HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
Vv/
\
w
O
For
U.S. Environmental Protection Agency
Region 2
and
U.S. Army Corps of Engineers
Kansas City District
Book 1 of 1
TAMS Consultants, Inc.
Menzie-Cura & Associates, Inc.
-------
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
REGION 2
? 290 BROADWAY
NEW YORK, NY 10007-1866
August 29,2000
To All Interested Parties:
The U.S. Environmental Protection Agency (USEPA) is pleased to release the
Responsiveness Summary for the baseline Ecological Risk Assessment for Future Risks in the
Lower Hudson River (ERA Addendum), which is part of Phase 2 of the Reassessment Remedial
Investigation/Feasibility Study for the Hudson River PCBs Superfund Site. For complete coverage,
the ERA Addendum and this Responsiveness Summary should be used together.
In the Responsiveness Summary, USEPA has responded to all significant written comments
received during the public comment period on the ERA Addendum. In addition, the Responsiveness
Summary contains revised calculations of ecological risks based on the January 2000 Revised
Baseline Modeling Report and comments received on the Ecological Risk Assessment and the ERA
Addendum. Importantly, the overall conclusions regarding the future risks to ecological receptors
due to PCBs in the Lower Hudson River remain unchanged.
If you need additional information regarding the Responsiveness Summary for the ERA
Addendum or the Reassessment RI/FS in general, please contact Ann Rychlenski, the Community
Relations Coordinator for this site, at (212) 637-3672.
Sincerely yours,
lichard L. Caspe, Director
A Emergency and Remedial Response Division
Internet Address (URL) • http://www.epa.gov
Recycled/Recyclable • Printed with Vegetable OD Based Inks on Recycled Paper (Minimum 30% Postconsumer)
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
For
U.S. Environmental Protection Agency
Region 2
and
U.S. Army Corps of Engineers
Kansas City District
Book 1 of 1
TAMS Consultants, Inc.
Menzie-Cura & Associates, Inc.
-------
HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
Page
TABLE OF CONTENTS i
LIST OF TABLES xiv
LIST OF FIGURES xix
LIST OF ACRONYMS xxi
I. INTRODUCTION AND COMMENT DIRECTORY
1. Introduction 1
2. Commenting Process 2
2.1 Distribution of ERA Addendum 2
2.2 Review Period and Public Availability Meetings 2
2.3 Receipt of Comments 2
2.4 Distribution of the Responsiveness Summary 2
3. Organization of ERA Addendum Comments and Responses to Comments 6
3.1 Identification of Comments 6
3.2 Location of Responses to Comments 6
4. Comment Directory 7
4.1 Guide to Comment Directory . 7
4.2 Comment Directory 7
II. RESPONSES TO COMMENTS ON THE ERA ADDENDUM FOR FUTURE RISKS
IN THE LOWER HUDSON RIVER
General Comments 13
EXECUTIVE SUMMARY 21
1.0 INTRODUCTION 21
1.1 Purpose of Report 22
1.2 Report Organization 22
2.0 PROBLEM FORMULATION 22
i TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
2.1 Site Characterization 23
2.2 Contaminants of Concern 23
2.3 Conceptual Model 23
2.3.1 Exposure Pathways in the Lower Hudson River Ecosystem 23
2.3.2 Ecosystems of the Lower Hudson River 23
2.3.3 Exposure Pathways 23
2.3.3.1 Aquatic Exposure Pathways 23
2.3.3.2 Terrestrial Exposure Pathways 23
2.4 Assessment Endpoints 24
2.5 Measurement Endpoints (Measures of Effect) 24
2.6 Receptors of Concern 25
2.6.1 Fish Receptors 25
2.6.2 Avian Receptors 25
2.6.3 Mammalian Receptors 25
2.6.4 Threatened and Endangered Species 25
2.6.5 Significant Habitats 25
3.0 EXPOSURE ASSESSMENT 25
3.1 Quantification of PCB Fate and Transport: Modeling Exposure Concentrations 31
3.1.1. Modeling Approach 31
3.1.1.1 Use of the Farley Model 31
3.1.1.2 Use of FISHRAND 32
3.1.1.3 Comparison to the March 1999 Farley Model (1987-1997) 32
3.1.1.4 Comparison Between Model Output and Sample Data 37
3.1.1.5 Comparison of White Perch Body Burden between the Farley
Model (Using Upper River Loads from HUDTOX) and
FISHRAND 38
3.1.1.6 Comparison Between FISHRAND Output and Sample Data ... 38
3.1.2. Model Results 38
3.1.2.1 Farley Model Forecast Water Column and Sediment
Concentrations 38
3.1.2.2 Farley Model Forecast Fish Body Burdens 38
11 TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
Page
3.1.2.3 FISHRAND Forecast Fish Body Burdens 39
3.1.3 Modeling Summary 39
3.2 Exposure Point Concentrations 39
3.2.1 Modeled Water Concentrations 46
3.2.2 Modeled Sediment Concentrations 46
3.2.3 Modeled Benthic Invertebrate Concentrations 47
3.2.4 Modeled Fish Concentrations 47
3.3 Identification of Exposure Pathways 48
3.3.1 Benthic Invertebrate Exposure Pathways 48
3.3.2 Fish Exposure Pathways 48
3.3.3 Avian Exposure Pathways, Parameters, Daily Doses, and Egg
Concentrations 48
3.3.3.1 Summary of ADD^p^d, ADD9S%UCL, and Egg Concentrations
for Avian Receptors 49
3.3.4 Mammalian Exposure Pathways, Parameters, and Daily Doses 49
3.3.4.1 Summary of ADDExpecled and ADD9JWJCL for Mammalian
Receptors 49
4.0 EFFECTS ASSESSMENT 49
4.1 Selection of Measures of Effects 58
4.1.1 Methodology Used to Derive TRVs 58
4.1.2 Selection of TRVs 59
5.0 RISK CHARACTERIZATION 59
5.1 Evaluation of Assessment Endpoint: Benthic Community Structure as a Food
Source for Local Fish and Wildlife 60
5.1.1 Do Modeled PCB Sediment Concentrations Exceed Appropriate
Criteria and/or Guidelines for the Protection of Aquatic Life and
Wildlife? 61
5.1.1.1 Measurement Endpoint: Comparisons of Modeled Sediment
Concentrations to Guidelines 61
,„ TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5.1.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria
and/or Guidelines for the Protection of Aquatic Life and Wildlife? — 61
5.1.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations of PCBs to Criteria 62
5.2 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Local Fish Populations 62
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? 62
5.2.1.1 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for Forage
Fish 62
5.2.1.2 Measurement Endpoint: Comparison of Modeled PCB TEQ
Fish Body Burdens to Toxicity Reference Values for Forage
Fish 62
5.2.1.3 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for Brown
Bullhead 62
5.2.1.4 Measurement Endpoint: Comparison of Modeled TEQ Basis
Fish Body Burdens to Toxicity Reference Values for Brown
Bullhead 64
5.2.1.5 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for White
and Yellow Perch 64
5.2.1.6 Measurement Endpoint: Comparison of Modeled TEQ Basis
Body Burdens to Toxicity Reference Values for White and
Yellow Perch 64
5.2.1.7 Measurement Endpoint: Comparison of Modeled Tri+ PCB
Fish Body Burdens to Toxicity Reference Values for Large-
mouth Bass 64
iv TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5.2.1.8 Measurement Endpoint: Comparison of Modeled TEQ Based
Fish Body Burdens to Toxicity Reference Values for Large-
mouth Bass 64
5.2.1.9 Measurement Endpoint: Comparison of Modeled Tri+ PCB
Fish Body Burdens to Toxicity Reference Values for Striped
Bass 64
5.2.1.10 Measurement Endpoint: Comparison of Modeled TEQ Based
Fish Body Burdens to Toxicity Reference Values for Striped
Bass 65
5.2.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria
and/or Guidelines for the Protection of Aquatic Life and Wildlife? .... 65
5.2.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations of PCBs to Criteria 65
5.2.3 Do Modeled PCB Sediment Concentrations Exceed Appropriate
Criteria and/or Guidelines for the Protection of Aquatic Life and
Wildlife? 65
5.2.3.1 Measurement Endpoint: Comparisons of Modeled Sediment
Concentrations to Guidelines 65
5.2.4 What Do the Available Field-Based Observations Suggest About the
Health of Local Fish Populations? 65
5.2.4.1 Measurement Endpoint: Evidence from Field Studies 66
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) 66
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? 66
5.3.1.1 Measurement Endpoint: Modeled Dietary Doses on a Tri+
PCB Basis to Insectivorous Birds (Tree Swallow) 66
5.3.1.2 Measurement Endpoint: Predicted Egg Concentrations on a Tri+
PCB Basis to Insectivorous Birds (Tree Swallow) 66
v TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5.3.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed on a TEQ Basis to Insectivorous Birds (Tree
Swallow) 66
5.3.1.4 Measurement Endpoint: Predicted Egg Concentrations
Expressed on a TEQ Basis to Insectivorous Birds (Tree
Swallow) 67
5.3.2 Do Modeled Water Concentrations Exceed Criteria for Protection
of Wildlife? 67
5.3.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations to Criteria for the Protection of
Wildlife 67
5.3.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Insectivorous Bird Populations? 67
5.3.3.1 Measurement Endpoint: Evidence from Field Studies 67
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) 67
5.4.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Waterfowl and
Egg Concentrations Exceed Benchmarks for Adverse Effects on
Reproduction? 67
5.4.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ PCBs
to Waterfowl (Mallard) 68
5.4.1.2 Measurement Endpoint: Predicted Egg Concentrations of
Tri+ PCBs to Waterfowl (Mallard) 68
5.4.1.3 Measurement Endpoint: Modeled Dietary Doses of TEQ-
Based PCBs to Waterfowl (Mallard) 68
5.4.1.4 Measurement Endpoint: Predicted Egg Concentrations of TEQ-
Based PCBs to Waterfowl (Mallard) 68
5.4.2 Do Modeled PCB Water Concentrations Exceed Criteria for the
Protection of Wildlife? 68
vi TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5.4.2.1 Measurement Endpoint: Comparison of Modeled Water Con-
centrations to Criteria 68
5.4.3 What Do the Available Field-Based Observations Suggest About the
Health of Lower Hudson River Waterfowl Populations? 68
5.4.3.1 Measurement Endpoint: Observational Studies 69
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) 69
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? 69
5.5.1.1 Measurement Endpoint: Modeled Dietary Doses of Total
PCBs for Piscivorous Birds (Belted Kingfisher, Great Blue
Heron, Bald Eagle) 69
5.5.1.2 Measurement Endpoint: Predicted Egg Concentrations
Expressed as Tri+ to Piscivorous Birds (Eagle, Great Blue
Heron, Kingfisher) 69
5.5.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed as TEQs to Piscivorous Birds (Belted Kingfisher,
Great Blue Heron, Bald Eagle) 69
5.5.1.4 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed as TEQs to Piscivorous Birds (Belted Kingfisher,
Great Blue Heron, Bald Eagle) 69
5.5.2 Do Modeled Water Concentrations Exceed Criteria for the Protection
of Wildlife? 70
5.5.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria 70
5.5.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Piscivorous Bird Populations? 70
5.5.3.1 Measurement Endpoint: Observational Studies 70
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
Page
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) 70
5.6.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to
Insectivorous Mammalian Receptors Exceed Benchmarks for Adverse
Effects on Reproduction? 70
5.6.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Insectivorous Mammalian Receptors (Little Brown Bat) ... 70
5.6.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Insectivorous Mammalian Receptors (Little
Brown Bat) 70
5.6.2 Do Modeled Water Concentrations Exceed Criteria for Protection of
Wildlife? 71
5.6.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 71
5.6.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Insectivorous Mammalian Populations? 71
5.6.3.1 Measurement Endpoint: Observational Studies 71
5.7 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Omnivorous Mammal Populations (as represented
by the Raccoon) 71
5.7 1 Do Modeled Total and TEQ-Based PCB Dietary Doses to
Omnivorous Mammalian Receptors Exceed Benchmarks for Adverse
Effects on Reproduction? 71
5.7.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Omnivorous Mammalian Receptors (Raccoon) 71
5.7.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Omnivorous Mammalian Receptors (Raccoon) 71
5.7.2 Do Modeled Water Concentrations Exceed Criteria for Protection of
Wildlife? 72
TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5.7.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 72
5.7.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Omnivorous Mammalian Populations? 72
5.7.3.1 Measurement Endpoint: Observational Studies 72
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) 72
5.8.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to
Piscivorous Mammalian Receptors Exceed Benchmarks for Adverse
Effects on Reproduction? 72
5.8.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Piscivorous Mammalian Receptors (Mink, River Otter) ... 72
5.8.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Piscivorous Mammalian Receptors (Mink, River
Otter) 73
5.8.2 Do Modeled Water Concentrations Exceed Criteria for the Protection
of Piscivorous Mammals? 73
5.8.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 73
5.8.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Mammalian Populations? 73
5.8.3.1 Measurement Endpoint: Observational Studies 73
5.9 Evaluation of Assessment Endpoint: Protection of Threatened and
Endangered Species 73
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? 73
5.9.1.1 Measurement Endpoint: Inferences Regarding Shortnose
Sturgeon Population 73
IX TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
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? 74
5.9.2.1 Measurement Endpoint: Inferences Regarding Bald Eagle and
Other Threatened or Endangered Species Populations 74
5.9.3 Do Modeled Water Concentrations Exceed Criteria for the Protection
of Wildlife? 74
5.9.3.1 Measurement Endpoint: Comparisons of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 74
5.9.4 Do Modeled Sediment Concentrations Exceed Guidelines for the
Protection of Aquatic Health? 74
5.9.4.1 Measurement Endpoint: Comparisons of Modeled
Sediment Concentrations to Guidelines 74
5.9.5 What Do the Available Field-Based Observations Suggest About the
Health of Local Threatened or Endangered Fish and Wildlife
Species Populations? 74
5.9.5.1 Measurement Endpoint: Observational Studies 74
5.10 Evaluation of Assessment Endpoint: Protection of Significant Habitats 75
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? 75
5.10.1.1 Measurement Endpoint: Inferences Regarding Receptor
Populations 75
5.10.2 Do Modeled Water Column Concentrations Exceed Criteria for the
Protection of Aquatic Wildlife? 75
5.10.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife 75
5.10.3 Do Modeled Sediment Concentrations Exceed Guidelines for the
Protection of Aquatic Health? 75
TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5.10.3.1 Measurement Endpoint: Comparison of Modeled Sediment
Concentrations to Guidelines for the Protection of Aquatic
Health 75
5.10.4 What Do the Available Field-Based Observations Suggest About the
Health of Significant Habitat Populations? 75
5.10.4.1 Measurement Endpoint: Observational Studies 75
6.0 UNCERTAINTY ANALYSIS 76
6.1 Conceptual Model Uncertainties 76
6.2 Toxicological Uncertainties 76
6.3 Exposure and Modeling Uncertainties 76
6.3.1 Natural Variation and Parameter Error 76
6.3.2 Model Error 76
6.3.2.1 Uncertainty in the Farley Model 76
6.3.2.2 Uncertainty in FISHRAND Model Predictions 77
6.3.3 Sensitivity Analysis for Risk Models for Avian and Mammalian
Receptors 78
7.0 CONCLUSIONS 78
7.1 Assessment Endpoint: Benthic Community Structure as a Food Source for
Local Fish and Wildlife 78
7.2 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth,
and Reproduction) of Local Fish (Forage, Omnivorous, and Piscivorous)
Populations 78
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) 79
X1 TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
Page
7.4 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth
and Reproduction) of Lower Hudson River Waterfowl (as Represented by
the Mallard) 79
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) 79
7.6 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Insectivorous Mammals (as represented by the Little Brown Bat) 79
7.7 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Local Omnivorous Mammals (as represented by the Raccoon) 79
7.8 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Local Piscivorous Mammals (as represented by the Mink and River Otter) 79
7.9 Assessment Endpoint: Protection of Threatened and Endangered Species .... 80
7.10 Assessment Endpoint: Protection of Significant Habitats 80
7.11 Summary 80
REFERENCES 93
APPENDICES 80
APPENDIX A - Conversion from Tri+ PCB Loads to Dichloro through
Hexachloro Homologue Loads at the Federal Dam 80
APPENDIX B - Effects Assessment 91
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
III. RISK ASSESSMENT REVISIONS
1. Summary
2. Introduction
2.1 Changes in the Modeled Concentrations of PCBs in Fish, Water and Sediment
2.1.1 Changes to the Farley Models between December 1999 and August 2000
2.1.2 Changes to FISHRAND between December 1999 and August 2000
2.2 Changes in Toxicity Reference Values
2.2.1 Changes in Fish TRVs
2.2.2 Changes in Avian TRVs
2.2.3 Changes in Mammalian TRVs
3. Results
3.1 Comparison/Discussion
IV. COMMENTS ON THE ERA ADDENDUM
Federal (EF-1)
State (ES-1)
Local (EL-1)
General Electric (EG-1)
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
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AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
LIST OF TABLES:
SECTION I
1 Distribution of ERA
2 Information Repositories
SECTION II
EL-1.8 Cumulative Loads Over the Troy Dam (kg)
EG-1.14 Comparison of Mean Striped Bass Body Burdens at Three Long-Term Monitoring
Locations (Data from NYSDEC)
SECTION III
3-5 Summary of Tri+ Whole Water Concentrations from the Farley Model and TEQ-Based
Predictions for 1993-2018 (Revised)
3-6 Summary of Tri+ Sediment Concentrations from the Farley Model and TEQ-Based
Predictions for 1993-2018 (Revised)
3-7 Organic Carbon Normalized Sediment Concentrations Based on USEPA Phase 2 Dataset
(Revised)
3-8 Summary of Tri+ Benthic Invertebrate Concentrations from the FISHRAND Model and
TEQ-Based Predictions for 1993-2018 (Revised)
3-9 Spottail Shiner Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
3-10 Pumpkinseed Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
3-11 Yellow Perch Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
3-12 White Perch Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
3-13 Brown Bullhead Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
3-14 Largemouth Bass Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
3-15 Striped Bass Predicted Tri+ Concentrations for 1993 - 2018 (Revised)
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
4-1 Toxicity Reference Values for Fish - Dietary Doses and Egg Concentrations of Total PCBs
and Dioxin Toxic Equivalents (TEQs) (Revised)
4-2 Toxicity Reference Values for Birds - Dietary Doses and Egg Concentrations of Total PCBs
and Dioxin Toxic Equivalents (TEQs) (Revised)
4-3 Toxicity Reference Values for Mammals - Dietary Doses of Total PCBs and Dioxin Toxic
Equivalents (TEQs) (Revised)
5-1 Ratio of Predicted Sediment Concentrations to Sediment Guidelines (Revised)
5-2 Ratio of Predicted Whole Water Concentrations to Criteria and Benchmarks (Revised)
5-3 Ratio of Predicted Pumpkinseed Concentrations to Field-Based NOAEL for Tri+ PCBs
(Revised)
5-4 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived NOAEL for Tri+
PCBs (Revised)
5-5 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived LOAEL for Tri+
PCBs (Revised)
5-6 Ratio of Predicted Pumpkinseed Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis (Revised)
5-7 Ratio of Predicted Pumpkinseed Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis (Revised)
5-8 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis (Revised)
5-9 Ratio of Predicted Spottail Shiner Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis (Revised)
5-10 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived NOAEL For Tri+
PCBs (Revised)
5-11 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived LOAEL For Tri+
PCBs (Revised)
5-12 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived NOAEL on a
TEQ Basis (Revised)
xv TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5-13 Ratio of Predicted Brown Bullhead Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis (Revised)
5-14 Ratio of Predicted White Perch Concentrations to Field-Based NOAEL for Tri+ PCBs
(Revised)
5-15 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived NOAEL for Tri+
PCBs (Revised)
5-16 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived LOAEL for Tri+
PCBs (Revised)
5-17 Ratio of Predicted White Perch Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis (Revised)
5-18 Ratio of Predicted White Perch Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis (Revised)
5-19 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived NOAEL on a TEQ
Basis (Revised)
5-20 Ratio of Predicted Yellow Perch Concentrations to Laboratory-Derived LOAEL on a TEQ
Basis (Revised)
5-21 Ratio of Predicted Largemouth Bass Concentrations to Field-Based NOAEL For Tri+ PCBs
(Revised)
5-22 Ratio of Predicted Largemouth Bass Concentrations to Laboratory-Derived NOAEL on a
TEQ Basis (Revised)
5-23 Ratio of Predicted Largemouth Bass Concentrations to Laboratory-Derived LOAEL on a
TEQ Basis (Revised)
5-24 Ratio of Predicted Striped Bass Concentrations to Tri+ and TEQ PCB-Based TRVs
(Revised)
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 (Revised)
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 (Revised)
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
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 (Revised)
5-28 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Tree Swallow
Using TEQ for the Period 1993 - 2018 (Revised)
5-29 Ratio of Modeled Dietary Dose for Female Mallard Based on FISHRAND Results for the
Tri+ Congeners (Revised)
5-30 Ratio of Egg Concentrations for Female Mallard Based on FISHRAND Results for the Tri+
Congeners (Revised)
5-31 Ratio of Modeled Dietary Dose to Benchmarks for Female Mallard for Period 1993 - 2018
on a TEQ Basis (Revised)
5-32 Ratio of Modeled Egg Concentrations to Benchmarks for Female Mallard for Period 1993
- 2018 on a TEQ Basis (Revised)
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 (Revised)
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 (Revised)
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 (Revised)
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 (Revised)
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 (Revised)
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 (Revised)
5-39 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Belted Kingfisher Using
TEQ for the Period 1993-2018 (Revised)
5-40 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Great Blue Heron Using
TEQ for the Period 1993 - 2018 (Revised)
XV11 TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
5-41 Ratio of Modeled Dietary Dose Based on FISHRAND for Female Bald Eagle Using TEQ
for the Period 1993-2018 (Revised)
5-42 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Belted Kingfisher
Using TEQ for the Period 1993-2018 (Revised)
5-43 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Great Blue Heron
Using TEQ for the Period 1993-2018 (Revised)
5-44 Ratio of Modeled Egg Concentrations Based on FISHRAND for Female Bald Eagle Using
TEQ for the Period 1993 - 2018 (Revised)
5-45 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Bat for Tri+ Congeners
for the Period 1993-2018 (Revised)
5-46 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Bat on a TEQ Basis for
the Period 1993-2018 (Revised)
5-47 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Raccoon for Tri+
Congeners for the Period 1993-2018 (Revised)
5-48 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Raccoon on a TEQ
Basis for the Period 1993 - 2018 (Revised)
5-49 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Mink for Tri+
Congeners for the Period 1993-2018 (Revised)
5-50 Ratio of Modeled Dietary Dose to Toxicity Benchmarks for Female Otter for Tri+ Congeners
for the Period 1993-2018 (Revised)
5-51 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Mink on a TEQ Basis
for the Period 1993-2018 (Revised)
5-52 Ratio of Modeled Dietary Doses to Toxicity Benchmarks for Female Otter on a TEQ Basis
for the Period 1993-2018 (Revised)
TAMS/MCA
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
BOOK 1 OF 1
TABLE OF CONTENTS
LIST OF FIGURES:
SECTION II
EL-1.8
EL-1.12
EL-14a
EL-14b
EL-14c
EL-1.23a
EL-1.23b
EL-1.26a
EL-1.26b
EL-1.26c
EG-1.12
Comparison of Cumulative PCB Loads at Waterford from Farley et
al.,1999andUSEPA,2000
Comparison Among the HUDTOX Upper River Load and Farley
Model Estimates Striped Bass Body Burdens in Food Web Region 2
(1987-2067)
Comparison Between FISHRAND Results and Measurements at RM
152 (Revised Figure 3-12a)
Comparison Between FISHRAND Results and Measurements at RM
113 (Revised Figure 3-12b)
Comparison Between FISHRAND Results and Measurements of
Pumpkinseed (Revised Figure 3-12c)
Relative Percent Difference for GE Water Column Sample Duplicates
at the TI Dam
Percent Similarity of GE Water Column Sample Duplicates for the
Tri through Hexa Homologues at the TI Dam
Total PCB Concentrations at the Thompson Island Dam (1991-2000)
Fort Edward Summer Average Flows
Tri+ Loads at the TI Dam Compared to Flow at Fort Edward
Relationship Between the TI Dam West and Central Channel
Stations for Homologue to Tri+ Ratios GE Data (1997-1999)
XIX
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
TABLE OF CONTENTS
BOOK 1 OF 1
SECTION III
III-l Comparison for Tri+ PCBs in the Dissolved Phase of the Water
Column Between the Revised Model Output versus the Data
Presented in the Lower River Ecological Risk Assessment
III-2 Comparison for Total PCBs in the Water Column (Whole Water)
Between the Revised Model Output versus the Data Presented in the
Lower River Ecological Risk Assessment
1II-3 Comparison for Total PCBs in the Sediment (0-2.5 cm) Between the
Revised Model Output versus the Data Presented in the Lower River
Ecological Risk Assessment
xx TAMS/MCA
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Acronyms
BSAF
CIP
DEIR
ERA
ERASOW
GE
HHRA
LOAEL
NOAA
NOAEL
NYSDEC
PCB
RBMR
RI/FS
RM
RI/FS
STC
SWEM
TCDD
TEQ
TI
TOC
TRY
TQ
UCL
USEPA
USAGE
USFWS
USGS
BIOTA-SEDIMENT ACCUMULATION FACTOR
COMMUNITY INTERACTION PROGRAM
DATA INTERPRETATION AND EVALUATION REPORT
ECOLOGICAL RISK ASSESSMENT
ECOLOGICAL RISK ASSESSMENT SCOPE OF WORK
GENERAL ELECTRIC
HUMAN HEALTH RISK ASSESSMENT
LOWEST-OBSERVED-ADVERSE-EFFECT-LEVEL
NATIONAL OCEANIC AND ATMOSPHERIC ADMINISTRATION
NO-OBSERVED-ADVERSE-EFFECT-LEVEL
NEW YORK STATE DEPARTMENT OF ENVIRONMENTAL CONSERVATION
POLYCHLORINATED BlPHENYL
REVISED BASELINE MODELING REPORT
REMEDIAL INVESTIGATION/FEASIBILITY STUDY
RIVER MILE
REMEDIAL INVESTIGATION/FEASIBILITY STUDY
SCIENCE AND TECHNICAL COMMITTEE
SYSTEM-WIDE EUTROPHICATION MODEL
2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN
TOXICITY EQUIVALENCY
THOMPSON ISLAND
TOTAL ORGANIC CARBON
TOXICITY REFERENCE VALUE
TOXICITY QUOTIENT
UPPER CONFIDENCE LIMIT
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
UNITED STATES ARMY CORPS OF ENGINEERS
UNITED STATES FISH AND WILDLIFE SERVICE
UNITED STATES GEOLOGICAL SURVEY
XXI
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Introduction
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HUDSON RIVER PCBs REASSESSMENT RI/FS
RESPONSIVENESS SUMMARY FOR
VOLUME 2E-A BASELINE ECOLOGICAL RISK ASSESSMENT
FOR FUTURE RISKS IN THE LOWER HUDSON RIVER
AUGUST 2000
I. INTRODUCTION AND COMMENT DIRECTORY
1. Introduction
The U.S. Environmental Protection Agency (USEPA) has prepared this Responsiveness
Summary to address comments received during the public comment period on the Phase 2
Ecological Risk Assessment for Future Risks in the Lower Hudson River (ERA Addendum) for the
Hudson River PCBs Reassessment Remedial Investigation/Feasibility Study (RI/FS), dated
December 1999.
For the Reassessment RI/FS, USEPA has established a Community Interaction Program
(CIP) to elicit feedback through regular meetings and discussion and to facilitate review of and
comment upon work plans and reports prepared during all phases of the Reassessment RI/FS.
The ERA Addendum is incorporated by reference and is not reproduced herein. The
comment responses and revisions noted herein are considered to amend the ERA Addendum. For
complete coverage, the ERA Addendum and this Responsiveness Summary must be used together.
The first part of this Responsiveness Summary is entitled, "Introduction and Comment
Directory." It describes the ERA Addendum review and commenting process, explains the
organization and format of comments and responses, and contains a comment directory.
The second part, entitled "Responses to Comments on the ERA for Future Risks in the Lower
Hudson River," contains USEPA's responses t o all significant comments received on the ERA
Addendum. Responses are grouped according to the section number of the ERA Addendum to
which they refer. For example, responses to comments on Section 2.2 of the ERA Addendum are
found in Section 2.2 of the Responsiveness Summary. Additional information about how to locate
responses to comments is contained in the Comment Directory.
The third part, entitled "Risk Assessment Revisions," presents the revised results for the ERA
Addendum, incorporating the modified forecast concentrations of PCBs in fish, sediments, and river
water from the Revised Baseline Modeling Report (USEPA, 2000a) and other revisions based on
comments received on the ERA Addendum. To facilitate comparison to the December 1999 ERA
Addendum, all table and figure numbers have retained their original designations.
The fourth part, entitled "Comments on the ERA Addendum," contains copies of the
comments on the ERA Addendum submitted to USEPA. Not all references provided by the
commenters are reproduced in this document. The comments are identified by .commenter and
comment number, as further explained in the Comment Directory.
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2. Commenting Process
This section documents and explains the commenting process and the organization of
comments and responses in this document. Readers interested in finding responses to their
comments may skip this section and go directly to the tab labeled "Comment Directory."
2.1 Distribution of ERA
The ERA Addendum, issued in December 1999, was distributed to federal and state agencies
and officials, participants in the CIP and General Electric Company (GE), as shown in Table 1.
Distribution was made to approximately 100 agencies, groups, and individuals. Copies of the ERA
Addendum were also made available for public review in 16 Information Repositories, as shown in
Table 2 and on the USEPA Region 2 Internet web page, entitled "Hudson River PCBs Superfund
Site Reassessment," at www.epa.gov/hudson.
2.2 Review Period and Public Availability Meetings
USEPA held a formal comment period on the ERA Addendum from December 29,1999 to
January 28,2000. USEPA held a Joint Liaison Group meeting on January 11,2000 in Poughkeepsie,
New York that was open to the public to present the ERA Addendum. Subsequently, USEPA
sponsored an availability session to answer questions on January 18,2000 in Poughkeepsie, New
York. These meetings were conducted in accordance with USEPA's "Community Relations in
Superfund: Handbook, Interim Version" (1998a). Minutes of the Joint Liaison Group meeting are
available for public review at the Information Repositories listed in Table 2.
As stated in USEPA's letter transmitting the ERA Addendum, all citizens were urged to
participate in the Reassessment process and to join one of the Liaison Groups formed as part of the
CIP.
2.3 Receipt of Comments
Comments on the ERA were received in two ways: letters submitted to USEPA and oral
statements made at the January 11, 2000 Joint Liaison Group meeting. USEPA's responses to oral
statements made at the Joint Liaison Group meetings are provided in the meeting minutes. Written
comments were received from four commenters; total comments number 100. All significant written
comments received on the ERA Addendum are addressed in this Responsiveness Summary.
2.4 Distribution of Responsiveness Summary
This Responsiveness Summary is being distributed to, among others, the Liaison Chairs and
Co-Chairs and interested public officials. This Responsiveness Summary is also being placed in the
16 Information Repositories and is part of the Administrative Record.
TAMS/MCA
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TABLE 1
DISTRIBUTION OF ERA ADDENDUM
HUDSON RIVER PCBs OVERSIGHT COMMITTEE MEMBERS
USEPA ERRD Deputy Division Director (Chair)
USEPA Project Managers
USEPA Community Relations Coordinator, Chair of the Steering Committee
NYSDEC Division of Hazardous Waste Management representative
NYSDEC Division of Construction Management representative
National Oceanic and Atmospheric Administration (NOAA) representative
Agency for Toxic Substances and Disease Registry (ATSDR) representative
US Army Corps of Engineers representative
New York State Thruway Authority (Department of Canals) representative
USDOI (US Fish and Wildlife Service) representative
NYSDOH representative
GE representative
Liaison Group Chairpersons
Scientific and Technical Committee representative
SCIENTIFIC AND TECHNICAL COMMITTEE MEMBERS
The members of the Science and Technical Committee (STC) are scientists and technical
researchers who provide technical input by evaluating the scientific data collected on the
Reassessment RI/FS, identifying additional sources of information and on-going research relevant
to the Reassessment RI/FS, and commenting on USEPA documents. Members of the STC are
familiar with the site, PCBs, modeling, toxicology, and other relevant disciplines.
Dr. Daniel Abramowicz
Dr. Donald Aulenbach
Dr. James Bonner, Texas A&M University
Dr. Richard Bopp, Rensselaer Polytechnic Institute
Dr. Brian Bush
Dr. Lenore Clesceri, Rensselaer Polytechnic Institute
Mr. Kenneth Darmer
Mr. John Davis, New York State Dept. of Law
Dr. Robert Dexter, EVS Consultants, Inc.
Dr. Kevin Farley, Manhattan College
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Dr. Jay Field, National Oceanic and Atmospheric Administration
Dr. Ken Pearsall, U.S. Geological Survey
Dr. John Herbich, Texas A&M University
Dr. Behrus Jahan-Parwar, SUNY - Albany
Dr. Nancy Kim, New York State Dept. of Health
Dr. William Nicholson, Mt. Sinai Medical Center
Dr. George Putman, SUNY - Albany
Dr. G-Yull Rhee, New York State Dept. of Health
Dr. Francis Reilly, Jr., The Reilly Group
Ms. Anne Secord, U.S. Fish and Wildlife Service
Dr. Ronald Sloan, New York State Dept. of Environmental Conservation
STEERING COMMITTEE MEMBERS
USEPA Community Relations Coordinator (Chair)
Governmental Liaison Group Chair and two Co-chairs
Citizen Liaison Group Chair and two Co-chairs
Agricultural Liaison Group Chair and two Co-chairs
Environmental Liaison Group Chair and two Co-chairs
USEPA Project Managers
NYSDEC Technical representative
NYSDEC Community Affairs representative
FEDERAL AND STATE REPRESENTATIVES
Copies of the ERA Addendum were sent to relevant federal and state representatives who have been
involved with this project. These include, in part, the following:
The Hon. Daniel P. Moynihan - The Hon. Michael McNulty
The Hon. Charles E. Schumer - The Hon. Sue Kelly
The Hon. John E. Sweeney - The Hon. Benjamin Oilman
The Hon. Nita Lowey - The Hon. Richard Brodsky
The Hon. Maurice Hinchey - The Hon. Bobby D'Andrea
The Hon. Ronald B. Stafford
16 INFORMATION REPOSITORIES
Copies of the ERA Addendum were placed in 16 Information Repositories (see Table 2).
TAMS/MCA
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TABLE 2
INFORMATION REPOSITORIES
Adriance Memorial Library
93 Market Street
Poughkeepsie, NY 12601
Catskill Public Library
1 Franklin Street
Catskill, NY 12414
AComell Cooperative Extension
Sea Grant Office
74 John Street
Kingston, NY 12401
Crandall Library
City Park
Glens Falls, NY 12801
County Clerk's Office
Washington County Office Building
Upper Broadway
Fort Edward, NY 12828
*AMarist College Library
Marist College
290 North Road
Poughkeepsie, NY 12601
*New York State Library
CEC Empire State Plaza
Albany, NY 12230
New York State Department
of Environmental Conservation
Division of Hazardous Waste Remediation
50 Wolf Road, Room 212
Albany, NY 12233
*A R. G. Folsom Library
Rensselaer Polytechnic Institute
Troy, NY 12180-3590
Saratoga County EMC
50 West High Street
Ballston Spa, NY 12020
* Saratoga Springs Public Library
49 Henry Street
Saratoga Springs, NY 12866
*ASUNY at Albany Library
1400 Washington Avenue
Albany, NY 1222
*ASojourner Truth Library
SUNY at New Paltz
NewPaltz,NY 12561
Troy Public Library
100 Second Street
Troy, NY 12180
U. S. Environmental Protection Agency
290 Broadway
New York, NY 10007
White Plains Public Library
100 Martine Avenue
White Plains, NY 12601
* Repositories with Database Report
CD-ROM (as of 10/98)
A Repositories without Project Documents
Binder (as of 10/98)
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3. Organization of ERA Addendum Comments and Responses to Comments
3.1 Identification of Comments
Each submission commenting on the ERA Addendum was assigned a letter "E" and one of
the following letter codes:
F - Federal agencies and officials;
S - State agencies and officials;
L - Local agencies and officials; and
G - GE.
The letter codes were assigned for the convenience of readers and to assist in the organization
of this document. Priority or special treatment was neither intended nor given in the responses to
comments.
Once a letter code was assigned, each submission was then assigned a number, in the order
that it was received and processed such as EF-1, EF-2 and so on. Each different comment within a
submission was assigned a separate sub-number. Thus, if a federal agency submission contained
three different comments, they are designated as EF-1.1, EF-1.2 and EF-1.3. Comment letters are
reprinted in the fourth section of this document.
The alphanumeric code associated with each reprinted written submission is marked at the
top right corner of the first page of the comment letter. The sub-numbers designating individual
comments are marked in the margin. Comment submissions are reprinted in numerical order by letter
code in the following order: EF, ES, EL, and EG.
3.2 Location of Responses to Comments
The Comment Directory, following this text, contains a complete listing of all commenters
and comments. This directory allows readers to find responses to comments and provides several
items of information.
The first column lists the names of commenters. Comments are grouped first by: EF
(Federal), ES (State), EL (Local) or EG (GE).
The second column identifies the alphanumeric comment code, e.g., EF-1.1, assigned
to each comment.
The third column identifies the location of the response by the ERA Addendum
Section number. For example, comments raised in Section 2.1 of the ERA
Addendum can be found in the corresponding Section 2.1 of the Responses,
following the third tab of this document.
The fourth, fifth and sixth columns list key words that describe the subject matter of
each comment. Readers will find these key works helpful as a means to identify
subjects of interest and related comments.
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Responses are grouped and consolidated by section number in order of the ERA Addendum
so that all responses to related comments appear together for the convenience of the reader interested
in responses to related or similar comments.
4. Comment Directory
4.1 Guide to Comment Directory
This section contains a diagram illustrating how to find responses to comments. The
Comment Directory follows. As stated in the Introduction, this document does not reproduce the
ERA Addendum. Readers are urged to utilize this Responsiveness Summary in conjunction with
the ERA Addendum.
4.2 Comment Directory
STEP1
Find the commenter or the key words
of interest in the Comment Directory.
STEP 2
Obtain the alphanumeric
comment codes and the
corresponding ERA Addendum
Section.
STEP 3
Find the responses following the
Responses tab. See the Table of
Contents to locate the page of
the Responsiveness Summary
for the ERA Addendum Section.
Key to Comment Codes:
Comment codes are in this format EX-a.b
X=Commenter Group (F=Federal, S=State, L=Local, G=General Electric)
a=Numbered letter containing comments
b=Numbered comment
Example:
COMMENT RESPONSE ASSIGNMENT FOR THE ERA
AGENCY/
Name
COMMENT
CODE
REPORT
SECTION
KEY WORDS
1
2
3
NOAA/Rosman EF-1.1
General Fate/Transport Bioaccumulation
BMR
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THIS PAGE LEFT BLANK INTENTIONALLY
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Comment Directory
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4.2 Comment Directory - Lower Hudson River Future Risks
AGENCY/ NAME
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
COMMENT
CODE
EF-1.1
EF-12
EF-1.3
EF-1.4
EF-I.5
EF-1.6
EF-17
EF-18
EF-1.9
EF-1.10
EF-1 1 1
EF-1.12
EF-I 13
EF-1. 14
EF-1.15
EF-1. 16
EF-1 17
EF-1. 18
EF-1. 19
EF-1.20
EF-1 21
EF-1 22
EF-1 23
EF-1. 24
EF-1. 25
EF-1 26
EF-1 27
EF-1 28
EF-1 29
REPORT
SECTION
General
Appendix A
General
General
General
General
General
General
Exec Sum
3.1 1 1
3.1 1 1
3.0
31 1.2
3.1 1.4
3116
32.2
3.2.2
3.2.3
323
332
4.1.2
41.2
51 1.1
5.1.1.1
5111
5121
521
5212
5213
1
Fate and Transport
Farley Model
Water Column and
Sediment Data
Food Chain
Modebng
Water and Sediment
Values
TRV's
Field Duplicate
Data
TRVs
Bald Eagles
Upstream Boundry
Farley Model
Striped Bass
Farley Model
Figure 3-7
Modeled Fish
Concentrations
TOC
Table 3-7
Predicted Benthic
PCB Concentrations
Striped Bass
NOAA 1999
TRV Laboratory and
Field Studies
TEFs and TEQs
Table Reference
Forecasted Sediment
Concentrations
Table 5-1
NYSDEC Surface
Water Criterion
Forage Fish
PCB Partitioning
TRV
KEYWORDS
2
Bioaccumulaaon
HUDTOX
Nearshore Areas
Fish Concentrations
Model Output
Study Selection
NOAELs and
LOAELs
Underestimate Risk
Future Risks
BMR High Flow
Conditions
FISHRAND
Body Burdens
Overestimates Water
Column Loss
Two Locations
Outside Boundaries
NYSDEC 1998 Data
Sediment Guidelines
EPA 1993
Guidelines Based on
TCDD
Empirical Data
Food Web Region 1
Fish PCB
Concentrations
Hansen et al, and
Bengtsson
Data Quality Issues
Underesomate
Sediment PCBs
NYSDEC Benthic
Criterion
Wrong SEL
Repraducnve Effects
Eggs vs Fish Tissue
Whole Body
Concentration
Studies
3
Baseline Modeling
Repon Revisions
Model Uncertainty
Food Web Pathways
Model Prediction
TOC - %Lipid
Uncertainty Factors
Modeling Results
Breeding Success
Underestimate Risk
Output Parameters
Normalization
Effects on Fish
Uptake of PCBs
PCB TEQs
Reasonable Estimate
Stnped Bass to
Largemouih Bass
Ratio
Reproduction and
Development
Other Effects than
Reproduction and
Development
PCBs BZ# 8 land
126
Farley Model TOC
TRV
Lipid Normalized vs
Wet Weight
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4.2 Comment Directory - Lower Hudson River Future Risks
AGENCY/ NAME
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NOAA/Rosman
NYSDEOPorts
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodison
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodeson
COMMENT
CODE
EF-1 30
EF-1 31
EF-1 32
EF-1 33
EF-1 34
EF-1 35
EF-1 36
EF-1 37
EF-1 38
EF-1 39
EF-1 40
EF-1 41
EF-1 42
EF-1 43
EF-1 44
EF-1 45
EF-1 46
EF-1 47
EF-1 48
ES-1 1
EL-1 1
EL-1 2a-h
EL-1 3
EL-1 4
EL-1 5
EL-1 6
EL-1 7a-e
EL-1 8
REPORT
SECTION
5213
5215
54
62
6321
6321
6322
71
73
Appendix A
Appendix B
Appendix B
Appendix B
Appendix B
Appendix B
Appendix B
Appendix B
Appendix B
Appendix B
5121
General
General
General
General
Appendix A
3111
3113
3113
1
Measurement
Endpomts
Field NOAEL
Current Trend in
Bird Usage
Dioxm Like Effects
Based on Congeners
Model • General
Trends
Uncertainty
Sensitivity Analysis
Uncertainty
Sensitivity of Tree
Swallows
Text Edit
See Comment
Section 5212
TRVs
Bengtsson et al
NOAEL and
LOAEL Transposed
Tables B-5 and B-6
El-No Effect and EL
Effeci
Table B-7
Figure B-2
Figure B-3
NYSDEC W[|dlife
Criterion
FISHRAND
Farley Model
PCB Source
Exposure
Assessment
Tree Swallows
Farley Model
Stnped Bass
Table 3-3
KEYWORDS
2
TEQ Concentrations
White Perch -
Yellow Perch
Historical Bird
Usage
Other Effects of
Congeners
Year to Year
Changes in Fish
Concentrations
Upstream Boundary
BMRand
FISHRAND
Sediment and Water
Forcasts
Compare to other
insectivorous Birds
Hexachloro
Homologue Ratio
Hansenetal, 1971
vs 1974
TRVs
Text Edit
Include USAGE
1988
NOAELs and
LOAELs
Lipid Values
Reported by Study or
Esimaied
Referenced NOAEL
Toxicity Endpomts
for Egg Dioxm
Equivalencies
Revised Value
Farley Model
EPA Review
Upper Hudson Load
Toxicitv Assessment
15 Year Increments
HHRA
HUDTOX
3
EPA 1993
Application of
Uncertainty Factor
GE'sUseofPCBs
Underestimates Risk
Underestimate Risk
Releases From
Remnant Deposits
Other Factors
Underestimate Risk
Other Effects than
Reproduction and
Development
Estimation
Procedure
Figure Focus -
Laboratory Studies
Figure Focus -
Laboratory Studies
Risks Based on
Various Loads
'revious Comments
on August Upper
River
Future Predictions
10
TAMS/MCA
-------
4.2 Comment Directory - Lower Hudson River Future Risks
AGENCY/ NAME
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
SCEMC/Hodgson
GE
GE
GE
GE
GE
GE
GE
GE
GE
GE
GE
COMMENT
CODE
EL-19
EL-1 10
EL-1 11
EL-1 12
EL-1 13
EL-1 14
EL-1 15
EL-1 16
EL-1 17
EL-1 IS
EL-1 19
EL-1 20
EL-1 21
EL-1 22
EL-1 23
EL-1 24
EL-1 25
EL-1 26
EL-1 27
EL-1 28
EG-1 1
EG-12
EG-1 3
EG-1 4
EG-1 5
EG-1 6
EG-1 7
EG-1 8
EG-1 9
EG-1 10
EG-1 11
REPORT
SECTION
3114
3115
3116
3122
3123
32
324
33
40
50
5219
5241
543
5731
Appendix A
Appendix A
Appendix A
Appendix A
Appendix A
Appendix A
General
General
General
20
25
30
40
40
40
40
30
1
Stuped Bass
Figure 3- 10
Figure 3- 12
Figures 3- 16 and 3
17
Figures 3- 16 to 3-
19
Selection or River
Mile Range
Brown Bullhead
See Comments on
August ERA
See Comments on
August ERA
See Comments on
August ERA
Measurement
Endpomts
Body Burden
Trends in Christmas
Bird Counts
Raccoon
Duplicate Sample
Figures A- 1 to A-5
TID Values
Citation Tor Data
Figures A-l to A-5
Bakers Falls
Releases
PCB Levels in the
Hudson River
Histoncal Databases
Future Ecological
Risk
Population vs
Individual Risks
Site Specific Data
Conservative
Assumptions
NOAA SECs
TEQ Approach
NOAA Review of
PCB Effects on Fish
TQ Approach
Model Deficiencies
KEYWORDS
2
Model vs Data
Farley vs
FISHRAND
Stnped Bass
Explanation
Internal consistency
Fish Body Burdens
Shortnosed Sturgeon
Bengtsson
Stnped Bass
Shortnosed Sturgeon
Health of Bird
Populations
Potential Risk
GE Samples
Geocherrucal Trends
GE Data
Decline in PCB
Loads
Validity of Factors
in Table A-2
Post 1990
Environmental
Health of River
Ecological Risk
Assessment
Remedial Action
GE's Previous
Comments
Biological Surveys
TQ Approach
Inherent Limitations
Screening Approach
Effects Exceeding 5
ppm
Not Scientifically
Defensible
HUDTOX
3
Food Web Region 1
Data Comparisons
Figure 3- 17 Region 2
only
Use of Farley or
FISHRAND
Selected Mile vs
Area Average
Uncertainty
Chlophen
River Miles 152 and
113
Uncertainty
EPA Samples
Factor I and Factor 2
EPA Data
Factors Constant for
40 vears
Post 1990 Releases
not of Concern
Ecological Risks
Fish and Wildlife
Populations
Toxicity Tests
Clich Rjver
Similarities
Primary Measure of
Risks
NOAA 1999
Comments on TEQ
Approach
Other Similar
Assessments
FISHRAND
II
TAMS/MCA
-------
4.2 Comment Directory - Lower Hudson River Future Risks
AGENCY/ NAME
GE
GE
GE
GE
GE
GE
GE
GE
GE
GE
GE
GE
GE
COMMENT
CODE
EG-1.12a
EG-I 12b
EG-1.13
EG-1.14
EG-1 15a-c
EG-1.16a&b
EG-1 17
EG-1 18a&b
EG-1. 19
EG- 1.20
EG-1 21
EG-1 22
EG-1. 23
REPORT
SECTION
30
Appendix A
30
3.0
30
3.0
50
4.0
40
40
4.0
40
70
1
HUDTOX Use in
Lower River
Farley Model
Farley Model
Lower River Fish
Concentrations
Farley Model
Fish modeled not
Representative of
Wildlife
Consumption
Condition of
Ecological Resources
EPAs Approach
Conservative
Fish TRVs
TRVs
TRVs
Limitations of TRVs
and TQ Approach
Reliance on Models
KEYWORDS
2
PCB Homologs
Model Uncertainty
Prediction of Lower
River Water and
Sediment
FISHRAND vs
Farley
Temporal Trends
Growth Rates not
Site-Specific
Benthic Community
Relies on Small
Subset of Data
Values Developed by
Monosson
Birds
Mammals
Selecave Treatment
of Literature
Ignore site-specific
Data
3
Application to
historical period and
Future
TI Dam Station
Cnucle Review of
Model
Food Web Structure
Striped Bass
Fish and Wildlife
Populations
Impropper
Interpretation of
Data
Uncertainty Factor
Uncertainty Factor
Scientific
Foundation of Lower
River BERA
12
TAMS/MCA
-------
Responses
-------
II. RESPONSIVENESS SUMMARY FOR THE ERA ADDENDUM FOR FUTURE
RISKS IN THE LOWER HUDSON RIVER
General Comments
Response to EF-1.1
Exposure modeling for the ERA Addendum combines the output of HUDTOX with the
Farley model and FISHRAND. Uncertainties and potential limitations in HUDTOX and
FISHRAND forecasts are discussed in the Revised Baseline Modeling Report (RBMR) (USEPA,
2000a). Discussions of potential limitations of the Farley model approach are provided in Farley et
al. (1999). Limitations inherent in USEPA's use of the Farley model approach, and general
modeling uncertainties associated with the Lower Hudson ERA, are discussed in the ERA
Addendum (see, USEPA, 1999c, pp. 70-73).
