
1.01E-03
1.10E-03
1 .02E-03
1.10E-03
1 .03E-03
8.97E-04
7.52E-04
7.16E-04
7.27E-04
2.54E-03
2.30E-03
1.88E-03
1.95E--03
1.82E-03
1.50E-03
1.32E-03
1.22E-03
1.27E-03
1.24E-03
1.12E-03
9.60E-04
8.92E-04
9.16E-04
8.79E-04
8.26E-04
7.53E-04
7.41E-04
7.84E-04
7.69E-04
7.85E-04
7.50E-04
7.00E-04
6.32E-04
5.86E-04
5.69E-04
2.04E-03
1.84E-03
1.61E-03
1.51E-03
1.40E-03
1.29E-03
1.12E-03
1.02E-03
9.68E-04
9.44E-04
8.89E-04
8.15E-04
7.48E-04
7.07E-04
6.79E-04
6.57E-04
6.14E-04
5.75E-04
5.80E-04
5.85E-04
5.85E-04
5.63E-04
5.43E-04
5.16E-04
4.83E-04
4.64E-04
1.93E-03
1.72E-03
1.53E-03
1.40E-03
1.28E-03
1.19E-03
1.06E-03
9.68E-04
8.95E-04
8.52E-04
8.07E-04
7.50E-04
6.95E-04
6.48E-04
6.16E-04
5.90E-04
5.59E-04
5.30E-04
5.16E-04
5.I3E-04
5.08E-04
4.88E-04
4.77E-04
4.60E-04
4.39E-04
4.25E-04
TAMS/MCA
-------
TABLE 3-43: SUMMARY OF ADDExpected AND EGG CONCENTRATIONS FOR
FEMALE EAGLE FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
3.60E-04
2.61E-04
2.24E-04
2.66E-04
2.42E-04
1.90E-04
1.64E-04
1 .54E-04
1.75E-04
1.60E-04
1.43E-04
1.15E-04
1.10E-04
1.23E-04
1.14E-04
1.06E-04
9.19E-05
9.65E-05
1 .08E-04
9.66E-05
1.05E-04
9.59E-05
8.86E-05
8.10E-05
7.29E-05
7.11E-05
Average Dietary Dose
(mg/Kg/day)
113 90
2.41E-04
2.10E-04
1.87E-04
1.75E-04
1.69E-04
1.53E-04
1.28E-04
1.17E-04
1.18E-04
1.17E-04
1.08E-04
9.54E-05
8.78E-05
8.67E-05
8.49E-05
8.16E-05
7.46E-05
7.13E-05
7.58E-05
7.40E-05
7.64E-05
7.24E-05
6.9IE-05
6.43E-05
5.93E-05
5.63E-05
5.52E-05
5.03E-05
4.53E-05
4.13E-05
3.86E-05
3.55E-05
3.20E-05
2.90E-05
2.76E-05
2.68E-05
2.55E-05
2.36E-05
2.18E-05
2.07E-05
1 .99E-05
1.92E-05
1 .82E-05
1.72E-05
1.71E-05
1.69E-05
1.74E-05
1 .64E-05
1 .60E-05
1.53E-05
1.45E-05
1.37E-05
50
5.21E-05
4.66E-05
4.19E-05
3.79E-05
3.50E-05
3.22E-05
2.93E-05
2.69E-05
2.51E-05
2.39E-05
2.27E-05
2.12E-05
1.97E-05
1.85E-05
1.76E-05
1.69E-05
1.61E-05
1.53E-05
1.48E-05
1.45E-05
1 .48E-05
1 .40E-05
I.37E-05
1.33E-05
I.28E-05
1.21E-05
152
5.37E-02
3.89E-02
3.34E-02
3.96E-02
3.60E-02
2.83E-02
2.45E-02
2.30E-02
2.61E-02
2.39E-02
2.13E-02
1.72E-02
1.63E-02
1.84E-02
1 .70E-02
1.58E-02
1 .37E-02
1.44E-02
1.61E-02
1 .44E-02
1 .57E-02
1.43E-02
1 .32E-02
1.21E-02
1 .09E-02
1 .06E-02
Average Egg Concentration
(mg/Kg)
113 90
3.59E-02
3.14E-02
2.79E-02
2.60E-02
2.51E-02
2.29E-02
1.92E-02
1.74E-02
1.76E-02
1.75E-02
1.60E-02
1.42E-02
1.31E-02
1.29E-02
1.27E-02
1 .22E-02
1.I1E-02
1.06E-02
1 . 1 3E-02
1.10E-02
1.14E-02
1.08E-02
1 .03E-02
9.59E-03
8.84E-03
8.39E-03
8.23E-03
7.49E-03
6.76E-03
6.15E-03
5.75E-03
5.30E-03
4.77E-03
4.33E-03
4.11E-03
4.00E-03
3.80E-03
3.53E-03
3.25E-03
3.08E-03
2.97E-03
2.87E-03
2.71E-03
2.57E-03
2.55E-03
2.53E-03
2.59E-03
2.45E-03
2.38E-03
2.28E-03
2.16E-03
2.04E-03
50
7.76E-03
6.95E-03
6.25E-03
5.65E-03
5.22E-03
4.79E-03
4.37E-03
4.00E-03
3.74E-03
3.57E-03
3.38E-03
3.16E-03
2.94E-03
2.76E-03
2.63E-03
2.52E-03
2.39E-03
2.28E-03
2.21E-03
2.16E-03
2.21E-03
2.09E-03
2.04E-03
1 .98E-03
1.91E-03
1.81E-03
TAMS/MCA
-------
TABLE 3-44: SUMMARY OF ADD95%UCL AND EGG CONCENTRATIONS FOR
FEMALE EAGLE FOR THE PERIOD 1993 - 2018 ON TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
3.68E-04
2.67E-04
2.29E-04
2.72E-04
2.47E-04
1.94E-04
1 .68E-04
1.57E-04
1.79E-04
1.63E-04
1.46E-04
1.18E-04
1.12E-04
1.26E-04
1.16E-04
1.08E-04
9.40E-05
9.86E-05
1.11E-04
9.88E-05
1.08E-04
9.81E-05
9.06E-05
8.28E-05
7.45E-05
7.28E-05
95% UCL Dietary Dose
(mg/Kg/day)
113 90
2.46E-04
2.15E-04
1.91E-04
1.78E-04
1.72E-04
1.57E-04
1.31E-04
1.20E-04
1.21E-04
1.20E-04
1.10E-04
9.76E-05
8.98E-05
8.87E-05
8.68E-05
8.34E-05
7.63E-05
7.29E-05
7.75E-05
7.57E-05
7.82E-05
7.41E-05
7.06E-05
6.58E-05
6.06E-05
5.76E-05
5.61E-05
5.11E-05
4.61E-05
4.20E-05
3.92E-05
3.61E-05
3.26E-05
2.96E-05
2.81E-05
2.73E-05
2.59E-05
2.41E-05
2.22E-05
2.10E-05
2.03E-05
1.96E-05
1 .85E-05
1 .76E-05
1 .74E-05
1.72E-05
1 .77E-05
1 .67E-05
1.62E-05
1.55E-05
1 .47E-05
1 .40E-05
50
5.29E-05
4.74E-05
4.26E-05
3.85E-05
3.56E-05
3.27E-05
2.98E-05
2.73E-05
2.55E-05
2.43E-05
2.31E-05
2.16E-05
2.01E-05
1 .89E-05
1.80E-05
1.72E-05
1.63E-05
1.56E-05
1.51E-05
1.48E-05
1.51E-05
1 .43E-05
1.39E-05
1 .35E-05
1.31E-05
1 .23E-05
95% UCL Egg Concentration
(mg/Kg)
152 113 90 50
5.48E-02
3.97E-02
3.41E-02
4.05E-02
3.68E-02
2.89E-02
2.50E-02
2.35E-02
2.67E-02
2.44E-02
2.17E-02
1.75E-02
1.67E-02
1.88E-02
1.74E-02
1.61E-02
1.40E-02
I.47E-02
1.65E-02
1.47E-02
1.61E-02
1 .46E-02
1 .35E-02
1.23E-02
1.I1E-02
1.08E-02
3.67E-02
3.21E-02
2.85E-02
2.66E-02
2.57E-02
2.34E-02
1 .96E-02
1 .78E-02
1.80E-02
1.78E--02
1 .64E-02
1.45E-02
1.34E-02
1 .32E-02
1.29E-02
1.24E-02
1.14E-02
1.09E-02
1.16E-02
1.13E-02
I.17E-02
1.10E-02
1 .05E-02
9.8IE-03
9.04E-03
8.59E-03
8.37E-03
7.62E-03
6.87E-03
6.26E-03
5.85E-03
5.39E-03
4.85E-03
4.41E-03
4.18E-03
4.07E-03
3.86E-03
3.59E-03
3.31E-03
3.14E-03
3.03E-03
2.92E-03
2.76E-03
2.62E-03
2.60E-03
2.57E-03
2.64E-03
2.50E-03
2.42E-03
2.32E-03
2.20E-03
2.08E-03
7.88E-03
7.06E-03
6.35E-03
5.74E-03
5.31E-03
4.87E-03
4.44E-03
4.07E-03
3.80E-03
3.63E-03
3.44E-03
3.22E-03
2.99E-03
2.81E-03
2.68E-03
2.56E-03
2.44E-03
2.32E-03
2.25E-03
2.20E-03
2.25E-03
2.13E-03
2.07E-03
2.01E-03
1.95E-03
1.84E-03
TAMS/MCA
-------
TABLE 3-45: SUMMARY OF ADDExpected FOR FEMALE BAT
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
6.18E-01
5.54E-01
5.36E-01
5.29E-01
5.01E-01
4.79E-01
4.55E-01
4.57E-01
4.47E-01
4.27E-01
4.01E-01
3.95E-01
3.84E-01
3.69E-01
3.64E-01
3.52E-01
3.44E-01
3.39E-01
3.25E-01
3.16E-01
3.10E-01
3.06E-01
2.97E-01
3.00E-01
2.97E-01
2.89E-01
Total Average Dietary Dose
(mg/Kg/day)
113 90
4.90E-01
4.59E-01
4.41E-01
4.23E-01
4.06E-01
3.95E-01
3.83E-01
3.67E-01
3.62E-01
3.49E-01
3.33E-01
3.21E-01
3.18E-01
3.09E-01
3.03E-01
2.91E-01
2.82E-01
2.77E-01
2.74E-01
2.68E-01
2.62E-01
2.56E-01
2.47E-01
2.40E-01
2.38E-01
2.37E-01
3.98E-01
3.78E-01
3.54E-01
3.37E-01
3.27E-01
3.11E-01
3.00E-01
2.92E-01
2.83E-01
2.76E-01
2.70E-01
2.56E-01
2.46E-01
2.36E-01
2.29E-01
2.23E-01
2.18E-01
2.14E-01
2.07E-01
2.02E-01
1.96E-01
1.91E-01
1.87E-01
1.83E-01
1.81E-01
1.78E-01
50
2.93E-OI
2.75E-01
2.61E-01
2.51E-01
2.43E-01
2.30E-01
2.23E-01
2.16E-01
2.10E-OI
2.06E-01
.99E-01
.90E-01
.83E-01
.75E-01
.70E-01
.65E-01
.62E-01
.59E-01
.56E-01
.53E-01
1.48E-01
1.44E-01
1.41E-01
1.38E-01
1.34E-01
1.31E-01
TAMS/MCA
-------
TABLE 3-46: SUMMARY OF ADD95%IJCL FOR FEMALE BAT
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
6.64E-01
5.94E-01
5.74E-01
5.67E-01
5.37E-01
5.14E-01
4.88E-01
4.90E-01
4.79E-01
4.59E-01
4.32E-01
4.25E-01
4.I4E-0!
3.97E-01
3.92E-01
3.79E-01
3.71E-OI
3.64E-01
3.49E-01
3.40E-01
3.33E-01
3.29E-01
3.21E-01
3.25E-01
3.21E-01
3.13E-01
Total 95% UCL Dietary Dose
(mg/Kg/day)
113 90
5.26E-01
4.92E-01
4.72E-01
4.54E-01
4.35E-01
4.23E-01
4.10E-01
3.94E-01
3.89E-01
3.75E-01
3.58E-01
3.45E-01
3.42E-01
3.32E-01
3.25E-01
3.14E-01
3.04E-01
2.98E-01
2.95E-01
2.89E-01
2.82E-01
2.75E-01
2.65E-01
2.59E-01
2.57E-01
2.56E-01
4.28E-01
4.05E-01
3.80E-01
3.61E-01
3.50E-01
3.34E-01
3.21E-OI
3.12E-01
3.03E-01
2.96E-01
2.90E-01
2.75E-01
2.65E-01
2.53E-01
2.47E-01
2.39E-01
2.34E-01
2.30E-01
2.22E-01
2.17E-01
2.10E-01
2.05E-01
2.0IE-01
I.97E-OI
1.95E-01
1.91E-01
50
3.14E-01
2.95E-01
2.80E-01
2.69E-01
2.60E-01
2.46E-01
2.39E-01
2.32E-OI
2.24E-01
2.21E-01
2.13E-01
2.04E-01
.96E-01
.88E-01
.82E-OI
.77E-01
.74E-01
.70E-01
.68E-01
.64E-01
.59E-01
.55E-01
.51E-01
.48E-01
.45E-OI
.41E-01
TAMS/MCA
-------
TABLE 3-47: SUMMARY OF ADDExpe<.led FOR FEMALE RACCOON
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.13E-01
9.99E-02
9.60E-02
9.59E-02
9.03E-02
8.52E-02
8.04E-02
8.07E-02
7.95E-02
7.57E-02
7.11E-02
6.93E-02
6.74E-02
6.53E-02
6.40E-02
6.17E-02
6.02E-02
5.94E-02
5.73E-02
5.57E-02
5.48E-02
5.40E-02
5.22E-02
5.23E-02
5.16E-02
5.04E-02
Average Dietary Dose
(mg/Kg/day)
113 90
8.84E-02
8.27E-02
7.86E-02
7.58E-02
7.26E-02
7.00E-02 .
6.75E-02
6.46E-02
6.39E-02
6. 1 7E-02
5.88E-02
5.64E-02
5.57E-02
5.42E-02
5.30E-02
5.10E-02
4.94E-02
4.84E-02
4.81E-02
4.70E-02
4.60E-02
4.49E-02
4.32E-02
4.19E-02
4.14E-02
4.12E-02
7.17E-02
6.78E-02
7.10E-02
6.80E-02
5.83E-02
5.54E-02
5.32E-02
5.15E-02
5.00E-02
4.87E-02
4.76E-02
4.51E-02
4.33E-02
4.14E-02
4.03E-02
3.91E-02
3.82E-02
3.74E-02
3.62E-02
3.54E-02
3.44E-02
3.36E-02
3.29E-02
3.21E-02
3.17E-02
3.11E-02
50
5.36E-02
5.01E-02
4.74E-02
4.54E-02
4.38E-02
4.14E-02
4.00E-02
3.87E-02
3.73E-02
3.66E-02
3.53E-02
3.37E-02
3.24E-02
3.09E-02
3.00E-02
2.92E-02
2.86E-02
2.80E-02
2.75E-02
2.69E-02
2.61E-02
2.54E-02
2.48E-02
2.43E-02
2.37E-02
2.31E-02
TAMS/MCA
-------
TABLE 3-48: SUMMARY OF ADD95%UCL FOR FEMALE RACCOON
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.21E-01
1.07E-01
1.03E-01
1.03E-01
9.70E-02
9.19E-02
8.72E-02
8.73E-02
8.56E-02
8.19E-02
7.71E-02
7.55E-02
7.36E-02
7.11E-02
6.96E-02
6.75E-02
6.62E-02
6.49E-02
6.23E-02
6.07E-02
5.95E-02
5.88E-02
5.71E-02
5.77E-02
5.70E-02
5.56E-02
95% UCL Dietary Dose
(mg/Kg/day)
113 90
9.49E-02
8.87E-02
8.45E-02
8.15E-02
7.81E-02
7.54E-02
7.30E-02
7.00E-02
6.92E-02
6.68E-02
6.37E-02
6.14E-02
6.07E-02
5.90E-02
5.78E-02
5.57E-02
5.41E-02
5.30E-02
5.25E-02
5.12E-02
5.02E-02
4.89E-02
4.72E-02
4.59E-02
4.55E-02
4.55E-02
7.70E-02
7.28E-02
6.81E-02
6.48E-02
6.26E-02
5.97E-02
5.72E-02
5.55E-02
5.39E-02
5.26E-02
5.14E-02
4.88E-02
4.70E-02
4.50E-02
4.38E-02
4.26E-02
4.16E-02
4.08E-02
3.95E-02
3.86E-02
3.74E-02
3.65E-02
3.58E-02
3.50E-02
3.45E-02
3.39E-02
50
5.75E-02
5.38E-02
5.09E-02
4.88E-02
4.7IE-02
4.45E-02
4.30E-02
4.16E-02
4.03E-02
3.96E-02
3.82E-02
3.65E-02
3.51E-02
3.36E-02
3.27E-02
3.18E-02
3.11E-02
3.04E-02
2.99E-02
2.93E-02
2.84E-02
2.77E-02
2.70E-02
2.65E-02
2.58E-02
2.52E-02
TAMS/MCA
-------
TABLE 3-49: SUMMARY OF ADDExpected FOR FEMALE MINK
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.37E-01
1.09E-01
9.99E-02
1.12E-01
9.86E-02
8.09E-02
7.38E-02
7.29E-02
7.67E-02
7.01E-02
6.51E-02
5.68E-02
5.54E-02
5.96E-02
5.33E-02
5.04E-02
4.77E-02
4.95E-02
5.06E-02
4.82E-02
4.96E-02
4.75E-02
4.37E-02
4.05E-02
3.93E-02
3.91E-02
Average Dietary Dose
(mg/Kg/day)
113 90
9.75E-02
8.94E-02
7.78E-02
7.84E-02
7.38E-02
6.53E-02
6.02E-02
5.65E-02
5.74E-02
5.56E-02
5.17E-02
4.70E-02
4.52E-02
4.50E-02
4.37E-02
4.15E-02
3.91E-02
3.85E-02
3.93E-02
3.85E-02
3.84E-02
3.71E-02
3.52E-02
3.31E-02
3.18E-02
3.14E-02
7.84E-02
7.22E-02
6.62E-02
6.24E-02
5.79E-02
5.41E-02
4.91E-02
4.63E-02
4.44E-02
4.32E-02
4.15E-02
3.86E-02
3.63E-02
3.46E-02
3.34E-02
3.24E-02
3.10E-02
2.98E-02
2.94E-02
2.91E-02
2.87E-02
2.78E-02
2.70E-02
2.61E-02
2.51E-02
2.44E-02
50
6.79E-02
6.16E-02
5.60E-02
5.22E-02
4.88E-02
4.56E-02
4.20E-02
3.93E-02
3.71E-02
3.58E-02
3.42E-02
3.21E-02
3.03E-02
2.86E-02
2.74E-02
2.65E-02
2.55E-02
2.45E-02
2.40E-02
2.37E-02
2.32E-02
2.25E-02
2.19E-02
2.13E-02
2.06E-02
2.00E-02
TAMS/MCA
-------
TABLE 3-50: SUMMARY OF ADD9S%IJCL FOR FEMALE MINK
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.42E-01
1.14E-01
1.05E-01
1..17E-01
1.03E-01
8.50E-02
7.77E-02
7.66E-02
8.04E-02
7.38E-02
6.85E-02
6.00E-02
5.85E-02
6.27E-02
5.63E-02
5.34E-02
5.06E-02
5.23E-02
5.33E-02
5.08E-02
5.22E-02
5.01E-02
4.62E-02
4.31E-02
4.19E-02
4.16E-02
95% UCL Dietary Dose
(mg/Kg/day)
113 90
1.02E-01
9.34E-02
8.16E-02
8.20E-02
7.73E-02
6.85E-02
6.34E-02
5.95E-02
6.04E-02
5.86E-02
5.45E-02
4.97E-02
4.77E-02
4.76E-02
4.61E-02
4.40E-02
4.I5E-02
4.07E-02
4.I6E-02
4.07E-02
4.06E-02
3.92E-02
3.73E-02
3.51E-02
3.38E-02
3.34E-02
8.21E-02
7.56E-02
6.82E-02
6.41E-02
6.07E-02
5.67E-02
5.16E-02
4.86E-02
4.66E-02
4.55E-02
4.37E-02
4.07E-02
3.83E-02
3.65E-02
3.53E-02
3.42E-02
3.28E-02
3.15E-02
3.11E-02
3.08E-02
3.03E-02
2.93E-02
2.86E-02
2.76E-02
2.66E-02
2.59E-02
50
7.08E-02
6.44E-02
5.85E-02
5.45E-02
5.10E-02
4.77E-02
4.40E-02
4.12E-02
3.88E-02
3.75E-02
3.59E-02
3.38E-02
3.18E-02
3.00E-02
2.89E-02
2.79E-02
2.68E-02
2.59E-02
2.53E-02
2.50E-02
2.45E-02
2.37E-02
2.3 1 E-02
2.25E-02
2.17E-02
2. 11 E-02
TAMS/MCA
-------
TABLE 3-51: SUMMARY OF ADDExpected FOR FEMALE OTTER
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
.83E+00
.33E+00
.14E+00
.35E+00
.23E+00
9.66E-01
8.35E-01
7.83E-01
8.90E-01
8.14E-01
7.25E-01
5.85E-01
5.57E-01
6.28E-01
5.79E-OI
5.38E-01
4.68E-01
4.91E-01
5.50E-01
4.92E-01
5.36E-01
4.88E-01
4.51E-01
4.12E-01
3.71E-01
3.62E-01
Average Dietary Dose
(mg/Kg/day)
113 90
1.23E+00
1.07E+00
9.52E-01
8.88E-01
8.57E-01
7.81E-01
6.54E-01
5.95E-01
6.02E-01
5.96E-01
5.48E-01
4.86E-01
4.47E-01
4.41E-01
4.32E-01
4.15E-01
3.79E-01
3.63E-01
3.86E-01
3.77E-01
3.89E-01
3.69E-01
3.51E-01
3.27E-01
3.02E-01
2.86E-01
2.81E-01
2.56E-01
2.31E-01
2.11E-01
1.96E-01
.81E-01
.63E-01
.48E-01
.40E-01
.37E-01
1.30E-01
1.20E-OI
1.11E-01
1.05E-01
1.02E-01
9.80E-02
9.26E-02
8.78E-02
8.71E-02
8.63E-02
8.86E-02
8.38E-02
8.13E-02
7.78E-02
7.37E-02
6.98E-02
50
2.65E-01
2.37E-01
2.13E-01
.93E-01
.78E-01
.64E-01
.49E-01
.37E-01
.28E-01
.22E-01
.15E-01
1.08E-01
l.OOE-01
9.44E-02
8.98E-02
8.60E-02
8.18E-02
7.79E-02
7.56E-02
7.39E-02
7.55E-02
7.14E-02
6.95E-02
6.75E-02
6.54E-02
6.17E-02
TAMS/MCA
-------
TABLE 3-52: SUMMARY OF ADD9S%UCL FOR FEMALE OTTER
BASED ON TRI+ PREDICTIONS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.87E+00
1.36E+00
1.16E+00
1.38E+00
1.26E+00
9.87E-01
8.53E-01
8.0IE-01
9.IOE-01
8.32E-01
7.42E-01
5.98E-01
5.70E-01
6.42E-01
5.92E-01
5.50E-OI
4.78E-01
5.02E-01
5.62E-01
5.03E-01
5.48E-01
4.99E-01
4.6IE-01
4.21E-01
3.79E-01
3.70E-01
95% UCL Dietary Dose
(mg/Kg/day)
113 90
1.25E+00
1.09E+00
9.72E-01
9.07E-01
8.76E-01
7.98E-01
6.68E-OI
6.08E-01
6.16E-01
6.09E-01
5.60E-01
4.97E-0!
4.57E-01
4.51E-01
4.41E-01
4.25E-01
3.88E-01
3.71E-OI
3.94E-01
3.85E-01
3.98E-01
3.77E-OI
3.59E-01
3.35E-01
3.09E-01
2.93E-01
2.86E-01
2.60E-01
2.35E-01
2.14E-01
2.00E-01
.84E-01
.66E-01
.5IE-01
.43E-01
.39E-OI
.32E-OI
1.23E-01
.I3E-01
1.07E-01
1.03E-01
9.98E-02
9.43 E-02
8.95E-02
8.88E-02
8.79E-02
9.02E-02
8.53E-02
8.28E-02
7.92E-02
7.51 E-02
7.12E-02
50
2.69E-OI
2.41E-01
2.17E-01
1.96E-01
.81E-01
.66E-01
.52E-01
.39E-01
.30E-OI
.24E-01
.I7E-01
1.10E-01
1.02E-OI
9.61 E-02
9.15E-02
8.76E-02
8.33E-02
7.93E-02
7.70E-02
7.52E-02
7.68E-02
7.27E-02
7.08E-02
6.87E-02
6.66E-02
6.29E-02
TAMS/MCA
-------
TABLE 3-53: SUMMARY OF ADDExpected FOR FEMALE BAT
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
6.67E-05
5.98E-05
5.78E-05
5.71E-05
5.40E-05
5.17E-05
4.90E-05
4.93E-05
4.82E-05
4.61E-05
4.33E-05
4.26E-05
4.15E-05
3.98E-05
3.93E-05
3.79E-05
3.72E-05
3.65E-05
3.50E-05
3.41E-05
3.34E-05
3.31E-05
3.21E-05
3.24E-05
3.20E-05
3.12E-05
Total Average Dietary Dose
(mg/Kg/day)
113 90
5.29E-05
4.96E-05
4.75E-05
4.57E-05
4.38E-05
4.26E-05
4.13E-05
3.96E-05
3.90E-05
3.77E-05
3.59E-05
3.47E-05
3.43E-05
3.33E-05
3.26E-05
3.14E-05
3.05E-05
2.99E-05
2.96E-05
2.89E-05
2.83E-05
2.76E-05
2.66E-05
2.59E-05
2.56E-05
2.56E-05
4.30E-05
4.08E-05
3.82E-05
3.64E-05
3.53E-05
3.36E-05
3.24E-05
3.15E-05
3.06E-05
2.98E-05
2.91E-05
2.76E-05
2.66E-05
2.54E-05
2.48E-05
2.40E-05
2.35E-05
2.31E-05
2.23E-05
2.18E-05
2.11E-05
2.06E-05
2.02E-05
1 .98E-05
1.95E-05
1 .92E-05
50
3.16E-05
2.96E-05
2.81E-05
2.71E-05
2.62E-05
2.48E-05
2.41E-05
2.33E-05
2.26E-05
2.22E-05
2.14E-05
2.05E-05
1 .97E-05
1.88E-05
1 .83E-05
1.78E-05
1 .74E-05
1.71E-05
1 .68E-05
1.65E-05
1.60E-05
1.56E-05
1 .52E-05
1 .49E-05
1.45E-05
1 .42E-05
TAMS/MCA
-------
TABLE 3-54: SUMMARY OF ADD95%UCL FOR FEMALE BAT
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
7.16E-05
6.41E-05
6.20E-05
6.12E-05
5.79E-05
5.55E-05
5.27E-05
5.29E-05
5.17E-05
4.95E-05
4.66E-05
4.59E-05
4.46E-05
4.28E-05
4.23E-05
4.09E-05
4.01E-05
3.93E-05
3.76E-05
3.67E-05
3.59E-05
3.55E-05
3.46E-05
3.51E-05
3.47E-05
3.38E-05
Total 95% UCL Dietary Dose
(mg/Kg/day)
113 90
5.68E-05
5.31E-05
5.10E-05
4.90E-05
4.69E-05
4.56E-05
4.43E-05
4.25E-05
4.19E-05
4.04E-05
3.86E-05
3.73E-05
3.69E-05
3.58E-05
3.51E-05
3.38E-05
3.28E-05
3.22E-05
3.18E-05
3.11E-05
3.05E-05
2.97E-05
2.86E-05
2.79E-05
2.77E-05
2.77E-05
4.62E-05
4.37E-05
4.10E-05
3.90E-05
3.78E-05
3.60E-05
3.47E-05
3.37E-05
3.27E-05
3.19E-05
3.13E-05
2.97E-05
2.86E-05
2.73E-05
2.66E-05
2.58E-05
2.53E-05
2.48E-05
2.40E-05
2.34E-05
2.27E-05
2.22E-05
2.17E-05
2.13E-05
2.10E-05
2.06E-05
50
3.39E-05
3.18E-05
3.02E-05
2.90E-05
2.8 IE-OS
2.65E-05
2.53E-05
2.50E-05
2.42E-05
2.33E-05
2.30E-05
2.20E-05
2.12E-05
2.02E-05
1.97E-05
1.9 IE-OS
I.88E-05
1 .84E-05
1.8 IE-OS
1 .77E-05
1.72E-05
1.67E-05
1 .63E-05
1.60E-05
1.56E-05
1 .53E-05
TAMS/MCA
-------
TABLE 3-55: SUMMARY OF ADDExpected FOR FEMALE RACCOON
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
.47E-05
.31E-05
.24E-05
.25E-05
.18E-05
.11E-05
1.04E-05
.04E-05
1 .03E-05
9.83E-06
9.26E-06
8.97E-06
8.72E-06
8.49E-06
8.31E-06
8.01E-06
7.80E-06
7.70E-06
7.45E-06
7.25E-06
7.12E-06
7.00E-06
6.77E-06
6.74E-06
6.64E-06
6.48E-06
Total Average Dietary Dose
(mg/Kg/day)
113 90
1.15E-05
1.08E-05
1.02E-05
9.85E-06
9.46E-06
9.09E-06
8.74E-06
8.38E-06
8.27E-06
8.00E-06
7.64E-06
7.33E-06
7.21E-06
7.02E-06
6.87E-06
6.62E-06
6.40E-06
6.27E-06
6.22E-06
6.08E-06
5.96E-06
5.81E-06
5.60E-06
5.44E-06
5.36E-06
5.32E-06
9.31E-06
8.82E-06
1.32E-05
1.29E-05
7.59E-06
7.22E-06
6.92E-06
6.69E-06
6.48E-06
6.33E-06
6.16E-06
5.85E-06
5.63E-06
5.39E-06
5.25E-06
5.09E-06
4.97E-06
4.86E-06
4.71E-06
4.60E-06
4.47E-06
4.37E-06
4.27E-06
4.17E-06
4.11E-06
4.02E-06
50
6.95E-06
6.52E-06
6.17E-06
5.91E-06
5.69E-06
5.39E-06
5.20E-06
5.01E-06
4.84E-06
4.74E-06
4.57E-06
4.37E-06
4.20E-06
4.02E-06
3.91E-06
3.80E-06
3.71E-06
3.62E-06
3.56E-06
3.48E-06
3.38E-06
3.29E-06
3.21E-06
3.14E-06
3.06E-06
2.99E-06
TAMS/MCA
-------
TABLE 3-56: SUMMARY OF ADD,5%UCL FOR FEMALE RACCOON
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.58E-05
.42E-05
.37E-05
.36E-05
.29E-05
.23E-05
.18E-05
.18E-05
.15E-05
.11E-05
.05E-05
.04E-05
.02E-05
9.74E-06
9.53E-06
9.35E-06
9.28E-06
8.97E-06
8.55E-06
8.34E-06
8.16E-06
8.05E-06
7.84E-06
8.02E-06
7.97E-06
7.73E-06
Total 95% UCL Dietary Dose
(mg/Kg/day)
113 90
1.25E-05
1.17E-05
1.12E-05
1.08E-05
1.04E-05
1.01E-05
9.83E-06
9.46E-06
9.32E-06
9.03E-06
8.65E-06
8.41E-06
8.32E-06
8.09E-06
7.91E-06
7.66E-06
7.51E-06
7.34E-06
7.24E-06
7.03E-06
6.87E-06
6.70E-06
6.53E-06
6.38E-06
6.33E-06
6.34E-06
1.01E-05
9.55E-06
9.03E-06
8.63E-06
8.32E-06
7.95E-06
7.64E-06
7.44E-06
7.23E-06
7.06E-06
6.90E-06
6.61E-06
6.38E-06
6.14E-06
6.01E-06
5.84E-06
5.71E-06
5.60E-06
5.43E-06
5.30E-06
5.15E-06
5.02E-06
4.91E-06
4.80E-06
4.73E-06
4.64E-06
50
7.51E-06
7.07E-06
6.72E-06
6.45E-06
6.24E-06
5.93E-06
5.75E-06
5.57E-06
5.40E-06
5.30E-06
5.14E-06
4.93E-06
4.76E-06
4.58E-06
4.47E-06
4.36E-06
4.26E-06
4.17E-06
4.10E-06
4.01E-06
3.89E-06
3.80E-06
3.70E-06
3.62E-06
3.53E-06
3.45E-06
TAMS/MCA
-------
TABLE 3-57: SUMMARY OF ADDExpected FOR FEMALE MINK
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
1.59E-05
1.27E-05
1.16E-05
1.30E-05
1.15E-05
9.41E-06
8.57E-06
8.46E-06
8.91E-06
8.14E-06
7.57E-06
6.59E-06
6.43E-06
6.93E-06
6.19E-06
5.85E-06
5.53E-06
5.75E-06
5.88E-06
5.60E-06
5.76E-06
5.52E-06
5.08E-06
4.69E-06
4.55E-06
4.52E-06
Average Dietary Dose
(mg/Kg/day)
113 90
1.13E-05
1.04E-05
9.05E-06
9.12E-06
8.58E-06
7.59E-06
6.99E-06
6.57E-06
6.66E-06
6.46E-06
6.01E-06
5.46E-06
5.24E-06
5.22E-06
5.07E-06
4.82E-06
4.54E-06
4.46E-06
4.56E-06
4.47E-06
4.46E-06
4.30E-06
4.09E-06
3.84E-06
3.69E-06
3.63E-06
9.12E-06
8.40E-06
8.33E-06
7.89E-06
6.73E-06
6.29E-06
5.71E-06
5.38E-06
5.15E-06
5.02E-06
4.81E-06
4.49E-06
4.22E-06
4.02E-06
3.89E-06
3.76E-06
3.60E-06
3.46E-06
3.41E-06
3.38E-06
3.33E-06
3.23E-06
3.14E-06
3.03E-06
2.91E-06
2.83E-06
50
7.90E-06
7.18E-06
6.52E-06
6.08E-06
5.67E-06
5.31E-06
4.89E-06
4.57E-06
4.31E-06
4.16E-06
3.97E-06
3.73E-06
3.52E-06
3.32E-06
3.19E-06
3.07E-06
2.96E-06
2.85E-06
2.79E-06
2.75E-06
2.69E-06
2.6IE-06
2.55E-06
2.47E-06
2.39E-06
2.32E-06
TAMS/MCA
-------
TABLE 3-58: SUMMARY OF ADD95%UCL FOR FEMALE MINK
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
.66E-05
.33E-05
.22E-05
.36E-05
.20E-05
9.94E-06
9.10E-06
8.96E-06
9.40E-06
8.64E-06
8.03E-06
7.06E-06
6.89E-06
7.37E-06
6.61E-06
6.29E-06
5.98E-06
6.16E-06
6.27E-06
5.98E-06
6.13E-06
5.89E-06
5.44E-06
5.08E-06
4.95E-06
4.91E-06
95% UCL Dietary Dose
(mg/Kg/day)
113 90
1.19E-05
1 .09E-05
9.53E-06
9.58E-06
9.04E-06
8.01E-06
7.42E-06
6.97E-06
7.07E-06
6.86E-06
6.39E-06
5.83E-06
5.61E-06
5.59E-06
5.42E-06
5.17E-06
4.89E-06
4.80E-06
4.89E-06
4.78E-06
4.77E-06
4.61E-06
4.39E-06
4.14E-06
3.99E-06
3.94E-06
9.55E-06
8.81E-06
7.95E-06
7.49E-06
7.08E-06
6.63E-06
6.03E-06
5.69E-06
5.45E-06
5.32E-06
5.1IE-06
4.77E-06
4.50E-06
4.28E-06
4.15E-06
4.02E-06
3.86E-06
3.70E-06
3.65E-06
3.62E-06
3.56E-06
3.45E-06
3.36E-06
3.24E-06
3.12E-06
3.04E-06
50
8.25E-06
7.50E-06
6.83E-06
6.37E-06
5.95E-06
5.58E-06
:5.14E-06
4.81E-06
4.54E-06
4.39E-06
4.20E-06
3.95E-06
3.73E-06
3.53E-06
3.39E-06
3.27E-06
3.15E-06
3.04E-06
2.98E-06
2.93E-06
2.87E-06
2.78E-06
2.72E-06
2.64E-06
2.55E-06
2.48E-06
TAMS/MCA
-------
TABLE 3-59: SUMMARY OF ADDExpected FOR FEMALE OTTER
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
2.14E-04
.55E-04
.33E-04
.58E-04
.44E-04
.13E-04
9.76E-05
9.I6E-05
1.04E-04
9.52E-05
8.48E-05
6.85E-05
6.52E-05
7.34E-05
6.77E-05
6.30E-05
5.47E-05
5.74E-05
6.43E-05
5.75E-05
6.27E-05
5.70E-05
5.27E-05
4.82E-05
4.34E-05
4.23E-05
Total Average Dietary Dose
(mg/Kg/day)
113 90
1.43E-04
1.25E-04
1.11E-04
1.04E-04
l.OOE-04
9.13E-05
7.64E-05
6.95E-05
7.04E-05
6.97E-05
6.40E-05
5.68E-05
5.22E-05
5.16E-05
5.05E-05
4.86E-05
4.44E-05
4.24E-05
4.51E-05
4.41E-05
4.55E-05
4.31E-05
4.11E-05
3.83E-05
3.53E-05
3.35E-05
3.30E-05
3.00E-05
2.76E-05
2.52E-05
2.31E-05
2.12E-05
.91E-05
.74E-05
.65E-05
.60E-05
.52E-05
.42E-05
.31E-05
.24E-05
1.19E-05
1.15E-05
1.09E-05
1.03E-05
1 .02E-05
1.02E-05
1.04E-05
9.85E-06
9.56E-06
9.14E-06
8.67E-06
8.22E-06
50
3.10E-05
2.78E-05
2.50E-05
2.26E-05
2.09E-05
1.92E-05
1.75E-05
1.60E-05
1.50E-05
1.43E-05
1.35E-05
1 .27E-05
1.18E-05
1.11E-05
1.06E-05
1.01E-05
9.61E-06
9.15E-06
8.88E-06
8.68E-06
8.87E-06
8.39E-06
8.17E-06
7.93E-06
7.68E-06
7.25E-06
TAMS/MCA
-------
TABLE 3-60: SUMMARY OF \DD9S%VCL FOR FEMALE OTTER
ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
152
2.19E-04
.58E-04
.36E-04
.61E-04
.47E-04
.15E-04
9.98E-05
9.36E-05
1 .06E-04
9.73E-05
8.68E-05
7.01E-05
6.68E-05
7.52E-05
6.93E-05
6.44E-05
5.60E-05
5.88E-05
6.58E-05
5.88E-05
6.42E-05
5.84E-05
5.40E-05
4.93E-05
4.45E-05
4.34E-05
Total 95% UCL Dietary Dose
(mg/Kg/day)
1 13 90
1 .46E-04
1.28E-04
1.14E-04
1.06E-04
1.02E-04
9.33E-05
7.82E-05
7.12E-05
7.20E-05
7.13E-05
6.55E-05
5.81E-05
5.35E-05
5.28E-05
5.17E-05
4.97E-05
4.55E-05
4.34E-05
4.62E-05
4.5IE-05
4.66E-05
4.4IE-05
4.21E-05
3.92E-05
3.62E-05
3.44E-05
3.35E-05
3.05E-05
2.75E-05
2.51E-05
2.35E-05
2.16E-05
1.95E-05
.77E-05
.68E-05
.64E-05
.55E-05
.44E-05
.33E-05
.26E-05
.22E-05
.18E-05
.IIE-05
.06E-05
.05E-05
.04E-05
.06E-05
.01E-05
9.77E-06
9.35E-06
8.87E-06
8.41E-06
50
3.15E-05
2.83E-05
2.54E-05
2.30E-05
2.I3E-05
1.95E-05
.78E-05
.63E-05
.53E-05
.46E-05
.38E-05
.29E-05
.20E-05
.13E-05
.08E-05
.03E-05
9.81E-06
9.35E-06
9.07E-06
8.87E-06
9.05E-06
8.57E-06
8.34E-06
8.10E-06
7.84E-06
7.41E-06
TAMS/MCA
-------
TABI.K4-I
TOXICITY RKKF.RKNCK VALURS FOR FISH
DIKTAKY IXJSKS AND K(i(i CONCKNTRATIONS OHTOTAI. PCBs AND D1OX1N TOXIC EQUIVAI.KNTS (TEQs)
TRVs
Pumpklnseed
(Lepomis gibbttsus )
Tissue Concentration
Lab-based TRVs for PCBs (mg/kg wel wl.)
Field-based TRVs for PCBs (mg/kg wel wt.)
I.OAKI.
NOAKI.
I.OAKI.
NOAHI.
1.5
0.16
NA
0.5
Spottail
Shiner
(Nolropis
hudsttnius )
15
1.6
NA
NA
Brown Bullhead
(Ictalurus nebulosus )
Yellow Pen*
(Perca Jlavescens )
1.5
0.16
NA
NA
1.5
0.16
NA
NA
While Perch
[Morone americana )
Largemouth Bass
(Micropterus
salmoides )
1.5
0.16
NA
3.1
1.5
0.16
NA
0.5
Egg Concentration
Lab-based TRY for TEQs (ug/kg lipid)
from salmonids
Lab-based TRY for TEQs (ug/kg lipid)
1'roni non-salmcmids
Field-based TRVs for TEQs (ug/kg lipid)
I.OAKI.
NOAKI.
I.OAKI.
NOAKI.
I.OAKI.
NOAKI.
0.6
0.29
1(1.3
0.54
NA
NA
Nut derived
Nol derived
103
5.4
NA
NA
18
8.0
Nol derived
Nol derived
NA
NA
0.6
0.29
10.3
(1.54
NA
NA
0.6
0.29
10.3
0.54
NA
NA
0.6
0.29
10.3
0.54
NA
NA
Striped Bass
(Morone saxatilus )
Sbortnose Sturgeon
(Adpenser
brevifostrunt )
1.5
(1.16
NA
3.1
/.5
0.16
NA
NA
References
Dengisson(1980)
While perch and striped bass: Weslin et
al. (1983)
Pumpkinseed and Largcmoulh bass:
Adams el al. (1989, 1990, 1992)
0.6
0.29
10.3
0.54
NA
NA
0.6
0.29
10.3
0.54
NA
NA
Brown Bullhead: Eloncneial. ( 1998)
All olhers: Walker el al. (1994)
Oliveri and Cooper ( 1 997)
Note:
* I'umpkinseed (LrjHtinh gibbimis) and spollail shiner (Notrapis hu
Uniis vary lor I'CUs and TKQ.
NA = Nol available
Selected TRVs are boldid and italicized.
TAMS/MCA
-------
TABLE 4-2
TOXICITY REFERENCE VALUES FOR BIRDS
DIETARY DOSES AND EGG CONCENTRATIONS OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
TRVs
Tree Swallow
(Tachycineta bicolor )
Mallard Duck
(Anas platyrhychos )
Belted Kingfisher
(Ceryle alcyon )
Great Blue Heron
(Ardea herodias )
Bald Eagle
(Haliaeetus
leucocephalus )
References
Dietary Dose
Lab-based TRVs for PCBs (mg/kg/day)
Field-based TRVs for PCBs (mg/kd/day)
Lab-based TRVs for TEQs (ug/kg/day)
Field-based TRVs for TEQs (ug/kg/day)
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
0.07
0.01
NA
16.1
0.014
0.0014
NA
4.9
2.6
0.26
NA
NA
0.014
0.0014
NA
NA
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
Mallard: Custer and Heinz (1980)
All others: Scott (1977)
Tree Swallow: US EPA Phase 2 Database (1998)
Noseketal. (1992)
US EPA Phase 2 Database (1998)
Egg Concentration
Lab-based TRVs for PCBs (mg/kg egg)
Field-based TRVs for PCBs (mg/kg egg)
Lab-based TRVs for TEQs (ug/kg egg)
Field-based TRVs for TEQs (ug/kg egg)
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
2.21
0.33
NA
26.7
0.02
0.01
NA
13
2.21
0.33
NA
NA
0.02
0.01
0.02
0.005
2.27
0.33
NA
NA
0.02
0.01
NA
NA
2.21
0.33
NA
NA
NA
2
0.5
0.3
2.21
0.33
NA
3.0
0.02
0.01
NA
NA
Scott (1977)
Bald Eagle: Wiemeyer(1984, 1993)
Tree Swallow: US EPA Phase 2 Database (1998)
Great Blue Heron: Janz and Bellward (1996)
Others: Powell et al. (I996a)
Mallard: While and Segniak (1994); White and Hoffman (1995)
Great Blue Heron: Sanderson et al. (1994)
Tree Swallow: US EPA Phase 2 Database (1998)
Note: Unils vary for I'CBs and TKQ.
NA = Not Available
Selected TRVs are bolded and italicized.
TAMS/MCA
-------
TABLE 4-3
TOXICITV REFERENCE VALUES FOR MAMMALS
DIETARY DOSES OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
TRVs
Lab-based TRVs for PCBs (mg/kg/day)
Field-based TRVs for PCBs (mg/kg/day)
Lab-based TRVs for TEQs (ug/kg/day)
Field-based TRVs for TEQs (ug/kg/day)
I.OAEL
NOAEL
IOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
Little Brown Bat
(Myotis lucifugus )
ft/5
0.032
NA
NA
0.001
0.0001
NA
NA
Raccoon
(Procyon lotor )
ft/5
0.032
NA
NA
0.001
0.0001
NA
NA
Mink
(Muslela
vison)
0.07
0.01
0.13
0.004
0.001
0.0001
0.00224
0.00008
River Otter
(Lutra canadensis )
0.07
0.01
0.13
0.004
0.001
0.0001
0.00224
0.00008
References
Mink and ulten Aulcrich and Ringer ( 1 977)
Raccixm and hilt: Under et al. (1984)
Hcau>nelal. (1995)
Murray el al. (1979)
I'i Hill clal. (1996)
Note: Units vary for PCBs and TEQ.
Note: TRVs for raccoon and bat are based on mulit-generational studies to which interspecies uncertainty factors are applied.
NA = Not Available
Final selected TRVs are bolded and italicized.
TAMS/MCA
-------
TABLE 4-4
WORLD HEALTH ORGANIZATION - TOXIC EQUIVALENCY FACTORS (TEFs)
FOR HUMANS, MAMMALS, FISH, AND BIRDS
Congener
Non-ortho PCBs
3,4,4',5-TetraCB(81)
3,3',4,4'-TetraCB (77)
3,3',4,4',5-PentaCB(126)
3,3',4,41,5,5'-HexaCB(169)
Mono-ortho PCBs
2,3,3',4,4'-PentaCB(105)
2,3,4,4',5-PentaCB(114)
2,3',4,4',5-PentaCB(118)
2',3,4,4',5-PentaCB(123)
2,3,3',4,4',5-HexaCB(156)
2,3,3',4,4',5'-HexaCB(157)
2,3',4,4',5,5'-HexaCB (167)
2,3,3',4,4',5,5'-HeptaCB (189)
Toxic 1
Humans/Mammals
0.0001
0.0001
0.1
0.01
0.0001
0.0005
0.0001
0.0001
0.0005
0.0005
0.00001
0.0001
Equivalency Facto
Fish
0.0005
0.0001
0.005
0.00005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
r
Birds
0.1
0.05
0.1
0.001
0.0001
0.0001
0.00001
0.00001
0.0001
0.0001
0.0000 1
0.00001
Notes: CB = chlorinated biphenyls
Reference: van den Berg et al. 1998. Toxic Equivalency Factors (TEFs) for PCBs, PCDDs,
PCDFs for Humans and Wildlife. Environmental Health Perspectives, 106:12, 775-791.
TAMS/MCA
-------
TABLE 5-1: RATIO OF PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
Average PCB Results
152 Total 113 Total 90 Total 50 Total 152 Total
Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
NOAA Consensus-Based Total PCB TEC:
24
22
20
20
20
18
17
17
17
16
15
15
14
14
14
13
13
13
12
12
12
11
11
11
10
10
19 15
18 15
17 55
16 54
16 13
15 12
14 11
14 11
13 11
13 10
13 10
12 9.7
12 9.3
11 9.0
11 8.7
11 8.5
11 8.2
10 8.0
10 7.8
9.9 7.6
9.7 7.4
9.4 7.3
9.2 7.1
8.9 6.9
8.7 6.7
8.5 6.5
11
11
10
9.7
9.3
8.9
8.5
8.2
7.9
7.6
7.4
7.1
6.9
6.7
6.5
6.3
6.1
5.9
5.8
5.6
5.5
5.3
5.2
5.1
5.0
4.8
27
26
25
24
24
24
23
23
22
22
21
22
22
20
20
20
21
19
18
17
17
17
16
18
18
17
Fri+ 95% UCL Results
113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone
0.04 mg/kg dry weight
21 17
21 16
20 16
20 16
19 15
19 15
19 14
19 14
18 14
18 14
17 13
17 13
17 13
17 13
16 13
16 12
16 12
16 12
15 11
15 11
14 11
14 11
14 10
14 10
14 9.9
14 9.7
13
12
12
11
11
11
11
11
10
10
9.9
9.7
9.5
9.3
9.3
9.1
8.9
8.7
8.5
8.3
8.1
7.9
7.7
7.5
7.3
7.2
Average PCB Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
NOAA Consensus-Based Total PCB MEC: 0.4 mg/kg dry weight
2.4
2.2
2.0
2.0
2.0
1.8
1.7
1.7
1.7
1.6
1.5
1.5
1.4
1.4
1.4
1.3
1.3
1.3
1.2
1.2
1.2
1.1
1.1
1.1
1.0
1.0
.9 1.5
.8 1.5
.7 5.5
.6 5.4
.6 1.3
1.5 1.2
1.4 1.1
1.4 1.1
.3 1.1
.3 1.0
.3 1.0
.2 1.0
.2 0.9
.1 0.9
.1 0.9
.1 0.8
.1 0.8
.0 0.8
1.0 0.8
1.0 0.8
1.0 0.7
0.9 0.7
0.9 0.7
0.9 0.7
0.9 0.7
0.8 0.7
1.1
1.1
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
2.7
2.6
2.5
2.4
2.4
2.4
2.3
2.3
2.2
2.2
2.1
2.2
2.2
2.0
2.0
2.0
2.1
1.9
1.8
1.7
1.7
1.7
1.6
1.8
1.8
1.7
2.1 1.7 .3
2.1 1.6 .2
2.0 1.6 .2
2.0 1.6 .1
1.9 1.5 .1
1.9 1.5 .1
1.9 1.4 .1
1.9 1.4 .1
1.8 1.4 .0
1.8 1.4 .0
1.7 1.3 .0
1.7 1.3 .0
1.7 1.3 .0
1.7 1.3 0.9
1.6 1.3 0.9
1.6 1.2 0.9
1.6 1.2 0.9
1.6 1.2 0.9
1.5 1.1 0.9
1.5 1.1 0.8
1.4 1.1 0.8
1.4 1.1 0.8
1.4 1.0 0.8
1.4 1.0 0.8
1.4 1.0 0.7
1.4 1.0 0.7
exceedances are bolded
Page 1 of 5
TAMS/MCA
-------
TABLE 5-1: RATIO OF PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES (CONT.)
Average PCB Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Year Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
NOAA Consensus-Based Total PCB EEC: 1 .7 mg/kg dry weight
1993 0.6 0.4 0.4 0.3 0.6 0.5 0.4 0.3
1994 0.5 0.4 0.3 0.3 0.6 0.5 0.4 0.3
1995 0.5 0.4 1.3 0.2 0.6 0.5 0.4 0.3
1996 0.5 0.4 1.3 0.2 0.6 0.5 0.4 0.3
1997 0.5 0.4 0.3 0.2 0.6 0.5 0.4 0.3
1998 0.4 0.4 0.3 0.2 0.6 0.5 0.3 0.3
1999 0.4 0.3 0.3 0.2 0.6 0.4 0.3 0.3
2000 0.4 0.3 0.3 0.2 0.5 0.4 0.3 0.2
2001 0.4 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2002 0.4 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2003 0.4 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2004 0.3 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2005 0.3 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2006 0.3 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2007 0.3 0.3 0.2 0.2 0.5 0.4 0.3 0.2
2008 0.3 0.3 0.2 0. 0.5 0.4 0.3 0.2
2009 0.3 0.2 0.2 0. 0.5 0.4 0.3 0.2
2010 0.3 0.2 0.2 0. 0.5 0.4 0.3 0.2
2011 0.3 0.2 0.2 0. 0.4 0.4 0.3 0.2
2012 0.3 0.2 0.2 0. 0.4 0.3 0.3 0.2
2013 0.3 0.2 0.2 0. 0.4 0.3 0.3 0.2
2014 0.3 0.2 0.2 0. 0.4 0.3 0.2 0.2
2015 0.3 0.2 0.2 0. 0.4 0.3 0.2 0.2
2016 0.3 0.2 0.2 0. 0.4 0.3 0.2 0.2
2017 0.2 0.2 0.2 0.1 0.4 0.3 0.2 0.2
2018 0.2 0,2 0.2 0.1 0.4 0.3 0.2 0.2
Average PCB Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
"4YSDEC Benthic Chronic Total PCB 19.3 mg/Kg OC (0.482 mg/kg using 2.5% OC
2.0 1.6 1.3 0.9 2.2 .8
1.4 1.0
1.8 1.5 1.2 0.9 2.1 .7 1.4 1.0
1.7 1.4 4.5 0.8 2.1 .7
1.4 1.0
.7 1.3 4.5 0.8 2.0 .6 1.3 1.0
.6 1.3 1.0 0.8 2.0 .6
.5 1.2 1.0 0.7 2.0 .6
.4 1.2 1.0 0.7 .9 .6
.4 1.1 0.9 0.7 .9 .5
.4 1.1 0.9 0.7 .8 .5
.3 1.1 0.9 0.6 .8 1.5
.3 1.0 0.8 0.6 .8 1.4
.2 1.0 0.8 0.6 .8 1.5
.2 1.0 0.8 0.6 1.8 1.4
.2 0.9 0.7 0.6 1.7
.1 0.9 0.7 0.5 1.6
.1 0.9 0.7 0.5 1.7
.1 0.9 0.7 0.5 1.7
.0 0.9 0.7 0.5 1.6
.0 0.8 0.6 0.5 1.5
.0 0.8 0.6 0.5 1.4
.0 0.8 0.6 0.5 1.4
0.9 0.8 0.6 0.4 1.4
0.9 0.8 0.6 0.4 1.4
0.9 0.7 0.6 0.4 1.5
0.9 0.7 0.6 0.4 1.5
0.8 0.7 0.5 0.4 1.4
.4
.4
.3
1.3 0.9
1.2 0.9
.2 0.9
.2 0.9
.1 0.9
.1 0.8
.1 0.8
.1 0.8
.1 0.8
.0 0.8
.0 0.8
.0 0.8
.4 1.0 0.7
.3 1
1.0 0.7
.3 0.9 0.7
.2 0.9 0.7
.2 0.9 0.7
.2 0.9 0.7
.2 0.9 0.6
.2 0.8 0.6
.2 0.8 0.6
.2 0.8 0.6
Page 2 of 5
TAMS/MCA
-------
TABLE 5-1: RATIO OF PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES (CONT.)
Average PCB Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Year Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
NYSDEC Wildlife Total PCB 1.4 mg/Kg OC (0.035 mg/kg using 2.5% OC)
1993 28
1994 25
1995 23
1996 23
1997 22
1998 21
1999 19
2000 19
2001 19
2002 18
2003 18
2004 17
2005 16
2006 16
2007 16
2008 15
2009 15
2010 14
201 1 14
2012 14
2013 13
2014 13
2015 13
2016 12
2017 12
2018 12
22 17
21 17
19 62
19 62
18. 14
17 14
16 13
16 13
15 12
15 12
14 11
14 11
13 11
13 10
13 10
12 9.7
12 9.4
12 9.1
12 8.9
11 8.7
11 8.5
11 8.3
10 8.1
10 7.9
9.9 7.7
9.7 7.5
13 31
12 29
12 29
11 28
11 27
10 27
9.7 27
9.3 26
9.0 25
8.7 25
8.4 24
8.2 25
7.9 25
7.6 23
7.4 23
7.2 23
7.0 24
6.8 22
6.6 20
6.4 20
6.3 19
6.1 19
6.0 19
5.8 20
5.7 20
5.5 19
25 19
24 19
23 19
23 18
22 17
22 17
22 16
21 16
21 16
20 15
20 15
20 15
20 15
19 14
19 14
18 14
19 14
18 13
18 13
17 13
16 12
16 12
16 12
16 12
16 11
16 11
14
14
14
13
13
13
12
12
12
11
11
11
11
11
11
10
10
9.9
9.7
9.5
9.2
9.0
8.8
8.6
8.4
8.2
152 Total
Sed Cone
97
88
81
81
79
73
68
67
67
65
62
59
57
56
55
53
51
50
49
48
47
46
44
43
42
41
Average PCB Results
113 Total 90 Total
Sed Cone Sed Cone
Persaud Total
76
72
68
65
63
60
57
55
54
52
51
49
47
46
45
43
42
41
40
39
39
38
37
36
35
34
61
58
218
218
50
48
46
44
42
41
40
39
37
36
35
34
33
32
31
30
30
29
28
28
27
26
Tri+ 95% UCL Results
50 Total 152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
PCB NEL 0.01 mg/Kgjlry weight
45
43
41
39
37
36
34
33
31
31
30
29
28
27
26
25
24
24
23
22
22
21
21
20
20
19
107
102
100
98
95
94
94
91
87
87
85
87
87
81
79
81
84
77
71
70
68
67
66
71
71
68
86 68
84 66
82 65
79 63
78 61
77 59
76 57
74 57
73 55
71 54
70 53
70 52
69 51
67 50
66 50
65 49
66 48
64 47
62 46
59 44
57 43
56 42
56 41
56 40
56 39
56 39
50
49
47
46
45
44
43
42
41
40
40
39
38
37
37
36
36
35
34
33
32
32
31
30
29
29
Page 3 of 5
TAMS/MCA
-------
TABLE 5-1: RATIO OF PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES (CONT.)
152 Total
Year Sed Cone
1993 14
1994 13
1995 12
1996 12
1997 11
1998 10
1999 10
2000 10
2001 10
2002 9
2003 9
2004 8
2005 8
2006 8
2007 8
2008 8
2009 7
2010 7
2011 7
2012 7
2013 7
2014 7
2015 6
2016 6
2017 6
2018 6
Average PCB Results
113 Total 90 Total
Sed Cone Sed Cone
Persaud Total
11
10
10
9
9
9
8
8
8
7
7
7
7
7
6
6
6
6
6
6
6
5
5
5
5
5
9
8
31
31
7
7
7
6
6
6
6
6
5
5
5
5
5
5
4
4
4
4
4
4
4
4
50 Total
Sed Cone
PCB LEL
6
6
6
6
5
5
5
5
4
4
4
4
4
4
4
4
3
3
3
3
3
3
3
3
3
3
Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone
0.07 mg/Kg dry weight
15
15
14
14
14
13
13
13
12
12
12
12
12
12
11
12
12
11
10
10
10
10
9
10
10
10
12
12
12
11
11
11
11
11
10
10
10
10
10
10
9
9
9
9
9
8
8
8
8
8
8
8
10
9
9
9
9
8
8
8
8
8
8
7
7
7
7
7
7
7
7
6
6
6
6
6
6
6
7
7
7
7
6
6
6
6
6
6
6
6
5
5
5
5
5
5
5
5
5
5
4
4
4
4
Average PCB Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
Persaud Total PCB SEL 530 mg/Kg OC (1.3 mg/kg using 2.5% OC)
0.
0.
0.
0.
0.
0.
0.
0.
0.1
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.1 0.0
0.1 0.0
0.1 0.2
0.0 0.2
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0 0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.0
0.1 0.1 0.1 0.0
0.1 0.1 0.0 0.0
0.1 0.1 0.0 0.0
0.1 0.1 0.0 0.0
0.1 0.1 0.0 0.0
0.1 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.1 0.0 0.0
0. 0.0 0.0 0.0
0. 0.0 0.0 0.0
0.1 0.0 0.0 0.0
0.1 0.0 0.0 0.0
0. 0.0 0.0 0.0
0. 0.0 0.0 0.0
0. 0.0 0.0 0.0
0. 0.0 0.0 0.0
0.0 0.0 0.0 0.0
0. 0.0 .0.0 0.0
0. 0.0 0.0 0.0
0. 0.0 0.0 0.0
Page 4 of 5
TAMS/MCA
-------
TABLE 5-1: RATIO OF PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES (CONT.)