The ERA Addendum was completed in December 1999, prior to the revisions in the
HUDTOX modeling of the Upper Hudson, which are presented in the RBMR (USEPA, 2000a). The
revisions result in some relatively small changes in the forecasts of Tri+ PCB load across Federal
Dam. The Farley model for the Lower River was subsequently re-run using the output from the
revised HUDTOX model. The resulting forecasts for the Lower Hudson River are presented in
Section HI of this Responsiveness Summary. The revised risk results do not change the overall
conclusions of the ERA Addendum.
Response to EF-1.3
The USEPA agrees that the effect of daily water level changes on PCB exchange and non-
scour related movement of PCB-contaminated sediments may add important additional loads that
are not specified by any HUDTOX model mechanism. As a result, these mechanisms will not be
represented directly in the estimation of PCB loads to the Lower Hudson. However, the HUDTOX
model is not strictly mechanistic and incorporates several empirical and semi-empirical components.
Specifically, the process or processes responsible for the non-scour PCB loads identified in the
RBMR (USEPA, 2000a) have been empirically represented in the model based on the observed data.
To the extent that daily water level changes and non-scour related movement are important, much
of their effect is captured by this empirical representation. Similarly, the mechanistic components
of transport represented in the HUDTOX model rely on long-term records for calibration. To the
extent that either process suggested by the commenter is important to the mechanistic components
of transport, its effect will be reflected in the adjustments to the model parameters incorporated in
the mechanistic expression. Given that both the mechanistic and empirical components of the
HUDTOX model are based on long-term records, it is unlikely that any major PCB load has not been
represented in the model calibration.
USEPA recognizes that although the model will effectively capture all major loads by the
empirical and semi-empirical components, the issue of the exact sources of these loads is less certain.
13 TAMS/MCA
-------
This uncertainty affects the reliability of the model forecasts to the extent that the release process
may change relative to the representation derived from the model. Near-shore/shoreline regions and
the possibility of non-scour-related sediment movement can be considered independently of the
model results when decision making occurs.
Response toEF-1.4
In the RMBR (USEPA, 2000a), USEPA did not use a generic fish growth rate for lake trout
in the FISHRAND model, but rather used species-specific growth rates (see Section in of this
Responsiveness Summary). Growth rate is not the most sensitive parameter in the FISHRAND
model, but was considered sensitive enough to focus on for calibration. The spottail shiner growth
rate was not a sensitive parameter, and there are virtually no data available for this species. Thus,
the spottail shiner growth rate is the "generic" growth rate used in the Gobas model. Note that the
calibrated growth rates for the other species modeled (e.g., pumpkinseed, brown bullhead,
largemouth bass) are very close to the "generic" growth rate used m the Gobas model for all species.
Response to EF-1.5
The fate and transport portion of the Farley et al. (1999) model represents the Lower Hudson
River on a finite segment basis, with each one-dimensional segment being 10 miles in length and
occupying the whole width of the river. Given this model segmentation and the use of seasonal
average flows, the Farley model is appropriate for the calculation of long-term, segment-wide
average environmental concentrations, and not appropriate for the fine spatial scale estimates of
concentrations within specific habitat types identified by the commenter. Given the scale of the
Farley model, it is appropriate to use averaged exposure concentrations within a segment to assess
the accumulation of PCBs in biota. The fit between model predictions and observation data could
be improved by use of a more detailed spatial and temporal representation of exposure
concentrations and individual species feeding patterns, but this is not possible within the context of
the Farley et al. (1999) model.
Sediment TOC was assumed to be constant throughout the Lower Hudson. TOC actually
varies; however, use of a constant average value is consistent with the segment-averaged nature of
the Farley model. Lipid content of fish was also set to a constant value by Farley et al. (1999). It
should be noted, however, that in the version of the Farley model used by USEPA, the lipid content
offish was modified from that given in Farley et al. (1999) to reflect the lipid content in fish sampled
by NYSDEC in the 1990s (Cooney, 1999).
These simplifying assumptions reflect the fact that the Farley model is designed for
prediction of long-term trends in average fish tissue concentrations. USEPA does not consider the
model to be appropriate for predicting short-term variability in fish concentrations or response to
transient events. Variability in these factors will affect predictions of long-term trends and
associated ecological risks only to the extent that the selected average values are biased.
14 TAMSMCA
-------
Lipid content in the FISHRAND model is described by a distribution. TOC was set to the
single value that is used by the Farley model in order to be consistent with the assumptions of that
model. TOC is a relatively sensitive parameter, and shows an inverse relationship with predicted
body burdens (i.e., increases in TOC lead to decreases in predicted fish body burden).
Response to EF-1.6
Based on the comments received on the ERA(USEPA, 1999c) and the ERA Addendum
(USEPA, 1999c), the studies that were used to derive TRVs were reexamined. This reexamination
is detailed in Section HI of this document.
Based on the reexamination, the laboratory-based TRVs were revised for all fish receptors
(i.e., pumpkinseed, spottail shiner, brown bullhead, yellow perch, white perch, largemouth bass,
striped bass, shortnose sturgeon). The sheepshead minnow study by Hansen et al. (1974) was
selected for development of the laboratory TRY, instead of Bengsston (1980). Hansen et al. (1974)
established a NOAEL for exposure to Aroclor 1254 of 1.9 mg PCBs/kg and a LOAEL of 9.3 mg
PCBs/kg for adult female fish. The values for adult fish determined in Hansen et al. (1974) are more
appropriate for comparison to measured and modeled concentrations in adult Hudson River fish than
the Bengsston (1980), which examined hatchability in minnows exposed to Clophen A50. Because
the sheepshead minnow is not in the same taxonomic family as any of the Hudson River fish
receptors, an interspecies uncertainty factor of 10 is applied to develop TRVs for all fish.
Therefore, on the basis of laboratory toxicity studies:
• The LOAEL TRV for the pumpkinseed, spottail shiner, brown bullhead, yellow perch,
white perch, largemouth bass, striped bass, and shortnose sturgeon is: 0.93 mg PCBs/kg
tissue.
• The NOAEL TRV for the pumpkinseed, spottail shiner, brown bullhead, yellow perch,
white perch, largemouth bass, striped bass, and shortnose sturgeon is: 0.19 mg/kg
PCBs/kg tissue.
The field-based TRVs of the pumpkinseed, spottail shiner, and largemouth bass were also
revised. For the pumpkinseed and largemouth bass, the field studies by Adams et al. (1989,1990,
1992) on the redbreast sunfish, a species in the same family as the pumpkinseed, were retained as
the studies to establish TRVs. However, the growth endpomt, rather than the higher fecundity
endpomt initially selected, was used to establish a TRV. The NOAEL for growth was reported as
being sigmficandy different from one downstream location, but no comparison to the reference sites
was provided. Growth is a relevant endpoint, and the NOAEL for growth, 0.3 mg/kg, was selected.
The sunfish (Lepomis auritus) in the studies were exposed to PCBs and mercury in the field.
However, because 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 TRV,
but not a LOAEL TRV, for the pumpkinseed and largemouth bass. An interspecies uncertainty
factor is not applied because these three species are all in the same family (Centrachidae). Because
15 TAMS/MCA
-------
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 and largemouth bass is: 0.3 mg PCBs/kg tissue.
The previous NOAEL TRY for the pumpkinseed and largemouth bass was 0.5 mg PCBs/kg tissue
based upon the fecundity endpoint in Adams et al. (1992).
In the ERA Addendum, no field-based TRY was selected for the spottail shiner. However,
upon re-examination, the study by USAGE (1988) using fathead minnow is considered to be a field-
related study, rather than a laboratory study, because the sediments to which the fathead minnow
were exposed were field-collected sediments (instead of spiked sediments). This study was selected
for development of a field-based TRY for the spottail shiner, a species in the same family as the
fathead minnow.
On the basis of the field study:
• The final NOAEL TRY for the spottail shiner is: 5.25 mg PCBs/kg wet wt tissue.
The field-based TRY was selected for use, rather than the laboratory-based TRVs used in the ERA
Addendum.
The associated future risks to fish in the Lower Hudson River using these revised TRVs are
presented in Section EL
Response to EF-1.7
The LOAEL and NOAEL values was used together to bracket risk.
As outlined in Section B.2.1 of the ERA Addendum (USEPA, 1999c), both laboratory and
field studies have advantages and disadvantages for the purpose of deriving TRVs. For example, a
controlled laboratory study can test the effect of a single formulation or congener on the test species
in the absence of the effects of other co-occurring contaminants or confounding field conditions.
Therefore, greater confidence can be placed in the conclusion that observed effects are related to
exposure to the test compound and both NOAEL- and LOAEL-based TRVs can be developed Field
studies have the advantage that organisms are exposed to a more realistic mixture of PCB congeners
(with different toxic potencies) than laboratory studies. Both types of studies may have the
disadvantage that they are conducted on species that are not closely related to the receptors of
concern at the Hudson River. Because each approach has both advantages and disadvantages, the
ERA Addendum developed TRVs and evaluated risk based on both laboratory and field studies.
16 TAMS/MCA
-------
If an appropriate field study is available for a species in the same taxonomic family as the
receptor of concern, that field study is used to derive a NOAEL TRV. However, in many cases,
appropriate field studies were not available for any species that were closely related (e.g., in the same
taxonomic family) as the receptor of concern. In such cases, the advantages of a controlled laboratory
study, in particular the ability to derive both LOAEL- and NOAEL-based TRVs, were felt to
outweigh those of field studies conducted on less closely related species, and a field-based TRV was
not developed.
Response to EF-1.8
The TRV selection process specifically focused on the toxicity endpoints of greatest
population relevance, i.e, survival, growth, and reproductive capacity. Although immune
suppression may have implications on a population level (e.g., reduced capacity to recover from
stress), the link to population level effects is far less direct than reproductive and growth effects, such
as survival and reproductive capacity.
The FISHRAND model does not consistently underestimate risks; typically predicted results
are within the error bars of the data. If there is an underestimate in concentration, it is generally
within a factor of two, which would not change the overall conclusions of the ERA Addendum.
Response to EL-1.1
The RBMR (USEPA, 2000a) was not available when the ERA Addendum was released.
USEPA provided the commenter a copy of Farley et al. (1999) report and added a copy to each of
the 16 information repositories. The final models are documented in the RBMR(USEPA, 2000a).
The only documentation for the Farley model is in the March 1999 report (Farley et al., 1999). An
update regarding lipid content was made to the Farley et al. (1999) model prior to its use in the ERA
Addendum; this update (Cooney, 1999) is noted in the ERA Addendum, the Responsiveness
Summary for the ERA, and this Responsiveness Summary for the ERA Addendum. The change to
the Farley et al. (1999) model used in the ERA Addendum is minor and does not affect the overall
conclusions of the ERA Addendum.
Response to EL-1.2
USEPA reviewed the Farley model to ensure that it was an appropriate tool for use in
predicting future risks in the Lower Hudson River. USEPA is not arranging a peer review of the
Farley model, although it is USEPA's understanding that the authors of the Farley et al. (1999) report
intend to submit their work in a paper to be published in a peer reviewed scientific journal.
Response to EL-1.2a
The data presented in Figure 1-1 of the Farley et al. (1999) report represent results from
individual cores. Two results are shown for RM 159 because two different cores are available.
17 . TAMS/MCA
-------
Throughout the freshwater portion of the Hudson there is a significant amount of variability in core
profiles from nearby locations, due to local heterogeneity in depositional patterns and different
histories of dechlorination. The first core shown in Figure 1-1 appears to reflect a more
contaminated location in which a significant amount of dechlorination has occurred. The second
core, with lower total PCB concentration in the surface layer, does not have a strong dechlorination
signature.
Response to EL-1.2b
USEPA agrees that the Farley model contains a number of simplifying assumptions, such as
the use of a constant seasonal pattern of flows based on a typical flow year. These simplifying
assumptions are appropriate for the purpose for which the model was built, which was to examine
the effects of PCB loading rates on the long-term trajectory of PCB concentrations in fish. Year-to-
year variability in flow and sedimentation will cause short-term fluctuations in PCB concentrations,
but will have a lesser effect on long term trends. This type of approach is particularly appropriate
for the Lower Hudson, which is far removed from the major upstream source of PCBs, resulting in
a smoothing out of temporal variations in loads. Further, mixing and sedimentation in much of the
Lower Hudson are strongly affected by tidal influences, which are relatively constant from year to
year. Regarding model-assigned sediment thicknesses, it should be noted that these are designed to
be representative of the vertical profile that potentially interacts with the water column, and not
actual sediment thickness. The assumptions are considered to be appropriate, and changing these
values by small amounts would have little effect on model results. It is incorrect to say that
sedimentation rates and sediment loads are assigned by Farley "with little or no justification." The
sources of data are cited on pp. 26-27 of Farley et al. (1999), along with a discussion of the rationale
for extrapolating or interpolating results. The resulting parameters are reasonable based on the best
available data; however, it is true that detailed data were not available to fully constrain specific
components, such as annual sediment loads from individual tributaries. It is also incorrect to imply
there is "little or no justification" for the model representation of organic carbon, as this was taken
from the detailed System-Wide Eutrophication Model (SWEM) effort completed by HydroQual.
Finally, USEPA agrees that the PCB loads from the New Jersey tributaries are subject to high degree
of uncertainty. These loads, however, appear to be of minor significance relative to loads from other
sources, particularly for Food Web Regions 1 and 2 of the Farley model, which are upstream of the
New Jersey tributaries.
Response to EL-1.2c
It is misleading to say that these components were specified "rather than" modeled. Page 18
of Farley et al. (1999) report states that these components were not directly simulated within the
Farley model itself. Instead, they were "specified based on field observations, other modeling work,
or simple mass balance calculations...". In fact, hydrodynamics and organic carbon were represented
in the Farley model based on the detailed SWEM modeling effort. The sediment transport
component combines SWEM hydrodynamics with a solids mass balance
18 TAMS/MCA
-------
Response to EL-1.2d and EL-1.2e
It is a common feature of essentially all modeling efforts that more data are desired than are
available. The Farley model lacks field data for comparison at certain locations, for certain
conditions, and times. However, available data were used to evaluate calibration of the model. The
lack of additional data for calibration increases uncertainty in model results, but does not invalidate
the approach.
Response to EL-1.2f
The Farley model is based on the earlier modeling effort of Thomann et al. (1989). For the
recalibration of the bioaccumulation model, Farley et al. (1999) adjusted only three parameters in
the model: the volatilization rate coefficient, the gill transfer efficiency, and the phytoplankton
uptake rate/growth rate ratio. (Several other parameters were adjusted relative to the original
Thomann Model, c.f. Table 2-2, but based directly on new data rather than calibration.) USEPA
disagrees with the comment that this constitutes a "large number of parameter adjustments;" rather
it represents a parsimonious set of parameters.
Response to EL-1.2g
Results presented by Farley et al. (1999) show some discrepancies in homologue distribution
between model results and high-resolution sediment core data collected by USEPA in 1992 (Figure
3-5), although total surface sediment concentrations are well replicated (Figure 3-4). In part, the
small error in replicating concentrations in individual cores is due to the expected spatial variability
among sediment samples. Despite this spatial variability, the Farley et al. (1999) results appear to
underestimate the dichlorobiphenyl and trichlorobiphenyl sediment concentrations on a fairly
consistent basis (e.g., Figure 3-5). These discrepancies are believed to be largely due to Farley's
assumptions for upstream loads over Federal Dam, which were replaced with HUDTOX generated
loads by USEPA (USEPA, 2000a). As shown in Table 3-3 and Figure 3-2 of the ERA Addendum
(USEPA, 1999c), the HUDTOX model output produces higher loads of dichlorobiphenyl and
trichlorobiphenyl than the estimates used by Farley et al. (1999) for the period prior to 1992.
Response to EL-1.2h
Figure 3-14 in Farley et al. (1999) shows a fairly good fit between model predictions and
observations in white perch, with most model predictions lying within confidence bounds on sample
data. For Food Web Region 1, it appears that the model overpredicts more than underpredicts
NYSDEC white perch data, while closely replicating the USEPA/NOAA white perch data. This is
in large part due to the fact that Farley et al. (1999) did not account for the effects of different
analytical methods between the USEPA/NOAA and NYSDEC results. In any case, the results
reported by Farley et al. have been superseded by results using the revised HUDTOX load estimates
across Federal Dam. Farley model predictions using the HUDTOX-generated load are compared to
observed body burdens in white perch in Figure 3-8 of the ERA Addendum.
19 TAMS/MCA
-------
Response to EL-1.3
The USEPA fate, transport and bioaccumulation models are designed to capture the general
trends in the concentrations of PCBs in river water, sediment, and fish. This is demonstrated by
comparing the model output to calibration data. There is good agreement between the model hindcast
and observed data. These models do not examine cases such as the nearshore environment or PCB
load variations caused by underestimating resuspension. This is an added uncertainty to the risk
assessments, not a modeling uncertainty.
Additionally, the USEPA has already recognized the importance of Lower Hudson River
contributions of PCBs (USEPA, 1997). However, as also noted by Farley et al. (1999), lower river
contributions are restricted to the region below the salt front. Thus, USEPA's statements that the
Upper Hudson River is the only significant source of PCBs to the freshwater Lower Hudson are
correct.
Response to EL-1.4
USEPA agrees that the framework and methodology used for the Upper Hudson and Lower
Hudson exposure and toxicity assessments are consistent. USEPA's response to comments on the
ERA regarding items such as exposure and the toxicity assessment are addressed in the
Responsiveness Summary for the ERA (USEPA, 2000b).
Response to EG-1.1
Although historic data for some species exist from the last 25 years, changes in populations
and communities can not viewed from a strictly PCB-oriented perspective. The fishing ban and
overall improvement in water quality have contributed to population increases in some fish species;
however, this does not indicate that there are no adverse effects from PCBs.
The Hudson River is a large and complex ecosystem influenced by a variety of factors. Some
clear correlations can be seen in the Hudson River ecosystem, such an increase in some fish
populations due to the fishing ban or an increase in pollution-intolerant filter feeding
macromvertebrates resulting from improved water quality. More subtle effects, including those of
PCBs, are difficult to discern amid the natural variability of the ecosystem. The kinds of effects
expected from PCBs include reduced fecundity, decreased hatching success, and similar kinds of
reproductive impairment indicators, which are often difficult to detect. The gradient of PCB
concentrations along the roughly 200 miles of river being examined in the Reassessment RJ/FS also
increases the difficulty of ascribing particular effects to PCBs. Therefore, the ERA Addendum
discusses the potential for adverse effects even in apparently healthy receptor populations.
Part of the difficulty of assessing receptor populations is that there are no data against which
to measure abundance. Limited breeding success in bald eagle nests along the Hudson does not
establish that the bald eagle is re-established there. Since these eagles are the first to breed in
20 TAMS/MCA
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approximately one hundred years, there is no appropriate reference population for comparison. In
addition, all eagles now breeding in NYS are the result of NYSDEC or other direct
release/restoration programs (Nye, 2000).
As noted in the Responsiveness Summary for the ERA (see, USEPA, 2000b response to EL-
1.1), population-specific information prior to PCBs, the fishing ban, and other anthropogenic
influences would be valuable in assessing the effects of PCBs on today's population. However, the
time frame necessary for these data is far longer than the nine years since the initiation of the ERA.
Response to EG-1.2
This comment on USEPA's ecological risk assessment approach has been addressed in the
Responsiveness Summary for the ERA (USEPA, 2000b). Responses to general points provided in
the Responsiveness Summary for the ERA also apply to the ERA Addendum (see responses to the
following comments: EG-1.38 (general response), EG-1.2 and EG-1.4 (inadequate consideration of
population versus individual-level effects), EG-1.12 (improper use of weight of evidence approach),
EG-1.5, EG-1.8, EG-1.9, and EG-1.11, EG-1.-31, EG-1.32, and EG-1.33 (ignoring or dismissing
site-specific data), EG-1.6 (use of sediment guidelines and criteria as a measurement endpoint), EG-
1.1, EG-1.13, EG-1.15 and EG-1.18 (conservative exposure and effects assumptions), EG-1.27 and
EG-1.29 (TEQ Approach), EG-1.19 (NOAA's expert review of PCB effects on fish)).
Response to EG-1.3
Consistent with USEPA guidance, the purpose of the ERA Addendum is to assess ecological
risk posed by site-related contaminants, and does not include decision-making or risk management.
USEPA will evaluate risk reduction in the Feasibility Study and Proposed Plan for the Reassessment.
EXECUTIVE SUMMARY
Response to EF-1.9
For clarity, p. ES-9 is corrected to read "limited breeding success" (of bald eagles) rather than
"lack of breeding success" (fourth paragraph, last sentence). The first and second sentences of the
fifth paragraph on p. ES-11 is corrected to read, "Collectively the evidence indicates that future PCB
exposures (predicted from 1993 to 2018) may impair reproduction or recruitment of threatened or
endangered species. Using the TEQ-based..."
1.0 INTRODUCTION
No significant comments were received on Section 1.0.
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1.1 Purpose of Report
No significant comments were received on Section 1.1.
1.2 Report Organization
No significant comments were received on Section 1.2.
2.0 PROBLEM FORMULATION
Response to EG-1.4
USEPA's bottom-up approach uses data on individuals in order to predict potential effects
on local populations and communities that occur or could occur at the site. USEPA's Risk
Management Guidance (OSWER Directive 9285.7-28P, p. 3) states,
"Levels that are expected to protect local populations and communities can be estimated by
extrapolating from effects on individuals and groups of individuals using a lines-of-evidence
approach. The performance of multi-year field studies at Superfund sites to try to quantify
or predict long-term changes in local populations is not necessary for appropriate risk
management decisions to be made."
USEPA used, among other things, observed concentrations of PCBs in benthic invertebrates and fish
in the Hudson River and field studies of birds and mammals in and along the Hudson River, in a
weight-of-evidence approach to characterize risks to ecological receptors (see, ERA Addendum,
Section 5.0: Risk Characterization. Also see response to EG-1.1 in USEPA, 2000b).
The life span of each receptor examined in the ERA Addendum is less than 25 years (with
the exception of the shortnose sturgeon), indicating that the 25-year modeling duration covers the
life span of most receptors. Life span information is presented below.
Fish: largemouth bass - up to 15 years (Smith, 1985); pumpkinseed sunfish - 8 to 10 years
(in Canadian populations) (Scott and Grossman, 1975); brown bullhead - 6 to 7 years (Smith,
1985); yellow perch - 9 years (Smith, 1985); white perch - 5 to 7 years; some live 14 to 17
years (Smith, 1985); spottail shiner - 4 years (Pflieger, 1997); striped bass - Smith (1988)
reports the oldest fish studied at 14 to 18 years and Cooper (1983) reports a single female
estimated to be 30 years; shortnose sturgeon - 25 to 30 years, sometimes more (Dovel, 1981).
Birds (maximum longevity): tree swallow -10 years; mallard - 26 years; belted kingfisher -
16 years (note: no species-specific information was available so the oldest nonpasserine land
bird was used [red-cockaded woodpecker]); great blue heron - 23 years; and bald eagle - 22
years (Klimkiewicz, 1997).
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Mammals: little brown bat - 6 to 7 years; raccoon less than 5 years; mink - up to 10 years;
and river otter - up to 23 years (Walker, 1997).
The toxicity quotient (TQ) exceeds one (on a Tri+ and/or TEQ basis) in the Lower Hudson
River for the life span of the pumpkinseed, brown bullhead, yellow perch, white perch, largemouth
bass, striped bass, mallard, belted kingfisher, great blue heron, bald eagle, little brown bat, raccoon,
mink, and otter. These exceedances indicate that population level effects are possible in these
species.
2.1 Site Characterization
No significant comments were received on Section 2.1.
2.2 Contaminants of Concern
No significant comments were received on Section 2.2.
2.3 Conceptual Model
No significant comments were received on Section 2.3.
2.3.1 Exposure Pathways in the Lower Hudson River Ecosystem
No significant comments were received on Section 2.3.1.
2.3.2 Ecosystems of the Lower Hudson River
No significant comments were received on Section 2.3.2.
2.3.3 Exposure Pathways
No significant comments were received on Section 2.3.3.
2.3.3.1 Aquatic Exposure Pathways
No significant comments were received on Section 2.3.3.1.
2.3.3.2 Terrestrial Exposure Pathways
No significant comments were received on Section 2.3.3.2.
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2.4 Assessment Endpoints
No significant comments were received on Section 2.4.
2.5 Measurement Endpoints (Measures of Effect)
Response to EG-1.5
The comparison between the Hudson River ERA and the Clinch River ERA was addressed
in the response to comment EG-1.1 in the Responsiveness Summary for the ERA (USEPA, 2000b).
Although the Hudson River and the Clinch River are both large contaminated sites, they are not
directly comparable. The Oak Ridge Reservation (ORR) is owned and administered by the
Department of Energy (DOE) and has fewer outside influences than the Hudson River. Performing
top-down studies that start with field population and community information may not accurately
represent the effects of PCBs, since others factors (e.g., fishing ban) may mitigate some of the effects
of PCBs (see, Responsiveness Summary for the ERASOW [USEPA, 1999b] at p. 13). Due to
concerns that a top-down approach would not be protective of biological resources of the Hudson
River, more weight was placed on use of toxicity quotients.
USEPA's bottom-up approach uses data on individuals in order to predict potential effects
on local populations and communities that occur or could occur at the site. USEPA's Risk
Management Guidance (OSWER Directive 9285.7-28P, p. 3) states,
"Levels that are expected to protect local populations and communities can be estimated by
extrapolating from effects on individuals and groups of individuals using a lines-of-evidence
approach. The performance of multi-year field studies at Superfund sites to try to quantify
or predict long-term changes in local populations is not necessary for appropriate risk
management decisions to be made."
USEPA used, among other things, observed concentrations of PCBs in benthic invertebrates and fish
in the Hudson River and field studies of birds and mammals in and along the Hudson River, in a
weight-of-evidence approach to characterize risks to ecological receptors (see. ERA Addendum,
Section 5.0: Risk Characterization. Also see response to EG-1.1 in USEPA, 2000b).
Conducting various studies on the Lower Hudson River beyond what NYSDEC, US Fish and
Wildlife Service, and others are already conducting would have provided more elements to the
weight of evidence approach, but also would have introduced such broad uncertainties of their own
that they are unlikely to have reduced general uncertainty in the assessment, as most of the data used
were collected in 1993. The decision not to conduct new site-specific toxicity studies was described
in the Responsiveness Summary for the ERA Scope of Work (see. USEPA 1999a, p. 27).
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2.6 Receptors of Concern
No significant comments were received on Section 2.6
2.6.1 Fish Receptors
No significant comments were received on Section 2.6.1.
2.6.2 Avian Receptors
No significant comments were received on Section 2.6.2.
2.6.3 Mammalian Receptors
No significant comments were received on Section 2.6.3.
2.6.4 Threatened and Endangered Species
No significant comments were received on Section 2.6.4.
2.6.5 Significant Habitats
No significant comments were received on Section 2.6.5.
3.0 EXPOSURE ASSESSMENT
Response to EF-1.12
Estimation of striped bass body burdens was based on wet weight.
Response to EG-1.6
The TQ approach used in the ERA Addendum was described in the September 1998 Scope
of Work for the ERA (USEPA, 1998b). USEPA noted that in the ERA it would address the
uncertainty associated with using reference concentrations derived from the scientific literature,
rather than from site-specific lexicological studies (see. Responsiveness Summary for the ERA
Scope of Work [USEPA, 1999a] at p. 27). This issue is addressed in the ERA (USEPA, I999b) at
pp. 157-158.
The use of TQs is part of USEPA's bottom-up approach that uses data on individuals in order
to predict potential effects on local populations and communities that occur or could occur at the site.
A recent USEPA directive (OSWER Directive 9285.7-28P, p. 3) states, "Levels that are expected
to protect local populations and communities can be estimated by extrapolating from effects on
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individuals and groups of individuals using a lines-of-evidence approach. The performance of multi-
year field studies at Superfund sites to try to quantify or predict long-term changes in local
populations is not necessary for appropriate risk management decisions to be made."
Response to EG-1.11
USEPA disagrees that the models used to predict future PCB concentrations are deficient.
The revised HUDTOX and FISHRAND models (USEPA, 2000a) were peer reviewed by a panel of
independent experts and found to be acceptable with revisions. USEPA is currently evaluating the
recommendations of the peer review panel and will issue a written response to the reviewers'
recommendations.
Response to EG-1.12a
The commenter raises several concerns regarding the use of the HUDTOX model to estimate
PCB loads delivered to the Lower Hudson via the Troy dam. These comments were addressed in the
Responsiveness Summary for the BMR (USEPA, 2000c). Of relevance are the responses to GE's
comments on PCB fate and transport issues (BG-17 through BG-20); description of model
development (BG-21 through 28); and prediction of water column levels (BG-1.32). Also, the
commenter questions whether the HUDTOX model is consistent in its use of equations and
coefficients, whether USEPA converted Tri+ loads to individual homologue groups on an adequate
data set, and whether some of the data used may be limited in usefulness. With regard to the first
concern, the HUDTOX model consistently applies a set of mathematical formulations to the entire
Upper Hudson. Different equations are not used in different nver regions. However, the model does
apply different coefficients among the various reaches to account for the observed differences in
PCB loading and other conditions that vary as a function of river reach. These adjustments were
derived during the calibration of the HUDTOX model to the data (see. Sections 6 and 7 of the
RBMR (USEPA, 2000a) for discussions of the data used and the model calibration, respectively).
Response to EG-1.13
The commenter asserts that the Farley model is largely untested and therefore the uncertainty
associated with the model's veracity undermines its application in the ERA Addendum. However,
USEPA reviewed the Farley model specifically for use in the ERA Addendum. The USEPA
acknowledges that the data set available to calibrate a PCB fate and transport model in the Lower
Hudson is limited. However other data and analyses are available (USEPA, 1997; USEPA 1999c)
that independently confirm the conclusions drawn from the modeling analysis. For example, the
conclusion that the principal source of PCBs to the Lower Hudson is the Upper Hudson is directly
supported by the high-resolution core analysis presented in the DEIR (USEPA, 1997). Similarly, the
gradual decline in concentration of PCBs in surface sediment as shown in Figure 3-7 of the ERA
Addendum is confirmed by the analysis of the high-resolution cores presented in USEPA (1997).
Additionally, although this version of the model has not yet been subjected to peer review, earlier
versions of the model developed by Thomann were peer reviewed and published (see, Thomann et
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al., 1989 and Thomann, 1989). It is USEPA's understanding that the authors of Farley et al. (1999)
will submit their work in a paper for publication in a peer reviewed scientific journal.
The comment that the model is biased toward the lower chlorinated congeners ignores the
application of the model in the ERA Addendum. The original model presentation is not "biased"
but simply based on the measured loads at the TI Dam as reported by GE. As documented in the
ERA Addendum (USEPA, 1999c) and the RBMR (USEPA, 2000a), the load measured at the TI
Dam is not the same as that delivered at Waterford. A significant change takes place during transit,
largely resulting from the loss of lighter congeners from the water column. To account for this, in
the ERA Addendum USEPA utilized loads derived from the HUDTOX model runs, which addressed
these losses and more correctly estimated the homologue patterns at Waterford. The effect of this
correction can be seen in Figure 3-2 of the ERA Addendum, which compares the cumulative loads
as estimated from HUDTOX and by Farley et al. (1999). As a result of the correction, the proportion
of dichloro homologue in the Waterford load is greatly decreased. Figure 3-6 shows that while the
model estimates of water column concentration are low, the proportions of the various homologues
as predicted by the model are very similar to those measured in the water column. Thus the model
is not "biased toward lower chlorinated congeners."
The USEPA acknowledges that the period of data available for the fate and transport model
is largely limited to one year. However, an integrative test of the model for the purposes of risk
assessment is provided by the long-term record of fish body burdens obtained by NYSDEC. This
data set suggests that the combined fate, transport, and bioaccumulation models are able to predict
fish body burdens within an acceptable range of accuracy for the ERA Addendum, given that risk
is largely resolved on an order-of-magnitude basis.
Lastly, USEPA disagrees that high resolution cores should be incorporated in Figure 3-7 of
the ERA Addendum. The high resolution cores represent unusual depositional environments,
whereas the purpose of Figure 3-7 is to depict more spatially representative sediment samples. For
this reason, the sediment samples collected for the ERA and ERA Addendum are included in Figure
3-7. Nonetheless, even these data set are not truly representative of all sediment environments.
While these data show general agreement with modeled concentrations of PCBs in sediment, the
data are not intended as a calibration point for the model.
Response to EG-1.14
USEPA does not agree that differences between the Farley model and FISHRAND must be
reconciled before reliable forecasts can be made. Any model of PCB bioaccumulation involves
approximations (i.e., empirical assumptions and coefficients) regardless of the degree of mechanistic
representation. As demonstrated in Appendix K of the ERA (USEPA, 1999b), fish body burdens
resulting from Lower Hudson River exposures cannot be exclusively linked to either their sediment
or water exposures. As noted in the DEIR (USEPA, 1997), sediment and water column
concentrations are not independent because resuspension and sediment release yields PCBs to
overlying water, which can then serve to redeposit these PCBs in quiescent regions downstream.
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Given that these matrices are linked, it is possible that the use of the semi-empirically determined
coefficients would yield similar results from two different mechanistic representations, especially
when considered over the long term. The only reliable test as to the veracity of the model forecasts
is the degree to which the model replicates existing data to within an acceptable level of error. Both
model forms (FISHRAND and Farley) have been shown to do this.
The commenter also suggests that the striped bass body burdens as estimated from the
largemouth bass underestimate contributions from the saline portion of the Lower Hudson and
consequently over predict site-related risks. As noted by the commenter, the FISHRAND model was
not used to characterize striped bass body burdens directly. Rather, an order-of-magnitude
approximation was used based on an empirical, data-based approach (i.e., ratios of PCB
concentrations in striped bass and largemouth bass from fish data for these two species sampled at
the same location and same time frame). This empirical ratio approach captures the current
relationship between the freshwater and saline regions without making an explicit assumptions
regarding the striped bass migratory patterns with a level of precision that is appropriate for use in
the ERA Addendum. In the freshwater region of the Lower Hudson, striped bass body burdens are
clearly dominated by PCBs from the Upper Hudson, as shown by an increase in body burden with
increasing river mile towards the GE plants, which are the source of the PCBs in the Upper Hudson
River sediments (see Table EG-1.14 below). Given that the Upper Hudson source dominates in the
Farley Food Web Region 1, the relative proportion of PCBs from the freshwater and saline Hudson
is unimportant in this region.
Table EG-1.14
Comparison of Mean Striped Bass Body Burdens at Two Long-Term Monitoring Locations
(Data from NYSDEC)
Year
1990
1991
1992
1993
1994
1995
1996
RM113
4.64
NA
2.94
3.27
2.30
1.11
1.66
RM152
9.02
NA
15.32
10.92
5.61
NA
4.28
28
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Response to EG-1.15a
As shown in Figure 3-2 of the ERA Addendum (USEPA, 1999c), the Farley model results
derived using HUDTOX loads to the Lower Hudson tend to yield slightly lower values before 1992
and slightly higher values after 1992 relative to the available data. Given that the Farley model was
calibrated using the PCBs loads estimated by Farley et al. (1999) rather than HUDTOX, a difference
in values would be expected. The slight lack of agreement in PCB loads has a minimal effect on the
estimate of risks to striped bass because the uncertainty introduced is minimal and does not change
the overall conclusions regarding risk to striped bass. Moreover, the model output for striped bass
aged 6-16 years in Food Web Region 2 is not the only metric used to measure risks to striped bass
in this region (body burden estimates for striped bass aged 2-6 years also were used). Thus, USEPA's
decision to use the Farley model results with the HUDTOX PCB loads was appropriate in the ERA
Addendum.
Response to EG-1.15b
The commenter suggests that the fact that the FISHRAND model does not yield fish body
burdens reflecting an impact of the 1991-1992 Allen Mills event, which in turn indicates that
exposure point concentrations and food web structure may be inaccurate in the model. However,
the data referred to by the commenter do not present a consistent picture of response to an increased
load from the Allen Mills event. Specifically, body burdens (wet-weight basis) for white perch and
brown bullhead peak in 1992, largemouth bass peak in 1993 and yellow perch steadily increase from
1991 to 1996. Young-of-the-year pumpkinseed at RM 142-152 (Figure 3-12c) show only a gradual
decline in both wet-weight and lipid-normalized PCBs from 1988 to 1996, with 1993 being the
lowest value recorded. Lipid-normalized results reflect similar disagreement among the species:
largemouth bass data from 1992 are not different from 1990 data; white perch, brown bullhead and
yellow perch peak in 1992 but are similar to the values seen in prior and subsequent years. These
data suggest that the Allen Mills pulse is of relatively minor importance over the long-term.
Predicted concentrations from the USEPA's FISHRAND model represent annualized
concentrations, which tend to average out varying concentrations throughout the year. Thus, the
response to the pulse loads may be partially smoothed by the approach. (The fish data represent
conditions during a few brief weeks of sample collection.) Additionally, the Farley model is not
designed to represent short-term behavior of PCB fate and transport. Rather, it is focused on long-
term trends. Nonetheless, USEPA is satisfied that the model is reasonably able to capture the long-
term trend and it therefore appropriate for use in the ERA Addendum. USEPA does not plan to use
the Farley model to quantitatively assess the effects of remedial activities in the Upper Hudson on
the Lower Hudson.
Response to EG-1.15c
In the ERA Addendum, FISHRAND model comparisons to data on a wet-weight basis were
emphasized over lipid-normalized predictions (Figures 3-12a,b, c, and d). This is because risk
29 TAMS/MCA
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estimates for all avian and mammalian receptors attributable to ingestion of PCB-contaminated fish
are based on predicted wet-weight concentrations. (Lipid-normalized results are used to evaluate
TEQ-based risks to fish). Additionally, a single version of the model for largemouth bass was
applied to all of the freshwater Lower Hudson. In this manner, the model was calibrated to meet
available data at both RMs 152 and 113. The model provides acceptable output on a wet-weight
basis at both locations and tends to underestimate lipid-normalized results, particularly at RM 152.
The lipid contents used in the model were obtained from NYSDEC and were based on monitoring
data. The potential underprediction in comparisons to data suggests that true TEQ-based risks to fish
may be underestimated at RM 152.
Response to EG-1.16a
The largest piscivorous ecological receptors, represented by bald eagle and river otter, will
consume any fish within a particular size range and favor large fish. To estimate the dietary dose
of PCBs for piscivorous birds, the data were divided into smaller (< 10 cm) and larger (> 25 cm)
fish, where smaller fish included minnows and sunfish while larger fish included catfish and bass
(see. USEPA 1999c, p. 52). The FISHRAND model predicts population distributions of PCB
concentrations in several larger fish species, including largemouth bass, brown bullhead, white perch
and yellow perch. Because there were no data available to suggest preferential fish selection by these
two species (i.e., no data to suggest that otter or eagle preferentially consume white perch, for
example), the largemouth bass was used as a surrogate species to represent larger piscivorous species
likely to be consumed by otter and eagle. Largemouth bass are the most abundant species in the river
and are more likely to inhabit the nearshore areas that serve as foraging areas for the otter and the
eagle.
The remaining ecological receptors have been shown to consume a much smaller size of fish.
For these species, the predicted population distributions of PCB concentrations in the forage fish
were used, as modeled for pumpkinseed and spottail shiner. Again, as there was no information
available to suggest preferential selection of these fish by any of the ecological receptors, the spottail
shiner was chosen as the representative surrogate forage fish. Most of the data in the Lower Hudson
River were available for the spottail shiner, and the spottail shiner has been shown to be one of the
most abundant forage fish species. Note that predicted PCB concentrations in spottail shiner are very
similar to those predicted for the pumpkinseed, suggesting it is appropriate to use one species as
representative of typical expected PCB concentrations in forage fish generally.
Response to EG-1.16b
The FISHRAND model was calibrated in the Upper Hudson River using growth rate
coefficients for each species as a calibration parameter. Consequently, the growth rates used in the
FISHRAND model were not the generic growth rates of the Gobas Model (except for spottail shiner,
which is not sensitive to this parameter and for which no data are available, see Response to EF-1.4).
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3.1 Quantification of PCB Fate and Transport: Modeling Exposure Concentrations
No significant comments were received on Section 3.1.
3.1.1. Modeling Approach
No significant comments were received on Section 3.1.1.
3.1.1.1 Use of the Farley Model
Response to EF-1.10
The models presented in the RBMR (USEPA, 2000a) address the potential impact of high
flow events on the Remnant Deposits and other areas of relatively high concentrations of PCBs
upstream from Rogers Island by modeling 0 ng/L and 30 ng/L upstream boundary conditions in
addition to 10 ng/L. The largemouth bass body burdens for the 10 ng/L and 30 ng/L conditions are
given in Tables 7-10 and 7-13 of the RBMR (USEPA, 2000a). These figures show approximately
a threefold increase in concentration in the long term "steady-state" value (asymptote) between
upstream boundary conditions of 10 ng/L and 30 ng/L. Risks vary in direct proportion to
concentration, resulting in a threefold increase in risk at 30 ng/L. These increases in concentration
and risk are also seen for the brown bullhead, Tables 7-11 and 7-14 of the RBMR (USEPA, 2000a).
In the Lower Hudson, the system response is expected to be less than this one-to-one response
estimated for the Upper Hudson. That is, body burdens will not respond in a strict proportional
fashion to changes in the Upper Hudson load, because the local sediment inventory resulting from
previous GE related contamination is expected to continue to effect fish levels for the entire forecast
period.
Response to EF-1.11
The Farley model output is sediment, water column, and fish body burdens. The modeled
species are white perch in Food Web Regions 1 and 2 and striped bass in Food Web Region 2 only.
The FISHRAND model output is for pumpkinseed, spottail shiner, yellow perch, white perch, brown
bullhead and largemouth bass. Striped bass concentrations in Food Web Region 1 were estimated
from FISHRAND forecasts for PCBs in largemouth bass, based on the ratio between the two species
in sample data for the same location and time period (see. Section 3.1.1.2 of the ERA Addendum).
PCB concentrations in white perch from the Farley model were used in the ERA Addendum in
preference to the FISHRAND output, because the Farley model has been designed specifically for
migratory species.
Response to EL-1.6
As discussed in Section 3.1.1.1 of the ERA Addendum, the Farley model was used with few
adjustments to predict future concentrations in Lower Hudson River sediment, water, and fish. The
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PCB loads from HUDTOX were converted from Tri+ PCBs to di through hexa homologues. These
homologue values replaced the Upper Hudson River PCB loads that were used in Farley et al. (1999)
and provide consistency between the Upper and Lower Hudson for purposes of the ERA and ERA
Addendum. Because the Farley model was designed to run for a period of up to 15 years, the models
were run four times to generate a 40-year forecast. The Upper Hudson River PCB loads from the
HUDTOX model for the appropriate time period were put into each of the four sets of input files.
The initial conditions for each model segment for the fate and transport models and species for the
bioaccumulation model were updated with data from the previous model run. The final
concentration in a segment (or species) becomes the initial concentration in the same segment for
the next model run. Each initial condition is determined by the previous model run, which is no
different than running the model on a continuous basis for the 40-year simulation period. This 15-
year step approach is simply a requirement of the original model coding and does not affect the
model results.
3.1.1.2 Use of FISHRAND
Response to EF-1.13
USEPA's comparison of the Farley model output to August and September 1993 water
column data suggests that the model underpredicts late summer water column concentrations (but
not spring concentrations). A possible reason for this discrepancy may be an overestimation of water
column losses to volatilization. However, this inference is based on only three observations (i.e.,
data points), which are not sufficient to clearly diagnose the existence of a bias, given the temporal
variability typically observed in water column concentrations. Different PCB loads at the Federal
Dam upper boundary loads is also a possible explanation. It should also be noted that the Farley
model is not strictly mechanistic, does not use fine-scale dynamic simulation, and is calibrated to the
available fish data, which integrate over time. Thus, while the Farley model may not represent
seasonal changes in water column concentrations, it does appear to do a reasonable job in replicating
average concentrations in fish.
In addition, overestimating loss of PCBs from the water column is likely to reduce water
column concentrations, but this loss is likely to occur for the lighter chlorinated and less toxic
congeners, which do not tend to accumulate in tissue. Thus, although the loss may be overestimated
(which would underestimate nsk), if the loss occurs primarily for these lighter chlorinated and less
toxic congeners, as would be expected, then there would be little or no effect on estimated risks.
3.1.1.3 Comparison to the March 1999 Farley Model (1987-1997)
Response to EL-1.7a
All fish body burdens, except for striped bass and white perch, are modeled using the
FISHRAND model. The FISHRAND model provides species-specific results for several different
species, while the Farley model only provides two to three estimates, depending on river region.
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Specifically, the Farley model provides estimates for white perch in all regions, a generic.forage fish
in all regions, and striped bass in Food Web Regions 2 to 5. As noted by the commenter, the Farley
model does not provide striped bass estimates in Food Web Region 1, which is between RMs 152
and 73.5. Yet, because PCB concentrations in striped bass were needed in this region, it was
necessary to apply the FISHRAND model in this region (using a ratio approach based on largemouth
bass). The fact that Farley et al. (1999) did not develop a striped bass estimate for Food Web Region
1 is likely a result of the model's focus on the saline portion of the Lower Hudson and New York
Harbor. It is not for lack of data above RM 73.5, as NYSDEC collected over 200 adult striped bass
samples from 1977 to 1997 at RM 152 and RM 113. As to its use in the ERA Addendum, the
FISHRAND model provides a finer spatial scale than does the Farley model, and to be consistent
with how body burdens are presented for the other species, the ERA Addendum uses the
FISHRAND results. Although the Farley model provides sufficient detail for sediment and water
column estimates, it does not provide sufficient fish body burden data to support the requirements
of the ERA Addendum (and the Mid-Hudson Human Health Risk Assessment).
Response to EL-1.7b
As discussed above and in response EL-1.14, the locations at RM 152 and 113 represent the
primary fish sampling locations for all species in the freshwater Lower Hudson and thus form the
main calibration and forecast locations in this region of the river.
Response to EL-1.7c
The bald eagle and river otter, representing the largest top-level piscivores, are assumed to
eat the fish portion of their diet as largemouth bass, as these are the fish most likely to be caught by
these receptors. All other ecological receptors that consume fish (i.e., mink, raccoon, great blue
heron, belted kingfisher) consume a smaller fish, which was assumed to be the pumpkinseed. Thus,
adult striped bass are not included in subsequent higher trophic level avian and mammalian
exposures.
Response to EL-1.7d
The ratio approach was designed to approximate PCB concentrations in striped bass because
the Farley model did not provide striped bass results for Region 1 despite the documented occurrence
of striped bass at this location (i.e., RM 152, RM 113) and because the FISHRAND model was not
parameterized for striped bass. The ratio approach was based on the observed relationship between
measured PCB concentrations in striped bass and in largemouth bass from the historical NYSDEC
data. There are not sufficient data available for striped bass and largemouth bass to compare
concentrations at RMs 90 and 50. The Farley model provides predicted concentrations of PCBs in
striped bass for the region between RMs 73 and 33, and thus the ratio approach was not required
there.
33 TAMS/MCA
-------
With regard to the use of the largemouth bass, Figure K-17 of the ERA (USEPA, 1999b)
shows that the largemouth bass (piscivores) has the most gradual decline of any feeding guild. Thus,
while the use of the ratio approach adds uncertainty, it is unlikely that this approach is unduly
conservative.
In constructing the ratios, only adult fish for each species were used. The difference between
the ratios for RMs 152 and 113 may be attributable to the migratory behavior of the striped bass
relative to the resident behavior of largemouth bass. However, given that this approach is purely
empirical, it is of little importance as long as the body burdens are correlated (see also, response to
EL-1.14).
Response to EL-1.7e
The FISHRAND model also addresses the age of the animal by approximating the
distribution of fish sizes and ages in the population. This is described in detail in the RBMR
(USEPA, 2000a).
Response to EL-1.8
Revised Table 3-3 and Figure 3-2 are given in Table EL-1.8 and Figure EL-1.8, respectively,
which show consistent units for the Upper Hudson River PCS loads presented in the RBMR
(USEPA, 2000a). Revised text for Section 3.1.1.3, paragraph 4 of the ERA Addendum is as follows:
Comparison of HUDTOX and Farley et al. (1999) PCS 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 Farley
model compare favorably with the estimates from HUDTOX for the period 1987-
1997. The largest difference among the tri to hexa homologues is 278 kg for the tetra
homologue, representing a cumulative difference of about 13 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, 428 kg (36 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 di homologue, which matters least to Lower Hudson fish body
34 TAMS/MCA
-------
Table EL-1.8
Cumulative Loads Over the Troy Dam (kg)
Homologue
Di
Tn
Tetra
Penta
Hexa
Total 1987-1997
Farley Model
1182
2320
1664
715
270
6151
HUDTOX
Converted
According to
Appendix A
1610
2097
1386
617
220
5930
Difference
428
-224
-278
-98
-50
-222
Homologue
Di
Tn
Tetra
Penta
Hexa
Total 4/91-2/96
Farley Model
857
1645
1081
406
145
4134
HUDTOX
Converted
According to
Appendix A
566
856
593
249
83
2348
TI Dam Estimate
from the DEIR1
638
1072
672
214
65
2662
Note.
1. Homologue loads were recalculated using the averaging estimator formula descnbed
in the DEIR (USEPA, 1997) as originally given in Dolan et al (1981):
Lm =
7=1
where.
L m is the load estimate for month m;
Nm is the number of days in month m,
q, is the daily mean flow on day j,
n „ is the number of days on which PCB observations were
made during month m. and
c, is a measured concentration within the month
Reanalyzed GE water column data are used (GE, 2000) In addition the TI Dam
bias is accounted for as described in Appendix C of the Responsiveness
to the Low Resolution Sediment Coring Report (USEPA, 1998)
-------
harleycl »!.. 1999
USEPA.2000
I i i i | i i l I i i i | i ' i I i i i I
1992 1993 1994 1995 1996 1997 1998
1987 1988 1989 1990 1991
Farley a il.
USEPA.2000
1992 1993 1994 1995 1996 1997 1998
1988 1989 1990 1991
Parity el il., 1999
USEPA.2000
.
1992 1993 1994 1995 1996 1997 1998
600 -.
500 4
400 4
FlrieyetlL.
USEPA, 2000
Farley il al.
USEPA.2000
1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998
Sources: Farley et al., 1999 and USEPA, 2000
Figure EL-1.8 (Revised Figure 3-2)
Comparison of Cumulative PCB Loads at Waterford from Farley et al., 1999 and
USEPA, 2000
TAMS/MCA
-------
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 second part of Table 3-3 compares three estimates of the PCB loads from April 1991
through February 1996 all of which are derived from data obtained by GE's federally mandated Post-
Construction Remnant Deposit Monitoring Program. The estimates are from Farley et al. (1999), a
conversion of HUDTOX model output (ERA Addendum) and the DEIR (USEPA, 1997). The DEIR
estimate in Table EL-1.8 has been updated to account for the TI Dam bias issue (USEPA, 1999d).
The Farley Model and DEIR estimates are of loads over the TI Dam assuming no gain or loss of
PCBs between the TI Dam and the Troy Dam while the HUDTOX estimate uses the modeled
concentrations over the Troy Dam. The HUDTOX estimate is the lowest estimate for di through hexa
homologues at 2,348 kg versus 4,134 kg for the Farley Model estimate and 2,662 kg for the DEIR
estimate. For the di homologue, the converted HUDTOX value is lower (72 kg) than the DEIR
estimate and substantially less than the Farley estimate (291 kg). For tri, tetra and penta homologues,
the converted HUDTOX estimates are considerably less than the Farley Model estimate, but similar
to the DEIR estimate. The higher Farley Model load estimate is based on the original GE data which
was not corrected for the TI Dam bias. The corrected data (QEA, 1998) were used for the HUDTOX
and revised DEIR estimates.
3.1.1.4 Comparison Between Model Output and Sample Data
Response to EF-1.14
The text on p. 20, paragraph 3 of the ERA Addendum is corrected to read:
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 59, which falls above and below the modeled data and for RM 47, which falls
above. At RM 47, 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.
Response to EL-1.9
Concentrations of PCBs in striped bass are not explicitly modeled for Food Web Region 1
because the striped bass has not been parameterized within the FISHRAND model (that is, it was
not one of the six fish species targeted for modeling). Consequently, the approach to estimating
concentrations of PCBs in striped bass was to develop a ratio based on observed concentrations of
PCBs in largemouth bass (wet weight basis), which is a target species for FISHRAND, with
observed concentrations of PCBs in striped bass at the same location and within the same time
period.
37 TAMS/MCA
-------
3.1.1.5 Comparison of White Perch Body Burden between the Farley
Model (Using Upper River Loads from HUDTOX) and
FISHRAND
Response to EL-1.10
The concentrations of PCBs in white perch estimated by FISHRAND at each RM are
presented in Figure 3-10 of the ERA Addendum. While the average of the FISHRAND data from
each station would provide a more concise comparison between the FISHRAND and Farley models,
the values shown in the figure were presented because they were the values used to calculate risk in
the ERA Addendum.
3.1.1.6 Comparison Between FISHRAND Output and Sample Data
Response to EF-1.15
The NYSDEC 1998 data were not available in the database at the time the ERA Addendum
was issued.
Response to EL-1.11
Because PCB concentrations in striped bass were not directly modeled, it is not possible to
compare modeled results to data, as was done for the other fish species.
3.1.2. Model Results
No significant comments were received on Section 3.1.2.
3.1.2.1 Farley Model Forecast Water Column and Sediment
Concentrations
No significant comments were received on Section 3.1.2.1.
3.1.2.2 Farley Model Forecast Fish Body Burdens
Response to EL-1.12
Figures 3-16 and 3-17 show the fish body burden bimonthly concentrations of Tri+ PCBs
modeled by the Farley bioaccumulation model (represented by "x") versus monthly load of Tri+
PCBs from the Upper Hudson River as modeled by HUDTOX. The range of values show the
seasonal variation. Figure EL-1.12 shows the revised Figure 3-17 with the corrected title (i.e.,
Striped Bass Body Burdens in Food Web Region 2).
38 TAMS/MCA
-------
3.1.2.3 FISHRAND Forecast Fish Body Burdens
Response to EL-1.13
The FISHRAND model compares favorably to data and provides concentrations of PCBs in
at finer spatial scales than does the Farley model. Consequently, the risk estimates presented in the
ERA Addendum rely primarily on the results from the FISHRAND model. Because striped bass
were not directly modeled, it is not possible to compare modeled results to data, as was done for
other fish species.
3.1.3 Modeling Summary
No significant comments were received on Section 3.1.3.
3.2 Exposure Point Concentrations
Response to EL-1.14
The selection of the specific river miles for the purposes of the FISHRAND simulations in
the Lower Hudson was based partially on the available data for the Lower Hudson. The RMs 152
and 113 correspond to the locations used by NYSDEC in its long-term (20 years or so) fish
monitoring. Therefore, these locations represent the best locations for the calibration of the
FISHRAND model. Because the model has been calibrated at these locations, they are also the best
locations for forecasting fish body burdens, and thus were used in the ERA Addendum. The other
locations, RMs 90 and 50, were added to achieve approximately equal spacing among the simulated
locations. For each of these locations, FISHRAND is used with the sediment and water results from
the Farley model to estimate fish body burdens in the resident fish species (i.e., all species except
striped bass). In the ERA Addendum, striped bass are only estimated for RMs 152 and 113 because
these are the only locations with sufficient striped bass and largemouth bass data to establish a ratio
between the two species. Thus striped bass are calculated from the FISHRAND output for
largemouth bass by the application of the ratios described in Section 3.2. (For the Mid-Hudson
Human Health Risk Assessment, striped bass body burdens also were estimated at RM 90 by
applying the ratio from RM 113 to the forecast concentrations of PCBs in largemouth bass at RM
90.)
Overall, the range of FISHRAND stations spans from RM 152 to 50, which is comparable
to the sediment and water ranges simulated by the model, RM 153.5 to 33.5. The differences in
historical fish body burdens between adjacent monitoring stations are generally less than the
uncertainty associated with the median value (compare single species for individual years among
Figures EL-14a, b and c [revised Figures 3-12a to 3-12c]; see also Figure 12 in the report
39 TAMS/MCA
-------
ti
_
HUDTOX Federal Dam Flux (kg/d)
o -
1980
2000
2020
2040
2060
2080
~ 4 -
5
•
° ±
"" u
3 -
2 -
1 -
Striped Bass Age Class 2-6 Years
1980
2000
2020
2040
2060
2080
3.5 -
C
£ DC
•5 £ 2.5 -.
3 J2
a a
2
~ + 1-5
03 .—
1 -
0.5 -i
0
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 EL-1.12 (Revised Figure 3-17)
Comparison Among the HUDTOX Upper River Load and Farley Model Estimates
Striped Bass Body Burdens in Food Web Region 2
(1987-2067)
TAMS/MCA
-------
FIGURE EL-1.14a (Revised Figure 12a): Comparison Between FISHRAND Results and Measurements at RM 152
Comparison to Data for Largemouth Bass at 152: wet
weight
20
I15'
•w
•&,„
2
•l-l
0 -
•
T
/I — T— r1
1986 1988 1990 1992
Year
r~~ — *T~
1994 1996 1998
Comparison to Data for White Perch at 152: wet
weight
""5
a. 20 .
r
M\ f
0 -
—
, . _
1989 1991 1993
Year
T,
i-*i
1995 1997
Cor
1400
i io°°
el,80
Z a 600
I 400
3 200
0
nparison to Data for Largemouth Bass at 152: lipid-
normalized
-
«
^x*
•
•
4
1
: T
f^^^ih—T— ~
L j 1 i
1987 1990
1993 1996
Year
Comparison to Data for White Perch at 152: lipid-
normalized
enn
i
*e3
| E.,nrt
« ^300 •
I 200-
a 100 =
•
T
-4
1
1
»
T
I T,
1 , * * *-
1989 1991
1993 1995 1997
Year
MCA/TAMS
-------
FIGURE EL-1.14a (Revised Figure 12a): Comparison Between FISHRAND Results and Measurements at RM 152
40
1.30
£25
fao
!£ 15
|l 10
^ 5
0
1
Comparison to Data for Brown Bullhead at 152: wet
weight
i
987 1989
i
T
1
i
_ T
1991 1993 1995 1997
Year
Comparison to Data for Yellow Perch
weight
67 .
'
•S "5 -
JS J
9>
15 2 -
^ 1 -
0 -
at 152: wet
•
4
- ^^^ 1
M
1989 1991
i •
• )
•
4
I
X
^^
1993 1995 1997
Year
Comparison to Data for Brown Bullhead at 152: lipid-
normalized
W OCA .
_g ZJU
| B?°
^^ 1 AA
•s
50 ~
•
— • — '
i
«
i^J
r—
m *
0.
1987 1989 1991 1993 1995
Year
1997
Comparison to Data for Yellow Perch at 152: lipid-
normalized
rn Rfifl
S3 £AA -
us ofln .
Ij A .
,
r ^l^~-
P .
. «
TT =
1989 1991 1993 1995 1997
Year
MCA/TAMS
-------
FIGURE EL-1.14b (Revised Figure 12b):
Wet Weight ppm
..--•»
Comparison Between FISHRAND Results and Measurements at RM 113
Comparison to Data for Largemouth Bass at 113: wet
weight
5
TT
T
^~~~~flt T 1
f\ * 1 _
^^^T— *^M I
X * • x
1987 1989 1991 1993 1995 1997
Year
Comparison to Data for Largemouth Bass at 113:
lipid-normalized
"3 r\ f \
B C •^^^' *k ,^^^J^^^_ ^L
o B. \_
Z ™ "00 — -^
s I t^
.g< 100 x
0 1 1 1
1987 1989 1991 1993 1995 1997
Year
Comparison to Data for White Perch at 113: wet
weight
t"
•§D4
j*J T
•^ 2
« i
^
•
1991
5'
T
= * J
X
•
>
1992 1993 1994 1995 1996 1997
Year
Comparison to Data for White Perch at 113
normalized
•a 250 -i
S 200 • »
P 850
o o« IT •
ZO*««* T 1
TOO Jt 4
s — • — * 1
a
0 1 1 1 1
lipid-
»
1991 1992 1993 1994 1995 1996 1997
Year
f^
MCA/TAMS
-------
FIGURE EL-1.14b (Revised Figure 12b): Comparison Between FISHRAND Results and Measurements at RM 113
_ 1
1"
£4-
^ 1 -
\ -
0 -
Comparison to Data for Spottail Shiner Wet Weight:
1993 US EPA Dataset
• D.
•r HFI
• '
-T 1I"
_! ft *I i
"" X *•
0
3
1.25
"5>. ,
1
£ }
0
50 100 150
River Mile
Jta
SHRAND
200
Comparison to Data for Yellow Perch at 113: wet
weight
— — • — T
t •
1991 1992 1993 1994 1995 1996 1997
Year
Comparison to Data for Spottail Shii
Normalized: 1993 US EPA Dat
&
I250 •
*3 onn
z 15°
"o inn H
.g 1UU
13 sn -
0.
•
J<
. 1
* n1
0 50 100
River Mile
ier Lipid
aset
i
•
i
i
150
• Data
• FISHRAND
200
Comparison to Data for Yellow Perch at 113: lipid
normalized
Ann
'js inn .
E ^?sn -
3 l50'
"5< inn -
J 50-
0 -
t *"• — -
i
1991 1992 1993 1994 1995
Year
— —
~
1996 1997
MCA/TAMS
-------
FIGURE EL-1.14c (Revised Figure 12c): Comparison Between FISHRAND Results and Measurements of Pumpkinseed
Comparison to Data for Pumpkinseed at 60: wet
weight
7
fi .
E 5 .
a
S 4 •
as
'3
^ 3 •
0.
\
\
X T ^^:L
I * IT-
I I >-" j * \T
m i *
1986 1988 1990 1992 1994 1996
Year
7 -
6.
i,4'
|3-
I2-
1 -
0 -
19
Comparison to Data for Pumpkinseed at 142 - 152:
wet weight
i ' '"""" NS^ _^-*5v
^ ± X N^^
T «
87 1989 1991 1993 1995 1997
Year
Comparison to Data for Pumpkinseed at 60: lipid-
normalized
S
"Sinn ^^
a100 ^^
M nrx ^ ^^«
p S I
§ 60 j J
Z 40 ^
P ""*
32°
i
T
t-k-q
4
I J
1986 1988 1990 1992
Year
•
1 J
•
1994 1996
Comparison to Data for Pumpkinseed at 142 -
152: lipid-normalized
nnn
•g 150
rt
1987 1989
^^
X
jp
•
1991 1993 1995 1997
Year
MCA/TAMS
-------
by Farley et al., 1999). Thus, the selection of these RM stations does not introduce excessive
conservatism.
3.2.1 Modeled Water Concentrations
No significant comments were received on Section 3.2.1.
3.2.2 Modeled Sediment Concentrations
Response to EF-1.16
The'Farley model predicts sediment concentrations on a dry weight basis. Organic carbon-
normalized concentrations must then be estimated from the dry weight concentration and an assumed
TOC. In fact, lipophilic compounds such as PCBs are strongly sorbed to paniculate organic carbon,
and, among sediments of the same depositional age, the concentration normalized to organic carbon
is often a more consistent and stable measure than dry weight concentration. Indeed, much of the
variability among dry weight concentration measurements is due to variability in TOC content.
Therefore, to back-calculate an average organic carbon-normalized PCB sediment concentration, it
is important to use the same estimate of TOC as was used in the Farley et al. (1999) report, which
is 2.5%. Given that the average sediment PCB concentration at a given location and time is really
an estimate of the expected (dry weight) concentration for an average TOC, it would not be
appropriate to compute organic carbon-normalized concentrations using extreme values from the
observed range of TOC concentrations in the Lower Hudson. In general, sediment areas with very
low TOC concentrations (e.g., sands) will have correspondingly low dry weight PCB concentrations,
although organic carbon-normalized concentrations may be similar to other locations.
Response to EF-1.17
The TCDD sediment guidelines developed by USEPA (1993) for the protection of fish, birds,
and mammals are:
• Sediment concentrations associated with low risk: 60 pg/g dry weight (dw) for fish, 21
pg/g dw for birds, and 2.5 pg/g dw for mammals; and
• Sediment concentrations associated with high risk to sensitive species: 100 pg/g dw for
fish, 210 pg/g dw for birds, and 25 pg/g dw for mammals.
However, because these TCDD sediment guidelines associated with nsk to Great Lakes
receptors (i.e., fish, birds, and mammals) were back-calculated using measured biota-to-sediment-
accumulation factors (BSAF) specific to the Great Lakes, these concentrations are not directly
comparable to risk-based concentrations that would be calculated for the Hudson River. These
measured BSAFs are based on lipid content offish and organic carbon content of sediment that are
measured in the Great Lakes. In addition, these BSAFs are specific for the Great Lakes food chain
and are likely to be different from those measured for a riverine food chain. Based on these
46 TAMS/MCA
-------
differences, the risk-based sediment concentrations for the Great Lakes (pg TCDD/g dry weight of
sediment) are not directly comparable to the Hudson River.
3.2.3 Modeled Benthic Invertebrate Concentrations
Response to EF-1.18
Benthic invertebrate data from 1993 are available for the following RMs: 122.4,100, 88.9,
47.3, and 25.8. No comparisons are available for river mile 152. River miles 122.4 and 100 were
averaged to obtain a comparable estimate for RM 113. River miles 88.9 and 100 were averaged to
obtain a comparable estimate for RM 90, and RM 47.3 was used to compare for RM 50.
Concentrations are shown as wet weight in mg/kg PCBs. These comparisons, which are provided
below, show that the concentrations of PCBs in benthic invertebrates are reasonably estimated.
RM113 RM90 RM50
Data Model Data Model Data Model
Mean 0.82 0.98 0.40 0.80 0.60 0.58
Standard Error 0.14 0.10 0.21 0.20 0.20 0.11
Response to EF-1.19
The FISHRAND model was not designed to capture bioaccumulation in highly migratory fish
species such as the striped bass, nor was it parameterized for this species (e.g., growth rate). The
Farley model explicitly models striped bass, but concentrations were only available for locations
below RM 60. Because striped bass are known to occur throughout the Lower Hudson River, and
direct modeled concentrations were not available the Lower Hudson River from the FISHRAND
model or from the Farley model above RM 60, the ratio approach was used for RMs 152 and 113.
The Farley model provides predictions of PCB concentrations in striped bass at greater spatial scales
than the FISHRAND model results for all the other species.
3.2.4 Modeled Fish Concentrations
Response to EL-1.15
There are no PCB body burden data available for shortnose sturgeon (an endangered species).
Thus, there are no data to calibrate a model. The approach taken in the ERA Addendum
approximates concentrations of PCBs in shortnose sturgeon based on the similarity to the brown
bullhead (feeding preferences, etc.).
Development of the FISHRAND model (and all of the bioaccumulation models) is
constrained by data availability. All of the monitoring data from 1977 to 1997 for largemouth bass,
yellow perch, white perch, brown bullhead are expressed on a fillet basis. To calibrate the model
47 TAMS/MCA
-------
and demonstrate model functionality, predictions from the FISHRAND model are also expressed as
a concentration of PCBs in fillet.
3.3 Identification of Exposure Pathways
Response to EL-1.16
The responses to comments on the on exposure pathways of the August 1999 ERA, which
were not reiterated in the comment, have been provided in the Responsiveness Summary for the
ERA (USEPA, 2000b, see responses in Section 3.4, Identification of Exposure Pathways).
3.3.1 Benthic Invertebrate Exposure Pathways
No significant comments were received on Section 3.3.1.
3.3.2 Fish Exposure Pathways
Response to EF-1.20
The NOAA (1999) report considered reproductive, developmental, and immunotoxic effects
on fish. These effects were selected as biological endpoints that are both sensitive to anthropogenic
contaminants and ecologically relevant. The ERA (USEPA, 1999b) developed a more narrow
definition of ecologically relevant endpoints, which included reproductive and developmental effects
but not immunotoxic effects. Immunotoxic effects were not included because such effects are often
less clearly related to the assessment endpoints than are developmental and reproductive effects.
In addition, NOAA (1999) reported data that were measured or converted into concentrations
in adult liver tissue. The relationship between the concentration in liver tissue and the concentration
in whole fish has not been well studied for most species. Therefore, the ERA Addendum gives
preference to studies that measured concentrations in whole fish. For dioxin-like compounds, most
studies examined effects on the basis of concentrations in eggs. However, the relationship between
concentrations in eggs and whole fish is better characterized than the relationship between liver
concentration and adult tissue concentration. In addition, more data are generally available for
effects associated with concentrations of PCBs in whole tissue.
3.3.3 Avian Exposure Pathways, Parameters, Daily Doses, and Egg
Concentrations
No significant comments were received on Section 3.3.3.
48 TAMS/MCA
-------
3.3.3.1 Summary of ADDEjlpected, ADD9S%UCL, and Egg Concentrations for
Avian Receptors
No significant comments were received on Section 3.3.3.1.
3.3.4 Mammalian Exposure Pathways, Parameters, and Daily Doses
No significant comments were received on Section 3.3.4.
3.3.4.1 Summary of ADDExp€£ted and ADD9S%UCL for Mammalian
Receptors
No significant comments were received on Section 3.3.4.1.
4.0 EFFECTS ASSESSMENT
Response to EL-1.17
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), which General Electric Company released into the Hudson River. The
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 (see, Appendix
K of USEPA, 1999b). Therefore, it is believed that Clophen A50 is a reasonable surrogate for the
composition of PCBs in Hudson River fish.
The revisions to the TRVs based on the Bengsston study followed a review of the comments
received on the ERA (USEPA, 2000b). The Bengsston study presented results for three dose groups
and a control group. Originally, the values of 170 mg/kg and 15 mg/kg were selected based on a
hatchability endpoint. For this endpoint, the high dose group was significantly different from the
control group, but not the low and medium dose groups. However, for the hatching time endpoint,
the medium and high dose groups were significantly different from the control group. Hatching time
is a less relevant endpoint than hatchability, however, Bengsston (1980) noted that premature
hatching "resulted in premature death of the fry" and that "very few survived for more than one week
after hatching." Because there were no formal statistics conducted in association with this statement,
the TRVs selected for the ERA were based on hatchability. However, USEPA received numerous
comments that, given the observed premature death of the fry, hatching time was a relevant endpoint.
As a result, USEPA revised the TRVs to reflect the concentrations derived from the hatching time
endpoint in the Bengsston (1980) study.
Based on additional comments received on the ERA Addendum, USEPA has selected the
Hansen et al. (1974) study rather than the Bengsston (1980) study, based on the rationale explained
in the response to comment EF-1.6 and Section ffl of this Responsiveness Summary.
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Response to EG-1.7
The NOAA Sediment Effects Concentrations (SECs) were used as sediment guidelines, not
TRVs, in the ERA Addendum (see, USEPA, 1999c, p. 38). Their use as guidelines is consistent with
accepted scientific practice. A detailed response on the development and use of the SECs by the
author of the document is contained in response to EG-1.40 of the Responsiveness Summary for the
ERA (see, USEPA, 2000b, pp. 75-80).
Response to EG-1.8
For the TEQ analysis, BZ#126 was used at the detection limit to compensate for not having
quantitated BZ#81, as described in the ERA (see. USEPA, 1999b, pp. 38-40). An analysis
evaluating the proportion of TEQ congeners in USEPA Phase 2 data and USFWS tree swallow data
showed that the proportion of BZ#126 in the Phase 2 dataset was approximately equal to the sum
of the BZ#126 and BZ#81 in the USFWS dataset (see. USEPA, 1999b, Appendix J). This approach
does not produce an overly conservative estimate of TEQ risks, because they are dominated by
BZ#126 (and presumably BZ#81) and thus may be too high by at most a factor of two. This is a
relatively small margin of error, given that the calculated risk levels exceed USEPA's levels of
concern by orders of magnitude.
The NOAA report (1999) applies the TEQ approach for estimating risk to Hudson River fish,
but identifies two areas of uncertainty. First, the NOAA (1999) report notes that the available data
indicate large interspecies differences in early life stage toxicity. Second, the report notes the risk
estimates are based on data for measured concentrations of only two dioxin-like compounds of PCBs
in Hudson River fish, and that these results may underestimate the concentrations and effects of total
dioxin equivalents in these fish. The ERA Addendum addresses the first source of uncertainty by
presenting interspecies differences in sensitivity of early life stages to dioxin-like compounds in
Table B-7 and acknowledging that salmonid species are the most sensitive group thus far tested.
The ERA Addendum accounts for this uncertainty by using two sets of TRVs, one based on data for
salmonids and one based on data for non-salmonids, to bracket the potential range of sensitivity of
Hudson River fish. The ERA Addendum addresses the second source of uncertainty by
acknowledging that concentrations of some dioxin-like compounds may be underestimated in
measurements of Hudson River fish and using a modeling approach to approximate the concentration
of total dioxin equivalents in these fish.
Response to EG-1.9
The review conducted by Dr. Emily Monosson for NOAA is cited m the ERA (and thereby
for the ERA Addendum) as NOAA (1999b). The NOAA report was reviewed during the process
of selection of TRVs for the ERA and ERA Addendum. However, results presented in the NOAA
report and in USEPA's ERA and ERA Addendum are based on different approaches and assertions
about the type of studies that are most appropriate for assessment of risk to Hudson River fish. The
study by NOAA reported measured or estimated concentrations in eggs, larvae, or in liver of adult
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fish that result in adverse reproductive, developmental, or immunotoxic effects. In contrast, USEPA
developed TRVs for total PCBs on the basis of studies that report measured concentrations of total
PCBs in tissue of larval and adult fish which result in adverse effects on survival growth or
reproduction, but not on sublethal immunotoxic or biochemical effects. USEPA believes that effect
levels reported as measured whole-body concentrations are most appropriate for comparison to
concentrations of PCBs in Hudson River fish.
As noted by the commenter, the NOAA (1999) report finds that concentrations of PCBs as
low as 5 ppm (whole body, wet wt.) in larvae can impair survival. This finding is based on studies
by Hansen et al. (1974) and Shimmel et al. (1974). NOAA (1999) reports a measured concentration
from the Hansen et al. study and an estimated concentration in larvae from the Shimmel et al. (1974)
study. Because the factor used by NOAA to estimate the concentration in larvae is highly uncertain
(e.g. NOAA reports ratios ranging from 0.1 to 545 in other fish), the study by Shimmel et al. (1974)
was not used by USEPA to develop TRVs.
The study by Hansen et al. (1974) was selected by USEPA for development of a TRY.
USEPA notes, however, that the concentration measured by Hansen et al. (1974) and reported by
NOAA as 5.1 mg/kg in larvae, was actually measured in eggs. USEPA did not use this concentration
for development of a TRY because no other studies were identified that examined concentrations
of total PCBs in eggs. Rather, USEPA developed a TRY from the effective concentration in adult
tissue that was reported in the same paper (Hansen et al. 1974). Effect concentrations determined
as concentrations in tissue of adult fish are believed to be more directly comparable to PCB
concentrations in adult Hudson River fish than are effect concentrations determined in eggs.
Therefore, USEPA used the NOAEL (1.9 mg/kg) and LOAEL (9.3 mg/kg) effect levels reported in
Hansen et al. (1974) for development of TRVs.
The comment states that adverse effects on adult fish might occur at concentrations
exceeding 12.5 ppm (whole body, wet weight). Actually, NOAA (1999) reports that effects may
occur at greater than 25 ppm in adult liver, a concentration that is expected to be equivalent to a
concentration of 12.5 ppm in fillet, not whole body as stated by the commenter. The NOAA report
compiled concentrations that were measured or estimated in liver of adult fish. Studies that reported
concentrations in liver were not used by USEPA for development of TRVs because uncertainty in
the ratio of concentration in liver to concentration in whole body or fillet was believed to be too
great. As noted in the NOAA report (1999), the liver/muscle ratio varies with a number of factors,
and can range from <1 to 77. USEPA selected studies that reported actual measured concentrations
of PCBs in whole body fish tissue.
Response to EG-1.10
The issues of the TQ approach and comparison to the Clinch River Assessment are addressed
in the responses to EG-1.1, EG-1.5, and EG-1.6 of this Responsiveness Summary.
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Response to EG-1.18a
The study by Bengsston (1980) was replaced by a study by Hansen et al. (1974) that was not
identified for the ERA. As described in the response to comment EF-1.6, the Hansen et al. (1974)
study was selected because it examined reproductive effects of Aroclor 1254, rather than Clophen,
a mixture of PCBs that is similar in chlorine content but which was not used in the United States.
USEPA conducted an extensive review of the available literature on the effects of PCBs on
wildlife species. All of the studies presented in Tables B-4 through B-22 in the ERA Addendum
(USEPA, 1999c) were considered in the development of the TRVs. Only after consideration of all
of the studies were the most appropriate individual studies selected for development of the TRVs.
In deriving the TRVs, in cases for which there is no appropriate information available on the
sensitivity of a receptor of concern, it is conservatively protective to assume that the receptor could
be as sensitive as the most sensitive species tested. However, based on comments received from the
peer reviewers, the sensitivity of wild birds is expected to be less than that of gallinaceous birds,
such as the chicken, which are often used as test species. USEPA is evaluating how best to revise
its selection of TRVs on the basis of the peer reviewers' comments.
Response to EG-1.18b
In regard to benthic invertebrate community endpoints, the ERA Addendum (USEPA, 1999c)
examined future risk to the benthic invertebrate community of the Lower Hudson River and therefore
used sediment guidelines and water quality criteria as measurement endpoints. The ERA (USEPA,
1999b) used 1993 data on macroinvertebrate communities as a measurement endpoint to evaluate
current risk to the Lower Hudson River benthic invertebrate community.
Response to EG-1.19
The studies by Adams et al. (1989, 1990, 1992) are mistakenly listed in Table B-6 as EL-
effect and EL-no effect, meaning that they examined a single effect level rather than a range of doses.
In fact, a range of doses was examined and the NOAEL is not unbounded (Adams et al., 1992).
Adverse effects that could be attributed to PCBs (or other co-occurring contaminants) were observed.
USEPA acknowledges the uncertainty associated with using the unbounded field study by
Westin et al. (1983). However, the study is used to develop TRVs because it was conducted on
striped bass from the Hudson River.
As stated in response to comment EG-1.9, the review conducted by NOAA is cited in the
ERA as NOAA (1999b). The NOAA report was reviewed during the process of selection of TRVs
for the ERA and ERA Addendum. However, results presented NOAA report and in the ERA and
ERA Addendum are based on different approaches and assertions about the type of studies that are
most appropriate for assessment of nsk to Hudson River fish. For example, the study by NOAA
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reported measured or estimated concentrations in eggs, larvae, or in liver of adult fish that result in
adverse reproductive, developmental, or immunotoxic effects. In contrast, USEPA developed TRVs
for total PCBs on the basis of studies that report measured concentrations of total PCBs in tissue of
larval and adult fish that result in adverse effects on survival, growth, or reproduction but not on
sublethal immunotoxic or biochemical effects. In addition, the NOAA report compiled
concentrations that were measured or estimated in liver of adult fish. Studies that reported
concentrations in liver were not used by USEPA for development of TRVs because uncertainty in
the ratio of concentration in liver to concentration in whole body or fillet was believed to be too
great. As noted in the NOAA report (1999), the liver/muscle ratio varies with a number of factors,
and can range from <1 to 77. USEPA selected studies that reported actual measured concentrations
of PCBs in whole body fish tissue.
The NOAA report (1999) applies the TEQ approach for estimating risk to Hudson River fish,
but identifies two areas of uncertainty. First, the NOAA (1999) report notes that the available data
indicate large interspecies differences in early life stage toxicity. Second, the reports notes the risk
estimates are based on data for measured concentrations of only two dioxin-like compounds of PCBs
in Hudson River fish, and that these results may underestimate the concentrations and effects of total
dioxin equivalents in these fish. The ERA Addendum addresses the first source of uncertainty by
presenting interspecies differences in sensitivity of early life stages to dioxin-like compounds in
Table B-7 and acknowledging that salmonid species are the most sensitive group thus far tested.
The ERA Addendum accounts for this uncertainty by using two sets of TRVs,-one based on data for
salmonids and one based on data for non-salmonids, to bracket the potential range of sensitivity of
Hudson River fish. The ERA Addendum addresses the second source of uncertainty by
acknowledging that concentrations of some dioxin-like compounds may be underestimated in
measurements of Hudson River fish and using a modeling approach to approximate the concentration
of total dioxin equivalents in these fish.
Niimi (1996) noted in his review, "These estimates represent the threshold concentrations
that were derived from a limited information base that may not be representative of the more
sensitive species and should be interpreted accordingly." In fact, the Niimi (1996) review does not
include or consider the studies that were found to be most relevant to the assessment of risk to
Hudson River fish. For example, it does not consider the results of the studies by Bengtsson (1980)
or Hansen et al. (1974) (see, response to comment EF-1.6). Therefore, the review by Niimi (1996)
cannot be considered to be a comprehensive overview of the relevant studies that can be used to
assess risk or develop TRVs.
Response to EG-1.20
As the commenter notes, all things being equal among several studies, selection of the
highest NOAEL for the development of TRVs would be appropriate. However, many different types
of studies varying in aspects such as exposure time, exposure route, lexicological endpoint examined
and species examined were reviewed for the ERA and ERA Addendum. Therefore, studies were
evaluated based on many criteria in order to select the most appropriate study. The selected study
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was often lower than other available endpoints, but was not selected solely on that basis. For
example, a chronic study on a sensitive reproductive endpoint may report a lower NOAEL/LOAEL
than a subchronic study that examined adult mortality. In this example, it is appropriate to select the
lower value because it is a more appropriate study.
The ERA Addendum did not use uncertainty factors in an overly conservative fashion. For
example, if a TRY was developed on the basis of a study that was conducted on a very sensitive
species, an additional interspecies uncertainty factor was not applied because it is unlikely that the
receptor of concern would be more sensitive than the highly sensitive species. If, however, a TRY
was developed on the basis of a study that was conducted on a species that was known to be of
intermediate sensitivity, an uncertainty factor of 10 was applied in case the receptor of concern is
more sensitive than the test species.
The rationale for the use of uncertainty factors is documented in the USEPA report, Great
Lakes Water Quality Initiative Technical Support Document for Wildlife Criteria (USEPA, 1995).
This report summarizes several studies that analyzed the variability in acute sensitivity of birds and
mammals to a variety of chemicals. For the effect of an individual chemical on birds, the ratio of
the LC50for the least sensitive species to that of the most sensitive species was usually less than 10.
For mammals, the ratio for the least sensitive species to the most sensitive species was usually less
than 100. These analyses of variability in the acute sensitivity provided support for the use of a
recommended range in interspecies uncertainty factors of 1 to 100. In addition, a smaller set of data
on chronic exposures indicated that an interspecies sensitivity ratio of 100 encompasses 84% of the
chronic data. Therefore, in cases for which no appropriate information is available on the chronic
sensitivity of a receptor of concern, it is conservatively protective to assume that the receptor could
be 10 times more sensitive than the species used to establish the TRY. However, as previously
noted, when the test species that was used to establish the TRY is known to be a highly sensitive
species, the ERA Addendum did not apply an interspecies uncertainty factor in order to estimate the
TRY for the receptor of concern.
Similarly, USEPA (1995) provides support for the conceptual basis for use of a subchronic-
to-chronic uncertainty factor. The report summarizes the results of three studies that examined the
ratio of subchromc-to-chronic toxicity endpoints. The first study reported that 97% of the ratios
were 9 or below. The second study reported that 98% of the ratios were less than 4 and all of the
ratios were less than 7. The third study reported that 90% of the ratios were within a factor of 5.
Therefore, use of an uncertainty factor of 10 in the ERA and ERA Addendum to estimate chronic
toxicity endpoints from sub-chronic studies is a conservatively protective approach.
Tree Swallow
Although tree swallows inhabiting areas along the Lower Hudson River are likely to be less
exposed to PCBs than those along the Upper Hudson River, the USFWS findings (summarized in
Secord and McCarty, 1997 and McCarty and Secord, 1999a,1999b) are relevant for the Lower
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Hudson River because they can be used to estimate a NOAEL for field exposure of tree swallows
to PCBs.
As stated in the ERA Addendum, USEPA agrees that the data by McCarty and Secord do not
demonstrate a consistent relationship between exposure to PCBs and adverse reproductive effects
in the tree swallow. Although significant adverse effects on reproduction were observed in the first
year of the study, significant adverse effects on reproduction were not observed in the second year
of the study. Reproductive success in the first year may have been influenced by the large number
of young females that typically inhabit nest boxes in the first year that the boxes are placed in the
field. These data were therefore used to establish a NOAEL, but not LOAEL TRV (see. USEPA,
1999c, Appendix B, p. 26).
Mallard
Five studies were identified that examined effects of PCBs on the mallard. The study by Hill
et al. (1975) was not selected for development of TRVs for exposure of mallards to PCBs because
it examined mortality as an endpoint, which is not expected to be as sensitive an endpoint as growth
and reproduction. The studies by Riseborough and Anderson (1975), Custer and Heinz (1980), and
Heath et al. (1972) found no effects on various reproductive endpoints based on exposure to a single
dose (40 ppm, 25 ppm, and 25 ppm in diet, respectively). Haseltine and Prouty (1980) observed no
adverse effects on reproductive endpoints after a 12-week exposure to 150 ppm Aroclor 1242 in
food, but did observe significantly reduced weight gain in adults. Therefore, the study by Haseltine
and Prouty (1980) was selected as the most appropriate study because it reports a LOAEL on an
ecologically relevant endpoint from which a NOAEL can be estimated. Because only a single dose
was tested, a LOAEL to NOAEL uncertainty factor of 10 was applied to estimate a NOAEL from
this study. The study was conducted over a 12 week period, so a sub-chronic to chronic uncertainty
factor was not applied.
Based on the results of Haseltine and Prouty (1980) on growth:
The LOAEL TRV for growth effects would be: 16 mg/kg/day
The NOAEL TRV for growth effects would be: 1.6 mg/kg/day.
Great Blue Heron
The study by Speich et al. (1992) was designed to examine eggshell thinning rather than the
more sensitive endpoint of egg mortality. The author reports, "In this study, we found no evidence
of current reproductive failure, high incidence of eggshell breakage, eggshell flaking, or low hatching
success (S.M. Speich, unpubl. field notes)." However, the study provides no data on egg mortality,
other than this reference to unpublished field notes. Unpublished field results that have not tested
statistically are not considered an appropriate basis for the development of TRVs.
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The description of the development of the TRVs for TEQs in eggs of the great blue heron is
revised to better explain how the TRY was developed. The description is revised to note that the
TRY was developed on the basis of both the study by Sanderson et al. (1994) and the study by Hart
et al. (1991). As noted, Sanderson et al. (1994) do not present data on reduced growth rate.
Sanderson et al. (1994) presents data on the concentration of TEQs in eggs that were collected from
a highly contaminated site (Crofton) and a less contaminated site (Vancouver) in 1988. Hart et al.
(1991) report that the yolk-free body weights of chicks collected in 1998 from Crofton were
significantly different from a reference site, but that the weights of chicks from Vancouver were not
different from the reference site. Therefore, the data from both Hart et al. (1991) and Sanderson et
al. (1994) were used in development of the TRVs for the great blue heron:
The LOAEL TRY for the great blue heron is 0.5 ug TEQs/kg egg.
The NOAEL TRY for the great blue heron is 0.3 ug TEQs/kg egg.
Belted Kingfisher
Taxonomic similarity is considered to be a better predictor of sensitivity to PCBs and dioxin-
like compounds than is similarity m feeding habits. Because no information is available on the
sensitivity of the belted kingfisher or for a species in the same family as the belted kingfisher, the
assessment conservatively assumes that the kingfisher could be as sensitive as the most sensitive
species tested.
Bald Eagle
USEPA agrees that the data by Wiemeyer et al. (1993) do not support the development of a
NOAEL TRY of 3.0 mg/kg for the bald eagle. However, USEPA does not agree with the
commenter's assertion that because mean five-year production was not significantly reduced for the
residue interval ranging from 5.6-<13 mg PCBs/kg, a NOAEL of 13 mg/kg is appropriate. It would
be more appropriate to take the average value of the data in the 5.6-<13 mg/kg interval as a measure
of the average concentration for which production was not significantly impacted as compared to
higher concentrations. However, those data are not reported by Wiemeyer et al. (1993). As an
alternative, USEPA is using the average PCB concentration in eggs from successful nests (5.5
mg/kg), which was shown to be sigmficandy lower than the concentration measured in unsuccessful
nests (8.7 mg/kg) (Wiemeyer et al. 1993, p. 224), as the NOAEL TRY for bald eagles.
Based on the study by Wiemeyer et al. (1993):
The NOAEL TRY for the bald eagle is 5.5 mg/kg egg.
As noted in the ERA Addendum, USEPA agrees that because of the presence of co-occurring
contaminants, adverse effects observed in field studies cannot be attributed solely to the presence
of PCBs. Therefore, USEPA did not use any field studies to establish LOAELs, the concentrations
or doses at which adverse effects are expected to occur. USEPA did, however, use field studies to
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establish NOAELs, the concentration or doses below which adverse effects are not expected to
occur. USEPA acknowledges that because of the confounding influence of co-occurring
contaminants, that actual NOAEL TRVs could be higher than those observed in field-based studies.