Average PCS Results Tri+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Year Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
Washington State Total PCB PAET Hyalella azteca 0.45 mg/Kg dry weight
1993 2.1
1994 2.0
1995 1.8
1996 1.8
1997 1.7
1998 1.6
1999 1.5
2000 1.5
2001 1.5
2002 1.4
2003 1.4
2004 1.3
2005 1.3
2006 1.2
2007 1.2
2008 1.2
2009 1.1
2010 1.1
2011 1.1
2012 1.1
2013 1.0
2014 1.0
2015 1.0
2016 1.0
2017 0.9
2018 0.9
1.7 1.4
1.6 1.3
1.5 4.8
1.4 4.8
1.4 1.1
1.3 1.1
1.3 1.0
1.2 1.0
1.2 0.9
1.2 0.9
1.1 0.9
1.1 0.9
1.0 0.8
1.0 0.8
1.0 0.8
1.0 0.8
0.9 0.7
0.9 0.7
0.9 0.7
0.9 0.7
0.9 0.7
0.8 0.6
0.8 0.6
0.8 0.6
0.8 0.6
0.8 0.6
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
' 0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
2.4
2.3
2.2
2.2
2.1
2.1
2.1
2.0
.9
.9
.9
.9
.9
.8
.8
.8
.9
.7
.6
.6
.5
.5
.5
.6
.6
.5
.9 1.5
.9 1.5
.8 .4
.8 .4
.7 .3
.7 .3
.7 .3
.7 .3
.6 .2
1.6 .2
1.5 .2
1.6 .2
1.5 .1
1.5 1.1
1.5 1.1
1.4 1.1
1.5 1.1
.4 1.0
.4 1.0
.3 1.0
.3 1.0
.2 0.9
.2 0.9
.2 0.9
.2 0.9
.2 0.9
1.1
1.1
.1
.0
.0
.0
.0
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.6
152 Total
Sed Cone
46
42
38
39
37
35
32
32
32
31
29
28
27
27
26
25
24
24
24
23
22
22
21
20
20
19
Average PCB Results Tri+ 95% UCL Results
113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone Sed Cone
Washington Total PCB PAET Microtox 0.021 mg/Kg dry weight
36 29
34 28
32 104
31 104
30 24
29 23
27 22
26 21
26 20
25 20
24 19
23 18
22 18
22 17
21 17
21 16
20 16
20 15
19 15
19 15
18 14
18 14
17 14
17 13
17 13
16 12
21
20
19
18
18
17
16
16
15
15
14
14
13
13
12
12
12
11
11
11
10
10
9.9
9.7
9.4
9.2
51
49
48
47
45
45
45
43
41
41
40
42
42
39
38
39
40
37
34
33
32
32
31
34
34
32
41 32
40 31
39 31
38 30
37 29
36 28
36 27
35 27
35 26
34 26
33 25
33 25
33 24
32 24
31 24
31 23
31 23
30 22
29 22
28 21
27 21
27 20
27 20
27 19
26 19
27 18
32
31
31
30
29
28
27
27
26
26
25
25
24
24
24
23
23
22
22
21
21
20
20
19
19
18
Page 5 of 5
TAMS/MCA
-------
TABLE 5-2: RATIO OF PREDICTED WHOLE WATER CONCENTRATIONS TO CRITERIA AND BENCHMARKS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
Tri+ Average PCB Results Tri+ 95% UCL Results
152 113 152 113
Whole Whole 90 Whole 50 Whole Whole Whole 90 Whole 50 Whole
Water Water Water Water Water Water Water Water
Cone Cone Cone Cone Cone Cone Cone Cone
USEPA/NYSDEC Total PCB - FW Benthic Aquatic Life 0.014 ug/L
3.1 2.2 1.6 1.3 4.4 2.7 2.0 1.6
2.9 .9 1.4 1.1 3.5 2.2 1.7 1.4
1.2 .2 I.I 1.0 1.3 1.3 1.4 1.2
3.4 .9 1.3 1.0 5.0 2.3 1.5 1.1
2.2 .5 1.1 0.9 2.9 1.8 1.4 I.I
1.3 .1 1.0 0.8 1.4 1.3 1.1 0.9
1.1 0.9 0.8 0.7 1.2 1.0 1.0 0.8
1.8 .1 0.8 0.6 2.2 1.3 1.0 0.8
2.0 .2 0.8 0.6 2.9 1.5 1.0 0.7
1.2 0.9 0.7 0.6 1.5 1.1 0.9 0.7
1.3 0.9 0.7 0.5 1.8 1.1 0.8 0.6
0.7 0.6 0.6 0.5 0.8 0.7 0.7 0.6
1.0 0.6 0.5 0.4 1.3 0.8 0.6 0.5
1.3 0.8 0.5 0.4 1.8 0.9 0.6 0.5
1.4 0.8 0.5 0.4 2.3 1.0 0.6 0.5
0.6 0.5 0.4 0.4 0.6 0.6 0.5 0.4
0.6 0.5 0.4 0.3 0.7 0.5 0.5 0.4
1.1 0.6 0.4 0.3 1.6 0.8 0.5 0.4
1.1 0.7 0.4 0.3 1.8 0.8 0.5 0.4
0.7 0.5 0.4 0.3 0.9 0.7 0.5 0.4
1.0 0.6 0.4 0.3 . 1.4 0.7 0.5 0.4
0.8 0.5 0.4 0.3 0.9 0.6 0.5 0.4
0.7 0.5 0.4 0.3 0.9 0.6 0.5 0.4
0.4 0.4 0.3 0.3 0.4 0.4 0.4 0.3
0.4 0.3 0.3 0.2 0.4 0.4 0.3 0.3
0.5 0.4 0.3 0.2 0.8 0.5 0.4 0.3
Tri+ Average PCB Results Tri+ 95% UCL Results
152 113 152 113
Whole Whole 90 Whole 50 Whole Whole Whole 90 Whole 50 Whole
Water Water Water Water Water Water Water Water
Cone Cone Cone Cone Cone Cone Cone Cone
NYSDEC Total PCB - Wildlife Bioaccumulation 0.001 ug/L
44 30 23 18 61 38 28 22
40 26 20 16 49 31 24 19
16 16 16 14 18 19 19 16
47 26 18 13 69 32 21 16
31 21 16 12 40 25 19 15
18 15 13 11 20 18 16 13
16 13 11 10 17 15 14 11
26 15 11 9.0 31 18 13 11
29 17 12 8.7 40 21 14 10
17 13 10 8.0 20 15 12 10
19 13 10 7.5 25 15 12 9.0
10 8.6 7.8 6.5 11 10 9.3 7.8
14 9.1 7.2 6.0 18 11 8.5 7.0
19 II 7.5 5.8 26 13 8.8 6.8
19 II 7.4 5.5 32 14 8.7 6.5
7.9 7.0 6.1 5.0 8.7 8.0 7.2 5.9
8.5 6.5 5.6 4.6 10 7.6 6.6 5.5
15 8.8 6.1 4.6 23 11 7.2 5.5
15 9.1 6.2 4.6 25 12 7.3 5.4
10 7.7 5.9 4.5 13 9.2 7.1 5.4
14 8.6 6.0 4.4 20 10 7.1 5.2
11 7.5 5.7 4.3 13 8.6 6.7 5.1
10 7.1 5.4 4.1 12 8.1 6.4 4.9
5.4 5.0 4.6 3.8 5.9 5.7 5.4 4.5
5.1 4.4 4.1 3.5 5.7 5.0 4.8 4.1
7.6 5.4 4.3 3.4 11 6.8 5.2 4.1
Tri+ Average PCB Results Tri+ 95% UCL Results
152 113 152 113
Whole Whole 90 Whole 50 Whole Whole Whole 90 Whole 50 Whole
Water Water Water Water Water Water Water Water
Cone Cone Cone Cone Cone Cone Cone Cone
USEPA Total PCB - Wildlife Criterion 1 .2E-04 affl
37 25 19 15 51 32 23 18
33 22 17 13 41 26 20 16
13 14 13 11 15 16 16 14
39 22 IS 11 58 27 17 13
26 18 13 10 34 21 16 12
15 13 11 9.1 17 15 13 11
13 11 10 8.0 14 12 11 9.6
21 13 9.5 7.5 26 15 11 8.9
24 14 9.6 7.3 34 17 11 8.7
14 11 8.5 6.7 17 13 10 8.0
16 11 8.1 6.3 21 13 10 7.5
8.4 7.2 6.5 5.4 9.3 8.2 7.7 6.5
12 7.6 6.0 5.0 15 8.8 7.1 5.9
16 8.9 6.2 4.8 22 11 7.3 5.7
16 9.1 6.1 4.6 26 12 7.3 5.5
6.6 5.8 5.1 4.2 7.2 7 6.0 4.9
7.1 5.4 4.6 3.8 8.5 6.3 5.5 4.6
13 7.3 5.1 3.9 19 9.1 6.0 4.6
13 7.6 5.1 3.8 20 9.6 6.1 4.5
8.7 6.4 4.9 3.7 11 7.7 5.9 4.5
11 7.2 5.0 3.7 17 8.7 5.9 4.4
9.0 6.3 4.7 3.6 11 7.2 5.6 4.3
8.7 5.9 4.5 3.5 10.2 6.7 5.3 4.1
4.5 4.2 3.8 3.1 4.9 4.8 4.5 3.7
4.3 3.7 3.4 2.9 4.7 4.2 4.0 3.4
6.4 4.5 3.6 2.9 8.8 5.7 ' 4.3 ' 3.4
excecdanccs arc bolded
-------
TABLE 5-3: RATIO OF PREDICTED PUMPKINSEED CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg w
weight)
2.3
1.7
1.5
1.8
1.6
1.1
0.9
1.0
1.1
0.9
0.9
0.6
0.7
0.8
0.6
0.6
0.5
0.6
0.6
0.6
0.6
0.6
0.5
0.4
0.4
0.4
River Mile
Median
et (mg/kg w
weight)
3.1
2.3
2.1
2.5
2.1
1.5
1.4
1.3
1.5
1.3
1.2
0.9
0.9
1.1
0.9
0.8
0.7
0.8
0.9
0.8
0.9
0.8
0.7
0.6
0.6
0.6
152
95th
Percentile
et (mg/kg wet
weight)
5.1
3.7
3.4
4.1
3.4
2.6
2.3
2.2
2.4
2.2
2.0
1.6
1.5
1.8
.5
.4
.3
.4
.5
.4
.5
1.4
1.2
1.0
1.0
1.0
25th
(mg/kg w
weight)
1.5
1.3
1.1
1.2
1.0
0.8
0.7
0.7
0.7
0.7
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
River Mile 1
Median
et (mg/kg we
weight)
2.1
.9
.5
.6
.5
.2
.1
.0
1.0
1.0
0.9
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
13
95th
Percentile
t (mg/kg wet
weight)
3.5
3.1
2.6
2.7
2.5
2.0
1.8
1.7
1.7
1.7
1.5
1.3
1.2
1.3
1.2
1.1
1.0
1.0
1.1
1.1
1.1
1.0
1.0
0.9
0.8
0.8
25th
(mg/kg w«
weight)
1.2
1.1
0.9
0.9
0.8
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
River Mile
Median
;t (mg/kg w
weight)
1.7
1.5
1.3
1.2
1.1
1.1
0.9
0.8
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
90
95th
Percentile
;t (mg/kg wet
weight)
2.7
2.5
2.2
2.0
.9
.7
.5
.4
.3
.3
.2
1.1
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
25th
(mg/kg w<
weight)
1.1
1.0
0.9
0.8
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
River Mile
Median
:t (mg/kg w<
weight)
1.6
1.4
1.3
1.2
1.1
1.0
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
50
95th
Percentile
;t (mg/kg wet
weight)
2.6
2.3
2.1
1.9
1.7
1.6
1.4
1.3
1.2
1.2
1.1
1.0
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-4: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
20 n
2014
2015
2016
2017
2018
0.22
0.18
0.14
0.18
0.16
0.11
0.10
0.10
0.12
0.09
0.08
0.07
0.07
0.08
0.06
0.06
0.05
0.07
0.06
0.06
0.06
0.05
0.05
0.04
0.04
0.04
0.29
0.25
0.18
0.25
0.20
0.14
0.12
0.14
0.15
0.12
0.11
0.09
0.09
0.11
0.09
0.07
0.07
0.09
0.08
0.08
0.08
0.07
0.07
0.05
0.05
0.05
0.48
0.39
0.32
0.41
0.32
0.22
0.20
0.22
0.24
0.19
0.18
0.14
0.15
0.18
0.14
0.12
0.11
0.13
0.13
0.13
0.14
0.12
0.12
0.09
0.08
0.09
River Mile 1 1
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.16
0.14
0.11
0.12
0.11
0.09
0.08
0.07
0.08
0.07
0.07
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.05
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.21
0.19
0.15
0.17
0.14
0.12
0.10
0.10
0.11
0.09
0.09
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.04
0.04
3
95th
Percentile
(mg/kg wet
weight)
0.31
0.28
0.22
0.25
0.21
0.18
0.16
0.16
0.16
0.14
0.14
0.12
0.11
0.11
0.11
0.10
0.09
0.09
0.10
0.10
0.10
0.09
0.09
0.08
0.07
0.07
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.13
0.12
0.10
0.10
0.09
0.08
0.07
0.06
0.06
0.06
0.05
0.05
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.17
0.15
0.13
0.12
0.11
0.10
0.09
0.08
0.08
0.08
0.07
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.24
0.22
0.19
0.18
0.17
0.15
0.13
0.13
0.12
0.12
0.11
0.10
0.09
0.09
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.13
0.11
0.10
0.09
0.08
0.07
0.07
0.06
0.06
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.16
0.14
0.13
0.11
0.11
0.09
0.09
0.08
0.07
0.07
0.07
0.06
0.06
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.23
0.21
0.18
0.17
0.15
0.14
0.13
0.12
0.11
0.10
0.10
0.09
0.08
0.08
0.08
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.06
TAMS/MCA
-------
TABLE 5-5: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
0.002
0.002
0.001
0.002
0.001
0.001
0.0009
0.0009
0.001
0.0009
0.0008
0.0006
0.0006
0.0007
0.0006
0.0005
0.0004
0.0006
0.0005
0.0005
0.0006
0.0005
0.0005
0.0004
0.0004
0.0004
0.031
0.027
0.019
0.027
0.021
0.015
0.013
0.015
0.016
0.013
0.012
0.009
0.010
0.012
0.010
0.008
0.007
0.009
0.009
0.008
0.009
0.008
0.008
0.006
0.005
0.006
0.051
0.042
0.034
0.044
0.034
0.023
0.021
0.023
0.026
0.020
0.020
0.015
0.015
0.019
0.015
0.013
0.012
0.014
0.014
0.014
0.015
0.013
0.012
0.009
0.009
0.009
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.017
0.015
0.012
0.013
0.011
0.009
0.008
0.008
0.008
0.008
0.007
0.006
0.005
0.006
0.005
0.005
0.004
0.005
0.005
0.005
0.005
0.004
0.004
0.004
0.003
0.003
0.022
0.021
0.016
0.018
0.015
0.013
0.011
0.011
0.011
0.010
0.010
0.008
0.008
0.008
0.008
0.007
0.006
0.006
0.007
0.007
0.007
0.006
0.006
0.005
0.005
0.005
0.033
0.030
0.023
0.027
0.022
0.019
0.017
0.017
0.017
0.015
0.014
0.012
0.012
0.012
0.012
0.010
0.010
0.010
0.010
0.011
0.010
0.010
0.009
0.008
0.008
0.008
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.014
0.013
0.011
0.010
0.009
0.008
0.007
0.007
0.006
0.006
0.006
0.005
0.005
0.004
0.004
0.004
0.004
0.003
0.004
0.004
0.003
0.003
0.003
0.003
0.003
0.003
0.018
0.016
0.014
0.013
0.012
0.010
0.010
0.009
0.009
0.008
0.008
0.007
0.006
0.006
0.006
0.006
0.005
0.005
0.005
0.005
0.005
0.005
0.005
0.004
0.004
0.004
0.026
0.024
0.020
0.019
0.018
0.016
0.014
0.013
0.013
0.012
0.012
0.011
0.010
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.008
0.007
0.007
0.007
0.006
0.006
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.014
0.012
0.011
0.010
0.009
0.008
0.007
0.007
0.006
0.006
0.005
0.005
0.005
0.004
0.004
0.004
0.004
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003.
0.003
0.017
0.015
0.013
0.012
0.011
0.010
0.009
0.008
0.008
0.008
0.007
0.007
0.006
0.006
0.005
0.005
0.005
0.005
0.005
0.005
0.004
0.004
0.004
0.004
0.004
0.004
0.025
0.022
0.020
0.018
0.016
0.015
0.014
0.012
0.012
0.011
0.010
0.010
0.009
0.009
0.008
0.008
0.007
0.007
0.007
0.007
0.007
0.006
0.006
0.006
0.006
0.006
TAMS/MCA
-------
TABLE 5-6: RATIO OF PREDICTED PUMPKINSEED CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.8
0.6
0.6
0.7
0.6
0.4
0.3
0.4
0.4
0.3
0.3
0.2
0.2
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
1.1
0.9
0.8
1.0
0.8
0.6
0.5
0.5
0.6
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
95th
Percentile
(mg/kg wet
weight)
1.9
1.5
1.4
1.7
1.4
1.0
0.9
0.9
1.0
0.8
0.8
0.6
0.6
0.7
0.6
0.6
0.5
0.6
0.6
0.6
0.6
0.6
0.5
0.4
0.4
0.4
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
0.6
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.8
0.7
0.6
0.6
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
1.4
1.2
1.0
1.1
1.0
0.8
0.7
0.7
0.7
0.7
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.4
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.6
0.6
0.5
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
1.1
1.0
0.9
0.8
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.4
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.
0.
0.
0.
0.
0.09
0.09
0.6
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
1.0
0.9
0.8
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-7: RATIO OF PREDICTED PUMPKINSEED CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg wet
weight)
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.09
0.07
0.07
0.07
River Mile 152
Median
(mg/kg wet
weight)
0.5
0.4
0.4
0.5
0.4
0.3
0.2
0.2
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
95th
Percentile
(mg/kg wet
weight)
0.9
0.7
0.7
0.8
0.7
0.5
0.4
0.4
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
0.3
0.2
0.2
0.2
0.2
0.2
0.
0.
0.
0.
0.
0.
0.09
0.10
0.09
0.08
0.08
0.08
0.08
0.08
0.08
0.08
0.07
0.06
0.06
0.06
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.08
0.7
0.6
0.5
0.5
0.5
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.08
0.08
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.09
0.09
0.09
0.08
0.08
0.08
0.07
0.07
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.08
0.08
0.07
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.06
0.5
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.
0.
0.
0.
0.
0.
0.1
0.1
0.1
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-8: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.05
0.05
0.03
0.05
0.04
0.03
0.02
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.02
0.02
0.01
0.02
0.01
0.01
0.01
0.01
0.01
0.07
0.07
0.05
0.07
0.05
0.04
0.03
0.04
0.04
0.03
0.03
0.02
0.02
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
95th
Percentile
(mg/kg wet
weight)
0.1
0.1
0.09
0.1
0.09
0.07
0.06
0.06
0.07
0.06
0.05
0.04
0.04
0.05
0.04
0.03
0.03
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
25th (mg/kg
wet weight)
0.04
0.04
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
River Mile 113
95th
Median Percentile
(mg/kg wet (mg/kg wet
weight) weight)
0.05
0.05
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.1
0.09
0.07
0.08
0.07
0.06
0.05
0.05
0.05
0.05
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.009
0.009
0.009
0.008
0.008
0.007
0.04
0.04
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.08
0.07
0.06
0.06
0.05
0.05
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.007
0.007
0.04
0.04
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.07
0.07
0.06
0.05
0.05
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-9: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.003
0.003
0.002
0.003
0.002
0.001
0.001
0.001
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.004
0.004
0.003
0.003
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
95th
Percentile
(mg/kg wet
weight)
0.007
0.006
0.005
0.006
0.005
0.004
0.003
0.003
0.004
0.003
0.003
0.002
0.002
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.001
25th (mg/kg
wet weight)
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.000
0.000
0.000
River Mile 113
95th
Median Percentile
(mg/kg wet (mg/kg wet
weight) weight)
0.003
0.003
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.005
0.005
0.004
0.004
0.004
0.003
0.003
0.002
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.004
0.004
0.003
0.003
0.003
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.000
0.000
0.004
0.003
0.003
0.003
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.001
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-10: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
15
13
12
12
11
11
9.5
9.3
9.4
9.0
8.4
7.8
7.6
7.7
7.3
7.0
6.7
6.6
6.7
6.5
6.4
6.2
5.9
5.6
5.5
5.3
21
18
17
17
16
15
14
14
14
13
12
11
11
11
11
10
9.8
9.8
9.7
9.5
9.3
9.0
8.6
8.3
8.0
7.8
34
31
28
29
27
25
23
23
23
22
21
19
19
19
18
17
17
17
16
16
16
15
15
14
13
13
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
11
10.4
9.7
9.3
8.9
8.4
7.8
7.5
7.4
7.2
6.8
6.5
6.2
6.1
6.0
5.8
5.6
5.4
5.4
5.3
5.2
5.0
4.9
4.7
4.6
4.4
16
15
14
13
13
12
11
11
11
10
10
9.5
9.1
8.9
8.7
8.4
8.1
8.0
7.8
7.7
7.6
7.4
7.1
6.9
6.7
6.5
27
25
23
22
22
20
19
18
18
18
17
16
15
15
15
14
14
13
13
13
13
12
12
12
11
11
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
8.9
8.4
7.9
7.5
7.1
6.8
6.4
6.0
5.8
5.7
5.4
5.2
5.0
4.8
4.7
4.5
4.4
4.3
4.2
4.1
4.0
3.9
3.8
3.7
3.6
3.4
13
12
11
11
10
9.8
9.3
8.8
8.5
8.3
8.0
7.6
7.3
7.1
6.9
6.6
6.4
6.2
6.1
6.0
5.8
5.7
5.5
5.4
5.2
5.1
21
20
19
18
17
16
16
15
14
14
13
13
12
12
11
11
11
10
10
10
9.8
9.5
9.3
9.0
8.8
8.6
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
6.9
6.5
6.1
5.7
5.4
5.2
4.9
4.6
4.4
4.3
4.1
4.0
3.8
3.7
3.5
3.4
3.3
3.2
3.1
3.1
3.0
2.9
2.8
2.8
2.7
2.6
9.8
9.2
8.7
8.2
7.8
7.4
7.0
6.7
6.4
6.2
6.0
5.8
5.5
5.3
5.2
5.0
4.8
4.7
4.6
4.5
4.4
4.2
4.1
4.0
3.9
3.8
16
15
14
14
13
12
12
11
11
10
10
9.6
9.3
8.9
8.6
8.4
8.1
7.8
7.6
7.5
7.3
7.1
6.9
6.7
6.5
6.4
TAMS/MCA
-------
TABLE 5-11: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
1.6
1.4
1.3
1.3
1.2
1.1
1.0
1.0
1.0
1.0
0.9
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
2.2
2.0
1.8
.8
.8
.6
.5
.4
.4
.4
1.3
1.2
1.2
1.2
1.1
1.1
1.0
1.0
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.8
3.7
3.3
3.0
3.1
2.9
2.7
2.5
2.4
2.4
2.3
2.2
2.1
2.0
2.0
.9
.8
.8
.8
.7
.7
.7
.6
.6
.5
.4
.4
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
1.2
1.1
1.0
1.0
1.0
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
1.7
1.6
1.5
1.4
1.4
1.3
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
2.9
2.7
2.5
2.4
2.3
2.2
2.0
2.0
1.9
1.9
1.8
1.7
1.6
1.6
1.6
1.5
1.5
1.4
1.4
1.4
1.4
1.3
1.3
1.2
1.2
1.2
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
1.0
0.9
0.8
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
1.4
1.3
1.2
1.1
1.1
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
2.3
2.2
2.0
1.9
1.8
1.8
1.7
1.6
.5
.5
.4
.4
.3
.3
.2
.2
.1
1.1
1.1
.1
.0
.0
.0
.0
0.9
0.9
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
1.0
1.0
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
1.7
1.6
1.5
1.5
1.4
1.3
1.3
1.2
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
TAMS/MCA
-------
TABLE 5-12: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.04
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.0!
0.01
0.01
0.01
0.01
0.05
0.05
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
95th
Percent! le
(mg/kg wet
weight)
0.09
0.08
0.07
0.08
0.07
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0,04
0.04
0.04
0.04
0.03
25th (mg/kg
wet weight)
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
River Mile 113
95th
Median Percentile
(mg/kg wet (mg/kg wet
weight) weight)
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.07
0.07
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.009
0.01
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.02
0.06
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.03
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.02
0.02
0.01
0.01
0.01
0.01
0.01 '
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.007
0.007
0.006
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-13: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.008
0.008
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.007
0.007
0.006
0.006
0.006
0.006
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
95th
Percentile
(mg/kg wet
weight)
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
25th (mg/kg
wet weight)
0.01
0.01
0.01
0.01
0.01
0.009
0.008
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.006
0.006
0.006
0.006
0.006
0.006
0.006
0.005
0.005
0.005
0.005
0.005
River Mile 113
95th
Median Percentile
(mg/kg wet (mg/kg wet
weight) weight)
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.007
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
River Mile 90
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
0.01
0.01
0.01
0.008
0.008
0.007
0.007
0.006
0.006
0.006
0.006
0.006
0.005
0.005
0.005
0.005
0.005
0.005
0.005
0.004
0.004
0.004
0.004
0.004
0.004
0.005
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.007
0.007
0.006
0.006
0.006
0.006
0.007
95th
Percentile
(mg/kg wet
weight)
0.025
0.024
0.022
0.021
0.020
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.007
0.007
0.007
0.006
0.006
0.006
0.005
0.005
0.005
0.005
0.004
0.004
0.004
0.004
0.004
0.004
0.004
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.01
0.01
0.01
0.009
0.009
0.008
0.008
0.008
0.007
0.007
0.007
0.006
0.006
0.006
0.006
0.006
0.005
0.005
0.005
0.005
0.005
0.005
0.005
0.005
0.004
0.004
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.008
0.007
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-14: RATIO OF PREDICTED WHITE PERCH CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Pcrcentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
0.9
0.7
0.7
0.7
0.7
0.6
0.6
0.5
0.6
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.9
0.8
0.7
0.8
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
1.1
0.9
0.9
0.9
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-15: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
5.3
4.4
4.2
4.4
4.1
3.6
3.3
3.1
3.2
3.1
2.9
2.6
2.5
2.6
2.5
2.4
2.2
2.2
2.3
2.2
2.2
2.1
2.0
1.9
1.8
1.7
6.2
5.3
5.0
5.2
4.9
4.4
3.9
3.7
3.9
3.7
3.4
3.2
3.0
3.1
2.9
2.9
2.7
2.6
2.7
2.6
2.6
2.5
2.4
2.3
2.2
2.1
8.0
6.9
6.5
6.6
6.3
5.8
5.2
4.9
5.1
4.9
4.5
4.2
4.0
4.1
3.9
3.8
3.5
3.5
3.6
3.5
3.4
3.3
3.2
3.0
2.9
2.8
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
4.0
3.6
3.4
3.3
3.1
2.9
2.7
2.5
2.5
2.4
2.3
2.2
2.1
2.0
2.0
1.9
1.8
1.8
1.8
1.7
1.7
1.7
1.6
1.5
1.5
1.4
4.7
4.3
4.0
3.8
3.7
3.5
3.2
3.0
3.0
2.9
2.8
2.6
2.5
2.4
2.4
2.3
2.2
2.1
2.1
2.1
2.1
2.0
2.0
1.9
1.8
1.7
6.1
5.6
5.3
5.0
4.9
4.6
4.2
4.0
4.0
3.9
3.7
3.5
3.3
3.3
3.2
3.1
2.9
2.9
2.9
2.8
2.8
2.7
2.6
2.5
2.4
2.3
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
3.2
3.0
2.8
2.6
2.5
2.3
2.2
2.1
2.0
1.9
1.9
1.8
1.7
.6
.5
.5
.4
.4
.4
.3
1.3
1.3
1.3
1.2
1.2
1.1
3.7
3.5
3.3
3.1
2.9
2.8
2.6
2.5
2.4
2.3
2.2
2.1
2.0
1.9
.9
.8
.8
.7
.7
.6
1.6
1.6
1.5
1.5
1.4
1.4
4.9
4.6
4.3
4.0
3.9
3.7
3.4
3.3
3.1
3.1
3.0
2.8
2.7
2.6
2.5
2.4
2.3
2.3
2.2
2.2
2.1
2.1
2.0
2.0
1.9
1.8
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
2.5
2.4
2.2
2.1
1.9
1.8
1.7
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
3.0
2.8
2.6
2.4
2.3
2.2
2.0
1.9
1.8
1.8
1.7
1.6
1.6
.5
.5
.4
.4
.3
.3
1.2
1.2
. 1.2
1.2
1.1
1.1
1.1
3.8
3.5
3.3
3.1
3.0
2.8
2.7
2.5
2.4
2.3
2.2
2.2
2.1
2.0
1.9
1.9
1.8
1.7
.7
.6
.6
.6
.5
.5
.4
.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-16: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL FOR TRI+ PCBS
River Mile 152
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
0.6
0.5
0.4
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.7
0.6
0.5
0.6
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.9
0.7
0.7
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
O.I
O.I
O.I
O.I
O.I
O.I
O.I
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
O.I
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.
0.1
0.1
0.1
0.
0.
0.
0.
0.
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-17: RATIO OF PREDICTED WHITE PERCH CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg we
weight)
1.5
1.3
1.2
1.3
1.2
1.0
1.0
0.9
1.0
0.9
0.8
0.8
0.7
0.8
0.7
0.7
0.6
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
River Mile 152
Median
t (mg/kg wet
weight)
2.0
1.7
1.6
1.7
1.6
1.4
1.3
.2
.3
.2
.1
.0
.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.7
0.7
0.7
95th
Percentile
(mg/kg wet
weight)
3.7
3.1
3.0
3.0
2.9
2.6
2.3
2.3
2.3
2.2
2.0
1.9
.8
.9
.8
.7
.6
1.6
1.6
1.6
1.6
1.5
1.4
1.3
1.3
1.3
25th (mg/kg
wet weight)
1.2
1.1
1.0
1.0
0.9
0.9
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
River Mile 1
Median
(mg/kg we
weight)
.5
.4
.3
.2
.2
.1
.0
.0
.0
1.0
0.9
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
13
95th
Percentile
t (mg/kg wet
weight)
2.8
2.5
2.4
2.3
2.2
2.0
1.9
1.8
1.8
1.7
1.6
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.3
1.3
1.2
1.2
1.2
1.1
1.1
1.1
25th
(mg/kg wet
weight)
0.9
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
River Mile
Median
(mg/kg w
weight)
1.2
1.1
1.1
1.0
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
90
95th
Percentile
;t (mg/kg wet
weight)
2.2
2.1
1.9
1.8
1.7
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.8
25th
(mg/kg we
weight)
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
River Mile 5(
Median
t (mg/kg wet
weight)
1.0
0.9
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
)
95th
Percentile
(mg/kg wet
weight)
1.7
1.6
1.5
1.4
1.3
1.3
1.2
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.6
0.6
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-18: RATIO OF PREDICTED WHITE PERCH CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg wet
weight)
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
River Mile 152
Median
(mg/kg wet
weight)
1.0
0.8
0.8
0.8
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
95th
Percentile
(mg/kg wet
weight)
1.8
1.5
1.4
1.5
1.4
1.2
1.1
1.1
1.1
1.1
1.0
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.6
0.6
25th (mg/kg
wet weight)
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
River Mile 1
Median
(mg/kg we
weight)
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0,3
0.3
0.3
0.3
0.3
13
95th
Percentile
t (mg/kg wet
weight)
1.3
1.2
1.1
1.1
1.1
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
25th
(mg/kg we
weight)
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
River Mile
Median
t (mg/kg w
weight)
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
90
95th
Percentile
:t (mg/kg wet
weight)
1.1
1.0
0.9
0.9
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
25th
(mg/kg we
weight)
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.
0.
0.
0.
0.
0.
River Mile
Median
t (mg/kg w
weight)
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
50
95th
Percentile
2t (mg/kg wet
weight)
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-19: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg wet
weight)
1.6
1.3
1.2
1.3
1.2
1.1
1.0
0.9
0.9
0.9
0.8
0.8
0.7
0.8
0.7
0.7
0.6
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
River Mile 152
Median
(mg/kg wet
weight)
1.8
1.6
1.5
1.6
1.4
1.3
1.2
1.1
1.2
1.1
1.0
1.0
0.9
1.0
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.6
95th
Percentile
(mg/kg wet
weight)
3.4
2.8
2.7
2.7
2.6
2.3
2.1
2.0
2.0
2.0
1.9
1.7
.6
.7
.6
.5
.4
.4
.5
.4
.4
.4
.3
1.2
1.2
1.1
25th (mg/kg
wet weight)
1.2
1.1
1.0
0.9
0.9
0.9
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
River Mile 1
Median
(mg/kg we
weight)
1.4
1.3
1.2
1.2
1.1
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
13
95th
Percentile
t (mg/kg wet
weight)
2.5
2.3
2.2
2.1
2.0
1.9
1.7
1.6
1.6
1.6
1.5
1.4
1.3
1.3
1.3
1.3
1.2
1.2
1.2
1.1
1.1
1.1
1.1
1.0
1.0
0.9
25th
(mg/kg we
weight)
0.9
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
River Mile
Median
t (mg/kg w
weight)
1.1
1.1
1.0
0.9
0.9
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
90
95th
Percentile
;t (mg/kg wet
weight)
2.0
1.9
1.8
1.7
1.6
1.5
1.4
1.3
1.3
1.2
1.2
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.7
25th
(mg/kg we
weight)
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
River Mile
Median
t (mg/kg w
weight)
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
50
95th
Percentile
2t (mg/kg wet
weight)
1.6
1.5
1.4
1.3
1.2
1.2
1.1
1.0
1.0
1.0
0.9
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-20: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg we
weight)
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.4
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
River Mile 152
Median
t (mg/kg wet
weight)
0.9
0.8
0.7
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
95th
Percentile
(mg/kg wet
weight)
.6
.4
.3
.3
.3
.1
.0
.0
.0
.0
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
25th (mg/kg
wet weight)
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
River Mile
Median
(mg/kg w
weight)
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
113
95th
Percentile
et (mg/kg wet
weight)
1.2
1.1
1.1
1.0
1.0
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
25th
(mg/kg we
weight)
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
River Mile
Median
,t (mg/kg w
weight)
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
90
95th
Percentile
st (mg/kg wet
weight)
1.0
0.9
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
25th
(mg/kg we
weight)
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.
0.
0.
0.
0.
0.
0.1
River Mile
Median
,t (mg/kg w
weight)
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
50
95th
Percentile
et (mg/kg wet
weight)
0.8
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-21: RATIO OF PREDICTED LARGEMOUTH BASS CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
25th
(mg/kg w
Year weight)
1993 23
1994 16
1995 14
1996 16
1997 15
1998 12
1999 10
2000 10
2001 11
2002 10
2003 8.7
2004 7.2
2005 6.7
2006 7.7
2007 7.0
2008 6.6
2009 5.6
2010 6.0
2011 6.6
2012 6.0
2013 6.4
2014 5.9
2015 5.4
2016 5.1
2017 4.5
2018 4.3
River Mile
Median
et (mg/kg w
weight)
29
21
18
21
19
15
13
12
14
13
11
9.1
8.7
9.8
9.0
8.4
7.3
7.7
8.6
7.7
8.4
7.6
7.0
6.4
5.8
5.6
152
95th
Percentile
et (mg/kg wet
weight)
43
31
27
32-
29
23
20
19
21
19
17
14
13
15
14
13
11
12
13
12
13
12
11
10
8.9
8.6
25th
(mg/kg w
weight)
15
13
12
11
11
9
8
7.1
7.3
7.2
6.6
5.9
5.4
5.3
5.2
5.1
4.6
4.4
4.6
4.5
4.7
4.4
4.2
4.0
3.6
3.4
River Mile 1
Median
et (mg/kg we
weight)
19
17
15
14
13
12
10
9.3
9.4
9.3
8.5
7.6
7.0
6.9
6.7
6.5
5.9
5.7
6.0
5.9
6.1
5.7
5.5
5.1
4.7
4.5
13
95th
Percentile
t (mg/kg wet
weight)
29
25
22
21
20
18
15
14
14
14
13
12
11
10
10
9.9
9.1
8.6
9.1
9.0
9.2
8.8
8.3
7.8
7.2
6.8
25th
(mg/kg w<
weight)
3.7
3.4
3.0
2.7
2.6
2.4
2.1
1.9
1.8
1.8
1.7
.6
.4
.3
.3
.3
.2
.1
.1
.1
.1
.1
.0
.0
0.9
0.9
River Mile
Median
it (mg/kg w
weight)
4.5
4.1
3.7
3.3
3.1
2.9
2.6
2.3
2.2
2.2
2.1
.9
.8
.7
.6
.5
.5
.4
.4
1.4
1.4
.3
.3
.2
.2
.1
90
95th
Percentile
et (mg/kg wet
weight)
6.1
5.6
5.1
4.6
4.3
4.0
3.6
3.3
3.1
3.0
2.9
2.7
2.5
2.3
2.3
2.2
2.1
2.0
1.9
1.9
2.0
1.9
1.8
1.7
1.6
1.6
25th
(mg/kg w(
weight)
3.5
3.1
2.8
2.6
2.3
2.2
2.0
1.8
1.7
1.6
1.5
1.4
1.3
1.2
1.2
1.1
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.8
0.8
River Mile
Median
;t (mg/kg w
weight)
4.2
3.8
3.4
3.1
2.8
2.6
2.4
2.2
2.0
1.9
1.8
.7
.6
.5
.4
.4
.3
.2
.2
.2
.2
.1
.1
.1
.0
.0
50
95th
Percentile
st (mg/kg wet
weight)
5.7
5.1
4.6
4.2
3.8
3.6
3.2
3.0
2.8
2.6
2.5
2.4
2.2
2.1
2.0
1.9
1.8
1.7
1.7
1.6
1.6
1.6
1.5
1.5
1.4
1.4
TAMS/MCA
-------
TABLE 5-22: RATIO OF PREDICTED LARGEMOUTH BASS CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg
wet
weight)
3.1
2.3
2.0
2.3
2.1
1.6
1.4
1.4
1.5
1.4
1.2
1.0
1.0
1.1
1.0
0.9
0.8
0.9
0.9
0.9
0.9
0.8
0.8
0.7
0.6
0.6
River Mile 152
Median
(mg/kg wet
weight)
3.9
3.0
2.5
3.0
2.7
2.1
1.8
1.8
2.0
1.7
1.6
1.3
1.3
1.4
1.3
1.2
.0
.1
.2
.1
.2
.1
i ft
A* V
0.9
0.8
0.8
95th
Percentile
(mg/kg wet
weight)
6.0
4.4
3.8
4.5
4.0
3.2
2.7
2.7
2.9
2.6
2.4
2.0
1.9
2.2
.9
.8
.6
.7
.8
.7
1.8
1.6
1.5
1.4
1.3
1.2
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
2.1
.9
.6
.5
.5
.3
.1
.0
1.0
1.0
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.6
0.7
0.6
0.7
0.6
0.6
0.6
0.5
0.5
2.7
2.4
2.1
2.0
1.9
1.7
1.4
1.3
1.3
1.3
1.2
1.1
1.0
1.0
1.0
0.9
0.8
0.8
0.9
0.8
0.9
0.8
0.8
0.7
0.7
0.6
4.1
3.6
3.2
3.0
2.9
2.5
2.2
2.0
2.0
2.0
.9
.7
.5
.5
.5
.4
.3
.2
1.3
1.3
1.3
1.2
1.2
1.1
1.0
1.0
River Mile 90
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
1.6
1.5
1.3
1.2
1.1
1.1
0.9
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.5
2.1
1.9
1.7
1.6
1.5
1.3
1.2
1.1
1.0
1.0
1.0
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.7
0.6
0.6
0.6
0.5
0.6
95th
Percentile
(mg/kg wet
weight)
3.1
2.9
2.6
2.4
2.2
2.0
1.8
1.7
1.6
1.5
1.5
1.4
1.3
1.2
1.1
1.1
1.0
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.8
1.0
River Mile 50
25th Median
(mg/kg wet (mg/kg wet
weight) weight)
1.6
1.4
1.3
1.1
1.0
1.0
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4 .
0.4
2.0
1.8
1.6
1.4
1.3
1.2
1.1
1.0
0.9
0.9
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
95th
Percentile
(mg/kg wet
weight)
3.0
2.7
2.4
2.2
2.0
1.8
1.7
1.5
1.4
1.4
1.3
1.2
1.1
1.1
1.0
1.0
0.9 '
0.9
0.9
0.8
0.9
0.8
0.8
0.8
0.7
0.7
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-23: RATIO OF PREDICTED LARGEMOUTH BASS CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
25th
(mg/kg wet
weight)
1.5
1.1
1.0
1.1
1.0
0.8
0.7
0.7
0.7
0.7
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
River Mile 152
Median
(mg/kg wet
weight)
1.9
1.4
1.2
1.5
1.3
1.0
0.9
0.9
1.0
0.8
0.8
0.6
0.6
0.7
0.6
0.6
0.5
0.6
0.6
0.5
0.6
0.5
0.5
0.4
0.4
0.4
95th
Percentile
(mg/kg wet
weight)
2.9
2.1
1.9
2.2
1.9
1.5
1.3
1.3
1.4
1.3
1.2
1.0
0.9
1.0
0.9
0.9
0.8
0.8
0.9
0.8
0.9
0.8
0.7
0.7
0.6
0.6
25th (mg/kg
wet weight)
1.0
0.9
0.8
0.7
0.7
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
River Mile 113
95th
Median Percentile
(mg/kg wet (mg/kg wet
weight) weight)
1.3
1.1
1.0
1.0
0.9
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
2.0
1.7
1.5
1.4
1.4
1.2
1.1
1.0
1.0
1.0
0.9
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.8
0.7
0.7
0.6
0.6
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
1.0
0.9
0.8
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
1.5
1.4
1.2
1.1
1.1
1.0
0.9
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.5
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.8
0.7
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
1.0
0.9
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
1.4
1.3
1.2
1.0
1.0
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-24: RATIO OF PREDICTED STRIPED BASS CONCENTRATIONS TO
TRI+ AND TEQ PCB-BASED TRVs
River Mile
Tri+-based
Field-derived TRY
25th
(mg/kg
wet
Year weight)
1993 92,
1994 6.6
1995 5.8
1996 6.8
1997 62
1998 5.0
1999 4.1
2000 3.9
2001 4.4
2002 42
2003 3.6
2004 2.9
2005 2.7
2006 3.1
2007 2.9
2008 2.7
2009 23
2010 2.4
201 1 2.7
2012 2.4
2013 2.6
2014 2.4
2015 2.2
2016 2.1
2017 1.9
2018 1.8
NOAEL
95th
Median Percentile
(mg/kg (mg/kg
wet wet
weight) weight)
12 18
8.5 13
73 11
8.7 13
7.9 12
62 9.5
5.3 8.0
5.0 7.6
5.7 8.5
5.2 7.9
4.6 7.0
3.7 5.7
3.6 5.4
4.0 6.1
3.7 5.6
3.4 52
3.0 4.6
3.1 4.8
33 5.3
3.1 4.8
3.4 52
3.1 4.7
2.9 4.4
2.6 4.1
2.4 3.6
2.3 33
25th
(mg/kg
wet
weight)
3.7
2.7
23
2.7
2.5
2.0
1.7
1.6
1.8
1.7
1.4
1.2
1.1
U
12
1.1
0.9
1.0
1.1
1.0
1.1
1.0
0.9
0.8
0.8
0.7
LOAEL
Median
(mg/kg
wet
weight)
4.7
3.4
2.9
3.5
3.2
23
22
2.0
23
2.1
1.9
1.5
1.4
1.6
1.5
1.4
1.2
1-3
1.4
13
1.4
13
1.2
1.1
1.0
0.9
152
TEQ-based
Laboratory-derived TRY
95th
Percentile 25th
(mg/kg (mg/kg
wet wet
weight) weight)
7.1 7.7
5.1 5.5
4.5 4.8
5.2 5.6
4.8 5.2
3.8 4.1
3.2 3.5
3.1 33
3.4 3.6
3.2 3.5
2.8 3.0
2.3 2.5
2.2 23
2.5 2.6
22 2.4
2.1 2.3
1.8 1.9
1.9 2.0
2.1 2.2
1.9 2.0
2.1 2.2
1.9 2.0
1.8 1.8
1.6 1.8
1.5 1.6
1.4 1.5
NOAEL
Median
(mg/kg
wet
weight)
9.8
7.1
6.1
72
6.6
5.2
4.5
42
4.8
4.4
3.9
3.1
3.0
3.3
3.1
2.9
2.5
2.6
2.9
2.6
2.9
2.6
2.4
2.2
2.0
1.9
95th
Percentile
(mg/kg
wet
weight)
15
11
92
11
9.9
7.9
6.7
63
7.1
6.6
5.8
4.8
4.5
5.1
4.6
4.4
3.8
4.0
4.4
4.0
43
4.0
3.7
3.4
3.0
2.9
Tri+-based
River Mile
Field-derived TRY
25th
(mg/kg
wet
weight)
13
1.1
1.0
0.9
0.9
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
NOAEL
95th
Median Percentile
(mg/kg (mg/kg
wet wet
weight) weight)
1.6 2.4
1.4 2.1
13 1.9
12 1.7
1.1 1.7
1.0 1.5
0.9 13
0.8 12
0.8 \2
0.8 12
0.7 1.1
0.6 1.0
0.6 0.9
0.6 0.9
0.6 0.9
0.5 0.8
0.5 0.8
0.5 0.7
0.5 0.8
0.5 0.8
0.5 0.8
0.5 0.7
0.5 0.7
0.4 0.7
0.4 0.6
0.4 0.6
LOAEL
25th Median
(mg/kg (mg/kg
wet wet
weight) weight)
0.5 0.6
0.4 0.6
0.4 0.5
0.4 0.5
0.4 0.5
0.3 0.4
0.3 0.3
0.2 0.3
0.2 0.3
0.2 0.3
0.2 0.3
0.2 0.3
0.2 0.2
0.2 0.2
0.2 0.2
0.2 0.2
0.2 0.2
0.1 0.2
0.2 0.2
0.2 0.2
0.2 0.2
0.1 0.2
0.1 0.2
0.1 0.2
0.1 0.2
0.1 0.2
113
TEQ-based
Laboratory-derived TRY
95th
Percentile 25th
(mg/kg (mg/kg
wet wet
weight) weight)
1.0 1.0
0.9 0.9
0.8 0.8
0.7 0.8
0.7 0.7
0.6 0.7
0.5 0.6
0.5 0.5
0.5 0.5
0.5 0.5
0.4 0.5
0.4 0.4
0.4 0.4
0.4 0.4
0.3 0.4
0.3 0.4
0.3 0.3
0.3 0.3
0.3 0.3
0.3 0.3
0.3 0.3
0.3 0.3
0.3 0.3
0.3 0.3
0.2 0.3
0.2 0.2
NOAEL
95th
Median Percentile
(mg/kg (mg/kg
wet wet
weight) weight)
13 2.0
12 IX
1.0 1.6
1.0 1.5
0.9 1.4
0.9 13
0.7 1.1
0.6 1.0
0.7 1.0
0.7 1.0
0.6 0.9
0.5 0.8
0.5 0.7
0.5 0.7
0.5 0.7
0.5 0.7
0.4 0.6
0.4 0.6
0.4 0.6
0.4 0.6
0.4 0.6
0.4 0.6
0.4 0.6
0.4 0.5
0.3 0.5
0.3 0.5
Note a Tri+ LOAEL was not determined
Bold values indicate exceedances
-------
TABLE 5-25: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR FEMALE
TREE SWALLOWS BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
152
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
152
Average
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.05
0.04
0.04
NOAEL
152
95% UCL
0.1
0.09
0.09
0.09
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.05
LOAEL
113
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
113
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
113
Average
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
NOAEL
113
95% UCL
0.08
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
LOAEL
90
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
90
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
90
Average
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
NOAEL
90
95% UCL
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
LOAEL
50
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
50
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
50
Average
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
NOAEL
50
95% UCL
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
Bold value indicates exceedances
TAMS/MCA
-------
TABLE 5-26 : RATIO OF MODELED EGG CONCENTRATIONS TO BENCHMARKS FOR FEMALE
TREE SWALLOWS BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
152
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
152
Average
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
NOAEL
152
95% UCL
0.1
0.1
0.
0.
0.
0.
0.
0.
0.
0.
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.07
LOAEL
113
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
113
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
113
Average
0.1
0.1
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
NOAEL
113
95% UCL
0.1
0.1
0.10
0.10
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
LOAEL
90
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
90
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
90
Average
0.08
0.08
0.08
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
NOAEL
90
95% UCL
0.09
0.09
0.08
0.08
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
LOAEL
50
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
50
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
50
Average
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
NOAEL
50
95% UCL
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
Bold value indicates exceedances
TAMS/MCA
-------
TABLE 5-27: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE TREE SWALLOW USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
152
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
152
Average
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
NOAEL
152
95% UCL
0.05
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
LOAEL
113
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
113
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
113
Average
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
NOAEL
113
95% UCL
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
LOAEL
90
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
90
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
90
Average
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
NOAEL
90
95% UCL
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
LOAEL
50
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
50
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
50
Average
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
NOAEL
50
95% UCL
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
TAMS/MCA
-------
TABLE 5-28: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE TREE SWALLOW USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
152
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
152
Average
0.1
0.
0.
0.
0.
0.
0.
0.
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
NOAEL
152
95% UCL
0.1
0.
0.
0.
0.
0.
0.
0.
0.
0.1
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.07
LOAEL
113
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
113
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
113
Average
0.1
0.1
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
NOAEL
. 113
95% UCL
0.1
0.1
0.1
0.1
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
LOAEL
90
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
90
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
90
Average
0.08
0.08
0.08
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
NOAEL
90
95% UCL
0.09
0.09
0.08
0.08
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
LOAEL
50
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
50
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
50
Average
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
0.03
NOAEL
50
95% UCL
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04-
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
TAMS/MCA
-------
TABLE 5-29: RATIO OF MODELED DIETARY DOSE FOR FEMALE MALLARD BASED ON
FISHRAND RESULTS FOR THE TRI+ CONGENERS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.09
0.09
0.08
0.07
0.08
0.08
0.08
0.08
0.07
0.07
0.06
0.06
0.07
LOAEL
152
95% UCL
0.2
0.2
0.2
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.08
0.09
0.08
0.08
0.09
0.08
0.08
0.07
0.07
0.07
NOAEL
152
Average
2.2
1.9
1.6
2.0
1.7
1.4
1.2
1.3
1.4
1.2
1.1
0.9
0.9
0.9
0.9
0.8
0.7
0.8
0.8
0.8
0.8
0.7
0.7
0.6
0.6
0.7
NOAEL
152
95% UCL
2.3
2.1
1.7
2.1
1.8
1.5
1.3
1.4
1.5
1.3
1.1
1.0
1.0
1.0
0.9
0.9
0.8
0.9
0.8
0.8
0.9
0.8
0.8
0.7
0.7
0.7
LOAEL
113
Average
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.09
0.08
0.07
0.07
0.07
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.05
0.05
0.05
LOAEL
113
95% UCL
0.2
0.2
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.09
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.05
0.06
NOAEL
113
Average
1.7
1.6
1.3
.4
.3
.1
.0
.0
.0
0.9
0.9
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
NOAEL
113
95% UCL
1.9
1.7
1.4
1.5
1.4
1.2
1.1
1.0
1.1
1.0
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.5
0.6
LOAEL
90
Average
0.1
0.1
0.1
0.1
0.1
0.09
0.08
0.08
0.08
0.07
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
LOAEL
90
95% UCL
0.2
0.
0.
0.
0.
0.