The study by Elliott et al. (1996) was included in USEPA's review (see. ERA Addendum,
Table B-14). USEPA notes, however, that the study by Wiemeyer et al. (1993) was determined to
be a more appropriate study for the development of a TRY for the bald eagle because the Wiemeyer
et al. (1993) study examined numerous eggs in 15 states over a period of many years, whereas the
Elliott et al. (1996) study examined only 16 eggs from a contaminated area and eight eggs from
reference areas.
USEPA inadvertently excluded the study by Elliott et al. (1996) from Table B-16, the
compilation of field studies on the effects of dioxin-like contaminants on bird eggs. The study by
Elliott et al. (1996) reports data for TEQ in the yolk sac of the bald eagle egg. The authors do report
a concentration of TEQs of 210 ng/kg wet weight in eggs for the Powell River, a contaminated site
with a concentration that is slightly less than the other contaminated site (East Vancouver Island).
If the concentration of TEQs at the East Vancouver Island site is estimated as 13,000 ng TEQs/kg
lipid, the estimated wet weight concentration would be approximately 217 ng/kg ww for East
Vancouver eggs. Because no significant difference was observed between the average hatching rate
of the eggs collected from the pulp mill sites (East Vancouver and Powell River) and the non-pulp
mill sites, a NOAEL based on the average egg concentration of these two sites could be developed.
Based on the results of Elliott et al. (1996), an average field based NOAEL of 214 ng TEQs/kg ww
would be established for the bald eagle. This is lower than the value of 400 ng/kg ww that was
suggested by the commenter, the derivation of which is unclear.
The field based NOAEL for the bald eagle eggs would be 0.214 ug/kg egg.
The study by Donaldson et al. (1999) was not available when the literature search was
conducted for the ERA Addendum. If the study were available, USEPA would still have selected
the study by Wiemeyer et al. (1993) for development of TRVs because this study examined many
more eggs. Donaldson et al. (1999) examined concentrations of PCBs in 6 eggs from a contaminated
site, whereas Wiemeyer et al. (1993) examined numerous eggs in 15 states over a period of many
years.
Response to EG-1.21
Mammals
USEPA acknowledges that limited data are available to assess the potential for risk to the
little brown bat or the raccoon and that laboratory studies conducted on rats must be used to make
conservative estimates risk to these organisms. However, as noted in Table B-27 of the ERA
Addendum, field studies on the mink, rather than laboratory studies, are used to develop final TRVs
for the mink and the otter.
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Mink
USEPA agrees that a LOAEL should not be established from the Tillitt et al. (1996) field
study. The revised risk estimates remove this comparison.
USEPA does not concur with the assertion of Sample et al. (1996) that an exposure over a
longer period would not result in a lower effective dose. As described in Section B.2.1, USEPA's
approach follows the approach used by the Great Lakes Water Quality Initiative (USEPA, 1995).
This approach uses uncertainty factors to account for the well-recognized observation that
subchronic toxicity studies may be of insufficient length to measure adverse effects that would be
observed in chronic tests of longer duration.
USEPA examined the available data on body burdens of PCBs and dioxin-like compounds
in mink of the Hudson River area and found that insufficient data were available to assess risk on
this basis. Therefore, the study by Leonards et al. (1995) was not used to develop TRVs for the
assessment.
River Otter
The paper by Harding et al. (1999) was not available when the literature search was done for
the ERA and ERA Addendum. However, the study would not be selected to develop a TRY because
it reports exposure on the basis of concentration in liver and reports effects on baculum length in
juvenile males.
Response to EG-1.22
The TQ approach isolates the effects of PCBs versus other confounding influences from
field-based studies. The TQ approach suggests the potential for risk attributable to PCBs alone. As
explained in the response to EG-20, the ERA Addendum did not rely on overly conservative
application of uncertainty factors in deriving TRVs and evaluated available literature in deriving
TRVs.
4.1 Selection of Measures of Effects
No significant comments were received on Section 4.1.
4.1.1 Methodology Used to Derive TRVs
No significant comments were received on Section 4.1.1.
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4.1.2 Selection of TRVs
Response to EF-1.21
The study by USAGE (1988), which examined field-collected sediments, was inadvertently
excluded from Appendix B. Results from the USAGE (1988) study, the lab studies by Hansen et al.
(1974) and Bengtsson (1980), and field studies by Adams et al. (1989, 1990, 1992) yield similar
toxicity values (see. USEPA, 1999c, Table B-5), thereby providing further weight of evidence to
support the selection of the TRVs.
The field-based NOAEL (0.5 mg/kg) reported in the ERA Addendum for pumpkinseed and
largemouth bass was based on a reproductive endpoint (Adams et al. 1989,1990,1992). USEPA
is revising the ERA Addendum to use the NOAEL of 0.3 mg/kg based on a growth endpoint from
Adams et al. (1989,1990,1992). It should be noted that Adams et al. (1989,1990,1992) examined
other endpoints that occurred at concentrations below 0.5 mg/kg. Adverse effects were associated
with DNA integrity, detoxification enzymes, lipid metabolism, community structure and histological
indices.
Response to EF-1.22
For the TEQ analysis, BZ#126 was used at the detection limit to compensate for not having
quantitated BZ#81, as described in the ERA (see. USEPA, 1999b, pp. 38-40). An analysis
evaluating the proportion of TEQ congeners in USEPA Phase 2 data and USFWS tree swallow data
showed that the proportion of BZ#126 in the Phase 2 dataset was approximately equal to the sum
of the BZ#126 and BZ#81 in the USFWS dataset (see. USEPA, 1999b, Appendix J). This approach
does not produce an overly conservative estimate of TEQ risks because they are dominated by
BZ#126 (and presumably BZ#81), and thus may be too high by at most a factor of two. This is a
relatively small margin of error considering that the calculated risk levels exceed USEPA's levels
of concern by orders of magnitude.
Direct water column exposures represent a tiny fraction of the overall daily dose for all
receptors, thus, the issue of detection limits is far less important for this medium.
5.0 RISK CHARACTERIZATION
Response to EL-1.18
The responses to comments on the risk characterization presented in the ERA can be found
in the Responsiveness Summary for the ERA (see. USEPA, 2000b, Section 5.0 Risk
Characterization).
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Response to EG-1.17
The condition of the ecological resources of the Lower Hudson River in the relation to PCBs
cannot be evaluated simply by examining trends over the last 30 years (see, response to EG-1.1).
Although macroinvertebrate communities in the Hudson River have improved since the 1970s, much
of the change can be attributed to improvements in treatment of municipal and industrial wastes
rather than to a direct response to lower PCB concentrations (see. ERA Responsiveness Summary
[USEPA, 2000b] responses to EG-1.7 and EG-1.34).
Fish population trends have also been addressed in the Responsiveness Summary for the
ERA (see, responses to comments EG-1.9 and EG1-34 in USEPA, 2000b). The kinds of effects due
to PCBs expected in the field include reduced fecundity, decreased hatching success, and similar
kinds of reproductive impairment indicators, which are often difficult to discern, particularly against
the background of the fishing ban.
Tree swallows are present throughout the Lower Hudson River Valley (no adverse effects
were predicted for the tree swallow). Waterfowl are abundant, which is expected given the high
habitat quality of many areas of the Hudson River (see. Section 2.6.5 of ERA Addendum). The
presence of one breeding colony of great blue herons does not indicate that they are breeding
throughout the Lower Hudson River Valley. Similarly, the mixed success of bald eagle nests along
the Hudson River in the last several years does not indicate that the bald eagle is re-established along
the Hudson River. Certainly, it is encouraging to see some successful nesting, but it is too early to
call the Hudson River population re-established. NYSDEC has been collecting eagle serum, prey
and unhatched eggs for several years to evaluate contaminant loads throughout the eagles ecosystem
(Nye, 2000). Preliminary PCB results from only two samples are high enough to be of concern, and
more data on PCB concentrations in birds along the Hudson River are expected to be available in
late 2000/early 2001 (Secord, 2000).
The abundance of raccoons along the Hudson River is addressed in the response to EL-1.22.
Although mink and river otter are present along the Hudson River, their numbers are generally low.
Preliminary results from a NYSDEC study (Mayack, 1999) indicate that PCBs may adversely affect
litter size and possibly kit survival of river otter in the Hudson River.
5.1 Evaluation of Assessment Endpoint: Benthic Community Structure as a Food
Source for Local Fish and Wildlife
No significant comments were received on Section 5.1.
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5.1.1 Do Modeled PCB Sediment Concentrations Exceed Appropriate
Criteria and/or Guidelines for the Protection of Aquatic Life and
Wildlife?
No significant comments were received on Section 5.1.1.
5.1.1.1 Measurement Endpoint: Comparisons of Modeled
Sediment Concentrations to Guidelines
Response to EF-1.23
Tables 3-2 and 3-3 are revised to read, 'Tables 3-6 and 3-7." Although the predicted
concentrations of PCBs in sediment consistently underestimate the mean concentrations measured
at RMs 152,113,90 and 50, they do fall within the range of the sampled concentrations for all RMs
except for RM 47 (see. Figure 3-7). In addition, the predicted mean sediment concentrations are
based on dicholoro to hexachloro homologues and therefore are expected to be slightly lower than
total PCB concentrations. Although average TOC values were generally greater than the TOC of
2.5% used by Farley et al. (1999), a TOC of 2.5% was used to provide consistency in the model (see
response to EF-1.16).
Response to EF-1.24
The first complete sentence is revised to read, "Forecast sediment concentrations exceed the
NYSDEC benthic aquatic life chronic toxicity criterion at RMs 152 and 113 for the duration of the
modeling period based on the 95% UCL."
Response to EF-1.25
The correction of the organic carbon-normalized SEL (Persaud et al., 1993) from 1.3 mg/kg
to 13 mg/kg in Table 5-1 and the associated text is noted. Ratios in Table 5-1 were calculated using
the correct organic carbon-normalized SEL of 13 mg/kg.
5.1.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria
and/or Guidelines for the Protection of Aquatic Life and Wildlife?
No significant comments were received on Section 5.1.2.
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5.1.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations of PCBs to Criteria
Response to ES-1.1 and EF-1.26
The change in the NYSDEC surface water standard for the protection of wildlife from 0.001
Mg/L total PCBs to 1.2 x 10"4 |ag/L in 1998 (6 NYCRR Part 703) is noted. Table 5-2 (see. Section
HI) is revised accordingly. Use of the earlier standard underestimated the ratio of predicted whole
water concentrations to the wildlife standard by an order of magnitude.
5.2 Evaluation of Assessment Endpoint: Protection and Maintenance (i.e.,
Survival, Growth, and Reproduction) of Local Fish Populations
No significant comments were received on Section 5.2.
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?
Response to EF-1.27
The assumption that measurements of young-of-year spottail shiner and age 1 pumpkinseed
are equivalent to concentrations in mature adults may underestimate concentrations of PCBs in those
species and animals that feed on them. The TRY used in the ERA Addendum for the spottail shiner
on a NOAEL basis is 1.6 mg/kg, not 15 mg/kg. Therefore, if comparisons are made between field
and laboratory based NOAELs, the difference is reduced to three-fold, as stated by the commenter.
5.2.1.1 Measurement Endpoint: Comparison of Modeled Total
PCB Fish Body Burdens to Toxicity Reference Values for
Forage Fish
No significant comments were received on Section 5.2.1.1.
5.2.1.2 Measurement Endpoint: Comparison of Modeled PCB
TEQ Fish Body Burdens to Toxicity Reference Values for
Forage Fish
Response to EF-1.28
The FISHRAND model generates lipid-normalized (and wet weight) fillet concentrations.
The model does not explicitly model an egg concentration, which was developed based on the
correlation between lipid normalized egg and whole body PCB concentrations (Niimi, 1983).
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5.2.1.3 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for Brown
Bullhead
Response to EF-1.29
RM 133 is revised to read 113. The first sentence of Section 5.2.1.3 is revised to read, "As
literature-derived TRVs were based on whole body concentration studies, the fish fillets were
converted to whole body for direct comparison."
Response to EF-1.30
The comparison of the concentrations of TCDD in fish tissue associated with low and high
levels of risk to Great Lakes receptors (EPA, 1993) is provided below. The study used by US EPA
(1993) to establish concentrations of TCDD associated with low risk to piscivorous fish, Walker et
al. 1992, is presented in Table B-7. This study found that for waterbome exposures, a residue of
0.034 ug TCDD/kg ww in lake trout eggs (estimated by USEPA to be about 0.050 ug TCDD/kg ww
in adult fish) did not exhibit significant effects relative to controls. The study used by USEPA to
establish the concentration of TCDD associated with high risk to piscivorous fish is also presented
in Table B-7. Walker et al. (1992) reported effects on fry survival at 0.055 ug TCDD/kg ww in trout
eggs (estimated by USEPA to be about 0.075 ug/kg ww in parent fish). The ERA Addendum reports
these egg concentrations in Table B-7 as both wet weight and as lipid normalized concentrations
(0.43 and 0.7 ug TEQ/kg lipid, respectively). However, in the Responsiveness Summary for the
ERA, USEPA selected a more recent study for the development of TRVs, Walker et al. (1994). This
study is selected for development of TRVs for salmonids because it reports a NOAEL that was
measured using a different and more realistic exposure route, maternal transfer of TCDD to eggs.
This study reported a NOAEL of 0.023 ug TCDD/kg ww egg (0.29 ug /kg lipid) and a LOAEL of
0.05 ug TCDD/kg ww (0.6 ug/kg lipid). Thus, the NOAEL TRV developed for the ERA Addendum
is slightly lower than that developed in USEPA (1993) and the LOAEL TRV is similar.
For mammals and birds, the ERA Addendum estimated risk on the basis of dietary dose (mg
TCDD/kg body weight/day), rather than as concentration in diet (mg TCDD/ kg fish). However, the
studies used in the USEPA (1993) report to develop risk-based concentrations in prey were the same
studies that were used to develop TRVs for some receptors in the present risk assessment. These are
Murray et al. (1979) and Nosek et al (1992), which are included in Tables B-ll and B-18. The
USEPA (1993) approach assumed that avian and mammalian receptors consume 100% fish, whereas
the ERA Addendum assumes that most receptors consume a variety of prey types, which include fish
in most cases. Therefore, although the same studies were used in some cases, different assumptions
were used about the types of prey consumed by Hudson River receptors in comparison to Great
Lakes receptors, and conclusions about protective levels in fish prey items are not directly
comparable.
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5.2.1.4 Measurement Endpoint: Comparison of Modeled TEQ Basis
Fish Body Burdens to Toxicity Reference Values for Brown
Bullhead
No significant comments were received on Section 5.2.1.4.
5.2.1.5 Measurement Endpoint: Comparison of Modeled Total PCB
Fish Body Burdens to Toxicity Reference Values for White
and Yellow Perch
Response to EF-1.31
The rationale for not applying interspecies uncertainty factors to field studies is provided in
the response to comment EF-1.7. Because white perch and yellow perch are not in the same
taxonomic family, if the NOAEL TRY for the white perch were used to develop a NOAEL TRY for
the yellow perch, an interspecies uncertainty factor of 10 would be applied. In that hypothetical case,
the NOAEL TRY for the yellow perch would be 0.31 mg PCBs/kg tissue, rather than the 0.16 mg/kg
laboratory-based NOAEL-based TRY, and the NOAEL-based toxicity quotients would be
approximately half of what was reported in the ERA Addendum.
5.2.1.6 Measurement Endpoint: Comparison of Modeled TEQ Basis
Body Burdens to Toxicity Reference Values for White and
Yellow Perch
No significant comments were received on Section 5.2.1.6.
5.2.1.7 Measurement Endpoint: Comparison of Modeled Tri+ PCB Fish Body
Burdens to Toxicity Reference Values for Large-mouth Bass
No significant comments were received on Section 5.2.1.7.
5.2.1.8 Measurement Endpoint: Comparison of Modeled TEQ Based Fish Body
Burdens to Toxicity Reference Values for Large- mouth Bass
No significant comments were received on Section 5.2.1.8.
5.2.1.9 Measurement Endpoint: Comparison of Modeled Tri+ PCB
Fish Body Burdens to Toxicity Reference Values for Striped Bass
Response to EL-1.19
Striped bass are known to occur throughout the upper portion of the Lower Hudson River
(NOAA, 1985). Concentrations of PCBs in striped bass were related to concentrations in largemouth
64 TAMS/MCA
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bass in the absence of explicitly modeled results from either the Farley or FISHRAND models.
There are no monitoring data available for largemouth bass at RMs 90 and 50; thus, ratios could not
be estimated for these locations. Although the Farley model provides results for Food Web Region
2, this area is a much larger area than that used for the remaining fish species. Note also that the
ERA provides risk estimates based on observed concentrations in striped bass, and that these results
suggest risk to the striped bass at some locations in the Lower Hudson River (see, Table 5-36
[unchanged] in ERA and ERA Responsiveness Summary [USEPA, 1999b and 2000b]).
5.2.1.10 Measurement Endpoint: Comparison of Modeled TEQ Based
Fish Body Burdens to Toxicity Reference Values for Striped
Bass
No significant comments were received on Section 5.2.1.10.
5.2.2 Do Modeled PCB Water Concentrations Exceed Appropriate Criteria
and/or Guidelines for the Protection of Aquatic Life and Wildlife?
No significant comments were received on Section 5.2.2.
5.2.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations of PCBs to Criteria
No significant comments were received on Section 5.2.2.1.
5.2.3 Do Modeled PCB Sediment Concentrations Exceed Appropriate
Criteria and/or Guidelines for the Protection of Aquatic Life and
Wildlife?
No significant comments were received on Section 5.2.3.
5.2.3.1 Measurement Endpoint: Comparisons of Modeled Sediment
Concentrations to Guidelines
No significant comments were received on Section 5.2.3.1.
5.2.4 What Do the Available Field-Based Observations Suggest About the
Health of Local Fish Populations?
No significant comments were received on Section 5.2.4.
65 TAMS/MCA
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5.2.4.1 Measurement Endpoint: Evidence from Field Studies
Response to EL-1.20
As discussed in the Responsiveness Summary for the ERA (see. USEPA, 2000b, responses
to EG-1.9 and EG-1.34, EG-1.38, EP-1.1, EP-2.10, and EL-1.46), the presence of healthy
populations does not indicate that PCBs have no adverse effect on local fish and wildlife.
Improvements in water quality and the fishing ban have undoubtably assisted the recovery and
maintenance of many species. The shortnose sturgeon in particular has benefitted from being listed
as an endangered species and the fishing ban in the Hudson River. These factors have allowed the
population of Hudson River shortnose sturgeon to increase despite any potential adverse effects from
PCB exposure.
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)
No significant comments -were received on Section 5.3.
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?
No significant comments were received on Section 5.3.1.
5.3.1.1 Measurement Endpoint: Modeled Dietary Doses on a Tri+ PCB
Basis to Insectivorous Birds (Tree Swallow)
No significant comments were received on Section 5.3.1.1.
5.3.1.2 Measurement Endpoint: Predicted Egg Concentrations on a
Tri+ PCB Basis to Insectivorous Birds (Tree Swallow)
No significant comments were received on Section 5.3.1.2.
5.3.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed on a TEQ Basis to Insectivorous Birds (Tree
Swallow)
No significant comments were received on Section 5.3.1.3.
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5.3.1.4 Measurement Endpoint: Predicted Egg Concentrations
Expressed on a TEQ Basis to Insectivorous Birds (Tree
Swallow)
No significant comments were received on Section 5.3.1.4.
5.3.2 Do Modeled Water Concentrations Exceed Criteria for Protection
of Wildlife?
No significant comments were received on Section 5.3.2.
5.3.2.1 Measurement Endpoint: Comparison of Modeled Water
Column Concentrations to Criteria for the Protection of
Wildlife
No significant comments were received on Section 5.3.2.1.
5.3.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Insectivorous Bird Populations?
No significant comments were received on Section 5.3.3.
5.3.3.1 Measurement Endpoint: Evidence from Field Studies
No significant comments were received on Section 5.3.3.1.
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)
Response to EF-1.32
The comparison of current trends in bird usage to historical usage (e.g., prior to GE's use of
PCBs at its two Hudson River facilities) would be of limited use in assessing ecological risk due to
PCBs due to the changes in habitat use that have occurred along the Hudson River (and, for
migratory species, other areas as well) over the last 50 years. The complexity of the ecosystem and
number of variables affecting bird usage does not allow direct effects to be determined based on PCB
concentrations.
5.4.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Waterfowl and
Egg Concentrations Exceed Benchmarks for Adverse Effects on
Reproduction?
No significant comments were received on Section 5.4.1.
67 TAMS/MCA
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5.4.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ PCBs
to Waterfowl (Mallard)
No significant comments were received on Section 5.4.1.1.
5.4.1.2 Measurement Endpoint: Predicted Egg Concentrations of Tri+
PCBs to Waterfowl (Mallard)
No significant comments were received on Section 5.4.1.2.
5.4.1.3 Measurement Endpoint: Modeled Dietary Doses of TEQ-Based
PCBs to Waterfowl (Mallard)
No significant comments were received on Section 5.4.1.3.
5.4.1.4 Measurement Endpoint: Predicted Egg Concentrations of
TEQ-Based PCBs to Waterfowl (Mallard)
No significant comments were received on Section 5.4.1.4.
5.4.2 Do Modeled PCB Water Concentrations Exceed Criteria for the
Protection of Wildlife?
No significant comments were received on Section 5.4.2.
5.4.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria
No significant comments were received on Section 5.4.2.1.
5.4.3 What Do the Available Field-Based Observations Suggest About the
Health of Lower Hudson River Waterfowl Populations?
Response to EL-1.21
The Christmas bird count species records are known to contain errors (Cornell University,
1999). Therefore, the database cannot be used for scientific studies until it has been reviewed and
corrected. In addition, count efforts (e.g., number of participants, skill level) are not consistent
between years or count circles. Based on these factors, it is difficult to discern any meaningful trends
in the data without intensive data analyses. In general, the greatest number of species was observed
near the mouth of the Hudson River, which is consistent with the locations of various habitats.
68 TAMS/MCA
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5.4.3.1 Measurement Endpoint: Observational Studies
No significant comments were received on Section 5.4.3.1.
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)
No significant comments were received on Section 5.5.
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?
No significant comments were received on Section 5.5.1.
5.5.1.1 Measurement Endpoint: Modeled Dietary Doses of Total PCBs
for Piscivorous Birds (Belted Kingfisher, Great Blue Heron,
Bald Eagle)
No significant comments were received on Section 5.5.1.1.
5.5.1.2 Measurement Endpoint: Predicted Egg Concentrations
Expressed as Tri+ to Piscivorous Birds (Eagle, Great Blue
Heron, Kingfisher)
No significant comments were received on Section 5.5.1.2.
5.5.1.3 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed as TEQs to Piscivorous Birds (Belted Kingfisher,
Great Blue Heron, Bald Eagle)
No significant comments were received on Section 5.5.1.3.
5.5.1.4 Measurement Endpoint: Modeled Dietary Doses of PCBs
Expressed as TEQs to Piscivorous Birds (Belted Kingfisher,
Great Blue Heron, Bald Eagle)
No significant comments were received on Section 5.5.1.4.
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5.5.2 Do Modeled Water Concentrations Exceed Criteria for the Protection of
Wildlife?
No significant comments were received on Section 5.5.2.
5.5.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria
No significant comments were received on Section 5.5.2.1.
5.5.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Piscivorous Bird Populations?
No significant comments were received on Section 5.5.3.
5.5.3.1 Measurement Endpoint: Observational Studies
No significant comments were received on Section 5.5.3.1.
5.6 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Insectivorous Mammal Populations (as represented by
the Little Brown Bat)
No significant comments were received on Section 5.6
5.6.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Insectivorous
Mammalian Receptors Exceed Benchmarks for Adverse Effects on
Reproduction?
No significant comments were received on Section 5.6.1.
5.6.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+ to
Insectivorous Mammalian Receptors (Little Brown Bat)
No significant comments were received on Section 5.6.1.1.
5.6.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Insectivorous Mammalian Receptors (Little Brown
Bat)
No significant comments were received on Section 5.6.1.2.
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5.6.2 Do Modeled Water Concentrations Exceed Criteria for Protection of
Wildlife?
No significant comments were received on Section 5.6.2.
5.6.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife
No significant comments were received on Section 5.6.2.1.
5.6.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Insectivorous Mammalian Populations?
No significant comments were received on Section 5.6.3.
5.6.3.1 Measurement Endpoint: Observational Studies
No significant comments were received on Section 5.6.3.1.
5.7 Evaluation of Assessment Endpoint: Protection (i.e., Survival and
Reproduction) of Local Omnivorous Mammal Populations (as represented by
the Raccoon)
No significant comments were received on Section 5.7.
5.7.1 Do Modeled Total and TEQ-Based PCB Dietary Doses Omnivorous
Mammalian Receptors Exceed Benchmarks for Adverse Effects on
Reproduction?
No significant comments were received on Section 5.7.1.
5.7.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Omnivorous Mammalian Receptors (Raccoon)
No significant comments were received on Section 5.7.1.1.
5.7.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Omnivorous Mammalian Receptors (Raccoon)
No significant comments were received on Section 5.7.1.2.
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5.7.2 Do Modeled Water Concentrations Exceed Criteria for Protection of
Wildlife?
No significant comments were received on Section 5.7.2.
5.7.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife
No significant comments were received on Section 5.7.2.1.
5.7.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Omnivorous Mammalian Populations?
No significant comments were received on Section 5.7.3.
5.7.3.1 Measurement Endpoint: Observational Studies
Response to EL-1.22
The ERA Addendum focuses on fish and wildlife found along the Hudson River. The
raccoon was selected to represent omnivorous mammal populations living near the Hudson River.
Although a large proportion of the raccoon population in the Lower Hudson River obtains food from
sources other than the Hudson River, those individuals using the Hudson River as their primary food
source may experience adverse effects.
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)
No significant comments were received on Section 5.8.
5.8.1 Do Modeled Total and TEQ-Based PCB Dietary Doses to Piscivorous
Mammalian Receptors Exceed Benchmarks for Adverse Effects on
Reproduction?
No significant comments were received on Section 5.8.1.
5.8.1.1 Measurement Endpoint: Modeled Dietary Doses of Tri+
to Piscivorous Mammalian Receptors (Mink, River Otter)
No significant comments were received on Section 5.8.1.1.
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5.8.1.2 Measurement Endpoint: Modeled Dietary Doses on a TEQ
Basis to Piscivorous Mammalian Receptors (Mink, River Otter)
No significant comments were received on Section 5.8.1.2.
5.8.2 Do Modeled Water Concentrations Exceed Criteria for the Protection of
Piscivorous Mammals?
No significant comments were received on Section 5.8.2.
5.8.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife
No significant comments were received on Section 5.8.2.1.
5.8.3 What Do the Available Field-Based Observations Suggest About the
Health of Local Mammalian Populations?
No significant comments were received on Section 5.8.3.
5.8.3.1 Measurement Endpoint: Observational Studies
No significant comments were received on Section 5.8.3.1.
5.9 Evaluation of Assessment Endpoint: Protection of Threatened and
Endangered Species
No significant comments were received on Section 5.9.
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?
No significant comments were received on Section 5.9.1.
5.9.1.1 Measurement Endpoint: Inferences Regarding Shortnose
Sturgeon Population
No significant comments were received on Section 5.9.1.1.
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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?
No significant comments were received on Section 5.9.2.
5.9.2.1 Measurement Endpoint: Inferences Regarding Bald Eagle and
Other Threatened or Endangered Species Populations
No significant comments were received on Section 5.9.2.1.
5.9.3 Do Modeled Water Concentrations Exceed Criteria for the Protection of
Wildlife?
No significant comments were received on Section 5.9.3.
5.9.3.1 Measurement Endpoint: Comparisons of Modeled Water
Concentrations to Criteria for the Protection of Wildlife
No significant comments were received on Section 5.9.3.1.
5.9.4 Do Modeled Sediment Concentrations Exceed Guidelines for the
Protection of Aquatic Health?
No significant comments were received on Section 5.9.4.
5.9.4.1 Measurement Endpoint: Comparisons of Modeled
Sediment Concentrations to Guidelines
No significant comments were received on Section 5.9.4.1.
5.9.5 What Do the Available Field-Based Observations Suggest About the
Health of Local Threatened or Endangered Fish and Wildlife Species
Populations?
No significant comments were received on Section 5.9.5.
5.9.5.1 Measurement Endpoint: Observational Studies
No significant comments were received on Section 5.9.5.1.
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5.10 Evaluation of Assessment Endpoint: Protection of Significant Habitats
No significant comments were received on Section 5.10.
5.10.1 Do Modeled Total and TEQ-Based PCB Body Burdens/Egg
Concentrations in Receptors Found in Significant Habitats Exceed
Bench-marks for Adverse Effects on Reproduction?
No significant comments were received on Section 5.10.1.
5.10.1.1 Measurement Endpoint: Inferences Regarding Receptor
Populations
No significant comments were received on Section 5.10.1.1.
5.10.2 Do Modeled Water Column Concentrations Exceed Criteria for the
Protection of Aquatic Wildlife?
No significant comments were received on Section 5.10.2.
5.10.2.1 Measurement Endpoint: Comparison of Modeled Water
Concentrations to Criteria for the Protection of Wildlife
No significant comments were received on Section 5.10.2.1.
5.10.3 Do Modeled Sediment Concentrations Exceed Guidelines for the
Protection of Aquatic Health?
No significant comments were received on Section 5.10.3.
5.10.3.1 Measurement Endpoint: Comparison of Modeled Sediment
Concentrations to Guidelines for the Protection of Aquatic
Health
No significant comments were received on Section 5.10.3.1.
5.10.4 What Do the Available Field-Based Observations Suggest About the
Health of Significant Habitat Populations?
No significant comments were received on Section 5.10.4.
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5.10.4.1 Measurement Endpoint: Observational Studies
No significant comments were received on Section 5.10.4.1.
6.0 UNCERTAINTY ANALYSIS
No significant comments were received on Section 6.0.
6.1 Conceptual Model Uncertainties
No significant comments were received on Section 6.1.
6.2 Toxicological Uncertainties
Response to EF-1.33
The ERA Addendum does not attempt to examine effects from congeners that have different
mechanisms of action from the dioxin-like congeners because much less data are available on the
non-dioxin effects. The effect of those congeners on risk is unknown.
6.3 Exposure and Modeling Uncertainties
No significant comments were received on Section 6.3.
6.3.1 Natural Variation and Parameter Error
No significant comments were received on Section 6.3.1.
6.3.2 Model Error
No significant comments were received on Section 6.3.2.
6.3.2.1 Uncertainty in the Farley Model
Response to EF-1.34
The goal of the fate, transport and bioaccumulation models is to capture the general, long-
term trend of PCBs in water, sediment, and fish. Capturing the year-to-year variability is not a goal
of the ERA Addendum or the Farley modeling effort. The Phase 2 sediment data shown in Figure
3-7 of the ERA Addendum is for the 0-5 cm layer. The majority of the data falls between the
modeled results for the 0-2.5 cm and 2.5-5 cm layers, capturing the trend in the data with the
exception of RM 47, which falls above the modeled data. The average of the sediment sample data
would fall near the average of the modeled results from the 0-2.5 cm and 2.5-5 cm layers. This is
76 TAMS/MCA
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good agreement, particularly because the sediment data are not spatially representative of the Lower
Hudson River sediments and show the heterogeneity of the sediments. This heterogeneity results
from the significant variability in deposition history for sections of the river. The Phase 2 high
resolution cores show a fourfold decline between the Albany Turning basin and Lents Cove (see
Figure 3-60 of the DEIR, USEPA, 1997). The cesium-normalized PCB concentrations show a three-
fold decline in the Lower Hudson River in Figure 3-64 of the DEIR (USEPA, 1997).
Response to EF-1.35
USEPA agrees that there is uncertainty associated with changes in the upstream boundary
conditions and potential releases from the remnant deposits. This is addressed in the response to
' comment EF-1.10.
6.3.2.2 Uncertainty in FISHRAND Model Predictions
Response to EF-1.36
The sensitivity analyses presented in the RBMR (USEPA, 2000a) addressed parameters in
the FISHRAND model (e.g., growth rate, lipid). These parameters were adjusted in the FISHRAND
model to optimize the fit between predicted body burdens and observed body burdens for the period
of the hindcast (i.e, calibration). In terms of TRVs, lipid normalization, which is how the egg versus
tissue TRVs were developed, represents a standardization of the TRY results. Because the
normalization is based on observed lipid in the egg and tissue, respectively, there is no sensitivity
to evaluate. The fillet to whole body ratios, derived on the basis of large datasets of observed ratios
between percent lipid in the fillet and the whole body, which is estimated at 1.5 for brown bullhead
and 2.5 for largemouth bass, would reduce estimated risks by these factors if the "true" ratio were
1:1. If the "true" ratio were higher, risks would accordingly increase, but the increase is unlikely to
be even a factor of 2 (i.e., from 2.5 to 5). Only the river otter and bald eagle risk estimates rely on
predicted body burdens using the 2.5 ratio (otter and eagle are assumed to consume largemouth
bass), and given the magnitude of risks for these receptors, a decrease in risk by a factor of 2.5 would
not change the overall conclusions of the ERA Addendum.
Regarding exposure media concentrations, the FISHRAND model incorporates annual
average sediment and monthly average water concentrations as inputs. Sediment concentrations are
stable and show little temporal variability. That is, monthly average sediment concentrations are not
significantly different from annual average concentrations. Water concentrations are much more
variable, and consequently, the FISHRAND model explicitly incorporates this variability by
characterizing water concentrations on a monthly basis. For the avian and mammalian receptors,
which are assumed to be exposed to summer average water concentrations, this period of exposure
coincides with the typical length of the toxicity study. This is also the period of greatest feeding,
particularly for migratory or hibernating species, and the period for which concentrations of PCBs
in water are highest. Thus, an annual average (which is longer than the period of exposure in the
77 TAMS/MCA
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toxicity study) would decrease risks, but given the magnitude of the predicted TQ, again, the overall
conclusions of the ERA Addendum would not change.
6.3.3 Sensitivity Analysis for Risk Models for Avian and Mammalian
Receptors
No significant comments were received on Section 6.3.3.
7.0 CONCLUSIONS
Response to EG-1.23
The ERA and ERA Addendum are based on USEPA policy and guidance and standard
ecological risk assessment practices. To the extent the ERA and ERA Addendum are used in
decision-making, along with the Human Health Risk Assessment, the Data Evaluation and
Investigation Report, and the results of the modeling, USEPA will document that use in the FS, the
Proposed Plan, and the Record of Decision (see also, responses to EG-1.1, EG-1.3, and EG-1.4).
Moreover, USEPA disagrees with the commenter's statement that the conclusions of the ERA
Addendum are "unambiguously contradicted" by data (see, response to EG-1.20).
7.1 Assessment Endpoint: Benthic Community Structure as a Food Source for
Local Fish and Wildlife
Response to EF-1.37
The uncertainty in sediment and water forecasts is estimated to be on the order of a factor of
two. This value is based on professional judgment. The parameterized model shows agreement
between the model predictions and the calibration data of a factor of two or better (see, Figures 3-5
and 3-7 of the ERA Addendum). As stated in the ERA Addendum (p. 71), "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."
7.2 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth, and
Reproduction) of Local Fish (Forage, Omnivorous, and Piscivorous)
Populations
No significant comments were received on Section 7.2.
78 TAMS/MCA
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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)
Response to EF-1.38
The sensitivity of tree swallows to PCBs as compared to other insectivorous birds is not well
documented. Other insectivorous bird species may be more sensitive to PCBs; however, even with
an uncertainty factor of ten to account for interspecies variation, most TQs would still fall below one.
The tree swallows and other insectivores in the Upper Hudson River may experience reproductive
impairment due to PCB exposure.
7.4 Assessment Endpoint: Protection and Maintenance (i.e., Survival, Growth and
Reproduction) of Lower Hudson River Waterfowl (as Represented by the
Mallard)
No significant comments were received on Section 7.4.
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)
No significant comments were received on Section 7.5.
7.6 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of
Insectivorous Mammals (as represented by the Little Brown Bat)
No significant comments were received on Section 7.6.
7.7 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of Local
Omnivorous Mammals (as represented by the Raccoon)
No significant comments were received on Section 7.7.
7.8 Assessment Endpoint: Protection (i.e., Survival and Reproduction) of Local
Piscivorous Mammals (as represented by the Mink and River Otter)
No significant comments were received on Section 7.8.
79 TAMS/MCA
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7.9 Assessment Endpoint: Protection of Threatened and Endangered Species
No significant comments were received on Section 7.9.
7.10 Assessment Endpoint: Protection of Significant Habitats
No significant comments were received on Section 7.10.
7.11 Summary
No significant comments were received on Section 7.11.
APPENDICES
APPENDIX A - Conversion from Tri+ PCB Loads to Dichloro through Hexachloro
Homologue Loads at the Federal Dam
Response to EF-1.2. EL-1.5 and EG-1.12b
The state variable modeled by HUDTOX is tri and higher PCBs (Tri+ PCBs). This variable
was chosen for the Hudson River because historic data exist for this form of PCBs, primarily from
the 1977 and 1984 NYSDEC sediment surveys and the USGS water column monitoring program
(1977 to the present). There is little historical homologue data on which to base a model calibration.
The Farley models are based on di through hexa homologues, requiring a conversion from Tri+
PCBs to homologues.
USEPA's conversion uses HUDTOX load estimates at Federal Dam (not Thompson Island
Dam) to predict the upstream boundary loads for the Farley model during the forecast period (only)
The processes used to convert HUDTOX Tri+ output to estimates of dichloro through hexachloro
homologues are explained in detail in Appendix A of the ERA Addendum (USEPA, 1999c), and are
based on observed data stratified by season. As with any forecast, there is uncertainty in these
estimates; however, uncertainty in the ratios between trichloro through hexachloro homologues is
expected to be much less than the uncertainty in forecasting total Tri+ PCB load, due to the
uncertainty in future contributions from the GE Bakers Falls source. Uncertainty is greatest for the
dichloro homologue, but this is the homologue with least significance to bioaccumulation in fish.
Variability in the homologue composition of future loads would have little effect on total PCB
concentrations in water and sediment, but would have an impact on bioaccumulation, as the more
chlorinated homologues generally have greater apparent bioaccumulation factors.
on
OU TAMS/MCA
-------
Thompson Island Dam and Waterford data were used to estimate the change in homologue
ratios relative to Tri+. This approach is reasonable for the following two reasons. As documented
in Figures A-24 to A-27 of Appendix A of the ERA Addendum (USEPA, 1999c), the ratios of the
homologue groups trichloro to hexachlorobiphenyl to the Tri+ sum have not varied greatly over time.
Indeed, most of the variation seen is due to seasonal changes that were addressed in the Appendix
A. Additionally, the results show very consistent trends over the period 1996-1998. This period was
utilized for the ratio estimates delivered at Thompson Island Dam. These results suggest that the
variations in these ratios have been well characterized and can be extrapolated without introducing
large amounts of uncertainty. Notably, the dichloro homologue has a much greater degree of
variability compared to the other four groups. However, its importance to downstream exposures
is much less, given that the dichloro homologue group does not tend to bioaccumulate and thus
constitutes a negligible portion of fish body burdens (see, Appendix K of the ERA, USEPA, 1999b).
As a result, human and ecological exposures to this homologue group are minimal. Thus, the greater
uncertainty in the dichloro homologue loads does not limit the usefulness of the loading calculation.
As to the examination of the changes in load between TI Dam and Waterford and the effect
on the homologue ratios, the presence or absence of a large upstream load above Rogers Island does
not affect the nature of the transport processes downstream. Specifically, the processes of sediment-
to-water exchange, gas exchange, and similar geochemical processes will occur in any event. Thus
the 1993 data are not inappropriate for examining the effects of transport between TI Dam and
Waterford. Recognizing that the geochemical processes will vary temporally, the results have been
grouped according to season. Additionally, the 1993 USEPA data have the distinction of either
tracking or integrating PCB loads in such a fashion so as to closely document the changes in
homologue ratios between these stations. Specifically, the transect data tracks and monitors a single
water parcel through the Upper Hudson during each sampling event. The flow-averaged samples
integrate PCB loads on the basis of flow over a 15-day period. Thus, a large number of randomly
collected data points is not necessary to establish the degree of change between the stations.
Regarding the representativeness of the USEPA Thompson Island Dam monitoring station,
USEPA has previously acknowledged that its Thompson Island Dam monitoring location as well as
the west wing station occupied by GE do not match the PCB load estimates obtained from a center
channel monitoring location (USEPA, 1999b). However, the USEPA does not agree that the center
channel is the true measure of the load at the Thompson Island Dam but rather, that the center
channel load is probably closer to the true value, which lies between the loads derived from the
center channel and west-wing-wall locations. Nonetheless, the use of the long-term wing wall
station in Appendix A does not examine load but simply the ratios among the congeners. Ratios at
this station are similar but not identical to those of the center channel. Figure EG-1.12 illustrates
this with data from GE, showing the ratio of each homologue group to Tri+ at the Thompson Island
Dam west wall and center channel stations. These results show the west wing wall station to have
a higher proportion of trichloro homologues and lower proportions of tetra through hexachloro
homologues relative to the center channel station. Thus, the use of the west wing wall station may
slightly underestimate the Upper Hudson contribution of the heaviest congeners. More importantly,
these figures show that each of the homologue groups represents a fairly consistent proportion of the
81 TAMS/MCA
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Ratio of Tri to Tri+
Ratio of Tetra to Tri+
0.7
0.6 -
0.5-
0.4 -
0.3 -
0.2 -
0.1 -
00
I f
1 2 .3 .4
TID West
0.5
04 -
0.3 -
a:
o_
Q 0.2 -
i—
0.1 -
00
.2 3
TID West
Ratio of Penta to Tri+
Ratio of Hexa to Tri+
0.3
0.2 -
0.1 -I
0.0
.1 2
TID West
0.3
0.2 -
0.1 -
0.0
.1 2
TID West
Notes:
Linear Fit
— Bivariate Normal Ellipse P=0 950
TID West = GE's Thompson Island Dam West-wing-wall station
TID PRW2 = GE's central channel station near the Thompson Island Dam
Dash line represents the 1:1 unity line.
Figure EG-1.12
Relationship Between the TI Dam West and Central Channel Stations for
Homologue to Tri+ Ratios
GE Data (1997- 1999)
-------
Tri+ sum regardless of the choice of monitoring location. Additionally, the difference between the
stations for each homologue group is less than or equal to the variability of the homologue-to-Tri+
ratio. This is illustrated in each instance by the 95* percentile ellipse that is elongated and close to
the line of perfect agreement. Lack of correlation between the stations would tend to yield a more
circular ellipse, indicating lack of correlation.
Ultimately, it must be noted that the ratios developed from the Thompson Island Dam west
station were only used to determine the mean proportion of each homologue in the Tri+ sum,
corrected on the basis of flow or season. The correlations behind these ratios have uncertainties
associated with them but these uncertainties should be relatively small and unlikely to affect the
long-term forecast results. Specifically, the differences in the ratios seen at the center channel and
west wing wall stations represent likely bounding values. Given that these differences were smaller
than the actual variations in the ratios as a function of flow or season, the uncertainty derived from
the use of the Thompson Island Dam west wing wall monitoring station does not introduce an
important additional degree of uncertainty.
Response to EF-1.39
The text on p. A-6 paragraph 1 of the ERA Addendum is revised to read:
The TI Dam data 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. The ratio variations among these groups are relatively small, typically only a few percent
of the total Tri-f 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 fall-winter change in the hexachloro homologue ratio
represents nearly a 50 percent decline in the ratio from fall-winter to summer. These results should
be compared to the dichloro homologue results, which show large changes on both absolute and
relative scales.
Response to EL-1.23
A total of 49 field duplicate samples at the Thompson Island Dam and Waterford stations are
given in the GE database (QEA, 1999). These samples are from the end of 1995 through 1999.
Averages of the values were not used because the concentration and homologue pattern of the
samples are similar. Using the sample as opposed to the average of the sample pair would not
change the conclusions of the analysis. This is seen in Figure EL-23a which shows the relative
percent difference (RPD) for the concentration of tri through hexa homologues. The differences are
small for the majority of samples with a median RPD of only 6%. Because the homologue patterns
were used in the conversion analysis, a comparison of the homologue distributions is more relevant.
83 TAMS/MCA
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Relative Percent Difference
1.5 -
Quantiles
maximum 100.0%
99.5%
97.5%
90.0%
quartile
median
quartile
minimum
75.0%
50.0%
25.0%
10.0%
2.5%
0.5%
0.0%
Moments
Mean
Std Dev
Std Error Mean
Upper 95% Mean
Lower 95% Mean
N
Sum Weights
167
167
167
33
12
5.5
2.5
1.1
0.18
0.08
0.08
15
34
4.8
25
58
49
49
Sources: GE, 1999
Relative Percent Difference (RPD) =
Measurement - Measurement!.
(Measurementl + Measurement!^)
~~
xlOO%
Figure EL- 1.23a
Relative Percent Difference for GE Water Column Sample Duplicates at the TI Dam
TAMS/MCA
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% Similarity Tri-Hexa
Quantiles
maximum 100.0%
99.5%
97.5%
90.0%
quanile 75.0%
median 50.0%
quartile 25.0%
10.0%
minimum
0.5%
0.0%
Moments
Mean
Std Dev
Std Error Mean
Upper 95% Mean
Lower 95% Mean
N
Sum Weights
99.7
99.7
99.7
99.1
98.0
96.8
94.4
89.0
59.6
51.3
51.3
95.0
7.23
1.03
97.2
93.0
49
49
Sources: GE, 1999
%Similarity = \(MinimumValueof[Homologue]i,[Homologue]2)
1=3
For example: Sample 1 Sample 2 Lesser
Tri(%) 45 48 45
Terra 22 20 20
Penta 18 19 18
Hexa 15 13 B
100% 100% 96% = 96% Similarity
Figure EL-1.23b
Percent Similarity of GE Water Column Sample Duplicates for the Tri through Hexa
Homologues at the TI Dam
TAMS/MCA
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Figure EL-1.23b shows the percent similarity for the tri through hexa homologues. Percent similarity
is a means of comparing distributions. The lower value for each of the homologue percent of Tri+
PCBs is summed. The closer the sum is to 100%, the more similar are the distributions. The
agreement between the homologue distributions also is satisfactory with a median value of 96.8%
and a mean value of 95.1%.
Response to EL-1.24
The conversion from Tri+ PCBs to homologues assumes that the factors can be applied to
a 40-year forecast because the geochemical processes creating the ratios among the homologues are
unlikely to change without remediation. The homologue patterns in the water column are generated
from the sediment inventories. Without altering the sediment inventory, the patterns in the water
column should remain relatively constant. Note the small variation in the mean mass percent of Tri+
PCBs using Thompson Island Dam data for tri through penta given by +/- two standard errors. More
variation is evident for di and hexa, but this is of less concern because Tri+ PCBs are modeled for
the fish body burdens and hexa is a small fraction of Tri+ PCBs. The mean mass percent ratio for
Waterford/TID is less well constrained. However, the di homologue is less important to the ERA
Addendum, which is primarily based on exposure to Tri+ PCBs through the food chain (see. EG-
1.12 for further discussion). The commenter is correct in noting that the discussion presents the
estimation of the Thompson Island Dam to Waterford correction first and the Thompson Island Dam
ratio estimates second.
Response to EL-1.25
USEPA Phase 2 water column samples were used to determine a correction to the PCS load
between the Thompson Island Dam and Waterford. 12 samples from the Thompson Island Dam and
12 samples from the Waterfbrd station were used in the calculation. There are a total of 53 water
column samples taken at Waterford by GE in 1991 and 1992. Although the GE data set provides
more than four times the number of USEPA samples at the Waterfbrd station, the collection method
that generates the data is inappropriate for this analyses. This is discussed in the ERA Addendum,
Appendix A, p. A-3, as follows:
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 en route. 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
86 TAMS/MCA
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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.
Response to EL-1.26
Figure EL-1.26a shows the most recent GE water column data at the Thompson Island Dam
(QEA, 2000). A decline in PCB concentrations at the Thompson Island Dam is not evident in the
data presented in this figure. This is further discussed in the Responsiveness Summary for the LRC
(USEPA, 1999d). While a decline in summer loads has been noted, this can be largely ascribed to
a decline in flow (see. Figure EL-1.26b). The fact that water column concentrations have not
declined with time after 1996 suggests a mechanism that releases PCBs from the sediment and
establishes a constant water column condition regardless of flow. Note how loads correlate with
flow in Figure EL-1.26c.
Response to EL-1.27
It is likely that these factors will remain constant for decades to come because the
homologue patterns found in the water column are a reflection of the patterns found in the source
of the contamination. The primary source of contamination is the river sediments. Without
remediation, PCBs will continue to be released from the sediment in the appropriate proportions
found there. Additionally, there are few data to establish an a priori basis for estimating a change
in these ratios. Thus, these values are assumed constant. See, response to comment EG-1.12 and
EL-1.24.
Response to EL-1.28
Between 1987 and 1990, there are no homologue data available because the GE monitoring
program had not begun. The 1991 GE sample data was used to represent this time period. Releases
from the Bakers Falls area, which occurred in 1991, might have yielded a different homologue
pattern than was actually present in 1987 to 1990. Thus, the 1991 pattern might not be representative
of the prior three years. However, the model output between 1987 and 1990 is not used in the ERA
Addendum. Thus, the lack of data between 1987 and 1990 does not effect the results of the ERA
Addendum.
87 TAMS/MCA
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TIDam West Station
s
5
u
CO
U
-
c
1000
800 -
600
400
200 -
Total PCB with Cubic Spline Fit
1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001
Year
Source: GE, 2000
Figure EL-1.26a
Total PCB Concentrations at the Thompson Island Dam (1991-2000)
TAMS/MCA
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4000
1 3500 -
!X]
£ 3000
o
a.
1
2500
2 2000
1500
1996
' ' ' I ' ' ' ' I ' ' ' ' I '
1996 1996 1997
1998
Year
1998
1998
1999
1999
Source: Hudson River Database Release 4.2
Figure EL-1.26b
Fort Edward Summer Average Flows
TAMS/MCA
-------
TIDWest
u
c.
+
£
o
S
50
40
30
20
10
y =-2.87 + 0.00756* R2= 0.631
' y = 7.99 + 0.002!! R2=0.116
y = -5.47 + 0.0067* RJ= 0.522 +
y = -40.9 + 0.0189x R2= 0.463
i 1 1 1 1 1 1 1 1 1 ! 1 1 1 1 1 1 1 1 1 r
1000 2000 3000 4000 5000
Monthly Average Flow Fort Edward (cfs)
6000
Sources: Hudson River Database Release 4.2 and GE, 2000
Figure EL-1.26c
Tri+ Loads at the TI Dam Compared to Flow at Fort Edward
TAMS/MCA
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APPENDIX B - Effects Assessment
Response to EF-1.40
As discussed in the response to EF-1.6, Hansen et al. (1974) and USAGE (1988) were
reexamined and selected to develop TRVs.
Response to EF-1.41
As discussed in the response to EF-1.6, Hansen et al. (1974), Adams et al. (1989, 1990,
1992), and USAGE (1988) were reexamined and selected to develop TRVs.
Endpoints of greatest relevance to the population and the ability of the population to
successfully reproduce were considered the most significant endpoints in developing TRVs.
Although other significant effects may have been observed at lower concentrations than those of the
selected TRVs, these endpoints have less direct relevance to population-level endpoints.
Response to EF-1.42
Comment acknowledged. See response to EF-1.6.
Response to EF-1.43
Comment acknowledged. The revised sentences read:
The LOAEL TRV for white perch is 0.6 ug TEQs/kg lipid (Table B-25).
The NOAEL TRV for white perch is 0.29 ug TEQs/kg lipid (Table B-25).
Response to EF-1.44
Comment acknowledged. Tables B-5 and B-6 are revised to include USAGE (1988).
Response to EF-1.45
The studies should be listed as dose-response studies, which estimate a NOAEL/LOAEL
rather than an EL-no effect or EL-effect.
91 TAMS/MCA
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Response to EF-1.46
The table states in a footnote that the lipid values were reported by a referenced study for that
species.
Response to EF-1.47
These toxicity endpoints are from the USAGE (1988) study and were inadvertently included
in Figure B-2. The sediments for this study were collected from the field and this study is not
considered a laboratory study. The purpose of the figure was to visually illustrate the range of
endpoints that were identified from the literature. Therefore, not all studies were included, but a few
representative studies were selected to show this range.
Response to EF-1.48
The purpose of Figure B-3 was to visually illustrate the range of toxicity endpoints that were
identified in the literature. Therefore, all studies were included in the evaluation, but very few
studies were selected to visually represent this range.
92 TAMS/MCA
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Risk Assessment
Revision
-------
III. RISK ASSESSMENT REVISIONS
1. Summary
This section of the Responsiveness Summary presents the revised results of the baseline
Ecological Risk Assessment for Future Risks in the Lower Hudson River (ERA Addendum). The
revisions are based on modified forecast concentrations of PCBs in fish, sediment, and water, which
in turn result from the revised upstream PCB boundary load into the Lower Hudson that was
presented in the Revised Baseline Modeling Report (RBMR)(USEPA, 2000a), and its subsequent
effects on output of the modeling for the Lower Hudson River. The revisions in this section also
incorporate changes to the toxicity reference values (TRVs), based on comments received on the
ERA and ERA Addendum. This section also compares the revised ecological risk results and
associated conclusions with those of the ERA Addendum.
The overall conclusions from the ERA Addendum (USEPA, 1999c) remain unchanged. The
revised calculations for the ERA Addendum show that there are ecological risks to receptors of
concern (except the tree swallow) above USEPA levels of concern. In addition, site-related risks due
to PCBs in Lower Hudson River are greatest for the top-level piscivorous receptors, such as the river
otter and the bald eagle.
2. Introduction
Part III of this Responsiveness Summary summarizes the modifications made and presents
the results of the revised risk calculations for the ERA Addendum. Those tables and figures that
were modified are labeled "Revised." To facilitate in the ease of comparing revised results with the
ERA Addendum results (USEPA, 1999c), all tables and figures have retained their number
designations.
2.1 Changes in the Modeled Concentrations of PCBs in Fish, Water and Sediment
The RBMR (USEPA, 2000a) contains the results of the recalibration of the HUDTOX and
FISHRAND models. Because these recalibrations yielded revised values for sediment, water, and
fish in the forecast results, it was necessary to revise the ERA Addendum to reflect these new values.
The changes in the HUDTOX and FISHRAND models reflected in the RBMR (USEPA, 2000a)
include the following:
• Use of a revised sediment resuspension model component in HUDTOX;
• Use of the 1998 surface (0-5 cm) sediment data obtained by GE as part of the
calibration;
• An extension of the model forecast to a 70 year period (1998 to 2067);
• Use of the 1991 sediment conditions as the initial conditions for the HUDTOX model
forecasts (i.e., after calibration, the model was initialized with the 1991 sediment
conditions and run to the year 2067);
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Recalibration of the FISHRAND model using Bayesian updating techniques; and
• Incorporation of individual species growth rates in the FISHRAND model.
The revised HUDTOX model results indicate that sediment concentrations increased slightly
(10-30%) or remained the same (see Appendix A of the RBMR, USEPA 2000a). The largest
difference was in the period 1993 to 1999, for which predicted sediment concentrations are now
higher than in the initial modeling results reported in the BMR (USEPA, 1999a). After 1999,
predicted sediment concentrations are approximately the same as they were previously. Predicted
water concentrations were more or less consistent between the BMR and RBMR. However, in the
ERA Addendum, Tri+ PCB concentrations were used to predict sediment and water concentrations,
while in this Responsiveness Summary, total PCB concentrations were used.
The models for the Lower Hudson River have been rerun due to changes made to the
HUDTOX forecast discussed above. The current forecast of PCB loads from the Upper Hudson
River given by HUDTOX are used. The revised Lower River fate and transport model is also used
(Cooney, 1999). The minor changes made to the Farley models since the March 1999 Report (Farley
et al., 1999) are discussed below.
2.1.1 Changes to the Farley Models between December 1999 and August 2000
Revised Farley fate, transport and bioaccumulation models (Cooney, 1999) were used for this
Responsiveness Summary for the ERA Addendum. The changes to the fate and transport model are
as follows (Cooney, 2000):
• The AUOC for sediment was changed from 1 to 0.1; and
A slight change to the solids balance that affects other flow parameters such as the
settling and resuspension rates.
Changes to the bioaccumulation model are:
The chemical and food assimilation for zooplankton have been set to 0.3;
• Striped bass and white perch lipid content are the average lipid content given by
NYSDEC fish samples taken in the 1990s; and
• A minor correction to the prey pattern has been made. Striped bass in one
compartment of the model fed upon white perch that were a year younger.
Comparison of Forecasted Waterr Sediment and Fish Data from the ERA Addendum and this
Responsiveness Summary
Comparisons were made between the Tri+ PCBs in the dissolved phase of the water column
between the revised model output and the data presented in the ERA Addendum. As seen in Figure
III-1, the R2 ranged from 0.996 at RM 152 to 1.0 at RMs 90 and 50, indicating that there is virtually
no difference between the original and revised results.
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Comparisons were also made between the total PCBs in the water column (whole water)
between the revised model output and the data presented in the ERA Addendum for future risks in
the Lower Hudson River. As seen in Figure III-2, the R2 ranged from 0.994 at RM 152 to 1.0 at
RMs 90 and 50, indicating that there was again virtually no difference between the original and
revised results.
The final comparisons were also made between the total PCBs in the sediment (0-2.5 cm)
between the revised model output and the data presented in the ERA Addendum. As seen in Figure
III-3, the R2 ranged from 0.998 at RM 152 to 1.0 at RMs 113, 90, and 50, indicating that there is
virtually no difference between the original and revised results of sediment concentrations.
*
These three comparisons show that although the Farley et al. (1999) was revised slightly,
changes were minimal and did not significantly change calculated water and sediment
concentrations.
2.1.2 Changes to FISHRAND between December 1999 and August 2000
In the RBMR (USEPA, 2000a), the FISHRAND model was formally recalibrated using
Bayesian updating. Growth rate coefficients, TOC, lipid content, and K,,w distributions were all
optimized within the constraints of the data.
Table 3-8 (Revised) shows the revised Tri+ PCB average and 95% UCL concentrations for
benthic invertebrates for the duration of the modeling period. Overall, concentrations in benthic
invertebrates are similar to those uses in the ERA Addendum. From 1993 to 2010-2011,
concentrations are 0.8 to 1.0 times the prediction used in the ERA Addendum. In the later portion
of the modeling period, concentrations are calculated to be up to 1.2 times higher than predicted
earlier. This pattern was fairly consistent between locations.
Revised predictions for the two forage fish modeled, the spottail shiner and the pumpkinseed,
are provided in Tables 3-9 and 3-10 (Revised), respectively. Predicted concentrations for the spottail
shiner are about 4.4 to 9.0 times higher than those used in the ERA Addendum. The spottail shiner
is the smaller forage fish species considered as a prey item by the other fish-eating wildlife receptors.
Revised concentrations for the pumpkinseed are 2.0 to 3.1 times higher than those used in the ERA
Addendum. Revised pumpkinseed concentrations were fairly consistent between locations and over
the duration of the sampling period.
Revised predictions for the yellow perch and white perch were 0.8 to 2.9 times the values
used in the ERA Addendum (Tables 3-11 and 3-12 Revised, respectively). There were fewer
changes in the values predicted in the 25th percentile (0.8 to 1.4 and 0.7 to 1.0 times the original
prediction for the yellow and white perch, respectively) than in the 95* percentile (1.4 to 2.9 and 1.6
to 2.2 times the original prediction for the yellow and white perch, respectively). The values for the
median were between the two percentiles (0.9 to 1.6 and 0.9 to 1.3 times the original prediction for
the yellow and white perch, respectively).
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Revised concentrations for brown bullhead (Table 3-13 Revised) were higher than those used
in the ERA Addendum. Concentrations were 1.1 to 1.9 times greater than earlier predictions.
Concentrations were higher at all locations for the duration of the modeling period. The greatest
increases were seen in the 25* percentile.
Revised concentrations for largemouth bass ranged from 0.45 to 3.17 times the values used
in the ERA Addendum(Table 3-14 Revised). Revised largemouth bass concentrations were lower
than initial predictions at RMs 152 and 113, with the exception of the 25th percentile at RM 152 after
1998. Revised concentrations at RMs 90 and 50 were 2.5 to 3.2 times higher than the initial
predictions. The largemouth bass is used as the "large" piscivorous fish consumed by the river otter
and bald eagle. Striped bass concentrations are calculated by applying a factor to the largemouth
bass concentrations (see Table 3-2), thus the changes in concentration for the striped bass are
proportional to the changes in concentration for the largemouth bass.
Revised exposure concentrations tables for avian and mammalian receptors based on the
revised forecasts are provided in Tables 3-25 to 3-60 (Revised).
2.2 Changes in Toxicity Reference Values
Toxicity Reference Values (TRVs) for fish, mallard, bald eagle, mink, and river otter were
revised from the ERA (USEPA, 1999b) and ERA Addendum (1999c) based on a reevaluation of
toxicity studies, as discussed in the following paragraphs.
2.2.1 Changes in Fish TRVs
The laboratory-based TRVs were revised for all fish receptors (i.e., pumpkinseed, spottail
shiner, brown bullhead, yellow perch, white perch, largemouth bass, striped bass, shortnose sturgeon.
The study by Hansen et al. (1974) was selected for development of the TRV for PCBs, instead of
the study by Bengsston (1980). Hansen et al. established a NOAEL for exposure to Aroclor 1254
of 1.9 mg/kg and a LOAEL of 9.3 mg/kg for adult female fish. The values for adult fish determined
in this study are more appropriate for comparison to measured and modeled concentrations in adult
Hudson River fish than the study by Bengsston (1980), which examined hatchability in minnows
exposed to Clophen A50. Because the sheepshead minnow is not in the same taxonomic family as
any Hudson River receptors, an interspecific uncertainty factor of 10 was applied to develop TRVs
for all fish.
Therefore, on the basis of laboratory toxicity studies:
• The LOAEL TRV for the pumpkinseed, spottail shiner, brown bullhead, yellow perch,
white perch, largemouth bass, spottail shiner, striped bass, and shortnose sturgeon is: 0.93
mg PCBs/kg tissue (Table 4-1).
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• The NOAEL TRY for the pumpkinseed, spottail shiner, brown bullhead, yellow perch,
white perch, largemouth bass, striped bass, and shortnose sturgeon is: 0.19 mg/kg PCBs/kg
tissue (Table 4-1).
The field-based TRVs for the pumpkinseed, spottail shiner, and largemouth bass were revised
from the ERA Addendum. For the pumpkinseed and largemouth bass, the field studies by Adams
et al. (1989,1990,1992) on the redbreast sunfish, a species in the same family as the pumpkinseed
and largemouth bass, were retained as the studies to establish TRVs. However, the growth endpoint,
rather than the reduced fecundity endpoint initially selected, was used to establish a TRY. The
NOAEL for growth was reported as being significantly different from a downstream location.
Growth is a relevant endpoint, and the NOAEL for growth, 0.3 mg/kg, is used in this assessment.
The sunfish (Lepomis auritus) in the studies were exposed to PCBs and mercury in the field.
However, because 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 TRY,
but not a LOAEL TRY, for the pumpkinseed and largemouth bass. An interspecies uncertainty
factor is not applied, because these three species are all in the same family (Centrachidae). 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 was not applied.
On the basis of the field studies:
• The NOAEL TRY for the pumpkinseed and largemouth bass is: 0.3 mg PCBs/kg tissue
(Table 4-1).
The previous NOAEL TRY for the pumpkinseed and largemouth bass was 0.5 mg PCBs/kg tissue
based upon the fecundity endpoint in Adams et al. (1992).
In the ERA Addendum, no field-based TRY was selected for the spottail shiner. However, upon
re-examination, the study by USACE (1988) using fathead minnow is considered to be a field-related
study, rather than a laboratory study, because the sediments to which the fathead minnow were
exposed were field-collected sediments (instead of spiked sediments). This study was selected for
development of a field-based TRY for the spottail shiner, a species in the same family as the fathead
minnow.
On the basis of the field study:
• The final NOAEL TRY for the spottail shiner is: 5.25 mg PCBs/kg wet wt tissue (Table 4-1).
The field-based TRY was selected for use, rather than the laboratory-based TRVs used in the August
1999 ERA Addendum.
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2.2.2 Changes in Avian TRVs
The total (Tri+) PCB daily dose TRY in the diet was revised for the mallard duck, as were the
total (Tri+) PCB and TEQ concentrations in bald eagle eggs. These changes are discussed below.
Mallard Duck
The development of TRVs for exposure of mallards to PCBs was re-examined with
consideration of two additional studies that were not identified in the literature studies that were
conducted for the ERA Addendum. A total of five studies were identified that examined the effects
of PCBs on mallards (Hill et al. 1975, Riseborough and Anderson 1975, Custer and Heinz 1980,
Heath et al. 1972, and Haseltine and Prouty 1980).
The study by Hill et al. (1975) was not selected for development of TRVs because it examined
mortality as an endpoint, which is not expected to be as sensitive an endpoint as growth and
reproduction. The studies by Riseborough and Anderson (1975), Custer and Heinz (1980), and
Heath et al. (1972) found no effects on various reproductive endpoints based on exposure to a single
dose (40 ppm, 25 ppm, and 25 ppm in diet, respectively). Haseltine and Prouty (1980) observed no
adverse effects on reproductive endpoints after a 12-week exposure to 150 ppm Aroclor 1242 in
food, but did observe significantly reduced weight gain in adults. Therefore, the study by Haseltine
and Prouty (1980) was selected as the most appropriate study, given that it is a dose response study
that reports a LOAEL on an ecologically relevant endpoint from which a NOAEL can be estimated.
Because only a single dose was tested, a LOAEL-to-NOAEL uncertainty factor often was applied
to estimate a NOAEL from this study. Because the study was conducted over a 12 week period, a
sub-chronic to chronic uncertainty factor is not applied.
Based on the results of Haseltine and Prouty (1980) on growth:
The LOAEL TRV for mallard (growth effects) is: 16 mg/kg/day (Table 4-2).
The NOAEL TRV for the mallard (growth effects) is: 1.6 mg/kg/day (Table 4-2).
Previously, a LOAEL of 2.6 mg/kg/day and a NOAEL of 0.26 mg/kg/day were used based on
Custer and Heinz (1980):
Bald Eagle
Upon reexamination, USEPA agrees that the data collected by Wiemeyer et al. (1993) does not
support the development of the previous NOAEL total PCB TRV of 3.0 mg/kg for bald eagle egg
concentrations. However, USEPA does not agree that because mean five-year production was not
significantly reduced for the residue interval ranging from 5.6 to <13 mg PCBs/kg, a NOAEL of 13
mg/kg is appropriate. It would be more appropriate to take the average value of the data in the 5.6
to <13 mg/kg interval as a measure of the average concentration for which production was not
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significantly impacted, as compared to higher concentrations. However, those data are not reported
in this paper. As an alternative, the average PCB concentration in eggs from successful nests (5.5
mg/kg), which was shown to be significantly lower than the concentration measured in unsuccessful
nests (8.7 mg/kg) (Wiemeyer et al. 1993, p. 224), is selected as the NOAEL-TRV for bald eagles.
Based on the study by Wiemeyer et al. (1993):
• The NOAEL TRY for PCBs in bald eagle eggs is: 5.5 mg PCBs/kg egg (Table 4-2).
Based on the same study, the previous NOAEL TRY for the bald eagle was 3.0 mg/kg egg.
To determine TEQ-based TRVs PCBs for bald eagle eggs, a study by Elliott et al. (1996) that
reports data for TEQ in the yolk sac of the bald eagle egg was used. This study reports a
concentration of TEQs of 210 ng/kg wet weight in eggs for the Powell River, a contaminated site
with a concentration that is slightly less than another nearby contaminated site, East Vancouver
Island. Based on Figure 4 in Elliott et al. (1996) the concentration of TEQs in the East Vancouver
Island site is estimated as 13,000 ng TEQs/kg lipid. Using the ratio between wet weight and lipid
at the Powell River site, the weight wet concentration at East Vancouver Island is approximately 217
ng/kg. Because no significant difference was observed between the average hatching rate of the eggs
collected from these two contaminated sites and the reference sites, the average concentration in eggs
from the contaminated sites (214 ng/kg wet weight) was selected as the NOAEL for this study.
The field based NOAEL TRV for TEQs in bald eagle eggs is: 0.214 ng/kg egg (Table
4-2).
Based on Powell et al. (1996), the previous laboratory-based NOAEL and LOAEL TRVs for
the bald eagle were 0.02 ug/kg egg and 0.01 ng/kg egg, respectively.
2.2.3 Changes in Mammalian TRVs
USEPA acknowledges that for TEQ-based PCBs in the diet, a LOAEL should not be established
from the Tillett et al. (1996) field study for the mink and river otter. In keeping with accepted
scientific practice, only NOAEL TRVs are developed from field studies in the ERA because other
contaminants or stressors may be contributing to observed effects. The revised risk estimates
remove this comparison (see Table 4-3).
3. Results
The overall conclusions drawn from the results of the ERA Addendum do not change as a result
of the revised risk calculations. Tables from Chapter 3,4, and 5 have been revised to reflect the
changes due to modeling and TRVs. In some cases, toxicity quotients (TQs) have increased or
decreased slightly, but these revisions do not affect the general text or the overall conclusions of the
ERA Addendum. Specific changes to risk characterization tables are as follows:
107 TAMS/MCA
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Table 5-1: Predicted sediment concentrations (Tri+) were adjusted to reflect total PCB
concentrations. None of the guidelines changed. Only average, rather than average and 95% UCL
results are calculated. Conclusions are unchanged but ratios increase slightly at all river miles (i.e.,
RM 152,113,90, and 50).
Table 5-2: Predicted water concentrations (Tri+) were adjusted to reflect total PCB concentrations.
The NYSDEC wildlife bioaccumulation criterion comparison was removed, since it is now the same
as the USEPA criterion (1.2 x 10"4 ug/L). Conclusions are unchanged but risks have decreased
slightly at all locations.
Table 5-3: Pumpkinseed field-based NOAEL changed from 0.5 to 0.3 mg/kg (based on Adams et
al., 1992; same study but different value). Toxicity quotients at all locations increased to 1.0 or
higher for the duration of the modeling period (1993 to 2018). Conclusions for pumpkinseed remain
unchanged, but predicted toxicity quotients increased slightly. Previously, all locations had predicted
toxicity quotients below one for a portion of the modeling period.
Table 5-4: Spottail shiner laboratory-based TRVs changed to a single field-based NOAEL based
on the USAGE study (previous lab-based NOAEL was 15 mg/kg while field-based NOAEL is 5.25
mg/kg). Predicted spottail shiner body burdens increased slightly or remained the same at all river
miles. Conclusions did not change, as all toxicity quotients remained below one, except for the 95th
percentile at RM 152 in 1993.
Table 5-5: This table is obsolete, as no LOAEL is derived from the field-based study.
Table 5-6: TEQ-based TRVs have not changed. Predicted pumpkinseed concentrations increased
slightly or remained the same at all river miles. Conclusions have changed slightly, because revised
risk estimates show that predicted toxicity quotients exceed one for a greater proportion of the
modeling period at all river miles.
Table 5-7: TEQ-based TRVs have not changed. Predicted pumpkinseed concentrations increased
slightly or remained the same at all river miles. Conclusions have changed slightly, previously all
predicted toxicity quotients fell below one at all locations. Revised risk estimates show that predicted
toxicity quotients are above one at RM 152 for the median in 1993, and above one but below ten for
the 95* percentile until 2003. At RM 113 revised risk estimates exceed one for the 95th percentile
until 1998. At RMs 90 and 50, revised risk estimates exceed one for the 95* percentile until 1996.
Tables 5-8 and 5-9: TEQ-based TRVs have not changed. Predicted spottail shiner concentrations
increased slightly at all river miles. Conclusions are unchanged (predicted toxicity quotients below
one for all locations and years).
Table 5-10: Laboratory-based TRVs for brown bullhead have changed: original NOAEL was 1.5
mg/kg based on Bengsston (1980) and revised NOAEL is 0.19 mg/kg based on Hansen et al. (1974).
108 TAMS/MCA
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Overall conclusions have changed slightly: predicted toxicity quotients have increased at all
locations.
Table 5-11: Laboratory-based TRVs for brown bullhead have changed: The original LOAEL was
1.5 based on Bengsston (1980) and the revised LOAEL is 0.93 based on Hansen et al. (1974).
Overall conclusions have changed slightly, as predicted toxicity quotients have increased and exceed
one at all locations for the duration of the modeling period.
Tables 5-12 and 5-13: The TEQ-based laboratory-based NOAEL and LOAEL for brown bullhead
have not changed. Predicted concentrations for brown bullhead have remained the same or changed
slightly at all locations. Overall conclusions have not changed: predicted toxicity quotients fall
below one for all locations and years.
Table 5-14: The field-based TRY for white perch has not changed. Conclusions have changed
slightly, previously only the 95* percentile toxicity quotient at RM 152 in 1993 was greater than one.
Revised numbers predict the 95th percentile to exceed one at RM 152 until 2015, at RM 113 until
1999, and at RM 90 until 1995. The median TQ at RM 152 also was greater than one in 1993.
Table 5-15: The laboratory-derived NOAEL for the yellow perch increased slightly to 0.19 mg/kg
based on the Hansen study from 0.16 mg/kg based on the Bengsston study. Overall conclusions
have not changed, but predicted toxicity quotients for the yellow perch have decreased for the
median and 25th percentile values for the later part of the modelling period.
Table 5-16: The laboratory-derived LOAEL for the yellow perch decreased to 0.93 mg/kg based on
the Hansen study from 1.5 mg/kg based on the Bengsston study. Conclusions have changed slightly,
previously all predicted toxicity quotients fell below one at all locations for the duration of the
modeling period. Revised risk estimates show that predicted toxicity quotients are above one at RM
152 for the 25th percentile until 1997, the median until 1998, and generally above one for the 95th
percentile until 2015. At RM 113 revised risk estimates exceed one for the median until 1995 and
the 95th percentile until 2004. At RMs 90 and 50, revised risk estimates exceed one for the 95*
percentile until 2000 and 1999, respectively.
Table 5-17: The TEQ-based NOAEL for the white perch has not changed. Predicted concentrations
have increased slightly for the 95* percentile at RMs 113,90 and 50. Conclusions have not changed.
Table 5-18: The TEQ-based LOAEL for the white perch has not changed. Predicted toxicity
quotients exceed one for the 95* percentile at all river miles for a greater proportion of the modeling
duration than predicted previously. Conclusions have not changed.
Table 5-19: The TEQ-based NOAEL for the yellow perch has not changed. Predicted concentrations
for yellow perch have increased slightly at all locations and for all percentiles. Predicted toxicity
quotients for the 95* percentile exceed one for the duration of the sampling period at all locations.
109 TAMS/MCA
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Predicted toxicity quotients for the median and 25* percentiles also exceed one for a portion of the
modeling period at all locations.
Table 5-20: The TEQ-based LOAEL for the yellow perch has not changed. Predicted concentrations
for yellow perch have increased slightly at all locations and for all percentiles. Predicted toxicity
quotients for the 95th percentile exceed one for the duration of the sampling period at RM 152 and
for a portion of the modeling period at other locations. Predicted toxicity quotients for the median
and 25th percentiles also exceed one for a portion of the modeling period at all locations.
Table 5-21: The field-based total PCB NOAEL for largemouth bass has decreased to 0.3 mg/kg from
0.5 mg/kg based on the Adams study. Predicted largemouth bass concentrations have decreased
slightly at RM 152 and increased slightly at RMs 113,90, and 50. All toxicity quotients exceed one
(and sometimes ten) at all river miles for the duration of the modeling period at the 25th percentile,
median concentration, and 95th percentile. Overall conclusions have not changed.
Tables 5-22 and 5-23: TRVs on a TEQ basis for largemouth bass have not changed. Revised risk
estimates show that predicted toxicity quotients exceed one on a NOAEL basis at all river miles for
the duration of the modeling period using the 95th percentile concentration. On a LOAEL basis,
toxicity quotients also slightly increased at all locations. Toxicity quotients exceed one for a greater
proportion of the modeling time frame than in the ERA Addendum.
Table 5-24: This table has not changed.
Table 5-25: The total PCB dietary dose TRY for the tree swallow has not changed. Overall, there
are slight decreases in predicted toxicity quotients, and all toxicity quotients remain below one for
the duration of the modeling period at all river miles. Conclusions have not changed.
Table 5-26: Total PCB egg concentration TRY for the tree swallow have not changed. Overall, there
are very slight decreases in predicted toxicity quotients, and all toxicity quotients remain below one
for the duration of the modeling period at all river miles.
Tables 5-27 and 5-28: TEQ-based TRVs for the tree swallow have not changed. The toxicity
quotients in these tables have decreased slightly.
Table 5-29: The dietary dose TRVs for the mallard have changed. The laboratory-based body
burden total PCB NOAEL is 1.6 mg/kg/day, and the laboratory based LOAEL is 16 mg/kg/day
based on Haseltine and Prouty (1980). The original NOAEL was 0.26 mg/kg/day and LOAEL 2.6
mg/kg/day based on Custer and Heinz (1980). Predicted toxicity quotients based on dietary dose
have decreased slightly and do not exceed one for any location, concentration, or time period.
Previously, calculated toxicity quotients exceeded one for a portion of the modeling period at all
locations.
110 TAMS/MCA
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Table 5-30: The mallard egg-based TRVs have not changed. Predicted toxicity quotients have
decreased slightly during the first part of the modeling period, but the conclusions have not changed.
Revised TQs exceed one on a NOAEL and LOAEL basis at RMs 152 and 113 for the duration of
the modeling period (1993-2018). NOAELs are exceeded at all RMs for the duration of the modeling
period.
Tables 5-31 and 5-32: Mallard TEQ-based TRVs have not changed. Toxicity quotients have
decreased slightly during the first part of the modeling period and increased slightly during the later
period. All TQs still exceed one at all locations for the duration of the modeling period.
Tables 5-33 and 5-34: Dietary dose TRVs did not change for the belted kingfisher and great blue
heron. Predicted concentrations of prey (spottail shiner) did increase at all locations, resulting in an
increase of the calculated toxicity quotients (generally less than a factor of two).
Table 5-35: Dietary dose TRVs did not change for the bald eagle. Predicted concentrations of prey
(largemouth bass) generally increased at RMs 152 and 113 and decreased at RM 90 and 50. Revised
toxicity quotients reflect these changes, decreasing slightly up river and increasing slightly down
river.
Tables 5-36 and 5-37: Egg concentration TRVs did not change for the belted kingfisher and great
blue heron. Predicted concentrations of prey (spottail shiner) did increase at all locations, resulting
in an increase of the calculated toxicity quotients (generally less than a factor of two). Calculated
TQs remain well above one.
Table 5-38: The field-based NOAEL for egg-based concentrations for the bald eagle has changed
from 3.0 mg/kg wet weight to 5.5 mg/kg wet weight. Conclusions have not changed although
predicted toxicity quotients have decreased by about a factor of two at RMs 152 and 113 and
increased by less than a factor of two at RMs 90 and 50 because of changes in prey concentration.
Tables 5-39: TEQ-based TRVs for the kingfisher have not changed. Dietary doses have increased
by roughly a factor of two, but the conclusions of risk remain unchanged.
Tables 5-40: TEQ-based TRVs for the great blue heron have not changed. Dietary doses have
increased by roughly a factor of three. All LOAEL-based toxicity quotients exceed one (with the
exception of the LOAEL at RM 50 in 2017 and 2018), in contrast to the earlier numbers where the
LOAEL-based TQs only exceeded one for a portion of the modeling period.
Table 5-41: TEQ-based TRVs for the bald eagle have not changed. Conclusions have not changed
although predicted toxicity quotients have decreased at RMs 152 and 113 and increased by more
than a factor of two at RMs 90 and 50 because of changes in prey concentration. All NOAEL and
LOAEL-based toxicity quotients exceed one for the duration of the modeling period.
111 TAMS/MCA
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Tables 5-42 and 5-43: TEQ-based egg concentration TRVs for the belted kingfisher and great blue
heron have not changed. Dietary doses have increased (within a factor of two) resulting in increases
in TQs. All NOAEL and LOAEL-based toxicity quotients exceed one by up to three orders of
magnitude. Conclusions do not change from the ERA Addendum.
Table 5-44: TEQ-based egg concentration TRVs for the bald eagle have changed. The revised
field-based NOAEL is 0.000214 mg/kg. It is not appropriate to develop a LOAEL from a field-
based study, thus, these comparisons have been removed. TQs have decreased at RMs 152 and 113
and increased at RMs 90 and 50 owing to changes in prey concentration. Conclusions have not
changed and TQs exceed one by up to four orders of magnitude at all locations.
Tables 5-45 and 5-46: TRVs for the little brown bat have not changed. TQs reflect changes in
benthic invertebrate concentrations, and are slightly lower during the first portion of the modeling
period and slightly higher at the end of the modeling period. Conclusions are unchanged.
Tables 5-47 and 5-48: TRVs for the raccoon have not changed. TQs have decreased slightly during
the first portion of the modeling period and increased slightly in later years. Conclusions remain
unchanged.
Table 5-49: TRVs for the mink have not changed. TQs have increased throughout the modeling
period because of increases in prey (forage fish) concentrations. Conclusions remain unchanged.
Table 5-50: TRVs for the river otter have not changed. TQs have decreased at RMs 152 and 113 and
increased at RMs 90 and 50 throughout the modeling period because of increases in prey
(piscivorous fish) concentrations. TQs now exceed one for both the NOAEL and LOAEL at all
locations for the duration of the modeling period.
Tables 5-51: The LOAEL-based comparisons for the mink have been removed since it is not
appropriate to develop a LOAEL from a field-based study. Consequently, only NOAEL-based
comparisons are provided on a TEQ basis. TQs for all NOAEL comparisons have increased because
of increases in prey (forage fish).
Table 5-52: The LOAEL comparisons on a TEQ basis for the river otter have been removed since
it is not appropriate to develop a LOAEL from a field-based study. Consequently, only NOAEL-
based comparisons are provided. TQs have decreased at RMs 152 and 113 and increased at RMs
90 and 50 throughout the modeling period because of increases in prey (piscivorous fish)
concentrations. TQs now exceed one by up to four orders of magnitude at all locations.
3.1 Comparison/Discussion
Several revisions were made to the HUDTOX model (input into the Farley model), FISHRAND
model, Farley model, and toxicity reference values that required recalculation of risks to receptors
evaluated in the ERA Addendum. None of the changes resulted in any significant changes to the
112 TAMS/MCA
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conclusions reached in the ERA Addendum for future risks in the Lower Hudson River. Years for
which predicted toxicity quotients fall above or below one, have changed slightly (in both directions)
based on the recalculated risks, as have the toxicity quotients.
The major findings of the ERA Addendum continue to indicate that receptors in close contact
with the Hudson River are at an increased ecological risk as a result of exposure to PCBs in
sediments, water, and/or prey. This conclusion is based on a toxicity quotient approach, in which
modeled body burdens, dietary doses, and egg concentrations of PCBs were compared to toxicity
reference values, and on field observations. On the basis of these comparisons, all receptors of
concern 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. Omnivorous
and piscivorous fish (i.e., which are higher on the food chain) are at greater risk than
forage fish. PCBs may adversely affect fish survival, growth, and reproduction.
• 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 River, 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 avion 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.
• 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 river designated as significant habitats may be
adversely affected by the PCBs.
113 TAMS/MCA
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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).
114 TAMS/MCA
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References
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McCarthy. 1989. The use of bioindicators for assessing the effects of pollutant stress on fish. Marine
Environmental Research. 28:459-464.
Adams, S.M., L.R. Shugart, G.R. Southworth and D.E. Hinton. 1990. Application of bioindicators
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communication to C. Hunt, TAMS Consultants, Inc. March 2,2000.
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(Anas platyrhynchos). Environmental Research. 23:29-34.
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Heath, R.G., J.W. Spann, J.F. Kreitzer, and C. Vance. 1972. Effects of polychlorinated biphenyls
on birds, pp. 475-485. In: The XVth International Ornithological Congress. K.H. Voous (ed).
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Hill, E., R.G Heath, J.W. Spann, and J.D. Williams. 1975. Lethal Dietary Toxicities of
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(TCDD) injected into the yolks of chicken (Callus domesticus) eggs prior to incubation. Arch.
Environ. Contam. Toxicol. 31:404-409.
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eggs. J. Wild. Manage. 39: 508-513.
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Bursian, T.J. Kubiak, J.P. Giesy, and R.J. Aulerich. 1996. Dietary exposure of mink to carp from
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equivalents, and biomagnification. Environmental Science & Technology. 30(1):283-291.