0.09
0.08
0.08
0.08
0.08
0.07
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.1
0.0
0.0
0.0
0.0
0.0
NOAEL
90
Average
1.4
1.3
1.1
1.1
1.0
0.9
0.8
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
NOAEL
90
95% UCL
1.5
1.4
1.2
1.2
1.1
1.0
0.9
0.8
0.8
0.8
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
LOAEL
50
Average
0.1
0.1
0.1
0.09
0.09
0.08
0.07
0.07
0.07
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.03
LOAEL
50
95% UCL
0.1
0.1
0.1
0.10
0.09
0.09
0.08
0.07
0.07
0.07
0.06
0.06
0.06
0.05
0.05
0.05
0.05
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
NOAEL
50
Average
1.3
1.1
1.0
0.9
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
NOAEL
50
95% UCL
1.4
1.2
1.1
1.0
0.9
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0:4
0.4
0.4
0.4
0.4
0.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-30: RATIO OF EGG CONCENTRATIONS FOR FEMALE MALLARD BASED ON
FISHRAND RESULTS FOR THE TRI+ CONGENERS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
2.4
2.1
2.1
2.0
1.9
1.8
1.8
1.8
1.7
1.6
1.5
1.5
1.5
1.4
1.4
1.4
1.3
1.3
1.3
1.2
1.2
1.2
I.I
1.2
1.1
1.1
LOAEL
152
95% UCL
2.6
2.3
2.2
2.2
2.1
2.0
1.9
1.9
1.8
1.8
1.7
1.6
1.6
1.5
1.5
1.5
.4
.4
.3
.3
.3
1.3
1.2
1.3
1.2
1.2
NOAEL
152
Average
15.9
14.3
13.8
13.7
12.9
12.4
11.7
11.8
11.5
11.0
10.4
10.2
9.9
9.5
9.4
9.1
8.9
8.7
8.4
8.2
8.0
7.9
7.7
7.8
7.7
7.5
NOAEL
152
95% UCL
17.1
15.3
14.8
14.6
13.9
13.3
12.6
12.7
12.4
11.8
11.1
11.0
10.7
10.2
10.1
9.8
9.6
9.4
9.0
8.8
8.6
8.5
8.3
8.4
8.3
8.1
LOAEL
113
Average
1.9
1.8
1.7
1.6
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.2
1.2
1.2
1.2
1.1
1.1
1.1
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
LOAEL
113
95% UCL
2.0
1.9
1.8
1.7
1.7
1.6
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.3
1.3
1.2
1.2
1.1
1.1
1.1
1.1
1.1
1.0
1.0
1.0
1.0
NOAEL
113
Average
12.7
11.9
11.4
10.9
10.5
10.2
9.9
9.5
9.3
9.0
8.6
8.3
8.2
8.0
7.8
7.5
7.3
7.1
7.1
6.9
6.8
6.6
6,4
6.2
6.1
6.1
NOAEL
113
95% UCL
13.6
12.7
12.2
11.7
11.2
10.9
10.6
10.2
10.0
9.7
9.2
8.9
8.8
8.6
8.4
8.1
7.9
7.7
7.6
7.5
7.3
7.1
6=9
6.7
6.6
6.6
LOAEL
90
Average
1.5
1.5
1.4
1.3
1.3
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
LOAEL
90
95% UCL
1.6
1.6
1.5
1.4
1.3
1.3
1.2
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.7
NOAEL
90
Average
10.3
9.8
9.1
8.7
8.4
8.0
7.7
7.5
7.3
7.1
7.0
6.6
6.4
6.1
5.9
5.8
5.6
5.5
5.3
5.2
5.1
4.9
4.8
4.7
4.7
4.6
NOAEL
90
95% UCL
11.0
10.5
9.8
9.3
9.0
8.6
8.3
8.1
7.8
7.6
7.5
7.1
6.8
6.5
6.4
6.2
6.0
5.9
5.7
5.6
5.4
5.3
5.2
5.1
5.0
4.9
LOAEL
50
Average
1.1
1.1
1.0
1.0
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
LOAEL
50
95% UCL
1.2
1.1
1.1
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
NOAEL
50
Average
7.6
7.1
6.7
6.5
6.3
5.9
5.8
5.6
5.4
5.3
5.1
4.9
4.7
4.5
4.4
4.3
4.2
4.1
4.0
3.9
3.8
3.7
3.6
3.6
3.5
3.4
NOAEL
50
95% UCL
8.1
7.6
7.2
6.9
6.7
6.4
6.2
6.0
5.8
5.7
5.5
5.3
5.1
4.8
4.7
4.6
4.5
4.4
4.3
4.2
4.1
4.0
3.9
3.8
3.7
3.7
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-31: RATIO OF MODELED DIETARY DOSE TO BENCHMARKS
FOR FEMALE MALLARD FOR PERIOD 1993 - 2018 ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
16
14
11
15
12
8.9
7.5
8.5
9.4
7.7
6.5
5.3
5.2
5.5
4.9
4.6
3.7
4.7
4.2
4.5
5.3
4.4
4.2
3.3
3.2
3.5
LOAEL
152
95% UCL
17.4
15.0
11.8
16.3
13.0
9.6
8.0
9.1
10.1
8.2
7.0
5.7
5.6
6.0
5.3
4.9
4.0
5.1
4.5
4.8
5.7
4.7
4.5
3.6
3.4
3.8
NOAEL
152
Average
162
140
110
152
121
89
75
85
94
77
65
53
52
55
49
46
37
47
42
45
53
44
42
33
32
35
NOAEL
152
95% UCL
174
150
118
163
130
96
80
91
101
82
70
57
56
60
53
49
40
51
45
48
57
47
45
36
34
38
LOAEL
113
Average
13
12
9
10
8.9
7.1
6.0
6.0
6.3
5.7
5.3
4.3
4.1
4.2
4.0
3.6
3.2
3.5
3.5
3.6
3.7
3.4
3.3
2.7
2.5
2.6
LOAEL
113
95% UCL
13.9
12.4
9.7
10.5
9.6
7.6
6.5
6.4
6.7
6.2
5.7
4.6
4.4
4.5
4.3
3.9
3.5
3.8
3.8
3.8
4.0
3.7
3.5
2.9
2.7
2.8
NOAEL
113
Average
130
116
91
98
89
71
60
60
63
57
53
43
41
42
40
36
32
35
35
36
37
34
33
27
25
26
NOAEL
113
95% UCL
139
124
97
105
96
76
65
64
67
62
57
46
44
45
43
39
35
38
38
38
40
37
35
29
27
28
LOAEL
90
Average
11
9.4
8.5
7.7
6.9
6.2
5.7
5.1
4.8
4.6
4.3
4.0
3.7
3.4
3.3
3.1
2.9
2.8
2.7
2.7
2.6
2.6
2.5
2.4
2.3
2.2
LOAEL
90
95% UCL
11
10
8.5
8.1
7.5
6.5
5.7
5.3
5.1
4.9
4.6
4.0
3.7
3.6
3.5
3.2
3.0
3.0
2.9
3.0
2.9
2.8
2.7
2.5
2.3
2.3
NOAEL
90
Average
107
94
85
77
69
62
57
51
48
46
43
40
37
34
33
31
29
28
27
27
26
26
25
24
23
22
NOAEL
90
95%UCL
112
101
85
81
75
65
57
53
51
49
46
40
37
36
35
32
30
30
29
30
29
28
27
25
23
23
LOAEL
50
Average
16
14
11
15
12
8.4
7.0
8.0
9.0
7.2
6.1
4.7
4.7
5.1
4.4
4.1
3.1
4.3
3.7
4.0
4.9
3.9
3.8
2.8
2.7
3.1
LOAEL
50
95% UCL
11
9.7
8.6
7.7
7.0
6.3
5.8
5.2
4.8
4.6
4.3
4.0
3.7
3.4
3.2
3.1
2.9
2.8
2.7
2.7
2.6
2.6
2.5
2.4
2.3
2.2
NOAEL
50
Average
159
138
105
151
118
84
70
80
90
72
61
47
47
51
44
41
31
43
37
40
49
39
38
28
27
31
NOAEL
50
95% UCL
110
97
86
77
70
63
58
52
48
46
43
40
37
34
32
31
29
28
27
27
26
26
25
24
23
22
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-32: RATIO OF MODELED EGG CONCENTRATION TO BENCHMARKS FOR
FEMALE MALLARD FOR PERIOD 1993 - 2018 ON A TEQ BASIS
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
340
305
295
291
276
264
250
252
246
235
221
218
212
203
201
194
190
187
179
174
171
169
164
165
163
160
LOAEL
152
95% UCL
366
327
317
312
296
283
269
270
264
253
238
234
228
219
216
209
205
201
192
187
183
181
177
179
177
173
NOAEL
152
Average
1362
1221
1181
1166
1104
1057
1002
1007
985
941
885
871
847
814
803
775
759
747
715
697
682
675
655
662
654
638
NOAEL
152
95% UCL
1463
1308
1266
1249
1183
1133
1076
1081
1055
1011
951
937
911
875
864
836
819
803
769
750
733
726
707
716
708
691
LOAEL
113
Average
270
253
243
233
224
217
211
202
199
192
183
177
175
170
167
160
156
153
151
148
145
141
136
132
131
130
LOAEL
113
95% UCL
290
271
260
250
240
233
226
217
214
207
197
190
189
183
179
173
168
164
163
159
156
152
146
142
141
141
NOAEL
113
Average
1081
1012
971
933
895
870
843
808
797
769
734
708
701
681
667
642
622
610
605
591
578
564
543
528
524
522
NOAEL
113
95% UCL
1160
1085
1041
1000
959
932
905
868
856
826
788
761
754
732
717
691
670
657
651
636
622
607
585
570
566
565
LOAEL
90
Average
219
208
195
186
180
171
165
161
156
152
149
141
136
130
126
123
120
118
114
111
108
105
103
101
100
98
LOAEL
90
95% UCL
236
223
209
199
193
184
177
172
167
163
160
151
146
140
136
132
129
127
122
120
116
113
111
109
107
105
NOAEL
90
Average
878
833
780
743
720
686
661
643
624
609
595
564
543
520
506
491
480
472
456
445
432
421
413
404
399
392
NOAEL
90
95%UCL
943
894
837
796
772
735
708
688
668
652
639
606
583
558
544
528
516
507
490
478
464
453
444
434
429
422
LOAEL
50
Average
161
151
144
138
134
126
123
119
115
113
109
105
101
96
93
91
89
87
86
84
81
79
78
76
74
72
LOAEL
50
95% UCL
173
162
154
148
143
136
132
128
124
122
118
112
108
103
101
98
96
94
92
90
88
86
83
82
80
78
NOAEL
50
Average
645
605
575
553
535
506
492
477
462
454
438
418
402
385
374
364
356
350
344
336
326
318
310
304
296
290
NOAEL
50
95% UCL
693
649
616
593
573
542
526
510
495
487
470
449
432
413
402
391
383
376
370
362
351
342
334
327
319
312
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-33: RATIO OF MODELED DIETARY DOSE TO BENCHMARKS BASED ON FISHRAND FOR FEMALE KINGFISHER
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
10
7.5
6.8
7.8
6.8
5.4
4.9
4.8
5.1
4.6
4.3
3.6
3.6
3.9
3.4
3.2
3.0
3.2
3.3
3.1
3.3
3.1
2.8
2.5
2.5
2.5
LOAEL
152
95% UCL
10
7.7
7.1
8.1
7.0
5.6
5.1
5.0
5.3
4.9
4.5
3.8
3.7
4.1
3.6
3.4
3.2
3.4
3.5
3.3
3.4
3.3
3.0
2.7
2.6
2.6
NOAEL
152
Average
67
52
47
54
47
38
34
34
36
32
30
25
25
28
24
23
21
22
23
22
23
22
20
18
17
17
NOAEL
152
95% UCL
69
54
49
57
49
39
36
35
37
34
31
27
26
29
25
24
22
23
24
23
24
23
21
19
18
18
LOAEL
113
Average
6.7
6.1
5.2
5.3
5.0
4.3
3.9
3.7
3.8
3.7
3.4
3.0
2.9
2.9
2.8
2.7
2.5
2.4
2.5
2.5
2.5
2.4
2.3
2.1
2.0
2.0
LOAEL
113
95% UCL
7.0
6.4
5.5
5.5
5.2
4.5
4.1
3.9
3.9
3.8
3.5
3.2
3.0
3.0
2.9
2.8
2.6
2.6
2.6
2.6
2.6
2.5
2.4
2.2
2.1
2.1
NOAEL
113
Average
47
43
37
37
35
30
28
26
26
26
24
21
20
20
20
19
17
17
18
17
17
17
16
15
14
14
NOAEL
113
95% UCL
49
45
38
39
36
32
29
27
28
27
25
22
21
21
21
20
18
18
19
18
18
18
17
16
15
15
LOAEL
90
Average
5.4
4.9
4.5
4.2
3.9
3.6
3.2
3.0
2.9
2.8
2.7
2.5
2.3
2.2
2.2
2.1
2.0
1.9
.9
.9
.8
.8
.7
.7
.6
.5
LOAEL
90
95% UCL
5.6
5.1
4.6
4.3
4.1
3.8
3.4
3.2
3.0
3.0
2.8
2.6
2.5
2.3
2.3
2.2
2.1
2.0
2.0
2.0
1.9
1.9
1.8
1.8
1.7
1.6
NOAEL
90
Average
38
34
31
30
27
25
23
21
20
20
19
18
16
16
15
15
14
13
13
13
13
12
12
12
11
11
NOAEL
90
95% UCL
39
36
32
30
28
26
24
22
21
21
20
18
17
16
16
15
15
14
14
14
14
13
13
12
12
11
LOAEL
50
Average
4.8
43
3.9
3.6
3.4
3.1
2.9
2.7
2.5
2.4
2.3
2.1
2.0
1.9
1.8
1.7
1.7
1.6
1.6
1.5
1.5
1.5
1.4
1.4
1.3
1.3
LOAEL
50
95% UCL
5.0
4.5
4.1
3.8
3.5
3.3
3.0
2.8
2.6
2.5
2.4
2.2
2.1
2.0
1.9
1.8
1.8
1.7
1.6
1.6
1.6
1.5
1.5
1.5
1.4
1.4
NOAEL
50
Average
33
30
27
25
23
22
20
19
17
17
16
15
14
13
13
12
12
11
11
11
11
10
10
10
9.4
9.1
NOAEL
50
95% UCL
35
31
28
26
24
23
21
19
18
18
17
16
15
14
13
13
12
12
12
11
11
11
11
10
10
10
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-34: RATIO OF MODELED DIETARY DOSE TO BENCHMARKS BASED ON FISHRAND FOR FEMALE BLUE HERON
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
3.7
2.8
2.5
3.0
2.5
1.9
1.6
1.6
1.8
1.6
1.4
1.1
1.1
1.3
1.1
1.0
0.9
1.0
1.1
1.0
1.1
1.0
0.9
0.7
0.7
0.7
LOAEL
152
95% UCL
3.8
2.9
2.5
3.1
2.6
1.9
1.7
1.7
1.8
1.6
1.5
1.2
1.1
1.4
1.1
1.0
0.9
1.0
1.1
1.1
1.1
1.1
0.9
0.8
0.7
0.8
NOAEL
152
Average
26
19
17
21
18
13
11
11
12
11
10
7.8
7.7
9.2
7.5
6.9
6.3
7.0
7.6
7.1
7.6
7.1
6.2
5.2
4.9
5.0
NOAEL
152
95% UCL
27
20
18
22
18
13
12
12
13
11
11
8.2
8.0
10
7.8
7.3
6.6
7.3
7.9
7.4
7.9
7.4
6.5
5.5
5.2
5.3
LOAEL
113
Average
2.5
2.3
1.8
1.9
1.8
1.5
1.3
1.2
1.2
1.2
1.1
0.9
0.9
0.9
0.9
0.8
0.7
0.7
0.8
0.8
0.8
0.7
0.7
0.6
0.6
0.6
LOAEL
113
95% UCL
2.6
2.3
1.9
2.0
1.9
1.5
1.4
1.2
1.3
1.3
1.2
1.0
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.6
0.6
NOAEL
113
Average
18
16
13
13
13
10
9.1
8.4
8.7
8.5
7.7
6.6
6.2
6.3
6.1
5.7
5.2
5.1
5.4
5.3
5.4
5.2
4.8
4.4
4.0
3.9
NOAEL
113
95% UCL
18
16
13
14
13
11
9.5
8.7
9.1
8.9
8.1
6.9
6.4
6.6
6.3
6.0
5.5
5.4
5.7
5.6
5.7
5.4
5.1
4.6
4.3
4.1
LOAEL
90
Average
2.0
1.8
2
1.5
1.4
1.3
1.1
1.0
1.0
0.9
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
LOAEL
90
95% UCL
2.1
1.9
1.6
1.5
1.4
1.3
1.1
1.0
1.0
1.0
0.9
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
NOAEL
90
Average
14
13
11
11
10
8.9
7.7
7.1
6.7
6.5
6.1
5.6
5.2
4.9
4.7
4.5
4.2
4.0
4.0
4.0
4.0
3.9
3.8
3.6
3.3
3.2
NOAEL
90
95% UCL
14
13
12
11
10
9.2
8.0
7.3
7.0
6.8
6.4
5.9
5.4
5.1
4.9
4.7
4.4
4.2
4.2
4.2
4.2
4.1
3.9
3.7
3.5
3.4
LOAEL
50
Average
1.9
1.7
1.5
1.4
1.3
1.2
1.0
1.0
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
LOAEL
50
95% UCL
2.0
1.7
1.6
1.4
13
1.2
1.1
1.0
0.9
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
NOAEL
50
Average
13
12
11
9.7
8.8
8.2
73
6.7
6.2
5.9
5.6
5.2
4.8
4.5
43
4.1
3.9
3.7
3.6
3.5
3.5
3.4
33
3.2
3.0
2.9
NOAEL
50
95% UCL
14
- 12
11
10
9.1
8.5
7.6
6.9
6.4
6.1
5.8
5.4
5.0
4.6
4.4
4.2
4.0
3.8
3.7
3.7
3.6
3.5
3.4
33
3.2
3.1
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-35: RATIO OF MODELED DIETARY DOSE TO BENCHMARKS BASED ON FISHRAND FOR FEMALE BALD EAGLE
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
27
20
17
20
18
14
12
12
13
12
11
8.7
8.3
9.3
8.6
8.0
6.9
7.3
8.1
7.3
7.9
7.2
6.7
6.1
5.5
5.4
LOAEL
152
95% UCL
28
20
17
20
19
15
13
12
13
12
11
8.9
8.4
10
8.8
8.1
7.1
7.4
8.3
7.4
8.1
7.4
6.8
6.2
5.6
5.5
NOAEL
152
Average
190
138
118
140
127
100
86
81
92
84
75
61
58
65
60
56
48
51
57
51
56
51
47
43
38
37
NOAEL
152
95% UCL
194
140
121
143
130
102
88
83
94
86
77
62
59
67
61
57
50
52
58
52
57
52
48
44
39
38
LOAEL
113
Average
18
16
14
13
13
12
10
8.8
8.9
8.8
8.1
7.2
6.6
6.5
6.4
6.1
5.6
5.4
5.7
5.6
5.8
5.5
5.2
4.8
4.5
4.2
LOAEL
113
95% UCL
19
16
14
13
13
12
10
9.0
9.1
9.0
8.3
7.3
6.8
6.7
6.5
6.3
5.7
5.5
5.8
5.7
5.9
5.6
5.3
5.0
4.6
4.3
NOAEL
113
Average
127
111
99
92
89
81
68
62
62
62
57
50
46
46
45
43
39
38
40
39
40
38
36
34
31
30
NOAEL
113
95% UCL
130
113
101
94
91
83
69
63
64
63
58
51
47
47
46
44
40
38
41
40
41
39
37
35
32
30
LOAEL
90
Average
4.2
3.8
3.4
3.1
2.9
2.7
2.4
2.2
2.1
2.0
.9
.8
.6
.6
.5
.4
1.4
1.3
1.3
1.3
1.3
1.2
1.2
1.1
1.1
1.0
LOAEL
90
95% UCL
4.2
3.8
3.5
3.2
3.0
2.7
2.5
2.2
2.1
2.1
2.0
1.8
1.7
1.6
1.5
1.5
1.4
1.3
1.3
1.3
1.3
1.3
1.2
1.2
1.1
1.1
NOAEL
90
Average
29
26
24
22
20
19
17
15
15
14
13
12
11
11
11
10
10
9.1
9.0
8.9
9.2
8.7
8.4
8.0
7.6
7.2
NOAEL
90
95% UCL
30
27
24
22
21
19
17
16
15
14
14
13
12
11
11
10
10
9.3
9.2
9.1
9.3
8.8
8.6
8.2
7.8
7.4
LOAEL
50
Average
3.9
3.5
3.2
2.9
2.6
2.4
2.2
2.0
1.9
1.8
1.7
1.6
1.5
1.4
13
13
1.2
1.2
.1
.1
.1
.1
.0
.0
1.0
0.9
LOAEL
50
95% UCL
4.0
3.6
3.2
2.9
2.7
2.5
2.2
2.1
1.9
1.8
1.7
1.6
1.5
1.4
1.4
13
1.2
1.2
1.1
1.1
1.1
1.1
1.0
1.0
1.0
0.9
NOAEL
50
Average
27
25
22
20
18
17
15
14
13
13
12
11
10
9.8
93
8.9
8.5
8.1
7.8
7.6
7.8
7.4
7.2
7.0
6.8
6.4
NOAEL
50
95% UCL
28
25
22
20
19
17
16
14
13
13
12
11
11
9.9
9.5
9.1
8.6
8.2
8.0
7.8
7.9
7.5
7.3
7.1
6.9
6.5
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-36: RATIO OF MODELED EGG CONCENTRATIONS TO BENCHMARKS FOR FEMALE KINGFISHER
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
23
18
16
19
16
13
12
11
12
11
10
8.7
8.5
9.4
8.2
7.7
7.2
7.6
7.9
7.5
7.8
7.4
6.7
6.1
5.9
5.9
LOAEL
152
95% UCL
24
19
17
19
17
13
12
12
13
12
11
9.1
8.9
10
8.6
8.1
7.6
8.0
8.3
7.8
8.1
7.8
7.1
6.4
6.2
6.2
NOAEL
152
Average
153
119
109
125
108
86
78
77
82
74
69
58
57
63
55
52
48
51
53
50
52
50
45
41
39
39
NOAEL
152
95% UCL
159
124
113
129
112
90
81
80
85
78
72
61
59
66
57
54
51
53
55
52
55
52
47
43
42
41
LOAEL
113
Average
16
15
12
13
12
10
9.4
8.8
9.0
8.7
8.1
7.2
6.9
6.9
6.7
6.3
5.9
5.8
6.0
5.9
5.9
5.7
5.4
5.0
4.8
4.7
LOAEL
113
95% UCL
17
15
13
13
12
11
10
9.2
9.4
9.1
8.4
7.6
7.2
7.2
7.0
6.7
6.2
6.1
6.3
6.2
6.2
6.0
5.6
5.3
5.0
4.9
NOAEL
113
Average
107
98
84
85
80
69
63
59
60
58
54
48
46
46
45
42
40
39
40
39
40
38
36
34
32
31
NOAEL
113
95% UCL
111
102
87
89
83
72
66
62
63
61
57
51
48
49
47
45
42
41
42
41
41
40
38
35
34
33
LOAEL
90
Average
13
12
11
10
9.3
8.6
7.8
7.3
6.9
6.8
6.5
6.0
5.6
5.3
5.1
5.0
4.7
4.5
4.5
4.5
4.4
4.3
4.1
4.0
3.8
3.7
LOAEL
90
95% UCL
13
12
11
10
10
9.0
8.1
7.6
7.2
7.1
6.7
6.3
5.9
5.6
5.4
5.2
5.0
4.7
4.7
4.7
4.6
4.5
4.3
4.2
4.0
3.9
NOAEL
90
Average
86
79
71
66
62
58
52
49
46
45
43
40
37
36
34
33
32
30
30
30
29
29
28
27
25
25
NOAEL
90
95% UCL
90
82
73
69
65
60
54
51
48
47
45
42
39
37
36
35
33
32
31
31
31
30
29
28
27
26
LOAEL
50
Average
11
10
9.3
8.6
8.0
7.5
6.8
6.3
6.0
5.7
5.5
5.1
4.8
4.5
4.3
4.2
4.0
3.8
3.7
3.7
3.6
3.5
3.4
3.3
3.2
3.1
LOAEL
50
95% UCL
12
11
10
9.0
8.3
7.8
7.1
6.6
6.2
6.0
5.7
5.3
5.0
4.7
4.5
4.3
4.2
4.0
3.9
3.9
3.8
3.7
3.6
3.5
3.3
3.3
NOAEL
50
Average
77
69
62
58
54
50
46
42
40
38
37
34
32
30
29
28
27
26
25
25
24
24
23
22
21
21
NOAEL
50
95% UCL
80
72
65
60
56
52
48
44
42
40
38
36
34
32
30
29
28
27
26
26
25
25
24
23
22
22
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-37: RATIO OF MODELED EGG CONCENTRATIONS TO BENCHMARKS FOR FEMALE BLUE HERON
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
23
17
15
18
15
11
10
10
11
9.5
8.7
6.8
6.6
8.0
6.5
6.0
5.4
6.0
6.5
6.1
6.6
6.1
5.3
4.5
4.2
4.3
LOAEL
152
95% UCL
23
17
15
19
16
12
10
10
11
10
9.0
7.0
6.8
8.2
6.7
6.2
5.6
6.2
6.7
6.3
6.8
6.3
5.5
4.6
4.4
4.5
NOAEL
152
Average
151
112
100
121
102
75
66
64
72
64
58
45
44
53
43
40
36
40
44
41
44
41
36
30
28
29
NOAEL
152
95% UCL
155
115
103
124
105
77
68
66
74
65
60
47
46
55
45
41
37
41
45
42
45
42
37
31
29
30
LOAEL
113
Average
15
14
11
12
11
8.9
7.9
7.2
7.5
7.3
6.7
5.7
5.3
5.4
5.2
4.9
4.5
4.4
4.7
4.6
4.7
4.5
4.2
3.8
3.5
3.4
LOAEL
113
95% UCL
16
14
12
12
11
9.2
8.1
7.4
7.8
7.6
6.9
5.9
5.5
5.6
5.4
5.1
4.6
4.5
4.8
4.7
4.8
4.6
4.3
3.9
3.6
3.5
NOAEL
113
Average
101
92
75
78
72
60
53
48
51
49
45
38
35
36
35
33
30
29
31
31
31
30
28
25
23
23
NOAEL
113
95% UCL
104
94
77
80
75
61
54
50
52
51
46
39
37
38
36
34
31
30
32
32
32
31
29
26
24
23
LOAEL
90
Average
12
11
10
9.0
8.3
7.7
6.6
6.1
5.8
5.6
5.3
4.8
4.4
4.2
4.0
3.9
3.6
3.4
3.5
3.5
3.5
3.3
3.2
3.1
2.9
2.8
LOAEL
90
95% UCL
12
11
10
9.2
8.6
7.9
6.8
6.3
5.9
5.8
5.4
5.0
4.6
4.3
4.2
4.0
3.8
3.5
3.6
3.6
3.6
3.4
3.3
3.2
3.0
2.8
NOAEL
90
Average
81
73
64
60
56
51
44
41
39
38
35
32
30
28
27
26
24
23
23
23
23
22
22
21
19
18
NOAEL
90
95% UCL
84
76
66
62
57
53
46
42
40
39
36
33
31
29
28
27
25
24
24
24
24
23
22
21
20
19
LOAEL
50
Average
11
10
9.1
8.3
7.6
7.1
63
5.8
53
5.1
4.8
4.5
4.1
3.9
3.7
3.5
3.3
3.2
3.1
3.1
3.0
2.9
2.8
2.7
2.6
2.5
LOAEL
50
95% UCL
12
11
9.4
8.6
7.9
73
6.5
5.9
5.5
5.2
4.9
4.6
4.3
4.0
3.8
3.6
3.4
3.2
3.2
3.1
3.1
3.0
2.9
2.8
2.7
2.6
NOAEL
50
Average
77
69
61
56
51
48
42
39
36
34
32
30
28
26
25
23
22
21
21
20
20
19
19
18
17
17
NOAEL
50
95% UCL
79
71
63
57
53
49
43
40
37
35
33
31
28
27
25
24
23
22
21
21
21
20
20
19
18
17
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-38: RATIO OF MODELED EGG CONCENTRATIONS TO BENCHMARKS FOR FEMALE BALD EAGLES
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
152
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
152
Average
139
101
86
103
93
73
63
59
68
62
55
44
42
48
44
41
35
37
42
37
41
37
34
31
28
27
NOAEL
152
95% UCL
142
103
88
105
95
75
65
61
69
63
56
45
43
49
45
42
36
38
43
38
42
38
35
32
29
28
LOAEL
113
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
113
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
113
Average
93
81
72
67
65
59
50
45
46
45
42
37
34
33
33
32
29
28
29
29
29
28
27
25
23
22
NOAEL
113
95% UCL
95
83
74
69
66
61
51
46
47
46
42
38
35
34
33
32
29
28
30
29
30
29
27
25
23
22
LOAEL
90
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
90
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
90
Average
20
18
16
15
- 14
12
11
10
10
9.2
8.8
8.2
7.6
7.2
6.8
6.5
6.2
5.9
5.7
5.6
5.7
5.4
5.3
5.1
5.0
4.7
NOAEL
90
95% UCL
20
18
16
15
14
13
11
11
10
9.4
8.9
8.3
7.7
7.3
6.9
6.6
6.3
6.0
5.8
5.7
5.8
5.5
5.4
5.2
5.0
4.8
LOAEL
50
Average
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
LOAEL
50
95% UCL
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NOAEL
50
Average
20
18
16
15
14
12
11
10
9.7
9.2
8.8
8.2
7.6
7.2
6.8
6.5
6.2
5.9
5.7
5.6
5.7
5.4
5.3
5.1
5.0
4.7
NOAEL
50
95% UCL
20
18
16
15
14
13
11
11
9.8
9.4
8.9
8.3
7.7
73
6.9
6.6
6.3
6.0
5.8
5.7
5.8
5.5
5.4
5.2
5.0
4.8
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-39: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE BELTED KINGFISHER USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
8.6
6.7
6.0
7.0
6.0
4.7
4.3
4.2
4.5
4.1
3.8
3.2
3.1
3.4
3.0
2.8
2.6
2.8
2.9
2.7
2.9
2.7
2.4
2.2
2.1
2.1
LOAEL
152
95% UCL
8.9
6.9
6.3
7.3
6.3
5.0
4.5
4.4
4.7
4.3
4.0
3.4
3.3
3.6
3.2
3.0
2.8
2.9
3.1
2.9
3.0
2.9
2.6
2.4
2.3
2.3
NOAEL
152
Average
86
67
60
70
60
47
43
42
45
41
38
32
31
34
30
28
26
28
29
27
29
27
24
22
21
21
NOAEL
152
95% UCL
89
69
63
73
63
50
45
44
47
43
40
34
33
36
32
30
28
29
31
29
30
29
26
24
23
23
LOAEL
113
Average
6.0
5.5
4.6
4.7
4.4
3.8
3.4
3.2
3.3
3.2
3.0
2.6
2.5
2.5
2.4
2.3
2.1
2.1
2.2
2.1
2.1
2.1
1.9
1.8
1.7
1.7
LOAEL
113
95% UCL
6.2
5.7
4.8
4.9
4.6
4.0
3.6
3.4
3.5
3.4
3.1
2.8
2.7
2.7
2.6
2.5
2.3
2.3
2.3
2.3
2.3
2.2
2.1
1.9
1.8
1.8
NOAEL
113
Average
60
55
46
47
44
38
34
32
33
32
30
26
25
25
24
23
21
21
22
21
21
21
19
18
17
17
NOAEL
113
95% UCL
62
57'
. 48
49
46
40
36
34
35
34
31
28
27
27
26
25
23
23
23
23
23
22
21
19
18
18
LOAEL
90
Average
4.8
4.4
4.3
4.1
3.4
3.2
2.9
2.7
2.5
2.5
2.4
2.2
2.0
1.9
1.9
1.8
1.7
1.6
1.6
1.6
1.6
1.5
1.5
1.4
1.4
1.3
LOAEL
90
95% UCL
12
11
11
10.3
9.9
9.6
9.2
8.8
8.4
8.3
8.1
8.1
7.9
7.4
7.2
7.3
7.4
6.8
6.4
6.3
6.2
6.1
6.0
6.2
6.2
5.9
NOAEL
90
Average
48
44
43
41
34
32
29
27
25
25
24
22
20
19
19
18
17
16
16
16
16
15
15
14
14
13
NOAEL
90
95% UCL
121
113
107
103
99
96
92
88
84
83
81
81
79
74
72
73
74
68
64
63
62
61
60
62
62
59
LOAEL
50
Average
4.3
3.9
3.5
3.2
3.0
2.8
2.5
2.4
2.2
2.1
2.0
1.9
1.8
1.7
1.6
1.5
1.5
1.4
1.4
1.4
13
1.3
13
1.2
1.2
1.1
LOAEL
50
95% UCL
10
10
9.0
8.6
83
8.0
7.7
7.4
7.1
6.9
6.7
6.6
6.4
6.2
6.0
5.9
5.9
5.7
5.5
53
5.2
5.0
5.0
5.0
4.9
4.9
NOAEL
50
Average
43
39
35
32
30
28
25
24
22
21
20
19
18
17
16
15
15
14
14
14
13
13
13
12
12
11
NOAEL
50
95% UCL
102
96
90
86
83
80
77
74
71
69
67
66
64
62
60
59
59
57
55
53
52
50
50
50
49
49
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-40: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE GREAT BLUE HERON USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
2.2
1.7
1.5
1.8
1.5
1.1
1.0
1.0
1.1
0.9
0.9
0.7
0.7
0.8
0.6
0.6
0.5
0.6
0.7
0.6
0.7
0.6
0.5
0.5
0.4
0.4
LOAEL
152
95% UCL
2.3
1.7
1.5
1.8
.6
.2
.0
.0
.1
.0
0.9
0.7
0.7
0.8
0.7
0.6
0.6
0.6
0.7
0.6
0.7
0.6
0.6
0.5
0.5
0.5
NOAEL
152
Average
22
17
15
18
15
11
10
10
11
9.4
8.7
6.8
6.6
7.9
6.5
6.0
5.5
6.0
6.5
6.1
6.5
6.1
5.3
4.5
4.3
4.3
NOAEL
152
95% UCL
23
17
15
18
16
12
10
10
11
10
9.2
7.3
7.1
8.4
6.9
6.5
5.9
6.5
6.9
6.5
6.9
6.5
5.7
4.9
4.7
4.7
LOAEL
113
Average
1.5
1.4
1.1
1.2
1.1
0.9
0.8
0.7
0.8
0.7
0.7
0.6
0.5
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
LOAEL
113
95% UCL
1.6
1.4
1.2
1.2
1.1
0.9
0.8
0.8
0.8
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
NOAEL
113
Average
15
14
11
12
11
9.0
8.0
7.4
7.7
7.5
6.8
5.9
5.5
5.6
5.4
5.1
4.6
4.6
4.8
4.7
4.8
4.6
4.3
3.9
3.6
3.5
NOAEL
113
95% UCL
16
14
12
12
11
9.3
8.3
7.7
7.9
7.7
7.1
6.1
5.7
5.8
5.6
5.3
4.9
4.8
5.0
4.9
5.0
4.8
4.5
4.1
3.8
3.7
LOAEL
90
Average
1.2
1.1
1.1
1.0
0.8
0.8
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
LOAEL
90
95% UCL
1.2
1.1
1.0
0.9
0.9
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
NOAEL
90
Average
12
11
11
9.9
8.4
7.7
6.7
6.2
5.9
5.7
5.4
5.0
4.6
43
4.2
4.0
3.8
3.6
3.6
3.6
3.6
3.5
3.3
3.2
3.0
2.9
NOAEL
90
95% UCL
12
11
9.9
9.3
8.6
8.0
7.0
6.4
6.1
5.9
5.6
5.1
4.8
4.5
4.3
4.2
3.9
3.7
3.7
3.7
3.7
3.6
3.5
3.3
3.1
3.0
LOAEL
50
Average
1.1
1.0
0.9
0.8
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
LOAEL
50
95% UCL
1.2
1.0
0.9
0.9
0.8
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
NOAEL
50
Average
11
10
9.0
8.3
7.6
7.1
63
5.8
5.4
5.1
4.9
43
4.2
3.9
3.7
3.6
3.4
3.2
3.1
3.1
3.1
3.0
2.9
2.8
2.7
2.6
NOAEL
50
95% UCL
12
10
93
83
7.8
73
6.5
6.0
53
S3
5.0
4.7
43
4.1
3.9
3.7
33
3.3
3-3
3.2
3.2
3.1
3.0
2.9
2.8
2.7
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-41: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE BALD EAGLE USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
26
19
16
19
17
14
12
11
12
11
10
8.2
7.8
8.8
8.1
7.6
6.6
6.9
7.7
6.9
7.5
6.8
6.3
5.8
5.2
5.1
LOAEL
152
95% UCL
26
19
16
19
18
14
12
11
13
12
10
8.4
8.0
9.0
8.3
7.7
6.7
7.0
7.9
7.1
7.7
7.0
6.5
5.9
5.3
5.2
NOAEL
152
Average
257
186
160
190
173
136
117
110
125
114
102
82
78
88
81
76
66
69
77
69
75
68
63
58
52
51
NOAEL
152
95% UCL
263
190
163
194
176
139
120
112
128
117
104
84
80
90
83
77
67
70
79
71
77
70
65
59
53
52
LOAEL
113
Average
17
15
13
12
12
11
9.2
8.3
8.5
8.4
7.7
6.8
6.3
6.2
6.1
5.8
5.3
5.1
5.4
5.3
5.5
5.2
4.9
4.6
4.2
4.0
LOAEL
113
95% UCL
18
15
14
13
12
11
9.4
8.5
8.6
8.6
7.9
7.0
6.4
6.3
6.2
6.0
5.4
5.2
5.5
5.4
5.6
5.3
5.0
4.7
4.3
4.1
NOAEL
113
Average
172
150
134
125
120
110
92
83
85
84
77
68
63
62
61
58
53
51
54
53
55
52
49
46
42
40
NOAEL
113
95% UCL
176
154
137
127
123
112
94
85
86
86
79
70
64
63
62
60
54
52
55
54
56
53
50
47
43
41
LOAEL
90
Average
3.9
3.6
3.2
2.9
2.8
2.5
2.3
2.1
2.0
1.9
1.8
1.7
1.6
1.5
1.4
1.4
1.3
1.2
1.2
1.2
1.2
1.2
1.1
1.1
1.0
1.0
LOAEL
90
95% UCL
4.0
3.6
3.3
3.0
2.8
2.6
2.3
2.1
2.0
1.9
1.9
1.7
1.6
1.5
1.4
1.4
1.3
1.3
1.2
1.2
1.3
1.2
1.2
1.1
1.1
1.0
NOAEL
90
Average
39
36
32
29
28
25
23
21
20
19
18
17
16
15
14
14
13
12
12
12
12
12
11
11
10
10
NOAEL
90
95% UCL
40
36
33
30
28
26
23
21
20
19
19
17
16
15
14
14
13
13
12
12
13
12
12
11
11
10
LOAEL
50
Average
3.7
3.3
3.0
2.7
2.5
2.3
2.1
1.9
1.8
1.7
1.6
1.5
1.4
1.3
1.3
1.2
1.1
1.1
1.1
1.0
1.1
1.0
1.0
0.9
0.9
0.9
LOAEL
50
95% UCL
3.8
3.4
3.0
2.8
2.5
23
2.1
1.9
1.8
1.7
1.6
1.5
1.4
1.3
1.3
1.2
1.2
1.1
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
NOAEL
50
Average
37
33
30
27
25
23
21
19
18
17
16
15
14
13
13
12
11
11
11
10
11
10
10
9.5
9.2
8.7
NOAEL
50
95% UCL
38
34
30
28
25
23
21
19
18
17
16
15
14
13
13
12
12
11
11
11
11
10
10
10
9.3
8.8
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-42: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE BELTED KINGFISHER USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
281
217
196
229
197
154
138
136
147
132
122
101
99
112
96
90
84
89
94
88
93
88
79
70
68
68
LOAEL
152
95% UCL
291
225
204
237
204
160
144
141
152
138
127
106
103
116
100
94
88
93
97
92
96
91
82
74
71
71
NOAEL
152
Average
563
434
393
457
394
308
277
272
294
264
244
203
198
223
192
180
167
178
187
176
185
175
158
140
135
136
NOAEL
152
95% UCL
582
450
408
473
408
321
289
283
305
275
254
212
207
232
200
188
175
186
195
184
193
183
165
147
142
142
LOAEL
113
Average
195
178
150
154
144
123
112
104
107
104
95
85
80
81
78
74
69
68
70
69
69
66
63
58
55
54
LOAEL
113
95% UCL
203
185
156
160
150
128
116
108
111
108
100
88
84
85
82
77
72
71
73
72
72
69
66
61
58
57
NOAEL
113
Average
391
356
301
308
288
247
223
208
214
207
191
169
161
162
156
148
138
135
140
137
138
133
126
116
110
108
NOAEL
113
95% UCL
405
369
313
319
299
257
233
217
222
216
199
177
168
169
163
155
144
142
146
144
145
139
131
121
115
113
LOAEL
90
Average
142
128
115
106
98
91
83
77
72
69
65
61
57
54
51
49
47
45
44
44
43
42
41
39
38
37
LOAEL
90
95% UCL
147
132
119
110
102
95
86
80
74
71
68
64
60
56
54
51
49
47
46
46
45
43
42
41
39
38
NOAEL
90
Average
283
255
229
212
196
183
166
153
143
138
131
122
114
107
103
99
94
90
88
87
86
83
81
79
75
73
NOAEL
90
95% UCL
293
264
238
220
203
190
172
159
149
143
136
127
119
112
107
103
98
94
92
91
90
87
85
82
79
76
LOAEL
50
Average
142
128
115
106
98
91
83
77
72
69
65
61
57
54
51
49
47
45
44
44
43
42
41
39
38
37
LOAEL
50
95% UCL
147
132
119
110
102
95
86
80
74
71
68
64
60
56
54
51
49
47
46
46
45
43
42
41
39
38
NOAEL
50
Average
283
255
229
212
196
183
166
153
143
138
131
122
114
107
103
99
94
90
88
87
86
83
81
79
75'
73
NOAEL
50
95% UCL
293
264
238
220
203
190
172
159
149
143
136
127
119
112
107
103
98
94
92
91
90
87
85
82
79
76
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-43: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE GREAT BLUE HERON USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
. 7.4
5.5
4.9
5.9
5.0
3.7
3.2
3.1
3.5
3.1
2.8
2.2
2.2
2.6
2.1
2.0
1.8
2.0
2.1
2.0
2.1
2.0
1.7
1.5
1.4
1.4
LOAEL
152
95% UCL
7.6
5.6
5.0
6.1
5.1
3.8
3.3
3.2
3.6
3.2
2.9
2.3
2.2
2.7
2.2
2.0
1.8
2.0
2.2
2.0
2.2
2.1
1.8
1.5
1.4
1.5
NOAEL
152
Average
12
9.1
8.1
10
8.3
6.1
5.4
5.2
5.9
5.2
4.7
3.7
3.6
4.3
3.5
3.3
3.0
3.3
3.6
3.3
3.6
3.3
2.9
2.4
2.3
2.3
NOAEL
152
95% UCL
13
9.4
8.3
10
8.5
6.3
5.6
5.4
6.0
5.3
4.9
3.8
3.7
4.5
3.6
3.4
3.0
3.4
3.7
3.4
3.7
3.4
3.0
2.5
2.4
2.4
LOAEL
113
Average
5.0
4.5
3.6
3.8
3.5
2.9
2.6
2.4
2.5
2.4
2.2
1.9
1.7
1.8
1.7
1.6
1.5
1.4
1.5
1.5
1.5
1.5
1.4
1.2
1.1
1.1
LOAEL
113
95% UCL
5.1
4.6
3.8
3.9
3.6
3.0
2.6
2.4
2.5
2.5
2.2
1.9
1.8
1.8
1.8
1.7
1.5
1.5
1.6
1.5
1.6
1.5
1.4
1.3
1.2
1.1
NOAEL
113
Average
8.3
7.5
6.1
6.3
5.9
4.8
4.3
3.9
4.1
4.0
3.6
3.1
2.9
3.0
2.8
2.7
2.4
2.4
2.5
2.5
2.5
2.4
2.3
2.0
1.9
1.8
NOAEL
113
95% UCL
8.5
7.7
6.3
6.5
6.1
5.0
4.4
4.1
4.2
4.1
3.7
3.2
3.0
3.1
2.9
2.8
2.5
2.5
2.6
2.6
2.6
2.5
2.3
2.1
2.0
1.9
LOAEL
90
Average
4.0
3.6
3.1
2.9
2.7
2.5
2.2
2.0
1.9
1.8
1.7
1.6
1.5
1.4
1.3
1.3
1.2
1.1
1.1
1.1
1.1
1.1
1.1
1.0
0.9
0.9
LOAEL
90
95% UCL
4.1
3.7
3.2
3.0
2.8
2.6
2.2
2.0
1.9
.9
.8
.6
.5
.4
.4
.3
.2
.2
.2
.2
.2
.1
1.1
1.0
1.0
0.9
NOAEL
90
Average
6.6
6.0
5.2
4.9
4.5
4.2
3.6
3.3
3.1
3.1
2.9
2.6
2.4
2.3
2.2
2.1
2.0
.9
.9
.9
.9
.8
1.8
1.7
1.6
1.5
NOAEL
90
95% UCL
6.8
6.1
5.4
5.0
4.7
4.3
3.7
3.4
3.2
3.1
3.0
2.7
2.5
2.4
2.3
2.2
2.0
1.9
1.9
2.0
2.0
1.9
1.8
1.7
1.6
1.5
LOAEL
50
Average
3.8
3.4
3.0
2.7
2.5
23
2.1
1.9
1.7
1.7
1.6
1.5
13
1.3
1.2
1.1
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.8
LOAEL
50
95% UCL
3.9
3.4
3.1
2.8
2.6
2.4
2.1
1.9
1.8
1.7
1.6
1.5
1.4
1.3
1.2
1.2
1.1
1.1
1.0
1.0
1.0
1.0
1.0
0.9
0.9
0.8
NOAEL
50
Average
6.3
5.6
5.0
4.5
4.2
3.9
3.4
3.1
2.9
2.8
2.6
2.4
2.2
2.1
2.0
1.9
1.8
1.7
1.7
1.7
1.6
1.6
1.5
1.5
1.4
1.4
NOAEL
50
95% UCL
6.4
5.7
5.1
4.7
4.3
4.0
3.5
3.2
3.0
2.8
2.7
2.5
23
2.2
2.1
2.0
1.9
1.8
1.7
1.7
1.7
1.6
1.6
1.5
1.5
1.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-44: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE BALD EAGLE USING TEQ FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
2683
1944
1669
1982
1802
1416
1223
1148
1304
1193
1063
858
817
920
849
789
685
719
806
720
786
715
660
604
544
530
LOAEL
152
95% UCL
2741
1986
1705
2024
1840
1447
1250
1173
1333
1218
1087
877
835
941
868
806
701
735
824
736
803
731
675
617
556
542
NOAEL
152
Average
5367
3889
3338
3963
3604
2832
2446
2295
2608
2386
2125
1715
1633
1841
1698
1577
1370
1439
1611
1440
1572
1430
1320
1207
1087
1060
NOAEL
152
95% UCL
5482
3973
3409
4049
3680
2893
2499
2346
2665
2437
2173
1753
1670
1882
1735
1612
1401
1471
1648
1472
1607
1462
1351
1234
1111
1085
LOAEL
113
Average
1795
1569
1395
1301
1256
1144
958
871
882
873
802
711
654
646
633
608
556
531
565
552
570
540
515
480
442
420
LOAEL
113
95% UCL
1834
1603
1425
1329
1284
1169
979
891
902
892
820
727
669
661
647
622
568
543
578
564
583
552
526
490
452
429
NOAEL
113
Average
3590
3138
2790
2602
2513
2288
1916
1743
1765
1746
1604
1423
1309
1293
1265
1217
1112
1062
1130
1104
1139
1080
1029
959
884
839
NOAEL
113
95% UCL
3668
3206
2850
2658
2567
2339
1958
1782
1804
1785
1641
1455
1339
1322
1293
1244
1137
1086
1155
1129
1165
1104
1053
981
904
859
LOAEL
90
Average
412
375
338
308
288
265
239
217
206
200
190
176
163
154
149
143
135
129
127
126
130
123
119
114
108
102
LOAEL
90
95% UCL
418
381
343
313
292
269
243
220
209
203
193
179
165
157
151
146
138
131
130
129
132
125
121
116
110
104
NOAEL
90
Average
823
749
676
615
575
530
477
433
411
400
380
353
325
308
297
287
271
257
255
253
259
245
238
228
216
204
NOAEL
90
95% UCL
837
762
687
626
585
539
485
441
418
407
386
359
331
314
303
292
276
262
260
257
264
250
242
232
220
208
LOAEL
50
Average
388
347
313
283
261
240
218
200
187
178
169
158
147
138
132
126
120
114
111
108
111
105
102
99
96
90
LOAEL
50
95% UCL
394
353
318
287
265
243
222
203
190
181
172
161
150
141
134
128
122
116
113
110
112
106
104
101
97
92
NOAEL
50
Average
776
695
625
565
522
479
437
400
374
357
338
316
294
276
263
252
239
228
221
216
221
209
204
198
191
181
NOAEL
50
95% UCL
788
706
635
574
531
487
444
407
380
363
344
322
299
281
268
256
244
232
225
220
225
213
207
201
195
184
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5^5: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE BAT FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
4.1
3.7
3.6
3.5
33
3.2
3.0
3.0
3.0
2.8
2.7
2.6
2.6
2.5
2.4
2.3
2.3
23
2.2
2.1
2.1
2.0
2.0
2.0
2.0
1.9
LOAEL
152
95% UCL
4.4
4.0
3.8
3.8
3.6
3.4
33
33
3.2
3.1
2.9
2.8
2.8
2.6
2.6
2.5
2.5
2.4
23
2.3
2.2
2.2
2.1
2.2
2.1
2.1
NOAEL
152
Average
19
17
17
17
16
15
14
14
14
13
13
12
12
12
11
11
11
11
10
10
10
10
93
9.4
93
9.0
NOAEL
152
95% UCL
21
19
18
18
17
16
15
15
15
14
13
13
13
12
12
12
12
11
11
11
10
10
10
10
10
10
LOAEL
113
Average
3.3
3.1
2.9
2.8
2.7
2.6
2.6
2.4
2.4
23
2.2
2.1
2.1
2.1
2.0
1.9
1.9
1.8
1.8
1.8
1.7
1.7
1.6
1.6
1.6
1.6
LOAEL
113
95% UCL
3.5
33
3.1
3.0
2.9
2.8
2.7
2.6
2.6
2.5
2.4
23
2.3
2.2
2.2
2.1
2.0
2.0
2.0
1.9
1.9
1.8
1.8
1.7
1.7
1.7
NOAEL
113
Average
15
14
14
13
13
12
12
11
11
11
10
10
10
10
9.5
9.1
8.8
8.6
8.6
8.4
8.2
8.0
7.7
7.5
7.4
7.4
NOAEL
113
95% UCL
16
15
15
14
14
13
13
12
12
12
11
11
11
10
10
10
10
93
9.2
9.0
8.8
8.6
83
8.1
8.0
8.0
LOAEL
90
Average
2.7
23
2.4
2.2
2.2
2.1
2.0
1.9
1.9
1.8
1.8
1.7
1.6
1.6
1.5
1.5
1.5
1.4
1.4
13
13
1.3
1.2
1.2
1.2
1.2
LOAEL
90
95% UCL
2.9
2.7
2.5
2.4
23
2.2
2.1
2.1
2.0
2.0
1.9
1.8
1.8
1.7
1.6
1.6
1.6
1.5
1.5
1.4
1.4
1.4
13
1.3
13
13
NOAEL
90
Average
12
12
11
11
10
10
9.4
9.1
8.9
8.6
8.4
8.0
7.7
7.4
7.2
7.0
6.8
6.7
6.5
63
6.1
6.0
5.9
5.7
5.7
5.6
NOAEL
90
95%UCL
13
13
12
11
11
10
10
10
9.5
9.2
9.1
8.6
83
7.9
7.7
7.5
73
7.2
6.9
6.8
6.6
6.4
63
6.2
6.1
6.0
LOAEL
50
Average
20
18
17
17
16
15
15
14
14
14
13
13
12
12
11
11
11
11
10
10
10
10
9.4
9.2
9.0
8.8
LOAEL
50
95% UCL
21
20
19
18
17
16
16
15
15
15
14
14
13
13
12
12
12
11
11
11
11
10
10
10
10
9.4
NOAEL
50
Average
9.1
8.6
8.2
7.8
7.6
7.2
7.0
6.8
6.5
6.4
6.2
5.9
5.7
5.5
53
5.2
5.1
5.0
4.9
4.8
4.6
43
4.4
43
4.2
4.1
NOAEL
50
95% UCL
10
9.2
8.7
8.4
8.1
7.7
7.5
7.2
7.0
6.9
6.7
6.4
6.1
5.9
5.7
5.5
5.4
53
5.2
5.1
5.0
4.9
4.7
4.6
4.5
4.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-46: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE BAT ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
67
60
58
57
54
52
49
49
48
46
43
43
41
40
39
38
37
37
35
34
33
33
32
32
32
31
LOAEL
152
95% UCL
72
64
62
61
58
55
53
53
52
49
47
46
45
43
42
41
40
39
38
37
36
36
35
35
35
34
NOAEL
152
Average
667
598
578
571
540
517
490
493
482
461
433
426
415
398
393
379
372
365
350
341
334
331
321
324
320
312
NOAEL
152
95% UCL
716
641
620
612
579
555
527
529
517
495
466
459
446
428
423
409
401
393
376
367
359
355
346
351
347
338
LOAEL
113
Average
53
50
48
46
44
43
41
40
39
38
36
35
34
33
33
31
30
30
30
29
28
28
27
26
26
26
LOAEL
113
95% UCL
57
53
51
49
47
46
44
43
42
40
39
37
37
36
35
34
33
32
32
31
30
30
29
28
28
28
NOAEL
113
Average
529
496
475
457
438
426
413
396
390
377
359
347
343
333
326
314
305
299
296
289
283
276
266
259
256
256
NOAEL
113
95% UCL
568
531
510
490
469
456
443
425
419
404
386
373
369
358
351
338
328
322
318
311
305
297
286
279
277
277
LOAEL
90
Average
43
41
38
36
35
34
32
31
31
30
29
28
27
25
25
24
24
23
22
22
21
21
20
20
20
19
LOAEL
90
95% UCL
46
44
41
39
38
36
35
34
33
32
31
30
29
27
27
26
25
25
24
23
23
22
22
21
21
21
NOAEL
90
Average
430
408
382
364
353
336
324
315
306
298
291
276
266
254
248
240
235
231
223
218
211
206
202
198
195
192
NOAEL
90
95%UCL
462
437
410
390
378
360
347
337
327
319
313
297
286
273
266
258
253
248
240
234
227
222
217
213
210
206
LOAEL
50
Average
32
30
28
27
26
25
24
23
23
22
21
20
20
19
18
18
17
17
17
16
16
16
15
15
15
14
LOAEL
50
95% UCL
34
32
30
29
28
27
26
25
24
24
23
22
21
20
20
19
19
18
18
18
17
17
16
16
16
15
NOAEL
50
Average
316
296
281
271
262
248
241
233
226
222
214
205
197
188
183
178
174
171
168
165
160
156
152
149
145
142
NOAEL
50
95% UCL
339
318
302
290
281
265
258
250
242
238
230
220
212
202
197
191
188
184
181
177
172
167
163
160
156
153
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-47: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE RACCOON FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
LOAEL
152
95% UCL
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
NOAEL
152
Average
3.5
3.1
3.0
3.0
2.8
2.7
2.5
2.5
2.5
2.4
2.2
2.2
2.1
2.0
2.0
1.9
1.9
1.9
1.8
1.7
1.7
1.7
1.6
1.6
1.6
1.6
NOAEL
152
95% UCL
3.8
3.3
3.2
3.2
3.0
2.9
2.7
2.7
2.7
2.6
2.4
2.4
2.3
2.2
2.2
2.1
2.1
2.0
1.9
1.9
1.9
1.8
1.8
1.8
1.8
1.7
LOAEL
113
Average
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
LOAEL
113
95% UCL
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
NOAEL
113
Average
2.8
2.6
2.5
2.4
2.3
2.2
2.1
2.0
2.0
1.9
1.8
1.8
1.7
1.7
1.7
1.6
1.5
1.5
1.5
1.5
1.4
1.4
1.4
1.3
1.3
1.3
NOAEL
113
95% UCL
3.0
2.8
2.6
2.5
2.4
2.4
2.3
2.2
2.2
2.1
2.0
1.9
1.9
1.8
1.8
1.7
1.7
1.7
.6
.6
.6
.5
.5
1.4
1.4
1.4
LOAEL
90
Average
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
LOAEL
90
95% UCL
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
NOAEL
90
Average
2.2
2.1
2.2
2.1
1.8
1.7
1.7
1.6
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.2
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
1.0
1.0
NOAEL
90
95% UCL
2.4
2.3
2.1
2.0
2.0
1.9
1.8
1.7
1.7
1.6
1.6
1.5
1.5
1.4
1.4
1.3
13
1.3
1.2
1.2
1.2
.1
.1
.1
.1
.1
LOAEL
50
Average
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
LOAEL
50
95% UCL
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
NOAEL
50
Average
1.7
1.6
1.5
1.4
1.4
13
1.2
1.2
1.2
1.1
1.1
1.1
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.7
0.7
NOAEL
50
95% UCL
1.8
1.7
1.6
1.5
1.5
1.4
13
13
13
1.2
1.2
1.1
1.1
1.0
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
Bold values indicate exceedances
MCA/TAMS
-------
TABLE 5-48: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE RACCOON ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
15
13
12
12
12
11
10
10
10
10
9.3
9.0
8.7
8.5
8.3
8.0
7.8
7.7
7.5
7.2
7.1
7.0
6.8
6.7
6.6
6.5
LOAEL
152
95% UCL
16
14
14
14
13
12
12
12
11
11
10
10
10
10
10
9.3
9.3
9.0
8.6
8.3
8.2
8.1
7.8
8.0
8.0
7.7
NOAEL
152
Average
147
131
124
125
118
111
104
104
103
98
93
90
87
85
83
80
78
77
75
72
71
70
68
67
66
65
NOAEL
152
95% UCL
158
142
137
136
129
123
118
118
115
111
105
104
102
97
95
93
93
90
86
83
82
81
78
80
80
77
LOAEL
113
Average
12
11
10
10
9.5
9.1
8.7
8.4
8.3
8.0
7.6
7.3
7.2
7.0
6.9
6.6
6.4
6.3
6.2
6.1
6.0
5.8
5.6
5.4
5.4
5.3
LOAEL
113
95% UCL
12
12
11
11
10
10
10
9.5
9.3
9.0
8.7
8.4
8.3
8.1
7.9
7.7
7.5
7.3
7.2
7.0
6.9
6.7
6.5
6.4
6.3
6.3
NOAEL
113
Average
115
108
102
99
95
91
87
84
83
80
76
73
72
70
69
66
64
63
62
61
60
58
56
54
54
53
NOAEL
113
95% UCL
125
117
112
108
104
101
98
95
93
90
87
84
83
81
79
77
75
73
72
70
69
67
65
64
63
63
LOAEL
90
Average
9
9
13
13
7.6
7.2
6.9
6.7
6.5
6.3
6.2
5.9
5.6
5.4
5.2
5.1
5.0
4.9
4.7
4.6
4.5
4.4
4.3
4.2
4.1
4.0
LOAEL
90
95% UCL
10
10
9.0
8.6
8.3
8.0
7.6
7.4
7.2
7.1
6.9
6.6
6.4
6.1
6.0
5.8
5.7
5.6
5.4
5.3
5.1
5.0
4.9
4.8
4.7
4.6
NOAEL
90
Average
93
88
132
129
76
72
69
67
65
63
62
59
56
54
52
51
50
49
47
46
45
44
43
42
41
40
NOAEL
90
95% UCL
101
95
90
86
83
80
76
74
72
71
69
66
64
61
60
58
57
56
54
53
51
50
49
48
47
46
LOAEL
50
Average
7.0
6.5
6.2
5.9
5.7
5.4
5.2
5.0
4.8
4.7
4.6
4.4
4.2
4.0
3.9
3.8
3.7
3.6
3.6
3.5
3.4
3.3
3.2
3.1
3.1
3.0
LOAEL
50
95% UCL
7.5
7.1
6.7
6.5
6.2
5.9
5.8
5.6
5.4
53
5.1
4.9
4.8
4.6
4.5
4.4
4.3
4.2
4.1
4.0
3.9
3.8
3.7
3.6
3.5
3.5
NOAEL
50
Average
70
65
62
59
57
54
52
50
48
47
46
44
42
40
39
38
37
36
36
35
34
33
32
31
31
30
NOAEL
50
95% UCL
75
71
67
65
62
59
58
56
54
53
51
49
48
46
45
44
43
42
41
40
39
38
37
36
35
35
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-49: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE MINK FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
1.1
0.8
0.8
0.9
0.8
0.6
0.6
0.6
0.6
0.5
0.5
0.4
0.4
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
LOAEL
152
95% UCL
1.1
0.9
0.8
0.9
0.8
0.7
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
NOAEL
152
Average
34
27
25
28
25
20
18
18
19
18
16
14
14
15
13
13
12
12
13
12
12
12
11
10
10
10
NOAEL
152
95% UCL
36
28
26
29
26
21
19
19
20
18
17
15
15
16
14
13
13
13
13
13
13
13
12
11
10
10
LOAEL
113
Average
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
LOAEL
113
95% UCL
0.8
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
NOAEL
113
Average
24
22
19
20
18
16
15
14
14
14
13
12
11
11
11
10
10
10
10
10
10
9.3
8.8
8.3
8.0
7.8
NOAEL
113
95% UCL
25
23
20
21
19
17
16
15
15
15
14
12
12
12
12
11
10
10
10
10
10
10
93
8.8
8.4
8.3
LOAEL
90
Average
0.6
0.6
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
LOAEL
90
95% UCL
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
NOAEL
90
Average
20
18
17
16
14
14
12
12
11
11
10
10
9.1
8.6
8.4
8.1
7.8
7.5
7.3
7.3
7.2
6.9
6.8
6.5
6.3
6.1
NOAEL
90
95% UCL
21
19
17
16
15
14
13
12
12
11
11
10
10
9.1
8.8
8.6
8.2
7.9
7.8
7.7
7.6
7.3
7.1
6.9
6.6
6.5
LOAEL
50
Average
0.5
0.5
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
LOAEL
50
95% UCL
0.5
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
NOAEL
50
Average
17
15
14
13
12
11
11
10
93
8.9
8.5
8.0
7.6
7.1
6.9
6.6
6.4
6.1
6.0
5.9
5.8
5.6
5.5
53
5.1
5.0
NOAEL
50
95% UCL
18
16
15
14
13
12
11
10
10
9.4
9.0
8.4
8.0
7.5
7.2
7.0
6.7
6.5
63
6.2
6.1
5.9
5.8
5.6
5.4
53
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-50: RATIO OF MODELED DIETARY DOSE TO TOXICITY BENCHMARKS
FOR FEMALE OTTER FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
14
10
8.8
10
9.5
7.4
6.4
6.0
6.8
6.3
5.6
4.5
4.3
4.8
4.5
4.1
3.6
3.8
4.2
3.8
4.1
3.8
3.5
3.2
2.9
2.8
LOAEL
152
95% UCL
14
10
8.9
11
10
7.6
6.6
6.2
7.0
6.4
5.7
4.6
4.4
4.9
4.6
4.2
3.7
3.9
4.3
3.9
4.2
3.8
3.5
3.2
2.9
2.8
NOAEL
152
Average
458
332
285
338
307
242
209
196
223
204
181
146
139
157
145
135
117
123
137
123
134
122
113
103
93
90
NOAEL
152
95% UCL
468
339
291
345
314
247
213
200
227
208
185
150
143
161
148
138
120
125
141
126
137
125
115
105
95
93
LOAEL
113
Average
9.4
8.2
7.3
6.8
6.6
6.0
5.0
4.6
4.6
4.6
4.2
3.7
3.4
3.4
3.3
3.2
2.9
2.8
3.0
2.9
3.0
2.8
2.7
2.5
2.3
2.2
LOAEL
113
95% UCL
10
8.4
7.5
7.0
6.7
6.1
5.1
4.7
4.7
4.7
4.3
3.8
3.5
3.5
3.4
3.3
3.0
2.9
3.0
3.0
3.1
2.9
2.8
2.6
2.4
2.3
NOAEL
113
Average
306
268
238
222
214
195
163
149
151
149
137
121
112
110
108
104
95
91
96
94
97
92
88
82
75
72
NOAEL
113
95% UCL
313
273
243
227
219
200
167
152
154
152
140
124
114
113
110
106
97
93
99
96
99
94
90
84
77
73
LOAEL
90
Average
2.2
2.0
1.8
1.6
1.5
1.4
1.3
1.1
1.1
1.1
1.0
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.5
LOAEL
90
95% UCL
2.2
2.0
1.8
1.6
1.5
1.4
1.3
1.2
1.1
1.1
1.0
0.9
0.9
0.8
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.6
. 0.6
0.6
0.5
NOAEL
90
Average
70
64
58
53
49
45
41
37
35
34
32
30
28
26
25
25
23
22
22
22
22
21
20
19
18
17
NOAEL
90
95% UCL
71
65
59
53
50
46
41
38
36
35
33
31
28
27
26
25
24
22
22
22
23
21
21
20
19
18
LOAEL
50
Average
2.0
1.8
1.6
1.5
1.4
1.3
1.1
1.1
1.0
0.9
0.9
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
LOAEL
50
95% UCL
2.1
1.9
1.7
1.5
1.4
1J
1.2
1.1
1.0
1.0
0.9
0.8
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
NOAEL
50
Average
66
59
53
48
45
41
37
34
32
30
29
27
25
24
22
22
20
19
19
18
19
18
17
17
16
15
NOAEL
50
95% UCL
67
60
54
49
45
42
38
35
32
31
29
27
26
24
23
22
21
20
19
19
19
18
18
17
17
16
Bold values indicate excecdanccs
TAMS/MCA
-------
TABLE 5-51: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE MINK ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
7.1
5.7
5.2
5.8
5.1
4.2
3.8
3.8
4.0
3.6
3.4
2.9
2.9
3.1
2.8
2.6
2.5
2.6
2.6
2.5
2.6
2.5
2.3
2.1
2.0
2.0
LOAEL
152
95% UCL
7.4
5.9
5.4
6.1
5.4
4.4
4.1
4.0
4.2
3.9
3.6
3.2
3.1
3.3
3.0
2.8
2.7
2.7
2.8
2.7
2.7
2.6
2.4
2.3
2.2
2.2
NOAEL
152
Average
199
158
145
163
143
118
107
106
111
102
95
82
80
87
77
73
69
72
74
70
72
69
63
59
57
57
NOAEL
152
95% UCL
207
166
153
171
150
124
114
112
117
108
100
88
86
92
83
79
75
77
78
75
77
74
68
64
62
61
LOAEL
113
Average
5.1
4.6
4.0
4.1
3.8
3.4
3.1
2.9
3.0
2.9
2.7
2.4
2.3
2.3
2.3
2.2
2.0
2.0
2.0
2.0
2.0
1.9
1.8
1.7
1.6
1.6
LOAEL
113
95% UCL
5.3
4.9
4.3
4.3
4.0
3.6
3.3
3.1
3.2
3.1
2.9
2.6
2.5
2.5
2.4
2.3
2.2
2.1
2.2
2.1
2.1
2.1
2.0
1.8
1.8
1.8
NOAEL
113
Average
142
130
113
114
107
95
87
82
83
81
75
68
66
65
63
60
57
56
57
56
56
54
51
48
46
45
NOAEL
113
95% UCL
148
136
119
120
113
100
93
87
88
86
80
73
70
70
68
65
61
60
61
60
60
58
55
52
50
49
LOAEL
90
Average
4.1
3.8
3.7
3.5
3.0
2.8
2.5
2.4
2.3
2.2
2.1
2.0
1.9
1.8
1.7
1.7
1.6
1.5
1.5
1.5
1.5
1.4
1.4
1.4
1.3
1.3
LOAEL
90
95% UCL
4.3
3.9
3.6
3.3
3.2
3.0
2.7
2.5
2.4
2.4
2.3
2.1
2.0
1.9
1.9
1.8
1.7
1.7
1.6
1.6
1.6
1.5
1.5
1.4
1.4
1.4
NOAEL
90
Average
114
105
104
99
84
79
71
67
64
63
60
56
53
50
49
47
45
43
43
42
42
40
39
38
36
35
NOAEL
90
95% UCL
119
110
99
94
89
83
75
71
68
66
64
60
56
54
52
50
48
46
46
45
44
43
42
41
39
38
LOAEL
50
Average
3.5
3.2
2.9
2.7
2.5
2.4
2.2
2.0
1.9
1.9
1.8
1.7
1.6
1.5
1.4
1.4
1.3
13
1.2
13
1.2
1.2
1.1
1.1
1.1
1.0
LOAEL
50
95% UCL
3.7
3.3
3.0
2.8
2.7
2.5
2.3
2.1
2.0
2.0
1.9
1.8
1.7
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.3
1.2
1.2
1.2
1.1
1.1
NOAEL
50
Average
99
90
82
76
71
66
61
57
54
52
50
47
44
41
40
38
37
36
35
34
34
33
32
31
30
29
NOAEL
50
95% UCL
103
94
85
80
74
70
64
60
57
55
52
49
47
44
42
41
39
38
37
37
36
35
34
33
32
31
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-52: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE OTTER ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
95
69
59
71
64
50
44
41
46
42
38
31
29
33
30
28
24
26
29
26
28
25
24
22
19
19
LOAEL
152
95% UCL
98
71
61
72
66
52
45
42
47
43
39
31
30
34
31
29
25
26
29
26
29
26
24
22
20
19
NOAEL
152
Average
2674
1938
1664
1975
1796
1412
1220
1145
1300
1189
1060
856
815
918
847
787
684
718
804
719
784
713
659
602
543
529
NOAEL
152
95% UCL
2732
1981
1700
2018
1835
1443
1247
1171
1330
1216
1085
876
835
939
866
805
700
735
823
735
802
730
675
617
556
542
LOAEL
113
Average
64
56
50
46
45
41
34
31
31
31
29
25
23
23
23
22
20
19
20
20
20
19
18
17
16
15
LOAEL
113
95% UCL
65
57
51
47
46
42
35
32
32
32
29
26
24
24
23
22
20
19
21
20
21
20
19
18
16
15
NOAEL
113
Average
1789
1564
1391
1297
1253
1141
955
869
880
871
800
710
653
645
631
607
555
530
564
551
568
539
514
479
441
419
NOAEL
113
95% UCL
1828
1598
1421
1326
1280
1167
977
890
900
891
819
727
669
660
646
621
568
543
577
564
582
552
526
490
452
429
LOAEL
90
Average
15
13
12
11
10
9.5
8.5
7.8
7.4
7.2
6.8
6.3
5.8
5.5
5.3
5.1
4.9
4.6
4.6
4.5
4.7
4.4
4.3
4.1
3.9
3.7
LOAEL
90
95% UCL
15
14
12
11
10
10
8.7
7.9
7.5
7.3
6.9
6.4
6.0
5.6
5.4
5.3
5.0
4.7
4.7
4.6
4.8
4.5
4.4
4.2
4.0
3.8
NOAEL
90
Average
412
375
345
315
288
266
239
217
206
200
190
177
163
155
149
144
136
129
128
127
130
123
119
114
108
103
NOAEL
90
95% UCL
419
382
344
314
293
270
244
221
210
204
194
180
167
158
153
147
139
132
131
130
133
126
122
117
111
105
LOAEL
50
Average
14
12
11
10
9.3
8.6
7.8
7.2
6.7
6.4
6.0
5.7
5.3
4.9
4.7
4.5
4.3
4.1
4.0
3.9
4.0
3.7
3.6
3.5
3.4
3.2
LOAEL
50
95% UCL
14
13
11
10
9.5
8.7
7.9
7.3
6.8
6.5
6.2
5.8
5.4
5.0
4.8
4.6
4.4
4.2
4.0
4.0
4.0
3.8
3.7
3.6
3.5
3.3
NOAEL
50
Average
388
347
313
283
261
240
219
200
187
179
169
159
147
139
132
126
120
114
111
109
111
105
102
99
96
91
NOAEL
. 50
95% UCL
394
353
318
288
266
244
222
204
191
182
173
162
150
141
135
129
123
117
113
111
113
107
104
101
98
93
Bold values indicate exceedances
TAMS/MCA
-------
THIS PAGE LEFT INTENTIONALLY BLANK
-------
Figures
-------
me No.: 48 FDe Nome: U:\proJect3J\Hud-cadd\Reports\Ecosow\Fig1.1.dwg DolK 3-8-99 User: CM
UNITED STATES
Lake
Cha.mpla.in
LAKE ONTARIO
NEW YORK
.MASSACHUSETTS
PENNSYLVANIA
CONNECTICUT
SOURCES:
1. NEW YORK STATE DEPARTMENT OF TRANSPORTATION. 1987
2. NYSOEC, 1978.
--- DRAINAGE BASIN
EW YORK
CIT
-------
Figure 1-2
Eight-Step Ecological Risk Assessment Process for Superfund
Hudson River PCB Reassessment
Ecological Risk Assessment
60
c
'•^ c
.2 o
X '*3
W 2
0 c
^ .o
U
o
*4_»
O
—
"o
U
STEP 1: SCREENING-LEVEL:
• Site Visit
• Problem Formulation
• Toxicity Evaluation
STEP 2: SCREENING-LEVEL:
• Exposure Estimate
• Risk Calculation
STEP 3: PROBLEM FORMULATION
Toxicity Evaluation
Assessment
Endpoints
-^ ^-
Conceptual Model
Exposure Pathways
t
Questions/Hypotheses
STEP 4: STUDY AND DESIGN DQO PROCESS
• Lines of Evidence
• Measurement Endpoints
Work Plan and Sampling and Analysis Plan
STEP 5: VERIFICATION OF FIELD
SAMPLING DESIGN
STEP 6: SITE INVESTIGAITON
AND DATA ANALYSIS
Risk Assessor and
Risk Manager
Agreement
H SMDP
SMDP
H SMDP
SMDP
SMDP
STEP 7: RISK CHARACTERIZATION
STEP 8: RISK MANAGEMENT
SMDP
-------
.Fl«namc U'\fROCCn.I3\HUO-CAOO\fCOSOW\nC2-2.d«g / OclK J-g-gt / lrt«: Ql
Figure 2-1
Baseline Ecological Risk Assessment
Lower Hudson River Sampling Stations
OH uu>-
ALBANY
150-
Troy
-ECO-STATION 10 (RM 143.5)
CCO-STATION 11A (RM 137.2)-
ECO-STATION 118 (RM 136.7)-
ECO-STAT10N 12A (RM 122.7}
ECO-STATION 12B (RM 122.4) X
125-
-ECO-STATION 13 (RM 113.8)
Brld9.
TOO
-ECO-STATION 14 (RM 100.0)
-ECO-STAHON 15A (RM 89.4)
-ECO-STATION 158 (RM 88.9)
%=s
ECO-STATION 16 (RM 58.7)-
LEGEND
190 RIVER MILE
ECOLOGICAL SAMPLING
LOCATION
-ECO-STATION 17 (RM 47.3)
ECO-STATION 18 (RM 25.8)-
Newark
NEW YORK
CITY
i Sottvy
TANS /MCA
-------
Figure 2-2
Hudson River PCB Reassessment
Conceptual Model Diagram Including Floodplain Soils
Floodplain Soils
HUDSON RIVER
SEDIMENTS
Exposure point for
macroinvertebrates
and fish living in the
PCB
SOURCES
GE FACILITIES
HUDSON RIVER
WATER COLUMN
Direct exposure point
for macroinvertebrates
and fish living in the
river and for birds and
mammals drinking from
the river
Trophic Level
Receptors
* Aquatic
invertebrates
Trophic Level 1 -3
Receptors
Terrestrial
invertebrates
Amphibians
and reptiles
Burrowing
animals
Trophic Level 2-3
Receptors
1 Terrestrial omnivores
Trophic Level 2
Receptors
• Forage fish
Trophic Level 2-3
Receptors
Piscivorous Fish
Trophic Level 3-5
Receptors
•Consumption of a
variety of invertebrate,
fish, avian and
mammalian prey
Trophic Level 2
Receptors
Insectivorous birds
and mammals
Trophic Level 3-5
Receptors
• Piscivorous birds
• Piscivorous mammals
Notes:
1. All receptors may be directly exposed to river water and sediments.
2. Trophic levels are provided as a general guide to bioaccumulation
potential, but vary according to species and food availability.
-------
VERTICAL SEGMENTATION
w mm <•__ . vi
ALBANY l£IROY /"MA
->*Fm '
WATER -
SEDIMENT-
• POUGHKEEPSIE
CONN. R.
V^t. NEW LONDON
HOUSATONICR.
BRIDGEPORT
MONTAUK PT
FOOD WEB
REGION #3
xifi
16
ASBURY PARK
FOOD WEB
FOODWEB
LAKEWOOD.
SEAWARD
STUDY LIMIT
SCAtE= STATUTE MULES
!....» I I I I
0 10 ZO 30 40 50 —
72°
' ATLANTIC
CITY
Source: Farley et al., 1999
Note: Model segment numbers 1-30 pertain to the Fate and transport model. Model segments are
combined into five food web regions for the bioaccumulation model calculations
Figure 3-1
Revised Segments and Regions of the Farley Model for PCBs in Hudson River Estuary
and Surrounding Waters
TAMS/MCA
-------
3500
3000 -
~ 2500 -i
(A
§ 2000 -.
g 1500 -i
** 1000 -j
500 -I
0
6000 -7
Di
0
4000
3500
^ 3000
| 2500
^ 2000
0 1500
1000
500
0
Tetra
•/> 800 -
CO
U 600 -_
600
500 ^
^ 400 -
&
^ 300 -
CD
a! 200 -3
100
0
Hexa
+
V
Farley el al.. 1 999
USEPA. 2000
-f-
V
l;arleyet al.. 1999
USKPA. 2000
******
#*«"**
+,.,+«+(+K'
,,^^^
^&-
+-
V
l-'arlcyci al.. 1999
USI-PA, 2000
4-+
4l,_,
Sources: Farley et al., 1999 and USEPA, 2000
Figure 3-2
Comparison of Cumulative PCB Loads at WaterFord from Farley et al., 1999 and
USEPA, 2000
TAMS/MCA
-------
10
"2 oo
O 3
** *
oa > 6
.S Dfi
C "•
U 0.
"2 -1
5 X 4
o Q
DD 3
co ^
U -o 2
Region 1-White Perch (Age Class 1-7 Years)
Tri+ PCBs Fish Body Burdens
y= 1.68 + 0.264x R= 0.191
1:1 Line
1
1
1
1
2468
PCB Body Burden as Estimated by Farley et al., 1999 (ug/g)
10
u -
T3 00
O 3
>• "S
7 ~
6 1
5 ~
4 ~
3 "
•a fr
o d
CO O
OJ I
U -a
Region 2-White Perch (Age Class 1-7 Years)
Tri+ PCBs Fish Body Burdens
y = -0.565 + 0.725x R = 0.944
1:1 Line
T~i—i~i—r~r~r~r~r
123456
PCB Body Burden as Estimated by Farley et al,, 2000
Sources: Farley et al., 1999 and USEPA, 2000
Figure 3-3
Comparison Between the White Perch Body Burdens Using the March, 1999 Model and
the Farley Model Run with HUDTOX Upper River Loads (1987-1997)
TAMS/MCA
-------
Region 2-Striped Bass (Agw Class 2-6 Years)
Tri+ PCBs Fish Body Burdens
c o.
u o.
^ 3 „
is
•o H
o Q
CQ 3
ca x
U T3 1
Q. c
y = -0.183 + 1.03x R = 0.958
T
T
1234
PCB Body Burden as Estimated by Farley et al., 1999 (ug/g)
Region 2-Striped Bass (Age Class 6-16 Years)
Tri+ PCBs Fish Body Burdens
y = -0.611 + 1.07x R =0.978
1234
PCB Body Burden as Estimated by Farley et al., 1999 (ug/g)
Sources: Farley et al., 1999 and USEPA, 2000
Figure 3-4
Comparison Between the Striped Bass Body Burdens Using the March, 1999 Model and the Farley Model
Run with HUDTOX Upper River Loads
TAMS/MCA
-------
o -c
C bo
o a
<•> s
Is
(rt O
5 s
0.06
0.05
0.04
0.03
0.02
0.01
160
Dissolved PCB Concentration April 1993
Model Apr-93
Apr-93 USEPA Data
140
120
I
100
80
River Mile
\
60
I
40
20
Dissolved PCB Concentration August 1993
O 3
0 o
•o -c
5 E
0.03
0.025
0.02
0.015
0.01
0.005
e— Model Aug-93
• Aug/Sep-93 USEPA Data
160 140 120 100 80 60 40 20
River Mile
Sources: Farley et a!., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-5
Comparison Between Field Data and Model Estimates for 1993 Dissolved PCB
Concentrations (Farley Model with HUDTOX Upper River Loads)
TAMS/MCA
-------
Water Field Data
RM 151.7 (Aug '93)
*S. n no °
solved Concentration (u§
0 0
b P b ?
o o — c
3 <-n — O) h
i i i , 1 , , , , 1 , , , , 1 , , , ,
5
3
solved Concentration (uj
o o
b P b <
o o — c
D V\ — Oi t
i i i i I i i i i 1 i i i i 1 i i 1 1
B
§
D RM 151.7 (Aug '93)
0 Model Results 153.5
/
/
/
' '!%?/ I 1
' ///// i i .,.,.,.,.„
Di ' Tri ' Tetra ' Penta ' Hexa
Water Field Data
RM 125 (Sep '93)
H
D RM 125 (Sep '93)
0 Model Results 133.5
/
/
; ^ i — L
Ul \J
g Di Tri Tetra Penta Hexa
Water Field Data
3 RM 77 (Sep '93)
">- rv no
solved Concentration (u§
o o
b P b ?
o o — c
D t-fl — Lft t-
1 1 1 ! 1 > I 1 1 ! 1 < 1 t 1 1 1 1 1
n
D RM 77 (Sep '9?)
0 Model Results 83.5
/
/
g " Di ' Tri ' Tetra ' Penta ' Hexa
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-6
Comparison of Model and Measured Homologue Pattern for 1993 Dissolved Phase PCB Concentrations
TAMS/MCA
-------
Surface Sediment PCB Cone. 1993
1.5
00
"So
o i
O '
0.5
• Aug-93 EPA Field Data
-a Model 0-2.5 cm 8/93
—X — Model 2.5-5 cm 8/9.1
•
•
Q •
N
\
I *-
>c
\
140 120 100 80 60 40 20 0
River Mile
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-7
Comparison of Model and Measured PCB Surface Sediment Concentration for 1993
TAMS/MCA
-------
Region 1 - White Perch
(Age Class 1-7 Years)
BO
"Sb
m
U
a.
3
CO
>-,
•a
O
CQ
60
50
40
30
20
10
— Model with HUDTOX Upper River Loads
NYSDEC Samples
USEPA and NOAA Samples
~ Mean
1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998
Year
Region 2 - White Perch
(Age Class 1-7 Years)
Model with HUDTOX Upper River Load
NYSDEC Samples
USEPA and NOAA Samples
~~i—|—i—i i—i—i—i—i | i i i—[—i—i—i—[ i i—i—r~r—i—i—|—i—i—i—i—i—i—i—]—i—i—r~
1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998
Year
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-8
Comparison Between Model and Measured White Perch Body Burdens
NYSDEC Fish Samples vs. Farley Model with HUDTOX Upper River Loads
TAMS/MCA
-------
Region 2 - Striped Bass
(Age Class 2-6 Years)
35 ~
"oo
"Sb 30 *
V)
S 25-
OL.
+ ;
(5 20-
C
U
•o
S 15 -
JO
>1
•o
£ 10 -
5 ~
-
0
C
)
• ?
/
r_
T
i 9
C
3
?/- '/" /& x* J
• » ' *^ 1/t
C
Model wilh HUDTOX Upper River Loads
O
O
Mean for NYSDEC Samples
Mean for USEPA and NOAA Samples
(Error bar shows the max and min)
)
-
o
•v —
1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1
" I
- I
) 0
-~1 — 1 —
T T
L T
AJ y^y-t
> <
r~ <
1 ' ' i ' ' ' i ' • '
1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998
Year
Region 2 - Striped Bass
(Age Class 6-16 Years)
"So 30
S 25
a.
o
oa
W .
5 '.
0 ~
5 1
0 ~.
5 "
0 ~
5 -
0
C
)
) ^
) (
)
(
c
?
)
<
c
i
Mode with HUDTOX Upper River Loads
0 Mean for NYSDEC Samples
O Mean for USEPA and NOAA Samples
(Error bar shows the max and min)
3
/ d
5
•II.
4.
y c
1 ' i
3
1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 199&
Year
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-9
Comparison Between Model and Measured Striped Bass Body Burdens
NYSDEC Fish Samples vs. Farley Model with HUDTOX Upper River Loads
TAMS/MCA
-------
2.5
0.5
3 1.5
u
cu
0.5
Region 1 RM 70-154
White Perch (Age Class 1-7 Years)
x FISHRAND RM 152
» FISHRAND RM 113
° FISHRAND RM 90
......... pariey Model Region 1-White Perch
*
D\
x FISHRAND RM 50
......... Regjon 2-White Perch
,t ., «—.^
*xxxxxxxxxxxxxxxx
"
1990 2000 2010 2020 2030 2040 2050 2060 2070
Year
Region 2 RM 14-70
White Perch (Age Class 1-7 Years)
1990 2000 2010 2020 2030 2040 2050 2060 2070
Year
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-10
Comparison of Model Estimates for White Perch Body Burdens
Farley Model with HUDTOX Upper River Loads vs. FISHRAND in Food Web Regions 1 and 2
TAMS/MCA
-------
Region 1-White Perch (Age Class 1-7 Years)
Annual Average Tri+ PCBs Fish Body Burdens
2.5
,.5
00
>%
•a
o
03
g '
<
Oi.
X
0.5
RM 152
y = 0.202 + 1.27x R = 0.862
1
I
0.5 I 1.5
Farley Model Body Burdens (ug/g)
Region 1-White Perch (Age Class 1-7 years)
Annual Average Tri+ PCBs Fish Body Burdens
1.5
CO
f '
CQ
Q
z
X
£ 0.5
u.
0
RM 113
2.5
T~ — i - 1 - 1
y = 0.199 + 0.988x R= 0.849
1 - 1 - 1 - 1 . | - 1 - 1 - 1 - 1 - 1 - 1 - 1 - 1 - r
0 0.5 1 1.5 2
Farley Model Body Burdens (ug/g)
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Notes: 1. Farley Model results represent the Farley Model with the upper river PCB loads from
HUDTOX.
2.The dashed line represents a 1 to 1 line.
Figure 3-11
Comparison of White Perch Body Burdens
(Farley Model vs. FISHRAND)
(page 1 of 2)
TAMS/MCA
-------
1.5
c
u
•o
00
Q
Z
I
2] 0.5
u.
oo
2? 1.5
•§ *
CO
D
Z
<
OS
X
22 0.5
u.
Region 1-White Perch (Age Class 1-7 Years)
Annual Average Tri+ PCBs Fish Body Burdens
RM 90
y = 0.113+ 0.825x R= 0.842
I I I I I i I I I I 1 I I I I
0.5 1 1.5
Farley Model Body Burdens (ug/g)
Region 2-White Perch (Age Class 1-7 Years)
Annual Average Tri+ PCBs Fish Body Burdens
RM 50
y = 0.0855+ 0.594x R= 0.916
0 0.5 1 1.5 2
Farley Model Body Burdens (ug/g)
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Notes: 1. Farley Model results represent the Farley Model with the upper river PCB loads from
HUDTOX.
2.The dashed line represents a 1 to 1 line.
Figure 3-11
Comparison of White Perch Body Burdens
(Farley Model vs. FISHRAND)
(page 2 of 2)
TAMS/MCA
-------
FIGURE 3-12a:
Comparison Between FISHRAND Results and Measurements at RM 152
Comparison to Data for Largemouth Bass at 152: wet
weight
a. .,
a. 15 -
JS
oc ,n
•3 10 -
0 -
«
T
^^H-^lJ
1986 1988 1990 1992
Year
i
i i
1994 1996 1998
Comparison to Data for White Perch at 152: wet
weight
oc
/J -
1.20
£2°-
•+^
!§>
^ 1U -
•M
0 -
<
^
' j
1
t^ J
1986 1988 1990 1992
Year
T T
\ \ 1«T
L 1 ±
1994 1996 1998
Comparison to Data for Largemouth Bass at 152: lipid-
normalized
i_inn
^ 1 200
3 1000
Stf snn
o a
Z ft Ann
S 400 ¥ • ' • — 1
.§• * 1 , T
[j °00 T B I
0 -J 1 r 1 1
1986 1989 1992 1995 1998
Year
Comparison to Data for White Perch at 152: lipid-
normalized
ROO
a
o o, 400
z T I
2 200 i . i -i — T*r
S L, » &
1 : — — - . ( — ,
1986 1988 1990 1992 1994 1996 1998
Year
Legend: ^
Median with 95% UCL and 95% LCL
FISHRAND
TAMS/MCA
-------
FIGURE 3-12a: Comparison Between FISHRAND Results and Measurements at RM 152
Comparison to Data for Brown Bullhead at 152: wet
weight
6
|D 6'
t& 4 ~
0 -
19
V <
^
•
«
«
L_ I «
» «
86 1988 1990 1992 1994
Year
Comparison to Data for Yellow Perch
weight
-
JS
•3 4 -
£
•
1
J
H
[ T
•
L ,
4
«
i
•
1996 1998
at
152: wet
t
I
-±-
»
i i i i
1986 1988 1990 1992 1994
Year
1996 1998
Comparison to Data for Brown Bullhead at 152: lipid-
normalized
"O ™n
_N
S_ 700 -
g zuu
o g.
100 -
.5* n -
•
K
- ,
1986 1988 1990 1992
Year
Comparison to Data for Yellow Perch
normalized
"8
.N 800 -
"I 600 -
!•* 5? Af\r\
o a 40°
y Q,
^ 200-
T3
'o, o -
3 19
T
1 1
T 1
— — i—
1 , , T 1 a
86 1988 1990 1992
Year
•
T
^
1994 1996 1998
at 152: lipid-
!f
T
U
— , — .L-, — ,
1994 1996 1998
Legend:
Median with 95% UCL and 95% LCL
FISHRAND
TAMS/MCA
-------
FIGURE 3-12b: Comparison Between FISHRAND Results and Measurements at RM 113
Comparison to Data for Largemouth Bass at 113: wet
weight
20
1
0. 15 -
Ml IQ
°3
*^ c
o>
^
"^ 0-1
»
^K^I { T
-H-^-J—
1986 1988 1990 1992 1994
Year
1996
1998
Comparison to Data for White Perch at 113: wet
weight
CL R
a
"So
2£ 2 -
«4
^ 0
^ ^ I
»
1986 1988 1990 1992 1994 1996 1998
Year
Comparison to Data for Largemouth Bass at 113:
lipid-normalized
e
a son
a
•s 40°-
N
'•5 30° -
| 200 -
o
Z 100 -
•a
'& 0 -
f| T
J J
"\.
~^^^^
^^^^^^^»_^ •• ^* ^f
J" ^—»«^fc^_ Hf
1-1 1986 1988 1990 1992 1994 1996 1998
Year
Comparison to Data for White Perch at 113: lipid-
normalized
.§
CO
C C i *\n
0 0.
Z °" 100 -
* P
T
^ ^ 4
1986 1988 1990 -1992 1994 1996 1998
Year
Legend: • Median with 95% UCL and 95% LCL
FISHRAND
TAMS/MCA
-------
FIGURE 3-12b: Comparison Between FISHRAND Results and Measurements at RM 113
Comparison to Data for Yellow Perch at 113: wet
weight
a. , g
o, i.j
*rf
r* 1
•a '
I? o s
£ 0.5-
_ ^ Y
^ *-?!
x
ftn
_N
"a 1 SO -
1 |
O O< 100 -
z *• 1UU
? 50-
o. J"
3 0-
^ I
^~?
X
1986 1988 1990 1992 1994 1996 1998
Year
Legend: • Median with 95% UCL and 95% LCL
FISHRAND
TAMS/MCA
-------
FIGURE 3-12c; Comparison Between FISHRAND Results and Measurements of Pumpkinseed
Comparison to Data for Pumpkinseed at RM 60: wet
weight
.5 *^~-
E 3 T T\T
L; * } "{
s 2 5 1 *\
* ,'
V
^ 1 ~ •
05
0_, ',_,.
_
T
" v. f
"^'^ i
• ' ^r~~~~-^ I
I —
^ •
1986 1988 1990 1992 1994 1996 1998
Year
Comparison to Data for Pumpkinseed at RM 142 -
152: wet weight
O,
M ^
> 3 I
^^ ~^
0 -) 1
fi
1
z
1986 1988 1990 1992 1994 1996 1998
Year
Comparison to Data for Pumpkinseed at RM 60:
lipid-normalized
140
§20 -
a
•Son _
o60 -
z
'a
n -
T
^ I I
s ^ t i
f <^
^- — .
4
I J
1986 1988 1990 1992
Year
Lipid Normalized ppm
— — N>
en O Ui O
O O O O O
•
T
"X ~*
1994 1996
1998
Comparison to Data for Pumpkinseed at RM 142 •
152: lipid-normalized
•^
x. . >
m - - •
986 1988 1990 1992
Year
1994 . 1996 1998
Legend:
• Median with 95% UCL and 95% LCL
FISHRAND
TAMS/MCA
-------
FIGURE 3-12d: Comparison Between FISHRAND Results and Measurements of Spottail Shiner
Comparison to Data for Spottail Shiner Wet Weight:
1993 US EPA Dataset
7
s7
S-fi
B.O
S -5
M5 "
'« 4 .
£1
•M 1
4> •>
£2
1
•
4
T
>
•
»
D* B *i ]
° In
0 50 100
River Mile
K
»
150 200
Comparison to Data for Spottail Shiner Lipid
Normalized: 1993 US EPA Dataset
E
Q- inn
cu Juu
"S 9SO -
.H
"3 900
O 1 "50 -
rf— 1
T3 inn -
.S"
j
-------
HUDTOX Federal Dam Flux (kg/d)
1980
2000
2020
0.03
8 ^
ca ao
JS 9
EL. •—'
4) »
»• u
s .±
*^ L-
0.025 -
0.02 ~
0.015 -;
0.01 -
0.005 -
Region 1
1980
2000
2020
% J
fl* """^
"S «
> u
0.03
0.025 -
0.02 -:
0.015
0.01 -
0.005 -
0
Region 2
1980
2000
2020
2040
2060
2080
x Dissolved Phase
2040
2060
2080
Dissolved Phase
2040
2060
2080
Year
Souces: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-13
Comparison Among the HUDTOX Upper River Load and Farley Model Estimates of
Dissolved Water Column Concentrations in Food Web Regions 1. and 2
(1987-2067)
TAMS/MCA
-------
x HUDTOX Federal Dam Flux (kg/d)
1980
2000
2020
2040
2060
2080
0.6
en
a S3
a U
o a,
'"£ .±
ca u
0.5 -
0.4
0.3 -
0.2 -
0.1
Water Column Particulate Phase
1980
2000
2020
2040
2060
2080
Whole Water Total PCBs
1980
2000
2020 2040
Year
2060
2080
Souces: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-14
Comparison Among the HUDTOX Upper River Load and Farley Model Estimates of
Particulate and Whole Water Column Concentrations in Food Web Region 1
(1987-2067)
TAMS/MCA
-------
5 -
L.
H
5
•* 4
•—• ^
I 3
2 -
1 -
HUDTOX Federal Dam Flux (kg/d)
* X
1980
2000
2020
2040
2060
2080
'3
"Si
S
"*—*
V)
03
U
ON
+
'u
H
"M
"oi
£
(A
03
U
0.
+
"u
H
^ -
-
1.5 -
.
.
1 -
.
0.5 -
-
0
x
Region 1 x River Mile 118.5
X
X
X
x
xXx
>0^0^Xx>0<>oooooox>o^^
1 I ' ' ' 1 ' l ' 1 ' ' ' 1 ' ' ( !
1980 2000 2020 2040 2060 2080
1.2
1 -
.
0.8 -
.
•
0.6 -
.
-
0.4 -
.
-
0.2 -
r\ —
Region 2 x River Mile 48.5
x
x
X
X
X
V
Xxxx
^^XXSGoc*°«ooooooc>ooo
-------
" a
le
i 2
5 -
4 •:
3 -
2
0
1980
HUDTOX Federal Dam Flux (kg/d)
2000
2020
2040
2060
2080
B •
-------
£ j?
5 Q
6
5
4
3 -
2 -
1 -
* HUDTOX Federal Dam Flux (kg/d)
* X
1980
2000
2020
2040
2060
2080
e -
£ es
3 («
a «
2*
4 -
3 -
i -
0
]
4
3.5
C M 3 "
i -s2.5
•"s ^
13 B«
,2 + 1.5 -:
0.5 -:
0
Striped Bass Age Class 2-6 Years
1980
2000
2020
2040
2060
2080
Striped Bass Age Class 6-16 Years
1980
2000
2020
2040
2060
2080
Year
Souces: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Figure 3-17
Comparison Among the HUDTOX Upper River Load and Farley Model Estimates
Striped Bass Body Burdens in Food Web Regions 1 and 2
(1987-2067)
TAMS/MCA
-------
RM 152
-W)
oo
u
Q.
H
c
•o
3
CO
o
09
-2?
3
CO
c;
OL.
C
U
•o
CO
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o
m
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CO
u
a.
j-J .
3 -_
2.5 -_
2 -_
1.5 -I
I _I
0.5 ~
0"
19
2.5 -|
1.5 -;
1 -_
0.5 -_
0~
19
0£ _
0.5 -_
~ ' Mean Value
"— ~ • •*• Upper 95% Confidence Limit
—
90 2000 2010 2020 2030 2040 2050 2060 20
RM 113
T • Mean Value
_ - Upper 95% Confidence Limit
"^^— — ^"'-^^ _ ^—_
90 2000 2010 2020 2030 2040 2050 2060 20
RM 90
•
^ * Mean Value
TT_ ~ Upper 95% Confidence Limit
T
70
70
o
03
CO
U
a.
09
• -o
o
CO
0.5 -_
0.4 -_
0.3 -.
0.2 -f
0.1 i
0 -
19
0.6 -n
0.5 -_
0.4 -:
0.3 -i
0.2 -_
O.I ^
t\ ~
~r
^^
90 2000 2010 2020
T
T
T
T
^TT^
Mean Value
Upper 95% Confidence Limit
— -_ __
•-^— ~-*^.
2030 2040 2050 2060 20
RM 50
Mean Value
Upper 95% Confidence Limit
"— ~
"•'"••'• •'***• »-y.
70
1990
2000 2010 2020 2030 2040 2050 2060 2070
Year
Figure 3-18
Forecasts of Large Mouth Bass Body Burdens from FISHRAND
TAMS/MCA
-------
r
RM 152
00
u
Qu
3
CO
>\
•o
o
00
•s?
Ml
oa
U
o.
o
OQ
CO
U
a.
o
cc
03
U
a.
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o
CO
?
1.5
1
0.5
7
1.5
1
0.5
1.2
1
0.8
0.6
OA
0.2
1
0.6
0.4
0.2
n
-
-
:
-_
-_
i
-
19
19
J
"I
-
-
~
1
:
— 1
H
19
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j
„
-
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,
_
-
-
,
TW
-W
'""-••^
i i i i 1 i i i > 1 i i i i I .
90 2000 2010 2020
^
"V^
""^
""""""'"'--w
90 2000 2010 2020
•••
•%.
"^-^T
T 1 , 1 | i | 1 1 | . , , 1 |
90 2000 2010 2020
T
T,
~~^
TT-»^
Mean Value
Upper 95% Confidence Limit
•**-•-
~~~~~*r '^^~^_
1 1 1 j . 1 11 | 1 1 1 ! , 1 1 1 1
2030 2040 2050 2
RM 113
Mean Value
Upper 95% Confidence Limit
2030 2040 2050
RM 90
Mean Value
~ Upper 95% Confidence Limit
"~^-— ~
1 . . | > 1 . 1 | . 1 t 1 | , 1 !
2030 2040 2050
RM 50
Mean Value
Upper 95% Confidence Limit
"~~~~~~~— - -—_„„_„
-v"-~--.^.
060 20
— — — ~~_
2060 2C
[ , i , i
2060 2C
— _
70
70
70
1990
2000
2010
2020 2030 2040 2050 2060
Year
2070
Figure 3-19
Forecasts of White Perch Body Burdens from FISHRAND
TAMS/MCA
-------
RM 152
1
Cfl
ffl
u
a.
+
H
c
u
13
3
03
>»
•a
o
CD
Dfi
"Sb
^
(/)
ca
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a.
+
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c
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3
CQ
>*
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CO
1
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a.
+
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c
u
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3
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>,
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CO
1
Wi
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U
a.
+
£
c
u
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3
CD
>v
T3
O
CQ
u.o .
0.7 -i
0.6 -E
0.5 -j
0.4 4
0.3 -j
0.2 -^
0.1 -E
T • Mean Value
"^ - . Upper 95% Confidence Limit
"•VV-.-
TT-^*VT
"•""""-»^"1V -»•*" "^-r-
1990 2000 2010 2020 2030 2040 2050 2060 2070
RM 113
Of,
0.5 -_
0.3 -,
0.2 -_
0.1 -:
T
TTT • Mean Value
T.. - Upper 95% Confidence Limit
T^v
ir___
TVT.
"*^T-»-
' , l , | 1 1 I 1 | 1 1 I I | . i : : 1 1 . 1 I | 1 1 I i | I •. l 1 , I : i .
1990 2000 2010 2020 2030 2040 2050 2060 2070
RM 90
0.4 -_
0.3 -;
0.2 -jj
0.1 -_
0~"
TT • Mean Value
TTT ~ Upper 95% Confidence Limit
T-»-T
TTT-*^TT
1990 2000 2010 2020 2030 2040 2050 2060 2070
RM 50
°'35 "3 V^ • Mean Value
0.3 -3 'r - Upper 95% Confidence Limit
0.25 -1
0.2-1 ^^
0.15-3 ^"^-^.^^
0.05 -|
1990 2000 2010 2020 2030 2040 2050 2060 2070
Year
Figure 3-20
Forecasts of Yellow Perch Body Burdens from FISHRAND
TAMS/MCA
-------
I
(A
00
u
a.
+
•c
H
3
03
>.
•o
o
CO
CO
U
a.
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>.
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00
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t/1
no
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a.
3
oa
o
oo
cc
u
Q.
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CO
1.5 -
1 -
0.5 -
0.2 -j
!
RM 152
Mean Value
Upper 95% Confidence Limit
1990 2000 2010 2020 2030 2040 2050 2060 2070
RM 113
0.5
1
06
0.4
0.2
0 8
0.6
0.4
-
'-
'-
-_
19
_
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.
:
_
•
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19
_
:
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~
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-
_
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90 2000 2010 2020
—
•T_
'~-_
•--_
* •—— _
••-v
90 2000 2010 2020
T_
'TT_
'---_
" "^TV
•T-c
Mean Value
- Upper 95% Confidence Limit
~*~~V-_—
2030 2040 2050
RM 90
Mean Value
- Upper 95% Confidence Limit
""""••*--* ~™_^
^^r^Vl^-^-,
2030 2040 2050
RM SO
Mean Value
- Upper 95% Confidence Limit
•"••"^w.
— _
2060 20
"*'• • •kr^-»^-w.
2060 20
70
70
I I I ' I I I I ' I I ' I ! I I ' I ' ' I I ' I I I ' I ' I I ' ' ' ' i I ' '
1990 2000 2010 2020 2030 2040 2050 2060 2070
Year
Figure 3-21
Forecasts of Brown Bullhead Body Burdens from FISHRAND
TAMS/MCA
-------
RM 152
09
U
O.
CO
>«
•o
o
CQ
CO
U
o.
3
ffl
O
CO
0.8 -
0.6 -
-
0.4 -
0.2 ^
• _ • Mean Value
~T - ~ Upper 95% Confidence Limit
T
TT ~ ~_~_
T -r T
1990 2000 2010 2020 2030 2040 2050 2060 2070
RM 113
07
0.6 -j
0.5 -E
0.4 -E
-
0.3 -_
0.2 -E
0.1 -i
0 -
—
~ - • Mean Value
' *_ - Upper 95% Confidence Limit
•_-T_
* _ _
" 'v
I^TW^'*V«---V'^V«V~-'*' TrV^rr^..,^-
—- ^ "-_
1990 2000 2010 2020 2030 2040 2050 2060 2070
RM 90
A f.
0.5 ^
—
- • Mean Value
j T- - Upper 95% Confidence Limit
ft 4 1 '-
UJ
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3
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o.
+
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c
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3
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3
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+
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e
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3
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>,
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0.5
0.4
0.3
0.2
0.1
0
0.6
0.5
0.4
0.3
0.2
0.1
n
-i ~
_i •
j '"^T
j "_
H *—
j '""TT-TT
3
|
^
1990 2000 2010 2020
_i
-; ~
"H -
--, •
H T
H 'T
j •_
J ^^^V,
-i
1
H
Mean Value
~ Upper 95% Confidence Limit
——-——_
2030 2040 2050 2060 2070
RM 50
Mean Value
- Upper 95% Confidence Limit
1990
2000 2010 2020 2030 2040 2050 2060 2070
Year
Figure 3-22
Forecasts of Pumpkinseed Body Burdens from FISHRAND
TAMS/MCA
-------
RM 152
-s?
Vi
CQ
O.
H
c
u
•a
CQ
T3
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CQ
-00
3
l/i
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c
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3
CQ
U
CL.
H
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TD
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0-)
0.15 T
O.I -_
0.05 -H
0"
19
07«
0.2 -j
0.15 ^
0.1 -_
0.05 ^
19
00
015 j
-
0.1 J
•
0.05 -
"
•
0 :
u n
19
A I £
0.14 T
0.12 -i
0.1 J
0.08 -^
0.06 -I
0.04 -f
0.02 -
-
0"
19
Mean Value
*_* ~ Upper 95% Confidence Limit
-*-~ ~
T^. T T
^ T W ^v»- ^T- " 'V-*V^r. ^ ^ ^
90 2000 2010 2020 2030 2040 2050 2060 20
RM 113
_ • Mean Value
'__v - Upper 95% Confidence Limit
"~
~~"*v
""— -r ^ ^^rr
90 2000 2010 2020 2030 2040 2050 2060 2C
RM 90
- • Mean Value
~_ - Upper 95% Confidence Limit
V
•5-
90 2000 2010 2020 2030 2040 2050 2060 20
RM 50
~
v • Mean Value
'-_ ~ Upper 95% Confidence Limit
T
' ' ' ' I ' ' ' ' 1 ' ' ' ' 1 . . : | i 1 . 1 1 | 1 1 1 1 |
90 2000 2010 2020 2030 2040 2050 2060 20
Year
70
70
70
70
Figure 3-23
Forecasts of Spottail Shiner Body Burdens from FISHRAND
TAMS/MCA
-------
THIS PAGE LEFT INTENTIONALLY BLANK
-------
Appendix A
-------
Appendix A
CONVERSION FROM TRI+ PCB LOADS TO DICHLORO THROUGH
HEXACHLORO HOMOLOGUE LOADS AT THE FEDERAL DAM
A.I Introduction
The fate and transport and bioaccumulation models of PCBs described in Farley etal.(\999)
(the Farley model) for the mid to lower regions of the Hudson River will be used to predict fish body
burdens for the Mid-Hudson Human Health Risk Assessment and the ERA Addendum. As originally
constructed, the Farley model relied on load estimates at Thompson Island (TI) Dam to directly
represent the loads delivered to the Lower Hudson. Future loads were assumed to be identical to that
measured at TI Dam in 1997. This assumption does not account for load variations between TI Dam
and Waterford nor the anticipated Upper Hudson load decline over time. Indeed, the forecast
prepared Farley et al. (1999) extended only to 2002. For the risk assessment requirements of the
Phase 2 investigation, a forecast beyond 2002 is required and so the Upper Hudson loads must be
adjusted to account for an expected decline in PCBs with time. Additionally, load estimates based
on TI Dam measurements do not account for the influences of the intervening 35 miles of river
between TI Dam and the Federal Dam at Troy, NY.
The preparation of the Upper Hudson model 70 year forecast also included estimates of
Upper Hudson loads at Waterford. Results from the Upper Hudson River model, HUDTOX,
developed by Limno-Tech, Inc. (LTI) will be used for the PCB loads coming over the Federal Dam
at Troy, NY. The HUDTOX model accounts anticipated declines in water column loads over time
as well as the riverine influences on these loads between TI Dam and Troy.
Dichloro through hexachloro PCB homologues are the state variables in the Farley model of
the Lower Hudson River but HUDTOX simulates total PCB and the sum of trichloro through
decachloro homologues (Tri+) for the Upper Hudson. Thus, a means of converting the data from
total or Tri+ PCBs to individual dichloro through hexachloro homologues is required.
A conversion algorithm was developed based on the available data. An extensive number of
samples are available from the TI Dam station, but relatively few are available from the lower station
at Waterford, NY and even fewer from Troy, NY. In this analysis, homologue patterns at the TI Dam
are compared to the patterns at Waterford to determine if a correction can be applied to the TI Dam
data so as to yield conditions at Waterford. Mean homologue mass fractions are calculated using data
collected at the TI Dam station and grouped to determine if the patterns should be adjusted for
season or flow rate. Through this effort, a means of conversion of the HUDTOX Tri+ sum is
developed. The conversion yields a daily load estimate of each of the homologue groups from
dichloro to hexachlorobiphenyl. Referenced tables and figures relating to this analysis follow the
text.
A-l TAMS/MCA
-------
A.2 Data Preparation
The data used for this memo are whole water data from USEPA, 1993 and General Electric
(QEA, 1999) from Waterford and TI Dam stations. The USEPA data are available in the Hudson
River Database, Release 4.1 (USEPA, 1998a). The GE data is from the March 1999 update to the
GE database. There are two important differences between the data sets, (1) homologue data from
the two data sets do not represent the same exact suite of congeners and (2) the analytical methods
are somewhat different. The USEPA homologue data is based on 126 congeners which are
individually measured and calibrated. The GE homologue mass fractions are taken directly from the
GE database file from March 1999 and are based on a smaller set of congeners and are calibrated to
Aroclor standards. Some congeners are unique to each data set.
In compiling the sample results for interpretation, field duplicates collected by GE are not
used. For the GE data, there are numerous instances of more than one sample per day per station,
obtained for Quality Assurance purposes. The first sample listed per day per station in the GE
database is used since the duplicate samples are equivalent. USEPA duplicates from the Phase 2
database were combined and averaged in the preparation of the database and were used as listed in
the database tables.
Two USEPA samples (transect 2) were excluded because of data quality issues. Eight GE
samples were excluded because the sum of the trichloro to hexachloro homologues was less 97
percent than the sum of the trichloro to decachloro homologues (Tri+). These samples were excluded
because it was deemed unlikely that estimates of the true value of the mass percent of heptachloro
through decachloro homologues would exceed a few percent of the Tri+ sum.
Samples are grouped by flow and season in several instances. High flow is defined as greater
than or equal to 4000 cubic feet per second (cfs); low flow is less than 4000 cfs as measured at the
USGS Fort Edward station. For the Waterford samples, flow data from the USGS Waterford station
was used in preference over the Fort Edward data to determine the flow condition when available.
The basis for defining the flowswith respect to 4000 cfs is discussed in the DEIR Responsiveness
Summary (USEPA, 1998b). The seasons are defined as follows: spring, 3/16-5/15; summer, 5/16-
10/31; and fall-winter, 11/1-3/15.
A.3 Dichloro Homologue
Optimally, to develop ratios to apply to the HUDTOX Tri-i- sum, a long-term record of the
homologue composition at Waterford is required. In this manner, a ratio could be developed for the
existing period of record, enabling an examination of the results during the 1987-1997 calibration
period. Similarly, the ratio could then be used to develop forecasts of Lower Hudson conditions.
Unfortunately, this information does not exist but a long-term record does exist at TI Dam. From this
information, an estimate of the homologue to Tri+ ratio at TI Dam could be obtained. This ratio is
an estimate of the average loading condition at TI Dam. However, this analysis does not yield the
homologue to Tri+ ratio at Waterford. Thus for each congener, the ratio at the TI Dam was examined
relative to Waterford for the period where data were available. This second factor represents the
A-2 TAMS/MCA
-------
effects of transport between TI Dam and Waterford. The first ratio would be expected to change
with changes in loads originating above TI Dam, as might arise from remediation at the GE facilities.
The second factor represents the impacts of water column transport and associated geochemical
processes occurring between TI Dam and Waterford. This factor would not be expected to change
with time because it is the cumulative result of geochemical processes (e.g., gas exchange, sediment-
water exchange, aerobic degradation) which should remain the same with time. This factor would
be expected to vary seasonally, however, because temperature and flow rate changes will affect the
rates of the various geochemical processes.
To determine the ratio of the dichloro homologue to the HUDTOX Tri+ load (di/Tri+) at
Waterford, the following steps were taken:
Comparison of the Waterford di/Tri+ ratio between the TI Dam and Waterford
stations. Homologue data for Waterford are limited, but are available for the TI Dam
from 1990-1998 using the GE data. A correction factor to relate these stations on
either a seasonal or flow basis is needed in order to use the long record of data at the
TI Dam. This factor represents the TI Dam-to-Waterford transport factor described
above.
Examination of the di/Tri+ PCBs ratio overtime to determined if the ratio has
changed substantially overtime. Data were grouped to determine the mean values of
the di/Tri+ ratio by period, season and flow. This represents the loading ratio
described above.
The data set to establish the TI Dam to Waterford ratio is limited. In particular, the 1991 GE
samples at TI Dam and Waterford were not timed to capture the same parcel of water as it traveled
from the TI Dam to Waterford. Thus, these samples do not directly track the changes to the water
column loads originating from the geochemical processes which occur enroute. Given the relatively
low number of samples collected at the two stations that year, there are not enough samples to
develop an average ratio to accurately represent the effects of the geochemical processes as a
function of flow and season. Table A-l lists the calculated time for each flow rate at Fort Edward
for water to travel from TI Dam to Waterford and the hours between sampling at these stations. None
of the travel times are similar to the sampling times, indicating that the sampling were not timed to
capture the same parcel of data. Because of this aspect of the GE sampling method, only the USEPA
Phase 2 samples, which were purposely timed to capture the same parcel of water, will be used to
compare TI Dam to Waterford. As discussed below, all of the GE and Phase 2 samples at TI Dam
will be used to examine the temporal changes in homologue percentages.
Figures A-l through A-5 show the di/Tri+ ratio (expressed as a percentage of the Tri+
concentration) grouped by station, season and flow rate for the USEPA data only. Figure A-l shows
a statistically significant difference in the di/Tri+ ratio at the two stations for all Phase 2 results. The
subsequent figures show how this difference correlates with flow and season. The grouping by flow
shows a significant difference of the means during low flow (Figure A-4) and no difference during
the high flow (Figure A-5). This suggests that during the typically low flow conditions of the warmer
A-3 TAMS/MCA
-------
months, there is time for the PCBs in the water column to interact with the sediments, altering the
homologue pattern. During the periods of high flow, the PCBs at TI Dam are translated to Waterfbrd
nearly unchanged. Flow was chosen as the main separation variable for this ratio because it yielded
the greatest separation among groups at low flow and no separation at high flow, as might be
expected.
To determine the loading ratio at TI Dam (the first factor discussed above), the di/Tri+ versus
time at the TI Dam and Waterford stations is shown in Figures A-6 and A-7, respectively. These
figures display both the USEPA and the GE data over the period 1991 to 1998. A change in the
pattern of the di/Tri+ ratio is evident starting in mid-1996 in the TI Dam results. (No data are
available for Waterford post-1993.) The range of di/Tri+ ratios is greater and the average value is
higher at the TI Dam after 1995. This is coincident with a drop in total PCB concentration as shown
in Figure A-8. This figure shows the total PCB concentration versus time at the TI Dam. The
decrease in concentration in 1996 and later is attributed to the 1993-1995 remediation efforts above
Rogers Island, which substantially reduced the Tri+ loading to the Hudson River. Little evidence of
subsequent decline in loads is evident post-1995. As a result of the GE remedial efforts, the
importance of the sediments to the water column loads was greatly increased while the sporadic,
large-scale releases above Rogers Island largely disappear. Based on these results, the data from
1996-1998 should be used to predict future conditions. Figure A-9 shows the TI Dam di/Tri+ ratio
grouped by years 1991-1995 and 1996-1998. The difference in means is clearly significant. Figures
A-10 through A-13 show the same data further grouped by season and flow. Of these, the best
separation of the means is seen using flow.
Table A-2 summarizes the basis for conversion for the di PCB homologue as well as the
other homologue groups, which are discussed below. The table is separated into the calibration perio,
(1987-1998) and the forecast period (1999 and later). The mean di/Tri+ ratios at the TI Dam are from
Figures A-12 and A-13. For low flow, the correction from the TI Dam to Waterford is 0.52 which
is the ratio of the means 45.5883/86.8350 given in Figure A-4. The correction during the high flow
is small (1.04) because, as shown in Figure A-5, there is no significant difference between the means.
Note that for the dichloro ratio only, the ratios developed here are applied throughout both the
calibration and forecast periods, as appropriate. For the period prior to 1991 where no congener data
exist, the ratios measured in 1991 are applied. In the forecast calculations, the ratios developed for
the period 1996-1998 at TI Dam are applied along with the TI Dam to Waterford transport
correction.
A.4 Trichloro through Hexachloro Homologues
Ratios for the trichloro to hexachloro homologues were developed in a fashion similar to that
used for the dichloro homologue. These ratios has the additional constraint that they must sum to 100
percent, representing the entire Tri+ load. The fractions of trichloro through hexachloro homologues
at Waterford are determined by two factors, as follows:
TI Dam-to-Waterford Correction: Comparison of the fractions of trichloro through
hexachloro homologues in Tri+ PCBs at Waterford to TI Dam. Because the number
A-4 TAMS/MCA
-------
of samples is limited at Waterford, the extensive data from the TI Dam can be used
with correction for the Waterford station. As was discussed in the DEIR (USEPA,
1997) and the LRC Responsiveness Summary (USEPA, 1999), the trichloro through
hexachloro homologues appear to be translated from the TI Dam to lower river
stations with little modification.
TI Dam-Loading Factor: Development of this factor was based on two steps:
• Principal components analysis to determine if the distribution of trichloro
through hexachloro homologues in Tri+ PCBs is significantly affected by
season, flow, etc.
• Examination of the TI Dam Tri+ PCB ratios to determined if the ratios have
changed substantially overtime. Data were grouped to determine the mean
values of the ratios by period, season and flow.
As in the examination of TI Dam-to-Waterford transport for the di homologue, the GE
samples were not timed to capture the same parcel of data (Table A-l). Thus, these samples were
excluded from the determination of the TI Dam-to-Waterford correction for the heavier homologues
as well.. Figures A-14 through A-21 show the USEPA data exclusively, grouped by season. The one
fall-winter sample is grouped with the spring data. A significant difference in the means is only
evident during the summer for the trichloro through pentachloro homologues. Notably, the fraction
of tri/Tri+ decreases from TI Dam to Waterford while the remaining heavier homologues all increase
relative to the TI Dam ratio. Mean ratios at TI Dam and Waterford are quite close during the
remainder of the year. Nonetheless, the ratios developed from this analysis were applied to the data
in order to represent the best estimate of the relative changes between TI Dam and Waterford. Use
of the entire suite of ratios also serves to maintain conservation of mass (i.e., one ratio cannot
decrease without corresponding increases in the remaining ratios). These are summarized in Table
A-2.
In the examination of the temporal variation of the homologue to Tri+ ratios, a principal
components analysis was undertaken. In this examination the mass fractions of trichloro through
hexachloro homologues were used as the primary variables. A principal components analysis using
the GE and USEPA data is shown in Figure A-22. The results of the analysis are presented in five
different ways, with indicators to denote sampling agency, season, flow, station and year (1991-1995
and 1996-1998). No significant separations among the data are seen using these groupings.
Although no evidence of the temporal variation was seen in the PCA analysis described
above, an examination of the trends of the various ratios with time suggests the occurrence of a
temporal change. A map of the GE TI Dam stations is shown in Figure A-23 with the coordinates
provided in the GE database. Data from these stations along with the USEPA Phase 2 results are
plotted against time as the mass fraction of trichloro through hexachloro homologues versus Tri+
PCBs in Figures A-24 through A-27. As with the di homologue fraction, a difference in the pattern
is seen beginning in 1996. This change in pattern (particularly evident in the tri/Tri+ and penta/Tri+
A-5 TAMS/MCA
-------
ratios) coincides with the decline in total PCB concentration seen in Figure A-8. Based on these
results, future conditions were predicted using the 1996 through 1998 data.
The TI Dam from 1996-1998 are grouped by season for each homologue of concern in
Figures A-28 through A-31. The data are grouped by flow in Figures A-32 through A-35. The best
separation (greatest distance between the Tukey-Kramer circles) of the means is given by grouping
on season. It should be noted, however, that the ratio variations among these groups are relatively
small, typically only a few percent of the total Tri+ mixture. The importance of these variations
increases as the fraction of the homologue decreases, as would be expected. Thus, the summer to
spring variation of 8 percent (54 - 46 percent) in the trichloro homologue percentage represents about
15 percent of the total trichloro mass. However, the 2.4 percent summer-to-spring change in the
hexachloro homologue ratio represents nearly a 50 percent decline in the ratio from spring to
summer. These results should be compared to the dichloro homologue results which show large
changes on both absolute and relative scales.
The final conversion factors for the trichloro through hexachloro homologues are shown in
Table A-2. The mean mass percent of trichloro to hexachloro homologues using the 1996-1998 TI
Dam data was obtained from Figures 29 through 32. The correction for transport from TI Dam to
Waterford is given as well. Before applying these two factors, a further step must be taken in order
to conserve mass in the calculation. This is done by assuming that the concentration of a homologue
at Waterford in 1996-98 is equal to the concentration at Fort Edward in 1996-98 times the ratio of
the 1993 concentrations observed at Waterford and Fort Edward. The ratio of concentrations between
Waterford and Fort Edward is assumed constant rather than the ratio of the mass percents. The
proper way to calculate the mean mass percent at Waterford in 1996-98 for homologue i is:
P(FE)i • K.
P(WATR). =
where:
P(WATR) is the mass percent relative to Tri+ at Waterford;
P(FE) is the mass percent at Fort Edward; and,
K is the ratio of the 1993 mass percent at Waterford to the 1993 mass percent at Fort Edward.
In this manner, the sum of the tri/Tri+ to hexa/Tri+ ratios will sum to 100 percent in all instances,
as it should. Without this correction, this last condition is not met.
A.5 Data Conversion Summary
Table A-2 provides a summary of the data conversion for all periods and flows. The
distributions will be applied to the Federal Dam loads generated by the May 1999 HUDTOX model
(both the calibration and forecast periods). For the period 1987-1990 where no homologue data are
available, the dichloro through hexachloro distribution for 1991 will be applied without correction.
A-6 TAMS/MCA
-------
Although PCB releases from the Bakers Falls area may have occurred, this is not of concern because
the 1987-1990 period will not be used in the ERA Addendum and Mid-Hudson HHRA and this
period does not weigh strongly in the calibration. For the dichloro homologue, the mean mass
percent of Tri+ PCBs calculated from the 1991-1995 TI Dam samples will be used for the Waterford
distribution during high flow with the TI Dam to Waterford correction. Starting in 1996 and
continuing for the remaining period of time to be modeled, the 1996-1998 mean mass percent of
di/Tri+ at TI Dam will be used.
For the trichloro through hexachloro homologues during 1991-1998, the distribution defined
by the mass percent of Tri+ PCBs from GE samples at the TI Dam was applied. For future
predictions of the trichloro through hexachloro homologues, the mean distribution defined by the
1996-1998 data at the TI Dam was used. Each of the mass percent values were corrected for the
measured difference between the TI Dam and Waterford to account for transport losses and then
adjusted to conserve mass.
A-7 TAMS/MCA
-------
A.6 References
Farley K.J., R.V Thomman, T.F. Cooney, D.R. Damiani, and J. R. Wand. 1999. An Integrated Model
of Organic Chemical Fate and Bioaccumulation in the Hudson River Estuary. Prepared for the
Hudson River Foundation. Manhattan College, Riverdale, NY.
Quantitative Environmental Analysis, LLC.(QEA) 1999. Database transmitted 3/1/99. Personal
communication from QEA to ED Garvey. March 2, 1999.
USEPA, 1997. Phase 2 Report, Further Site Characterization and Analysis, Volume 2C- Data
Evaluation and Interpretation Report, Hudson River PCBs Reassessment RI/FS. Prepared by
T AMS/Gradient/Cadmus.
USEPA, 1998a. Datbase for the Hudson River PCB Reassessment RI/FS. Release 4.1 (Compack
Disk) Prepared for USEPA, Region n and the US Army Corps of Engineers, Kansas City District,
Prepared by TAMS consultants, Inc, August, 1998.
USEPA, 1998b. Responsiveness Summary For Volume 2A: Database Report Volume 2B:
Preliminary Model Calibration Report Volume 2C: Data Evaluation and Interpretation Report,
Hudson River PCBs Reassessment RFFS. Prepared by TAMS/Tetra Tech, December, 1998.
USEPA, 1999. Responsiveness Summary for the Low Resolution Sediment Coring Report. USEPA,
Region 2, New York. February, 1999.
A-8 TAMS/MCA
-------
Table A-1. Time Between General Electric TID and Waterford Samples in 1991
TID Sample
Date Hour Minute
4/5/91
4/12/91
4/19/91
4/26/91
5/3/91
5/10/91
5/17/91
5/24/91
5/31/91
6/7/91
6/14/91
7/11/91
7/25/91
8/7/91
8/22/91
9/5/91
9/11/91
9/18/91
9/25/91
10/2/91
10/9/91
10/16/91
10/23/91
10/30/91
11/6/91
11/13/91
11/20/91
11/26/91
12/4/91
12/11/91
12/18/91
12/26/91
14
16
16
13
15
16
15
16
14
16
17
16
7
12
10
11
10
10
10
10
10
10
10
9
10
9
9
10
10
11
11
10
30
0
15
0
15
0
10
15
15
0
0
0
20
0
45
15
50
15
25
40
20
0
10
35
40
20
55
50
25
5
20
45
Waterford Sample
Date Hour Minute
4/5/91
4/12/91
4/19/91
5/3/91
5/17/91
5/31/91
6/7/91
6/14/91
7/11/91
7/25/91
8/7/91
8/22/91
9/5/91
9/11/91
9/18/91
9/25/91
10/2/91
10/9/91
10/16/91
10/23/91
10/30/91
11/6/91
11/13/91
11/20/91
11/26/91
12/4/91
12/11/91
12/18/91
12/26/91
17
18
19
17
17
17
18
19
18
14
14
13
15
13
12
12
13
13
12
12
12
13
12
12
13
13
14
14
14
30
15
15
20
15
10
0
0
10
10
30
0
25
30
45
50
30
0
45
40
15
30
0
30
30
10
20
20
10
Fort Edward
Flow Rate
6240
12900
4750
6820
4000
3310
2900
2210
2590
2210
2320
2450
2170
2890
3230
2710
2410
3340
3180
3110
2440
2590
3120
2870
3300
3700
4220
4200
3600
Interval Between Estimated Time from TID
Samples (hours) to Waterford (hours)
3.0
2.3
3.0
2.1
2.1
2.9
2.0
2.0
2.2
6.8
2.5
2.3
4.2
2.7
2.5
2.4
2.8
2.7
2.8
2.5
2.7
2.8
2.7
2.6
2.7
2.8
3.3
3.0
3.4
48
23
63
44
74
90
103
135
115
135
128
122
137
103
92
110
124
89
94
96
122
115
96
104
90
81
71
71
83
TAMS/MCA
-------
Table A-2. Summary of Conversion for the Di through Hexa Homologues
Homologue
Calibration
Di-Hexa
Tri-Hexa
Tri-Hexa
Tri-Hexa
Period
Period
1987-1990
Fall-winter 1991 -1998
Spring 1991 -1998
Summer 1991-1998
Mean Mass
Percent of
Tri+ Using
TID Data
Mean Mass
+2 Standard -2 Standard Percent Ratio
Errors
Repeat
GE TID Data
GE TID Data
GE TID Data
Errors
the 1991
Waterford/TID
Distribution
Same as below by
homologue.