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TABLE 3-5: SUMMARY OF TRI+ WHOLE WATER CONCENTRATIONS FROM THE FARLEY MODEL AND TEQ-BASED PREDICTIONS FOR 1993 - 2018
REVISED
Tr
152
Whole
Water
Year Cone
mg/1
1993 3.4E-05
1994 37E-05
1995 16E-05
1996 49E-05
1997 3 OE-05
1998 19E-05
1999 1 6E-05
2000 26E-05
2001 3 OE-05
2002 15E-05
2003 17&05
2004 10E-05
2005 I5E-05
2006 19E-OS
2007 18E-05
2008 75E-06
2009 78E-06
2010 15E-05
2011 14E-05
2012 90E-06
2013 13E-05
2014 1 OE-05
2015 97E-06
2016 47E-06
2017 46E-06
2018 S.2E-06
i+ Average PCB Results Tn+ 95% UCL Results
113 152 113 152
Whole 90 Whole 50 Whole Whole Whole 90 Whole 50 Whole Whole
Water Water Water Water Water Water Water Water
Cone Cone Cone Cone Cone Cone Cone Cone
mg/1 mg/1 raf>/l mgfl mg/1 me/I aw/1 me/1
24E-05
2.2E-05
I4E-05
24&05
18E-05
14E-05
1 IE-OS
14E-05
16E-05
1 OE-05
1 1E-05
73E-06
80E-06
94E-06
9.4E-06
57E-06
52E-06
7.6E-06
76E-06
60E-06
74E-06
62E-06
38E-06
38E-06
33E-06
37&06
19E-05
16E-05
13E-05
15E-05
13E-05
1.1E-05
92E-06
93E-06
95E-06
8.0E-06
77E-06
61EO6
57E-06
59E-06
58E-06
46E-05
41E-06
46&06
47E-06
43E-06
46E-06
42E-06
40EO6
32E-06
29EO6
29E-06
1.5E-05
13E-05
1 IE-OS
1.1&05
9.9E-06
86E-OS
7.5&06
71E-06
6.8E-06
6IEO6
58E-06
49E-06
4.5E-06
43B-06
4.1E-06
36E06
33E-06
33E-06
33E-06
31&06
3.2&06
30E-06
29B06
26E-06
2.3E-05
23E-06
2.9E-08
32E-08
14E-08
42&08
25&08
16E-OS
14E-08
22E-08
2.5E-08
13EO8
14E-08
8.5&09
13E-08
16E-08
15E-08
64E-09
66E-09
I3E-08
12E-08
76&09
1IE08
87E09
83&09
40&09
39EO9
44&09
Average Avian TEF 95% Avian TEF Average Mammalian TEF 95» UCL Mammalian TEF
113 152 113 152 113 152 113
Whole 90 Whole SO Whole Whole Whole 90 Whole 50 Whole Whole Whole 90 Whole 50 Whole Whole Whole 90 Whole 50 Whole
Water Water Water Water Water Water Water Water Water Water Water Water Water Water Water
Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone Cone
mg/1 me/1 me/1 mg/l me/1 mg/1 mg/1 mg/1 me/1 me/1 me/1 me/1 me/1 me/1 mo/1
2.0E-08
19E-08
12E-08
2 IE-OS
1.6E-08
1 IE-OS
95E-09
12E-08
13E-08
89E-09
90E-09
6.2E-09
68E-09
80E-09
80E-09
48E-09
45E-09
6.5E-09
64E-09
51E-09
63E-09
5.3E-09
49E-09
33E-09
28E-09
31E-09
I6E-08
14E-08
1 IE-OS
1.3&08
1.1E-08
92E-09
78E-09
79E-09
81E-09
68E-09
65E-09
52E-09
48E-09
50E-09
49E-09
39E-09
35E-09
39E-09
40E-09
37E-09
39E-09
3.6E-09
34E-09
2.8E-09
2.4E-09
25E4»
13E-08
1 IE-OS
94E-09
92&09
84E-09
73E-09
64E-09
60E-09
58E-09
52E-09
49E-09
42E-09
3SE-09
37E-09
35E-09
31E-09
28E-09
2.8E-09
28E-09
2.7E-09
2.7E-09
26E-09
24E-09
22E-09
2.0E-09
20E-09
2.2E-08
2.4E-08
1 IE-OS
32E-08
19E-08
12E-08
1 IE-OS
17E-08
19E-08
98E-09
LIE-OS
6.5E-09
97E-09
12E-08
12E-08
49E-09
5.1E-09
9.9E-09
89E-09
5.9E-09
8.7E-09
6.7E-09
6.3E-09
3.1E-09
30E-09
34E-O9
16E-08
1.4E-08
9.2E-09
I6E-08
1.2E-08
8.8E-09
7.3E-09
92E-09
10E-08
68E-09
69E-09
4.8E-09
52E-09
62E-09
62E-09
37E-09
34E-09
SOE-09
49E-09
39E-09
48E-09
41E-09
3.8E-09
2.5E-09
22E-09
24E-09
1.2E-08
1 IE-08
84E-09
9.6E-09
85E-09
71E-09
60E-09
61E-09
62E-09
52E-09
50E-09
40E-09
37E-09
39E-09
38E-09
30E-09
2.7E-09
30E-09
31E-09
28E-09
30E-09
2.8E-09
2.6E-09
21E-09
1.9E-09
19E-09
98&09
86E-09
7.2E-09
7.IE-09
6.5E-09
56E-09
49E-09
46E-09
4.5E-09
4.0E-09
38E-09
3.2E-09
29E-09
28E-09
27E-09
23E-09
2IE-09
2.2E-09
2IE-09
20E-09
2.1E-09
20E-09
19E-09
17E-09
1.5E-09
15E-09
TAMS/MCA
-------
TABLE 3-6: SUMMARY OF TRI+ SEDIMENT CONCENTRATIONS FROM THE FARLEY MODEL AND TEQ-BASED PREDICTIONS FOR 1993-2018
REVISED
Tn+ Average PCB Results Tn+ 95% UCL Results
152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total 152 Total
Year SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc
ran/kg me/kg mg/kg ing/kg mg/kg mg/kg me/kg mg/kg mg/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
1 106
1015
0929
0957
0942
0875
0820
0817
0838
0806
0771
0725
0705
0715
0706
0676
0646
0654
0657
0643
0638
0621
0603
0578
0560
0556
0843
0.805
0758
0740
0726
0695
0661
0643
0640
0630
0611
0.585
0564
0557
0.549
0536
0518
0512
0509
0503
0497
0488
0477
0463
0451
0.443
0.664
0.634
0603
0580
0563
0542
0520
0502
0490
0.482
0469
0454
0439
0428
0419
0410
0400
0392
0386
0381
0376
0370
0363
0355
0347
0340
0484
0461
0440
0422
0408
0.394
0380
0367
0356
0349
0340
0331
0321
0.313
0306
0.299
0.292
0286
0281
0276
0272
0267
0263
0258
0254
0248
94E-O4
86E-O4
79E-04
8 IE-04
80E-04
7.4E-04
70E-04
69E-04
7 IE-04
69E-04
66E-04
62E-04
60E-04
6 IE-04
60E-04
57E-04
55E04
56E-04
56E-04
55E-04
S4E-04
5.3E-04
5 IE-04
49E-04
48E-04
47E-O4
Avenge Avian TEF
113 Total 90 Total
SedConc SedConc
me/kg mg/kg
7.2E-04
68E-04
64E-04
6.3E-04
6.2E-04
59E-04
56E-04
55E-04
54E-04
54E-04
5.2E-04
50E-04
48E-04
4.7E-04
47E-04
46E-04
44E-04
44E-O4
4JE-04
43E-04
42E-04
4 IE-04
4.1E-04
39E04
38E-04
38E-04
56E-04
54E-04
5 IE-04
49E-04
48E-04
46E-04
4.4E-04
4.3E-04
4.2E-04
4 IE-04
40E-O4
3.9E-04
3.7E-O4
36E-04
36E-04
35EO4
3.4&04
3.3E-O4
33E-04
32E-04
32E-04
31E-O4
3 IE-04
30E-04
30E-04
29E-04
95% Avian I tl- Average Mammalian 'ItF 95% UCL Mammalian TEF
50 Total 152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total 152 Total 113 Total 90 Total 50 Total
SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc SedConc
mg/kg rag/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kx mg/ke
4.1E-04
39&O4
37E-04
3.6E-04
35E-04
33E-04
32E-04
3 IE-04
30E-04
30E-04
29E-04
28E-04
27E-04
27E-04
2.6E-04
25EO4
25E-04
24E-04
24E-04
23E-04
23E-04
23E04
2.2E-04
22E-04
22E-04
2.1&O4
72E-04
66E44
6.1E-04
62E-04
62E-04
57E-04
54E-04
53E-04
55E-04
53E-04
50E-04
4.7&04
46E04
4.7E-04
46E-04
44E-04
42E-04
43E-04
4.3E-04
4.2E-04
42E-04
4I&04
39E-04
3.8&04
37E-04
36E-04
5.5E-04
S.3E-04
49E-04
48E-04
47E-04
4.5E-04
43E-04
4.2E-04
4.2E-04
4 IE-04
4.0E-04
38E-04
3.7E-04
3.6E-04
36E-04
3SE-04
34E-04
33E-04
33E-04
33E-04
32E-04
32E-04
3.1E-04
30E-04
29E-04
29E-04
4.3E-04
4.1E-04
3.9E-04
38E-04
37E-04
35E-04
34E-04
3.3E-04
3.2E-04
3 IE-04
3.1E-04
30E-04
29E-04
28E-04
27E-O4
Z7E-04
26E-04
26E-04
25E-04
25E-04
25E-O4
24E-04
24E-04
23E-04
23E-04
22E-04
3.2E-04
30E-04
2.9E-04
28E-04
27E-04
26E-04
25E-04
2.4E-04
2.3E-04
2.3E-04
22E-04
2.2E-04
2.1E-04
20E-04
20E-O4
2.0E-04
19E-04
19E-O4
I8E-04
18E-04
18E-04
17E-04
17E-04
1.7E-04
17E-04
16E-04
TAMS/MCA
-------
TABLE 3-7: ORGANIC CARBON NORMALIZED SEDIMENT
CONCENTRATIONS BASED ON USEPA
PHASE2DATASET
REVISED
Tri+ Average PCB Results
152 Total Sed
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
Cone
mg/kg
44.25
40.62
37.17
38.27
37.70
34.99
32.79
32.66
33.52
32.25
30.86
29.02
28.22
28.59
28.25
27.03
25.85
26.16
26.29
25.72
25.51
24.82
24.12
23.11
22.41
22.24
1 13 Total Sed
Cone
mg/kg
33.73
32.20
30.31
29.59
29.06
27.81
26.45
25.72
25.60
25.19
24.42
23.40
22.56
22.27
21.97
21.42
20.73
20.47
20.37
20.12
19.87
19.51
19.09
18.52
18.03
17.71
90 Total Sed
Cone
mg/kg
26.54
25.35
24.13
23.19
22.51
21.70
20.80
20.07
19.61
19.28
18.77
18.17
17.56
17.13
16.78
16.41
15.99
15.67
15.44
15.23
15.02
14.79
14.53
14.21
13.89
13.59
50 Total Sed
Cone
mg/kg
19.34
18.42
17.60
16.90
16.31
15.75
15.19
14.68
14.26
13.96
13.60
13.22
12.85
12.52
12.23
11.96
11.69
11.44
11.23
11.04
10.86
10.69
10.51
10.32
10.14
9.94
average TOC from Farley model 2.5%
TAMS/MCA
-------
TABLE 3-8: SUMMARY OF TRI+ BENTH1C INVERTEBRATE CONCENTRATIONS FROM THE FISHRAND MODEL AND TEQ-BASED PREDICTIONS FOR 1993 - 2018
REVISED
Tn+ Average PCS Results
152 Total 113 Total 90 Total 50 Total
Benlhic Benduc Benlhic Bcnthic
Year Cone Cone Cone Cone
ins/kg nut/kg out/kg me/kit
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2003
2006
2007
2008
2009
2010
2011
2012
2013
2014
2013
2016
2017
2018
1530
1410
1354
1357
1289
1217
1 165
1 182
1 167
1 123
1066
1058
1046
1038
1021
1007
0991
0978
0969
0954
0938
0923
0912
0917
0917
0920
1209
1 146
1094
1056
1028
0982
0949
0926
0918
0889
0858
0855
0846
0829
0818
0809
0795
0787
0775
0765
0754
0744
0732
0721
0720
0728
0967
0923
0883
0847
0816
0786
0757
0731
0723
0.702
0.677
0656
0652
0639
0631
0620
0613
0607
0597
0590
0581
0574
0564
0554
0.548
0537
0713
0686
0654
0630
0.600
0.582
0566
0549
0534
0516
0505
0.486
0.488
0.478
0470
0465
0458
0454
0445
0439
0433
0427
0421
0415
0409
0402
TIM- 95* UCL Results
152 Total 113 Total 90 Total 50 Total
Benthic Bcnthic Benduc Benduc
Cone Cone Cone Cone
ing/kg me/Ice me/lot me/leu
1611
1.488
1430
1432
1363
1292
1238
1254
1236
1.193
1 138
1 133
1 120
1 110
1093
1081
1064
1048
1037
1022
1004
0.990
0979
0991
0993
0998
1275
1.209
1 154
1 116
1087
1040
1.008
0983
0975
0.946
0914
0914
0905
0.887
0876
0868
0854
0845
0832
0820
0809
0799
0786
0775
0777
0788
1018
0973
0932
0894
0862
0832
0802
0775
0767
0745
0720
0699
0697
0.684
0676
0.664
0657
0651
0.640
0633
0623
0615
0605
0595
0588
0577
0751
0723
0.690
0665
0634
0.616
0600
0582
0566
0549
0537
0518
0.522
0.512
0.504
0498
0491
0487
0477
0471
0.465
0.458
0.452
0.445
0439
0433
152 Total
Benthic
Cone
mg/kg
2.1E-04
2.0E44
19E-04
19E-04
18E-04
17E-04
16E-04
16E-04
1.6ED4
1.6E-04
1.5E-04
15E-04
14E-04
14E-04
1.4E-04
14E-04
14E-04
14E-04
13E-04
13E44
13E-O4
13E-04
13E-04
13E-04
13E-04
13E-04
Avenge Avian TEH
113 Total 90 Total
Benduc Benthic
Cone Cone
ing/kg mg/kg
17E-04
16E04
I5E-04
1.5E-04
1.4E-04
1.4E-04
1.3E-04
1.3E-04
I3E-04
I2E-04
1.2E-04
I.2E-04
12E-04
1.1E-04
.1E-04
.IE-04
.1E-04
.1E-04
.1E-04
1E-O4
l.OE-04
10E-04
10E-04
10E-04
l.OE-04
10E-04
1JE-04
1.3E-O4
12E-04
12E-04
1 1E-04
1 1E-04
10E-04
10E-04
10E-04
97E-05
94E-05
91E-05
9.0E-05
89E-05
88E-05
86E-05
85E-05
84E-05
83E-05
82E-05
80E-05
80E-05
78E-05
77E-05
7.6E-05
74E-OS
SO Total
Benthic
Cone
9.9E-05
95E-OS
9 IE-OS
8.7E-05
8.3E-05
8.1E-05
78E-05
7.6E-05
7.4E-05
7 IE-OS
70E-05
67E-OS
68E-05
66E-05
65E-05
64E-05
63E-05
63E-05
62E-05
6 IE-OS
6.0E-05
59E-05
5.8E-05
S7&05
57E-05
56E-05
152 Total
Benthic
Cone
22E-04
2 IE-04
20E-04
20E-04
I9E-04
18E-04
17E-04
17E-04
1.7E-04
17E-04
16E-04
16E-04
16E-04
15E-04
15E-04
15E-04
15E-04
15E-04
14E-04
14E-04
I4E-04
14E-04
14E-04
14E-04
14E-04
14E-04
95* Avian TEF
113 Total 90 Total
Benduc Benduc
Cone Cone
18E-04
17E-04
1.6E-04
15E-04
I5E-04
I4E-04
14E-04
I.4E-04
1.4E-04
1.3E-04
1.3E-04
13E-04
13EO4
12E-04
12E-04
12E-04
12E-04
12E-04
1.2E-04
.1E-04
.1EO4
1E-04
1E-04
.1E-04
1E-04
1E-04
1.4E-04
13E-04
13E-04
12E-04
12E-04
12E-04
1 1E-04
1.1E-04
1 1E-04
l.OE-04
10E-04
97E-05
9.7E-05
95E-OS
9.4E-05
92E-05
9 IE-OS
90E-OS
89E-05
88E-05
8.6E-05
8.5E-05
84E-05
82E-OS
82E-OS
80E-OS
50 Total
Benduc
Cone
ing/kg
10E-04
10E-04
96E-OS
9.2E-05
8.8E-OS
85E-05
8.3&05
8.IE-05
7.9E-05
76E-05
7.4M5
7.2E-05
7.2E-OS
1. IE-OS
70E-OS
6.9E-05
68E-OS
67E-OS
66E-OS
6.5E-OS
64EO5
64E-05
63E-05
6.2E-05
61E-05
6.0&O5
Avenge Manunalian TEF
152 Total 113 Total 90 Total 50 Total
Benduc Benthic Benthic Benthic
Cone Cone Cone Cone
nig/kg mg/kg mg/kg mg/kg
I7E-04
15E-04
15E-04
15&04
14E-04
IJE-04
13E44
I3EXW
13E-04
12E-04
1.2E-04
1IE-04
1.IE-04
11E-04
1 IE-04
1 1E-04
1 1E-04
11E-04
10E-04
10E-04
10E-04
10E-04
98E-OS
99E-OS
99E-05
9.9&OS
13E-04
1.2E-04
1.2E-04
1 1E-04
1 1E-04
1.IE-04
10E-04
10E-04
99E-05
96E-05
93E-05
92E-OS
91E-05
89E-05
88E-OS
8.7E-05
86E-05
85E-05
84E-05
83E-OS
8.IE-05
8.0E-05
79E-05
78E-OS
78E-OS
79E-05
10E-04
IOE-04
9.5E-05
9 IE-OS
88E-OS
85E-OS
82E-OS
79&OS
7.8E-05
76&05
7.3E-05
7 IE-OS
7.0E-05
69E-05
68E-OS
6.7E-05
66E-OS
66E-OS
64E-05
64E-OS
63E-05
6.2E-OS
6.1E-05
6.0E-05
5.9E-05
58E-05
7.7E-05
7.4&05
7 IE-OS
6 SB-OS
65E-OS
63E-05
6 IE-OS
5.9E-05
58E-05
56E-05
54E-OS
S2E-OS
5.3E-05
52BO5
5 IE-OS
SOE-05
49E-OS
4.9E/O5
48E-05
47E-O5
47E-05
4.6EO5
45E-O5
45EO5
44E-OS
43E05
951k UCL Mammalian TEI-
152 Total 113 Total 90 Total 50 Total
Benlhic Bcnthic Benthic Benthic
Cone Cone Cone Cone
ing/kg mg/kg mg/kg nig/kg
17E-04
16E-04
15E-04
1.5E-04
1.5E-04
1.4E-04
13E-04
14E-04
13E-04
13E-04
12E-04
12E-04
12E-04
12E-04
1.2E-04
12E-04
1E-04
.1E-04
1E-04
.1E-04
1E-04
1E-04
1E-04
IE-04
IE-04
IE-04
I4E-04
13E-04
1.2E-04
12E-04
12E-O4
1 IE-04
1 lEOt
I.1E-04
1 IE-04
l.OE-04
99E-05
99E-05
98E-05
96&05
95E-05
94E-05
92E-OS
9 IE-OS
90E-05
89E-OS
87E-05
86E-OS
8 SB-OS
84E-05
84E-05
85E-05
1 IE-04
IOE-04
IOE-04
96E-05
93E-OS
90E-05
86EO5
8.4E-05
8.3E-OS
80E-05
7.8E-05
7.5E-OS
7SE-05
74E-05
73E-OS
7.2E-05
7 IE-OS
70E-OS
69E-05
68E-OS
67E-OS
6.6E-05
6.SE-OS
6.4E-OS
6.3E-OS
62E-OS
8 IE-OS
78E-05
74E-05
72E-05
68E-OS
66E-OS
6SE-OS
63E-03
6 IE-OS
59E-OS
S SB-OS
56E-05
56E-05
5 SB-OS
S4E-05
54E-OS
53E-OS
S.3E-05
S IE-OS
5 IE-OS
5 OEMS
49E-05
49E-OS
48E-05
47E-OS
47E-05
TAMS/MCA
-------
TABLE 3-9: SPOTTAIL SHINER PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
REVISED
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 Percentil
(mg/kg (rag/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.66
1.20
1.11
1.35
1.10
0.91
0.81
0.77
0.89
0.82
0.64
0.54
0.57
0.65
0.53
0.48
0.41
0.54
0.50
0.49
0.53
0.45
0.42
0.38
0.37
0.38
2.34
1.73
1.54
1.99
1.61
1.24
1.17
1.06
1.23
1.09
0.95
0.76
0.77
0.88
0.76
0.69
0.58
0.72
0.72
0.71
0.76
0.67
0.60
0.52
0.52
0.54
5.25
3.54
3.22
4.17
3.45
2.63
2.37
2.15
2.47
2.29
1.94
1.49
1.48
1.72
1.47
1.36
1.17
1.40
1.46
1.40
1.51
1.36
1.18
1.05
1.05
1.05
River Mile 113
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.67
1.54
1.21
1.34
1.18
1.01
0.90
0.90
0.89
0.84
0.80
0.68
0.65
0.66
0.64
0.58
0.52
0.54
0.58
0.56
0.57
0.55
0.52
0.47
0.44
0.43
2.15
1.96
1.56
1.70
1.52
1.31
1.16
1.15
1.15
1.08
1.02
0.87
0.83
0.85
0.82
0.76
0.68
0.70
0.74
0.73
0.73
0.71
0.67
0.61
0.57
0.57
4.40
3.70
3.19
3.32
3.11
2.62
2.18
2.05
2.16
2.05
1.82
1.59
1.45
1.50
1.44
1.40
1.20
1.24
1.30
1.30
1.30
1.27
1.17
1.08
1.02
1.00
River Mile 90
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.33
1.21
1.05
0.98
0.91
0.83
0.74
0.69
0.66
0.64
0.61
0.55
0.51
0.49
0.48
0.45
0.42
0.42
0.42
0.41
0.41
0.40
0.39
0.36
0.34
0.33
.70
.54
.34
.25
.16
.07
0.95
0.87
0.85
0.81
0.77
0.70
0.65
0.63
0.62
0.58
0.54
0.54
0.54
0.54
0.53
0.52
0.51
0.48
0.45
0.44
3.35
3.00
2.70
2.42
2.24
2.04
1.78
1.59
1.50
1.50
1.40
1.29
1.18
1.11
1.06
1.05
0.97
0.91
0.93
0.92
0.92
0.91
0.88
0.85
0.80
0.76
River Mile 50
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.25
1.11
0.98
0.90
0.82
0.75
0.68
0.62
0.58
0.55
0.52
0.49
0.45
0.43
0.41
0.39
0.36
0.35
0.35
0.34
0.34
0.33
0.32
0.31
0.30
0.29
1.63
1.44
1.28
1.15
1.05
0.96
0.87
0.79
0.74
0.71
0.67
0.62
0.58
0.55
0.52
0.50
0.47
0.46
0.45
0.45
0.44
0.43
0.42
0.40
0.39
0.37
3.31
2.89
2.55
2.26
2.04
.87
.68
.49
.36
.30
.23
.16
.06
0.98
0.93
0.89
0.83
0.80
0.78
0.77
0.76
0.75
0.73
0.71
0.68
0.65
TAMS/MCA
-------
TABLE 3-10: PUMPKINSEED PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
REVISED
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 Percent!!
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
3.24
2.47
2.07
2.80
2.18
1.49
1.29
1.48
1.57
1.23
1.27
0.88
0.98
1.18
0.94
0.75
0.75
0.86
1.02
0.87
0.97
0.83
0.78
0.55
0.53
0.57
4.09
3.10
2.67
3.54
2.73
1.89
1.64
1.84
1.99
1.57
1.61
1.10
1.19
1.46
1.15
0.94
0.95
.06
.25
.08
.23
.05
0.96
0.69
0.67
0.69
7.73
5.16
4.58
5.91
5.11
3.80
3.14
2.99
3.41
3.13
2.66
1.92
1.93
2.31
1.81
1.73
1.46
1.81
1.92
1.77
1.96
1.74
1.48
1.23
1.17
1.18
River Mile 113
95th
25th Median Percent!)
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
2.04
.79
.38
.55
.37
.13
0.96
0.99
1.01
0.90
0.85
0.70
0.67
0.71
0.69
0.60
0.52
0.58
0.60
0.57
0.59
0.56
0.53
0.46
0.42
0.42
2.58
2.25
1.72
1.96
1.69
1.42
1.22
1.24
1.27
1.14
1.08
0.88
0.84
0.87
0.85
0.75
0.65
0.71
0.75
0.72
0.73
0.71
0.66
0.57
0.53
0.52
5.08
4.22
3.59
3.59
3.36
2.90
2.39
2.08
2.33
2.22
1.91
1.60
1.41
1.53
1.44
1.34
1.11
1.22
1.27
1.23
1.28
1.21
1.12
0.97
0.90
0.86
River Mile 90
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.58
1.41
1.21
1.12
1.04
0.92
0.81
0.75
0.72
0.69
0.65
0.59
0.54
0.51
0.49
0.47
0.42
0.43
0.43
0.42
0.42
0.41
0.40
0.37
0.35
0.33
1.99
1.77
1.50
1.41
1.30
1.17
1.03
0.94
0.90
0.87
0.81
0.73
0.66
0.64
0.62
0.58
0.53
0.53
0.53
0.52
0.53
0.51
0.49
0.46
0.42
0.41
3.74
3.31
2.93
2.64
2.41
2.21
1.91
1.64
1.58
1.51
1.41
1.30
1.15
1.07
1.03
0.98
0.90
0.85
0.86
0.85
0.86
0.84
0.81
0.78
0.72
0.67
River Mile 50
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.56
1.36
1.20
1.07
0.97
0.88
0.79
0.72
0.66
0.63
0.59
0.55
0.50
0.46
0.44
0.42
0.39
0.37
0.37
0.36
0.36
0.35
0.34
0.33
0.31
0.30
2.02
1.76
1.53
1.36
1.23
1.11
1.00
0.89
0.83
0.79
0.74
0.68
0.62
0.57
0.54
0.52
0.48
0.47
0.46
0.45
0.45
0.44
0.43
0.40
0.38
0.37
3.79
3.29
2.86
2.54
2.27
2.05
1.83
1.63
1.50
1.41
1.31
1.20
1.08
1.00
0.94
0.89
0.81
0.77
0.75
0.73
0.73
0.72
0.70
0.67
0.64
0.61
TAMS/MCA
-------
TABLE 3-11: YELLOW PERCH PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
REVISED
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 Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
0.98
0.94
0.70
0.99
0.72
0.57
0.50
0.62
0.61
0.48
0.48
0.37
0.41
0.46
0.41
0.32
0.30
0.35
0.38
0.34
0.37
0.33
0.31
0.25
0.24
0.24
1.35
1.30
0.98
1.36
0.99
0.78
0.68
0.83
0.82
0.65
0.66
0.49
0.54
0.62
0.54
0.43
0.41
0.46
0.52
0.47
0.50
0.44
0.42
0.34
0.33
0.33
3.21
2.64
2.41
2.92
2.33
.72
.47
.56
.67
.40
.35
.00
.06
1.30
1.02
0.88
0.86
0.97
1.07
1.01
0.99
0.94
0.87
0.68
0.65
0.66
River Mile 113
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
0.72
0.67
0.52
'0.58
0.51
0.44
0.40
0.40
0.40
0.37
0.35
0.30
0.29
0.30
0.29
0.25
0.23
0.24
0.25
0.25
0.25
0.24
0.23
0.20
0.19
0.19
1.00
0.93
0.72
0.80
0.70
0.61
0.53
0.54
0.54
0.50
0.48
0.40
0.39
0.40
0.38
0.35
0.31
0.33
0.34
0.33
0.34
0.33
0.31
0.28
0.26
0.26
2.32
2.06
1.62
1.75
1.58
1.31
1.13
1.09
1.11
1.04
0.95
0.79
0.76
0.79
0.76
0.71
0.62
0.64
0.68
0.67
0.66
0.65
0.62
0.55
0.51
0.51
River Mile 90
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
0.57
0.52
0.45
0.43
0.39
0.36
0.32
0.31
0.30
0.29
0.27
0.25
0.23
0.22
0.22
0.20
0.18
0.19
0.19
0.19
0.19
0.18
0.18
0.16
0.16
0.15
0.79
0.72
0.62
0.59
0.54
0.49
0.44
0.41
0.39
0.39
0.36
0.33
0.31
0.29
0.28
0.27
0.25
0.25
0.25
0.25
0.25
0.24
0.23
0.22
0.21
0.20
1.80
1.61
1.39
1.28
1.16
1.04
0.91
0.85
0.82
0.78
0.73
0.65
0.60
0.58
0.57
0.53
0.49
0.50
0.50
0.49
0.49
0.48
0.47
0.43
0.41
0.40
River Mile 50
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
0.54
0.48
0.42
0.39
0.36
0.32
0.29
0.27
0.26
0.25
0.23
0.22
0.20
0.19
0.18
0.18
0.16
0.16
0.16
0.15
0.15
0.15
0.15
0.14
0.13
0.13
0.75
0.67
0.59
0.54
0.49
0.45
0.40
0.37
0.35
0.33
0.32
0.29
0.27
0.26
0.24
0.23
0.21
0.21
0.21
0.20
0.20
0.20
0.19
0.19
0.18
0.17
1.78
1.56
1.36
1.22
1.10
0.99
0.87
0.79
0.74
0.70
0.65
0.60
0.55
0.52
0.49
0.47
0.43
0.43
0.42
0.41
0.41
0.40
0.39
0.38
0.36
0.35
TAMS/MCA
-------
TABLE 3-12: WHITE PERCH PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
REVISED
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
200S
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
River Mile 152
95th
25th Median Percentil
(mg/kg (rag/kg e (mg/kg
wet wet wet
weight) weight) weight)
2.19
2.00
1.67
1.93
1.74
1.53
1.41
1.48
1.47
1.36
1.29
1.18
1.20
1.25
1.19
1.11
1.06
1.10
1.11
1.06
1.07
1.03
1.01
0.96
0.94
0.93
3.20
2.84
2.40
2.83
2.48
2.17
2.01
2.09
2.10
1.94
1.84
1.70
1.71
1.76
1.70
1.58
1.53
1.57
1.57
1.50
1.54
1.48
1.45
1.37
1.34
1.35
7.03
5.88
5.24
5.50
5.21
4.60
4.23
4.17
4.42
4.11
3.83
3.52
3.44
3.47
3.42
3.27
3.12
3.18
3.22
3.09
3.13
3.02
2.94
2.87
2.77
2.72
River Mile 113
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
.40
.29
.18
.17
.12
.04
0.96
0.94
0.95
0.92
0.87
0.82
0.79
0.79
0.78
0.76
0.72
0.73
0.73
0.71
0.71
0.69
0.68
0.66
0.64
0.64
1.96
1.83
1.64
1.63
1.54
1.45
1.34
1.33
1.32
1.27
1.22
1.15
1.12
1.13
1.10
1.06
1.03
1.02
1.02
1.00
1.00
0.98
0.96
0.93
0.91
0.91
4.34
3.96
3.64
3.44
3.36
3.13
2.87
2.71
2.72
2.71
2.56
2.43
2.32
2.29
2.26
2.21
2.13
2.09
2.11
2.08
2.06
2.02
1.98
1.93
1.89
1.86
River Mile 90
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
1.09
1.02
0.95
0.90
0.86
0.81
0.76
0.73
0.71
0.70
0.67
0.64
0.62
0.60
0.59
0.58
0.56
0.55
0.55
0.54
0.53
0.53
0.52
0.50
0.49
0.49
1.55
1.44
.33
.26
.20
.14
.07
.03
.01
0.98
0.95
0.90
0.87
0.85
0.84
0.81
0.79
0.79
0.78
0.77
0.76
0.75
0.73
0.72
0.70
0.69
3.41
3.16
2.94
2.75
2.60
2.45
2.31
2.17
2.09
2.05
1.98
1.90
1.82
1.77
1.73
1.70
1.65
1.62
1.60
1.58
1.56
1.54
1.51
1.48
1.45
1.42
River Mile 50
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
0.92
0.84
0.78
0.73
0.69
0.65
0.61
0.58
0.56
0.55
0.53
0.51
0.49
0.48
0.46
0.45
0.44
0.43
0.42
0.42
0.41
0.41
0.40
0.39
0.38
0.38
1.30
1.20
1.10
1.03
0.97
0.91
0.86
0.82
0.79
0.77
0.74
0.72
0.69
0.67
0.65
0.64
0.62
0.61
0.60
0.59
0.58
0.57
0.56
0.55
0.54
0.53
2.94
2.70
2.45
2.27
2.13
2.00
.87
.76
.67
.62
.58
1.52
1.45
1.39
1.35
1.32
1.29
1.26
1.24
1.22
.20
.19
.17
.15
.12
.10
TAMS/MCA
-------
TABLE 3-13: BROWN BULLHEAD PREDICTED TRI+ CONCENTRATIONS FOR 1993 - 2018
REVISED
River Mile
25th Median
(mg/kg (mg/kg
wet wet
Year weight) weight)
1993 3.11
1994 2.84
1995 2.58
1996 2.73
1997 2.57
1998 2.38
1999 2.23
2000 2.26
2001 2.29
2002 2.18
2003 2.04
2004 1.95
2005 1.95
2006 1.97
2007 1.93
2008 1.84
2009 1.80
2010 1.84
2011 1.81
2012 1.76
2013 1.76
2014 1.70
2015 1.67
2016 1.63
2017 1.62
2018 1.60
3.83
3.51
3.19
3.43
3.16
2.94
2.78
2.82
2.84
2.68
2.56
2.47
2.46
2.46
2.40
2.33
2.28
2.27
2.24
2.20
2.18
2.15
2.13
2.09
2.06
2.05
152
95th
Percentil
e (mg/kg
wet
weight)
6.47
5.69
5.43
5.57
5.34
4.99
4.62
4.51
4.60
4.50
4.24
4.07
3.97
3.99
3.90
3.85
3.73
3.67
3.68
3.61
3.59
3.52
3.45
3.41
3.38
3.31
River Mile
25th Median
(mg/kg (mg/kg
wet wet
weight) weight)
2.39
2.25
2.09
2.06
1.98
1.88
1.79
1.76
1.75
1.70
1.65
1.58
1.55
1.54
1.52
1.48
1.44
1.44
1.42
1.40
1.38
1.36
1.34
1.31
1.29
1.28
2.94
2.76
2.57
2.55
2.44
2.32
2.23
2.20
2.18
2.12
2.05
.99
.95
.93
.91
.87
.83
.81
.79
.76
.74
.72
.70
.67
.65
.64
113 River Mile 90 River Mile 50
95th 95th 95th
Percentil 25th Median Percentil 25th Median Percentil
e (mg/kg (mg/kg (mg/kg e (mg/kg (mg/kg (mg/kg e (mg/kg
wet wet wet wet wet wet wet
weight) weight) weight) weight) weight) weight) weight)
4.90
4.60
4.32
4.18
4.06
3.90
3.67
3.57
3.53
3.49
3.38
3.27
3.19
3.15
3.10
3.07
2.99
2.94
2.92
2.88
.90
.80
.70
.62
.56
.49
.43
.38
.35
.32
.28
.24
.21
.18
.16
.14
.12
.10
.09
.07
2.85 1.06
2.81 1.04
2.75 1.03
2.72 1.01
2.68 0.99
2.65 0.98
2.33
2.21
2.08
2.00
1.92
1.84
1.77
1.72
1.68
1.65
1.60
1.55
1.51
1.48
1.46
1.43
1.41
1.39
1.37
1.35
1.34
1.32
1.30
1.28
1.26
1.24
3.88
3.66
3.46
3.30
3.18
3.04
2.90
2.80
2.74
2.69
.49
.40
.32
.25
.20
.15
.10
.06
.03
.00
2.62 0.98
2.56 0.95
2.48 0.92
2.43 0.90
2.38 0.88
2.35 0.87
2.31 0.85
2.26 0.84
2.24 0.82
2.21 0.81
2.18 0.80
2.15 0.79
2.11 0.78
2.08 0.76
2.05 0.75
2.01 0.74
1.82
1.72
1.62
1.54
1.47
1.41
1.35
1.31
1.28
1.25
1.21
1.18
1.15
1.12
1.10
1.09
1.07
1.06
1.04
1.03
1.01
1.00
0.98
0.97
0.95
0.94
3.06
2.86
2.69
2.55
2.43
2.33
2.24
2.15
2.09
2.04
1.99
1.94
1.89
1.85
1.82
1.79
1.76
1.72
1.70
1.67
1.65
1.63
1.60
1.58
1.55
1.53
TAMS/MCA
-------
TABLE 3-14: LARGEMOUTH BASS PREDICTED TR1+ CONCENTRATIONS FOR 1993 - 2018
REVISED
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 Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
9.23
6.71
5.49
7.39
6.08
4.94
5.63
5.98
6.20
6.51
4.72
4.23
4.63
5.06
4.84
3.99
3.23
3.87
4.33
4.64
4.86
5.17
3.76
3.47
4.04
4.41
10.61
7.54
6.26
8.22
6.97
4.89
4.16
4.07
4.54
4.10
3.68
2.76
2.77
3.23
2.78
2.51
2.28
2.44
2.85
2.44
2.74
2.48
2.22
1.99
1.82
1.81
15.61
10.49
9.30
11.08
10.25
5.60
4.73
4.58
5.11
4.67
4.13
3.14
3.15
3.58
3.14
2.87
2.58
2.72
3.20
2.74
3.14
2.76
2.53
2.27
2.05
2.02
River Mile 113
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
7.26
6.23
5.12
5.36
4.94
4.24
3.69
3.41
3.58
3.43
3.12
2.72
2.44
2.54
2.48
2.34
2.06
2.05
2.22
2.13
2.18
2.09
1.96
1.85
1.69
1.62
8.25
7.09
5.86
6.04
5.61
4.84
4.20
3.88
4.04
3.91
3.52
3.08
2.78
2.87
2.80
2.64
2.33
2.33
2.52
2.40
2.48
2.35
2.22
2.09
1.91
1.84
12.02
10.38
8.86
8.56
8.22
7.10
6.13
5.64
5.84
5.64
5.12
4.54
4.06
4.13
4.06
3.86
3.42
3.33
3.61
3.52
3.57
3.44
3.27
3.07
2.78
2.69
River Mile 90
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
5.59
5.04
4.42
4.00
3.74
3.40
3.03
2.72
2.57
2.53
2.39
2.21
2.00
1.87
1.81
1.76
1.65
1.56
1.57
1.56
1.55
1.53
1.49
1.43
1.34
1.27
6.34
5.72
5.04
4.52
4.23
3.85
3.43
3.07
2.93
2.85
2.69
2.49
2.25
2.12
2.06
1.99
1.86
1.77
1.78
1.77
1.76
1.72
1.67
1.61
1.51
1.43
9.11
8.28
7.38
6.53
6.09
5.60
4.98
4.46
4.23
4.11
3.89
3.63
3.28
3.10
3.00
2.88
2.70
2.59
2.61
2.59
2.57
2.54
2.46
2.34
2.20
2.11
River Mile 50
95th
25th Median Percentil
(mg/kg (mg/kg e (mg/kg
wet wet wet
weight) weight) weight)
5.45
4.81
4.24
3.78
3.44
3.13
2.82
2.55
2.36
2.25
2.12
1.97
1.82
1.70
1.61
1.54
1.43
1.37
1.34
.32
.31
.29
.26
.22
1.17
1.11
6.20
5.46
4.82
4.30
3.90
3.55
3.20
2.89
2.66
2.52
2.39
2.24
2.06
.91
.82
.73
.62
.55
.52
.49
.48
.45
,.43
.38
.32
.26
9.07
7.96
7.03
6.21
5.64
5.12
4.62
4.19
3.84
3.62
3.43
3.21
2.95
2.75
2.61
2.50
2.36
2.25
2.21
2.17
2.15
2.11
2.06
2.00
1.91
1.82
TAMS/MCA
-------
TABLE 4-1
TOXIOTY REFERENCE VALUES FOR FISH
DIETARY DOSES AND EGG CONCENTRATIONS OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
REVISED
TRVs
Tissue Concentration
Lab-based TRVs for PCBs (mg/kg wet WL)
Field-based TRVs for PCBs (mg/kg wel wt)
Lab-based TRY for TEQs (ug/kg Iipid)
from salmooids
Lab-based TRV for TEQs (ug/kg Iipid)
From non-solrnonids
Held-based TRVs for TEQs (ug/kg Iipid)
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
NOAEL
Pump Unseed
(Lepomls
giUonv)
SpottaU
Shiner
(Nctropli
hudsonbu)
093
019
NA
0.3
0.6
OJ9
103
OJ4
NA
NA
093
0.19
NA
5.25
Not denved
Not denved
103
5.4
NA
NA
Brown Bullhead
neiufoiuf)
Yellow Perch
(PmaflmtKcns )
0.93
0.19
NA
NA
IS
8.0
Noi denved
Nol denved
NA
NA
0.93
0.19
NA
NA
0.6
0.29
103
0.54
NA
NA
White Perch
(Moron*
amertcana)
093
019
NA
3.1
0.6
0.29
103
0.54
NA
NA
Largemouth Bass
(Micnptenu
salmolda)
093
019
NA
0.3
0.6
0.29
103
054
NA
NA
Striped Bass
(Marone
fondttu)
093
019
NA
3.1
0.6
039
10.3
DM
NA
NA
Shortnose Sturgeon
(Actpciuer
onvtfostnuJt )
References
0.93
0.19
NA
NA
0.6
0.29
10.3
OJ4
NA
NA
Himseneial (1974)
While perch and slnped bass* Wesun el
al. (1983). spouail shiner USAGE
(1988)
PuroploDseed nod Largernoutti bass.
Adams el al (1989. 1990. 1992)
Brown Bullhead Elonen et &1 ( 1998)
Alloiners Walker el al (1994)
Oliven and Cooper (1997)
Note:
' Pumpbnsced (Lepoma gibboaa) and spouail shiner (Nompis hadtanau)
Units vary for PCBs and TEQ
NA = Noi available
Selected TRVs are boldtd and Holloaed
5/8/00
Page 1 of 1
TAMS/MCA
-------
TABLE 4-2
TOXICITY REFERENCE VALUES FOR BIRDS
DIETARY DOSES AND EGO CONCENTRATIONS OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
REVISED
II It
I I i * ^
TRVs r ,
Dietary Dose ~ " " . r-a
Lab-based TRVs for PCBs (mg/kg/day)
Field-based TRVs for PCBs (mgflcd/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
^ ~ "* -^ "° crb-
* Tree Swallow a
(Tachycineia bicolor)
0.07
0.01
NA
16.1
0.014
0.0014
NA
4.9
; Mallard Duck
, Mhos plajyrhychos~]
i, jj "
16
1.6
NA
NA
0.014
0.0014
NA
• NA
^ I i L
~n
Belted Kingfisher '
[Ceiyh alcyon )""""
*JT ™ "t ~ u ™ ,
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
* i ^"T !
1 .r
Great Blue Heron
(Ardea herodias)
~i" " "
0.07
0.01
NA
NA
0.014
0.0014
NA
NA
i Bald Eagle '
(Haliaeetas
leucocephalus )
r* ^
0.07
0.07
NA
NA
' I
; .' , . V ; ,, References „ i' .
t „'-,-„ „ , ;
1 - ' - ' i
Mallard: Haseltine and Prouty (1980)
All others: Scon (1977)
Tree Swallow: US EPA Phase 2 Database (1998)
0.0/4
0.0074
NA
NA
Noseketal. (1992)
US EPA Phase 2 Database (1998)
Egg Concentration ' ' £ n „
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
NA
0.005
2.21
0.33
NA
NA
0.02
0.01
NA
NA
2.27
0.33
NA
NA
NA
2
0.5
0.3
2.21
0.33
NA
5.5
0.02
0.01
NA
0.274
Scott (1977)
Bald Eagle: Wiemeyer (1984, 1993)
Tree Swallow: US EPA Phase 2 Database (1998)
Great Blue Heron: Janz and Bell ward (1996)
Others: Powell etal. (1996a)
Mallard: White and Segniak (1994); White and Hoffman (1995)
Great Blue Heron: Sanderson et al. (1994)
Eagle: Bliotetal. (1996a)
Tree Swallow: US EPA Phase 2 Database (1998)
Note: Units vary for PCBs and TEQ.
NA = Not Available
Selected TRVs are bolded and italicized.
5/8/00
Page 1 of
TAMS/MCA
-------
TABLE 4-3
TOXICITY REFERENCE VALUES FOR MAMMALS
DIETARY DOSES OF TOTAL PCBs AND DIOXIN TOXIC EQUIVALENTS (TEQs)
REVISED
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
(Procyonlotor)
0.15
0.032
NA
NA
0.001
0.0001
NA
NA
Mink
(Mustela
vison)
0.07
0.01
0.13
0.004
0.001
0.0001
NA
0.00008
Otter
(Lutra canadensis }
0.07
0.01
0.73
0.004
0.001
0.0001
NA
0.00008
References
Mink and oner Autench and Ringer (1977)
Raccoon and bat: Linder et a! (1984)
Heatonetal. (1995)
Murray etal (1979)
Tilhttetal (1996)
Note: Units vary for PCBs and TEQ.
Note: TRVs for raccoon and bat are based on mulit-generauonal studies to which interspecies uncertainty factors are applied.
NA = Not Available
Final selected TRVs are balded and italicized.
5/8/00
Page 1 of 1
TAMS/MCA
-------
TABLE 5-1: RATIO OF FARLEY PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES
REVISED
Year
1993
1994
199S
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
Sed Cone Sed Cone Sed Cone Sed Cone
TEC. 0 04 mg/kg dry weight
28 21 17 12
25 20 16 12
23 19 IS 11
24 18 14 11
24 18 14 10
22 17 14 9.8
20 17 13 9.5
20 16 13 9.2
21 16 12 8.9
20 16 12 8.7
19 15 12 &5
18 15 11 S3
18 14 11 8.0
18 14 11 7.8
18 14 10 7.6
17 13 10 75
16 13 10 73
16 13 9.8 7.2
16 13 9.6 7.0
16 13 9.5 6.9
16 12 9.4 6.8
16 12 9.2 6.7
15 12 9.1 6.6
14 12 8.9 6.4
14 11 8.7 6.3
14 11 8£ 6.2
Average PCB Results
152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone
MEC: 0.4 mg/kg dry weight
2,8 2.1 1.7 1.2
2.5 2.0 1.6 1.2
23 1.9 1.5 1.1
2.4 1.8 1.4 1.1
2.4 1.8 1.4 1.0
2.2 1.7 1.4 1.0
2.0 1.7 13 0.9
2.0 1.6 13 09
2.1 1.6 1.2 09
2.0 1.6 1.2 0.9
1.9 1.5 1.2 0.8
1.8 IS 1.1 08
1.8 1.4 1.1 08
1.8 1.4 1.1 0.8
1.8 1.4 1.0 08
1.7 13 1.0 0.7
1.6 13 1.0 0.7
1.6 13 1.0 07
1.6 13 1.0 0.7
1.6 13 1.0 0.7
1.6 1.2 0.9 07
1.6 1.2 09 07
1.5 1.2 09 0.7
1.4 1.2 09 06
1.4 1.1 09 06
1.4 1.1 08 0.6
Average PCB Results
152Total 113 Total 90Total SOTotal
Sed Cone Sed Cone Sed Cone Sed Cone
EEC: 1 7 mg/kg dry weight
0.7 05 0.4 0.3
0.6 0.5 0.4 03
0.5 04 04 0.3
0.6 0.4 03 02
0.6 04 03 0.2
0.5 04 0.3 0.2
0.5 04 0.3 0.2
0.5 04 0.3 02
0.5 04 0.3 02
0.5 04 0.3 02
0.5 04 0.3 02
0.4 0.3 03 02
0.4 03 0.3 02
0.4 0.3 0.3 0.2
0.4 03 0.2 02
0.4 0.3 0.2 02
0.4 03 0.2 02
0.4 0.3 02 0.2
04 0.3 02 0.2
04 0.3 02 0.2
0.4 03 02 0.2
0.4 03 02 02
04 03 02 02
03 03 0.2 0.2
0.3 0.3 0.2 O.I
0.3 03 02 0.1
Average PCB Results
152 Total 113 Total 90 Total SOTotal
Sed Cone Sed Cone Sed Cone Sed Cone
NYSDEC Ben. Chr. 19 3 mg/Kg OC
23 1.7 1.4 1.0
2.1 1.7 13 1.0
1.9 1.6 13 0.9
2.0 1.5 1.2 0.9
2.0 1.5 1.2 0.8
1.8 1.4 1.1 0.8
1.7 1.4 1.1 08
1.7 13 1.0 0.8
1.7 13 1.0 07
1.7 13 1.0 0.7
1.6 13 1.0 0.7
1.5 1.2 09 0.7
1.5 1.2 09 0.7
1.5 1.2 0.9 06
1.5 1.1 0.9 0.6
1.4 1.1 0.9 06
13 1.1 0.8 0.6
1.4 1.1 0.8 06
1.4 1.1 0.8 06
13 1.0 0.8 0.6
13 1.0 0.8 06
13 1.0 0.8 06
1.2 1.0 08 0.5
1.2 1.0 0.7 0.5
1.2 09 0.7 0.5
1.2 09 0.7 0.5
Average PCB Results
152 Total 113 Total 90 Total SOTotal
Sed Cone Sed Cone Sed Cone Sed Cone
NYSDEC Wildlife 1 4 mg/Kg OC
32 24 19 14
29 23 18 13
27 22 17 13
27 21 17 12
27 21 16 12
25 20 15 11
23 19 15 10.8
23 18 14 103
24 18 14 10.2
23 18 14 10.0
22 17 13 9.7
21 17 13 9.4
20 16 13 9.2
20 16 12 8.9
20 16 12 8.7
19 15 11.7 8.5
18 15 11.4 83
19 15 11.2 '-8.2
19 15 11.0 8.0
18 14 10.9 7.9
18 14 10.7 7.8
18 14 10.6 7.6
17 14 10.4 • 7.5
17 13 10.1 7.4
16 12.9 9.9 7.2
16 12.6 9.7 7.1
exceedances are bolded
Page 11 of 13
TAMS/MCA
-------
TABLE 5-1: RATIO OF FARLEY PREDICTED SEDIMENT CONCENTRATIONS TO SEDIMENT GUIDELINES
CONTINUED -REVISED
Average PCB Results
152 Total 113 Total 90 Total 50 Total
Year Sed Cone Sed Cone Sed Cone Sed Cone
Persaud LEL 0 07 rag/Kg dw
1993 16 12 9 7
1994 IS 11 9 7
1995 13 11 9 6
1996 14 11 8 6
1997 13 10 8 6
1998 12 10 8 6
1999 12 9 7 5
2000 12 9 7 5
2001 12 9 7 5
2002 12 9 7 5
2003 11 9 7 5
2004 10 8 6 S
2005 10 8 6 5
2006 10 8 6
2007 10 8 6
2008 10 8 6
2009 976
2010 976
2011 9 7 6
2012 975
2013 975
2014 975
2015 975
2016 875
2017 865
2018 865
Average PCB Results
152 Total 113 Total 90 Total SO Total
Sed Cone Sed Cone Sed Cone Sed Cone
Persaud SEL S30 rag/Kg OC
0.08 0.06 005 004
0.08 0.06 0.05 0.03
007 006 0.05 0.03
007 0.06 004 0.03
007 0.05 004 0.03
007 0.05 0.04 0.03
0.06 0.05 004 0.03
006 0.05 004 0.03
0.06 005 0.04 0.03
0.06 0.05 0.04 0.03
0.06 0.05 004 0.03
005 0.04 0.03 0.02
005 004 0.03 002
0.05 0.04 003 0.02
0.05 0.04 003 0.02
0.05 0.04 003 0.02
0.05 0.04 0.03 002
0.05 0.04 003 002
005 0-04 0.03 0.02
0.05 0.04 0.03 0.02
0.05 004 0.03 0.02
0.05 0.04 003 0.02
005 0.04 003 0.02
0.04 003 0.03 0.02
0.04 0.03 0.03 002
0.04 003 0.03 002
Average PCB Results
152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone
WA PAET 1242 0 1 mg/Kg dw
11 8.4 6.6 4.8
10 8.0 63 4.6
9.3 7.6 6.0 4.4
9.6 7.4 5.8 4.2
9.4 73 5.6 4.1
8.7 7.0 5.4 3.9
8.2 6.6 5.2 3.8
8.2 6.4 5.0 3.7
8.4 6.4 4.9 3.6
8.1 6.3 4.8 3.5
7.7 6.1 4.7 3.4
73 5.9 4.5 33
7.1 5.6 4.4 3.2
7.1 5.6 43 3.1
7.1 53 4.2 3.1
6.8 5.4 4.1 3.0
63 5.2 4.0 2.9
63 5.1 3.9 2.9
6.6 5.1 3.9 2.8
6.4 5.0 3.8 2.8
6.4 5.0 3.8 2.7
6.2 4.9 3.7 2.7
6.0 4.8 3.6 2.6
5.8 4.6 3.6 2.6
5.6 43 33 23
5.6 4.4 3.4 2.5
Average PCB Results
152 Total 113 Total 90 Total 50 Total
Sed Cone Sed Cone Sed Cone Sed Cone
WA PAET Microtox 0.021 mg/Kg
53 40 32 23
48 38 30 22
44 36 29 21
46 35 28 20
45 35 27 19
42 33 26 19
39 31 25 18
39 31 24 17
40 30 23 17
38 30 23 17
37 29 22 16
35 28 22 16
34 27 21 15
34 27 20 15
34 26 20 15
32 26 20 14
31 25 19 14
31 24 19 14
31 24 18 13
31 24 18 13
30 24 18 13
30 23 18 13
29 23 17 13
28 22 17 12
27 21 17 12
26 21 16 12
Page 12 of 13
TAMS/MCA
-------
TABLE 5-2: RATIO OF FARLEY PREDICTED WHOLE WATER
CONCENTRATIONS TO BENCHMARKS - REVISED
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
USEPA/NYSDEC - Ben. Aqu. 0.014 ug/L
2.4
2.7
1.2
3.5
2.1
1.3
1.2
1.9
2.1
1.1
1.2
0.7
1.1
1.3
1.3
0.5
0.6
1.1
1.0
0.6
1.0
0.7
0.7
0.3
0.3
0.4
1.7
1.6
1.0
1.7
1.3
1.0
0.8
1.0
1.1
0.7
0.8
0.5
0.6
0.7
0.7
0.4
0.4
0.5
0.5
0.4
0.5
0.4
0.4
0.3
0.2
0.3
1.3
1.2
0.9
1.1
0.9
0.8
0.7
0.7
0.7
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
0.2
0.2
0.2
1.1
0.9
0.8
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
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
Tri+ Average PCB Results
152 113
Whole Whole 90 Whole 50 Whole
Water Water Water Water
Cone Cone Cone Cone
USEPA/NYSDEC - W. Bio. 1.2 E-04 ug/L
286
310
136
412
249
157
135
220
247
125
138
84
123
155
152
63
65
126
114
75
111
85
81
39
38
44
199
184
117
203
154
113
93
117
130
87
88
61
67
79
79
48
44
63
63
50
61
52
48
32
28
31
155
136
108
123
108
90
77
77
80
67
64
51
47
49
48
38
34
39
39
36
38
35
33
27
24
24
125
109
92
91
82
72
63
59
57
51
48
41
37
36
34
30
27
28
27
26
26
25
24
21
19
19
exceedances are bolded
-------
TABLE 5-3: RATIO OF PREDICTED PUMPKINSEED CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
REVISED
River Mile 152 River Mile 1 13 River Mile 90
95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight) weight) weight) weight) 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
10.8
8.2
6.9
9.3
7.3
5.0
4.3
4.9
5.2
4.1
4.2
2.9
3.3
3.9
3.1
2.5
2.5
2.9
3.4
2.9
3.2
2.8
2.6
1.8
1.8
1.9
13.6
10.3
8.9
11.8
9.1
6.3
5.5
6.1
6.6
5.2
5.4
3.7
4.0
4.9
3.8
3.1
3.2
3.5
4.2
3.6
4.1
3.5
3.2
2.3
2.2
2.3
25.8
17.2
15.3
19.7
17.0
12.7
10.5
10.0
11.4
10.4
8.9
6.4
6.4
7.7
6.0
5.8
4.9
6.0
6.4
5.9
6.5
5.8
4.9
4.1
3.9
3.9
6.8
6.0
4.6
5.2
4.6
3.8
3.2
3.3
3.4
3.0
2.8
2.3
2.2
2.4
2.3
2.0
-1.7
1.9
2.0
1.9
2.0
1.9
1.8
1.5
1.4
1.4
8.6
7.5
5.7
6.5
5.6
4.7
4.1
4.1
4.2
3.8
3.6
2.9
2.8
2.9
2.8
2.5
2.2
2.4
2.5
2.4
2.4
2.4
2.2
1.9
1.8
1.7
16.9
14.1
12.0
12.0
11.2
9.7
8.0
6.9
7.8
7.4
6.4
5.3
4.7
5.1
4.8
4.5
3.7
4.1
4.2
4.1
4.3
4.0
3.7
3.2
3.0
2.9
5.3
4.7
4.0
3.7
3.5
3.1
2.7
2.5
2.4
2.3
2.2
2.0
1.8
1.7
1.6
1.6
1.4
1.4
1.4
1.4
1.4
1.4
1.3
1.2
1.2
1.1
6.6
5.9
5.0
4.7
4.3
3.9
3.4
3.1
3.0
2.9
2.7
2.4
2.2
2.1
2.1
1.9
1.8
1.8
1.8
1.7
1.8
1.7
1.6
1.5
1.4
1.4
12.5
11.0
9.8
8.8
8.0
7.4
6.4
5.5
5.3
5.0
4.7
4.3
3.8
3.6
3.4
3.3
3.0
2.8
2.9
2.8
2.9
2.8
2.7
2.6
2.4
2.2
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
5.2
4.5
4.0
3.6
3.2
2.9
2.6
2.4
2.2
2.1
2.0
1.8
1.7
1.5
1.5
1.4
1.3
1.2
1.2
1.2
1.2
1.2
1.1
1.1
1.0
1.0
6.7
5.9
5.1
4.5
4.1
3.7
3.3
3.0
2.8
2.6
2.5
2.3
2.1
1.9
1.8
1.7
1.6
1.6
1.5
1.5
1.5
1.5
1.4
1.3
1.3
1.2
12.6
11.0
9.5
8.5
7.6
6.8
6.1
5.4
5.0
4.7
4.4
4.0
3.6
3.3
3.1
3.0
2.7
2.6
2.5
2.4
2.4
2.4
2.3
2.2
2.1
2.0
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-4: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
REVISED
River Mile 1 52 River Mile 1 13 River Mile 90
95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight) weight) weight) weight) 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
0.11
0.08
0.074
0.09
0.07
0.060
0.054
0.051
0.059
0.055
0.042
0.036
0.038
0.043
0.036
0.032
0.027
0.036
0.033
0.033
0.035
0.030
0.028
0.025
0.025
0.025
0.45
0.33
0.29
0.38
0.31
0.24
0.22
0.20
0.23
0.21
0.18
0.15
0.15
0.17
0.14
0.13
0.11
0.14
0.14
0.13
0.15
0.13
0.11
0.10
0.10
0.10
1.00
0.68
0.61
0.80
0.66
0.50
0.45
0.41
0.47
0.44
0.37
0.28
0.28
0.33
0.28
0.26
0.22
0.27
0.28
0.27
0.29
0.26
0.22
0.20
0.20
0.20
0.32
0.29
0.23
0.26
0.22
0.19
0.17
0.17
0.17
0.16
0.15
0.13
0.12
0.13
0.12
0.11
0.10
0.10
0.11
0.11
0.11
0.11
0.10
0.09
0.08
0.08
0.41
0.37
0.30
0.32
0.29
0.25
0.22
0.22
0.22
0.21
0.19
0.17
0.16
0.16
0.16
0.15
0.13
0.13
0.14
0.14
0.14
0.14
0.13
0.12
0.11
0.11
0.84
0.71
0.61
0.63
0.59
0.50
0.41
0.39
0.41
0.39
0.35
0.30
0.28
0.29
0.27
0.27
0.23
0.24
0.25
0.25
0.25
0.24
0.22
0.21
0.19
0.19
0.25
0.23
0.20
0.19
0.17
0.16
0.14
0.13
0.13
0.12
0.12
0.11
0.10
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.06
0.32
0.29
0.26
0.24
0.22
0.20
0.18
0.17
0.16
0.16
0.15
0.13
0.12
0.12
0.12
0.11
0.10
0.10
0.10
0.10
0.10
0.10
0.10
0.09
0.09
0.08
0.64
0.57
0.51
0.46
0.43
0.39
0.34
0.30
0.29
0.29
0.27
0.25
0.22
0.21
0.20
0.20
0.18
0.17
0.18
0.18
0.18
0.17
0.17
0.16
0.15
0.14
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.24
0.21
0.19
0.17
0.16
0.14
0.13
0.12
0.11
0.10
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
0.05
0.31
0.28
0.24
0.22
0.20
0.18
0.17
0.15
0.14
0.13
0.13
0.12
0.11
0.10
0.10
0.10
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.63
0.55
0.49
0.43
0.39
0.36
0.32
0.28
0.26
0.25
0.24
0.22
0.20
0.19
0.18
0.17
0.16
0.15
0.15
0.15
0.14
0.14
0.14
0.14
0.13
0.12
TAMS/MCA
-------
TABLE S-S: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL FOR TRI+ PCBS- OBSOLETE TABLE
REVISED
River Mile 152 River Mile 113 River Mile 90 River Mile SO
95th 95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight) weight) weight) weight) weight) weight) weight) weight) weight) weight)
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002 THIS SPECIES IS NOW COMPARED TO A FIELD-BASED NOAEL ONLY
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
TAMS/MCA
-------
TABLE 5-6: RATIO OF PREDICTED PUMPKINSEED CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
REVISED
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
1.4
1.1
0.9
1.2
0.9
0.6
0.6
0.6
0.7
0.5
0.5
0.4
0.4
0.5
0.4
0.3
0.3
0.4
0.4
0.4
0.4
0.4
0.3
0.2
0.22
0.24
2.1
1.5
1.3
1.7
1.4
1.0
0.8
0.9
1.0
0.8
0.8
0.5
0.6
0.7
0.6
0.5
0.5
0.5
0.6
0.5
0.6
0.5
0.5
0.3
0.3
0.3
5.4
3.6
3.3
4.2
3.5
2.6
2.2
2.1
2.4
2.1
1.9
1.4
1.4
1.6
1.3
1.2
1.0
1.2
1.4
1.3
1.4
1.2
1.1
0.8
0.8
0.8
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.9
0.8
0.6
0.7
0.6
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3 .