11
11
Corrected TID
Mass Percent
Varies
Varies
Varies
Mass Percent
of Tri+ at
Waterford
Varies
Varies
Varies
Forecast Period
Di
Di
Di
Di
Di
Di
Tri
Tri
Tri
Tetra
Tetra
Tetra
Penta
Penta
Penta
Hexa
Hexa
Hexa
Tri-Hexa
Tri-Hexa
Tri-Hexa
High Flow 1991 -1995
Low Flow 1991 -1995
High Flow 1996-1 998
Low Flow 1996-1 998
High Flow 1999+
Low Flow 1999+
Fall-winter 1 999+
Spring 1999+
Summer 1999+
Fall-winter 1999+
Spring 1999+
Summer 1999+
Fall-winter 1999+
Spring 1999+
Summer 1999+
Fall-winter 1999+
Spring 1999+
Summer 1999+
Fall-winter 1999+
Spring 1 999+
Summer 1999+
32.17
48.40
70.64
96.46
70.64
96.46
47.21
45.90
54.30
29.66
34.41
30.12
18.10
15.65
12.95
5.00
4.04
2.62
99.97
100.00
99.99
36.28
53.02
76.69
102.16
76.69
102.16
48.82
47.71
55.12
30.51
35.55
30.55
19.22
16.88
13.54
5.58
4.61
2.82
28.07
43.78
64.60
90.76
64.60
90.76
45.60
44.09
53.48
28.81
33.26
29.69
16.98
14.41
12.37
4.42
3.48
2.41
1.04
0.52
1.04
0.52
1.04
0.52
0.98
0.98
0.91
0.97
0.97
1.09
1.19
1.19
1.28
1.23
1.23
1.39
33.37
25.41
73.27
50.64
73.27
50.64
46.11
44.83
49.18
28.76
33.36
32.81
21.49
18.58
16.64
6.15
4.97
3.64
102.50
101.74
102.26
33.37
25.41
73.27
50.64
73.27
50.64
44.97
44.06
48.08
28.05
32.79
32.08
20.96
18.26
16.27
6.00
4.89
3.56
99.97
100.00
99.99
TAMS/MCA
-------
Di/Tri+ By Station
125
100 -
75 -
50 -
T
WTTRD
SFNS
AlPnh
Titey-Kroner
0.05
Level
TID
WTFRD
minimum
11.2
10.12
Level
TID
WTFRD
10.0%
12.649
10.984
Quantiles
25.0%
28.245
14.355
median
75.42
35.765
75.0%
100.2925
67.7975
90.0%
112.793
74.643
maximum
115.58
74.76
Means and Std Deviations
Number Mean Std Dev
12 68.6608 36.4512
12 39.5517 24.7008
Means Comparisons
Dif=Mean[i]-Mean[j] TID
TID 0.0000
WTFRD -29.1092
Std Err Mean
10.523
7.130
WTFRD
29.1092
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.07387
Abs(Dif)-LSD TID WTFRD
TID -26.3609 2.7483
WTFRD 2.7483 -26.3609
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-1
Di/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
TAMS/MCA
-------
Di/Tri+ By Station
110
100 -
90 -
80 -
70 -
60 -
50 -
to -
30
I
FD
TOO
Station
/MPnh
0.05
Level
TID
WTFRD
Quantiles
minimum 10.0% 25.0% median
59.64 59.64 65.66 83.21
32.57 32.57 32.75 40.58
Means and Std Deviations
Level Number Mean Std Dev
TID 7 82.7286 18.8217
WTFRD 7 49.7000 17.8718
Means Comparisons
Dif=Mean[i]-MeanO] TID
TID 0.0000
WTFRD -33.0286
75.0%
102.25
71.17
90.0% maximum
106.29 106.29
74.37 74.37
Std Err Mean
7.1139
6.7549
WTFRD
33.0286
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.17880
Abs(Dif)-LSD
TID
WTFRD
TID
-21.3741
11.6545
WTFRD
11.6545
-21.3741
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-2
Di/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Summer
TAMS/MCA
-------
Di/Tri+ By Station
125
100 -
75 -
50 -
TOO
Statin
Level
TID
WTFRD
minimum
11.2
10.12
Level
TID
WTFRD
10.0%
11.2
10.12
Quantiles
25.0%
13.615
11.56
median
17.78
14.29
75.0%
99.91
44.655
90.0%
115.58
74.76
maximum
115.58
74.76
Means and Std Deviations
Number Mean Std Dev
5 48.9660 47.8678
5 25.3440 27.6803
Means Comparisons
Dif=Mean[i]-Mean[j] TID
TID 0.0000
WTFRD -23.6220
Std Err Mean
21.407
12.379
WTFRD
23.6220
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2 30593
Abs(Dif)-LSD TID WTFRD
TID -57.0224 -33.4004
WTFRD -33.4004 -57.0224
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-3
Di/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Fall, Winter and Spring
TAMS/MCA
-------
Di/Tri+ By Station
100 -
BO -
•£
60 -
10 -
o
re
WTFRD
Stotw
flPnrs
Tttijf-KiciKf
0.05
Level
TID
WTFRD
minimum
59.64
32.57
Level
TID
WTFRD
10.0%
59.64
32.57
Quantiles
25.0%
66.1525
32.705
median
88.815
39.68
75.0%
105.28
61.0525
90.0%
115.58
71.17
maximum
115.58
71.17
Means and Std Deviations
Number Mean Std Dev
8 86.8350 20.9416
6 45.5883 15.5330
Means Comparisons
Dif=Mean[i]-MeanO] TID
TID 0.0000
WTFRD -41 .2467
Std Err Mean
7.4040
6.3413
WTFRD
41.2467
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.17880
Abs(Dif)-LSD TID WTFRD
TID -20.5649 19.0341
WTFRD 19.0341 -23.7463
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-4
Di/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Low Flow
TAMS/MCA
-------
Di/Tri+ By Station
9D
7D -
60 -
50 -
(0 -
30 -
JO -
10 -
WTRD
Stntbn
0.05
Level
TID
WTFRD
Quantiles
minimum 10.0% 25.0% median
11.2 11.2 12.4075 16.905
10.12 10.12 12.28 14.42
Means and Std Deviations
Level Number Mean Std Dev
TID 4 32.3125 34.7300
WTFRD 6 33.5150 31.8363
Means Comparisons
Dif=Mean[i]-Mean[j] WTFRD
WTFRD 0.00000
TID -1.20250
75.0%
67.625
74.4675
90.0% maximum
84.24 84.24
74.76 74.76
Std Err Mean
17.365
12.997
TID
1.20250
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.30593
Abs(Dif)-LSD WTFRD TID
WTFRD -43.8689 -47.8444
TID -47.8444 -53.7282
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-5
Di/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterfbrd Stations
- High Flow
TAMS/MCA
-------
6 USEPA
- » GE
Source: Hudson River Database Release 4.1
Figure A-6
Di/Tri+ Mass Ratio in USEPA and General Electric Water Column Samples
at the Thompson Island Dam
TAMS/MCA
-------
+
•c
t:
Q
1
0.8
0.6
0.4
0.2
o!
o>
c
co
73
CD
O
• B
en
CO
75
0>
Q
In
cn
O5
c
(0
to
cn
c»
C
03
—
e
to
en
c
(0
-5
b>
cn
O>
c
-------
0.01
o>
£. 0.001
at
m
o
0.
"S
o 0.0001
10"
CO
CM
S3
eg
Total PCBs
Source: Hudson River Database Release 4.1
Figure A-8
Total PCBs in General Electric Water Column Samples
at the Thompson Island Dam
TAMS/MCA
-------
Di/Tri+ By <>=1996
200 -
.£ 100 -
o-
<
1
1
a
199
|
i —
1
S
0=1996
s
>=199
6
O
O
AlPoK
riAey-Krantf
0.05
Level
<1996
>=1996
minimum
2.76
0
Level
<1996
>=1996
10.0%
12.816
29.514
Quantiles
25.0%
23.755
50.05
median
35.47
77.49
75.0%
56.73
107.76
90.0%
78.752
133.368
maximum
115.75
209.28
Means and Std Deviations
Number Mean Std Dev
225 42.1256 25.5812
293 79.9830 39.3904
Means Comparisons
Dif=Mean[i]-Mean[j] >=1996
>=1996 0.0000
<1996 -37.8574
Std Err Mean
1.7054
2.3012
<1996
37.8574
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.96457
Abs(Dif)-LSD
>=1996
<1996
>=1996
-5.5332
31.9209
<1996
31.9209
-6.3142
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-9
Di/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Years
TAMS/MCA
-------
Di/Tri+ By SEASON
1ZD -
110 -
100 -
90 -
80 -
70 -
.£ 60 -
(0 -
30 -
10 -
•
"
rfl
4-
FriNrlw
•
•
— ••
i
i
J:
Sprtfl
"
1
•
j
-••
j
•
Simw
SEASON
a
0
AlPm
0.05
Level
Fall-winter
Spring
Summer
minimum
2.76
9.34
4.72
Level
Fall-winter
Spring
Summer
10.0%
8.732
11.96
22.16
Quantiles
25.0%
15.8275
17.855
30.21
median
31.31
29.67
40.835
75.0%
55.4925
45.305
60.1625
90.0%
95.303
58.962
89.935
maximum
112.5
81.09
115.75
Means and Std Deviations
Number Mean
66 39.4198
45 32.3282
114 47.5595
Means Comparisons
Dif=Mean[i]-Mean[j] Summer
Summer 0.0000
Fall-winter -8.1396
Spring -15.2313
Std Dev
29.1977
17.9895
24.6685
Fall-winter
8.1396
0.0000
-7.0916
Std Err Mean
3.5940
2.6817
2.3104
Spring
15.2313
7.0916
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.35960
Abs(Dif)-LSD Summer Fall-winter
Summer -7.8040 -0.9735
Fall-winter -0.9735 -10.2565
Spring 4.8584 -4.2988
Spring
4.8584
-4.2988
-12.4212
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-10
Di/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Season (1991-1995)
TAMS/MCA
-------
Di/Tri+ By SEASON
200 -
.£ 10D -
o-
i
-1-
•
i
Fnfl-iiirfef
-§•
•
3
String Sumtf
SHSON
o
u
0
/UP*!
FiAty-Kftrrer
0.05
Level
Fall-winter
Spring
Summer
minimum
7.03
0
19.56
Level
Fall-winter
Spring
Summer
10.0%
33.186
7.504
57.832
Quantiles
25.0%
42.4
21.5025
73.93
median
71.555
33.395
90.51
75.0%
91 .6575
60.9475
119.395
90.0%
126.802
114.419
141.026
maximum
150.34
174.33
209.28
Means and Std Deviations
Number Mean
76 71.4116
56 45.8050
161 95.9172
Means Comparisons
Dif=Mean[i]-Mean[j] Summer
Summer 0.0000
Fall-winter -24.5056
Spring -50.1122
Std Dev
34.0453
38.7849
32.7418
Fall-winter
24.5056
0.0000
-25.6066
Std Err Mean
3.9053
5.1828
2.5804
Spring
50.1122
25.6066
0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.35588
Abs(Dif)-LSD Summer Fall-winter
Summer -9.0065 13.2594
Fall-winter 13.2594 -13.1087
Spring 37.5757 11.3755
Spring
37.5757
11.3755
-15.2712
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-11
Di/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Season (1996-1998)
TAMS/MCA
-------
Di/Tri+ By FLOW
110 -
100 -
90 -
BO -
70 -
.£ 60 -
§ 50 -
30 -
20 -
1ft —
o-
H
•
h
1
1
i
"*"
j
HiQn MM IDA Fluti
FLOW
O
o
AlPufi
Tiiiy-Kwrtr
0,05
Quantiles
Level minimum 10.0% 25.0% median 75.0% 90.0% maximum
High Flow 2.76 9.658 18.19 30.12 42.55 54.952 99.59
Low Flow 4.72 17.883 .27.2525 43 62.5625 93.017 115.75
Means and Std Deviations
Level Number Mean Std Dev Std Err Mean
High Flow 87 32.1717 19.1418 2.0522
Low Flow 138 48.4009 27.1546 2.3116
Means Comparisons
Dif=Mean[i]-Mean[j] Low Flow High Flow
Low Flow 0.0000 16.2291
High Flow -16.2291 0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.97066
Abs(Dif)-LSD Low Flow High Flow
Low Flow -5.78354 9.65241
High Flow 9.65241 -7.28406
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-12
Di/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Flow (1991-1995)
TAMS/MCA
-------
Di/Tri+ By FLOW
200 -
i •
o-
H
3
rtfbi
F
Lwi
LOW
•
1
— T-
•
I
Fbn
O
a Purs
Riiy- Kramer
0.05
Level
High Flow
Low Flow
minimum
0
34.84
Level
High Flow
Low Flow
Quantiles
10.0% 25.0% median 75.0% 90.0% maximum
23.702 36.07 66.1 100.16 131.4 174.44
69.141 76.41 89.6 111.5575 139.054 209.28
Means and Std Deviations
Number Mean Std Dev Std Err Mean
187 70.6428 41.3259 3.0220
106 96.4606 29.3286 2.8486
Means Comparisons
Dif=Mean[i]-Mean[j] Low Flow High Flow
Low Flow 0.0000 25.8177
High Flow -25.8177 0.0000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.96815
Abs(Dif)-LSD Low Flow High Flow
Low Flow -10.1225 16.8581
High Flow 16.8581 -7.6212
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-13
Di/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Flow (1996-1998)
TAMS/MCA
-------
Tri/Tri+ By Stations
Level
TID
WTFRD
6D.O -
57.5 -
.+ 550 -
|
52.5 -
50.0 -
• —
T _
1
J_
1
m mo
Stntiw
Quantiles
minimum 10.0% 25.0%
55.42 55.42 55.74
48.92 48.92 48.92
o
o
AlPors
FiA^- Kramer
0,05
median 75.0%
57.48 59.56
51 .6 54.62
90.0%
60.84
58.21
maximum
60.84
58.21
Means and Std Deviations
Level Number Mean Std Dev Std Err Mean
TID 7 57.5943 2.03177 0.7679
WTFRD 7 52.1600 3.55562 1.3439
Means Comparisons
Dif=Mean[i]-Mean[j] TID WTFRD
TID 0.00000 5.43429
WTFRD -5.43429 0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.17880
Abs(Dif)-LSD TID WTFRD
TID -3.37242 2.06187
WTFRD 2.06187 -3.37242
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-14
Tri/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Summer
TAMS/MCA
-------
Tetra/Tri+ By Stations
36 -
rio
State
AlPoirs
lUcy-
0,05
Level
TID
WTFRD
minimum
32.47
32.91
Level
TID
WTFRD
10.0%
32.47
32.91
Quantiles
25.0%
32.49
35.37
median
34.14
36.75
75.0%
36.03
39.46
90.0%
37.02
41.48
maximum
37.02
41.48
Means and Std Deviations
Number Mean Std Dev
7 34.2100 1.78649
7 37.2629 2.88486
Means Comparisons
Dif=Mean[i]-Mean[j] WTFRD
WTFRD 0.00000
TID -3.05286
Std Err Mean
0.6752
1.0904
TID
3.05286
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.17880
Abs(Dif)-LSD WTFRD TID
WTFRD -2.79434 0.25852
TID 0.25852 -2.79434
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-15
Tetra/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Summer
TAMS/MCA
-------
Penta/Tri+ By Stations
Pentan"ri+
i i i i i i i
Level minimum
TID 3.58
WTFRD 7.47
^_
T
I
1 ' •
rt
. —
j_
mo
Stntbns
Quantiles
10.0% 25.0%
3.58 5.61
7.47 8.14
Q
0
AlPors
FuliEy-K rarer
0.05
median 75.0% 90.0% maximum
7.1 8.27 8.43 8.43
8.26 9.73 10.15 10.15
Means and Std Deviations
Level Number Mean Std Dev Std Err Mean
TID 7 6.85143 1.76931 0.66874
WTFRD 7 8.79857 1.01821 0.38485
Means Comparisons
Dif=Mean[i]-Mean[j] WTFRD TID
WTFRD 0.00000 1.94714
TID -1.94714 0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.17880
Abs(Dif)-LSD WTFRD TID
WTFRD -1.68109 0.26606
TID 0.26606 -1.68109
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-16
Penta/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Summer
TAMS/MCA
-------
Hexa/Tri+ By Stations
2.5
1
^
+U
0.5
It
WTFliO
StntiOT
AlPoh
Tiiiy-Knniw
0,05
Level
TID
WTFRD
minimum
0.57
1.18
Level
TID
WTFRD
1 0.0%
0.57
1.18
Quantiles
25.0%
0.59
1.18
median
1.02
1.55
75.0%
1.43
1.78
90.0%
2.25
2.09
maximum
2.25
2.09
Means and Std Deviations
Number Mean Std Dev
7 1.10429 0.590815
7 1.53857 0.333038
Std Err Mean
0.22331
0.12588
Means Comparisons
Dif=Mean[i]-Mean[j] WTFRD TID
WTFRD 0.000000 0.434286
TID -0.43429 0.000000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.17880
Abs(Dif)-LSD
WTFRD
TID
WTFRD
-0.55852
-0.12423
TID
-0.12423
-0.55852
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-17
Hexa/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Summer
TAMS/MCA
-------
Tri/Tri+ By Stations
Level
TID
WTFRD
65 -
60 -
55 -
fe 50-
15 -
10 -
minimum
54.44
39.57
••"•""•"•""•"
T
^^^^^^^—
™
•
1
^—^—w^—
TO
Stntte
f^]
(J
ytlPurs
fiiiy-Knrrtr
0.05
Quantiles
10.0% 25.0% median 75.0%
54.44 55.04 55.71 60.09
39.57 45.48 56.27 66.025
90.0%
61.01
69.39
maximum
61.01
69.39
Means and Std Deviations
Level Number Mean Std Dev
TID 5 57.1940 2.7689
WTFRD 5 55.8560 11.3448
Means Comparisons
Dif=Mean[i]-Mean[j] TID
TID 0.00000
WTFRD -1.33800
Std Err Mean
1.2383
5.0735
WTFRD
1.33800
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2 30593
Abs(Dif)-LSD TID WTFRD
TID -12.0426 -10.7046
WTFRD -10.7046 -12.0426
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-18
Tri/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Fall, Winter and Spring
TAMS/MCA
-------
Tetra/Tri-f By Stations
37.5
35.0 -
32.5 -
i ,
•ffi 30.0 -
27.5 -
25.0
no
WTO
Stotiw
It tan
0.05
Level
TID
WTFRD
minimum
30.89
26.5
Level
TID
WTFRD
10.0%
30.89
26.5
Quantiles
25.0%
32.155
27.62
median
33.45
32.98
75.0%
34.235
36.49
90.0%
34.87
36.66
maximum
34.87
36.66
Means and Std Deviations
Number Mean Std Dev
5 33.2460 1.44787
5 32.2400 4.52570
Means Comparisons
Dif=Mean[i]-Mean[j] TID
TID 0.00000
WTFRD -1 .00600
Std Err Mean
0.6475
2.0240
WTFRD
1.00600
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.30593
Abs(Dif)-LSD TID WTFRD
TID -4.90012 -3.89412
WTFRD -3.89412 -4.90012
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A- 19
Tetra/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Fall, Winter and Spring
TAMS/MCA
-------
Penta/Tri+ By Stations
17.5
15,0 -
12.5 -
7.5 -
5.0 -
2.5
ro
WTRD
Stntbw
AlPois
riAey-Krtntr
0,05
Level
TID
WTFRD
minimum
4.84
3.71
Level
TID
WTFRD
10.0%
4.84
3.71
Quantiles
25.0%
5.435
5.6
median
8.54
9.34
75.0%
9.765
12.84
90.0%
10.53
15.91
maximum
10.53
15.91
Means and Std Deviations
Number Mean Std Dev
5 7.78800 2.30945
5 9.24400 4.42784
Means Comparisons
Dif=Mean[i]-Mean[j] WTFRD
WTFRD 0.00000
TID -1.45600
Std Err Mean
1.0328
1.9802
TID
1.45600
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.30593
Abs(Dif)-LSD
WTFRD
TID
WTFRD
-5.14996
-3.69396
TID
-3.69396
-5.14996
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-20
Penta/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Fall, Winter and Spring
TAMS/MCA
-------
Hexa/Tri+ By Stations
m
MTRD
Stows
AlPnn
0,05
Level
TID
WTFRD
minimum
0.66
0.38
Level
TID
WTFRD
10.0%
0.66
0.38
Quantiles
25.0%
0.73
0.645
median
0.82
1.19
75.0%
2.775
3.575
90.0%
4.2
5.35
maximum
4.2
5.35
Means and Std Deviations
Number Mean Std Dev
5 1.56600 1.49572
5 1.92600 1.98140
Std Err Mean
0.66891
0.88611
Means Comparisons
Dif=Mean[i]-Mean[j] WTFRD TID
WTFRD 0.000000 0.360000
TID -0.36 0.000000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.30593
Abs(Dif)-LSD
WTFRD
TID
WTFRD
-2.56012
-2.20012
TID
-2.20012
-2.56012
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-21
Hexa/Tri+ Mass Ratio in USEPA Phase 2 Samples at the TI Dam and Waterford Stations
- Fall, Winter and Spring
TAMS/MCA
-------
Correlations
Variable Tri/Tri+
Tn7Tri+ 1.0000
Tetra/Tri+ -0.4212
Penta/Tri+ -0.7396
Hexa/Tri+ -0.6588
Tetra/Tri+ Penta/Tri+
-0.4212
1.0000
-0.2161
-0.1544
-0.7396
-0.2161
1.0000
0.5716
Prin. Components / Factor Analysis
Principal Components
Eigenvalue: 2.3172
Percent: 57.9312
CumPercent: 57.9312
i -
3-
.
2-
1 -
CN
c. f) —
•c u
Q.
~
-3 "-
1.2456
31.1402
89.0713
0.4360
10.8997
99.9710
*•
f
* ; •.
4
*V.V.':;V'
-*^^^-
*:*^^P^^
•'^i
*. r.'-..-
+
-i -3 -2-10 1
PrM
* •
."•
•*' • .•
• %•• * •
•* •."
•
2 3 t 5 6
•
•
Hexa/Tri+
-0.6588
-0.1544
0.5716
1.0000
0.0012
0.0290
100.0000
General Electric-Plus Sign
USEPA-Squares
t
3-
2-
1 -
0-
-1-
*VA*.':v' •
* - s.-j "?
I'M I • I • I • I • I • I
-I -3 -I -\ D 1 1 3 t 5 6 1
PiM
Summer-Plus Sign
Fall, Winter and Spring-Squares
Figure A-22
Principal Components Analysis for USEPA and General Electric Water Column
Samples at TI Dam, Schuylerville, Stillwater and Waterford 1991-1998
Page 1 of 3
TAMS/MCA
-------
I
3'
2'
1 -
1 ••
-1-
-2-
-3-
(-3-2-10 1 2 3 ( 5 6
PrM
TID-X
Stillwater-Plus Sign
Schuylerville -Square
Waterford- Hollow Square
I • I ' I • I • I • I • I • I ' I • I • I
-* -3 -2 -1 0 1 2 3 ( 5 6 1
Pfhl
Low Flow-Plus Sign
HighFlow-Squares
Source: Hudson River Database Release 4.1
Figure A-22
Principal Components Analysis for USEPA and General Electric Water Column
Samples at TI Dam, Schuylerville, Stillwater and Waterford 1991-1998
Page 2 of 3
TAMS/MCA
-------
'• 5B5sW«. **:*
cSKS^v-''----
%$$%%'*••. •• *
V;? •*'•'*' *
-3 -J -1 D 1 2 3 t 5 6 7
PrM
1996-1998-Plus Sign
1991-1995-Squares
Source: Hudson River Database Release 4.1
Figure A-22
Principal Components Analysis for USEPA and General Electric Water Column
Samples at TI Dam, Schuylerville, Stillwater and Waterford 1991-1998
Page 3 of 3
TAMS/MCA
-------
'-EAST
E3
N
400
400
800 Feet
w
General Electric Water Column Stations
Shoreline
Source: Hudson River Database Release 4.1
Figure A-23
Location of General Electric Water Column Stations Near the Thompson Island Dam
TAMS/MCA
-------
0.7
0.6
0.5
0.4
0.3
0.2
0.1
CM
O)
O
c
(0
USEPA
TID-West
TID-East
co
73
co
Q
CO
05
O)
co
73
a
-3
CO
O5
o>
c
CD
- - X - - TID-PRW 1
+ TID-PRW 2
rr.- 1 ID-PRW 3
FS-18C
FS-18E
CO
73
o>
(0
-5
O
O
O
CM
-A---- TIP-18C
--» TIP-18SW
-•.> ViD-PRC 1
—S — TID-PRE 2
- -ffl - - TID-PRE 3
Source: Hudson River Database Release 4.1
Figure A-24
Tri/Tri+ Mass Ratio in USEPA and General Electric Water Column Samples
at the Thompson Island Dam
TAMS/MCA
-------
+
•c
CO
0.5
0.45
0.4
0.35
0.3
0.25
0.2
0.15
-6 USEPA
- = TID-West
o TID-East
- - X - - TID-PRW 1
+ TID-PRW 2
i TIO-PRW 3
-------
2
c:
C
ra
-6 USEPA
- » TID-West
o TID-East
IT"
O)
£2
o
O)
c
(0
O)
CO
c
(0
-5
C
(0
-A TIP-18C
--T TIP-18SW
•'.v TIO-PHF !
—S — TID-PRE 2
- ffl - - TID-PRE 3
Source: Hudson River Database Release 4.1
Figure A-26
Penta/Tri+ Mass Ratio in USEPA and General Electric Water Column Samples
at the Thompson Island Dam
TAMS/MCA
-------
CO
X
0.3
0.25
0.2
0.15
0.05
0
-0.05
CM
O)
0>
CM
01
m ,Un T .D ^a OH isa
inni^&n/»i4 oan
° a ti a
CO
°>
CO
73
S
CD
CD
°>
CO
73
00
O5
O>
c
C0
0>
O)
O>
o
o
o
CM
C
-------
Tri/Tri+ By SEASON
50 -
50 -
•£
"~ to -
•
30 -
•
fTj
•
i
LjJ
•
i
•
•
FnHrter
•
•
^
@"
I
•
•
•
3
"
:
i i
Sixirfl Simrtr
SEASON
O
0
AlPofs
Tiiej-Krnrei
0.05
Level
Fall-winter
Spring
Summer
minimum
28.78
25.47
36.45
Level
Fall-winter
Spring
Summer
10.0%
38.734
37.28
46.986
Quantiles
25.0%
40.69
42.935
50.92
median
47.355
46.265
55.24
75.0%
53.04
49.725
57.665
90.0%
56.456
53.652
60.39
maximum
61.36
61.99
63.3
Means and Std Deviations
Number Mean
76 47.2103
56 45.9007
161 54.3002
Means Comparisons
Dif=Mean[i]-Meanfj] Summer
Summer 0.00000
Fall-winter -7.08992
Spring -8.39947
Std Dev
7.02309
6.75424
5.19712
Fall-winter
7.08992
0.00000
-1.30955
Std Err Mean
0.80560
0.90257
0.40959
Spring
8.39947
1.30955
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.35588
Abs(Dif)-LSD Summer Fall-winter
Summer -1.58223 5.11422
Fall-winter 5.11422 -2.30291
Spring 6.19710 -1.19053
Spring
6.19710
-1.19053
-2.68281
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-28
Tri/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Season (1996-1998)
TAMS/MCA
-------
Tetra/Tri+ By SEASON
(0 -
s
tu
*~ 30 -
-
•
-4-
i
f-i-i
T
•4-
T
-i-
FoU-nrtK
^
4
$
1
1
•4
-1-
Scrho
•
9
i
•
.
Sumrrtr
SHSON
o
0
AJPors
riity-Krarrw
0.05
Level
Fall-winter
Spring
Summer
minimum
22.37
27.27
23.3
Level
Fall-winter
Spring
Summer
10.0%
25.087
28.957
26.822
Quantiles
25.0%
27.4375
30.91
28.51
median
29.57
34.195
29.86
75.0%
31.3825
37.285
31.56
90.0%
35.525
40.829
33.948
maximum
42.98
45.19
37.48
Means and Std Deviations
Number Mean
76 29.6617
56 34.4057
161 30.1178
Means Comparisons
Dif=Mean[i]-Mean[j] Spring
Spring 0.00000
Summer -4.28789
Fall-winter -4.74400
Std Dev
3.71522
4.29803
2.72116
Summer
4.28789
0.00000
-0.45612
Std Err Mean
0.42617
0.57435
0.21446
Fall-winter
4.74400
0.45612
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.35588
Abs(Dif)-LSD Spring Summer
Spring -1.48723 3.06698
Summer 3.06698 -0.87712
Fall-winter 3.35806 -0.63913
Fall-winter
3.35806
-0.63913
•1.27663
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-29
Tetra/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Season (1996-1998)
TAMS/MCA
-------
Penta/Tri+ By SEASON
30 -
1 J°~
£
•
10 -
•
•
j
ri
j
•
•
"d- nrler
i
•
T
-t
|
Sprig S
SttSON
a
—
limner
-
o
n
o
AlPoh
rj^-Krawr
0.05
Level
Fall-winter
Spring
Summer
minimum
8.76
6.3
6.26
Level
Fall-winter
Spring
Summer
10.0%
11.768
9.708
9.138
Quantiles
25.0%
14.9975
12.5925
10.38
median
17.35
15.655
12.31
75.0%
21.815
18.1875
15.125
90.0%
24.75
21.728
17.508
maximum
32.1
26.68
26.82
Means and Std Deviations
Number Mean
76 18.1020
56 15.6493
161 12.9545
Means Comparisons
Dif=Mean[i]-Mean[j] Fall-winter
Fall-winter 0.00000
Spring -2.45269
Summer -5.14750
Std Dev
4.89292
4.61842
3.70613
Spring
2.45269
0.00000
-2.69481
Std Err Mean
0.56126
0.61716
0.29208
Summer
5.14750
2.69481
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.35588
Abs(Dif)-LSD Fall-winter Spring
Fall-winter -1.61308 0.70150
Spring 0.70150 -1.87918
Summer 3.76361 1.15216
Summer
3.76361
1.15216
-1.10828
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-30
Penta/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Season (1996-1998)
TAMS/MCA
-------
Hexa/Tri+ By SEASON
M -
12 -
•
10 -
8_
1 6-
1 -
-
2-
i
.
-1-
• P
j
i
FoS-nriw
_;.
SI
tj
—4-
:
;
— 4 —
^
1
— I"
i i
Sorino Sunnw
SWSON
8
o
AlPois
riiEy-Krowr
0.05
Level
Fall-winter
Spring
Summer
minimum
0.55
1.08
0.84
Level
Fall-winter
Spring
Summer
10.0%
2.23
1.805
1.402
Quantiles
25.0%
3.235
2.5625
1.7
median
4.365
3.645
2.33
75.0%
6.4325
5.1575
3.265
90.0%
8.952
7.485
3.96
maximum
14.32
9.85
9.92
Means and Std Deviations
Number Mean
76 4.99842
56 4.04464
161 2.61528
Means Comparisons
Dif=Mean[i]-Mean[j] Fall-winter
Fall-winter 0.00000
Spring -0.95378
Summer -2.38314
Std Dev
2.53976
2.10051
1.30706
Spring
0.95378
0.00000
-1.42936
Std Err Mean
0.29133
0.28069
0.10301
Summer
2.38314
1.42936
0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
2.35588
Abs(Dif)-LSD Fall-winter Spring
Fall-winter -0.70961 0.18342
Spring 0.18342 -0.82667
Summer 1.77436 0.75074
Summer
1.77436
0.75074
-0.48754
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-31
Hexa/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Season (1996-1998)
TAMS/MCA
-------
Tri/Tri+ By FLOW
60 -
•
50 -
*~ (0 -
•
•
30 -
'
|
1
:
Hftfbn
j
— !
—
.
LDR fbfl
now
O
o
AlPrirs
riisy-Krarrtr
D.05
Level
High Flow
Low Flow
minimum
25.47
34.32
Level
High Flow
Low Flow
Quantiles
10.0% 25.0% median 75.0% 90.0% maximum
39.216 43.83 49.56 55.76 57.812 63.3
46.723 50.31 54.44 57.3375 60.507 62.11
Means and Std Deviations
Number Mean Std Dev Std Err Mean
187 49.2525 7.45478 0.54515
106 53.6843 5.48550 0.53280
Means Comparisons
Dif=Mean[i]-Mean[j] Low Flow High Flow
Low Flow 0.00000 4.43188
High Flow -4.43188 0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.96815
Abs(Dif)-LSD Low Flow High Flow
Low Flow -1.84111 2.80229
High Flow 2.80229 -1.38616
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-32
Tri/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Flow (1996-1998)
TAMS/MCA
-------
Tetra/Tri+ By FLOW
to -
30 -
n
O
LDfl
FLOW
AlPnrs
0.05
Quantiles
Level minimum 10.0% 25.0% median 75.0% 90.0% maximum
High Flow 22.74 26.808 28.46 30.6 33.5 37.146 45.19
Low Flow 22.37 26.64 28.44 29.76 31.28 33.962 37.48
Means and Std Deviations
Level Number Mean Std Dev Std Err Mean
High Flow 187 31.3062 4.16583 0.30464
Low Flow 106 29.9596 2.73741 0.26588
Means Comparisons
Dif=Mean[i]-Mean[j] High Flow Low Flow
High Flow 0.00000 1.34658
Low Flow -1.34658 0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.96815
Abs(Dif)-LSD High Flow Low Flow
High Flow -0.75602 0.45779
Low Flow 0.45779 -1.00415
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-33
Tetra/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Flow (1996-1998)
TAMS/MCA
-------
Penta/Tri+ By FLOW
3D -
1
10 -
.,
•
Hft Fta LMI
FLOW
3
:
Fbn
0
O
AlPoiri
Tlity-Krarrtr
0.05
Quantiles
Level minimum 10.0% 25.0% median 75.0%
High Flow 6.3 10.028 11.98 14.99 18.87
Low Flow 6.26 8.874 10.3725 12.545 15.9125
90.0%
22.76
17.912
maximum
32.1
26.41
Means and Std Deviations
Level Number Mean Std Dev Std Err Mean
High Flow 187 15.6719 4.97880 0.36409
Low Flow 106 13.2749 3.88450 0.37730
Means Comparisons
Dif=Mean[i]-Mean[j] High Flow Low Flow
High Flow 0.00000 2.39697
Low Flow -2.39697 0.00000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.96815
Abs(Dif)-LSD High Flow Low Flow
High Flow -0.93913 1.29290
Low Flow 1.29290 -1.24737
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-34
Penta/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Flow (1996-1998)
TAMS/MCA
-------
Hexa/Tri+ By FLOW
If -
12 -
10 -
6-
2-
o
LDB Fbn
FLOW
AlPoh
0.05
Quantiles
Level minimum 10.0% 25.0% median 75.0%
High Flow 0.55 1.602 2.21 3.12 4.82
Low Flow 1.02 1.407 1.665 2.75 3.66
90.0% maximum
6.88 14.32
5.237 9.92
Means and Std Deviations
Level Number Mean Std Dev Std Err Mean
High Flow 187 3.75824 2.26079 0.16533
Low Flow 106 3.06274 1.77186 0.17210
Means Comparisons
Dif=Mean[i]-Mean[j] High Flow Low Flow
High Flow 0.000000 0.695499
Low Flow -0.6955 0.000000
Alpha= 0.05
Comparisons for all pairs using Tukey-Kramer HSD
q*
1.96815
Abs(Dif)-LSD High Flow Low Flow
High Flow -0.42694 0.193583
Low Flow 0.193583 -0.56707
Positive values show pairs of means that are significantly different.
Source: Hudson River Database Release 4.1
Figure A-35
Hexa/Tri+ Mass Ratio in General Electric Samples at the TI Dam
Grouped by Flow (1996-1998)
TAMS/MCA
-------
THIS PAGE LEFT INTENTIONALLY BLANK
-------
Appendix B
-------
APPENDIX B
EFFECTS ASSESSMENT
This appendix provides a general overview of the toxicology of PCBs and describes the
methods used to characterize particular lexicological effects of PCBs on aquatic and terrestrial
organisms. Toxicity reference values (TRVs) used to estimate the potential risk to receptor species
resulting from exposure to PCBs are presented following the background on PCB toxicology. TRVs
are levels of exposure associated with either Lowest Observed Adverse Effects Levels (LOAELs)
or No Observed Adverse Effects Levels (NOAELs). They provide a basis for judging the potential
effects of measured or predicted exposures that are above or below these levels.
Use of both LOAELs and NOAELS provides perspective on the potential for risk as a result
of exposure to PCBs originating from the site. LOAELs are values at which effects have been
observed in either laboratory or field studies, while the NOAEL represents the lowest dose or body
burden at which an effect was not observed. Exceedance of a LOAEL indicates a greater potential
for risk.
B.I Poly chlorinated Biphenyl Structure and Toxicity
The toxicity of PCBs has been shown to manifest itself in many different ways, among
various species of animals. Typical responses to PCB exposure in animals include wasting
syndrome, hepatotoxicity, immunotoxicity, neurotoxicity, reproductive and developmental effects,
gastrointestinal effects, respiratory effects, dermal toxicity, and mutagenic and carcinogenic effects.
Some of these effects are manifested through endocrine disruption. Table B-l provides a summary
of the common effects documented to occur in animals as a result of PCB exposure.
PCBs are typically present in the environment as complex mixtures. These mixtures consist
of discrete PCB molecules that are individually referred to as PCB congeners. PCB congeners are
often introduced into the environment as commercial mixtures known as Aroclors. PCB toxicity
varies significantly among different congeners and is dependent on a number of factors. Two
significant factors relate to the chemical structure of the PCB congener (Figure B-l), including the
degree of chlorination and the position of the chlorines on the biphenyl structure (Safe et al., 1985a).
In general, higher chlorine content typically results in higher toxicity, and PCB congeners that are
chlorinated in the ortho position are typically less toxic than congeners chlorinated in the meta and
para positions. These differences are discussed in more detail in the following sections with a focus
on the metabolic processes involved in the activation of PCBs. Metabolic activation is believed to
be the major process contributing to PCB toxicity.
B.I.I Structure-Function Relationships of PCBs
PCB congeners have been shown to produce toxic effects similar to, although typically less
potent than, 2,3,7,8-tetrachlorodibenzo-/?-dioxin (2,3,7,8-TCDD), the most toxic member of all
B-l TAMS/MCA
-------
groups of halogenated aromatic hydrocarbons (Van den Berg et al. , 1998). The toxicity of these
hydrocarbons is thought to be related to their ability to induce cytochrome P450-dependent aryl
hydrocarbon metabolizing mixed-function oxidases (MFOs) (Safe et al. , 1985b; McFarland and
Clarke, 1989). Similar to 2,3,7,8-TCDD, a number of PCB congeners have been shown to induce
aryl hydrocarbon hydroxylase (AHH) activity, as well as ethoxyresorufin-O-deethylase (EROD)
activity. The potency and specificity of MFO induction of individual PCB congeners is directly
related to how closely they approach the molecular structure of 2,3,7,8-TCDD (Safe et al., 1985b;
McFarland and Clarke, 1989). The dioxin, 2,3,7,8-TCDD assumes a rigid coplanar configuration
which facilitates its binding to the cytosolic Ah (aryl hydrocarbon) receptor (AhR). Translocation
of the dioxin-A/z-receptor complex to the nuclear Ah locus is thought to initiate the synthesis of
enzymes that exhibit AHH and EROD activity (Safe et al., 1985a). The activation of these enzymes
may be involved in biotransformation, conjugation and removal, or metabolic activation of aryl
hydrocarbons to potentially toxic intermediates (McFarland and Clarke, 1989).
Studies of structure-function relationships for PCB congeners indicate that the location of
the chlorine substitution determines the type and intensity of the toxicity that can be elicited (Safe
etal. , 1985a). PCB congeners with substitutions at the meta- and para- positions as well as some
mono-ortho- substituted congeners assume a coplanar conformation similar to 2,3,7,8-TCDD, and
are typically more toxic than non-coplanar congeners with high orr/io-substitution. The phenyl rings
of PCB molecules are linked by a single carbon:carbon bond (Figure B-l), that, unlike the rigidly
bound phenyl rings of dioxins, allows relatively unconstrained freedom of rotation of one ring
relative to the other (Safe et al. , 1985a). When bulky chlorine atoms are substituted at certain
positions on the biphenyl nucleus they inflict certain constraints on rotational freedom. The greatest
effect is exerted by substitution of at least two opposing orjTzo-substitutions on opposite rings. The
energetic cost of maintaining a coplanar configuration becomes increasingly high as ortho
substitution increases. The release of steric hindrance, as a consequence of chlorine substitution in
ortho- positions, yields a non-coplanar molecular configuration, making it less "dioxin-like".
Moreover, since coplanarity facilitates binding to the AhR, which in turn effects the level of AHH
activity, metabolic activation, and potential toxicity of certain PCB congeners, the toxicity of PCB
congeners decreases as ortho substitution increases. PCB congeners with two chlorines in the ortho
position (di-ortho), or other highly ortho-substiluted congeners do not produce a strong, toxic,
"dioxin-like" response (McFarland and Clarke, 1989; Safe, 1990). Table B-2 lists the coplanar non-
ortho and mono-ortho congeners.
B.1.2 Metabolic Activation and Toxicity of PCBs
The toxicological effects of PCBs, as well as other halogenated aromatic hydrocarbons,
including dioxins, are correlated with their ability to induce the cytochrome P450-dependent mixed
function oxygenases (MFOs) (Safe et al., 1985b; McFarland and Clarke, 1989). MFOs are a group
of microsomal enzymes that catalyze oxidative biotransformation of aromatic ring-containing
compounds to facilitate conjugation and removal. This metabolic activation occurs mainly in the
liver and is a major mechanism of PCB metabolism and toxicity. The MFOs that are induced by
PCBs have been divided into three general groups: 3-methylcholanthrene-type (3-MC-type);
phenobarbital-type (PB-type); and mixed-type, possessing catalyzing properties of both. PB-induced
B-2 TAMS/MCA
-------
MFOs typically catalyze insertion of oxygen into conformationally nonhindered sites of non-coplanar
lipophilic molecules, such as orf/zo-substituted PCBs, and 3-MC-induced MFOs typically catalyze
insertion of oxygen into conformationally hindered sites of planar molecules, such as non-ortho-
substituted PCBs (McFarland and Clarke, 1989). The intermediate transition products typically
formed from these oxidations are reactive epoxides. Epoxide-derivatives of PCBs may be the
carcinogenic, mutagenic, or teratogenic metabolites of the parent compounds (McFarland and
Clarke, 1989). Ordinarily, reactions catalyzed by PB-induced MFOs go on to conjugation, which
generally increases their water solubility, making them more easily excreted. On the other hand, the
conformational hindrance of the oxygenated molecule subsequent to oxidation by 3-MC-induced
MFOs, provides stability of the intermediate and tends to inhibit conjugation and detoxification
(McFarland and Clarke, 1989). Thus, the potential for contributing to toxicity through bioactivation
via an epoxide-intermediate is considered to be much greater with 3-MC induced enzymic reactions.
This is reflected in the observed higher toxicity of the more "dioxin-like" coplanar PCBs, which are
potent inducers of AHH, a 3-MC-type MFO (McFarland and Clarke, 1989).
There is significant variability in MFO activity among species. MFO activity generally
decreases in the following order: mammals > birds and amphibians > fish (Walker et al. , 1984).
The levels in aquatic invertebrates were found to be even lower. In addition, the levels can vary
significantly even among closely related species (Knight and Walker, 1982). Low MFO activity may
be a significant contributing factor in the bioaccumulation of organochlorines in many organisms
(Fossietal. , 1990).
B.1.3 Estimating the Ecological Effects of PCBs
This ecological risk assessment focuses on effects that relate to the survival, growth, and
reproduction of individuals within the local populations of fish and wildlife species. Reproductive
effects are defined broadly herein to include egg maturation, spawning, egg hatchability, and survival
of fish larvae.
Reproductive effects tend to be the most sensitive endpoint for animals exposed to PCBs.
Indeed, toxicity studies in vertebrates indicate a relationship between PCB exposure, as demonstrated
by AHH induction, and functions that are mediated by the endocrine system, such as reproductive
success. A possible explanation for the relationship between AHH activity and reproductive success
may be due to a potential interference from the P450-dependent MFO with the ability of this class
of P450 proteins to regulate sex steroids. In fact, the induction of cytochrome P450 isozymes from
PCB exposure has been shown to alter patterns of steroid metabolism (Spies et al. , 1990). As
another example, the maternal hepatic AHH activity of the flatfish, Paralichthys stellatus, at the time
of spawning, was found to be inversely related to three reproductive functions: egg viability,
fertilization success, and successful development from fertilization through hatching (Long and
Buchman, 1990).
As discussed earlier, PCBs are often introduced into the environment as commercial PCB
congener mixtures, known as Aroclors. Historically, the most common approach for assessing the
ecological impact of PCBs has involved estimating exposure and effects in terms of totals or Aroclor
B-3 TAMS/MCA
-------
mixtures. It is important to note that, since different PCB congeners may be metabolized at different
rates through various enzymatic mechanisms, when subjected to processes of environmental
degradation and mixing, the identity of Aroclor mixtures is altered (McFarland and Clarke, 1989).
Therefore, depending on the extent of breakdown, the environmental composition of PCBs may be
significantly different from the original Aroclor mixture. Furthermore, commercial Aroclor mixtures
used in laboratory toxicity studies may not represent true environmental exposure to this Aroclor.
Thus, there are some uncertainties associated with estimating the ecological effects of PCBs in terms
of total PCBs or Aroclors. As a result, there has been a great emphasis on the development of
techniques that provide an assessment of potential risk from exposure to individual PCB congeners.
A methodology has been established, known as Toxic Equivalency (TEQ) Toxic Equivalency
Factors (TEF) methodology (TEQ/TEF), that quantifies the toxicities of PCB congeners relative to
the toxicity of the potent dioxin 2,3,7,8-TCDD (see van den Berg et al. , 1998 for review). It is
currently accepted that the carcinogenic potency of dioxin is effected by its ability to bind AhR. In
fact, dioxin is thought to be the most potent known AhR ligand (NOAA, 1999b). It is also generally
accepted that the dioxin-like toxicities of PCB congeners are directly correlated to their ability to
bind the AhR. Thus, the TEQ/TEF methodology provides a toxicity measurement for all AhR-
binding compounds based on their relative toxicity to dioxin. Since 2,3,7,8-TCDD has the greatest
affinity for the AhR, it is assigned a TCDD-Toxicity Equivalent Factor of 1.0. PCB congeners are
then assigned a TCDD-TEF relative to 2,3,7,8-TCDD, based on experimental evidence. For
example, if the relative toxicity of a particular congener is one-thousandth that of TCDD, it would
have a TEF of 0.001. The potency of a PCB congener is estimated by multiplying the tissue
concentration of the congener in question by the TEF for that congener to yield the toxic equivalent
(TEQ) of dioxin. Finally, a TEQ for the whole mixture can be determined from the sum of the
calculated TEQs for each AhR-binding congener. The World Health Organization has derived TEFs
for a number of PCB congeners (van den Berg et al. , 1998). These values are presented in Table
B-2.
An advantage of the TEQ/TEF approach is that it provides a basis for determining the
toxicity of a complex mixture of PCBs in media or tissues. The disadvantage of this approach is that
only AhR-active PCBs, and AhR-mediated endpoints, are considered for TEF calculations. For this
reason, it is useful to consider the TEQ/TEF method in concert with other methods for evaluating
toxicity.
Recent data suggest that non-AhR mediated side effects may be important contributors to
PCB toxicity. For example, Moore and Peterson (1996) suggest that PCBs may play a non-AhR
mediated role in the induction of neurotoxicity, hormonal effects, estrogenic effects, and infertility
in males. Although coplanar, "dioxin-like" congeners appear most toxic based on current evidence,
other congeners may have important non-AhR mediated toxic effects. Thus it is becoming
increasingly more important to examine the toxic effects of mixtures as well as individual congeners
of PCBs when evaluating the total ecological impact of PCBs.
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B.2 Selection of Measures of Effects
Many studies examine the effects of PCBs on aquatic and terrestrial organisms, and results
of these studies are compiled and summarized in several reports and reviews (e.g., Eisler and Belisle,
1996; Niimi, 1996; Hoffman et al. , 1998; ATSDR, 1996; Eisler, 1986; NOAA, 1999b). For the
present assessment, studies on the toxic effects of PCBs were identified by searching the National
Library of Medicine (NLM) MEDLENE and TOXLENE databases. Other studies were identified
from the reference section of papers that were identified by electronic search. Papers were reviewed
to determine whether the study was relevant to the topic.
Many different approaches and methodologies are used in these studies, some of which are
more relevant than others to the selection of toxicity reference values (TRVs) for the present risk
assessment. TRVs are levels of exposure associated with either LOAELs or NOAELs. They provide
a basis for judging the potential effects of measured or predicted exposures that are above or below
these levels. Some studies express exposures as concentrations or doses of total PCBs, whereas other
studies examine effects associated with individual congeners (e.g. PCB 126) or as total dioxin
equivalents (TEQs). This risk assessment develops separate TRVs for total PCBs and TEQs. This
chapter briefly describes the rationale that was used to select TRVs for various ecological receptors
of concern.
Some studies examine toxicity endpoints (such as lethality, growth, and reproduction) that
are thought to have greater potential for adverse effects on populations of organisms than other
studies. Other studies examine toxicity endpoints such as behavior, disease, cell structure, or
biochemical changes that affect individual organisms, but may not result in adverse effects at the
population level. For example, toxic effects such as enzyme induction may or may not result in
adverse effects to individual animals or populations. The present risk assessment selects TRVs from
studies that examine the effects of PCBs on lethality, growth or reproduction. Studies that examined
the effects of PCBs on other sublethal endpoints are not used to select TRVs. Lethality, growth, and
reproductive-based endpoints typically present the greatest risk to the viability of the individual
organism and therefore of the population's survival. Thus, these are considered to be the endpoints
of greatest concern relative to the stated assessment endpoints.
When exposures are expected to be long-term, data from studies of chronic exposure are
preferable to data from medium-term (subchronic), short-term (acute), or single-exposure studies
(USEPA, 1997). Because of the persistence of PCBs, exposure of ecological receptors to PCBs from
the Hudson River is expected to be long-term, and therefore studies of chronic exposure are used to
select TRVs for the present risk assessment. Long-term studies are also preferred because
reproductive effects of PCBs are typically studied after long-term exposure.
Dose-response studies compare the response of organisms exposed to a range of doses to that
of a control group. Ideally, doses that are below and above the threshold level that causes adverse
effects are examined. Toxicity endpoints determined in dose-response and other studies include:
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• NOAEL (No-Observed-Adverse-Effect-Level) is the highest exposure level shown to be
without adverse effect in organisms exposed to a range of doses. NOAELs may be expressed
as dietary doses (e.g., mg PCBs consumed/kg body weight/d), as concentrations in external
media (e.g., mg PCBs/kg food), or as concentrations in tissue of the effected organisms (e.g.,
mg chemical/kg egg).
• LOAEL (Lowest-Observed-Adverse-Effect-Level) is the lowest exposure level shown to
produce adverse effect in organisms exposed to a range of doses. LOAELs may also be
expressed as dietary doses (e.g., mg PCBs consumed/kg body weight/d), as concentrations
in external media (e.g., mg PCBs/kg food), or as concentrations in tissue of the effected
organisms (e.g., mg chemical/kg egg).
• LD50 is the Lethal Dose that results in death of 50% of the exposed organisms. Expressed
in units of dose (e.g., mg PCBs administered/kg body weight of test organism/d).
• LC50 is the Lethal Concentration in some external media (e.g. food, water, or sediment) that
results in death of 50% of the exposed organisms. Expressed in units of concentration (e.g.,
mg PCBs/kg wet weight food).
• ED50 is the Effective Dose that results in a sublethal effect in 50% of the exposed organisms
(mg/kg/d).
• EC50 is the Effective Concentration in some external media that results in a sublethal effect
in 50% of the exposed organisms (mg/kg).
• CBR or Critical Body Residue is the concentration in the organism (e.g., whole body, liver,
or egg) that is associated with an adverse effect (mg PCBs/kg wet wt tissue).
• EL-effect is the effect level that results in an adverse effect in organisms exposed to a single
dose, rather than a range of doses. Expressed in units of dose (mg/kg/d) or concentration
(mg/kg).
• EL-no effect is the effect level that does not result in an adverse effect in organisms
exposed to a single dose, rather than a range of doses. Expressed in units of dose (mg/kg/d)
or concentration (mg/kg).
Most USEPA risk assessments typically estimate risk by comparing the exposure of receptors
of concern to TRVs that are based on NOAELs. TRVs for the present baseline risk assessments are
developed on the basis of both NOAELs and LOAELs to provide perspective on the range of
potential effects relative to measured or modeled exposures.
Differences in the feeding behavior of aquatic and terrestrial organisms determine the type
of toxicity endpoints that are most easily measured and most useful in assessing risk. For example,
the dose consumed in food is more easily measured for terrestrial animals than for aquatic organisms
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since uneaten food can be difficult to collect and quantify in an aqueous environment. Therefore,
for aquatic organisms, toxicity endpoints are more often expressed as concentrations in external
media (e.g., water) or as accumulated concentrations in the tissue of the exposed organism (also
called a "body burden"). In some studies, doses are administered via gavage, intraperitoneal
injection into an adult, or injection into a fish or bird egg. If appropriate studies are available, TRVs
for the present baseline risk assessment are selected on the basis of the most likely route of exposure,
as described below:
• TRVs for benthic invertebrates are expressed as concentrations in external media (e.g.,
mg/kg sediment). Critical body burdens (e.g., mg/kg body weight) for benthic invertebrates
are presented, but a TRY is not selected due to limited data.
• TRVs for fish are expressed as critical body residues (CBR) (e.g., mg/kg whole body
weight and mg/kg lipid in eggs).
• TRVs for terrestrial receptors (e.g., birds and mammals) are expressed as daily dietary
doses (e.g., mg/kg whole body wt/d).
• TRVs for birds are also expressed as concentrations in eggs (e.g. mg/kg wet wt egg).
B.2.1 Methodology Used to Derive TRVs
The literature on toxic effects of PCBs to animals includes studies conducted solely in the
laboratory, as well as studies including a field component. Each type of study has advantages and
disadvantages for the purpose of deriving TRVs for a risk assessment. For example, a controlled
laboratory study can be designed to test the effect of a single formulation or congener (e.g. Aroclor
1254 or PCB 126) on the test species in the absence of the effects of other co-occurring
contaminants. This is an advantage since greater confidence can be placed in the conclusion that
observed effects are related to exposure to the test compound. However, laboratory studies are often
conducted on species that are easily maintained in the laboratory, rather than on wildlife species.
Therefore, laboratory studies may have the disadvantage of being conducted on species that are less
closely related to a particular receptor of concern. Field studies have the advantage that organisms
are exposed to a more realistic mixture of PCB congeners (with differences in toxic potencies), than,
for example, laboratory tests that expose organisms to a commercial mixture, such as Aroclor 1254.
Field studies have the disadvantage that organisms are usually exposed to other contaminants and
observed effects may not be attributable solely to exposure to PCBs. Field studies can be used most
successfully, however, to establish concentrations of PCBs or TEQs at which adverse effects are not
observed (e.g., a NOAEL). Because of the potential contribution of other contaminants (e.g. metals,
pesticides, etc.) to observed effects in field studies, the present risk assessment uses field studies to
establish NOAEL TRVs, but not LOAEL TRVs.
If appropriate field studies are available for species in the same taxonomic family as the
receptor of concern, those field studies will be used to derive NOAEL TRVs for receptors of
concern. Appropriateness of a field study will be based on the following considerations:
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• whether the study examines sensitive endpoints, such as reproductive effects, in a species
that is closely related (e.g. within the same taxonomic family) to the receptor of concern;
• whether measured exposure concentrations of PCBs or dioxin-like compounds are reported
for dietary doses, whole organisms, or eggs;
• whether the study establishes a dose-response relationship between exposure concentrations
of PCBs or dioxin-like contaminants and observed effects; and
• whether contributions of co-occurring contaminants are reported and considered to be
negligible in comparison to contribution of PCBs or dioxin-like compounds.
If appropriate field studies are not available for a test species in the same taxonomic family
as the receptor species of concern, laboratory studies will be used to establish TRVs for the receptor
species. The general methodology described in the following paragraphs will be used to derive TRVs
for receptors of concern from appropriate studies.
When appropriate chronic-exposure toxicity studies on the effects of PCBs on lethality,
growth, or reproduction are not available for a species of concern to the risk assessment,
extrapolations from other studies are made in order to estimate appropriate TRVs. For example, if
toxicity data is unavailable for a particular species of bird, toxicity data for a related species of bird
is used if appropriate information was available. Several methodologies have been developed for
deriving TRVs for wildlife species (e.g., Sample etal., 1996; California EPA, 1996; USEPA, 1996;
Menzie-Cura & Associates, 1997). The general methodology that is used to develop LOAEL and
NOAEL toxicity reference values (TRVs) for the present study is described below.
• If an appropriate NOAEL is unavailable for a phylogenetically similar species (e.g. within
the same taxonomic family), the assessment adjusts NOAEL values for other species (as
closely related as possible) by dividing by an uncertainty factor of 10 to account for
extrapolations between species. The lowest appropriate NOAEL is used whenever several
studies are available. However, if the surrogate test species is known to be the most sensitive
of all species tested in that taxonomic group (e.g. fish, birds, mammals), then an interspecies
uncertainty factor is not applied
• In the absence of an appropriate NOAEL, if a LOAEL is available for a phylogenetically
similar species, these may be divided by an uncertainty factor of 10 to account for a LOAEL
to NOAEL conversion. The LOAEL to NOAEL conversion is similar to USEPA's derivation
of human health RfD (Reference Dose) values, where LOAEL studies are adjusted by a
factor of 10 to estimate NOAEL values.
• When calculating chronic dietary dose-based TRVs (e.g. mg/kg/d) from data for sub-
chronic tests, the sub-chronic LOAEL or NOAEL values are divided by an additional
uncertainty factor of 10 to estimate chronic TRVs. The use of an uncertainty factor of 10 is
consistent with the methodology used to derive human health RfDs. These factors are applied
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to account for uncertainty in using an external dose (mg/kg/d in diet) as a surrogate for the
dose at the site of toxic action (e.g. mg/kg in tissue). Because organisms may attain a toxic
dose at the site of toxic action (e.g. in tissues or organs) via a large dose administered over
a short period, or via a smaller dose administered over a longer period, uncertainty factors
are used to estimate the smallest dose that, if administered chronically, would result in a
toxic dose at the site of action. USEPA has not established a definitive line between sub-
chronic and chronic exposures for ecological receptors. The present risk assessment follows
recently developed guidance (Sample et al., 1996) which considers 10 weeks to be the
minimum time for chronic exposure of birds and 1 year for chronic exposure of mammals.
• For studies that actually measure the internal toxic dose (e.g. mg PCBs/kg tissue), no sub-
chronic to chronic uncertainty factor is applied. This is appropriate since effects are being
compared to measured internal doses, rather than to external dietary doses that are used as
surrogates for the internal dose.
• In cases where NOAELs are available as a dietary concentration (e.g., mg contaminant per
kg food), a daily dose for birds or mammals is calculated on the basis of standard estimates
of food intake rates and body weights (e.g., USEPA, 1993).
The sensitivity of the risk estimates to the use of these various uncertainty factors is
examined in the uncertainty chapter (Chapter 6.0) of the ERA Addendum.
B.2.2 Selection of TRVs for Benthic Invertebrates
B.2.2.1 Sediment Guidelines
Various guidelines exist for concentrations of PCBs in sediment (Table B-3). Modeled
concentrations of PCBs in sediments of the Hudson River will be compared to the Sediment Effects
Concentrations (SEC) developed for this site (NOAA, 1999a).
B.2.2.2 Body Burden Studies
Relatively few studies were identified that examined the effects of PCBs or dioxin-like
compounds on the basis of body burdens in aquatic invertebrates. Concentrations of PCBs that are
without adverse effects range from 5.4 to 127 mg/kg wet wt (Table B-4). Lowest-observed-adverse-
effect-levels range from 27 to 1570 mg/kg wet wt. A body burden-based TRY is not developed
because of the limited amount of data that is available for benthic invertebrates.
B.2.3 Selection of TRVs for Fish
In this section, TRVs are developed for the forage fish receptors (pumpkinseed and spottail
shiner), as well as for fish that feed at higher trophic levels, such as the brown bullhead, yellow
perch, white perch, largemouth bass, striped bass, and shortnose sturgeon.
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Laboratory studies that examine the effects of total PCBs or Aroclors on fish are summarized
in Table B-5. Most of these studies report measured concentrations of PCBs in whole body fish
tissue, although one study (Black et ai, 1998a) reported a nominal injected dose. Laboratory studies
on the effects of dioxin-like compounds (TEQs) on fish (Table B-7) typically report concentrations
of TEQs in fish eggs, rather than in whole body, since eggs represent a more sensitive life stage.
Comparison of effect levels (e.g. NOAELs or LOAELs) reported as wet weight concentrations in
eggs to whole body tissue concentrations in adult Hudson River fish is complicated by the fact that
eggs and whole body adult fish tend to have different lipid contents and concentrations of lipophilic
contaminants, such as TEQs. However, if we assume that TEQs partition equally into the lipid phase
of the egg and into the lipids in the tissue of adult fish, then lipid-normalized concentrations in fish
eggs that are associated with adverse effects (ug TEQs/kg lipid) can be compared to lipid-normalized
tissue concentrations of TEQs in adult Hudson River fish. Therefore, this assessment establishes
TRVs for TEQs in fish on a lipid-normalized basis so that measured or predicted whole body
concentrations of TEQs in Hudson River fish can be compared to TRVs established from studies on
fish eggs.
B.2.3.1 Pumpkinseed (Lepomis gibbosus)
Total PCB Body Burden in Pumpkinseed
No laboratory studies were identified that examined toxicity of PCBs to the pumpkinseed
forage fish receptor, or to a fish species in the same family as the pumpkinseed (Table B-5, Figure
B-2). Two studies (Hansen et al., 1971 and Hansen et al., 1974) were identified that examined
toxicity of PCBs to species in the same order as the pumpkinseed (Table B-23). However, the studies
by Hansen et al. (1971, 1974) are not selected for the development of TRVs because these studies
examined adult mortality, which is not expected to be a sensitive endpoint. Therefore,
concentrations of PCBs in the pumpkinseed will be compared to the lowest appropriate NOAEL and
corresponding LOAEL from the available appropriate studies (Table B-5).
Hansen et al, (1974) established a NOAEL of 1.9 mg PCBs/kg and a LOAEL of 9.3 mg
PCBs/kg for adult female fish. This study was based on a flow-through bioassay of Aroclor 1254 on
sheepshead minnow. Fish were exposed for 28 days, and then egg production was induced. The
eggs were fertilized and placed in PCB-free flowing seawater and observed for mortality. The TRVs
resulting from this study are comparable to the TRVs for the study that was selected (Bengtsson,
1980).
The study by Black et al. (1998a) is not selected because it reports a nominal dose, rather
than a measured whole body concentration. The Hansen et al. (1974) study was not selected because
the Bengtsson study was more recent and of longer duration. The study by Bengtsson (1980) on the
minnow is selected as the lowest appropriate NOAEL for development of the TRV for pumpkinseed.
In this study, fish were exposed to Clophen A50 (a commercial mixture with a chlorine content of
50%) in food for 40 days. Although Clophen A50 was not used in the United States, the chlorine
content of Clophen A50 (50% chlorine) is reasonably similar to the chlorine content of Aroclor 1248
(48% chlorine) and Aroclor 1242 (42% chlorine) that were released into the Hudson River. The
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chlorine content of Hudson River fish resembles that of Aroclor 1254 (54% chlorine), which is more
similar to the chlorine content of Clophen A50, than to that of Aroclor 1248 or 1242 (Appendix K
USEPA, 1999). Therefore, it is believed that Clophen A50 is a reasonable surrogate of the actual
environmental composition of PCBs in Hudson River fish.
Hatchability was significantly reduced in fish with an average total PCB concentration of 170
mg/kg (measured on day 171 of the experiment), but not in fish with an average concentration of 15
mg/kg or 1.6 mg/kg. The only other reproductive endpoints that Bengsston et al. (1980) reported
to be significantly different in PCB-exposed fish as compared to control fish is the hatching time.
Fish in the medium and high exposure groups had significantly reduced hatching times compared
with the control group. Exposed fish that hatched prematurely all died within a week of hatching,
however, this result was not tested statistically. Nonetheless, because the prematurely hatched fry
all died, the low dose group is considered a NOAEL (1.6 mg/kg), and the medium dose group a
LOAEL( 15 mg/kg).
Because the experimental study measured the actual concentration in fish tissue, rather than
estimating the dose on the basis of the concentration in external media (e.g., food, water, or
sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied. An interspecies
uncertainty factor of 10 is applied to develop TRVs for the pumpkinseed.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the pumpkinseed is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the pumpkinseed is 0.16 mg PCBs/kg tissue (Table B-25).
Several field studies were identified that examined the effect of PCBs on the redbreast
sunfish, a species in the same family as the pumpkinseed (Tables B-6 and B-23). Field studies by
Adams et al. (1989, 1990, 1992) reported reduced fecundity, clutch size and growth in redbreast
sunfish (Lepomis auritus) that were exposed to PCBs and mercury in the field. However, since other
contaminants (e.g. mercury) were measured and reported in these fish and may have been
contributing to observed effects, these studies are used to develop a NOAEL TRVs, but not a
LOAEL TRY, for the pumpkinseed. An interspecies uncertainty factor is not applied since these
species are in the same family. Because the experimental study measured the actual concentration
in fish tissue, rather than estimating the dose on the basis of the concentration in external media (e.g.,
food, water, or sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied.
On the basis of the field studies:
The NOAEL TRY for the pumpkinseed is 0.5 mg PCBs/kg tissue (Table B-25).
As described previously, a LOAEL is not derived from the field studies because of the
potential for interactive effects of other contaminants in addition to PCBs.
Total Dioxin Equivalents (TEQs) in Eggs of the Pumpkinseed
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No laboratory studies were identified that examined toxicity of dioxin-like compounds to the
pumpkinseed or to a species in the same taxonomic family or order as the pumpkinseed (Tables B-7,
Figure B-3). Therefore, concentrations of TEQs in the pumpkinseed will be compared to the lowest
appropriate NOAEL and LOAEL from the selected studies (Table B-7). The study by Walker et al.
(1994) for the lake trout is selected as the lowest appropriate LOAEL and NOAEL from the selected
applicable studies (Table B-7). In that study, significant early life stage mortality was observed in
lake trout eggs with a concentration of 0.6 ug TEQs/kg lipid. This effect was not observed at a
concentration of 0.29 ug/kg lipid. Because the experimental study is based on the concentration in
the egg, rather than an estimated dose, a subchronic-to-chronic uncertainty factor is not applied.
Because salmonids, such as the lake trout, are among the most sensitive species tested (Table B-7),
an interspecies uncertainty factor is not applied.
On the basis of laboratory toxicity studies for salmonids:
The LOAEL TRY for the pumpkinseed is 0.6 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the pumpkinseed is 0.29 ug TEQs/kg lipid (Table B-25).
Because salmonids are known to be highly sensitive to effects of dioxin-like compounds
(Table B-7), alternative TRVs, developed from laboratory studies conducted on non-salmonid
species, are presented for comparison. (Uncertainty associated with comparison of Hudson River fish
to these TRVs is discussed in the uncertainty chapter). The lowest non-salmonid NOAEL (5.4 ug
TEQ/kg lipid) and LOAEL (103 ug TEQs/kg lipid) from the selected applicable studies (Table B-7)
for the fathead minnow, are used to derive alternative TRVs for the pumpkinseed. Because the
experimental study is based on the concentration in the egg, rather than an estimated dose, a
subchronic-to-chronic uncertainty factor is not applied. An interspecies uncertainty factor of 10 is
applied to account for potential differences between fathead minnow and pumpkinseed (Table B-25).
No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of the pumpkinseed or on a fish in the same taxonomic family as
the pumpkinseed (Table B-8).
B.2.3.2 Spottail Shiner (Notropis hudsonius)
Total PCB Body Burden in Spottail Shiner
Concentrations of PCBs in spottail shiner will be compared to the lowest appropriate NOAEL
and corresponding LOAEL from the selected applicable studies (Table B-5). The study by
Bengtsson (1980) on the minnow is selected as the lowest appropriate NOAEL (1.6 mg/kg) and
corresponding LOAEL (15 mg/kg) for development of the TRV for the spottail shiner because the
minnow is in the same family as the spottail shiner. Because the experimental study measured the
actual concentration in fish tissue, rather than estimating the dose on the basis of the concentration
in external media (e.g., food, water, or sediment, or injected dose), a subchronic-to-chronic
uncertainty factor is not applied. Because the spottail shiner and the minnow are in the same family,
an interspecies uncertainty factor is not applied.
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On the basis of laboratory toxicity studies:
The LOAEL TRY for the spottail shiner is 15 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the spottail shiner is 1.6 mg PCBs/kg tissue (Table B-25).
No field studies were identified that examined the effects of PCBs on the spottail shiner or
on a species in the same taxonomic family as the spottail shiner (Tables B-6 and B-23).
Total Dioxin Equivalents (TEQs) in Eggs of Spottail Shiner
Several laboratory studies were identified that examined toxicity of dioxin-like compounds
on fish in the same family as the spottail shiner (Tables B-7, Figure B-3). The study by Olivieri and
Cooper (1997) on the fathead minnow provides the lowest appropriate LOAEL and NOAEL from
the selected applicable studies (Table B-7). In that study, significant early life stage mortality was
observed in fathead minnow eggs with a concentration of 103 ug TEQs/kg lipid. This effect was not
observed at a concentration of 5.4 ug TEQs/kg lipid. The study did not report a lipid content for
fathead minnow eggs, so the 2.4% reported in Elonen et al. (1998) was used to obtain lipid
normalized results based on Olivieri and Cooper (1997). Because the experimental study is based
on the concentration in the egg, rather than an estimated dose, a subchronic-to-chronic uncertainty
factor is not applied. Because fathead minnow and spottail shiner are in the same taxonomic family,
an interspecies uncertainty factor is not applied.
Alternative TRVs for dioxin-like compounds are not developed for the spottail shiner since
the laboratory-based TRVs for the spottail shiner are not based on data for highly sensitive
salmonids.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the spottail shiner is 103 Mg TEQs/kg lipid (Table B-25).
The NOAEL TRY for the spottail shiner is 5.4 ^g TEQs/kg lipid (Table B-25).
No field studies were identified that examined the effects of dioxin-like compounds on
reproduction, growth or mortality of the spottail shiner or on a species in the same taxonomic family
as the spottail shiner (Table B-8).
B.2.3.3 Brown bullhead (Ameiurus nebulosus)
Total PCB Body Burden in the Brown Bullhead
No laboratory studies were identified that examined toxicity of PCBs to the brown bullhead
or to a species in the same taxonomic family or order as the brown bullhead (Table B-5, Figure B-2).
Therefore, concentrations of PCBs in the brown bullhead will be compared to the lowest appropriate
LOAEL and NOAEL from the selected applicable studies (Table B-5). The study by Black et al.
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(1998a) is not selected because it reports a nominal dose, rather than a measured whole body
concentration. The study by Bengtsson (1980) on the minnow is selected for development of the
TRY. Hatching time was significantly reduced in fish with an average total PCB concentration of
15 mg PCBs/kg, but not in fish with an average concentration of 1.6 mg PCBs/kg. Because the
experimental study measured the actual concentration in fish tissue, rather than estimating the dose
on the basis of the concentration in external media (e.g., food, water, or sediment, or injected dose),
a subchronic-to-chronic uncertainty factor is not applied. Because results of studies of PCBs and
dioxin-like compounds on fish eggs have shown that minnows are of intermediate sensitivity in
comparison to other fish (Tables B-5, B-7), an interspecies uncertainty factor of 10 is applied to
develop TRVs for the brown bullhead.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the brown bullhead is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the brown bullhead is 0.16 mg PCBs/kg tissue (Table B-25).
No field studies were identified that examined effects of PCBs on reproduction, growth or
mortality of the brown bullhead or on a species in the same taxonomic family as the brown bullhead
(Table B-6).
Total Dioxin Equivalents (TEQs) in Eggs of the Brown Bullhead
No laboratory studies were identified that examined toxicity of dioxin-like compounds on
the brown bullhead (Table B-7). The study by Elonen et al. (1998) on the channel catfish (Table B-
7) is selected for development of TRYs for the brown bullhead because the channel catfish and the
brown bullhead are in the same taxonomic family (Table B-23). In that study, significant early life
stage mortality was observed in catfish eggs having a concentration of 18 ug TEQs/kg lipid. This
effect was not observed at a concentration of 8.0 ug TEQs/kg lipid. Because the experimental study
is based on the concentration in the egg, rather than an estimated dose, a subchronic-to-chronic
uncertainty factor is not applied. An interspecies uncertainty factor is not applied because channel
catfish and brown bullhead are in the same taxonomic family.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the brown bullhead is 18 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the brown bullhead is 8.0 ug TEQs/kg lipid (Table B-25).
Because TRVs for effects of dioxin-like compounds on the brown bullhead were not based
on data for sensitive salmonid species, alternative TRVs are not derived.
B-14 TAMS/MCA
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No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of brown bullhead or a fish in the same taxonomic family as brown
bullhead (Table B-8).
B.2.3.4 Yellow Perch (Perca flavescens)
Total PCB Body Burden in the Yellow Perch
No laboratory studies were identified that examined toxicity of PCBs to the yellow perch
(Table B-5, Figure B-2). Two studies (Hansen et al, 1974 and Hansen et al., 1971) were identified
that examined toxicity of PCBs to species of the same order as the yellow perch. However, the
studies by Hansen et al. are not selected for the development of TRVs because these studies
examined adult mortality, which is not expected to be a sensitive endpoint. Therefore,
concentrations of PCBs in the yellow perch will be compared to the lowest appropriate NOAEL and
corresponding LOAEL from the selected applicable studies (Table B-5). The study by Black et al.
(1998a) is not selected because it reports a nominal dose, rather than a measured whole body
concentration. The study by Bengtsson (1980) on the minnow is selected as the lowest appropriate
NOAEL for development of the TRY. In this study, hatching time was significantly reduced in fish
with an average total PCB concentration of 15 mg/kg, but not in fish with an average concentration
of 1.6 mg PCBs/kg. Because the experimental study measured the actual concentration in fish tissue,
rather than estimating the dose on the basis of the concentration in external media (e.g., food, water,
or sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied. Because
results of studies of dioxin-like compounds and PCBs on fish eggs have shown another species of
minnow to be of intermediate sensitivity compared to all other fish species tested (Tables B-5, B-7),
an interspecies uncertainty factor of 10 is applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the yellow perch is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the yellow perch is 0.16 mg PCBs/kg tissue (Table B-25).
No field studies were identified that examined effects of PCBs on yellow perch or on a fish
in the same family as the yellow perch or on a species in the same family as the yellow perch (Tables
B-6 and B-23).
Total Dioxin Equivalents (TEQs) in Eggs of the Yellow Perch
No laboratory studies were identified that examined toxicity of dioxin-like compounds to the
yellow perch or to a species in the same taxonomic family or order as the yellow perch (Tables B-7,
Figure B-3). Therefore, concentrations of TEQs in the yellow perch will be compared to the lowest
appropriate NOAEL and corresponding LOAEL from the selected laboratory studies (Table B-7).
The study by Walker et al. (1994) reported significant early life stage mortality in lake trout eggs
with a concentration of 0.6 TEQs/kg lipid. This effect was not observed at a concentration of 0.29
ug/kg lipid. Because the experimental study is based on the concentration in the egg, rather than an
B-15 TAMS/MCA
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estimated dose, a subchronic-to-chronic uncertainty factor is not applied. Because lake trout are
among the most sensitive species tested (Table B-7), an interspecies uncertainty factor is not applied.
On the basis of laboratory toxicity studies for salmonids:
The LOAEL TRY for the yellow perch is 0.6 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the yellow perch is 0.29 ug TEQs/kg lipid (Table B-25).
Because salmonids are known to be highly sensitive to effects of dioxin-like compounds
(Table B-7), alternative TRVs, developed from studies conducted on non-salmonid species, are
presented for comparison. (Uncertainty associated with comparison of Hudson River fish to these
TRVs is discussed in Chapter 6 of the ERA Addendum.) The lowest NOAEL (5.4 ug TEQ/kg lipid)
and corresponding LOAEL (103 ug TEQs/kg lipid) for a non-salmonid species (Table B-7), the
fathead minnow, are presented as alternative TRVs for the yellow perch. An interspecies uncertainty
factor of 10 is applied to account for potential differences between the fathead minnow and the
yellow perch. Because the experimental study measured the concentration in the egg, rather than
estimating a dose, a subchronic-to-chronic uncertainty factor is not applied (Table B-25).
No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of the yellow perch or on a species in the same taxonomic family
as the yellow perch (Table B-8).
B.2.3.5 White Perch (Morone americana)
Total PCB Body Burden in the White Perch
No laboratory studies were identified that examined toxicity of PCBs to the white perch
(Table B-5, Figure B-2). Two studies (Hansen et al., 1974 and Hansen et al., 1971) were identified
that examined toxicity of PCBs to species of the same order as the white perch. However, the studies
by Hansen et al. are not selected for the development of TRVs because these studies examined adult
mortality, which is not expected to be a sensitive endpoint. Therefore, concentrations of PCBs in
the white perch will be compared to the lowest appropriate NOAEL and corresponding LOAEL from
the selected applicable studies (Table B-5). The study by Black et al. (1998a) is not selected
because it reports a nominal dose, rather than a measured whole body concentration. The study by
Bengtsson (1980) on the minnow is selected as the lowest appropriate NOAEL and corresponding
LOAEL for development of the TRV. In that study, hatching time was significantly reduced in fish
with an average total PCB concentration of 15 mg/kg, but not in fish with an average concentration
of 1.6 mg PCBs/kg. Because the experimental study measured the actual concentration in fish tissue,
rather than estimating the dose on the basis of the concentration in external media (e.g., food, water,
or sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied. Because
results of studies of dioxin-like compounds and PCBs on fish eggs have shown another species of
minnow to be of intermediate sensitivity compared to all other fish species tested (Tables B-5, B-7),
an interspecies uncertainty factor of 10 is applied.
B-16 TAMS/MCA
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On the basis of laboratory toxicity studies:
The LOAEL TRY for the white perch is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the white perch is 0.16 mg PCBs/kg tissue (Table B-25).
Two field studies were identified that examined the effects of PCBs on striped bass (Table
B-6). In one study, larval mortality was observed at concentrations of 0.1 to 10 mg PCBs/kg eggs,
but a NOAEL was not reported (Westin et al. , 1985). Another study found no adverse effect on
survival of striped bass larvae with average concentrations of 3.1 mg PCBs/kg larval tissue (Westin
et al., 1983). This study is selected for development of a NOAEL-based TRY for the white perch.
An interspecies uncertainty factor is not applied because white perch and striped bass are in the same
taxonomic family (Table B-23). Because the study measured the concentration in the larval tissue,
rather than estimating a dose, a subchronic-to-chronic uncertainty factor is not applied (Table B-25).
On the basis of the field study:
The NOAEL TRY for the white perch is 3.1 mg PCBs/kg tissue (Table B-25).
Total Dioxin Equivalents (TEQs) in Eggs of the White Perch
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
to the white perch or to a species in the same taxonomic family or order as the white perch (Tables
B-7, Figure B-3). Therefore, concentrations of TEQs in the white perch will be compared to the
lowest appropriate LOAEL and NOAEL from the selected studies (Table B-7). The study by Walker
et al. (1994) for the lake trout is selected as the lowest appropriate LOAEL and NOAEL from the
selected applicable studies (Table B-7). In that study, significant early life stage mortality was
observed in lake trout eggs with a concentration of 0.6 ug TEQs/kg lipid. This effect was not
observed at a concentration of 0.29 ug/kg lipid. Because the experimental study is based on the
concentration in the egg, rather than an estimated dose, a subchronic-to-chronic uncertainty factor
is not applied. Because lake trout are among the most sensitive species tested (Table B-7), an
interspecies uncertainty factor is not applied.
On the basis of laboratory toxicity for salmonid studies:
The LOAEL TRY for the white perch is 0.29 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the white perch is 0.6 ug TEQs/kg lipid (Table B-25).
Because salmonids are known to be highly sensitive to effects of dioxin-like compounds
(Table B-7), alternative TRVs, developed from studies conducted on non-salmonid species, are
presented for comparison. (Uncertainty associated with comparison of Hudson River fish to these
TRVs is discussed in Chapter 6 of the ERA Addendum.) The lowest NOAEL (5.4 ug TEQs/kg lipid)
and LOAEL (103 ug TEQs/kg lipid) for a non-salmonid species (Table B-7), the fathead minnow,
are used to develop alternative TRVs for the white perch (Olivieri and Cooper, 1997). An uncertainty
factor of 10 is applied to account for potential differences between the fathead minnow and the white
B-17 TAMS/MCA
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perch. Because the experimental study is based on the concentration in the egg, rather than an
estimated dose, a subchronic-to-chronic uncertainty factor is not applied (Table B-25).
No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of the white perch or on a species in the same taxonomic family
as the white perch (Table B-8).
B.2.3.6 Largemouth bass (Micropterus salmoides)
Total PCB Body Burden in the Largemouth Bass
No laboratory studies were identified that examined toxicity of PCBs to the largemouth bass
(Table B-5, Figure B-2). Two studies (Hansen et al., 1974 and Hansen et al., 1971) were identified
that examined toxicity of PCBs to species of the same order as the largemouth bass. However, the
studies by Hansen et al. are not selected for the development of TRVs because these studies
examined adult mortality, which is not expected to be a sensitive endpoint. Therefore,
concentrations of PCBs in the largemouth bass will be compared to the lowest appropriate NOAEL
and corresponding LOAEL from the selected applicable studies (Table B-5). The study by Black
et al. (1998a) is not selected because it reports a nominal dose, rather than a measured whole body
concentration. The study by Bengtsson (1980) on the minnow is selected as the lowest appropriate
NOAEL and corresponding LOAEL for development of the TRY. Hatching time was significantly
reduced in fish with an average total PCB concentration of 15 mg/kg, but not in fish with an average
concentration of 1.6 mg/kg. Because the experimental study measured the actual concentration in
fish tissue, rather than estimating the dose on the basis of the concentration in external media (e.g.,
food, water, or sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied.
Because results of studies of dioxin-like compounds and PCBs on fish eggs have shown another
species of minnow to be of intermediate sensitivity compared to all other fish species tested (Tables
B-5, B-7), an interspecies uncertainty factor of 10 is applied to the LOAEL (170 mg/kg) and NOAEL
(15 mg/kg) from this study to develop TRVs for the largemouth bass.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the largemouth bass is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the largemouth bass is 0.16 mg PCBs/kg tissue (Table B-25).
Several field studies were identified that examined effect of PCBs on the redbreast sunfish,
a species in the same family as the largemouth bass (Table B-6 and B-23). Field studies by Adams
et al. (1989, 1990, 1992) reported reduced fecundity, clutch size and growth in redbreast sunfish
(Lepomis auritus) that were exposed to PCBs and mercury in the field. However, since other
contaminants (e.g., mercury) were measured and reported in these fish and may have been
contributing to observed effects, these studies are used to develop a NOAEL TRVs, but not a
LOAEL TRY, for the largemouth bass. An interspecies uncertainty factor is not applied since these
species are in the same family. Because the experimental study measured the actual concentration
B-18 TAMS/MCA
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in fish tissue, rather than estimating the dose on the basis of the concentration in external media (e.g.,
food, water, or sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied.
On the basis of the field studies:
The NOAEL TRY for largemouth bass is 0.5 mg PCBs/kg tissue (Table B-25).
Total Dioxin Equivalents (TEQs) in Eggs of the Largemouth Bass
No laboratory studies were identified that examined toxieity of dioxin-like compounds to the
largemouth bass or to a species in the same taxonomic family or order as the largemouth bass (Table
B-7, Figure B-3). Therefore, concentrations of TEQs in the largemouth bass will be compared to
the lowest appropriate NOAEL and corresponding LOAEL from the selected studies (Table B-7).
The study by Walker et al. (1994) for the lake trout is selected as the lowest appropriate LOAEL and
NOAEL from the selected applicable studies (Table B-7). In that study, significant early life stage
mortality was observed in lake trout eggs with a concentration of 0.6 TEQs/kg lipid. This effect was
not observed at a concentration of 0.29 ug/kg lipid. Because the study is based on the concentration
in the egg, rather than an estimated dose, a subchronic-to-chronic uncertainty factor is not applied.
Because lake trout are among the most sensitive species tested (Table B-7), an interspecies
uncertainty factor is not applied.
On the basis of laboratory toxieity for salmonid studies:
The LOAEL TRY for the largemouth bass is 0.6 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the largemouth bass is 0.29 ug TEQs/kg lipid (Table B-25).
Because salmonids are known to be highly sensitive to effects of dioxin-like compounds
(Table B-7), alternative TRVs, developed from studies conducted on non-salmonid species, are
presented for comparison. (Uncertainty associated with comparison of Hudson River fish to these
TRVs is discussed in Chapter 6 of the ERA Addendum.) The lowest NOAEL (5.4 ug TEQ/kg lipid)
and corresponding LOAEL (103 ug TEQs/kg lipid) for a non-salmonid species, the fathead minnow,
are presented as alternative TRVs for the largemouth bass. An uncertainty factor of 10 is applied to
account for potential differences between the fathead minnow and the largemouth bass. Because the
experimental study is based on the concentration in the egg, rather than an estimated dose, a
subchronic-to-chronic uncertainty factor is not applied (Table B-25).
No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of the largemouth bass or on a species in the same taxonomic
family as the largemouth bass (Table B-8).
B.2.3.7 Striped bass (Morone saxatilis)
PCB Body Burdens in the Striped Bass
B-19 TAMS/MCA
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No laboratory studies were identified that examined toxicity of PCBs to the striped bass
(Table B-5, Figure B-2). Two studies were identified that examined toxicity of PCBs to species that
are in the same taxonomic order as the striped bass (Hansen et al., 1971, 1974). However, these
studies are not selected for the development of TRVs because they examined adult mortality, which
is not considered a sensitive endpoint. Therefore, concentrations of PCBs in the striped bass will be
compared to the lowest appropriate NOAEL and corresponding LOAEL from the selected applicable
studies (Table B-5). The study by Black et al. (1998a) is not selected because it reports a nominal
dose, rather than a measured whole body concentration. The study by Bengtsson (1980) on the
minnow is selected for development of the TRY. In this study, hatching time of eggs from adult fish
with an average total PCB concentration of 15 mg PCBs/kg was significantly reduced in comparison
to control fish. Hatching time was not reduced in eggs from adult fish with an average concentration
of 1.6 mg PCBs/kg. Because the study measured the actual concentration in fish tissue, rather than
estimating the dose on the basis of the concentration in external media (e.g., food, water, or
sediment, or injected dose), a subchronic-to-chronic uncertainty factor is not applied. Because results
of studies of dioxin-like compounds and PCBs on fish eggs have shown another species of minnow
to be of intermediate sensitivity compared to all other fish species tested (Table B-5, B-7), an
interspecies uncertainty factor of 10 is applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the striped bass is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the striped bass is 0.16 mg PCBs/kg tissue (Table B-25).
Two field studies were identified that examined the effects of PCBs on striped bass (Table
B-6). In one study, larval mortality was observed at concentrations of 0.1 to 10 mg PCBs/kg eggs,
but a NOAEL was not reported (Westin et al. , 1985). Another study found no adverse effect on
survival of striped bass larvae with average concentrations of 3.1 mg PCBs/kg larval tissue (Westin
et al. , 1983). This study is selected for development of a TRY for the striped bass. Because this
study measured the concentration in the larval tissue, rather than estimating a dose, a subchronic-to-
chronic uncertainty factor is not applied. An interspecies uncertainty factor is not applied (Table B-
25).
On the basis of the field study:
The NOAEL TRY for the striped bass is 3.1 mg PCBs/kg tissue (Table B-25).
Total Dioxin Equivalents (TEQs) in Eggs of Striped Bass
No laboratory studies were identified that examined toxicity of dioxin-like compounds to the
striped bass or to a species in the same taxonomic family or order as the striped bass (Table B-7,
Figure B-3). Therefore, concentrations of PCBs in the striped bass will be compared to the lowest
appropriate NOAEL and corresponding LOAEL from the selected applicable studies (Table B-7).
The study by Walker et al. (1994) for the lake trout is selected as having the lowest appropriate
LOAEL and NOAEL from the selected applicable studies (Table B-7). In that study, significant
B-20 TAMS/MCA
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early life stage mortality was observed in lake trout eggs with a concentration of 0.6 TEQs/kg lipid.
This effect was not observed at a concentration of 0.29 ug/kg lipid. Because the experimental study
is based on the concentration in the egg, rather than an estimated dose, a subchronic-to-chronic
uncertainty factor is not applied. Because lake trout are among the most sensitive species tested
(Table B-7), an interspecies uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the striped bass is 0.6 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the striped bass is 0.29 ug TEQs/kg lipid (Table B-25).
Because salmonids are known to be highly sensitive to effects of dioxin-like compounds
(Table B-7), alternative TRVs, developed from studies conducted on non-salmonid species, are
presented for comparison. (Uncertainty associated with comparison of Hudson River fish to these
TRYs will be discussed in the uncertainty chapter.) The lowest NOAEL (5.4 ug TEQ/kg lipid) and
corresponding LOAEL (103 ug TEQs/kg lipid) from the selected applicable studies (Table B-7) for
a non-salmonid species, the fathead minnow, are presented as alternative TRVs for the striped bass.
An uncertainty factor of 10 is applied to account for potential differences between the fathead
minnow and the striped bass. Because the study is based on the concentration in the egg, rather than
estimating a dose, a subchronic-to-chronic uncertainty factor is not applied (Table B-25).
No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of the striped bass or on a species in the same taxonomic family
as the striped bass (Table B-8).
B.2.3.8 Shortnose sturgeon (Acipenser brevirostrum)
Total PCB Body Burden in the Shortnose Sturgeon
No laboratory studies were identified that examined toxicity of PCBs to the shortnose
sturgeon or to a species in the same taxonomic family or order as the shortnose sturgeon (Table B-5,
Figure B-2). Therefore, concentrations of PCBs in the shortnose sturgeon will be compared to the
lowest appropriate LOAEL and NOAEL from the selected applicable studies (Table B-5). The study
by Black et al. (1998a) is not selected because it reports a nominal dose, rather than a measured
whole body concentration. The study by Bengtsson (1980) on the minnow is selected for
development of the TRY. In this study, hatching time of eggs from adult fish with an average total
PCB concentration of 15 mg PCBs/kg was significantly reduced. No effects were seen for fish with
an average concentration of 1.6 mg PCBs/kg. Because the experimental study measured the actual
concentration in fish tissue, a subchronic-to-chronic uncertainty factor is not applied. Because results
of studies of dioxin-like compounds and PCBs on fish eggs have shown another species of minnow
to be of intermediate sensitivity compared to all other fish species tested (Table B-5, B-7), an
interspecies uncertainty factor of 10 is applied.
On the basis of laboratory toxicity studies:
B-21 TAMS/MCA
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The LOAEL TRY for the shortnose sturgeon is 1.5 mg PCBs/kg tissue (Table B-25).
The NOAEL TRY for the shortnose sturgeon is 0.16 mg PCBs/kg tissue (Table B-25).
No field studies were identified that examined effects of PCBs on reproduction, growth or
mortality of the shortnose sturgeon or on a species in the same taxonomic family as the sturgeon
(Table B-6).
Total Dioxin Equivalents (TEQs) in Eggs of the Shortnose Sturgeon
No laboratory studies were identified that examined toxicity of dioxin-like compounds to the
shortnose sturgeon or to a species in the same taxonomic family or order as the shortnose sturgeon
(Table B-7, Figure B-3). Therefore, the lowest NOAEL and corresponding LOAEL from the
selected applicable studies (Table B-7) are selected for development of TRVs. Walker et al. (1994)
observed significant early life stage mortality in lake trout eggs with a concentration of 0.6 ug
TEQs/kg lipid. This effect was not observed at a body burden of 0.29 mg/kg lipid. Because the
study is based on the concentration in the egg, rather than estimating a dose, a subchronic-to-chronic
uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the shortnose sturgeon is 0.6 ug TEQs/kg lipid (Table B-25).
The NOAEL TRY for the shortnose sturgeon is 0.29 ug TEQs/kg lipid (Table B-25).
Because salmonids are known to be highly sensitive to effects of dioxin-like compounds
(Table B-7), alternative TRVs, developed from studies conducted on non-salmonid species, are
presented for comparison. (Uncertainty associated with comparison of Hudson River fish to these
TRVs is discussed in Chapter 6 of the ERA Addendum.) The lowest NOAEL (5.4 ug TEQ/kg lipid)
and corresponding LOAEL (103 ug TEQs/kg lipid) for a non-salmonid species, the fathead minnow,
are used to develop alternative TRVs for the shortnose sturgeon. An uncertainty factor of 10 is
applied to account for differences between the fathead minnow and the shortnose sturgeon. Because
the study is based on the concentration in the egg, rather than estimating a dose, a subchronic-to-
chronic uncertainty factor is not applied (Table B-25).
No field studies were identified that examined effects of dioxin-like compounds on
reproduction, growth or mortality of the shortnose sturgeon or on a species in the same taxonomic
family as the sturgeon (Table B-8).
B.2.4 Selection of TRVs for Avian Receptors
Toxicity studies for birds are typically based on dietary doses fed to the birds or on
concentrations of chemicals in eggs. Concentrations in eggs may be expressed as actual measured
concentrations, as is typical of field studies, or as nominal doses that are injected into the egg. TRVs
are developed for birds according to the methodology described previously.
B-22 TAMS/MCA
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B.2.4.1 Tree swallow (Tachycineta bicolor)
Total PCBs in the Diet of the Tree Swallow
No laboratory studies were identified that examined the toxicity of PCBs in the diet of the
tree swallow or a bird in the same taxonomic family or order as the tree swallow (Table B-9, Figure
B-4). Therefore, the lowest appropriate LOAEL and NOAEL from the selected studies, the LOAEL
(0.7 mg/kg/d) and NOAEL (0.1 mg/kg/d) for the domestic chicken (Scott, 1977), are used to develop
TRVs for the tree swallow. This study is selected for calculating TRVs for the tree swallow because
it shows a clear dose-response relationship with a meaningful endpoint. Scott (1977) found
significantly reduced hatchability in the eggs of hens that had been fed PCBs for a period of 4 or 8
weeks. A subchronic-to-chronic uncertainty factor of 10 is applied to the reported value to account
for the short-term exposure. Because gallinaceous birds, such as the chicken, are among the most
sensitive of avian species to the effects of PCBs, an interspecies uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the tree swallow is 0.07 mg PCBs/kg/day (Table B-26).
The NOAEL TRY for the tree swallow is 0.01 mg PCBs/kg/day (Table B-26).
Two field studies were identified that examined concentrations of PCBs in food of tree
swallows in comparison to measures of reproductive effects (Table B-10). Custer et al. (1998)
reported that measures of reproductive success (e.g., clutch and egg success) were not significantly
different for birds from a PCB-contaminated site in comparison to birds from a reference site. In that
study, dietary doses of PCBs, estimated on the basis of average measured food concentrations at the
site (2 samples) and a food ingestion rate of 0.9 kg food/kg body wt/day for the tree swallow, ranged
from 0.38 to 0.55 mg PCBs/kg/day.
Dietary doses of PCBs to tree swallows can also be estimated on the basis of composite
samples of food taken from feeding tree swallows on the Hudson River in 1995 (USEPA, 1998).
Dietary doses (estimated using the aforementioned food ingestion rate) for the tree swallow at
three locations on the Hudson River are 0.08, 6.0, and 16.1 mg PCBs/kg/day. The final TRY is
based on the highest concentration shown to be without adverse effects in both field studies, a
value of 16.1 mg PCBs/kg/day.
On the basis of field studies:
The NOAEL TRY for the tree swallow is 16.1 mg PCBs/kg/day (Table B-26).
Total Dioxin Equivalents (TEQs) in the Diet of the Tree Swallow
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the diet of the tree swallow or for a bird in the same taxonomic family or order as the tree swallow
B-23 TAMS/MCA
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(Tables B-l 1 and Figure B-5). Therefore, the lowest values from the selected applicable studies
(Table B-ll), the NOAEL (0.014 ug TEQs/kg/day) and corresponding LOAEL (0.0014 ug
TEQs/kg/day) for the pheasant (Nosek et al, 1992) are used to develop TRVs for the tree swallow.
Because gallinaceous birds, such as the pheasant, are among the most sensitive to 2,3,7,8-TCDD
(Table B-l 1), an interspecies uncertainty factor is not applied. Because of the short-term nature of
the exposure (10 weeks), a subchronic-to-chronic uncertainty factor of 10 is applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the tree swallow is 0.014 ug TEQs/kg/day (Table B-26).
The NOAEL TRY for the tree swallow is 0.0014 ug TEQs/kg/day (Table B-26).
Note that the study by Nosek et al. (1992) was also selected by the USEPA as the basis for
development of concentrations of 2,3,7,8-TCDD associated with risk to avian receptors (USEPA,
1993).
Two field studies were identified that examined the effects of dioxin-like compounds in the
diets of tree swallows (Table B-12). Custer et al. (1998) reported that measures of reproductive
success (e.g., clutch and egg success) were not significantly different for birds from a PCB-
contaminated site in comparison to birds from a reference site. In that study, dietary doses of dioxin-
like compounds were as high as 0.08 ug TEQs/kg/day.
Dietary doses of dioxin-like compounds to the tree swallow can also be estimated on the
basis of composite samples of food taken from feeding tree swallows on the Hudson River in 1995
(USEPA, 1998). Dietary doses .(estimated using the aforementioned food ingestion rate) for the tree
swallow at three locations on the Hudson River are: 0.12, 1.8, and 4.9 ug TEQs/kg/day. The final
TRY is based on the highest concentration shown to be without adverse effects in the 1995 field
study, a value of 4.9 ug TEQs/kg/day.
On the basis of the field studies:
The NOAEL TRY for the tree swallow is 4.9 jag TEQs/kg/day (Table B-26).
Total PCBs in Eggs of the Tree Swallow
No laboratory studies were identified that examined the toxicity of PCBs in eggs of the tree
swallow or for a bird in the same taxonomic family or order as the tree swallow (Table B-l3 and
Figure B-6). Therefore, the lowest appropriate NOAEL and corresponding LOAEL from the selected
applicable studies (Table B-l3) are used to develop TRVs for the tree swallow. The study by Scott
(1977) on chickens is selected for development of TRVs. This study is selected for calculating TRVs
for the tree swallow because it shows a clear dose-response with a meaningful endpoint. Scott (1977)
found significantly reduced hatchability in the eggs of hens that had been fed PCBs for a period of
4 or 8 weeks. Because gallinaceous birds, such as the chicken, are among the most sensitive of avian
species to the effects of PCBs, an interspecies uncertainty factor is not applied. Because the
B-24 TAMS/MCA
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experimental study measured actual concentrations in the egg, rather than reporting a surrogate dose,
a subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the tree swallow egg is 2.21 mg PCBs/kg egg (Table B-26).
The NOAEL TRY for the tree swallow egg is 0.33 mg PCBs/kg egg (Table B-26).
Several field studies were identified that examined effects of PCBs on eggs of the tree
swallow (Table B-14). Custer et al. (1998) found that clutch success (the probability of a clutch
hatching at least one young) and egg success (the probability of an egg hatching in a successful nest)
were not significantly lower at two contaminated sites in comparison to reference sites. Average
concentrations of total PCBs in eggs and pippers (newly hatched young) near a PCB contaminated
site ranged from 0.95 to 3.85 mg PCBs/kg and were significantly higher than concentrations from
the reference site, which ranged from 0.05 to 0.77 mg PCBs/kg.
The United States Fish and Wildlife Service (USFWS) studied the effects of PCB
contamination on tree swallows in the Upper Hudson River Valley in 1994 and 1995 (Secord and
McCarty, 1997, McCarty and Secord, 1999). Concentrations of PCBs were measured in tree
swallow eggs and nestlings from three sites on the Hudson River, one reference site on the
Champlain Canal, and one reference site in Ithaca, NY. Because concentrations of PCBs are not
usually measured in whole birds, concentrations of PCBs measured in whole bodies of Hudson River
tree swallows are not considered in this risk assessment.
In 1994, the mean mass of nestlings on the day of hatching from all of the Hudson River sites
combined was significantly less than the mean mass of nestlings from the Ithaca site. Reproductive
success at the Hudson sites was significantly impaired relative to other sites in New York due to
reduced hatchability and increased levels of nest abandonment during incubation, but clutch size,
nestling survival, and nestling growth and development were all normal. Average concentrations of
total PCBs in swallow eggs measured in 1994 were 11.7, 12.4, and 42.1 mg/kg wet wt for three
Hudson River sites, and 6.28 mg/kg wet wt for the Champlain Canal reference site (Secord and
McCarty, 1997).
In 1995 reproductive output of swallows at the Hudson sites was normal, but higher than
expected rates of abandonment and supernormal clutch size persisted. Growth and development of
nestlings was not significantly impaired. Average concentrations of PCBs in swallow eggs reported
in this subsequent study were 5.3, 24.1, and 26.7 mg/kg wet wt at the three Hudson sites, 5.9 mg/kg
at the Champlain Canal reference site, 1.85 mg/kg wet wt at an inland reference site, and 0.209
mg/kg wet wt at the Ithaca reference site.
Reproductive success in 1994 may have been influenced by the large number of young
females that typically inhabit nest boxes the first year that they are placed in the field (Secord and
McCarty, 1997). Because of the lack of a consistent pattern of reproductive success between the two
years of the study, these results are not used to establish a LOAEL TRY for the swallow. These
B-25 TAMS/MCA
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results do suggest, however, that tree swallows are more resistant to the effects of PCBs than are
many other species studied, and results can be used to derive a NOAEL TRY. Because of the
obvious relevance of the Hudson River study to the present assessment, the data from Secord and
McCarty are selected for development of a field-based TRY for the tree swallow. The highest
concentration from the year without significant effects is used to establish this field-based NOAEL
TRY for tree swallows.
On the basis of field toxicity studies:
The NOAEL TRY for tree swallows is 26.7 mg PCBs/kg egg (Table B-26).
Total Dioxin Equivalents (TEQs) in Eggs of the Tree Swallow
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the eggs of the tree swallow or for a bird in the same taxonomic family as the tree swallow (Table
B-15 and Figure B-7). Therefore, the lowest appropriate NOAEL (0.01 ug TEQs/kg egg) and
LOAEL (0.02 ug TEQs/kg egg) from the applicable studies are used to develop TRVs for the tree
swallow. Powell et al. (1996a) found significantly reduced hatchability in eggs of domestic
chickens that were injected with 0.2 ug PCB 126/kg egg. This effect was not observed in eggs
injected with 0.1 ug PCB 126/kg egg. The effective concentrations of BZ#126 are multiplied by the
TEF (0.1) for BZ#126 to estimate TRVs. Because gallinaceous birds, such as the chicken, are among
the most sensitive of avian species to the effects of dioxin-like compounds, an interspecies
uncertainty factor is not applied. Because by nature, a hatching period is a short-term event, a
subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the tree swallow is 0.02 ug TEQs/kg egg (Table B-26).
The NOAEL TRY for the tree swallow is 0.01 ug TEQs/kg egg (Table B-26).
Two field studies were identified that examined effects of dioxin-like compounds on tree
swallows (Table B-16). Field studies conducted in 1994 and 1995 reported elevated concentrations
of dioxin-like compounds in tree swallow eggs at contaminated Hudson River sites in comparison
to reference sites (USEPA, 1998). As noted in the discussion above regarding PCBs in tree swallow
eggs, reproductive success was significantly reduced in 1994, but not in 1995. Because of the lack
of a consistent pattern of reproductive success between the two years of the study, these results are
not used to establish a LOAEL TRY for the swallow. The results do suggest, however, that tree
swallows are more resistant to the effects of PCBs than are many other species studied, and the
results can be used to derive a NOAEL TRY. The highest average concentration from the year
without significant adverse effects on reproduction, growth, or mortality (13 ug TEQs/kg egg at the
Remnant Site in 1995) is used to establish this field-based NOAEL TRY for tree swallows.
On the basis of field toxicity studies:
B-26 TAMS/MCA
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The NOAEL TRY for the tree swallows is 13 ug TEQs/kg egg (Table B-26).
B.2.4.2 Mallard (Anas platyrhychos)
Total PCBs in Diet of the Mallard
Three laboratory studies were identified which examined effects of PCBs in the diet on
mallards (Table B-9, Figure B-4). The study that reported the lowest NOAEL is selected for
development of TRVs for the mallard. Custer and Heinz (1980) observed no adverse effects on
reproduction after approximately 1 month on a dosage of 2.6 mg Aroclor 1254/kg/day. Because of
the short-term exposure period of the experimental study (1 month), a subchronic-to-chronic
uncertainty factor of 10 is applied to the reported NOAEL. A LOAEL was not provided in this study,
so the LOAEL is assumed to be 10 times the estimated NOAEL for the mallard.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the mallard is 2.6 mg PCBs/kg/day (Table B-26).
The NOAEL TRY for the mallard is 0.26 mg PCBs/kg/day (Table B-26).
No field studies were identified that examined effects of dietary exposure to PCBs on
reproduction, growth or mortality of the mallard or on a species in the same taxonomic family as the
mallard (Table B-10).
Total Dioxin Equivalents (TEQs) in Diet of the Mallard
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the diet of the mallard or for a bird in the same taxonomic family or order as the mallard (Tables
B-ll and Figure B-5). Therefore, the lowest appropriate LOAEL (0.14 ug TEQs/kg/day) and
NOAEL (0.014 ug TEQs/kg/day) from the selected applicable studies (Table B-ll) (Noseketai,
1992) are used to develop TRVs for the mallard. Nosek et al. (1992) observed reduced fertility and
increased embryo mortality in ring-necked pheasants that received weekly intraperitoneal injections
of 2,3,7,8-TCDD over the course 10 weeks. It is generally acknowledged that intraperitoneal
injection and oral routes of exposure are similar because in both instances the chemical is absorbed
by the liver, thereby permitting first-pass metabolism (USEPA, 1995). Because data indicate that
the mallard (LD50 > 108 mg/kg/day for a single dose) is less sensitive than the pheasant (LD75 = 25
mg/kg/day for a single dose) to the acute effects of 2,3,7,8-TCDD (Table B-ll), an interspecies
uncertainty factor is not applied. Because of the short-term nature of the exposure in this study (10
weeks), a subchronic-to-chronic uncertainty factor of 10 is applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the mallard is 0.014 ug TEQs/kg/day (Table B-26).
The NOAEL TRY for the mallard is 0.0014 ug TEQs/kg/day (Table B-26).
B-27 TAMS/MCA
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No field studies were identified that examined effects of dietary exposure to dioxin-like
compounds on reproduction, growth or mortality of the mallard or on a species in the same
taxonomic family as the mallard (Table B-12).
Total PCBs in Eggs of the Mallard
No laboratory studies were identified that examined the toxicity of PCBs in eggs of the
mallard or for a bird in the same taxonomic family or order as the mallard (Table B-13 and Figure
B-6). Therefore, the lowest appropriate LOAEL and NOAEL from the selected applicable studies
(Table B-13) are used to develop TRVs for the mallard. The study by Scott (1977) on chickens is
selected for development of TRVs. This study is selected for calculating TRVs for the mallard
because it shows a clear dose-response with a meaningful endpoint. Scott (1977) found significantly
reduced hatchability in the eggs of hens that had been fed PCBs for a period of either 4 or 8 weeks.
Because gallinaceous birds, such as the chicken, are among the most sensitive of avian species to
the effects of PCBs, an interspecies uncertainty factor is not applied. Because the study measured
actual concentrations in the egg, rather than reporting a surrogate dose, a subchronic-to-chronic
uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRV for the mallard egg is 2.21 mg PCBs/kg egg (Table B-26).
The NOAEL TRV for the mallard egg is 0.33 mg PCBs/kg egg (Table B-26).
No field studies were identified that examined effects of PCBs in eggs of the mallard or in
eggs of a species in the same taxonomic family as the mallard (Table B-14).
Total Dioxin Equivalents (TEQs) in Eggs of the Mallard
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the eggs of the mallard or for a bird in the same taxonomic family as the mallard (Table B-15 and
Figure B-7). Therefore, the lowest appropriate NOAEL (0.01 jag TEQs/kg egg) and corresponding
LOAEL (0.02 jjg TEQs/kg egg) from the applicable studies are used to develop TRVs for the
mallard. Powell et al. (1996a) found significantly reduced hatchability in domestic chicken eggs
that were injected with 0.2 jag BZ#126/kg egg. This effect was not observed in eggs injected with
0.1 ng BZ#126/kg egg. The effective concentrations of BZ#126 are multiplied by the avian TEF for
BZ#126 (0.1) to estimate TRVs on a dioxin basis. Because gallinaceous birds, such as the chicken,
are among the most sensitive of avian species to the effects of dioxin-like compounds (Table B-15),
an interspecies uncertainty factor is not applied. Because the experimental study is based on an actual
measured dose to the egg, rather than on a surrogate dose, a subchronic-to-chronic uncertainty factor
is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRV for the mallard egg is 0.02 ug TEQs/kg egg (Table B-26).
B-28 TAMS/MCA
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The NOAEL TRY for the mallard egg is 0.01 ug TEQs/kg egg (Table B-26).
Two field studies were identified that examined effects dioxin-like compounds in eggs of the
wood duck, Aix sponsa, a species in the same family as the mallard (Tables B-16 and B-23). These
studies reported significant negative correlations between measures of reproductive effects and
concentrations of TEQs in eggs of wood ducks (White and Segniak, 1994 White and Hoffman,
1995). These studies reported substantially reduced nest success, hatching success, and duckling
production, at concentrations of 0.020 ug TEQs/kg egg. These effects were not observed at
concentrations of 0.005 ug TEQs/kg egg. Measured concentrations of organochlorine pesticides and
PCBs were low and were not believed to be biologically significant. Because of the relevance of this
study to the mallard, the LOAEL (0.02 ug TEQs/kg egg) and NOAEL (0.005 ug TEQs/kg egg) from
these studies are selected for development of a field-based TRY for the mallard. Note that this study
used TEFs provided by USEPA (1989) to calculate TEQs, which may differ slightly from TEFs used
in this report (Yan den Berg et al. , 1998). Potential differences in effect concentrations that are
based on use of differing TEFs are estimated at 12 to 30% (See sections on great blue herons and
mink). Because the mallard and the wood duck are in the same family, an interspecies uncertainty
factor is not applied. Because the LOAEL and NOAEL are based on measured concentrations, a
subchronic-to-chronic uncertainty factor is not applied.
On the basis of field studies:
The LOAEL TRY for the mallard egg is 0.02 ug TEQs/kg egg (Table B-26).
The NOAEL TRY for the mallard egg is 0.005 u TEgQs/kg egg (Table B-26).
B.2.4.3 Belted kingfisher (Ceryle alcyori)
Total PCBs in the Diet of the Belted Kingfisher
No laboratory studies were identified that examined the toxicity of PCBs in the diet of the
belted kingfisher or for a bird in the same taxonomic family or order as the kingfisher (Table B-9,
Figure B-4). Therefore, the lowest appropriate NOAEL (0.1 mg/kg/d) and corresponding LOAEL
(0.7 mg/kg/d) for the domestic chicken (Scott, 1977) are used to develop TRVs for the belted
kingfisher. This study is selected for calculating TRVs because it shows a clear dose-response
relationship with a meaningful endpoint. Scott (1977) found significantly reduced hatchability in the
eggs of hens that had been fed PCBs for a period of 4 or 8 weeks. A subchronic-to-chronic
uncertainty factor of 10 is applied to the reported value to account for the short-term exposure.
Because gallinaceous birds, such as the chicken, are among the most sensitive of avian species to
the effects of PCBs (Table B-9), an interspecies uncertainty factor is not applied. Because by nature
a hatching period is a short-term event, a subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the belted kingfisher is 0.07 mg PCBs/kg/day (Table B-26).
The NOAEL TRY for the belted kingfisher is 0.01 mg PCBs/kg/day (Table B-26).
B-29 TAMS/MCA
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No field studies were identified that examined effects of dietary exposure to PCBs on growth,
reproduction, or mortality of the belted kingfisher or to a species in the same taxonomic family as
the kingfisher (Table B-10).
Total Dioxin Equivalents (TEQs) in the Diet of the Belted Kingfisher
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the diet of the belted kingfisher or for a bird in the same taxonomic family or order as the
kingfisher (Tables B-l 1 and Figure B-5). Therefore, the lowest appropriate values from the selected
applicable studies (Table B-ll), the NOAEL (0.014 ug TEQs/kg/day) and LOAEL (0.14 ug
TEQs/kg/day) for the pheasant (Nosek et al. , 1992), are used to develop TRVs for the kingfisher.
Because gallinaceous birds, such as the pheasant, are among the most sensitive birds to the effects
of dioxin-like compounds (Table B-l 1), an interspecies uncertainty factor is not applied. Because
of the short-term nature of the exposure (10 weeks), a subchronic-to-chronic uncertainty factor of
10 is applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the belted kingfisher is 0.014 ug TEQs/kg/day (Table B-26).
The NOAEL TRY for the belted kingfisher is 0.0014 ug TEQs/kg/day (Table B-26).
No field studies were identified that examined effects of dietary exposure to dioxin-like
compounds on growth, reproduction, or mortality of the belted kingfisher or a species in the same
family as the kingfisher (Table B-l2).
Total PCBs in Eggs of the Belted Kingfisher
No laboratory studies were identified that examined the toxicity of PCBs in eggs of the belted
kingfisher or in eggs of a bird in the same order as the kingfisher (Tables B-l3 and Figure B-6).
Therefore, the lowest appropriate NOAEL and LOAEL from the selected applicable studies (Table
B-l3) are used to develop TRVs for the belted kingfisher. The study by Scott (1977) is selected for
development of TRVs since this study reports the lowest effect levels and provides both a NOAEL
and a LOAEL. Because gallinaceous birds, such as the chicken, are among the most sensitive of
avian species to the effects of PCBs, an interspecies uncertainty factor is not applied. Because by
nature, a hatching period is a short-term event, a subchronic-to-chronic uncertainty factor is not
applied.
On the basis of laboratory toxicity studies:
The LOAEL TRV for the belted kingfisher is 2.21 mg PCBs/kg egg (Table B-26).
The NOAEL TRV for the belted kingfisher is 0.33 mg PCBs/kg egg (Table B-26).
No field studies were identified that examined effects of PCBs in eggs of the belted
kingfisher or on a species in the same taxonomic family as the kingfisher (Table B-14).
B-30 TAMS/MCA
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Total Dioxin Equivalents (TEQs) in Eggs of the Belted Kingfisher
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the eggs of the belted kingfisher or for a bird in the same taxonomic family as the kingfisher
(Tables B-15 and Figure B-7). Therefore, the lowest appropriate NOAEL (0.01 ug TEQs/kg egg) and
LOAEL (0.02 ug TEQs/kg egg) from the applicable studies are used to develop TRVs for the belted
kingfisher. Powell et al. (1996a) found significantly reduced hatchability in domestic chicken eggs
that were injected with 0.2 ug PCB 126/kg egg. This effect was not observed in eggs injected with
0.1 ug BZ#126/kg egg. The effective concentrations of BZ#126 are multiplied by the avian TEF for
BZ#126 (0.1) to estimate TRVs on a dioxin basis. Because gallinaceous birds, such as the chicken,
are among the most sensitive of avian species to the effects of dioxin-like compounds (Table B-15),
an interspecies uncertainty factor is not applied. Because by nature a hatching period is a short-term
event, a subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the belted kingfisher egg is 0.02 ug TEQs/kg egg (Table B-26).
The NOAEL TRY for the belted kingfisher egg is 0.01 ug TEQs/kg egg (Table B-26).
No field studies were identified that examined effects of dioxin-like compounds on eggs of
the belted kingfisher or on a bird in the same taxonomic family as the kingfisher (Table B-16).
B.2.4.4 Great Blue Heron (Ardea herodias)
Total PCBs in the Diet of the Great Blue Heron
No laboratory studies were identified that examined the toxicity of PCBs in the diet of the
great blue heron or a bird in the same taxonomic family or order as the heron (Table B-9, Figure B-
4). Therefore, the lowest appropriate LOAEL and NOAEL from the applicable studies, the LOAEL
(0.7 mg/kg/d) and NOAEL (0.1 mg/kg/d) for the domestic chicken (Scott, 1977), are used to develop
TRVs for the great blue heron. Scott (1977) found significantly reduced hatchability in the eggs of
hens that had been fed PCBs for a period of 4 or 8 weeks. A subchronic-to-chronic uncertainty factor
of 10 is applied to the reported value to account for the short-term exposure. Because gallinaceous
birds, such as the chicken, are among the most sensitive of avian species to the effects of PCBs, an
interspecies uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the great blue heron is 0.07 mg PCBs/kg/day (Table B-26).
The NOAEL TRY for the great blue heron is 0.01 mg PCBs/kg/day (Table B-26).
No field studies were identified that examined effects of dietary exposure to PCB compounds
on growth, reproduction, or mortality of the great blue heron or on a species in the same taxonomic
family as the great blue heron (Table B-10).
B-31 TAMS/MCA
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Total Dioxin Equivalents (TEQs) in the Diet of the Great Blue Heron
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the diet of the great blue heron or for a bird in the same taxonomic family or order as the heron
(Tables B-l 1 and Figure B-5). Therefore, the lowest appropriate values from the selected applicable
studies (Table B-l 1), the NOAEL (0.014 ug TEQs/kg/day) and LOAEL (0.14 ug TEQs/kg/day) for
the pheasant (Nosek et al. , 1992), are used to develop TRVs for the great blue heron. Because
gallinaceous birds, such as the pheasant, are among the most sensitive birds to the effect 2,3,7,8-
TCDD (Table B-l 1), an interspecies uncertainty factor is not applied. Because of the short-term
nature of the exposure of the experimental study (10 weeks), a subchronic-to-chronic uncertainty
factor of 10 is applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the great blue heron is 0.014 ug TEQs/kg/day (Table B-26).
The NOAEL TRY for the great blue heron is 0.0014 ug TEQs/kg/day (Table B-26).
No field studies were identified that examined effects of dietary exposure to dioxin-like
compounds on growth, reproduction, or mortality of the great blue heron or on a species in the same
taxonomic family as the great blue heron (Table B-l2).
Total PCBs in Eggs of the Great Blue Heron
No laboratory studies were identified that examined the toxicity of PCBs in eggs of the great
blue heron or for a bird in the same taxonomic family or order as the heron (Tables B-l3 and Figure
B-6). Therefore, the lowest appropriate NOAEL and LOAEL (Scott, 1977) from the selected
applicable studies (Table B-l3) are used to develop TRVs for the great blue heron. Because
gallinaceous birds, such as the chicken, are among the most sensitive of avian species to the effects
of PCBs (Table B-l3), an interspecies uncertainty factor is not applied. Because by nature, a hatching
period is a short-term event, a subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for great blue heron eggs is 2.21 mg PCBs/kg egg (Table B-26).
The NOAEL TRY for great blue heron eggs is 0.33 mg PCBs/kg egg (Table B-26).
No field studies were identified that examined effects of PCBs to eggs of the great blue heron
or for eggs of a species in the same taxonomic family as the great blue heron (Table B-l4).
Total Dioxin Equivalents (TEQs) in Eggs of the Great Blue Heron
One laboratory study was identified that examined effects of dioxin-like compounds on eggs
of the great blue heron (Table B-15). Janz and Bellward (1996) found no substantial adverse effect
on hatchability or growth rate of chicks from great blue heron eggs that were injected with 2 ug
B-32 TAMS/MCA
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2,3,7,8-TCDD/kg egg. Because the study reports a measured dose to the egg rather than a surrogate
dose, no subchronic-to-chronic uncertainty factor is applied. Because the study was conducted on
the great blue heron, no interspecies uncertainty factor is applied.
On the basis of the laboratory toxicity study:
The NOAEL TRY for the great blue heron is 2.0 ug TEQs/kg egg (Table B-26).
Three field studies were identified that examined the effects of dioxins, furans, and PCBs in
field-collected eggs of the great blue heron at a site in British Columbia (Table B-16). One of the
studies documented complete reproductive failure in a colony of great blue herons with average egg
concentrations of 0.23 ug TEQs/kg egg in the 1986-1987 season (Elliott et al. , 1989). Average
concentrations of TEQs in great blue heron eggs from the same failed colony in 1988 were greater
than 0.5 ug TEQs/kg egg (Hart et al., 1991, Sanderson et al., 1994). The study by Sanderson et al
(1994) is selected for development of TRVs for the great blue heron because this study reported
concentrations of PCBs, in addition to concentrations of dioxins and furans. Sanderson et al. (1994)
reported no significant difference in hatchability of eggs, but a significant reduction in body weight
associated with egg concentrations greater than 0.5 ug TEQs/kg egg (Sanderson et al. , 1994). This
effect was not observed at egg concentrations of approximately 0.3 ug TEQs/kg egg (Sanderson et
al. , 1994). TEQs calculated by Sanderson et al. (1994) at the same site using the TEF values of
Safe et al. (1990) are estimated to be 30% lower than the concentration of TEQs that would be
calculated using the TEFs of Van den Berg et al. (1998) that are used in the present report. The
LOAEL (0.5 ug/kg egg) and NOAEL (0.3 ug TEQs/kg egg) from this study (Sanderson et al., 1994)
are selected for development of a field-based TRY for the great blue heron. Because the LOAEL and
NOAEL endpoints are based on measured concentrations, a subchronic-to-chronic uncertainty factor
is not applied.
On the basis of field toxicity studies:
The LOAEL TRY for the great blue heron is 0.5 ug TEQs/kg egg (Table B-26).
The NOAEL TRY for the great blue heron is 0.3 ug TEQs/kg egg (Table B-26).
B.2.4.5 Bald eagle (Haliaeetus leucocephalus)
Total PCBs in the Diet of the Bald Eagle
No laboratory studies were identified that examined the toxicity of PCBs in the diet of the
bald eagle or a bird in the same taxonomic family or order as the bald eagle (Table B-9, Figure B-4).
Therefore, the lowest appropriate the NOAEL (0.1 mg/kg/d) and corresponding LOAEL (0.7
mg/kg/d) for the domestic chicken (Scott, 1977), are used to develop TRVs for the great blue heron.
Scott (1977) found significantly reduced hatchability in the eggs of hens that had been fed PCBs for
a period of 4 or 8 weeks. A subchronic-to-chronic uncertainty factor of 10 is applied to the reported
value to account for the short exposure period of the experimental study (up to 8 weeks). Because
gallinaceous birds, such as the chicken, are among the most sensitive of avian species to the effects
of PCBs, an interspecies uncertainty factor is not applied.
B-33 TAMS/MCA
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On the basis of laboratory toxicity studies:
The LOAEL TRY for the bald eagle is 0.07 mg PCBs/kg/day (Table B-26).
The NOAEL TRY for the bald eagle is 0.01 mg PCBs/kg/day (Table B-26).
No field studies were identified that examined effects of dietary exposure to PCBs on growth,
reproduction, or mortality of the bald eagle or on a species in the same taxonomic family as the bald
eagle (Table B-10).
Total Dioxin Equivalents (TEQs) in the Diet of the Bald Eagle
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the diet of the bald eagle or for a bird in the same taxonomic family or order as the bald eagle
(Tables B-ll and Figure B-5). Therefore, the lowest values from the selected applicable studies
(Table B-ll), the NOAEL (0.014 ug TEQs/kg/day) and LOAEL (0.14 ug TEQs/kg/day) for the
pheasant (Nosek et al. , 1992) are used to develop TRVs for the bald eagle. Because gallinaceous
birds, such as the pheasant, are among the most sensitive birds to the effects 2,3,7,8-TCDD (Table
B-ll), an interspecies uncertainty factor is not applied. Because of the short-term nature of the
exposure (10 weeks), a subchronic-to-chronic uncertainty factor of 10 is applied. These TRVs are
expected to be protective of the bald eagle.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the bald eagle is 0.014 ug TEQs/kg/day (Table B-26).
The NOAEL TRY for the bald eagle is 0.0014 ug TEQs/kg/day (Table B-26).
No field studies were identified that examined effects of dietary exposure to dioxin-like
compounds on growth, reproduction, or mortality of the bald eagle or on a species in the same
taxonomic family as the bald eagle (Table B-12).
Total PCBs in Eggs of the Bald Eagle
No laboratory studies were identified that examined the toxicity of PCBs in eggs of the bald
eagle or for a bird in the same taxonomic family or order as the bald eagle (Table B-13 and Figure
B-6). Therefore, the lowest appropriate NOAEL and corresponding LOAEL from the selected
applicable studies (Table B-13) are used to develop TRVs for the bald eagle. The study by Scott
(1977) is selected for development of TRVs since this study reports a NOAEL and a LOAEL for a
meaningful reproductive endpoint. Because gallinaceous birds, such as the chicken, are among the
most sensitive of avian species to the effects of PCBs (Table B-13), an interspecies uncertainty factor
is not applied. Because by nature, a hatching period is a short-term event, a subchronic-to-chronic
uncertainty factor is not applied. These TRVs are expected to be protective of the bald eagle.