0.2
0.2
0.2
0.3
0.2
0.2
0.2
0.2
0.19
0.18
0.18
1.3
1.1
0.9
1.0
0.9
0.7
0.6
0.6
0.6
0.6
0.5
0.4
0.4
0.4
0.4
0.4
0.3
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
3.5
2.9
2.4
2.6
2.4
2.0
1.6
1.5
1.6
1.5
1.3
1.1
1.0
1.1
1.0
0.9
0.8
0.8
0.9
0.9
0.9
0.9
0.8
0.7
0.6
0.6
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.7
0.6
0.5
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.21
0.20
0.18
0.18
0.18
0.18
0.18
0.17
0.17
0.16
0.14
0.14
1.0
0.9
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.21
0.20
2.7
2.4
2.1
1.8
1.7
1.6
1.4
1.2
1.1
1.1
1.0
0.9
0.8
0.8
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 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.7
0.6
0.5
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.2
0.2
0.21
0.19
0.19
0.18
0.16
0.16
0.15
0.15
0.15
0.15
0.14
0.14
0.13
0.12
1.0
0.9
0.8
0.7
0.6
0.6
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
0.21
0.20
0.19
0.18
2.7
2.3
2.1
1.8
1.6
1.5
1.3
1.2
1.0
1.0
0.9
0.9
0.8
0.7
0.7
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-7: RATIO OF PREDICTED PUMPKINSEED CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
REVISED
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
0.7
0.5
0.4
0.6
0.5
0.3
0.3
0.3
0.3
0.3
0.3
0.18
0.20
0.24
0.19
0.15
0.15
0.18
0.21
0.18
0.20
0.17
0.16
0.11
0.11
0.11
1.0
0.7
0.6
0.8
0.7
0.5
0.4
0.4
0.5
0.4
0.4
0.3
0.3
0.3
0.3
0.23
0.23
0.3
0.3
0.3
0.3
0.3
0.23
0.17
0.16
0.17
2.6
1.8
1.6
2.0
1.7
1.3
1.0
1.0
1.2
1.0
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
0.4
0.4
0.4
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.4
0.4
0.3
0.3
0.3
0.2
0.20
0.20
0.21
0.18
0.17
0.14
0.14
0.14
0.14
0.12
0.11
0.12
0.12
0.12
0.12
0.11
0.11
0.09
0.09
0.08
0.6
0.6
0.4
0.5
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.21
0.20
0.21
0.21
0.18
0.16
0.17
0.18
0.17
0.18
0.17
0.16
0.14
0.13
0.12
1.7
1.4
1.2
1.2
1.1
0.9
0.8
0.7
0.8
0.7
0.7
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
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.3
0.3
0.2
0.2
0.21
0.19
0.17
0.15
0.15
0.14
0.13
0.12
0.11
0.10
0.10
0.09
0.09
0.09
0.09
0.09
0.09
0.08
0.08
0.08
0.07
0.07
0.5
0.4
0.4
0.3
0.3
0.3
0.2
0.2
0.22
0.21
0.20
0.18
0.16
0.15
0.15
0.14
0.13
0.13
0.13
0.13
0.13
0.12
0.12
0.11
0.10
0.10
1.3
1.1
1.0
0.9
0.8
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
0.3
0.3
0.3
0.28
0.26
0.24
0.23
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.3
0.3
0.3
0.22
0.20
0.18
0.16
0.15
0.13
0.13
0.12
0.11
0.10
0.09
0.09
0.09
0.08
0.08
0.07
0.07
0.07
0.07
0.07
0.07
0.06
0.06
0.5
0.4
0.4
0.3
0.3
0.3
0.2
0.21
0.20
0.19
0.18
0.16
0.15
0.14
0.13
0.12
0.11
0.11
0.11
0.11
0.11
0.10
0.10
0.10
0.09
0.09
1.3
1.1
1.0
0.9
0.8
0.7
0.6
0.6
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.28
0.27
0.26
0.25
0.25
0.25
0.24
0.23
0.22
0.21
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-8: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
REVISED
River Mile 152
25th Median
(mg/kg wet (mg/kg wet
Year 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
0.18
0.13
0.12
0.15
0.12
0.09
0.09
0.08
0.09
0.08
0.07
0.06
0.06
0.07
0.06
0.051
0.044
0.06
0.05
0.052
0.06
0.049
0.045
0.039
0.039
0.041
0.26
0.18
0.17
0.21
0.18
0.14
0.13
0.11
0.13
0.12
0.10
0.08
0.08
0.10
0.08
0.07
0.06
0.08
0.08
0.08
0.08
0.07
0.06
0.056
0.056
0.057
95th
Percentile
(mg/kg wet
weight)
0.64
0.41
0.40
0.50
0.41
0.32
0.29
0.25
0.29
0.28
0.22
0.17
0.18
0.21
0.17
0.16
0.13
0.17
0.17
0.16
0.18
0.15
0.14
0.12
0.12
0.12
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
0.06
0.05
0.04
0.05
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.024
0.023
0.024
0.023
0.021
0.019
0.020
0.021
0.020
0.020
0.020
0.019
0.017
0.016
0.015
0.09
0.08
0.06
0.07
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
0.03
0.03
0.024
0.022
0.022
0.20
0.17
0.15
0.16
0.14
0.12
0.10
0.09
0.10
0.10
0.09
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
0.04
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.05
0.04
0.04
0.04
0.03
0.03
0.03
0.02
0.02
0.02
0.022
0.020
0.018
0.018
0.017
0.016
0.015
0.015
0.015
0.015
0.015
0.014
0.014
0.013
0.012
0.012
0.07
0.06
0.05
0.05
0.05
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.02
0.024
0.023
0.021
0.021
0.021
0.021
0.021
0.020
0.020
0.019
0.018
0.017
0.16
0.14
0.13
0.11
0.11
0.10
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
0.03
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.04
0.04
0.04
0.03
0.03
0.03
0.02
0.02
0.021
0.020
0.019
0.017
0.016
0.015
0.015
0.014
0.013
0.013
0.012
0.012
0.012
0.012
0.012
0.011
0.011
0.010
0.06
0.06
0.05
0.05
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.02
0.023
0.021
0.020
0.020
0.018
0.018
0.018
0.017
0.017
0.017
0.016
0.016
0.015
0.015
0.16
0.14
0.12
0.11
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
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
0.03
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-9: RATIO OF PREDICTED SPOTTAIL SHINER CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
REVISED
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
0.009
0.007
0.006
0.008
0.006
0.005
0.005
0.004
0.005
0.004
0.004
0.003
0.003
0.004
0.003
0.003
0.002
0.003
0.003
- 0.003
0.003
0.003
0.002
0.002
0.002
0.002
0.014
0.010
0.009
0.011
0.009
0.007
0.007
0.006
0.007
0.006
0.005
0.004
0.004
0.005
0.004
0.004
0.003
0.004
0.004
0.004
0.004
0.004
0.003
0.003
0.003
0.003
0.033
0.022
0.021
0.026
0.021
0.017
0.015
0.013
0.015
0.014
0.011
0.009
0.009
0.011
0.009
0.009
0.007
0.009
0.009
0.009
0.009
0.008
0.007
0.006
0.006
0.006
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
0.003
0.003
0.002
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
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
0.002
0.002
0.002
0.001
0.001
0.002
0.001
0.001
0.001
0.001
0.001
0.001
0.001
0.011
0.009
0.008
0.008
0.008
0.006
0.005
0.005
0.005
0.005
0.004
0.004
0.003
0.004
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.003
0.002
0.002
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
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
0.001
0.001
0.001
0.001
0.001
0.004
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
0.001
0.008
0.007
0.007
0.006
0.006
0.005
0.004
0.004
0.004
0.004
0.003
0.003
0.003
0.003
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) 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
0.001
0.001
0.001
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
0.001
0.001
0.008
0.007
0.006
0.006
0.005
0.005
0.004
0.004
0.003
0.003
0.003
0.003
0.003
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
0.002
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-10: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL FOR TRI+ PCBS
REVISED
River Mile 152 River Mile 1 1 3 River Mile 90
95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight) weight) weight) weight) 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
16
15
14
14
14
13
12
12
12
11
11
10
10
10
10
9.7
9.4
9.7
9.5
9.2
9.3
8.9
8.8
8.6
8.5
8.4
20
18
17
18
17
15
15
15
15
14
13
13
13
13
13
12
12
12
12
12
11
11
11
11
11
11
34
30
29
29
28
26
24
24
24
24
22
21
21
21
21
20
20
19
19
19
19
19
18
18
18
17
13
12
11
11
10
9.9
9.4
9.3
9.2
8.9
8.7
8.3
8.2
8.1
8.0
7.8
7.6
7.6
7.5
7.4
7.3
7.1
7.0
6.9
6.8
6.7
15
15
14
13
13
12
12
12
11
11
11
10
10
10
10
9.8
9.7
9.5
9.4
9.3
9.2
9.0
8.9
8.8
8.7
8.6
26
24
23
22
21
21
19
19
19
18
18
17
17
17
16
16
16
15
15
15
15
15
14
14
14
14
10
9
9
9
8
7.8
7.5
7.3
7.1
7.0
6.8
6.5
6.3
6.2
6.1
6.0
5.9
5.8
5.7
5.6
5.6
5.5
5.4
5.3
5.2
5.2
12
12
11
11
10
9.7
9.3
9.0
8.9
8.7
8.4
8.2
8.0
7.8
7.7
7.6
7.4
7.3
7.2
7.1
7.0
6.9
6.8
6.7
6.6
6.5
20
19
18
17
17
16
15
15
14
14
14
13
13
13
13
12
12
12
12
12
11
11
11
11
11
11
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
8
7
7
7
6.3
6.0
5.8
5.6
5.4
5.3
5.1
5.0
4.8
4.7
4.6
4.6
4.5
4.4
4.3
4.3
4.2
4.2
4.1
4.0
4.0
3.9
10
9
9
8
8
7.4
7.1
6.9
6.7
6.6
6.4
6.2
6.0
5.9
5.8
5.7
5.6
5.6
5.5
5.4
5.3
5.3
5.2
5.1
5.0
5.0
16
15
14
13
13
12
12
11
11
11
10
10
10
9.7
9.6
9.4
9.2
9.1
8.9
8.8
8.7
8.6
8.4
8.3
8.2
8.0
TAMS/MCA
-------
TABLE 5-11: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL FOR TRI+ PCBS
REVISED
River Mile 152 River Mile 113
95th 95th
25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight) 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
3.3
3.0
2.8
2.9
2.8
2.6
2.4
2.4
2.5
2.3
2.2
2.1
2.1
2.1
2.1
2.0
1.9
2.0
1.9
1.9
1.9
1.8
1.8
1.8
1.7
1.7
4.1
3.8
3.4
3.7
3.4
3.2
3.0
3.0
3.1
2.9
2.8
2.7
2.6
2.6
2.6
2.5
2.5
2.4
2.4
2.4
2.3
2.3
2.3
2.2
2.2
2.2
7.0
6.1
5.8
6.0
5.7
5.4
5.0
4.9
4.9
4.8
4.6
4.4
4.3
4.3
4.2
4.1
4.0
3.9
4.0
3.9
3.9
3.8
3.7
3.7
3.6
3.6
2.6
2.4
2.2
2.2
2.1
2.0
1.9
1.9
1.9
1.8
1.8
1.7
1.7
1.7
1.6
1.6
1.6
1.5
1.5
1.5
1.5
1.5
1.4
1.4
1.4
1.4
3.2
3.0
2.8
2.7
2.6
2.5
2.4
2.4
2.3
2.3
2.2
2.1
2.1
2.1
2.0
2.0
2.0
1.9
1.9
1.9
1.9
1.8
1.8
1.8
1.8
1.8
5.3
4.9
4.6
4.5
4.4
4.2
3.9
3.8
3.8
3.8
3.6
3.5
3.4
3.4
3.3
3.3
3.2
3.2
3.1
3.1
3.1
3.0
3.0
2.9
2.9
2.8
River Mile 90 River Mile 50
95th 95th
25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight) weight) weight) weight)
2.0
1.9
1.8
1.7
1.7
1.6
1.5
1.5
1.5
1.4
1.4
1.3
1.3
1.3
1.2
1.2
1.2
1.2
1.2
1.2
1.1
1.1
1.1
1.1
1.1
1.1
2.5
2.4
2.2
2.1
2.1
2.0
1.9
1.8
1.8
1.8
1.7
1.7
1.6
1.6
1.6
1.5
1.5
1.5
1.5
1.5
1.4
1.4
1.4
1.4
1.4
1.3
4.2
3.9
3.7
3.6
3.4
3.3
3.1
3.0
2.9
2.9
2.8
2.7
2.7
2.6
2.6
2.5
2.5
2.4
2.4
2.4
2.3
2.3
2.3
2.2
2.2
2.2
1.6
1.5
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
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
2.0
1.9
1.7
1.7
1.6
1.5
1.5
1.4
1.4
1.3
1.3
1.3
1.2
1.2
1.2
1.2
1.
1.
1.
1.
1.
1.1
1.1
1.0
1.0
1.0
3.3
3.1
2.9
2.7
2.6
2.5
2.4
2.3
2.2
2.2
2.1
2.1
2.0
2.0
2.0
1.9
1.9
1.8
1.8
1.8
1.8
1.7
1.7
1.7
1.7
1.6
TAMS/MCA
-------
TABLE 5-12: RATIO OF PREDICTED BROWN BULLHEAD CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
REVISED
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
0.04
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
0.021
0.021
0.020
0.020
0.020
0.05
0.04
0.04
0.04
0.04
0.04
0.03
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
0.03
0.03
0.03
0.03
0.03
0.09
0.08
0.07
0.08
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
0.05
0.05
0.05
0.05
0.05
0.05
0.04
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
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
0.019
0.019
0.018
0.018
0.018
0.018
0.017
0.017
0.017
0.017
0.016
0.016
0.016
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
0.02
0.02
0.02
0.02
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
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
0.04
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
0.017
0.017
0.016
0.016
0.015
0.015
0.015
0.014
0.014
0.014
0.014
0.014
0.013
0.013
0.013
0.013
0.013
0.012
0.012
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
0.017
0.017
0.016
0.016
0.016
0.02
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
0.03
0.03
0.03
0.03
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
0.016
0.016
0.015
0.014
0.014
0.013
0.013
0.012
0.012
0.012
0.011
0.011
0.011
0.011
0.011
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.009
0.009
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.015
0.014
0.014
0.014
0.014
0.013
0.013
0.013
0.013
0.013
0.013
0.012
0.012
0.012
0.012
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
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
REVISED
River Mile 152
95th
25th Median Percentile
(mg/kg wet (rag/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
0.02
0.02
0.01
0.02
0.014
0.013
0.012
0.013
0.013
0.012
0.011
0.011
0.011
0.011
0.011
0.010
0.010
0.010
0.010
0.010
0.010
0.009
0.009
0.009
0.009
0.009
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.015
0.014
0.014
0.014
0.014
0.013
0.013
0.013
0.013
0.013
0.012
0.012
0.012
0.012
0.012
0.011
0.011
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
0.02
0.02
0.02
0.02
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
0.013
0.013
0.012
0.011
0.011
0.010
0.010
0.010
0.010
0.009
0.009
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.008
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.02
0.02
0.01
0.01
0.014
0.013
0.012
0.012
0.012
0.012
0.011
0.011
0.011
0.011
0.011
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.009
0.009
0.009
0.009
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
0.017
0.017
0.017
0.016
0.016
0.016
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.011
0.010
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
0.006
0.006
0.006
0.006
0.006
0.006
0.006
0.005
0.013
0.012
0.012
0.011
0.011
0.010
0.010
0.010
0.009
0.009
0.009
0.009
0.008
0.008
0.008
0.008
0.008
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.007
0.007
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.02
0.015
0.015
0.015
0.014
0.014
0.014
0.014
0.013
0.013
0.013
0.013
0.013
0.012
0.012
0.012
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
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
0.005
0.005
0.004
0.004
0.004
0.004
0.004
0.004
0.010
0.010
0.009
0.009
0.008
0.008
0.008
0.007
0.007
0.007
0.007
0.007
0.006
0.006
0.006
0.006
0.006
0.006
0.006
0.006
0.006
0.006
0.005
0.005
0.005
0.005
0.02
0.02
0.02
0.02
0.01
0.014
0.013
0.013
0.012
0.012
0.012
0.012
0.011
0.011
0.011
0.011
0.010
0.010
0.010
0.010
0.010
0.010
0.010
0.009
0.009
0.009
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-14: RATIO OF PREDICTED WHITE PERCH CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
REVISED
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
0.7
0.6
0.5
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
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
1.0
0.9
0.8
0.9
0.8
0.7
0.6
0.7
0.7
0.6
0.6
0.5
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
2.3
1.9
1.7
1.8
1.7
1.5
1.4
1.3
1.4
1.3
1.2
1.1
1.1
1.1
1.1
1.1
1.0
1.0
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
River Mile 1 13
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) 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
0.3
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
0.6
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
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
1.4
1.3
1.2
1.1
1.1
1.0
0.9
0.9
0.9
0.9
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
River Mile 90 River Mile 50
95th 95th
25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight) weight) weight) 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
0.2
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
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
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
0.5
0.5
0.5
0.5
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
O.I
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
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
0.9
0.9
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
0.4
0.4
0.4
0.4
0.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-15: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL FOR TRI+ PCBS
REVISED
River Mile 152
95th
25th Median Percentile
(ing/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
5.2
4.9
3.7
5.2
3.8
3.0
2.6
3.2
3.2
2.5
2.5
1.9
2.1
2.4
2.2
1.7
1.6
1.8
2.0
1.8
1.9
1.7
1.7
1.3
1.2
1.3
7.1
6.8
5.2
7.2
5.2
4.1
3.6
4.4
4.3
3.4
3.5
2.6
2.9
3.3
2.8
2.3
2.1
2.4
2.8
2.5
2.6
2.3
2.2
1.8
1.7
1.7
16.9
13.9
12.7
15.4
12.3
9.0
7.8
8.2
8.8
7.4
7.1
5.3
5.6
6.8
5.4
4.6
4.5
5.1
5.6
5.3
5.2
4.9
4.6
3.6
3.4
3.5
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
3.8
3.5
2.8
3.0
2.7
2.3
2.1
2.1
2.1
1.9
1.9
1.6
1.5
1.6
1.5
1.3
1.2
13
1.3
1.3
1.3
1.3
1.2
1.1
1.0
1.0
5.3
4.9
3.8
4.2
3.7
3.2
2.8
2.8
2.8
2.6
2.5
2.1
2.0
2.1
2.0
1.8
1.6
1.7
1.8
1.8
1.8
1.7
1.6
1.5
1.4
1.4
12.2
10.8
8.5
9.2
8.3
6.9
5.9
5.7
5.8
5.5
5.0
4.2
4.0
4.2
4.0
3.7
3.3
3.4
3.6
3.5
3.5
3.4
3.3
2.9
2.7
2.7
River Mile 90 River Mile 50
95th 95th
25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight) weight) weight) weight)
3.0
2.7
2.4
2.3
2.1
1.9
1.7
1.6
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
1.0
0.9
0.9
0.8
0.8
4.2
3.8
3.3
3.1
2.8
2.6
2.3
2.2
2.1
2.0
1.9
1.7
1.6
1.5
1.5
1.4
1.3
1.3
1.3
1.3
1.3
1.3
1.2
1.2
1.1
1.1
9.5
8.5
7.3
6.7
6.1
5.5
4.8
4.5
4.3
4.1
3.9
3.4
3.2
3.1
3.0
2.8
2.6
2.6
2.6
2.6
2.6
2.5
2.5
2.3
2.2
2.1
2.8
2.5
2.2
2.0
1.9
1.7
1.5
1.4
1.3
1.3
1.2
1.1
1.1
1.0
1.0
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
0.7
3.9
3.5
3.1
2.8
2.6
2.4
2.1
2.0
1.8
1.8
1.7
1.5
1.4
1.3
1.3
1.2
.1
.1
.1
.1
.1
.0
1.0
1.0
0.9
0.9
9.4
8.2
7.2
6.4
5.8
5.2
4.6
4.2
3.9
3.7
3.4
3.2
2.9
2.7
2.6
2.5
2.2
2.3
2.2
2.2
2.2
2.1
2.1
2.0
1.9
1.8
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-16: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL FOR TRI+ PCBS
REVISED
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
1.1
1.0
0.8
1.1
0.8
0.6
0.5
0.7
0.7
0.5
0.5
0.4
0.4
0.5
0.4
0.3
0.3
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
1.4
1.4
1.1
1.5
1.1
0.8
0.7
0.9
0.9
0.7
0.7
0.5
0.6
0.7
0.6
0.5
0.4
0.5
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
3.5
2.8
2.6
3.1
2.5
1.8
1.6
1.7
1.8
1.5
1.4
1.1
1.1
1.4
1.1
0.9
0.9
1.0
1.2
1.1
1.1
1.0
0.9
0.7
0.7
0.7
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.8
0.7
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
0.2
0.3
0.3
0.3
0.3
0.3
0.2
0.2
0.2
0.2
1.1
1.0
0.8
0.9
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
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
2.5
2.2
1.7
1.9
1.7
1.4
1.2
1.2
1.2
1.1
1.0
0.8
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
River Mile 90 River Mile 50
95th 95th
25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight) weight) weight) weight)
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
0.2
0.2
0.9
0.8
0.7
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
0.2
0.2
0.2
1.9
1.7
1.5
1.4
1.3
1.1
1.0
0.9
0.9
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
0.6
0.5
0.5
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
0.1
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
1.9
1.7
1.5
1.3
1.2
1.1
0.9
0.8
0.8
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
0.4
0.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-17: RATIO OF PREDICTED WHITE PERCH CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
REVISED
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
1.0
0.9
0.8
0.9
0.8
0.7
0.6
0.7
0.7
0.6
0.6
0.5
0.5
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
1.5
1.3
1.1
1.3
1.2
1.0
0.9
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
0.7
0.7
0.7
0.6
0.6
0.6
3.5
2.9
2.6
2.7
2.6
2.3
2.1
2.0
2.2
2.0
1.9
1.7
1.7
1.7
1.7
1.6
1.6
1.6
1.6
1.5
1.6
1.5
1.5
1.4
1.4
1.4
River Mile 1 13
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
1.0
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
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
1.5
1.4
1.2
1.3
1.2
1.1
1.0
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
0.8
0.7
0.7
0.7
0.7
0.7
3.5
3.1
2.9
2.7
2.7
2.5
2.3
2.2
2.1
2.2
2.0
1.9
1.9
1.8
1.8
1.8
1.7
1.7
1.7
1.7
1.7
1.6
1.6
1.6
1.5
1.5
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
0.7
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
0.4
0.4
0.4
0.4
0.4
0.4
1.2
1.1
1.0
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
0.6'
0.6
0.6
0.6
0.6
0.5
0.5
0.5
2.7
2.5
2.3
2.2
2.0
1.9
1.8
1.7
1.7
1.6
1.6
1.5
1.5
1.4
1.4
1.4
1.3
1.3
1.3
1.3
1.2
1.2
1.2
1.2
1.2
1.1
River Mile 50
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
0.4
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
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
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
0.4
0.4
2.3
2.1
1.9
1.8
1.7
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
1.0
1.0
0.9
0.9
0.9
0.9
0.9
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-18: RATIO OF PREDICTED WHITE PERCH CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
REVISED
River Mile 152
95th
25th Median Percentile
(rag/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
0.5
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
0.2
0.2
0.7
0.6
0.5
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
0.4
0.4
0.3
0.4
0.3
0.3
0.3
0.3
0.3
1.7
1.4
1.3
1.3
1.3
1.1
1.0
1.0
1.1
1.0
0.9
0.8
0.8
0.8
0.8
0.8
0.7
0.8
0.8
0.7
0.8
0.7
0.7
0.7
0.7
0.7
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
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
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.2
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
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
1.7
1.5
1.4
1.3
1.3
1.2
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
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
0.3
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
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
0.3
0.3
0.3
0.3
1.3
1.2
1.1
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
0.6
0.6
0.6
0.6
0.5
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.3
0.3
0.3
0.3
0.2
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
0.15
0.15
0.14
0.14
0.14
0.14
0.13
0.13
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
0.2
0.2
0.2
0.2
1.1
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
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-19: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
REVISED
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
3.1
2.9
2.2
3.1
2.2
1.8
1.6
1.9
1.9
1.5
1.5
1.1
1.3
1.4
1.3
1.0
0.9
1.1
1.2
1.1
1.1
1.0
1.0
0.8
0.7
0.7
4.3
4.1
3.1
4.3
3.1
2.5
2.2
2.6
2.6
2.1
2.1
1.5
1.7
2.0
1.7
1.4
1.3
1.5
1.7
1.5
1.6
1.4
1.3
1.1
1.0
1.1
10.2
8.6
7.8
9.5
7.4
5.5
4.8
5.0
5.3
4.6
4.3
3.2
3.4
4.1
3.3
2.8
2.7
3.1
3.4
3.2
3.2
3.0
2.8
2.2
2.1
2.1
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
1.9
1.8
1.4
1.6
1.4
1.2
1.0
1.1
1.1
1.0
0.9
0.8
0.8
0.8
0.8
0.7
0.6
0.6
0.7
0.7
0.7
0.6
0.6
0.5
0.5
0.5
2.7
2.5
2.0
2.2
1.9
1.6
1.4
1.5
1.5
1.4
1.3
1.1
1.1
1.1
1.0
0.9
0.8
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.7
0.7
6.3
5.6
4.5
4.8
4.3
3.7
3.1
3.0
3.1
2.9
2.6
2.2
2.1
2.2
2.1
2.0
1.7
1.8
1.9
1.8
1.8
1.8
1.7
1.5
1.4
1.4
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
1.5
1.4
1.2
1.1
1.1
1.0
0.9
0.8
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.4
2.1
1.9
1.7
1.6
1.5
1.3
1.2
1.1
1.1
1.0
1.0
0.9
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
0.5
4.9
4.4
3.8
3.6
3.2
2.9
2.5
2.3
2.3
2.2
2.0
1.8
1.7
1.6
1.6
1.5
1.3
1.4
1.4
1.4
1.4
1.3
1.3
1.2
1.1
1.1
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
1.4
1.3
1.1
1.0
1.0
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.5
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
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
0.5
0.5
0.5
0.5
0.5
0.5
0.5
4.9
4.3
3.7
3.3
3.0
2.7
2.4
2.2
2.0
1.9
1.8
1.6
1.5
1.4
1.4
1.3
1.2
1.2
1.2
1.1
1.1
1.1
1.1
1.0
1.0
1.0
Bold values indicate exceeoances
TAMS/MCA
-------
TABLE 5-20: RATIO OF PREDICTED YELLOW PERCH CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
REVISED
River Mile 152
95th
25th Median Percentile
(mg/kg wet (rag/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
1.5
1.4
1.1
1.5
1.1
0.9
0.8
0.9
0.9
0.7
0.7
0.6
0.6
0.7
0.6
0.5
0.5
0.5
0.6
0.5
0.5
0.5
0.5
0.4
0.4
0.4
2.1
2.0
1.5
2.1
1.5
1.2
1.1
1.3
1.2
1.0
1.0
0.7
0.8
0.9
0.8
0.7
0.6
0.7
0.8
0.7
0.8
0.7
0.6
0.5
0.5
0.5
4.9
4.1
3.8
4.6
3.6
2.7
2.3
2.4
2.6
2.2
2.1
1.6
1.6
2.0
1.6
1.4
1.3
1.5
1.7
1.6
1.5
1.5
1.3
1.1
1.0
1.0
River Mile 113
95th
Median Percentile
25th (mg/kg (mg/kg wet (mg/kg wet
wet weight) weight) weight)
0.9
0.9
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
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
0.2
1.3
1.2
0.9
1.0
0.9
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
3.1
2.7
2.2
2.3
2.1
1.8
1.5
1.5
1.5
1.4
1.3
1.1
1.0
1.0
1.0
0.9
0.8
0.9
0.9
0.9
0.9
0.9
0.8
0.7
0.7
0.7
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.7
0.7
0.6
0.6
0.5
0.5
0.4
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
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
2.4
2.1
1.8
1.7
1.6
1.4
1.2
1.1
1.1
1.0
1.0
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
0.5
River Mile 50
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
0.4
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
0.20
0.20
0.19
0.19
0.18
0.17
0.17
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
0.3
0.2
0.2
0.2
2.4
2.1
1.8
1.6
1.5
1.3
1.2
1.1
1.0
0.9
0.9
0.8
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
0.5
0.5
0.5
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-21: RATIO OF PREDICTED LARGEMOUTH BASS CONCENTRATIONS TO
FIELD-BASED NOAEL FOR TRI+ PCBS
REVISED
River Mile 152 River Mile 113 River Mile 90 River Mile 50
95th 95th 95th 95th
25th Median Percentile 25th Median Percentile 25th Median Percentile 25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet (rng/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet (mg/kg wet
Year weight) weight) weight) weight) weight) weight) weight) weight) weight) 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
31
22
18
25
20
16
14
14
15
14
12
9.2
9.2
11
9.3
8.4
7.6
8.1
9.5
8.1
9.1
8.3
7.4
6.6
6.1
6.0
35
25
21
27
23
19
16
15
17
16
14
10
10
12
10
9.6
8.6
9.1
11
9.1
10
9.2
8.4
7.6
6.8
6.7
52
35
31
37
34
28
23
21
24
23
19
15
14
17
15
14
12
13
15
13
14
13
12
11
9.9
9.4
24
21
17
18
16
14
12
11
12
11
10
9.1
8.1
8.5
8.3
7.8
6.9
6.8
7.4
7.1
7.3
7.0
6.5
6.2
5.6
5.4
27
24
20
20
19
16
14
13
13
13
12
10
9.3
9.6
9.3
8.8
7.8
7.8
8.4
8.0
8.3
7.8
7.4
7.0
6.4
6.1
40
35
30
29
27
24
20
19
19
19
17
15
14
14
14
13
11
11
12
12
12
11
11
10
9.3
9.0
19
17
15
13
12
11
10
9.1
8.6
8.4
8.0
7.4
6.7
6.2
6.0
5.9
5.5
5.2
5.2
5.2
5.2
5.1
5.0
4.8
4.5
4.2
21
19
17
15
14
13
11
10
9.8
9.5
9.0
8.3
7.5
7.1
6.9
6.6
6.2
5.9
5.9
5.9
5.9
5.7
5.6
5.4
5.0
4.8
30
28
25
22
20
19
17
15
14
14
13
12
11
10
10
9.6
9.0
8.6
8.7
8.6
8.6
8.5
8.2
7.8
7.3
7.0
18
16
14
13
11
10
9.4
8.5
7.9
7.5
7.1
6.6
6.1
5.7
5.4
5.1
4.8
4.6
4.5
4.4
4.4
4.3
4.2
4.1
3.9
3.7
21
18
16
14
13
12
11
9.6
8.9
8.4
8.0
7.5
6.9
6.4
6.1
5.8
5.4
5.2
5.1
5.0
4.9
4.8
4.8
4.6
4.4
4.2
30
27
23
21
19
17
15
14
13
12
11
11
9.8
9.2
8.7
8.3
7.9
7.5
7.4
7.2
7.2
7.0
6.9
6.7
6.4
6.1
TAMS/MCA
-------
TABLE 5-22: RATIO OF PREDICTED LARGEMOUTH BASS CONCENTRATIONS TO
LABORATORY-DERIVED NOAEL ON A TEQ BASIS
REVISED
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 95th
(mg/kg Median Percentile
wet (mg/kg wet (mg/kg wet
weight) weight) weight)
5.0
3.6
2.9
3.9
3.3
2.6
2.2
2.2
2.4
2.2
2.0
1.5
1.5
1.7
1.5
1.4
1.2
1.3
1.5
1.3
1.5
1.3
1.2
1.1
1.0
1.0
6.7
4.7
4.0
5.1
4.4
3.5
3.0
2.9
3.2
3.0
2.6
2.0
2.0
2.2
2.0
1.8
1.6
1.7
2.0
1.7
2.0
1.7
1.6
1.4
1.3
1.3
13.3
9.1
8.1
9.6
8.7
7.2
5.9
5.5
6.3
6.0
5.0
4.0
3.8
4.4
3.9
3.6
3.1
3.4
3.8
3.4
3.8
3.4
3.1
2.9
2.6
2.5
River Mile 113
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
2.4
2.1
1.7
1.8
1.7
1.4
1.2
1.1
1.2
1.1
1.1
0.9
0.8
0.9
0.8
0.8
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.6
0.6
0.5
3.3
2.8
2.3
2.4
2.2
1.9
1.7
1.5
1.6
1.6
1.4
1.2
1.1
1.1
1.1
1.1
0.9
0.9
1.0
1.0
1.0
0.9
0.9
0.8
0.8
0.7
6.7
5.8
4.9
4.7
4.6
4.0
3.4
3.1
3.2
3.1
2.9
2.5
2.3
2.3
2.2
2.1
1.9
1.9
2.0
2.0
2.0
1.9
1.8
1.7
1.5
1.5
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
1.9
1.7
1.5
1.3
1.3
1.1
1.0
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
0.5
0.4
2.5
2.3
2.0
1.8
1.7
1.5
1.4
1.2
1.2
1.1
1.1
1.0
0.9
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
5.1
4.6
4.1
3.7
3.4
3.1
2.7
2.5
2A
2.3
2.2
2.0
1.8
1.7
1.7
1.6
1.5
1.4
1.5
1.4
1.4
1.4
1.4
1.3
1.2
1.2
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
1.8
1.6
1.4
1.3
1.1
1.1
0.9
0.9
0.8
0.8
0.7
0.7
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
2.5
2.2
1.9
1.7
1.6
1.4
1.3
1.2
1.1
1.0
1.0
0.9
0.8
0.8
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
0.5
0.5
5.0
4.4
3.9
3.4
3.1
2.9
2.6
2.3
2.1
2.0
1.9
1.8
1.7
1.5
1.5
1.4
1.3
1.3
1.2
1.2
1.2
1.2
1.2
1.1
1.1
1.0
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-23: RATIO OF PREDICTED LARGEMOUTH BASS CONCENTRATIONS TO
LABORATORY-DERIVED LOAEL ON A TEQ BASIS
REVISED
River Mile 1S2
95th
25th Median Percentile
(mg/kg wet (rag/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
2.4
1.7
1.4
1.9
1.6
1.3
1.1
1.1
1.2
1.1
1.0
0.7
0.7
0.8
0.7
0.7
0.6
0.6
0.7
. 0.6
0.7
0.6
0.6
0.5
0.5
0.5
3.2
2.3
1.9
2.5
2.1
1.7
1.4
1.4
1.5
1.4
1.3
1.0
1.0
1.1
1.0
0.9
0.8
0.8
1.0
0.8
1.0
0.8
0.8
0.7
0.6
0.6
6.4
4.4
3.9
4.6
4.2
3.5
2.9
2.6
3.0
2.9
2.4
1.9
1.8
2.1
1.9
1.8
1.5
1.6
1.9
1.6
1.8
1.7
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)
1.2
1.0
0.8
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
'0.3
0.3
0.4
0.3
0.4
0.3
0.3
0.3
0.3
0.3
1.6
1.4
1.1
1.2
1.1
0.9
0.8
0.7
0.8
0.8
0.7
0.6
0.5
0.6
0.5
0.5
0.5
0.4
0.5
0.5
0.5
0.5
0.4
0.4
0.4
0.4
3.3
2.8
2.4
2.3
2.2
1.9
1.6
1.5
1.6
1.5
1.4
1.2
1.1
1.1
1.1
1.0
0.9
0.9
1.0
0.9
1.0
0.9
0.9
0.8
0.7
0.7
River Mile 90
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.9
0.8
0.7
0.6
0.6
0.6
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
1.2
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
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
2.5
2.2
2.0
1.8
1.6
1.5
1.3
1.2
1.1
1.1
1.1
1.0
0.9
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
River Mile 50
95th
25th Median Percentile
(mg/kg wet (mg/kg wet (mg/kg wet
weight) weight) weight)
0.9
0.8
0.7
0.6
0.6
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
1.2
1.1
0.9
0.8
0.8
0.7
0.6
0.6
0.5
0.5
0.5
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.3
0.2
2.4
2.1
1.9
1.7
1.5
1.4
1.2
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
Bold values indicate exceedances
TAMS/MCA
-------
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
REVISED
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
0.08
0.08
0.07
0.07
0.07
0.06
0.06
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
0.05
0.05
005
0.05
NOAEL
152
95% UCL
0.09
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
0.06
0.06
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
0.06
0.06
0.06
0.06
0.05
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
0.04
0.04
0.04
0.04
0.04
NOAEL
113
95% UCL
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
0.05
0.05
0.05
0.04
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
0.05
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
0.03
0.03
0.03
NOAEL
90
95% UCL
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
0.03
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
0.03
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
0.02
0.02
0.02
NOAEL
50
95% UCL
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
0.03
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.02
0.02
Bold value indicates exceedances
-------
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
REVISED
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
0.09
0.09
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.08
008
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.07
0.07
NOAEL
152
95% UCL
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
0.08
0.08
0.08
0.08
0.08
0.08
0.08
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
0.09
0.09
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
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
NOAEL
113
95% UCL
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
0.07
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
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
0.07
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
0.05
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
0.08
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
0.05
0.05
0.05
0.05
0.05
0.05
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
0.05
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
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% UCI
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
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
Bold value indicates exceedances
-------
TABLE 5-27: RATIO OF MODELED DIETARY DOSE BASED ON F1SHRAND FOR
FEMALE TREE SWALLOW USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
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
0.03
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
0.02
0.02
0.02
0.02
NOAEL
152
95%UCL
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
0.03
0.03
0.03
0.03
0.03
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
0.02
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
0.02
0.02
0.02
0.02
NOAEL
113
95% UCL
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
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
0.02
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
0.01
0.01
0.01
NOAEL
90
95% UCL
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
0.02
0.02
0.02
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
0.01
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
0.01
0.01
0.01
NOAEL
50
95% UCL
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
0.01
0.01
-------
TABLE 5-28: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE TREE SWALLOW USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
200S
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
0.09
0.08
0.08
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
0.07
0.07
NOAEL
152
95% UCL
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.08
0.08
0.08
0.08
0.08
0.08
0.08
0.08
0.08
0.08
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
0.08
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
0.06
0.06
0.06
0.06
0.05
0.05
0.05
0.05
NOAEL
113
95% UCL
0.1
0.1
0.1
0.1
0.08
0.08
0.08
0.07
0.07
0.07
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
0.06
0.06
0.06
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
0.07
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
0.05
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
0.08
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
0.05
0.05
0.05
0.05
0.05
0.05
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
0.05
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
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
0.06
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
0.04
0.04
0.04
0.04
0.04
0.04
0.03
0.03
0.03
0.03
0.03
0.03
-------
TABLE 5-29: RATIO OF MODELED DIETARY DOSE FOR FEMALE MALLARD BASED ON
FISHRAND RESULTS FOR THE TRI+ CONGENERS
REVISED
LOAEL
152
Year Average
1993 0.03
1994 0.03
1995 0.02
1996 0.03
1997 0.02
1998 0.02
1999 0.02
2000 0.02
2001 0.02
2002 0.02
2003 0.02
2004 0.01
2005 0.01
2006 0.01
2007 0.01
2008 0.01
2009 0.01
2010 0.01
2011 0.01
2012 0.01
2013 0.01
2014 0.01
2015 0.01
2016 0.01
2017 0.01
2018 0.01
LOAEL
152
95% UCL
0.03
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
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
0.01
NOAEL
152
Average
0.3
0.3
0.2
0.3
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
0.1
0.1
0.1
NOAEL
152
95% UCL
0.3
0.3
0.2
0.3
0.3
0.2
0.2
0.2
0.2
0.2
0.2
0.1
0.1
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
LOAEL LOAEL
113 113
Average 95% UCL
0.02 0.03
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
0.01 0.01
0.01 0.01
0.009 0.01
0.01 0.01
0.009 0.01
0.009 0.01
0.01 0.01
0.009 0.01
0.009 0.01
0.008 0.009
0.008 0.009
0.008 0.009
NOAEL
113
Average
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.