On the basis of laboratory toxicity studies:
B-34 TAMS/MCA
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The LOAEL TRY for the bald eagle is 2.21 mg PCBs/kg egg (Table B-26).
The NOAEL TRV for the bald eagle is 0.33 mg PCBs/kg egg (Table B-26).
Several field studies were identified that examined the effects of PCBs in eggs of bald eagles
(Table B-14). Clark et al. (1998) presented information on concentrations of total PCBs (range =
20 to 54 mg/kg egg) and TEQs in eggs from two sites in New Jersey where reproductive failures
have occurred, but the data could not be used to establish NOAEL or LOAELs. Studies by
Wiemeyer et al. (1984,1993) reported adverse effects on mean 5-year production in bald eagle with
egg concentrations greater than 3.0 mg PCBs/kg egg. Because significant intercorrelation of many
contaminants made it difficult to determine which contaminants had cause the adverse effects
(Wiemeyer, 1993), these studies can not be used to establish a field-based LOAEL for the effects of
PCBs. However, a field-based NOAEL of 3.0 mg PCBs/kg egg can be established on the basis of
this study for the bald eagle (Wiemeyer et al. , 1993). This NOAEL is expected to be protective of
the bald eagle.
On the basis of field toxicity studies:
The NOAEL TRV for the bald eagle is 3.0 mg PCBs/kg egg (Table B-26).
Total Dioxin Equivalents (TEQs) in Eggs of the Bald Eagle
No laboratory studies were identified that examined the toxicity of dioxin-like compounds
in the eggs of the bald eagle or for eggs of a bird in the same taxonomic family as the bald eagle
(Table B-15 and Figure B-7). Therefore, the lowest appropriate NOAEL (0.01 ug TEQs/kg egg) and
corresponding LOAEL (0.02 jag TEQs/kg egg) from the applicable studies (Table B-15) are used to
develop TRVs for the bald eagle. Powell et al. (1996a) found significantly reduced hatchability in
domestic chicken eggs that were injected with 0.2 jag BZ#126/kg egg. This effect was not observed
in eggs injected with 0.1 jag BZ#126/kg egg. The effective concentrations of BZ#126 are multiplied
by the avian TEF for BZ#126 (0.1) to estimate TRVs on a dioxin basis. Because gallinaceous birds,
such as the chicken, are among the most sensitive of avian species to the effects of dioxin-like
compounds (Table B-15), an interspecies uncertainty factor is not applied. Because by nature, a
hatching period is a short-term event, a subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRV for the bald eagle is 0.02 ug TEQs/kg egg (Table B-26).
The NOAEL TRV for the bald eagle is 0.01 pg TEQs/kg egg (Table B-26).
A field study by Clark et al. (1998) presented information regarding concentrations of TEQs
(range = 0.513 to 1.159 ug/kg) in bald eagle eggs from two sites in New Jersey where reproductive
failures have occurred. However, these data were not detailed enough to establish NOAEL TRV.
B-35 TAMS/MCA
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B.2.5 Selection of TRVs for Mammalian Receptors
B.2.5.1 Little brown bat (Myotis lucifugus)
Total PCBs in the Diet of the Little Brown Bat
No laboratory studies that examined the effects of PCBs on bats or on a species in the same
taxonomic family or order as the bat were identified (Table B-17 and Figure B-9). Therefore, the
lowest appropriate NOAEL (0.32 mg/kg/day) and corresponding LOAEL (1.5 mg/kg/day) from the
applicable studies (Table B-17) are selected for the development of TRVs for the little brown bat.
The study by Linder et al. (1974) is selected over other studies because it is a multigenerational
study, and thus more robust. In this study, mating pairs of rats and their offspring were fed PCBs in
the diet. Offspring of rats fed Aroclor 1254 at a dose of 1.5 mg/kg/day exhibited decreased litter size
in comparison to controls. This effect was not observed at a dose of 0.32 mg/kg/day. An uncertainty
factor of 10 is applied to account for potential differences in sensitivity to PCBs between the rat and
the little brown bat (Table B-27). Because of the extended duration of the experimental study (2
generations) a subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the little brown bat is 0.15 mg PCBs/kg/day (Table B-27).
The NOAEL TRY for the little brown bat is 0.032 mg PCBs/kg/day (Table B-27).
Several field studies were identified that examined the effects of PCBs on bats (Clark, 1978,
Clark and Krynitsky, 1978; Clark and Lament, 1976). However, these studies are not used to select
TRVs because effect endpoints in these studies are reported on the basis of concentrations of PCBs
in bat tissue, rather than as dietary doses. No field studies were identified that examined effects of
dietary exposure to PCBs on growth, reproduction, or mortality of the little brown bat or on a species
in the same family as the little brown bat. These studies are not presented in a table due to their
overall lack of relevance to the development of TRVs for mammals.
Total Dioxin Equivalents (TEQs) in the Diet of the Little Brown Bat
No laboratory studies were identified that examined effects of dioxin-like compounds on bats
bats or on a species in the same taxonomic family or order as the bat were identified (Tables B-18
and Figure B-10). Therefore, the multigenerational study by Murray et al. (1979) is selected to
derive the TRV for the little brown bat. The study by Murray et al. (1979) was selected over the
study of Bowman et al. , (1989b) on rhesus monkeys because the length of exposure was
significantly longer than that used in the rhesus monkey study. Murray et al. (1979) reported a
LOAEL of 0.01 |ag/kg/day and a NOAEL of 0.001 ug/kg/day for adverse reproductive effects in the
rat. An uncertainty factor of 10 is applied to account for potential differences between the rat and the
little brown bat in sensitivity to dioxin-like compounds. Because the experimental study examined
over three generations, a sub-chronic-to-chronic uncertainty factor is not applied.
B-36 TAMS/MCA
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On the basis of laboratory toxicity studies:
The LOAEL TRY for the little brown bat is 0.001 ug TEQs/kg/day (Table B-27).
The NOAEL TRY for the little brown bat is 0.0001 ug TEQs/kg/day (Table B-27).
Note that the study by Murray et al. (1979) was also selected by the USEPA as the basis for
development of concentrations of 2,3,7,8-TCDD associated with risk to mammalian receptors
(USEPA, 1993).
No field studies were identified that examined effects of dietary exposure to dioxin-like
compounds on growth, reproduction, or mortality of the little brown bat or on a species in the same
taxonomic family as the little brown bat.
B.2.5.2 Raccoon (Procyon lotor)
Total PCBs in the Diet of the Raccoon
One study was identified that examined acute effects (8-day exposure) of PCBs on the growth
of raccoons (Montz et al. , 1982). Because of the difficulty in estimating chronic LOAELs and
NOAELs from acute studies, this study is not used to estimate TRVs for the raccoon.
No appropriate experiments that examined the effects of PCBs on raccoons or on species in
the same taxonomic family or order were identified (Table B-17 and Figure B-9). Therefore, the
lowest appropriate NOAEL (0.32 mg/kg/day) and corresponding LOAEL (1.5 mg/kg/day) from the
selected applicable mammalian studies (Table B-17) are selected for the development of TRVs for
the raccoon. The study by Linder et al. (1974) is selected over other studies because it is a robust
multigenerational study, in which mating pairs of rats and their offspring were fed PCBs in their
diets. Offspring of rats fed Aroclor 1254 at a dose of 1.5 mg/kg/day exhibited decreased litter size
in comparison to controls. This effect was not observed at a dose of 0.32 mg/kg/day.
Because acute effects of PCBs on raccoons (Montz et al. 1982, Table B-17) are not directly
comparable to sub-chronic or chronic effects of PCBs on the rat, the sensitivities of the two species
to PCBs cannot be compared. Therefore, an uncertainty factor of 10 is applied to account for
potential differences in sensitivity to PCBs between the rat and the raccoon. Because of the extended
duration of the experimental study (two generations), a subchronic-to-chronic uncertainty factor is
not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the raccoon is 0.15 mg PCBs/kg/day (Table B-27).
The NOAEL TRY for the raccoon is 0.032 mg PCBs/kg/day (Table B-27).
B-37 TAMS/MCA
-------
No field studies were identified that examined effects of dietary exposure to PCBs on growth,
reproduction, or mortality of the raccoon or on a species in the same taxonomic family as the
raccoon.
Total Dioxin Equivalents (TEQs) in the Diet of the Raccoon
No studies were identified that examined effects of dioxin-like compounds on raccoons or
a species in the same taxonomic family as the racoon (Table B-18). Therefore, the multigenerational
study by Murray et al. (1979) is selected to derive the TRY for raccoons. Murray et al. (1979)
observed reduced reproductive capacity in two generations of offspring of the rats that were exposed
to 2,3,7,8-TCDD in the diet (Table B-18). Murray etal. (1979) reported a LOAEL of 0.01 Mg/kg/day
and a NOAEL of 0.001 ug/kg/day for these reproductive effects. An uncertainty factor of 10 is
applied to account for potential differences between the rat and the raccoon in sensitivity to dioxin-
like compounds. Because the experimental study examined exposure over three generations, a
subchronic-to-chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the raccoon is 0.001 ug TEQs/kg/day (Table B-27).
The NOAEL TRY for the raccoon is 0.0001 ug TEQs/kg/day (Table B-27).
No field studies were identified that examined effects of dietary exposure to dioxin-like
compounds on growth, reproduction, or mortality of the raccoon or on a species in the same
taxonomic family as the raccoon.
B.2.5.3 Mink (Mustela vison)
Total PCBs in the Diet of the Mink
Numerous studies have evaluated the effects of total PCBs on mortality, growth and
reproduction in mink (Table B-19 and Figure B-8). The lowest effective dose in the selected
applicable studies (Table B-19) (Platanow and Karstad, 1973) is not selected for development of
TRVs because that study compared growth and reproduction of PCB-treated mink to the
performance of an institutional herd of mink, rather than to a true experimental control group.
Instead, the study of Aulerich and Ringer (1977) is selected for calculating TRVs for the mink. In
this study, reproduction was markedly reduced when female mink were fed Aroclor 1254 at a dose
of 0.7 mg/kg/day for a period of 4 months. These effects were not observed at a dose of 0.1
mg/kg/day. A subchronic-to-chronic uncertainty factor of 10 is applied to the reported LOAEL and
NOAEL to account for the short exposure duration of the study.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the mink is 0.07 mg PCBs/kg/day (Table B-27).
The NOAEL TRY for the mink is 0.01 mg PCBs/kg/day (Table B-27).
B-38 TAMS/MCA
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Two field studies were identified that examined effects of PCBs in the diet of the mink
(Table B-20). The study that reported a lack of adverse reproductive effects at the lowest dose is used
to develop TRVs for the mink. Adult ranch mink were fed diets containing various amounts of PCB-
contaminated carp from Lake Michigan (Heaton et al. , 1995). Mink fed the contaminated diet
before and during reproduction had reduced reproduction and/or growth and survival of offspring.
Concentrations of other contaminants were measured and were substantially lower than
concentrations of PCBs. The dietary LOAEL was 0.13 mg PCBs/kg/day. The dietary NOAEL was
0.004 mg PCBs/kg/day. Because of the extended period of exposure (128 days) a subchronic-to-
chronic uncertainty factor is not applied.
On the basis of field toxicity studies:
The LOAEL TRY for the mink is 0.13 mg PCBs/kg/day (Table B-27).
The NOAEL TRY for the mink is 0.004 mg PCBs/kg/day (Table B-27).
This field study was accepted as appropriate for use in developing TRVs for the mink, and
these TRVs are accepted as final TRVs for the mink, rather than the laboratory-based TRVs.
Total PCBs in the Liver of the Mink
Two studies were identified that related concentrations of PCBs in the liver of mink to
adverse reproductive effects. Platanow and Karstad (1973) reported that a liver concentration of 1.23
mg/kg (weathered Aroclor 1254) corresponded to impaired reproductive success (as reported in
Wren, 1991). It should be noted, however, that reproductive success in the control group of that
study was also very poor in relation to that of control groups in other experiments. Reduced growth
of mink kits was observed in female mink with 3.1 mg Aroclor 1254/gm liver (Wren et al. , 1987).
Total Dioxin Equivalents (TEQs) in the Diet of the Mink
Two studies were identified that examined acute effects (12- and 28-day exposures) of
dioxin-like compounds on mink (Hochstein et al. , 1988, Aulerich et al. , 1988) (Table B-18).
Because of the difficulty in estimating chronic LOAELs and NOAELs from acutely lethal doses,
these studies are not used to derive TRVs for the effects of dioxin-like compounds on the mink.
Instead, the study by Murray et al (1979) is selected to derive TRVs for mink (Table B-18). Murray
et al. (1979) observed reduced reproductive capacity in two generations of the offspring of rats that
were exposed to 2,3,7,8-TCDD in the diet. This study was selected over the study of Bowman et al.
, (1989b) on rhesus monkeys because: (1) the length of exposure was significantly longer than that
used in the rhesus monkey study, and (2) information on the short-term toxicity (LD50) of 2,3,7,8-
TCDD to the rat and the mink (Tables B-18, B-21) helps indicate the sensitivity of these two animals
relative to one another. This data indicates that the mink is much more sensitive than the rat, so an
inter-order uncertainty factor should be applied. Murray et al. (1979) reported a LOAEL of 0.01
ug/kg/day and a NOAEL of 0.001 ug/kg/day for reproductive effects in rats. An uncertainty factor
of 10 is used to account for the extreme sensitivity of the mink in comparison to the rat. Because the
B-39 TAMS/MCA
-------
experimental studies examined exposure to 2,3,7,8-TCDD over three generations, a subchronic-to-
chronic uncertainty factor is not applied.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the mink is 0.001 jag TEQs/kg/day (Table B-27).
The NOAEL TRY is for the mink is 0.0001 ug TEQs/kg/day (Table B-27).
Two field studies were identified which examined effects of dioxin-like compounds on
reproduction and survival in mink (Table B-22). The study that reports adverse reproductive effects
at the lowest dose is used to develop TRVs for the mink. In this study, mink were fed diets
containing contaminated carp from Lake Michigan (Tillitt etal., 1996). Concentrations of TEQs in
the food was quantified by two methods: standard analytical chemistry and with a bioassay
conducted on an extract of the food. The growth rate of kits born to the adults that were fed the carp
diet were significantly reduced in comparison to controls. This effect was observed at a dose of
0.00224 ug/kg/day, but not at a dose of 0.00008 ug/kg/day. TEQs calculated by Tillitt et al (1996)
are estimated to be 12% higher than the concentration of TEQs that would be calculated using the
TEFs of van den Berg et al. (1998) that are used in the present report.
On the basis of field toxicity studies:
The LOAEL for the mink is 0.00224 ug TEQs/kg/day (Table B-27).
The NOAEL for the mink is 0.00008 ug TEQs/kg/day (Table B-27).
B.2.5.4 River Otter (Lutra canadensis)
Total PCBs in the Diet of the River Otter
No studies were identified that examined the toxic effects of PCBs on otters (Table B-17 and
Figure B-9). Because river otter and mink are in the same phylogenetic family (Table B-23), the
LOAEL TRY (0.07 mg Aroclor 1254/kg/day) and NOAEL TRY (0.01 mg Aroclor 1254/kg/day) for
the mink are used to develop TRYs for the otter. Since mink are generally considered to be among
the most sensitive of mammalian species and otter are not expected to be more sensitive, the
interspecies uncertainty factor is set to 1.
On the basis of laboratory toxicity studies:
The LOAEL TRY for the river otter is 0.07 mg PCBs/kg/day (Table B-27).
The NOAEL TRY for the river otter is 0.01 mg PCBs/kg/day (Table B-27).
Because river otters are closely related to mink, the field studies that examined effects of
dietary exposure to PCBs to mink are used to develop TRVs for the river otter. Two field studies
were identified that examined effects of PCBs in the diet of the mink (Table B-20). The study that
reported adverse reproductive effects at the lowest dose is used to develop TRVs for the mink and
the otter. Adult ranch mink were fed diets containing various amounts of PCB-contaminated carp
B-40 TAMS/MCA
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(Heaton et al., 1995). Mink fed the contaminated diet before and during reproduction had reduced
reproduction and/or growth and survival of offspring. Concentrations of other contaminants were
measured and were substantially lower than concentrations of PCBs. The dietary LOAEL was 0.13
mg PCBs/kg/day. The dietary NOAEL was 0.004 mg PCBs/kg/day.
On the basis of field studies:
The LOAEL TRY for the river otter is 0.13 mg PCBs/kg/day (Table B-27).
The NOAEL TRV for the river otter is 0.004 mg PCBs/kg/day (Table B-27).
Total Dioxin Equivalents (TEQs) in the Diet of the River Otter
No studies were identified that examined effects of dioxin-like compounds to otters or on a
species in the same taxonomic family as the otter (Table B-18 and Figure B-10). The multi-
generational study by Murray et al. (1979), which was selected as appropriate for the mink, is
selected to derive TRVs for the closely related river otter. The study of Murray et al., (1979) was
selected over the study of Bowman et al. (1989b) on rhesus monkeys because the length of exposure
was significantly longer than that used in the rhesus monkey study. Murray et al. (1979) reported a
LOAEL of 0.01 (ag/kg/day and a NOAEL of 0.001 ug/kg/day for adverse reproductive effects in the
rat. Because of the lack of any acute or chronic toxicity data for effects of dioxin-like compounds
on the river otter, an uncertainty factor of 10 is applied to account for potential differences in
sensitivity to dioxin-like compounds between the rat and the river otter. Because the experimental
study examined exposure over three generations, a subchronic-to-chronic uncertainty factor is not
applied.
On the basis of laboratory toxicity studies:
The LOAEL TRV for the river otter is 0.001 ug TEQs/kg/day (Table B-27).
The NOAEL TRV for the river otter is 0.0001 ug TEQs/kg/day (Table B-27).
Because otters are closely related to mink, the field studies that examined effects of dietary
exposure to dioxin-like compounds to mink are used to develop TRVs for the otter. Two field studies
were identified that examined effects of dioxin-like compounds on reproduction and survival in mink
(Table B-22). The study that reports adverse reproductive effects at the lowest dose is used to
develop TRVs for the otter. In this study, mink were fed diets containing contaminated carp from
Lake Michigan (Tillitt et al., 1996). Concentrations of TEQs in the food was quantified by two
methods: standard analytical chemistry and with a bioassay conducted on the extract of the food. The
growth rate of kits born to the adults that were fed the carp diet were significantly reduced in
comparison to controls. This effect was observed at a dose of 0.00224 ug/kg/day, but not at a dose
of 0.00008 ug/kg/day. TEQs calculated by Tillitt et al. (1996) are estimated to be 12% higher than
the concentration of TEQs that would be calculated using the TEFs of van den Berg et al. (1998)
that are used in the present report. Because mink and river otter are in the same taxonomic family,
an interspecies uncertainty factor is not applied. Because of the extended exposure period of the
study (182 days) a subchronic-to-chronic uncertainty factor is not applied.
B-41 TAMS/MCA
-------
On the basis of field toxicity studies:
The LOAEL TRY for the river otter is 0.00224 ng TEQs/kg/day (Table B-27).
The NOAEL TRV for the river otter is 0.00008 |ag TEQs/kg/day (Table B-27).
B-42 TAMS/MCA
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McFarland, V.A. and J.U. Clarke. 1989. Environmental occurrence, abundance and potential
toxicity of polychlorinated biphenyl congeners: considerations for a congener-specific analysis.
Environ. Health Perspectives 81:225-239.
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Montz, W.E., W.C. Card, and R.L. Kirkpatrick. 1982. Effects of polychlorinated biphenyls and
nutritional restriction on barbituate-induced sleeping times and selected blood characteristics in
raccoons (Procyon lotor). Bull. Environ. Contam. Toxicol. 28:578-583.
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Murray, F.J., F.A. Smith, K.D. Nitschke, C.G. Huniston, R.J. Kociba, and B.A. Schwetz. 1979.
Three-generation reproduction study of rats given 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in
the diet. Toxicol. Appl. Pharmacol. 50:241-252.
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Evaluation of Consensus-Based Sediment Effect Concentrations for PCBs in the Hudson River.
B-45 TAMS/MCA
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Prepared for NOAA Damage Assessment Center, Silver Spring, MD. Prepared through Industrial
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National Oceanographic and Atmospheric Administration (NOAA). 19995. Reproductive,
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Studies. Prepared for NOAA Damage Assessment Center, Silver Spring, MD. Prepared through
Industrial Economics Inc. by E. Monosson. March, 1999.
Niimi, A.J. 1996. PCBs in Aquatic Organisms. In Environmental Contaminants in Wildlife:
Interpreting Tissue Concentrations. W. Beyer, G.H. Heinz, and A.W. Redmon-Norwood (eds.).
Lewis Publishers. Boca Raton, FL.
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reproductive effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin toxicity in ring-necked pheasant hens.
J. Toxicol. Environ. Health. 35:187-198.
Olivieri, C.E., and K.R. Cooper. 1997. Toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)
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1150.
Platonow, N. and C. Karstad. 1973. Dietary effects of polychlorinated biphenyls on mink. Can. J.
Comp. Med. 37:391-400.
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B-46 TAMS/MCA
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B-47 TAMS/MCA
-------
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dibenzofurans on nesting wood ducks (Aix Sponsa) at Bayou Meto, Arkansas. Environmental Health
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585.
B-48 TAMS/MCA
-------
TABLE B-l
COMMON EFFECTS OF PCB EXPOSURE IN ANIMALS
Hepatotoxicity
Hepatomegaly; bile duct hyperplasia, proliferation of smooth ER
Focal necrosis; fatty degeneration
Induction of microsomal enzymes; implications for hormone imbalances, pancreas and reproductive effects
Depletion of fat soluble vitamins (predominantly vitamin A)
Porphyria
Immunotoxicity
Atrophy of lymphoid tissues
Reduction in circulating leukocytes and lymphocytes
Suppressed antibody responses
Enhanced susceptibility to viruses
Suppression of natural killer cells
Neurotoxicity
Impaired behavioral responses
Alterations in catecholamine levels
Depressed spontaneous motor activity
Developmental deficits
Numbness in extremities
Reproduction
Increased abortion; low birth weights
Decreased survival and mating success
Increased length of estrus
Embryo and fetal mortality
Gross teratogenic effects
Biochemical, neurological, and functional changes following in utero exposure (mammals)
Decreased libido, decreased sperm numbers and motility
Gastrointestinal
Gastric hyperplasia
Ulceration and necrosis
Respiratory
Chronic bronchitis
Decreased vital capacity
Dermal Toxicity
Chloracne
Hyperplasia and hyperkeratosis of epithelium
Edema
Mutagenic Effects
Commercial mixtures are weakly mutagenic
Carcinogenic Effects
Preneoplastic changes
Neoplastic changes
Promotion considered main contribution
Attenuation of other carcinogens under certain conditions
Source: Hansen, L. G.. 1987. Environmental Toxicology of Polychlorinated Biphenyls in Environmental
Toxin Series 1. eds. Safe, S. and Hutzinger, O., p. 32.
TAMS/MCA
-------
TABLE B-2
WORLD-HEALTH ORGANIZATION FOR TOXIC EQUIVALENCY FACTORS (TEFs) FOR HUMANS,
MAMMALS, FISH, AND BIRDS
Congener
Non-ort/io PCBs
3,4,4',5-TetraCB (81)
S.S'.M'-TetraCB (77)
3,3',4,4',5-PentaCB (126)
3,3',4,4',5,5'-HexaCB (169)
Toxic Equivalency Factor
Humans/Mammals
0.0001
0.0001
0.1
0.01
Fish
0.0005
0.0001
0.005
0.00005
Birds
0.1
0.05
0.1
0.001
Mono-ortho PCBs
2,3,3',4,4'-PentaCB (105)
2,3,4,4',5-PentaCB(114)
2,3',4,4',5-PentaCB(118)
2',3,4,4',5-PentaCB (123)
2,3,3',4,4',5-HexaCB (156)
2,3)3',4,4',5'-HexaCB (157)
2,3',4,41,5,5'-HexaCB (167)
2)3,3',4,4',5,5'-HeptaCB(18<
0.0001
0.0005
0.0001
0.0001
0.0005
0.0005
0.00001
0.0001
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
<0.000005
0.0001
0.0001
0.00001
0.00001
0.0001
0.0001
0.00001
0.00001
Notes: CB = chlorinated biphenyls
Reference: van den Berg, et al. (1998). Toxic Equivalency Factors (TEFs) for PCBs
PCDDs, PCDFs for Humans and Wildlife. Environmental Health Perspectives,
106:12,775-791.
TAMS/MCA
-------
TABLE B-3
SELECTED SEDIMENT SCREENING GUIDELINES: PCBs
Hudson River Sediment Effect Concentrations (mg/kg, orppm)
(MacDonald Env. Sci., 1999)
(Estuarine. freshwater, and saltwater)
Threshold Effect Concentration
Mid-range Effect Concentration
Extreme Effect Concentration
NYSDEC {1998) (Freshwater) (mg/kg organic carbon)
Benthic Aquatic Life Acute Toxicity
Benthic Aquatic Life Chronic Toxicity
Wildife Bioaccumulation
NYSDEC (1998) (Saltwater) (mg/kg organic carbon)
Benthic Aquatic Life Acute Toxicity
Benthic Aquatic Life Chronic Toxicity
Wildlife Bioaccumulation
Ontario Ministry of the Environment Sediment Guidelines (Freshw
(Persaud et ai, 1993)
No Effect Level (mg/kg)
Lowest Effect Level (mg/kg)
Severe Effect Level (rag/kg organic carbon)
Long el at. (1995) Sediment Guidelines (ug/kg)
(Marine and Estuarine)
Effects-Range-Low
Effects-Range-Median
Ingersoll et al. (1996) Sediment Guidelines (ug/kg, orppb)
(Freshwater)
(Derived from 28-day Hyalella azteca data)
Effects-Range-Low
Effects-Range-Median
Threshold Effect Level
Probable Effect Level
No Effect Concentration
Washington State Dep't of Ecology 1997 Sediment Guidelines
(Freshwater) ( ug/kg, orppb) '
Apparent Effects Threshold (Microtox)
Apparent Effects Threshold (Hyalella azteca )
Probable Apparent Effects Threshold (Microtox)
Probable Apparent Effects Threshold (Hyalella azteca )
Lowest Apparent Effects Threshold
(between Microtox and H. aiteca )
Florida Department of Environmental Protection (ug/kg, orppb)
(MacDonald, D.D., et ai, 1996) (Marine and Estuarine,
Threshold Effect Level
Probable Effect Level
Jones et al. (1997) (ug/kg, orppb)
EqP-derived; recommended TOC adjustment
Secondary Chronic Value
Smith etal. (1996) (ug/kg, orppb)
Threshold Effect Level
Probable Effect Level
Total Aroclor Aroclor Aroclor Aroclor
PCBs 1254 1248 1016 1260
0.04
0.4
1.7
2760.8
19.3
1.4
13803.3
41.4
1.4
ater)
0.01
0.07
530
22.7
180
50
730
32
240
190
21
820
21
450
21
21.6
189
34.1
277
0.06
34
7.3
350
7.3
240
7.3
810
0.03
150
21
21
1000
0.007
53
Aroclor
1242
IOC
IOC
0.005
24
450000
Note: All values arc dry weight unless noted.
Please note that for Washington state values, the Aroclor 1016
column becomes Aroclor 1242. This applies only to this one set
of values.
1 Some values also available in mg/kg organic carbon
TAMS/MCA
-------
TABLE B-4
TOX1CITY ENDPOINTS FOR BENTHIC INVERTEBRATES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs, AROCLORS, AND DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Amphipod
(Gammarus psrutlalimnaeux )
Amphipod
(Hyalflla autcu )
Amphipod
(Hyalella aaeca )
Amphipod
(Gammarus psfuthtlimnatus )
EXPOSURE
MEDIA
Water
Water
Water
Water
PCB TYPE
Aroclor 1248
PCB 52
Aroclorl242
Aroclor 1242
EXPOSURE
DURATION
2 months
>or = 10 weeks
>or= 10 weeks
2 months
EFFECT LEVEL
LDM
LD,,,,
LD,,»,
LD»
EFFECT CONC,
WHOLE BODY CONC.
(mg/kg wet wl)
552
180
100
316
EFFECT ENDPOINT
Mortality
Mortality
Mortality
Mortality
REFERENCE
Nebeker and Puglisi ( 1974)
Borgmann et al. (1990)
Borgmann el al. (1990)
Nebeker and Puglisi (1974)
Cladoceran
(Dafthnia magna )
Amphipod
(Gammurus pseutlolimnaeus )
Snail
(Physa spp.)
Amphipod
(Gammarus pseudfilimnaeus )
Oligochaete
(Lumhriculus varitgatus )
Oligochaete
(Lumhriculus variegalus )
Oligochaete
(Lumhriculus variegalus )
Oligochaete
(Lumhriculus varifgatus )
Oligochaete
(Lumhriculus varitgatus )
Oligochaete
(Lumhriculus varifgalus )
Grass shrimp
(Palafmttnrlrs pugiit )
Oligochaete
(Lumbriculus varifgalus )
Oligochaete
(Lumhriculus varifgatus )
Grass shrimp
(Palatmimetex pugiir )
Model ecosystem
Water
Water
Water
Algae (Food)
Algae (Food)
Algae (Food)
Algae (Food)
Algae (Food)
Algae (Food)
Water
Algae (Food)
Algae (Food)
Water
2,.1,7.8-TCDD
Aroclor 1248
2,3,7.8-TCDD
Aroclor 1242
PCB 153
PCB 153
PCB 15
PCB 15
PCB 47
PCB 47
Aroclor 1 254
PCB 1
PCB 1
Aroclor 1254
33 days
2 months
33 days
2 months
35 days
35 days
35 days
35 days
35 days
35 days
7 days
35 days
35 days
EL (no effect)
LOAEL
EL (no effect)
EL (effect)
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
16 days LOAEL
Amphipod
(Gammaru.1 pxeuilitlimnaeiu )
Amphipod
(Gammarux pseudtilimnatm )
Oligochaete
(Lumbriculus variegalu,t )
Oligochaete
(Lumhriculus varifgatux )
Oligochaete
(Lumhriculus variegatus )
Water
Water
Algae (Food)
Algae (Food)
Algae (Food)
Aroclot 1248
Aroclor 1242
PCB 153
PCB 153
PCB 15
2 months
2 months
35 days
35 days
35 days
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
1570
552
502
316
126
126
119
119
113
113
65
64
64
27
127
76
65
65
63.1
Mortality
Reproduction reduced by at least
50%
Mortality
No reproduction
Mortality
Weight loss
Mortality
Weight loss
Mortality
Weight loss
Mortality (60%)
Mortality
Weight loss
Mortality (45%)
Isensee and Jones (1975)
Nebeker and Puglisi (1974)
Isensee and Jones (1975)
Isensee (1978)
Nebeker and Puglisi (1974)
Fisher ctal. (1998)
Fisher etal.( 1998)
Fisher ctal. (1998)
Fisher etal. (1998)
Fisher etal. (1998)
Fisher etal. (1998)
Nimmo etal. (1974)
Fisher etal. (1998)
Fisher etal. (1998)
Nimmo etal. (1974)
Reproduction
Reproduction
Mortality
Weight loss
Mortality
Nebeker and Puglisi (1974)
Nebeker and Puglisi (1974)
Fisher etal. (1998)
Fisher etal. (1998)
Fisher etal. (1998)
Page I of 2
TAMS/MCA
-------
TABLE B-4
TOXIC1TY ENDPOINTS FOR BENTHIC INVERTEBRATES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs, AROCLORS, AND DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Oligochaete
(Lumbriculus variegatus )
Amphipod
(Hyalella azteca )
Oligochaele
(Lumbriculus varifgatux )
Oligochacte
(Lumbriculus variegatus )
Oligochaele
(Lumbriculus variegatus)
Oligochaete
(Lumbricutus varitgatus )
Amphipod
(Hyalella azteca )
Grass shrimp
(Palurmtmetex pugin )
Grass shrimp
(Palaenumetes pugio )
EXPOSURE
MEDIA
Algae (Food)
Water
Algae (Food)
Algae (Food)
Algae (Food)
Algae (Food)
Water
Water
Water
PCB TYPE
PCB 15
PCB 52
PCB 47
PCB 47
PCB 1
PCB 1
Aroclor 1242
Aroclor 1254
Aroclor 1255
EXPOSURE
DURATION
35 days
> or = 10 weeks
35 days
35 days
35 days
35 days
>or = 10 weeks
16 days
7 days
EFFECT LEVEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
EFFECT CONC,
WHOLE BODY CONC.
(mg/kg wel wl)
63.1
54
49.3
49.3
33.2
33.2
30
IS
5.4
EFFECT ENDPOINT
Weight loss
Mortality
Mortality
Weight loss
Mortality
Weight loss
Mortality
Mortality
Mortality
REFERENCE
Fisher etal. (1998)
Borgmann et al. ( 1 990)
Fisher etal. (1998)
Fisher etal. (1998)
Fisher etal. (1998)
Fisher etal. (1998)
Borgmann et al. ( 1 990)
Nimmo etal. (1974)
Nimmo etal. (1974)
Page 2 of 2
TAMS/MCA
-------
TABLE B-5
TOXICITY ENDPOINTS FOR FISH - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs AND AROCLORS
SPECIES
EXPOSURE MEDIA
PCB TYPE
EXPOSURE
DURATION
EFFECT LEVEL
EFFECT
CONCENTRATION
WHOLE BODY
CONCENTRATION
tttfflig wet wt.
EFFECT ENDPOINT
REFERENCE
Laboratory studies
Lake trout
(Salvelinux namayfush )
Chinook salmon
(Oncorhnchus tshawytscha)
Water
Water
PCB-153
PCB- 153
15 days
15 days
LD100
LDIOO
7.6
3.6
Fry modality
Fry mortality
Broyles and Noveck, 1979
Broyles and Noveck, 1979
Adult Fathead Minnow
(Pimephaltx promelas )
Adult Fathead Minnow
'Pimephalts promrlas )
Srook trout fry
(Salvr linu.1 funtinalis )
2
Brook trout fry
[Salvtlinwi fontinalLi )
Junvenile Spot
\Lewsntmus xanlhurus )
Adult pinfish
[Lugfuhtn rhttmhftiilrx )
Adult Minnow
(Pht)xinu.i phfixinux )
Killifish
(Fundulu\ htteroclitux )
Sheepshead minnow
(Cyprirwditn varitgatwi )
Lake trout fry
(Salnw gairdneri )
Killiflsh
(Fundulux hrtcnfclitus )
Adult Fathead Minnow
(Pimephales prnmelia )
Adult Fathead Minnow
(Pimfphalex promelwt )
Adult pinfish
(Lagiidtin rhttmbttidts )
Adult Fathead Minnow
(Pimtphale.1 priimrltu )
Brook trout fry
(Salvtliniu fiintinalix )
Juvenile Spot
(Lfiiistomux xanthurux )
Water
Water
Water
Water '
Water
Water
Water
Diet
Single inlraperitoneal
injection into adults
Water
Water
Single intraperitoneal
injection into adults
Water
Water
Water
Water
Water
Water
Aroclor 1254
Aroclor 1254
Aroclor 1254
Aroclor 1 254
Aroclor 1254
Aroclor 1254
Aroclor 1016
Clophcn A50
PCB mixture
Aroclor 1254
Aroclor 1254
PCB mixture
Aroclor 1242
Aroclor 1254
Aroclor 1016
Aroclor 1254
Aroclor 1254
Aroclor 1254
9 months
9 months
118 days
21 days
21 days
20 days
42 days
40 days; studied for
300 days
Single injection,
40 d of observation
28 days
48 days
Single injection. 40
days of observation
9 months
9 months
42 days
9 months
118 days
Lab Stu
LOAEL
LOAEL
LOAEL
EL-effect
EL-effect
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
EL-effect
LOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
999
429
125
32.8 in muscle
77.9 in eggs
46
42
15
19
(nominal dose)
9.3
4.5
3.8
(nominal dose)
436
429
170
105
71
27
Adult mortality
Spawning
Fry mortality
Egg hatchabilty
Egg hatchability
Adult mortality
Adult mortality
Hatching time; fry survival
Adult female mortality
Fry mortality
Fry mortality
Egg production and food
consumption
Adult mortality
Egg hatchability
Adult mortality
Spawning
Fry mortality
Adult mortality
Nebekeretal., 1974
Nebekeretal., 1974
Maucketal., 1978
Freeman and Idler. 1974
Freeman and Idler, 1 974
Hansen et al., 1971
Hansenetal., 1974
Bengtsson, B , 1980
Black etal., I998a
Hansenetal., 1974
Mac and Seelye, 1981
Black etal., 1998a
Nebekeretal., 1974
Nebekeretal., 1974
Hansen et al., 1974
Nebekeretal., 1974
Maucketal.. 1978
Hansen el al., 1971
12/27/99
Page 1 of 2
TAMS/MCA
-------
TABLE B-5
TOXICITY ENDPOINTS FOR FISH - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs AND AROCLORS
SPECIES
Killifish
(Furululus hetrntflitus )
Sheepshead minnow
(Cyprinttilim variegatus)
Adult Minnow
(Phttxinus phfixinu.i )
Killifish
(Fundulux heterttclitux )
EXPOSURE MEDIA
Single intraperitoneal
injection into adults
Water
Diet
Single intraperitoneal
injection into adults
PCB TYPE
PCB mixture
Aroclor 1254
Clophen A50
PCB mixture
EXPOSURE
DURATION
Single injection, 40
days of observation
28 days
40 days; studied for
300 days
Single injection, 40
days of observation
EFFECT LEVEL
NOAEL
NOAEL
NOAEL
NOAEL
EFFECT
CONCENTRATION
WHOLE BODY
CONCENTRATION
Rlg/ltK wrt wt.
3.8
(nominal dose)
1.9
1.6
0.76
(nominal dose)
EFFECT ENDPOINT
Adult female mortality
Fry mortality
Hatching time; fry survival
Egg production and food
consumption
REFERENCE
Black etal.. I998a
Hansenetal., 1974
Bengtsson, B., 1980
Black etal., 1998a
12/27/99
Page 2 of 2
TAMS/MCA
-------
TABLE B-6
TOXICITY ENDPOINTS FOR FISH - HELD STUDIES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs AND AROCLORS
SPECIES
FIELD COMPONENT
CONTAMINANT
TYPE
EFFECT LEVEL
EFFECT
CONCENTRATION
mg/kg wet wt
(or as noted bdow)
EFFECT ENDPOINT
REFERENCE
Field studies
Arctic chair
[Salvetinux atpinux )
Winter flounder
Pxeuditpleurtmectes amerii-anus )
Killifish
Fundulux helertfclitux )
Killiflsh
Fundulux heleroclitux )
English sole
Paruphryx vetulux )
Striped bass
Mitnme xaxatilix )
Chinook salmon
(Oncttrhynchux txnawytxcha )
Chinook salmon
(Ont-iirhynchux txhawyixcha )
Rainbow trout
(Salrrui gairdneri )
English sole
(Pantphryx vetulux )
Lake trout
(Sulvelinux namaycuxh )
Chinook salmon
(Oncitrhynchux ixhawytxcha )
Starry flounder
(Plalichthyx xltllalux )
ledbreast sunflsh
(Lfpumix auritux )
Baltic herring
(Clupea harengux )
Baltic flounder
(Platichlfiyx flexux )
Adult fish and eggs collected
from Lake Geneva
Adult and eggs collected
from New Bedford Harbor
Fish collected
from New Bedford Harbor
Fish collected
from New Bedford Harbor
Fish collected
from Puget Sound
Eggs from hatcheries. Larvae fed
naturally contaminated food.
Adult fish and eggs collected
from Lake Michigan
Adult fish and eggs collected from
Lake Michigan
Adult fish and eggs
hatchery
Adults and eggs collected
from Puget Sound
Adult fish and eggs collected
from Great Lakes
Adult fish and eggs collected
from Lake Michigan
Adult fish and eggs collected from
area of San Francisco Bay
Adult fish collected from
East Tennessee stream
Adult fish and eggs
collected from Baltic Sea
Adult fish and eggs
collected from Baltic Sea
PCBs
DDT
PCBs
PCBs
PCBs
PCBs, PAHs
PCBs. HCB,
pesticides
PCBs,
pesticides
PCBs
PCBs, DDT
PCBs
PCBs
PCBs,
pesticides
PCBs, HCB, Plhalates
PCBs. PAHs. metals,
chlorine
PCBs,
pesticides
PCBs.
pesticides, metals
Killiflsh
(Fundulux hrtcraclilus )
Striped bass
(Moron* xaxatilix )
Winter flounder
(Pseudoplturonectex americanus )
English sole
(Pantphryx vetulux )
Fish collected
from New Bedford Harbor
Eggs from Hudson River fish.
Larvae fed naturally contaminated
food
Adult and eggs collected
from New Bedford Harbor
Adults and eggs collected
from Pugel Sound
PCBs
PCBs
PCBs
PCBs
EL-cffect
EL-effcct
LOAEL
LOAEL
EL-effect
EL-effect
EL-effect
El-effect
EL-effect
LOAEL
EL-effect
EL-effect
EL-effect
EL-effect
EL-effect
EL-cffecl
NOAEL
EL-no effect
EL-no effect
NOAEL
10to78 mg/kg lipid
in eggs
39.6 mg/kg dry wt
in eggs
29.2 mg/kg dry wt
in liver
20.8 mg/kg dry wt
in liver
Approx. 10 mg/kg
in liver
0.1 to 10 in eggs
2.8 to 9.9
A- 1 254 in eggs
2.75 to 5.75 in eggs
2.7 in eggs
2.56 in liver
0.25 to 7.77
in eggs
0.322 to 2.6
A- 1260 in eggs
about 50 tO 200
in eggs
0.95
> 0.120
in ovaries
> 0.120
in ovaries
9.5 mg/kg dry wt
in liver
3.1 in
post yolk sac larvae
1 .08 mg/kg dry wt
in eggs
0.09 in liver
Embryomortality
Growth rate of larvae
Embryo and larval survival
Adult female mortality
Increased fecundity
Larval mortality
Hathcing success
Hatching success
Embryomortality
Prodcution of normal larvie
Egg mortality and
percent of normal fry hatching
Hathcing success
Hathcing success
Fecundity, clutch size, growth
Hathcing success
Hathcing success
Embryo and larval mortality
Larval mortality
Growth rate of larvae
Prodcution of normal larvie
Monod. 1985
Black etal., 1988b
Black etal., I998b
Black etal.. 1998b
Johnson et al., 1997
Westinetal., 1985
Giesyetal., 1986
Ankleyetal., 1981
Hogan and Braun, 1975
Casillasetal., 1991
Mac etal., 1993
Giesyetal., 1986
Spies and Rice, 1988
Adams etal.. 1989. 1990. 1992
Hansen et el., 1985
von Westernhagen et al.. 1981
Black etal.. 1998b
Westin et al., 1983
Black etal., 1988b
Casillasetal.. 1991
12/27/99
Page 1 of 2
TAMS/MCA
-------
TABLE B-6
TOXICITY ENDPOINTS FOR FISH - FIELD STUDIES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs AND AROCLORS
Redbreast sunfish
(Ltintmis auritus )
Killifish
(Fundulus helentclilux )
Arctic chair
(Salvetinux alpinus )
Fish from an East Tennessee
stream
Fish collected
from New Bedford Harbor
Adult fish and eggs
collected from Lake Geneva
PCBs. PAHs, metals.
chlorine
PCBs
PCBs
DDT
EL-no effect
NOAEL
EL- no effect
0.5
0.461 mg/kgdry wt
in liver
O.I to 0.31
in eggs
Fecundity, clutch size, growth
Adult female mortality
Embryomortality
Adams etal., 1989, 1990. 1992
Black etal., I998b
Monod, 1985
12/27/99
Page 2 of 2
TAMS/MCA
-------
TABLE B-7
TOXICTTY ENDPOINTS FOR FISH - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
EXPOSURE
MEDIA
EFFECT
LEVEL
TISSUE
CONTAMINANT
TYPE
EFFECT
CONC.
(ug/kg ww)
LIPID CONTENT
OF EGG
(g lipid/gww egg)
TEF
EFFECT CONC. DIOXIN
EQUIVALENTS
(ugTEQ/kglipid)
EFFECT ENDPOINT
REFERENCE
Laboratory studies*
Fathead minnow
(Pimephales promelas )
Zebrafish
(Danio danio )
Zebrafish
(Danio danio)
While sucker
(Carastomus commersuni )
Northern Pike
(Estix lucius )
Mcdaka
(Oryzias latiptx )
Fathead minnow
[Pimephates prontelax )
Lake herring
(Coregonus artedii )
Channel catfish
(Ictalurus punctalus )
Rainbow Trout
(Salmo gairderi ) - Erwin strain
Rainbow Trout
(Salmo gairderi ) - Erwin strain
Bnx>k Trout
(Salvcnius foatinatis)
Rainbow Trout
(Salmo gairderi ) - Erwin strain
Rainbow Trout
(Salmo gairderi )
Rainbow Trout
(Salmo gairderi )
Brmik Trout
(Salvenius fontinulix)
(Salmo gairdneri )
Erwin strain
Lake trout
(Salvenius namavcush)
Fathead minnow
(Pimcphalcs promclas)
Lake trout
(Salvenius natnaycuxh)
Lake trout
(Salvenius nama\cu.\h)
Lake trout
(Salvenius nama\cush)
Fathead minnow
(Pimcphalcs promclas)
Water
Water
. Water
Water
Water
Water
Water
Water
Water
Water
Injection
Water
Egg injection
Egg injection
Egg injection
Water
Egg injection
Water
Water
Water
Water
Injection
Water
LD50
LD50
LD50
LD50
LDSO
LD50
LD50
LD50
LDSO
LD50
LD50
LDIOO
LDSO
LDSO
LD50
LDSO
LD50
LDSO
LDSO
LDSO
LDSO
LDSO
LDIOO
Embryo
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Egg
Embryo
Egg
Egg
Egg
Larvae
2.3,7.8-TCDD
2.3.7,8-TCDD
2,3.7.8-TCDD
2.3,7.8-TCDD
2,3,7.8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
2.3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2.3.7.8-TCDD
2,3,7,8-TCDD
PCB 126
2.3.7,8-TCDD
2,3.7.8-TCDD
PCB 126
2,3.7.8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
2.3.7.K-TCDD
2,3.7.8-TCDD
25.7
2.61
2.5
1.89
2.46
l.ll
0.539
0.902
0.644
0.439
0.421
0.324
0.409
0.374
74
0.200
0.242
29
0.026
0.085
0.065
0.047
163
0.024
0.017
0.017
0.025
0.042
0.029
0.024
0.066
0.048
0.087
0.087
0.068
0.087
0.087
0.087
0.068
0.087
0.08
0.024
0.08
0.08
0.08
Not reported for larvae
1
1
1
1
1
1
1
1
1
1
1
1
1
1
0.005
1
1
0.005
1
1
1
1
I
1071
154
147
76
59
38
22
14
13
5.0
4.8
4.8
4.7
4.3
4.3
2.9
2.8
1.8
I.I
I.I
0.8
0.6
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Early life stage mortality
Olivier! and Cooper, 1997 k
Eloncn el al.. 1998
Henry eta!.. 1997
Elonen ct al.. 1998
Elonen ct al.. 1998
Elonen eta!., 1998
Eloncn ctal.. 1998
Elonen el al.. 1998
Elonen ctal.. 1998
Walker ctal., 1992
Walker el al., 1992
Walker and Peterson, 1994
Zabel & Peterson, 1996
Walker and Peterson, 1991
Walker and Pclcrson. 1991
Walker and Peterson, 1994
Zabcl & Pclcrson, 1996
Zabclclal.. 1995
Olivicri and Cooper, 1997
Zabel ct al., 1995
Walker clal., 1992
Walker clal.. 1992
Olivier! and Cooper, 1997
12/27/99
Pagel of 5
TAMS/MCA
-------
TABLE B-7
TOXICITY ENDPOINTS FOR FISH - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Fathead minnow
(Pimephales promelas)
EXPOSURE
MEDIA
Water
EFFECT
LEVEL
LD50
TISSUE
Larvae
CONTAMINANT
TYPE
2.3.7.K-TCDD
EFFECT
CONC.
(ug/kgww)
70.9
LIPID CONTENT
OF EGG
(glipid/gwwegg)
Not reported for larvae
TEF
1
EFFECT CONC. DIOXIN
EQUIVALENTS
(ugTEQ/kglipid)
EFFECT ENDPOINT
Early life stage mortality
REFERENCE
Olivicri and Cooper. 1997
12/27/99
Page2 of 5
TAMS/MCA
-------
TABLE B-7
TOXICITY ENDPOINTS FOR FISH - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF D1OXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Zchrafish
(Danio danio )
Fathead minmiw
(Pimtphatex prtimela.i )
While sucker
\Cataxlnmiu ctimmerximi )
Northern Pike
\Exox luciiu )
Mcdaka
[Oryziax latipfx )
Fathead minnow
(Pimephalex prumelas )
Channel catfish
(Ictalurux punctatux )
Lake herring
[Coregunux anedii )
Rainbow Trout
(Salmtf gairtieri )
Rainbow Trout
[Sulnw gairderi )
Brook Trout
(Salveniuxfuntinalix )
Lake trout
(Salvelinux namaycush )
Lake Uoul
(Sulvelinus namu\cuxh )
Lake trout
(Salvelinus nama\cu.ih )
Lake trout
(Sulvfflinux ntimaycuxh )
Lake trout
(Salvelinux natnuYcuxh )
Fathead minnow
(Pimephalex pmmelax )
White sucker
(Catastomux a>mmerx
-------
TABLE B-7
TOXICITY ENDPOINTS FOR FISH • LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Luke herring
(Cortgtmux artedii )
Rainbow Trout
(Salmt) gairtlcri ) *
Bnxik Trout
(Salveniusf(intinati\ )
EXPOSURE
MEDIA
Water
Injection
Water
EFFECT
LEVEL
NOAEL
NOAEL
NOAEL
TISSUE
Egg
Egg
Egg
CONTAMINANT
TYPE
2.3.7.K-TCDD
2.3.7.S-TCDD
2,3,7,8-TCDD
EFFECT
CONC.
(ug/kg ww)
0.175
0.291
0.135
LIPID CONTENT
OF EGG
(g lipid/gww egg)
0.066
0.087
0.06X
TEF
I
1
1
EFFECT CONC. DIOXIN
EQUIVALENTS
(ogTEQ/kglipid)
2.7
3.3
2.0
EFFECT ENDPOINT
Early life stage mortality
Early life stage mortality
Early life stage mortality
REFERENCE
Eloncnctal.. IV9X
Walker cl al., IW2
Walker and Peterson, 1994
12/27/99
Page4 of 5
TAMS/MCA
-------
TABLE B-7
Toxicrrv ENDPOINTS FOR FISH - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Lake troul
(Salvelinus namaycush )
Luke trout
(Salvelinus namaycush )
Lake Iniul
(Salvelinus namaycush )
Lake trout
(Salvelinus namaycush )
Lake Lroul
(Salvelinus namaycush )
Fathead minnow
(Pimephales promelas )
EXPOSURE
MEDIA
Injection
Injection
Water
Water
Maternal
transfer
Water
EFFECT
LEVEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
TISSUE
Egg
Egg
Egg
Egg
Egg
Larvae
CONTAMINANT
TYPE
2.3.7.K-TCDD
2.3.7,8-TCDD
2,3.7,8-TCDD
2.3,7,8-TCDD
2.3.7,8-TCDD
2,3,7.8-TCDD
EFFECT
CONC.
(ug/kg ww)
0.044
().()44
0.034
0.034
0.023
3.59
LIPID CONTENT
OF EGG
(g lipid/gww egg)
0.08
0.08
0.08
O.OX
O.OX
Not reported
for larvae
TEF
1
1 '
1
1
,
1
EFFECT CONC. DIOXIN
EQUIVALENTS
(ugTEQ/kglipId)
0.55
0.55
0.43
0.43
0.29
EFFECT ENDPOINT
Early life stage mortality
Early life stage mortality
Early life stage mortality
REFERENCE
Walker etal., 1992
Walker ct all., 1994
Walker ctal., 1992
Walker el all., 1994
Walker el all.. 1994
Olivicri and Cooper, 1997
Notes:
" No relevant Held studies were found.
k Fathead minnow cmhryo is assumed to have same lipid content as reported for eggs (Elimcn el al.. 1998)
12/27/99
Pages of 5
TAMS/MCA
-------
TABLE B-8
TOXICITY ENDPOINTS FOR FISH - FIELD STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
Rainbow Tmul - Artec strain
(Salmu gairtlneri )
Rainbow Trout - Erwin strain
(Stilnut gairdneri )
Rainhtiw Tnml - Lake Superior
(Stilmt ftainlneri )
Killifish
(Funtlulus heteruclitu.1 )
KiUilish
(Funilulu.1 hfttroclitus )
KiluTish
(Fwilulus hetewclitus )
Lake irnul
(Satvflimu nttmayi'uxh )
KillifKh
(Funilulus hftervditia >
EXPOSURE
MEDIA
Egg injection of
extract frtHn field-
collected fish
Egg injection of
extract from field-
collected fish
Egg injection of
extract from tield-
colleded fish
Fish collected
from New Bedford
Harbor
Fish collected
from New Bedford
Harbor
Fish collected
from New Bedford
Harbor
Fish collected from
Lake Ontario
Fish collected
from New Bedford
Harbor
EFFECT
LEVEL
LD50
LD50
LD50
LOAEL
LOAEL
NOAEL
EL-no effect
NOAEL
TISSUE
Eggs
Eggs
Eggs
Liver
Liver
Liver
Eggs
Liver
CONTAMINANT
TYPE
TEQs
TEQs
TEQs
EFFECT
CONC.
(o(ftf ww,
unless mud differently Mow)
0.514
0.206
1.43
LIPID CONTENT OF
EGG
(tlipid/jwwegg)
0.0117
O.OS7
O.OS7
TEQs
TEQs
TEQs
TEQs
TEQs
1.56 ug/kg dry et
0.543 ug/kg dry wl
0.1 32 ug/kg dry wt
0.011
0.00572 ug/kg dry wt
Not available
Not available
Not available
O.OS
Not available
EFFECT CONC.
(ne/kt lipld)
5.9
2.4
16.4
Not available
Not available
Not available
0.1
Not available
TEF
1
1
1
1
1
1
1
1
EFFECT CONC. DIOXIN
EQUIVALENTS
(ll(TEQAgtipid)
5.9
2.4
16.4
Mitt available
Not available
Sol available
O.I
Not available
EFFECT ENDPOINT
Embryomonality
Embryomonality
Embryomortalny
Embryo and larval survival
Adult female mortality
Embryo and larval survival
Early life stage mortality
Adult female mortality
REFERENCE
Wright and Tillilt , 1999
Wright and Tillitt. 1999
Wright and Tillitt, 1999
Black ctal.. 1998
Black el al.. 1998
Black etal., 1998
Guiney et al., 1996
Black el al.. 1998
12/27/99
Page 1 of 1
TAMS/MCA
-------
TABLE B-V
TOX1CITY ENDPO1NTS FOR AVIANS • LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
PCB TYPE
EFFECTIVE
DOSE
(mg/kg/day)
EFFECTIVE
FOOD CONC.
(rng/kg)
EFFECT ENDPOINT
REFERENCE
Laboratory studies
Mallard Duck
(Anas platrtiynchos)
Japanese Quail
(Cotumix cotumlx)
Bobwhite Quail
(Colinus virglnianus)
Brown-headed Cowbird
(Molothrus aler)
Red-winged Blackbird
(Ageloius phoeniceus)
Japanese Quail
(Cotumix cotumix)
Mallard Duck
(Anas platrtiynchos)
Domestic Chicken
(Ga//us domesticus)
Ring-Necked Pheasant
(Phosionus colchlcus)
Ring-Necked Pheasant
(Phas/anus colchicus)
Domestic Chicken
(Ga//us domesf/cus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Ringed Turtle Dove
(Streptopelia risoria)
Ringed Turtle Dove
(Streptopelia risoria)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domesticus)
Diet
Diet
Oral by syringe
Diet
Drinking water
Diet, in gelatin
capsules
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
5 day
5 day
5 day
7 days
6 days
7 days
12 weeks
6 weeks
Once per week for
1 7 weeks
Not available
9 weeks
9 weeks
9 weeks
9 weeks
9 weeks
9 weeks
3 months
6 weeks
8 weeks
8 weeks
8 weeks
LD50
LD50
LD50
EL-effect
EL-effect
LOAEL
EL-effect
EL-effect
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
EL-effect
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
Aroclorl254
Aroclorl254
Aroclor 1254
Aroclor 1254
Aroclor 1254
Aroclor 1260
Aroclor 1242
Aroclor 1254
Aroclor 1254
Aroclor 1254
Aroclor 1242
Aroclor 1248
Aroclor 1254
Aroclor1242
Aroclor 1254
Aroclor1248
Aroclor 1254
Aroclor 1254
Aroclor 1242
Aroclor 1242
Aroclor 1248
Aroclor 1248
Aroclor 1254
853
759
141
333
321
100
16
3.5
2.9
2.9
1.4
1.4
1.4
1.4
1.4
1.4
1.1
1.1
0.7
0.7
0.7
0.7
0.3
8122
6737
1516
1500
1500
888
150
50
50
50
20
20
20
20
20
20
10
10
10
10
10
10
5
Mortality
Mortality
Mortality
Hilletal., 1975
Hilletal.. 1975
Hilletal.. 1975
Mortality
Mortality
Weight loss
Decreased weight gain in hens,
eggshell thinning
Hatching success
Egg production
Female fertility
Egg production, hatching
success, chick growth
Egg production, hatching
success, chick growth
Egg production, hatching
success, chick growth
Hatching success
Hatching success
Hatching success
Hatching success
Hatching success
Hatching success
Hatching success
Hatching success
Hatching success
Fertility and egg production
Stickel et al., 1984
Stickel et al., 1984
Vosetal.. 1971
Haseltine and Prouty, 1980
Tumasonis et al., 1973
Dahlgren et al., 1972
Roberts etal.. 1978
Ullieetal.. 1974
Ullieetal.. 1974
Ullie et al., 1974
Cecil etal., 1974
Cecil et al.. 1974
Cecil etal., 1974
Peakoll et al, 1972
Peakall and Peakall,
1973
Brttton and Huston, 1973
Ullieetal., 1975
Ullieetal.. 1975
Scott, 1977
Platonow and Reinhart,
1973
Pafc I
TAMS/MCA
-------
TABLE B-9
TOXICITY ENDPOINTS FOR AVIANS - LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
PCB TYPE
EFFECTIVE
DOSE
(mg/kg/day)
EFFECTIVE
FOOD CONC.
(mg/kg)
EFFECT ENDPOINT
REFERENCE
Laboratory studies
European Starling
(Stemus vu/garts)
Common Grackle
(Quiscalus qu/scuto)
Mallard Duck
(Anas pfatrtiynchos)
Japanese Quail
(Cotum/x cotumix)
Mallard Duck
(Anas platymynchos)
Japanese Quail
(Cofurn/x cotumix)
Domestic Chicken
(Gallus domesf/cus)
Domestic Chicken
(Gallus domestlcus)
Domestic Chicken
(Gallus domesf/cus)
Domestic Chicken
(.Gallus domesf/cus)
Domestic Chicken
(Gallus domesticus)
Domestic Chicken
(Gallus domestlcus)
Ring-Necked Pheasant
(Phaslanus colchlcus)
Screech Owl
(Ofusasto)
Domestic Chicken
(Gallus domestlcus')
Domestic Chicken
(Gallus domest/cus)
Domestic Chicken
(Gallus domestlcus')
Domestic Chicken
(Gallus domestlcus)
Domestic Chicken
(Gallus domestlcus)
Domestic Chicken
(Gallus domestlcus)
Domestic Chicken
(Gallus domest/cus)
Domestic Chicken
(Gallus domestlcus)
Domestic Chicken
(Gallus domest/cus)
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet, in gelatin
capsules
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
4 days
8 days
12 weeks
14 weeks
Approx. 1 month
Not reported
8 weeks
8 weeks
9 weeks
9 weeks
9 weeks
9 weeks
Once per week for
1 7 weeks
> 8 weeks
6 weeks
8 weeks
8 weeks
9 weeks
9' weeks
9 weeks
9 weeks
9 weeks
9 weeks
El-effect
EL-effect
EL-no effect
EL-no effect
EL-no effect
NOAEL
NOAEL
NOAEL
EL-no effect
EL-no effect
EL-no effect
EL-no effect
NOAEL
EL-no effect
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
Aroclor 1254
Aroclor 1254
Aroclorl242
Aroclor 1254
Aroclor 1254
Aroclor 1248
Aroclor 1016
Aroclor 1254
Aroclor1221
Aroclor1232
Aroclor1268
Aroclor 5442
Aroclor 1254
Aroclor 1248
Aroclorl242
Aroclor 1242
Aroclor 1248
Aroclor 1242
Aroclor 1248
Aroclor 1254
Aroclor 1242
Aroclor1248
Aroclor 1254
Not available
Not available
16
5.6
2.6
2.3
1.4
1.4
1.4
1.4
1.4
1.4
0.7
0.4
0.3
0.3
0.3
0.1
0.1
0.1
0.1
0.1
0.1
1.500
1.500
150
50
25
20
20
20
20
20
20
20
12.5
3
5
5
5
2
2
2
2
2
2
Mortality
Mortality
Reproduction success.
hatching success, survival and
growth of chicks
Mortality and growth rates of
adufts
Reproduction success
Hatching success
Egg production
Egg production
Hatching success
Hatching success
Hatching success
Hatching success
Egg production
Egg prdoduction. hatching
success, fledging success
Hatching success
Hatching success
Hatching success
Egg production, hatching
success, chick growth
Egg production, hatching
success, chick growth
Egg production, hatching
success, chick growth
Hatching success
Hatching success
Hatching success
Stlckel et al.. 1984
Stickel et al.. 1984
Haseltine and Prouty. 1980
Chang and Stokstad.
1975
Custer and Heinz. 1980
Scon. 1977
Llllieetal.. 1975
Ullieetal.. 1975
Cecil el al.. 1974
Cecil el al.. 1974
Cecil el al.. 1974
Cecil el al.. 1974
Dahlgren et al., 1972
McLane and Hughes, 1980
Brltton and Huston. 1973
Ullieetal, 1975
Ullieetal.. 1975
Ullieetal.. 1974
Ullieetal.. 1974
Ullieetal.. 1974
Cecil etal., 1974
Cecil etal., 1974
Cedlelal., 1974
Page 2 of .1
TAMS/MCA
-------
TABLE B-9
TOXIC1TY ENDPO1NTS FOR AVIANS • LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
PCB TYPE
EFFECTIVE
DOSE
(mg/kg/day)
EFFECTIVE
FOOD CONC.
(mg/kg)
EFFECT ENDPOINT
REFERENCE
Laboratory studies
Domestic Chicken
(Ga//us domesticus)
Diet
8 weeks
NOAEL
Aroclorl248
0.1
1
Hatching success
Scott, 1977
Page 3 of 3
TAMS/MCA
-------
TABLE B-10
TOXICITY ENDPO1NTS FOR AVIANS - FIELD STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
FIELD
COMPONENT
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
DOSE
(mg/kg/day)
EFFECTIVE
FOOD CONC.
(mg/kg)
EFFECT ENDPOINT
REFERENCE
Field studies
Tree Swallow
(Tachycineta bicolor )
Tree Swallow
(Tach\cineta bicolor)
Populations in Fox
River and Green
Bay, Lake
Michigan, studied
Populations along
Hudson River
studied
NOAEL
NOAEL
PCBs, DDE
PCBs
0.55
16.1
up to 0.61
up to 17.9
Clutch and egg success
Growth, mortality, reproduction
Custeretal., 1998
US EPA
Phase 2 Database (1998)
TAMS/MCA
-------
TABLE B-ll
TOXICITY ENDPOINTS FOR AVIANS - LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF D1OXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
DOSE
DIOXIN
EQUIVALENTS
(ug/kg/day)
EFFECT ENDPOINT
REFERENCE
Laboratory studies*
Ringed turtle dove
(Streptopelia risor/o)
Mallard
(Anas platrhyncos)
Chicken
(Gallus domesficus)
Ring-necked pheasant
(Phasianus colchlcus)
Northern bobwhlte quail
(Colinus virgin/onus)
Oral
Oral
Oral •
Intraperitoneal
Oral
Singe dose
Single dose
21 days
Single dose
Single dose
LDso
LD100
LD75
LDso
2.3.7.8-TCDD
2.3.7.8-TCDD
2.3.7.8-TCDD
2,3.7.8-TCDD
2.3.7.8-TCDD
>810
>108
25-50
25
15
Mortality
Mortality
Mortality
Mortality
Mortality
Hudson etal., 1984
Hudson etal.. 1984
Greigetal.. 1973
Noseketal.. 1992
Hudson etal.. 1984
Chicken
(Ga//us domestlcus)
Ring-necked pheasant
(Phasianus colchlcus)
Oral
Intraperitoneal
21 days
10 weeks
LOAEL
LOAEL
2.3.7.8-TCDD
2.3.7.8-TCDD
1.0
0.14
Mortality
Fertility, embryo mortality
Schwetz et al.. 1973
Noseketal.. 1992
Chicken
(Gallus domestfcus)
Ring-necked pheasant
(Phasianus colchicus)
Oral
Intraperitoneal
21 days
10 weeks
NOAEL
NOAEL
2,3.7.8-TCDD
2.3.7.8-TCDD
0.1
0.014
Mortality
Fertility, embryo mortality
Schwetz et al.. 1973
Noseketal.. 1992
Notes:
* No relevant field studies were found.
Note units of ug/kg/day.
TAMS/MCA
-------
TABLE B-12
TOXICITY ENDPOINTS FOR AVIANS - FIELD STUDIES
EFFECTIVE DIETARY DOSES OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
FIELD
COMPONENT
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
DOSE
DIOXIN
EQUIVALENT
S
(ug/kg/day)
EFFECTIVE
FOOD
CONC.
(ue/kg)
EFFECT ENDPOINT
REFERENCE
Field studies
Tree Swallow
(Tachycincta hicolor )
Tree Swallow
(Tachycineta tricolor)
Populations
along Hudson
River studied
Populations in
Fox River and
Green Bay, Lake
Michigan,
EL-no effect
EL-no effect
TEQs
TEQs, DDE
4.9
0.08
up to 5.41
up to 0.091
Growth, mortality, reproduction
Clutch and egg success
US EPA
Phase 2 Database, 1998
Custeretal., 1998
TAMS/MCA
-------
TABLE B-I3
TOXICITY ENDPOINTS FOR AVIAN EGGS • LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs AND AROCLORS
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
PCBTYPE
EFFECTIVE
EGG CONC.
(mg/kg egg)
EFFECT ENDPOINT
REFERENCE
Laboratory studies
Chicken
(Callus domesticus )
Chicken
(Callus ditmcsticux )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus diimeslicus )
Drinking walcr
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Diet
Diet
6 weeks
6 weeks
4 weeks
EL-effecl
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
Aroclor 1254
Arocltir 1260
Aroclor 1254
Aroclor 1242
Aroclor 1254
Aroclor 1242
Aroclor 1242
Aroclor 1248
> 10-1 5 ppm in
yolk
10
6.7
5
5
5
5
3.7
2.21
Deformities
Growth rate of chicks
Growth and mortality of embryos
Hatching success
Hatching success
Growth mtc of chicks
Egg production and hutching success
Hatching success
Hatching success
Tumasnnis et al., 1973
Carlson and Duhy, 1973
Gould etal., 1997
Carlson and Duby. 1973
Carlson and Duhy. 1973
Carlson and Duby. 1973
Platanow and Reinhart, 1973
Britton and Huston, 1973
Scott, 1977
Chicken
(Callus domesticus )
Screech owl
(Otus asiti )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus domesticus )
Chicken
(Callus domeslicus )
Chicken
(Callus domesticus )
Chicken
(Callus dome sticus )
Chicken
(Callus domesticus )
Egg injection
Diet of hens
Egg injection
Egg injection
Egg injection
Egg injection
Diet
Egg injection
Diet
> 8 weeks
6 weeks
4 weeks
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
Aroclor 1260
Aroclor 1248
Aroclor 1260
Aroclor 1242
Aroclor 1254
Aroclor 1242
Aroclor 1242
Aroclor 1254
Aroclor 1248
10
7.1
5
2.5
2.5
2.5
1.7
0.67
0.33
Hatching success
Egg production, hatching success,
anil fledging success
Growth rale of chicks
Hatching success
Hatching success
Growth rate of chicks
Hatching success
Growth and mortality of cmhryos
Hatching success
Carlson and Duby. 1973
McLane and Hughes, 19X0
Carlson and Duby, 1973
Carlson and Duby, 1973
Carlson and Duby, 1973
Carlson and Duby, 1973
Britlon and Huston. 1973
Gould etal., 1997
Scott. 1977
TAMS/MCA
-------
TABLE B-14
TOXICITV ENDPOINTS FOR AVIAN EGGS - FIELD STUDIES
EFFECTIVE CONCENTRATIONS OF TOTAL PCBs AND AROCLORS
SPECIES
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
EGG CONC.
(mg/kgegg)
EFFECT ENDPOINT
REFERENCE
Field studies
Bald eagle
(Haliaeelux Icucticephttlus )
Double-crested cormorant
(Phaliicrocurax auritux )
Caspian tem
(Hydrtiptignc caxpia )
Forster's tern
(Sternafarxteri )
Common lern
(Sterna hirundu )
Common tern
(Sterna hirundti )
Bald eagle
(Haliueetux leucocephalux )
EL-Effcct level
EL-Effccl level
EL-Effcci level
LOAEL
LOAEL
LOAEL
LOAEL
PCBs. Pesticides
PCBs, Pesticides, Hg
PCBs, Pesticides
PCBs, Pesticides,
Dioxins, Furans
PCBs, Pesticides, Hg
PCBs, Pesticides, Hg
PCBs, Pesticides, Hg
20-54
23.8
4.2-18
22.2
7
9.8
3 - 5.6
Reproductive success
Hatching success and
fledging success
Increased rale of
cmhryo deformities
Hatching success
Hatching success
Hatching success
10 % reduction in
reproductive success
Clark el al., I9S8
Wcscloh el al., 19X3
Yamashita el al.. 1993
Kuhiak ct al.. 1989
Becker cl al., 1993
Hoffman ctal.. 1993
Wicmcycrctal.. 1984, 1993
Bald eagle
(Haliaeetus leutitcrphalux )
Tree swallow
(Tachycinfta bicotor)
Common lem
(Sterna hirunitti )
Common lem
(Sterna hirundu )
Forster's tern
(Sterna forsteri)
Tree swallow
(Tachycineta bittilur )
Bald eagle
(Haliaeetu.* Icui-ucephatus )
EL- No Effect
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
PCBs, TEQs, Pesticides
PCBs
PCBs, Pesticides, Hg
PCBs, Pesticides, Hg
PCBs, Pesticides.
Dioxins, Furans
PCBs. DDE
PCBs. Pesticides, Hg
33.2 - 64 in
yolk sac
26.7
6.7
5.2
4.5
3.24 in eggs
and pippers
<3
Hatching success
Reproductive output
Hatching success
Hatching success
Hutching success
Clutch success, egg success
Reproductive success
Elliott el al., 1996
Sccord and McCarty, 1997,
McCarty and Sccord, 1999.
U.S. EPA Phase 2 Datahasc Release 4.lb, 1998
Hoffman etal., 1993
Becker ctal., 1993
Kuhiak ctal.. 19X9
Custcrclal.. 1998
Wicmcycrctal., 1984. 1993
TAMS/MCA
-------
TABLE B-l!
TOXICITY ENDPOINTS FOR AVIAN EGGS - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
I laboratory studies
American kestrel
r*«/t» sparverius)
[>ouhle-crcsted cormorant
Phalacrocorax tluritux )
Common tem
Sterna hiruiitlit )
American kestrel
.Fillto xparverius )
Ring-necked pheasant
Phasianus calcHicux )
Chicken
(Callus dttmexticux )
Chicken
Callus aumestiLux )
Chicken
Callus ilimiesticux )
Chicken
(Callus tlttnusricus )
Chicken
(Callus gallux)
Chicken
(Callus gallus)
Chicken
(Callus iltmtexticus )
>>uhle-cre5lcd cormorant
(PhaliKracarax uuritus )
American kestrel
(Falco xparverius )
American kestrel
(Falca sparverius )
Common tern
(Sterna hiruntlo )
Double-crested cormorant
(PhalacriKorax aurilux)
Ring-necked pheasant
(Phaxianux colLhicus )
Chicken
(Callus aontexticus )
Chicken
(Callus ttnmexlicus)
Chicken
(Callus gallus)
Chicken
(Callus aoniestiL-ux )
Pidgcon
(Calumhti livia )
Chicken
(Callus tlimtexlicux )
Chicken
(Callus gallus)
Chicken
(Callus ilomeslicus )
Douhlc-c rested cormorant
(PhaloLTHctiriw auritux)
F.XPOSURE
MEDIA
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
EXPOSURE
DURATION
EFFECT
LEVEL
IK days
2 1 days
IK days
211 days
2«days
IX days
IK days
24 days
24 days
20 days
1 K days
IK days
21 days
20 days
20 days
IK days
2 1 days
21 days
IK days
IK days
IK days
24 days
Einhryonic Day 3
Ihnnigh hatch
IK days
IK days
24 days
21 days
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
EL-Eflecl
LOAEL
LOAEL
LOAEL
NOAEL
CONTAMINANT
TYPE
PCB77
PCB 126
PCB 126
PCB 126
2.3.7.K-TCDD
PCB 105
PCB 77
PCB 126
2.3.7.K-TCDD
PCB 77
PCB I2n
PCB 126
PCB 126
PCB 126
PCB 77
PCB 126
2.3.7.K-TCDD
2.3.7.K-TCDD
PCB 105
PCB 77
PCB 77
2.3.7.K-TCDD
2.3.7.K-TCDD
PCB 126
PCB 126
PCB 126
PCB 126
EFFECTIVE
EGG CONC.
(us/kg *0l)
TEF
EFFF.CTIVE
EGG CONC.
DIOXIN
EQUIVALENTS
(ugTEQ/kgegg)
EFFECT ENDPOINT
REFERENCE
316
I5K
104
65
1.35
5592
K.K
2.3
0.15
2.6
(1.4
0.6
KOO
233
100
44
4
1
KIOO
<)
6
0.16
1
H.I
0.5
0.2
4(X>
' 0.05
O.I
O.I
O.I
1
0.(XXH
0.05
O.I
1
0.05
O.I
O.I
O.I
0.1
0.05
O.I
1
1
O.IXXII
0.05
0.05
1
1
O.I
O.I
01
O.I
16
16
10
7
1
1
0.4
0.2
0.2
0.1
0.04
0.1
XO
23
5
4
4
1.0
,
0.5
0.3
0.2
1.0
0.09
0.05
0.02
40
Enthryo mortality
Emhryo mortality
Emhryo mortality
Emhryo mortality
Enihryo mortality
Emhryo mortality
Emhryo mortality
Emhryo nxmality
Emhryo mortality
Emhryo mortality
Emhryo mortality
Emhryo mortality
Enihryo mortality
Emhryo mortality
Emhryo mortality
Emhryo mortality
Enthryo mortality
Enihryo mortality
Emhryo mortality
Emhryo mortality
Emhryo mortality
Emhryo nxirtalily
Halchahilily
Emhryo mortality
Emhryo mortality
Emhryo mortality
Emhryo mortality
Hoffman el al., 1998
Powell ail.. 1997
Hoffman el al.. I99X
Hoffman el al.. 199!)
Noseketal.. 1993
Powell el al.. !9V6h
Powell etal.. I996h
Powell ail., IWa
Powell ail.. I996a
Hoffman et al.. 199K
Hoffman el al.. I99X
Powell el al., I996h
Powell a al.. 1997
Hoffman a al.. 1998
Hoffman el al.. 1998
Hoffman el al.. 1998
Powell ail., 1997
Nosck a al.. 1993
Powell a al.. I9%h
Powell el al.. I9%h
Hoffman el al.. 1998
Powell etal.. I996a
anz and Bcllward, 1996
Powell el al.. I996b
Hoffman el al.. IWH
Powell a al.. I996a
Powell el al.. 1997
-------
TABLE B-1S
TOXICITY ENDPOINTS FOR AVIAN EGGS • LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOX1N TOXIC EQUIVALENTS (TEQs)
SPECIES
Great Blue Heron
(Antea hcrtHliax)
American kestrel
(F ttko sparvfrius )
Double-crested ctirmorant
(Phaluzrocariu tluritux )
EXPOSURE
MEDIA
Egg injection
Egg injection
Egg injection
EXPOSURE
DURATION
Embryonic Day 9
through halch
20 days
21 days
EFFECT
I.EVEI,
EL-No effect
NOAEL
NOAEL
CONTAMINANT
TYPF.
2.3.7.K-TCDD
PCB 12ft
2.3.7.8-TCDD
EFFECTIVE
EGG CONC.
(UK/kg ea>)
2
23
1
TEF
1
O.I
1
EFFECTIVE
EGG CONC.
DIOXIN
EQUIVALENTS
(utTEQ/kxeKK)
2
2
1
EFFECT ENDPOINT
HatchaHlily
Enihryti iimnalily
Enihryo mortality
REFERENCE
Janz and Bellward. IV9A
Hnflnunelal., I99»
Powell el al., 1997
TAMS/MCA
-------
TABLE B-15
TOXICITY F.NDPO1NTS FOR AVIAN EGGS - LABORATORY STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEOs)
SPECIES
Chicken
((jalltts tlumcsticus)
Chicken
(Callus dvnusticus)
Ring-necked pheasant
(Phusianus culchicus )
Chicken
(Callus Jtuttfxticus)
Chicken
(Callus fallus)
Chicken
(Callus gallux )
Chicken
(Callus gallus )
Chicken
[Callus tlttmexticus )
Chicken
(Callus ilitnifxlicus )
EXPOSURE
MEDIA
EgB injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
Egg injection
F.XPOSURF.
DURATION
IX days
IX days
2X days
24 days
1 X days
Embryonic Day 4
through hatch
IX days
IX days
24 days
EFFECT
LEVEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
EL-No effect
NOAEL
NOAEL
NOAEL
CONTAMINANT
TYPE
PCB 1(15
PCB77
2.3.7.X-TCDD
2.3.7.8-TCDD
PCB 77
2.3.7.X-TCDD
PCB 126
PCB 126
PCB 126
EFFECTIVE
EGG CONC.
(tig/kg egg)
•2700
3
O.I
O.OX
1.2
0.1
0.3
0.3
0.1
TEF
0.0001
0.05
1
1
0.05
1
O.I
O.I
O.I
EFFECTIVE
EGG CONC.
DIOXIN
EQUIVALENTS
(at TF.O/kK tgg)
0.3
0.2
0.1
0.1
O.I
O.I
0.03
0.03
0.01
EFFECT ENDPOINT
Embryo mortality
Eniniyo mmalily
Embryo mortality
Embryo mortality
Embryo mortality
Halchahiliry
Embryo mortality
Embryo mortality
Embryo mortality
REFERENCE
Powell elal., IWfih
Powell el al.,IW6h
Noseketal., IW3
Powell elal.. IWna
Hoffman et al., 199X
Janz and Bellward, 1906
Hoffman etal..lV98
Powell et al.. l«%h
Powell et al.. lW6a
-------
TABLE B-16
TOXICITY ENDPOINTS FOR AVIAN EGGS - FIELD STUDIES
EFFECTIVE CONCENTRATIONS OF DIOXIN TOXIC EQUIVALENTS (TEQi)
SPECIES
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
EGG CONC.
DIOXIN
EQUIVALENTS
(ugTEQ/kgrgg)
EFFECT ENDPOINT
KEFKKKMCK
Field studies
Osprey
\Pandion hitliiicelux )
Bald eagle
(Halitieelux Ifucwephnlux )
Great blue heron
(Artleti henntitix )
Great blue heron
(Artlftt liewtlitix)
Cormorant
[Plitilticrttwrttx uuritux )
Great hluc heron
(Anlfil hertntitix )
Forelegs tern
(Slcmn jurxleri )
Former's lem
[Stenttt ftirxleri )
W,H«J duck
(Aix xpimsii )
EL-Etta level
EL-Etfcct level
LOAEL
EL-Effcct level
EL-cllect level
EL-EITcct level
EL-Ellcct
EL-EITccl level
LOAEL
TCDD
TEQs. DDE
TEQs
TEQs. pesticides
TEQ
TEQs. pesticides
TEQs. pesticides
TEQ
TEQs, pesticides
29- 162
0.51-1.2
0.5
0.5
0.0.15 - 0.344
0.23
2.20
0.21
0.02
Growth rate of chicks
Reproductive success
Growth rate
Growth rale
Egg mortality
Reproductive success
Hatching success.
growth rule of chicks
Hatching success
Scst success, hatching
success, duckling
production
WiKKJlurd. et al., 1998
Clark el al.. I99K
Sanderson el al.. 1994
Hart el al.. 1991
Tillilt cc al., 1992
Elliott el al., 19X9
Kuhiakclal.. 19X9
Tillilt cl al.. 1993
White and Scginak. 1994;
While and Hoffman. 1995
Tree swallow
(TaLhycinela bicotttr)
Tree .swallow
(Titchycineltl hicnlor)
Great blue henin
(Artlcti heruflittx)
Great hluc heron
(Artleti heruilinx )
Ftirstcr's lern
(Stf nut ftirxleri)
Great hluc heron
(Artlrti hrriHtitix)
Osprcy
(Ptintliim hiilincrlus)
Osprcy
(Piliitliim liiiliiitrtu.*)
Foster's tern
(Stemtlftirxleri)
Wivxl duck
(Aix XIHHIXll )
NOAEL
EL-No elTccl
NOAEL
NOAEL
EL-no ctfecl
EL-No elTccl
EL-no eltccl
EL-no eltccl
EL-no clfcd
NOAEL
TEQs
TEQs
TEQs
TEQs
TEQs. pesticides
TEQs, pesticides
TCDD, TEQs
TI-Qs
TEQs
TEQs, pesticides
13
0.5X9 in pippers
0.3
0.24
0.2
0.079
ND-23.K
0.1 3ft
(1.023
0.005
Reproductive success
Reproductive success
Reduced hudy weight
Growth rate
Hatchahilry.
growth rate of chicks
Reproductive success
Growth rale of chicks
Enihryo survival
Hatching success
Nest success, hatching
success, duckling
production
US EPA
Phase 2 Database (1998)
Cuslcretal., I99X
Sanderson cl al., 1994
Hanelal.. 199 1
Kuhiaketal.. 19X9
Elliot! cl al., 19X9
Woodford cl al.. I99X
Woodford cl al.. 1998
Tillill et al.. 1993
While and Scginak. 1994:
While and Hoffman. 1995
-------
TABLE B-17
TOXICITY ENDPO1NTS FOR OTHER MAMMALS • LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
Laboratory studies"
Osbome-Mendel Rut
Osbome-Mendcl Ral
Wistar Ral
Juvenile Mak Rat
Juvenile Mak Ral
Sherman Rut
Raccoon
(Pwcytin btvr)
Oshome-Mendcl Ral
Balh/c Miiu.se
Adull Female Rat
WislarRal
While-failed Mouse
(Pr romyxcux leucupus )
Wislar Ral
Mouse
Rabbit
PiS
New Zealand While Rabbit
Osbomc-Mendel Ral
Rhesus Monkey
(Macttca muhttlti )
Rhesus Monkey
(Macaco mulana )
Fischer Ral
Guinea Pig
Shcmian Rut
Wistar Ral
Oldllcld Mouse
( Peromyxcux ptiliutuitus )
Rhesus Monkey
(Mutual mulatta )
EXPOSURE
MEDIA
Orul-guvuge
Onil-gavugc
Diet
Single iruraperiloncul
injection
Single inlraperitoncal
injection
Diet
Diet
Diet
Oral
Oral
Oral-gavage
Diet
Diet
Diet
Oral-giivage
Diet
Did
Diet
Diet
Diet
Diet
Oral-gavugc
Did
Did
Diet
Diet
EXPOSURE
DURATION
2.5 wk, 2 d per week
2.5 wk. 2 d per week
Fnmi mating to weaning
olnuns
Observed after 14 days
Observed alter 1 4 days
X months
X days •
During pregnancy and
lactation
6 months
Day 1, 3,5.7 and 9 of
lactation
1 month
1 2 weeks
42 days
1 OX days
2X days
91 days
> 4 weeks
During pregnancy and
uctalion
2 months
2 months
105 weeks
Gcslalional day IX-60
Mulligcncrational
52 weeks
12 months
3X weeks
EFFECT
LEVEL
LDW
LD™
LDW
LOAEL
LOAEL
LOAEL
EL-effcct
LOAEL
LOAEL
LOAEL
LOAEL
EL-cffcct
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
EL-cflecl
LOAEL
PCB TYPE
Aroclor 1254
Aniclor 1254
Aroclor 1254
Aniclor I24X
Aroclur 12.12
Aniclor 1260
Arochor 1254
Not reported
Aniclor 1254
Aniclor 1254
Aniclor 1254
Aniclor 1254
Aniclor 1254
Aniclor 1254
Aroclor 1254
Aniclor 1242
Aniclor I24X
Not reported
Aroclor I24X
Aniclur I24X
Aniclur 1254
Clophcn A50
Aniclor 1254
Aniclor 1254
Aniclor 1254
Aniclor 1254
EFFECTIVE
DOSE
(mg/kg/day)
1530
1530
22
2000
2000
72.4
50
49.471
4X.75
32
30
17
13.5
12.5
12.5
9.2
X.9
4.947
4.3
4.3
2.5
2.5
1.5
1
0.6X
0.2
FOOD
INCESTION
RATE
(kg/kg/day)
0.099
0.099
0
O.OX
0.0X0
O.IX
O.OX
O.OX
O.OX
O.IX
0.034
0.0
0.0X0
0.2
0.2
O.OX
O.OX
O.OX
0.01
0.2
EFFECTIVE
FOOD CONC.
(nig/kg)
EFFECT ENDPOINT
269
SIX)
10
250
50
20
5
Mortality
Mortality
2 day postnatal mortality or offspring
Gntwth rale of juveniles
Growth rate of juveniles
Mortality
Decreased weigh! gain
Reduced litter si/e
Monality
Reduced growth rate of offspring
Decreased litter si/£. survival of
wcunlings
Reduced growth rate reproduction in
second generation
Nconatul death
Decreased conception
Felal death
Decreased weight gain
Reduced growth rule in offspring
Reduced growth rule of offspring
Decreased conception
Abortion
Decreased survival
Fetal death
Decreased litter si/£
Decreased growth rate
Decreased offspring bom per mated pair.
binh weight, % survival of offspring lo
weaning
No conception, ahonion
REFERENCE
Ganhoff el al.. 19X1 (ATSDR)
GunholTel al., 19X1 (ATSDR)
Overmann ct al., 1987
Harris elul., 1993
Harris el al.. 1993
Kimhniugh et al., 1972
(ATSDR)
Montzetul.. 1982
Collins & Capen, 19X0
Kollcretal.. 1977 (ATSDR)
Suger&Girard. 1994
Brczner el ul., 19X4 (ATSDR)
Limey, l98X(Goluh)
Overmann. 1987 (ATSDR)
Welsch. 1975 (ATSDR)
Villeneuve el al., 1971
(ATSDR)
Hanscn ct al.. 1976 (ATSDR)
Thomas and Hinsdill. 19X0
(Golub)
Collins & Capen. 19X0
Allen etal.. I974a (ATSDR)
Allen ctal.. I974a (ATSDR)
NCI, 197X (ATSDR)
Lundkvisl. 1990 (ATSDR)
Linderetal.. 1974
Phillips etal.. 1972 (ATSDR)
McCoy etal., 1995
Arnold ct al.. 1990 (ATSDR)
Page I of 2
TAMS/MCA
-------
TABLE B-17
TOXICITY ENDPOINTS FOR OTHER MAMMALS - LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
Rhesus Monkey
[Muctica mulatto )
Wislar Rul
Rhesus Monkey
[Mitcacti mulutta )
Rhesus Monkey
[Mtiaica mulatto )
Rhesus Monkey
[Muctiai mulatto )
"ynoniotgus Monkey
Rhesus Monkey
[Macaco multttta }
Rhesus Monkey
,Macaat mulattti )
Swine
EXPOSURE
MEDIA
Diet
Die)
Dicl
Diet
Dici
Die)
Dicl
Dicl
Dicl
EXPOSURE
DURATION
7 months
From mating lit weaning
of pups
2 months
1 .5 years
1 X months
23 X days
IX.2
> 8 months
Throughout gestation
EFFECT
LEVEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
EL-clTcd
PCB TYPE
Aniclor 1248
An>clor 1254
An>clor 1242
An>clor 1248
Aniclor 1248
Aniclor 1254
Aniclor 1248
Anicliir 1016
Aniclor 1242
EFFECTIVE
DOSE
(nig/kg/day)
0.2
0.2
0.12
0.12
O.I
O.I
0.08
0.04
Not uvuiljhlc
FOOD
INGESTION
RATE
(kg/kg/day)
0.2
0.08
0.2
0.2
0.2
0.2
0.2
EFFECTIVE
FOOD CONC.
(mg/kg)
2.5
5
1
20
EFFECT ENDPOINT
Dccreused conception
Reduced growth rjle in offspring
No weight gain
Reduced Wnh weight
Infant nuinulily
100% fetal death
Decreased hinh weight
Reduced hinh weight
Decreased litter size
REFERENCF.
Barscxti el al., 1976 (ATSDR)
Overmann el al., 1987
Becker el al., 1979 (ATSDR)
Allen and Barsolti, 1976 (G.iluh)
Allen end.. 1980 (ATSDR)
Truclove el al., 1982 (ATSDR)
Levin elal.. 1988 (ATSDR)
Barsotli and Van Miller. 1984
(Goluh)
Hanscnetal.. 1975(Golub)
Juvenile Male Rat
Juvenile Mute Rai
Wistar Ral
Rabbit
A dull Female Ral
New Zealand While Rabbit
Sherman Ral
Osbomc- Mendel Rul
Rhesus Monkey
,M(taicd multttta )
Wisiar Ral
Single intnipcritoncul
in Jed ion
Single inlrapcriloneal
injection
Dicl
Orol-gavage
Oral
Dicl
Diet
Dicl
Diet
Dicl
Observed uller 14 days
Observed alter 1 4 days
52 weeks
2X days
Day 1.3,5.7 and 9 of
lactation
> 4 weeks
Muliigcncralional
During pregnancy and
lactation
> X months
From muling to weaning
,,r>ps
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
Aroclor 1248
Aniclor 1232
Aniclor 1254
Aniclor 1254
Aniclor 1254
Anicliir 1248
An>clnr 1254
Anicliir 1254
Aroclor 1016
Anidiir 1254
480
480
10
10
8
3.6
0.32
0.059
0.01
0.0016
0.08
0.034
0.099
0.034
0.08
0.08
0.2
0.08
KM)
5
50
0.25
0.02
Gniwth rate of juveniles
Gniwth rale of juveniles
Decreased gniwlh rate
Fetal death
Gniwth rate of offspring
Reduced growth rale in offspring
Decreased litter size
Reduced litter size
Reduced birth weight
Reduced growth rate in offspring
Harris elal.. 1993
Harris ctal.. 1993
Phillips elal ..1972 (ATSDR)
Villencuve ct al., 1971
(ATSDR)
Sager&Girard. 1994
Thomas and Hinsdill, 1980
(Guluh)
Under ctal., 1974
Collins & Capcn. 1980
Barsolli and Van Miller. 1984
(Guluh)
Ovcnnann el al.. 1987
Notes;
*No relevant Held studies were found.
Dose to rhesus monkey calculated using fixid ingestion rale of 0.2 kg/day and hody weight of 5 kg (Sample el a]., IV96)
Page 2 of 2
TAMS/MCA
-------
TABLE B-18
TOXIC1TY ENDPOINTS FOR OTHER MAMMALS - LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
• DOSE
DIOXIN
EQUIVALENTS
(ugTEQ/kg/day)«
EFFECT ENDPOINT
REFERENCE
Laboratory studies
Hamster
Mouse
Dog
Rabbil
Rhesus monkey
'Macaca mulatra )
Rat
Guinea pig
Oral
Oral
Oral
Oral
Oral
Oral
Oral
Single dose
Single dose
Single dose
Single dose
Single dose
Single dose
Single dose
LD<,,
LDW
LD,,
LDS,
LDW
LDW
LD,,
2,3,7,8-TCDD
2,3.7.8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2.3,7.8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
1.160-5.050
1 14 -284
about 100 • 200
115
approx. 70
22-45
0.6-2.1
Mortality
Mortality
Mortality
Mortality
Mortality
Mortality
Mortality
Kociba and Schwctz, 1982
Kociba and Schwetz. 1982
Kociba and Schwetz, 1982
Schwetz et al., 1973
Kociba and Schwetz. 1982
Schwetz etal., 1973
Schwetz el al.. 1973
Rat
Ral
Rat
Rhesus monkey
(Macaca mulatta )
Rhesus monkey
Macaca mulatta )
Rat
Rat
Rat
Rhesus monkey
(Macaca mulatta )
Gestation days 6 to 15
2 years
3 generations
7 months
7-48 months, maternal
Gestation days 6 to 15
2 years
3 generations
7 to 48 months, maternal
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
NOAEL
NOAEL
NOAEL
NOAEL
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3.7.8-TCDD
0.25
O.I
0.01
0.0021
0.00059
0.125
0.01
0.001
0.00012
Litter size, pup weight
Female mortality
Reproductive capacity
Number of births
Reproductive
Litter size, pup weight
Female mortality
Reproductive capacity
Reproductive
Khera and Ruddick. 1973
Kociba et al.. 1978
Murray el al .,1979
Allen etal.. 1979
Bowman etal.. 1989b
Khera and Ruddick, 1973
Kociba etal.. 1978
Murray etal. 1979
Bowman etal.. 1989
TAMS/MCA
-------
TABLE B-19
TOXICITY ENDPOINTS FOR MINK - LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
EXPOSURE
MEDIA
EXPOSURE
DURATION
EFFECT
LEVEL
PCB TYPE
EFFECTIVE
DOSE
(mg/kg/day)
EFFECTIVE
FOOD CONC.
. (mg/kg)
EFFECT ENDPOINT
REFERENCE
Laboratory studies
Mink (Mustela vixifin )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Diet
Diet
Diet
Diet
Diet
4 weeks
4 weeks
4 weeks
4 weeks
9 months
LD50
LD50
LD50
LD50
LD50
Aroclor 1254
Aroclor 1 254
Aroclor 1254
Aroclor 1254-
(weathered)
Aroclor 1254
11.5
10.8
6.4
6.4
0.9
84
79
47
47
6.6
Adult mortality
Adult mortality
Adult mortality
Adult mortality
Mortality
Homshaw (1984), as cited in
Aulerich el al. (1986)
Aulerich et al. (1986)
Homshaw el al. (1986)
Aulerich etal. (1986)
Ringer et al. (1981)
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vision )
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
Diet
8 months
8 months
4 weeks
8 months
3 months
3 months
8 months
4 months
105 days
66 days
4 months
6 months
160 days
EL-effect
EL-effect
LOAEL
LOAEL
EL-effect
EL-effect
LOAEL
LOAEL
LOAEL
LOAEL
EL-effect
EL-effect
LOAEL
Aroclor 1016
Aroclor 1016
Aroclor 1254
Aroclor 1242
Clophen A-50
Aroclor 1254
Aroclor 1242
Aroclor 1254
Aroclor 1254
(weathered)
Not reported
Aroclor 1254
Aroclor 1254
Aroclor 1254
(weathered)
2.7
2.7
1.4
1.4
2
2
0.7
0.7
0.5
0.5
0.3
0.1
0.09
20
20
10
10
Not reported
Not reported
5
5
3.57
3.3
2.5
1
0.64
Reduced birth weight and growth rate of kits
Adult mortality
Reduced weight gain in juveniles
Adult mortality
Decreased number of kits born alive
Decreased number of kits bom alive
Reduced reproduction
Decreased number of kits bom alive
Adult mortality
Decreased number of kits born alive
Decreased number of kits born alive
Reduced growth rates of kits
Reduced number of kits bom alive
Bleavins et al., 1980
Bleavins et al., 1980
Homshaw etal. (1986)
Bleavins et al., 1980
Kihlstom etal., 1992
Kihlstometal., 1992
Bleavins etal., 1980
Aulerich and Ringer (1977)
Platonow & Karstad (1973)
Jensen etal. (1977)
Aulerich etal. (1985)
Wren etal., 1987
Platanow & Karstad (1973)
Mink (Mustela vision )
Mink (Mustela vision )
Diet
Diet
8 months
4 months
NOAEL
NOAEL
Aroclor 1242
Aroclor 1 254
0.9
0.1
5
1
Adult mortality
Decreased number of kits bom alive
Bleavins et al., 1980
Aulerich & Ringer (1977)
12/27/99
Page 1 of 1
TAMS/MCA
-------
TABLE B-20
TOXICITY ENDPOINTS FOR MINK - FIELD STUDIES
EFFECTIVE DIETARY DOSES OF TOTAL PCBs AND AROCLORS
SPECIES
FIELD
COMPONENT
STUDY DURATION
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
DOSE
(mg/kg/day)
EFFECTIVE
FOOD CONC.
(mg/kg)
EFFECT ENDPOINT
REFERENCE
Field studies
Mink (MiLiteta vixitm )
Mink (Muxtela visitm )
Mink (Mustcla visiiin )
Fed contaminated caip
from Saginaw Bay, Ml
Fed contaminated carp
from Saginaw Bay, Ml
Fed contaminated carp
from Saginaw Bay, Ml
Mink were fed prior to and
throughout the reproductive
period
Mink fed prior to breeding
and over two generations
Mink fed prior to breeding
and over two generations
LOAEL
LOAEL
LOAEL
PCBs, TEQs, others
PCBs, pesticides
PCBs, pesticides
0.1.1
0.08
0.04
N/A
0.5
0.25
Reproductive success, growth/survival of
offspring
Kit survival
Reduced growth rate of kits
Heatonetal. (1995)
Restumetal., 1998
Resrumetal., 1998
Mink (Mustcla vision )
Mink (Muslela vixiim )
Fed contaminated carp
from Saginaw Bay, Ml
Fed contaminated carp
from Saginaw Bay. Ml
Mink fed prior to breeding
and over two generations
Mink were fed prior to and
throughout the reproductive
period
LOAEL
NOAEL
PCBs, pesticides
PCBs, TEQs, others
0.04
0.004
0.25
N/A
Kit survival
Reproductive success, growth/survival of
offspring
Restumetal., 1998
Heatonetal. (1995)
12/27/99
Page 1 of 1
TAMS/MCA
-------
TABLE B-21
TOXICITY ENDPO1NTS FOR MINK - LABORATORY STUDIES
EFFECTIVE DIETARY DOSES OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
FIELD
COMPONENT
STUDY
DURATION
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
DOSE (mg/kg/day)
EFFECTIVE
DOSE
DIOXIN
EQUIVALENTS
(ue TEO/ke/day)
EFFECT
ENDPOINT
REFERENCE
Laboratory studies
Mink kits
(Mustela vixtm )
Mink males
(Muxtela vixtm )
Intraneritoncal
Oral
12 days
Single dose
LD,,
LD,,
2.3.7.K-TCDD
2.3.7.X-TCDD
<().()!
4.2
< 0.01
4.2
Mortality
Mortality
Aulcnch ct id., 1988
Hochstein el al., WHH
TAMS/MCA
-------
TABLE B-22
TOXICITY ENDPOINTS FOR MINK - FIELD STUDIES
EFFECTIVE DIETARY DOSES OF DIOXIN TOXIC EQUIVALENTS (TEQs)
SPECIES
FIELD
COMPONENT
STUDY DURATION
EFFECT
LEVEL
CONTAMINANT
TYPE
EFFECTIVE
DOSE
DIOXIN
EQUIVALENTS
(tig TEO/ke/dav)
EFFECT
ENDPOINT
REFERENCE
Field studies
Mink (Mustela vision )
Mink (Muxtelu vision )
Mink (Mustela vision )
Fed contaminated carp
frum Saginaw Bay, MI
Fed contaminated carp
from Suginaw Bay, MI
Fed contaminated carp
from Saginaw Bay, MI
Fed prior to and throughout
breeding period
Fed prior to and throughout
breeding period
Fed prior to and throughout
breeding period
LOAEL
LOAEL
LOAEL
TEQs. pesticides
TEQs
(chemically derived)
TEQs
(hioassay derived)
0.0036
0.00224
0.00027
Growth rate of kits
Growth and survival
rate of kits
Growth and survival
rate of kits
Hcatonetal. (1995)
Tillitt et al.. 1996
Tillittetal., 1996
Mink (Mustela vision )
Mink (Mustela vision )
Mink (Mustela vixitm )
Fed contaminated carp
from Saginaw Bay, MI
Fed cimtaminaled carp
from Saginaw Bay, MI
Fed contaminated carp
from Saginaw Bay. Ml
Fed prior to and throughout
breeding period
Fed prior to and throughout
breeding period
Fed prior to and throughout
breeding period
NOAEL
NOAEL
NOAEL
TEQs
(bioassay derived)
TEQs, pesticides
TEQs
(chemically derived)
0.00344
0.00025
O.OOTO8
Growth and survival
rale of kits
Growth rate of kits
Growth and survival
rate of kits
Tillilt et al., 1996
Hcatnnetal. (1995)
Tillilt el al.. 1996
TAMS/MCA
-------
TABLE B-23
TAXONOMY OF STUDIED ORGANISMS
Phylum
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Class
Mammalia
Mammalia
Mammalia
Mammalia
Mammalia
Mammalia
Mammalia
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Aves
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Subclass
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Aclinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Order
Carnivora
Carnivora
Carnivora
Chiroptera
Lagomorphus
Rodentia
Rodentia
Anseriformes
Anseriformes
Charadriiformes
Charadriiformes
Charadriiformes
Ciconiiformes
Coraciiformes
Falconiiformes
Falconiiformes
Falconiiformes
Galliformes
Galliformes
Galliformes
Galliformes
Passeriformes
Passeriformes
Passeriformes
Passeriformes
Passeriformes
Pelecaniformes
Strigiformes
Acipenseriformes
Beloniformes
Clupeiformes
Cypriniformes
Cypriniformes
Cypriniformes
Cypriniformes
Cypriniformes
Cypriniformes
Perci formes
Perciformes
Perciformes
Perciformes
Perciformes
Perciformes
Perciformes
Family
Mustelidae
Mustelidae
Procyonidae
Vespertilionidae
Leporidae
Muridae
Muridae
Anatidae
Anatidae
Laridae
Laridae
Laridae
Ardeidae
Alcedinidae
Accipitridae
Falconidae
Pandionidae
Phasianidae
Phasianidae
Phasianidae
Phasianidae
Hirundinidae
Icteridae
Icteridae
Icteridae
Sturnidae
Phalacrocoracidae
Strigidae
Acipenseridae
Adrianichthydiae
Clupeidae
Catostomidae
Cyprinidae
Cyprinidae
Cyprinidae
Cyprinidae
Cyprinodontidae
Centrarchidae
Centrarchidae
Centrarchidae
Moronidae
Moronidae
Percidae
Sciaenidae
Genus .
Lutra
Mustela
Procyan
Myotix
[Sylvilagusl
[Pertanyscus]
[Ratlux]
Aix
Anas
Hydropogne
Sterna
Sterna
Ardea
Cen/te
Haliaeetus
Falco
Pandiim
Colinus
Ciiturnix
Callus
Phaxianux
Tachycineta
Agelaius
Mnlothrus
Quiscalus
Sturnux
Phalacnicorax
Otux
Acipenser
Oryzias
Clupea
Cataxlomus
Daniit
Notropix
Ptuixinus
Pimephalus
Fundulux
Lepomis
Lepomis
Micrnpterus
Mitrtme
Morime
Perca
Leitistumus
Species
canaJensis
vision
lotor
lucifugux
[transitionalis]
[piiliiuwtus]
[rattus]
sponsa
plafyrhynchfts
caspia
hirundo
fnrsleri
herodias
alcyim
leucocephalus
xparvenius
haliaeetus
virginianux
coturnix
dftmexticux
cttlchicus
hicotttr
phoeniceux
ater
quiscula
vulgaris
auritus
axio
brevirostrum
lalipex
harengux
CDtnmersimi
danio
HuJxtinius
phoxinus
promelas
hetemclitux
gihhi>xu.i
aurilux
xulmniJes
americana
.taxalilix
flavexcenx
xanthurux
Common name
River Otter
Mink
Raccoon
Little Brown Bat
Rabbit [Eastern Cottontail]
Mouse [Oldfield Mouse]
Rat
Birds
Wood Duck
Mallard Duck
Caspian tem
Common tern
Forster's tern
Great Blue Heron
Kingfisher
Bald Eagle
American Kestrel
Osprey .
Northern Bobwhite
Japanese Quail
Domestic Chicken
Ring-Necked Pheasant
Tree Swallow
Red- Winged Blackbird
Brown-Headed Cowbird
Common Crackle
European Starling
Double-Crested Cormorant
Screech Owl
Fish
Shortnose Sturgeon
Medaka
Baltic Herring
White sucker
Zebrafish
Spottail Shiner
Minnow
Fathead Minnow
Killifish
Pumpkinseed
Redbreast Sunfish
Largemouth Bass
White Perch
Striped Bass
Yellow Perch
Spot
Page I of 2
TAMS/MCA
-------
TABLE B-23
TAXONOMY OF STUDIED ORGANISMS
Phylum
Chordata
Chorda ta
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata .
Chordata
Chordata
Chordata
Chordata
Chordata
Class
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Pisces
Subclass
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Actinopterygii
Order
Perci formes
Pleuronectiformes
Pleuronectiformes
Pleuronectiformes
Pleuronectiformes
Salmoniformes
Salmoniformes
Salmoniformes
Salmoniformes
Salmoniformes
Salmoniformes
Salmoniformes
Siluriformes
Siluriformes
Family
Sparidae
Pleuronectidae
Pleuronectidae
Pleuronectidae
Pleuronectidae
Esocidae
Salmonidae
Salmonidae
Salmonidae
Salmonidae
Salmonidae
Salmonidae
Ictaluridae
Ictaluridae
Genus
LugoJon
Parnphrys
Platichthys
Platichthys
Pseudopleuronectes
Esux
Coregimus
Oncorhynchus
Salmo
Salvelinus
Salveliniu
Salvelinus
Iclalurus
Ictalurus
Species
rhtimbmdes
velulus
flesus
xtellatus
americanus
lucius
arteJii
tshawylscha
gairdneri
alpinus
fftntinalis
namaycush
nehuloxux
punctatus
Common name
Pinfish
English Sole
Baltic Flounder
Starry Flounder
Winter Rounder
Northern Pike
Lake Herring
Chinook Salmon
Rainbow Trout
Arctic Charr
Brook Trout
Lake Trout
Brown Bullhead
Channel Catfish
Page 2 of 2
TAMS/MCA
-------
TABLE B-24
STANDARD ANIMAL BODY WEIGHTS AND FOOD INTAKE RATES
Animal
MAMMALS
Mink
Mouse
Mean Mouse
Mouse, Oldfield
Rabbit
Rhesus Monkey
Rat
Mean Rat
BIRDS
Blackbird, Red- Winged
Chicken, Domestic-adult
Mean Chicken, Domestic—adult
Chickens, Domestic-chick
Mean Chicken, Domestic-chick
Cowbird, Brown-headed
Dove, Ringed
Duck, Mallard-adult
Mean Duck, Mallard— adult
Duck, Mallard— duckling
Kestrel, American
Owl, Screech
Pheasant, Ring-necked
Quail, Japanese
Quail, Japanese— 3 months
Body Weight
(kg)
1
0.03
0.028
0.029
0.014
3.8
5
0.35
0.435
0.303
0.273
0.365
0.26
0.331
0.064
1.6
1.5
1.55
0.121
0.534
0.3275
0.049
0.155
1
1.153
1.15
1
1.17
1.0946
0.782
0.13
0.181
1
0.15
0.072
Food Ing.
Rate (g/d)
1.9
115
78.2
25
Food
Ingestion Rate
(kg/d)
0.137
0.0055
0.0019
0.135
0.2
0.028
0.0375
0.03275
0.0137
0.11
0.106
0.108
0.0126
0.044
0.0283
0.01087
0.017
0.1
0.11
0.115
0.128
0.121
0.1148
0.0782
0.01
0.025
0.0582
0.0169
Food/actor
(kg/kg body wt/d)
0.137
0.180
0.034
0.040
0.080
0.137
0.099
0.214
0.069
0.071
0.070
0.104
0.082
0.086
0.222
0.110
0.100
0.095
0.100
0.128
0.103
0.105
0.100
0.077
0.138
0.058
0.113
Note: All values are from Toxicological Benchmarks for Wildlife: 1996 Revision (USEPA, 1996) unless othi
TAMS/MCA
-------
TABLE B-25
TOXICITY REFERENCE VALUES FOR FISH
DIETARY DOSES AND HOG CONCENTRATIONS OF TOTAL PCBs AND DIOXIN TOXIC EQUIVAI.ENTS (TEQs)
TRVs
Tissue Concentration
Lab-based TRVs for PCBs (mg/kg wel wl.)
Field-based TRVs for PCBs (mg/kg wet wt.)
Pumpkinseed
(Lepoiais giooosus )
I.OAEI.
NOAEI.
LOAKL
NOAEI,
1.5
0.16
NA
0.5
Spotlail
Shiner
(Notropis
hudsonius )
/5
1.6
NA
NA
Brown Bullhead
(laalunu ncbulosus )
1.5
0.16
NA
NA
Yellow Pen*
(Perca/lavesceni )
1.5
0.16
NA
NA
(Morone americana )
1.5
0.16
NA
3.1
Largemoutb Bass
(Mifroptena
salmoutes)
1.5
0.16
NA
0.5
Egg Concentration
Lab-based TRV for TEQs dig/kg lipid)
from salmonids
Lab-based TRV for TEQs (ug/kg lipid)
from nun-salmonid.s
Field-based TRVs for TEQs (ug/kg lipid)
LOAEL
NOAEI.
I.OAEI.
NOAEI.
I.OAEI.
NOAEI.
0.6
0.19
10.3
0.54
NA
NA
Nol derived
Not derived
103
S.-l
NA
NA
18
8.0
Not derived
Not derived
NA
NA
0.6
0.29
10.3
0.54
NA
NA
0.6
0.29
10.3
0.54
NA
NA
0.6
0.29
10.3
0.54
NA
NA
Striped Ban
(Morone rarftf^itt )
Shortnose Sturgeon
(Acipenser
btevliustmn)
1.5
0.16
NA
3.1
I.S
0.16
NA
NA
0.6
0.29
10.3
0.54
NA
NA
0.6
0.29 .
10.3
0.54
NA
NA
References
Beng!sson(1980)
While perch and striped bass: Westin el
al. (1983)
Pumpkinseed and Largemouth bass:
Adams etal. (1989, 1990. 1992)
Brown Bullhead: Elonen et al. ( 1998)
All oihers: Walker et al. (1994)
Oliveri and Cooper (1997)
Note:
' I'umpkinseed (Leiximis gibbnsus ) and spottail shiner (Nolropis hudsonius)
Units vary for l>CBs and TEQ.
NA = Nol available
Selected TRVs are bolded and italicized.
TAMS/MCA
-------
TABLE B-26
TOXICITY REFERENCE VALUES FOR BIRDS
DIETARY DOSES AND EGG CONCENTRATIONS OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
TRVs
Dietary Dost
Lab-based TRVs for PCBs (mg/kg/day)
Field-based TRVs for PCBs (mg/kd/day)
Lab-based TRVs for TEQs (ug/kg/day)
Field-based TRVs for TEQs (ug/kg/day)
Tret Swallow
(Tachycineta bicolnr )
Mallard Duck
(Anax platyrhychos )
Belled Kingfisher
(Ceryle alcyon )
Great Blue Heron
(Ardea herodiax )
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
0.07
0.01
NA
16.1
0.014
0.0014
NA
4.9
2.6
0.26
NA
NA
0.014
0.0014
NA
NA
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
Bald Eagle
(Hattacau!
leucocephalux )
References
0.07
0.01
NA
NA
Mallard: Cuslcr and Heinz (19X0)
All others: Scou( 1977)
Trcc Swallow: US EPA Phase 2 Database ( 1 998)
0.014
0.0014
NA
NA
NiBcket al. (1992)
US EPA Phase 2 Database (1998)
Egg Concentration
Lab-based TRVs for PCBs (rag/kg egg)
Fidd-based TRVs for PCBs (mg/kg egg)
Lab-based TRVs for TEQs (ug/kg egg)
Field-based TRVs for TEQs (ug/kg egg)
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
2.21
0.33
NA
26.7
0.02
0.01
NA
13
2.21
0.33
NA
NA
0.02
0.01
0.02
0.005
2.2/
0.33
NA
NA
0.02
0.01
NA
NA
2.21
0.33
NA
NA
NA
2
0.5
0.3
2.21
0.33
NA
3.0
0.02
0.01
NA
NA
Saitt(l977)
Bald Eagle: Wiemcyer (1984, 1993)
Tree Swallow: US EPA Phase 2 Database (1998)
Great Blue Heron: Janz and Bellward (1996)
Others: Powell etal. (I996a)
Mallard: White and Segniak (1994); White and Hoffman (1995)
Great Blue Heron: Sanderson et al. (1994)
Tree Swallow: US EPA Phase 2 Database (1998)
Note: Units vary for PCBs and TEQ.
NA = Not Available
Selected TRVs arc balded and ilalkiud.
TAMS/MCA
-------
TABLE B-27
TOXICITY REFERENCE VALUES FOR MAMMALS
DIETARY DOSES OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
TRVs
Lab-based TRVs for PCBs (mg/kg/day)
Field-based TRVs for PCBs (mg/kg/day)
Lab-based TRVs for TEQs (ug/kg/day)
Field-based TRVs for TEQs (ug/kg/day)
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
Little Brown Bat
(Myotis lucifugus )
0.15
0.032
NA
NA
0.001
0.0001
NA
NA
Raccoon
(Procyon lotor)
0.15
0.032
NA
NA
0.001
0.0001
NA
NA
Mink
(Mustcla
vison)
0.07
0.01
0.13
0.004
0.00 1
0.0001
0.00224
0.00008
Otter
(Lutra canadensis )
0.07
0.01
0.13
0.004
0.001
0.0001
0.00224
0.00008
References
Mink amjullen Aulerich and Ringer (1977)
Race,. in and hat: Under el al. (1984)
Hcau.nelal. (1995)
Murray el al. ( 1 979)
Tillillctal. (1996)
Note: Units vary for PCBs and TEQ.
Note: TRVs for raccoon and bat are based on mulil-generalional studies to which interspecies uncertainly factors are applied.
NA = Not Available
Final selected TRVs are bolded and italicized.
TAMS/MCA
-------
TABLE B-28: WILDLIFE SURVEY RESULTS Amphibians
Hudson River
New York
Information Source .:.-,..• | Date | Contact | Response | Contact Information (Data Available (Information/Findings .,';*?*-C
Amphibians
Amphibian Expert
NYSDEC - Amphibian and
Reptile Atlas Project
NYS Department of
Environmental Conservation -
Endangered Species Unit
NYS Department of
Environmental Conservation
NYSDEC
Ndakinna Wilderness Project
l-Jun-99
3-Jun-99
8-Jun-99
8-Jun-99
16-Jun-99
6/3/1999
6/16/99
Email
Email
WWW
WWW
Call
Email
Call
Call
Yes
No
No
No
Yes
No
No Yes
Thomas Palmer, frog consultant
for Wellesley Project;
Ophis@world.std.com
herps @gw.dec.state.ny.us;
http://www.dec.state.ny.us/website
/dfwmr/wildlife/herp/index.html
www.dec.state.ny .us/website/dfw
mr/wildlife/endspec/enspamphib.ht
ml
www.dec.state.ny. us/website/dfw
mr/wildlife/herp/atproj.html
Mark Brown (518) 623-3671
Jim Brushek (518) 583-9980x3, 23
Middle Grove Road, Greenfield
Center, NY 12833; Received
address from Saratoga County
Information - Annamaria Dalton
(annamaria@spa.net)
He doesn't know anything
about PCB effects on frogs;
posted message on amphibian
web page
Brief summaries, listed by
species, for NY state.
10 year survey documenting
geographic distribution of
herpetofauna in NY state.
Familiar with the area
regarding mammals, birds, and
herps. Good source. See
General Info page.
Professional Tracker
Recommended the following website: http://cciw.ca/green-Iane/herptox/
Eurycea longicauda (Longtail Salamander): nocturnal salamander
which occupies shallow rocky streams and moist forested areas. Found
in Cattaraugus County and mid Hudson Valley. Very few in NY.
Status: Special Concern.
Common frogs and toads abundant, snapping turtles abundant, some
box turtles present.
Reports snapping and painted turtles, red back and two-line
salamanders. Frogs: bull, spring peepers, gray tree, northern leopard,
and pickeral. American toad. Garter and water snakes (none are
poisonous). Currently working on a herp survey.
Common amphibians present in strong numbers. Box, snapper, and
painted turtles. Some snakes which he could not identify.
Page I of 1
TAMS/MCA
-------
THIS PAGE LEFT INTENTIONALLY BLANK
-------
2 (ortho) 6' (ortho)
3 (meta) 5' (meta)
4 (para)
5 (meta)
4' (para)
6 (ortho) 2' (ortho)
' (meta)
Figure B-l: Shape of Biphenyl and Substitution Sites
TAMS/MCA
-------
100
GO
•a
&
o
o
C
o
o
o
10
1 -
0.1
Figure B-2
Selected Fish Aroclor and Total PCB Toxicity Endpoints
46: Juvenile Spot, LOAEL, 20 days, adult mortality
36: Fathead Minnow, 16 weeks, LOAEL, spawning and fecundity
27: Juvenile Spot, NOAEL, 56 days, adult mortality
•11.6: Fathead Minnow, NOAEL, 16 weeks, spawning and fecundity
-3.8 (nominal dose): Killifish, LOAEL, 40 days observation, egg production and food consumption
-0.76 (nominal dose): Killifish, NOAEL, 40 days of observation, egg production and food consumption
TAMS/MCA
-------
Equivalent (ug TEQ/kg lipid'
ox
ect Concentration
M
O
1
Figure B-3
Selected Fish Egg Dioxin Equivalent Toxicity Endpoints
Endpoint: Early Life Stage Mortality
-49: White sucker, LOAEL
-34: White sucker, NOAEL
-18: Channel catfish, LOAEL
-8.0: Channel catfish, NOAEL
-0.7: Lake trout, LOAEL
-0.43: Lake trout, NOAEL
TAMS/MCA
-------
Effective Dose (mg PCB/kg wet body wt./day)
5
> P H- 0 0
i h— 1 H- 1 O • O O
I I 1 1 1
Figure B-4
Selected Bird Diet Aroclor and Total PCB Toxicity Endpoints
i
[IS
r*f£.
Mj
1
m
n
i
^
«^
^'
i
s
• • 853- Mallard Duck LD 5 days
16.1: Tree swallow, NuAbL, held study, reproductive output
2 6* Mallard Duck EL (no effect) appro\ 1 month reproductive success
11- Ringed Turtle Dove EL (effect) hatching success
1 o 7i Domestic Chicken LOAEL hatching success
0.4: Screech Owl, EL (no effect), > 8 weeks, egg production, hatching success,
and fledging success
U.U1
TAMS/MCA
-------
Figure B-5
Selected Bird Diet Dioxin Equivalent Toxicity Endpoints
Dioxin Equivalents (ug TEQ/kg wet body wt./da
> H- S
* H* O O
1 1 1
-------
Figure B-6
Selected Bird Egg Aroclor and Total PCB Toxicity Endpoints
Effective Egg Concentration (mg PCB/kg egg)
I P 0
-» N- 1 1— ' O O
till 1
1
1
*
IS
OKt
•m
is
s
I
26.7: Tree swallnu/ NftAFI , rrprnrtucriv? niilput
5: Domestic chicken, LOAFL, hatching success
3.0: Bald eagle. NOAEL. reproductive success
1.7: Domestic chicken. LOAEL, hatching success
TAMS/MCA
-------
Figure B-7
Selected Bird Egg Dioxin Equivalent Toxicity Endpoints
Endpoint: Embryo Mortality
Effective Dose Dbxin Equivalents (ug TEQ/kg egg)
§ b p i
H^ H* h* 1— i O C
' i i i i 1
*
f &
M
1
i
P
it
i
i
1
f
80: Cormorant, LOAEL
40: Cormorant, NOAEL
23: American Kestrel, LOAEL
1 :»: Tree Svwillnw, NOAF.I .
• — 5: American Kestrel, LOAEL
1 4: Common Tern, LOAEL
1 — 4: Cormorant, LOAEL
1 2.3: American Kestrel, NOAEL,
1 1: Coiinuianl, NOAEL
Oft? WnnH Hii^lf I OAFI
TAMS/MCA
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Figure B-8
Selected Mink Aroclor and Total PCB Toxicity Endpoints
I 11.5: LD50,
1 10.8: LD,0,
4 weeks, adult mortality
4 weeks, adult mortality
-T 6.4: LD50, 4 weeks, adult mortality
6.4: LD50, 4 weeks, adult mortality (weathered PCBs)
-1.4: LOAEL, 4 weeks, reduced weight gain in juveniles
-0.91: LC50,9 months, mortality
-0.69: NOAEL, 4 months, decreased number of kits born live
-0.49: LOAEL, 105 days, adult mortality
-0.34: EL, 4 months, decreased number of kits born live
0.14: NOAEL, 4 months, decreased number of kits born live
_0.14: EL, 6 months, reduced growth rates of kits
-0.09: LOAEL, 160 days, reduced number of kits born alive
TAMS/MCA
-------
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Figure B- 10
Selected Mammal Dioxin Equivalent Toxicity Endpoints
s~\
>> -,
ffective Dose Dioxin Equivalents (ug TEQ/kg wet body wt./da
! 1 §
* h* h* h^ H
m U.UUUl J
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rf
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0.25: Rnt, l,r>AF,l gestation days fi- 1 ^, I'tfrsi/f and pup weight
.01; Rat 1 OAEL ? years female mortality
... . 00021' Rhesus Monkey LOAFI 7 months number of births
.. 0^001: Rat NOAFI ^ generations, reproductive capacity
, 000059' Rhesus Monkey LOAFI 7-48 months (maternal) reproductive
000012* Rhesus Monkey NOAEL 7 48 months (maternal) reproductive
TAMS/MCA
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