0.09
0.09
0.08
0.08
0.08
NOAEL LOAEL
113 90
95% UCL Average
0.3 0.02
0.2 0.02
0.2 0.02
0.2 0.01
0.2 0.01
0.2 0.01
0. 0.01
0. 0.01
0. 0.01
0. 0.01
0. 0.009
0. 0.009
0. 0.008
0. 0.008
0. 0.008
0. 0.008
0. 0.007
0. 0.007
0. 0.007
0. 0.007
0. 0.007
0. 0.007
0. 0.007
0.09 0.006
0.09 0.006
0.09 0.006
LOAEL NOAEL
90 90
95% UCL Average
0.02 0.2
0.02 0.2
0.02 0.2
0.02 0.
0.01 0.
0.01 0.
0.01 0.
0.01 0.
0.01 0.
0.01 0.
0.01 0.09
0.009 0.09
0.009 0.08
0.009 0.08
0.009 0.08
0.008 0.08
0.008 0.07
0.008 0.07
0.008 0.07
0.008 0.07
0.008 0.07
0.007 0.07
0.007 0.07
0.007 0.06
0.007 0.06
0.007 0.06
NOAEL
90
95% UCL
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
0.09
0.09
0.08
0.08
0.08
0.08
0.08
0.08
0.07
0.07
0.07
0.07
0.07
LOAEL
50
Average
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.009
0.009
0.008
0.008
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
LOAEL
50
95% UCL
0.02
0.02
0.01
0.01
0.01
0.01
0.01
0.01
0.009
0.009
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
NOAEL NOAEL
50 50
Average 95% UCL
0.2 0.2
0.2 0.2
0.1 0.
0.1 0.
0.1 0.
0.1 0.
0.1 0.
0.09 0.
0.09 0.09
0.08 0.09
0.08 0.08
0.07 0.08
0.07 0.08
0.07 0.07
0.06 0.07
0.06 0.07
0.06 0.07
0.06 0.06
0.06 0.06
0.06 0.06
0.06 0.06
0.05 0.06
0.05 0.06
0.05 0.06
0.05 0.05
0.05 0.05
Bold values indicate exceedances
-------
TABLE 5-30: RATIO OF EGG CONCENTRATIONS FOR FEMALE MALLARD BASED ON
FTSHRAND RESULTS FOR THE TRI+ CONGENERS
REVISED
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
2.1
1.9
1.8
1.8
1.7
1.7
1.6
1.6
1.6
1.5
1.4
1.4
1.4
1.4
1.4
1.4
1.3
1.3
13
1.3
1.3
1.3
1.2
1.2
1.2
1.2-
LOAEL
152
95%UCL
2.2
2.0
1.9
1.9
1.9
1.8
1.7
1.7
1.7
1.6
1.5
1.5
1.5
1.5
1.5
1.5
1.4
1.4
1.4
1.4
1.4
1.3
1.3
13
1.3
1.4
NOAEL
152
Average
14
13
12
12
12
11
11
11
11
10
9.7
9.6
9.5
9.4
93
9.2
9.0
8.9
8.8
8.7
8.5
8.4
83
83
83
8.4
NOAEL
152
95%UCL
15
14
13
13
12
12
11
11
11
11
10
10
10
10
9.9
9.8
9.7
9.5
9.4
93
9.1
9.0
8.9
9.0
9.0
9.1
LOAEL
113
Average
1.6
1.6
1.5
1.4
1.4
13
13
13
1.2
1.2
1.2
1.2
1.1
1.1
1.1
1.1
1.1
1.1
1.1
1.0
1.0
1.0
1.0
1.0
1.0
1.0
LOAEL
113
95%UCL
1.7
1.6
1.6
1.5
1.5
1.4
1.4
13
13
13
1.2
1.2
1.2
1.2
1.2
1.2
1.2
1.1
1.1
1.1
1.1
1.1
1.1
1.1
1.1
1.1
NOAEL
113
Average
11
10
9.9
9.6
93
8.9
8.6
8.4
83
8.1
7.8
7.8
7.7
7.5
7.4
7.4
7.2
7.2
7.0
7.0
6.9
6.8
6.7
6.6
6.5
6.6
NOAEL
113
95% UCL
12
11
10
10
9.9
9.5
9.2
8.9
8.9
8.6
83
83
8.2
8.1
8.0
7.9
7.8
7.7
7.6
7.5
7.4
73
7.1
7.0
7.1
7.2
LOAEL
90
Average
13
13
1.2
1.1
1.1
1.1
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.8
0.7
0.7
LOAEL
90
95% UCL
1.4
13
13
1.2
1.2
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
0.8
0.8
0.8
NOAEL
90
Average
8.8
8.4
8.0
7.7
7.4
7.1
6.9
6.6
6.6
6.4
6.2
6.0
5.9
5.8
5.7
5.6
5.6
5.5
5.4
5.4
53
5.2
5.1
5.0
5.0
4.9
NOAEL
90
95% UCL
93
8.8
8.5
8.1
7.8
7.6
73
7.0
7.0
6.8
6.5
6.4
63
6.2
6.1
6.0
6.0
5.9
5.8
5.8
5.7
5.6
5.5
5.4
53
5.2
LOAEL
50
Average
1.0
0.9
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
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.5
LOAEL
50
95% UCL
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
0.7
0.7
0.7
0.7
0.6
0.6
0.6
0.6
0.6
0.6
0.6
0.6
NOAEL
50
Average
6.5
6.2
5.9
5.7
5.5
5.3
5.1
5.0
4.9
4.7
4.6
4.4
4.4
43
43
4.2
4.2
4.1
4.0
4.0
3.9
3.9
3.8
3.8
3.7
3.7
NOAEL
50
95% UCL
6.8
6.6
6.3
6.0
5.8
5.6
5.5
53
5.1
5.0
4.9
4.7
4.7
4.7
4.6
4.5
4.5
4.4
43
43
4.2
4.2
4.1
4.0
4.0
3.9
Bold values indicate exceedances
-------
TABLE 5-31: RATIO OF MODELED DIETARY DOSE TO BENCHMARKS
FOR FEMALE MALLARD FOR PERIOD 1993 - 2018 ON A TEQ BASIS
REVISED
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
12
9.3
13
10
7.7
6.7
7.3
8.2
6.7
5.8
4.7
4.6
4.9
4.5
4.1
3.3
4.2
3.9
4.0
4.6
3.9
3.7
3.0
2.9
2.9
LOAEL
152
95%UCL
15
13
10.2
14
11
8.4
7.3
8.0
9.0
7.3
6.3
5.1
5.0
5.3
4.9
4.4
3.6
4.6
4.2
4.4
5.0
4.2
4.0
3.2
3.1
3.2
NOAEL
152
Average
137
121
93
128
103
77
67
73
82
67
58
47
46
49
45
41
33
42
39
40
46
39
37
30
29
29
NOAEL
152
95% UCL
150
133
102
140
113
84
73
80
90
73
63
51
50
53
49
44
36
46
42
44
50
42
40
32
31
32
LOAEL
113
Average
11
9.6
7.5
8.1
7.4
6.0
5.1
5.0
5.3
4.8
4.5
3.6
3.4
3.5
3.4
3.1
2.8
3.0
3.0
3.0
3.1
2.9
2.7
2.3
2.2
2.2
LOAEL
113
95% UCL
12
10
8.2
8.8
8.1
6.6
5.5
5.4
5.8
5.2
4.9
4.0
3.7
3.8
3.7
3.3
3.0
3.2
3.3
3.3
3.4
3.2
3.0
2.5
2.3
2.4
NOAEL
113
Average
107
96
75
81
74
60
51
50
53
48
45
36
34
35
34
31
28
30
30
30
31
29
27
23
22
22
NOAEL
113
95% UCL
117
104
82
88
81
66
55
54
58
52
49
40
37
38
37
33
30
32
33
33
34
32
30
25
23
24
LOAEL
90
Average
8.8
7.7
6.8
6.1
5.5
5.0
4.5
4.1
3.8
3.6
3.4
3.1
2.9
2.7
2.5
2.4
2.3
2.2
2.1
2.1
2.1
2.0
2.0
1.9
1.8
1.7
LOAEL
90
95% UCL
9.4
8.4
7.0
6.8
6.1
5.4
4.7
4.3
4.2
4.0
3.8
3.3
3.1
2.9
2.8
2.6
2.5
2.4
2.4
2.4
2.4
2.3
2.3
2.1
1.9
1.9
NOAEL
90
Average
88
77
68
61
55
50
45
41
38
36
34
31
29
27
25
24
23
22
21
21
21
20
20
19
18
17
NOAEL
90
95%UCL
94
84
70
68
61
54
47
43
42
40
38
33
31
29
28
26
25
24
24
24
24
23
23
21
19
19
LOAEL
50
Average
14
12
9.1
13
10.2
7.4
6.3
6.9
7.9
6.3
5.4
4.1
4.1
4.4
4.0
3.5
2.7
3.7
3.3
3.5
4.2
3.4
3.2
2.4
2.3
2.3
LOAEL
50
95% UCL
9.2
8.1
7.1
6.3
5.7
5.1
4.6
4.1
3.8
3.6
3.4
3.1
2.9
2.7
2.5
2.4
2.2
2.1
2.1
2.0
2.0
2.0
1.9
1.9
1.8
1.7
NOAEL
50
Average
137
121
91
128
102
74
63
69
79
63
54
41
41
44
40
35
27
37
33
35
42
34
32
24
23
23
NOAEL
50
95% UCL
92
81
71
63
57
51
46
41
38
36
34
31
29
27
25
24
22
21
21
20
20
20
19
19
18
17
Bold values indicate exceedances
-------
TABLE 5-32: RATIO OF MODELED EGG CONCENTRATION TO BENCHMARKS FOR
FEMALE MALLARD FOR PERIOD 1993 - 2018 ON A TEQ BASIS
REVISED
Year
1993
1994
199S
1996
1997
1998
1999
2000
2001
2002
2003
2004
2003
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
297
274
263
263
250
236
226
229
226
218
207
205
203
201
198
195
192
190
188
185
182
179
177
178
178
179
LOAEL
152
95% UCL
313
289
277
278
264
251
240
243
240
232
221
220
217
215
212
210
206
203
201
198
195
192
190
192
193
194
NOAEL
152
Average
1187
1094
1051
1053
1000
945
904
917
905
872
828
821
812
805
792
782
769
759
752
740
728
716
708
712
712
714
NOAEL
152
95% UCL
1250
1154
1110
1111
1058
1002
961
973
959
926
883
880
869
862
848
839
826
813
805
793
779
768
760
769
770
774
LOAEL
113
Average
235
222
212
205
199
191
184
180
178
172
167
166
164
161
159
157
154
153
150
148
146
144
142
140
140
141
LOAEL
113
95% UCL
247
235
224
217
211
202
196
191
189
183
177
177
176
172
170
168
166
164
161
159
157
155
152
150
151
153
NOAEL
113
Average
938
889
849
819
798
762
737
718
713
690
666
663
656
643
635
628
617
611
602
594
585
578
568
559
559
565
NOAEL
113
95% UCL
990
938
896
866
844
807
782
763
757
734
710
710
703
689
680
673
662
656
645
637
628
620
610
602
603
612
LOAEL
90
Average
188
179
171
164
158
153
147
142
140
136
131
127
126
124
123
120
119
118
116
115
113
111
109
108
106
104
LOAEL
90
95% UCL
197
189
181
173
167
161
156
150
149
145
140
136
135
133
131
129
128
126
124
123
121
119
117
115
114
112
NOAEL
90
Average
750
716
685
657
634
610
587
567
561
545
525
509
506
496
490
481
476
471
463
458
451
445
438
430
425
417
NOAEL
90
95%UCL
790
755
724
694
669
645
622
601
595
578
559
542
541
531
524
515
510
505
497
491
483
478
470
462
456
448
LOAEL
50
Average
138
133
127
122
116
113
110
106
104
100
98
94
95
93
91
90
89
88
86
85
84
83
82
80
79
78
LOAEL
50
95% UCL
146
140
134
129
123
120
116
113
110
106
104
100
101
99
98
97
95
94
93
91
90
89
88
86
85
84
NOAEL
50
Average
553
532
508
489
465
452
439
426
414
400
392
377
379
371
365
361
356
352
345
341
336
331
327
322
317
312
NOAEL
50
95% UCL
583
561
535
516
492
478
465
452
440
426
417
402
405
397
391
386
381
378
370
366
361
356
351
346
341
336
Bold values indicate exceedances
-------
TABLE 5-33: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR FEMALE KINGFISHER
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
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
9.2
12
9.6
7.6
7.2
6.6
7.4
6.7
6.1
5.0
5.0
5.6
5.0
4.7
4.1
4.7
4.8
4.7
4.9
4.5
4.1
3.7
3.7
3.8
LOAEL
152
95% UCL
14
11
9.6
12
10
7.9
7.5
6.9
7.7
7.0
6.3
5.3
53
5.8
5.2
4.9
4.3
4.9
5.0
4.9
5.1
4.7
4.2
3.9
3.9
4.0
NOAEL
152
Average
95
72
65
81
67
53
51
46
52
47
42
35
35
39
35
33
29
33
33
33
34
31
28
26
26
27
NOAEL
152
95% UCL
99
75
67
84
70
56
53
48
54
49
44
37
37
41
36
34
30
34
35
34
36
33
30
27
27
28
LOAEL
113
Average
12
11
8.9
9.4
8.6
7.5
6.7
6.5
6.6
6.2
5.8
5.2
5.0
5.0
4.9
4.6
4.2
43
4.5
4.4
4.4
4.3
4.1
3.8
3.6
3.6
LOAEL
113
95% UCL
12
11
9.2
9.7
8.9
7.8
6.9
6.8
6.8
6.5
6.1
5.4
5.1
5.2
5.0
4.8
4.4
4.4
4.6
4.6
4.5
4.5
4.2
3.9
3.8
3.7
NOAEL
113
Average
83
75
62
66
60
53
47
46
46
44
41
36
35
35
34
32
29
30
31
31
31
30
29
26
25
25
NOAEL
113
95% UCL
86
77
64
68
63
55
49
47
48
45
42
38
36
36
35
34
31
31
32
32
32
31
30
27
26
26
LOAEL
90
Average
9.3
8.4
7.5
7.0
6.5
6.0
5.5
5.0
4.9
4.7
4.5
4.1
3.9
3.8
3.7
3.5
33
33
33
33
3.2
3.2
3.1
2.9
2.8
2.7
LOAEL
90
95% UCL
9.6
8.7
7.8
7.2
6.8
63
5.7
5.2
5.0
4.9
4.6
43
4.0
3.9
3.8
3.7
3.4
3.4
3.4
3.4
3.4
3.3
3.2
3.1
2.9
2.9
NOAEL
90
Average
65
59
53
49
46
42
38
35
34
33
31
29
27
26
26
25
23
23
23
23
23
22
22
21
20
19
NOAEL
90
95% UCL
67
61
55
51
47
44
40
36
35
34
33
30
28
27
27
26
24
24
24
24
24
23
22
21
21
20
LOAEL
50
Average
8.7
7.8
6.9
63
5.7
5.3
4.8
4.4
4.2
4.0
3.8
3.5
33
3.2
3.1
2.9
2.8
2.7
2.7
2.6
2.6
2.6
2.5
2.4
2.3
23
LOAEL
50
95% UCL
9.0
8.0
1.2
6.5
5.9
5.5
5.0
4.6
43
4.1
3.9
3.7
3.5
33
3.2
3.1
2.9
2.8
2.8
2.7
2.7
2.7
2.6
2.5
2.4
23
NOAEL
50
Average
61
54
48
44
40
37
34
31
29
28
27
25
23
22
21
21
19
19
19
18
18
18
18
17
16
16
NOAEL
50
95% UCL
63
56
50
45
41
38
35
32
30
29
27
26
24
23
22
21
20
20
19
19
19
19
18
18
17
16
Bold values indicate exceedances
-------
TABLE 5-34: RATIO OF MODELED DIETARY DOSE (BASED ON FISHRAND) FOR FEMALE BLUE HERON
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
Year
1993
1994
199S
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
5.9
43
3.8
5.0
4.1
3.1
2.9
2.6
3.0
2.7
2.4
1.9
1.9
2.2
1.9
1.7
1.4
1.8
1.8
1.8
1.9
1.7
1.5
13
1.3
1.3
LOAEL
152
95%UCL
6.2
43
4.0
5.2
4.2
3.2
3.0
2.7
3.1
2.8
2.5
2.0
2.0
2.2
1.9
1.8
1.5
1.8
1.9
1.8
2.0
1.8
1.5
13
13
1.4
NOAEL
152
Average
42
30
27
35
28
22
20
18
21
19
17
13
13
15
13
12
10
12
13
12
13
12
10
9.0
9.0
9.4
NOAEL
152
95% UCL
43
31
28
36
30
22
21
19
22
20
17
14
14
16
14
13
11
13
13
13
14
12
11
9.4
9.4
9.8
LOAEL
113
Average
5.3
4.7
3.8
4.1
3.7
3.2
2.8
2.7
2.8
2.6
2.4
2.1
2.0
2.0
1.9
1.8
1.6
1.7
1.8
1.7
1.7
1.7
1.6
13
1.4
1.4
LOAEL
113
95% UCL
5.4
4.8
3.9
4.2
3.8
33
2.9
2.8
2.8
2.7
23
2.1
2.0
2.1
2.0
1.9
1.7
1.7
1.8
1.8
1.8
1.8
1.7
13
1.4
1.4
NOAEL
113
Average
37
33
27
29
26
22
20
19
19
18
17
15
14
14
14
13
11
12
12
12
12
12
11
10
9.6
93
NOAEL
113
95% UCL
38
34
28
30
27
23
20
20
20
19
17
15
14
15
14
13
12
12
13
13
13
12
12
10
9.9
9.8
LOAEL
90
Average
4.1
3.7
33
3.0
2.8
2.6
23
2.1
2.0
2.0
1.9
1.7
1.6
13
1.5
1.4
13
13
13
13
1.3
1.2
1.2
1.1
1.1
1.1
LOAEL
90
95% UCL
43
3.8
3.4
3.1
2.9
2.7
2.4
2.1
2.1
2.0
1.9
1.7
1.6
1.6
13
1.4
13
13
1.3
13
1.3
13
1.2
1.2
1.1
1.1
NOAEL
90
Average
29
26
23
21
20
18
16
15
14
14
13
12
11
11
10
9.8
9.1
9.0
9.1
8.9
8.9
8.7
83
8.0
7.6
7.4
NOAEL
90
95% UCL
30
27
24
22
20
19
17
15
15
14
13
12
11
11
11
10
93
9.2
93
9.2
9.2
9.0
8.7
8.2
7.8
7.6
LOAEL
50
Average
4.0
3.5
3.1
2.8
2.5
2.3
2.1
1.9
1.8
1.7
1.6
13
1.4
13
13
1.2
1.1
1.1
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
LOAEL
50
95% UCL
4.1
3.6
3.2
2.9
2.6
2.4
2.2
2.0
1.8
1.7
1.7
13
1.4
13
13
1.2
1.1
1.1
1.1
1.1
1.1
1.1
1.0
1.0
1.0
0.9
NOAEL
50
Average
28
25
22
20
18
16
15
13
12
12
11
10
9.7
9.2
8.8
8.4
7.8
7.7
7.5
7.4
7.4
7.2
7.1
6.8
6.5
6.2
NOAEL
50
95% UC1
29
25
22
20
18
17
15
14
13
12
12
11
10
9.4
9.0
8.6
8.0
7.9
7.8
7.6
73
7.4
73
7.0
6.7
6.4
BOIO values indicate exceedances
-------
TABLE 5-35: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR FEMALE BALD EAGLE
BASED ON THE SUM OF TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
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
20
14
12
15
13
8.9
8.1
7.9
8.7
8.1
7.1
5.4
5.5
63
5.6
4.9
4.4
4.8
5.5
5.0
5.5
5.2
4.4
4.0
3.9
4.0
LOAEL
152
95%UCL
20
14
12
15
13
9.0
7.8
7.7
8.5
7.8
6.9
5.2
53
6.1
5.4
4.8
43
4.6
5.4
4.8
53
5.0
43
3.9
3.7
3.8
NOAEL
152
Average
139
98
82
106
91
62
56
55
61
56
49
38
39
44
39
35
31
33
39
35
39
36
31
28
27
28
NOAEL
152
95% UCL
141
99
83
108
93
63
55
54
59
55
48
37
37
43
38
34
30
32
38
34
37
35
30
27
26
27
LOAEL
113
Average
15
13
11
11
10
9.0
7.8
7.2
7.5
73
6.6
5.8
5.2
53
5.2
4.9
4.4
43
4.7
4.5
4.6
4.4
4.1
3.9
3.6
3.4
LOAEL
113
95% UCL
16
13
11
11
11
9.2
8.0
7.4
7.7
7.4
6.7
5.9
53
5.4
53
5.0
4.4
4.4
4.8
4.5
4.7
4.5
4.2
4.0
3.6
3.5
NOAEL
113
Average
108
93
77
79
73
63
55
51
53
51
46
40
36
37
37
35
30
30
33
31
32
31
29
27
25
24
NOAEL
113
95% UCL
110
94
78
80
75
64
56
52
54
52
47
41
37
38
37
35
31
31
33
32
33
31
30
28
25
24
LOAEL
90
Average
12
11
9.4
8.4
7.9
7.2
6.4
5.7
5.5
53
5.0
4.6
4.2
3.9
3.8
3.7
3.5
33
33
33
33
3.2
3.1
3.0
2.8
2.7
LOAEL
90
95% UCL
12
11
10
8.6
8.0
73
6.5
5.8
5.5
5.4
5.1
4.7
43
4.0
3.9
3.8
3.5
3.4
3.4
3.4
33
33
3.2
3.0
2.9
2.7
NOAEL
90
Average
83
75
66
59
55
50
45
40
38
37
35
33
29
28
27
26
24
23
23
23
23
23
22
21
20
19
NOAEL
90
95% UCL
84
76
67
60
56
51
46
41
39
38
36
33
30
28
27
26
25
24
24
23
23
23
22
21
20
19
LOAEL
50
Average
12
10
9.0
8.0
73
6.6
6.0
5.4
5.0
4.7
4.4
4.2
3.8
3.6
3.4
3.2
3.0
2.9
2.8
2.8
2.8
2.7
2.7
2.6
2.5
23
LOAEL
50
95% UCL
12
10
9.1
8.2
7.4
6.7
6.1
5.5
5.0
4.8
4.5
4.2
3.9
3.6
3.4
3.3
3.1
2.9
2.9
2.8
2.8
2.8
2.7
2.6
2.5
2.4
NOAEL
50
Average
81
71
63
56
51
46
42
38
35
33
31
29
27
25
24
23
21
20
20
19
19
19
19
18
17
16
NOAEL
50
95% UCL
82
72
64
57
52
47
42
38
35
33
32
30
27
25
24
23
21
21
20
20
20
19
19
18
17
17
Bold values indicate exceedances
-------
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
REVISED
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
32
25
22
28
23
18
17
16
18
16
14
12
12
13
12
11
10
11
11
11
12
11
9.7
8.8
8.8
9.1
LOAEL
152
95% UCL
34
26
23
29
24
19
18
17
19
17
15
13
13
14
12
12
10
12
12
12
12
11
10
93
93
9.6
NOAEL
152
Average
218
164
148
185
154
122
116
106
119
108
97
81
81
89
80
75
66
75
76
75
79
72
65
59
59
61
NOAEL
152
95% UCL
227
171
154
193
161
127
121
111
124
112
101
84
84
93
83
78
69
79
80
78
82
75
68
62
62
64
LOAEL
113
Average
28
26
21
23
21
18
16
16
16
15
14
12
12
12
12
11
10
10
11
11
10
10
9.8
9.0
8.6
8.6
LOAEL
113
95% UCL
29
27
22
23
21
19
17
16
16
15
14
13
12
12
12
12
10
11
11
11
11
11
10
9.4
9.0
9.0
NOAEL
113
Average
190
172
142
151
138
121
107
105
105
100
94
83
79
80
78
74
67
68
72
70
70
69
65
60
58
58
NOAEL
113
95% UCL
197
178
148
156
143
125
111
109
109
104
97
86
82
83
81
77
70
71
74
73
73
71
68
63
60
60
LOAEL
90
Average
22
20
18
17
16
14
13
12
12
11
11
9.9
93
9.0
8.8
8.4
7.9
7.9
7.9
7.8
7.7
7.6
7.4
7.0
6.7
6.6
LOAEL
90
95% UCL
23
21
19
17
16
15
14
12
12
12
11
10
9.7
9.4
9.1
8.8
8.2
8.2
8.2
8.1
8.0
7.9
7.7
73
7.0
6.8
NOAEL
90
Average
150
136
121
112
105
97
88
80
78
76
72
66
62
61
59
57
53
53
53
52
52
51
50
47
45
44
NOAEL
90
95% UCL
155
140
125
116
109
100
91
83
81
78
74
69
65
63
61
59
55
55
55
54
54
53
52
49
47
46
LOAEL
50
Average
21
19
17
15
14
13
12
11
10
9.6
9.1
8.5
8.0
7.6
7.3
7.0
6.6
6.5
6.4
63
6.2
6.1
6.0
5.8
5.6
5.4
LOAEL
50
95% UCL
22
19
17
16
14
13
12
11
10
9.9
9.4
8.8
83
7.9
7.6
73
6.9
6.8
6.7
6.5
6.5
6.4
6.2
6.0
5.8
5.6
NOAEL
50
Average
140
125
111
101
92
85
78
71
67
64
61
57
53
51
49
47
44
44
43
42
42
41
40
39
37
36
NOAEL
50
95% UCL
145
129
115
104
95
88
80
74
69
66
63
59
55
53
51
49
46
45
45
44
43
43
42
40
39
38
Bold values indicate exceedances
-------
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
REVISED
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
?003
2004
200S
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
36
26
23
30
25
19
18
16
18
16
14
11
11
13
11
10
8.7
11
11
11
11
10
8.9
7.8
7.8
8.1
LOAEL
152
95%UCL
37
27
24
32
26
19
18
16
19
17
15
12
12
14
12
11
9.0
11
11
11
12
11
9.2
8.1
8.1
8.4
NOAEL
152
Average
241
175
155
203
165
125
119
106
123
110
97
76
76
88
76
70
58
71
73
71
77
68
60
52
52
54
NOAEL
152
95% UCL
251
182
161
211
172
130
123
109
127
114
101
79
79
91
78
72
61
74
75
74
80
71
62
54
54
56
LOAEL
113
Average
32
29
23
25
23
19
17
17
17
16
15
13
12
12
12
11
10
10
11
11
11
10
9.7
8.8
83
8.2
LOAEL
113
95% UCL
33
29
24
26
23
20
17
17
17
16
15
13
12
13
12
11
10
10
11
11
11
11
10
9.0
8.5
8.5
NOAEL
113
Average
214
192
155
167
151
130
114
111
112
106
98
84
80
82
79
75
66
68
72
71
70
69
65
59
56
55
NOAEL
113
95% UCL
221
197
160
172
157
134
117
114
115
109
101
87
82
84
81
77
68
69
74
73
72
71
67
61
57
57
LOAEL
90
Average
25
23
20
18
17
16
14
13
12
12
11
10
9.4
9.2
8.9
8.5
7.8
7.8
7.8
7.7
7.7
7.6
7.3
6.9
6.5
6.4
LOAEL
90
95% UCL
26
23
21
19
18
16
14
13
13
12
12
11
9.7
9.4
9.1
8.7
8.0
8.0
8.0
7.9
7.9
7.8
7.5
7.1
6.7
6.5
NOAEL
90
Average
168
151
133
123
114
105
93
85
82
79
75
68
63
61
60
57
52
52
52
52
52
51
49
46
44
43
NOAEL
90
95% UCL
173
156
137
127
118
108
96
87
84
81
77
70
65
63
61
58
54
53
54
53
53
52
50
48
45
44
LOAEL
50
Average
24
21
19
17
15
14
13
12
11
10
9.7
9.0
8.4
7.9
7.6
7.3
6.7
6.6
6.5
6.4
6.4
6.2
6.1
5.9
5.6
5.4
LOAEL
50
95% UCL
25
22
20
18
16
15
13
12
11
11
10
93
8.6
8.2
7.8
7.5
6.9
6.8
6.7
6.6
6.5
6.4
63
6.0
5.8
5.5
NOAEL
50
Average
162
143
127
114
103
94
85
77
72
69
65
60
56
53
51
49
45
45
44
43
43
42
41
39
38
36
NOAEL
50
95% UCL
168
148
131
117
106
97
88
80
74
71
67
62
58
55
52
50
46
46
45
44
44
43
42
40
39
37
Bold values indicate exceedances
-------
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
REVISED
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
55
39
33
43
36
29
25
24
27
24
21
16
16
19
16
15
13
14
17
14
16
14
13
12
11
10
NOAEL
152
95% UCL
56
40
33
43
37
30
25
24
27
25
22
17
17
19
17
15
14
14
17
14
17
15
13
12
11
11
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
43
37
31
31
29
25
22
20
21
20
18
16
14
15
15
14
12
12
13
12
13
12
12
11
9.9
9.6
NOAEL
113
95% UCL
44
38
31
32
30
26
22
21
21
21
19
16
15
15
15
14
12
12
13
13
13
12
12
11
10
9.8
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
32
28
25
22
20
18
17
15
14
13
12
12
11
10
9.5
9.0
8.4
8.1
7.9
7.8
7.7
7.6
7.4
7.2
6.9
6.5
NOAEL
90
95% UCL
33
29
26
23
21
19
17
15
14
13
13
, 12
11
10
9.6
9.2
8.6
8.2
8.1
7.9
7.8
7.7
7.6
73
7.0
6.7
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
32
28
25
22
20
18
17
15
14
13
12
12
11
10
9.5
9.0
8.4
8.1
7.9
7.8
7.7
7.6
7.4
7.2
6.9
6.5
NOAEL
50
95% UCL
33
29
26
23
21
19
17
15
14
13
13
12
11
10
9.6
9.2
8.6
8.2
8.1
7.9
7.8
7.7
7.6
73
7.0
6.7
Bold values indicate exceedances
-------
TABLE 5-39: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE BELTED KINGFISHER USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
201S
2016
2017
2018
LOAEL
152
Average
13
9.8
8.8
11
9.2
73
6.9
63
7.1
6.4
5.8
4.8
4.8
53
4.7
4.4
3.9
4.5
4.5
4.4
4.7
4.3
3.9
3.5
3.5
3.6
LOAEL
152
95% UCL
13
10
9.1
11
9.5
7.5
7.1
6.5
7.3
6.7
6.0
5.0
5.0
5.5
4.9
4.6
4.1
4.6
4.7
4.6
4.9
4.4
4.0
3.7
3.6
3.7
NOAEL
152
Average
129
98
88
110
92
73
69
63
71
64
58
48
48
53
47
44
39
45
45
44
47
43
39
35
35
36
NOAEL
152
95% UCL
134
101
91
114
95
75
71
65
73
67
60
50
50
55
49
46
41
46
47
46
49
44
40
37
36
37
LOAEL
113
Average
11
10
8.4
8.9
8.2
7.2
6.4
6.2
6.3
5.9
5.6
4.9
4.7
4.8
4.6
4.4
4.0
4.1
4.2
4.2
4.1
4.1
3.9
3.6
3.4
3.4
LOAEL
113
95% UCL
12
10
8.7
9.2
8.5
7.4
6.6
6.4
6.5
6.1
5.7
5.1
4.9
4.9
4.8
4.6
4.1
4.2
4.4
4.3
43
4.2
4.0
3.7
3.5
3.5
NOAEL
113
Average
113
102
84
89
82
72
64
62
63
59
56
49
47
48
46
44
40
41
42
42
41
41
39
36
34
34
NOAEL
113
95% UCL
117
105
87
92
85
74
66
64
65
61
57
51
49
49
48
46
41
42
44
43
43
42
40
37
35
35
LOAEL
90
Average
8.9
8.0
7.2
6.7
6.2
5.8
5.2
4.8
4.6
4.5
4.3
3.9
3.7
3.6
3.5
3.4
3.1
3.1
3.1
3.1
3.1
3.0
2.9
2.8
2.7
2.6
LOAEL
90
95% UCL
28
26
23
23
22
21
19
19
19
18
18
16
16
16
16
15
14
14
14
14
14
14
13
13
12
12
NOAEL
90
Average
89
80
72
67
62
58
52
48
46
45
43
39
37
36
35
34
31
31
31
31
31
30
29
28
27
26
NOAEL
90
95% UCL
279
256
232
231
224
208
193
188
190
183
175
164
158
159
156
149
142
143
144
141
140
137
133
127
123
121
LOAEL
50
Average
8.3
7.4
6.6
6.0
5.4
5.0
4.6
4.2
4.0
3.8
3.6
3.4
3.2
3.0
2.9
2.8
2.6
2.6
2.5
2.5
2.5
2.4
2.4
23
2.2
2.1
LOAEL
50
95% UCL
23
21
20
19
18
17
16
15
15
15
14
13
13
13
12
12
12
11
11
11
11
11
11
10
9.9
9.7
NOAEL
50
Average
83
74
66
60
54
50
46
42
40
38
36
34
32
30
29
28
26
26
25
25
25
24
24
23
22
21
NOAEL
50
95% UCL
229
213
197
187
180
170
160
153
150
146
141
134
129
126
123
120
115
114
113
111
110
108
106
102
99
97
Bold values indicate exceedances
-------
TABLE 5-40: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE GREAT BLUE HERON USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
Year
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
200S
2006
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
LOAEL
152
Average
5.8
4.2
3.7
4.9
4.0
3.0
2.9
2.6
3.0
2.7
2.4
1.9
1.9
2.1
1.9
1.7
1.4
1.8
1.8
1.7
1.9
1.7
1.5
13
13
1.3
LOAEL
152
95% UCL
6.0
4.4
3.9
5.0
4.1
3.1
3.0
2.7
3.1
2.8
2.4
1.9
1.9
2.2
1.9
1.8
1.5
1.8
1.8
1.8
1.9
1.7
1.5
13
13
IA
NOAEL
152
Average
58
42
37
49
40
30
29
26
30
27
24
19
19
21
19
17
14
18
18
17
19
17
IS
13
13
13
NOAEL
152
95% UCL
60
44
39
50
41
31
30
27
31
28
24
19
19
22
19
18
15
18
18
18
19
17
15
13
13
14
LOAEL
113
Average
5.1
4.6
3.7
4.0
3.6
3.1
2.7
2.7
2.7
2.5
2.4
2.0
1.9
2.0
1.9
1.8
1.6
1.6
1.7
1.7
1.7
1.7
1.6
1.4
1.4
13
LOAEL
113
95%UCL
5.2
4.7
3.8
4.1
3.7
3.2
2.8
2.7
2.8
2.6
2.4
2.1
2.0
2.0
2.0
1.9
1.7
1.7
1.8
1.8
1.8
1.7
1.6
1.5
1.4
1.4
NOAEL
113
Average
51
46
37
40
36
31
27
27
27
25
24
20
19
20
19
18
16
16
17
17
17
17
16
14
14
13
NOAEL
113
95% UCL
52
47
38
41
37
32
28
27
28
26
24
21
20
20
20
19
17
17
18
18
18
17
16
15
14
14
LOAEL
90
Average
4.0
3.6
3.2
2.9
2.7
2.5
2.2
2.0
2.0
1.9
1.8
1.7
1.5
1.5
1.4
1.4
13
13
13
13
13
1.2
13
1.1
1.1
1.0
LOAEL
90
95% UCL
4.1
3.7
33
3.0
2.8
2.6
23
2.1
2.0
2.0
1.9
1.7
1.6
1.5
1.5
1.4
13
13
13
13
13
13
1.2
13
1.1
1.1
NOAEL
90
Average
40
36
32
29
27
25
22
20
20
19
18
17
15
15
14
14
13
13
13
13
13
12
12
11
11
10
NOAEL
90
95% UCL
41
37
33
30
28
26
23
21
20
20
19
17
16
15
15
14
13
13
13
13
13
13
12
12
11
11
LOAEL
50
Average
3.8
3.4
3.0
2.7
2.4
2.2
2.0
1.9
1.7
1.7
1.6
1.5
1.4
13
13
13
1.1
1.1
1.1
1.0
1.0
1.0
1.0
1.0
0.9
0.9
LOAEL
50
95% UCL
3.9
3.5
3.1
2.8
2.5
23
2.1
1.9
1.8
1.7
1.6
1.5
1.4
1.3
13
1.2
1.1
1.1
1.1
1.1
1.1
1.0
1.0
1.0
0.9
0.9
NOAEL
50
Average
38
34
30
27
24
22
20
19
17
17
16
15
14
13
12
12
11
11
11
10
10
10
9.9
9.5
9.1
8.8
NOAEL
50
95% UCL
39
35
31
28
25
23
21
19
18
17
16
15
14
13
13
12
11
11
11
11
11
10
10
9.7
93
9.0
Bold values indicate exceedances
-------
TABLE 5-41: RATIO OF MODELED DIETARY DOSE BASED ON FISHRAND FOR
FEMALE BALD EAGLE USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
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
19
13
11
14
12
9.9
8.4
8.1
9.0
8.3
7.3
5.6
5.5
6.3
5.5
5.1
4.5
4.8
5.6
4.8
5.5
4.9
4.5
4.0
3.6
3.5
LOAEL
152
95% UCL
19
13
11
15
13
10
8.5
8.2
9.1
8.4
7.4
5.7
5.6
6.4
5.6
5.2
4.6
4.9
5.7
4.9
5.6
4.9
4.5
4.1
3.7
3.6
NOAEL
152
Average
188
133
111
144
124
99
84
81
90
83
73
56
55
63
55
51
45
48
56
48
55
49
45
40
36
35
NOAEL
152
95% UCL
191
135
113
146
126
101
85
82
91
84
74
57
56
64
56
52
46
49
57
49
56
49
45
41
37
36
LOAEL
113
Average
15
13
10
11
9.9
8.6
7.4
6.9
7.1
6.9
6.2
5.5
4.9
5.1
5.0
4.7
4.1
4.1
4.5
4.2
4.4
4.2
3.9
3.7
3.4
33
LOAEL
113
95% UCL
15
13
11
11
10
8.7
7.6
7.0
7.3
7.0
6.3
5.6
5.0
5.2
5.0
4.8
4.2
4.2
4.5
4.3
4.5
4.2
4.0
3.8
3.4
33
NOAEL
113
Average
146
126
104
107
99
86
74
69
71
69
62
55
49
51
50
47
41
41
45
42
44
42
39
37
34
33
NOAEL
113
95% UCL
148
128
106
108
101
87
76
70
73
70
63
56
50
52
50
48
42
42
45
43
45
42
40
38
34
33
LOAEL
90
Average
11
10
8.9
8.0
7.5
6.8
6.1
5.4
5.2
5.0
4.8
4.4
4.0
3.7
3.6
3.5
3.3
3.1
3.2
3.1
3.1
3.1
3.0
2.8
2.7
2.5
LOAEL
90
95% UCL
11
10
9.1
8.1
7.6
6.9
6.2
5.5
5.3
5.1
4.8
4.5
4.1
3.8
3.7
3.6
3.3
3.2
3.2
3.2
3.2
3.1
3.0
2.9
2.7
2.6
NOAEL
90
Average
112
101
89
80
75
68
61
54
52
50
48
44
40
37
36
35
33
31
32
31
31
31
30
28
27
25
NOAEL
90
95% UCL
114
103
91
81
76
69
62
55
53
51
48
45
41
38
37
36
33
32
32
32
32
31
30
29
27
26
LOAEL
50
Average
11
9.7
8.5
7.6
6.9
63
5.7
5.1
4.7
4.5
4.2
4.0
3.6
3.4
3.2
3.1
2.9
2.7
2.7
2.6
2.6
2.6
2.5
2.4
2.3
2.2
LOAEL
50
95% UCL
11
9.8
8.7
7.7
7.0
6.4
5.7
5.2
4.8
4.5
4.3
4.0
3.7
3.4
3.3
3.1
2.9
2.8
2.7
2.7
2.7
2.6
2.6
2.5
2.4
2.3
NOAEL
50
Average
110
97
85
76
69
63
57
51
47
45
42
40
36
34
32
31
29
27
27
26
26
26
25
24
23
22
NOAEL
50
95% UCL
112
98
87
77
70
64
57
52
48
45
43
40
37
34
33
31
29
28
27
27
27
26
26
25
24
23
Bold values indicate exceedances
-------
TABLE 5-42: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE BELTED KINGFISHER USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
Year
1993
1994
199S
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
412
308
277
350
289
227
215
196
222
200
179
147
147
164
145
136
118
137
139
136
144
131
117
106
106
109
LOAEL
152
95% UCL
430
321
288
365
302
236
224
204
230
208
187
153
153
171
151
142
123
143
145
142
151
137
123
111
111
115
NOAEL
152
Average
825
616
553
700
579
453
430
391
443
400
358
294
294
328
290
271
236
275
278
272
289
262
235
212
212
219
NOAEL
152
95% UCL
860
641
575
729
604
472
448
407
461
416
373
307
306
341
303
283
246
286
290
284
301
274
245
222
222
229
LOAEL
113
Average
362
326
269
286
261
227
201
197
198
187
175
154
147
149
144
137
124
126
132
130
129
127
121
111
106
105
LOAEL
113
95% UCL
375
337
279
296
271
236
208
203
205
194
181
159
152
154
149
142
129
131
137
135
134
132
' 125
115
110
109
NOAEL
113
Average
724
652
538
572
522
455
402
393
396
375
350
308
294
298
288
275
248
252
265
260
259
254
241
221
211
210
NOAEL
113
95% UCL
751
674
558
592
542
471
416
406
409
388
362
319
304
308
298
285
257
261
274
270
268
264
250
230
219
219
LOAEL
90
Average
268
239
213
192
175
161
147
134
126
120
114
106
100
95
91
88
82
81
80
78
78
76
75
72
69
67
LOAEL
90
95% UCL
278
247
220
199
181
166
152
139
130
124
118
110
103
98
94
91
85
84
82
81
80
79
77
75
72
69
NOAEL
90
Average
537
477
425
384
350
322
294
269
252
241
229
213
200
190
182
176
164
162
159
157
155
153
149
144
139
134
NOAEL
90
95% UCL
556
494
440
397
362
333
304
278
260
249
236
220
207
197
189
182
170
168
165
162
160
158
155
149
144
139
LOAEL
50
Average
268
239
213
192
175
161
147
134
126
120
114
106
100
95
91
88
82
81
80
78
78
76
75
72
69
67
LOAEL
50
95% UCL
278
247
220
199
181
166
152
139
130
124
118
110
103
98
94
91
85
84
82
81
80
79
77
75
72
69
NOAEL
50
537
477
425
384
350
322
294
269
252
241
229
213
200
190
182
176
164
162
159
157
155
153
149
144
139
134
NOAEL
50
95% UC1
556
494
440
397
362
333
304
278
260
249
236
220
207
197
189
182
170
168
165
162
160
158
155
149
144
139
cola values indicate exceeaances
-------
TABLE 5-43: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE GREAT BLUE HERON USING TEQ FOR THE PERIOD 1993 - 2018
REVISED
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
19
14
12
16
13
9.9
9.4
8.4
9.7
8.7
7.7
6.0
6.1
7.0
6.0
5.5
4.6
5.7
5.8
5.6
6.1
5.4
4.7
4.1
4.1
4.3
LOAEL
152
95% UCL
20
14
13
17
14
10
9.8
8.7
10
9.0
8.0
6.3
6.3
7.2
6.2
5.7
4.8
5.9
6.0
5.8
6.3
5.6
4.9
4.3
4.3
4.5
NOAEL
152
Average
32
23
21
27
22
17
16
14
16
14
13
10
10
.12
10
9.2
7.7
9.4
9.6
9.4
10
9.0
7.9
6.9
6.9
7.2
NOAEL
152
95% UCL
33
24
21
28
23
17
16
14
17
15
13
10
10
12
10
9.5
8.0
9.8
10
9.7
11
9.4
8.2
7.1
7.1
7.4
LOAEL
113
Average
17
15
12
13
12
10
9.0
8.8
8.9
8.4
7.8
6.7
6.4
6.5
6.3
5.9
5.2
5.4
5.7
5.6
5.6
5.5
5.2
4.7
4.4
4.4
LOAEL
113
95% UCL
18
16
13
14
12
11
9.3
9.1
9.2
8.6
8.0
6.9
6.5
6.7
6.4
6.1
5.4
5.5
5.9
5.8
'5.7
5.6
53
4.8
4.5
4.5
NOAEL
113
Average
28
25
21
22
20
17
15
15
15
14
13
11
11
11
10
9.9
8.7
8.9
9.5
9.3
9.3
9.1
8.6
7.8
7.3
7.3
NOAEL
113
95% UCL
29
26
21
23
21
18
15
15
15
14
13
11
11
11
11
10
9.0
9.2
9.8
9.6
9.6
9.4
8.8
8.0
7.5
7.5
LOAEL
90
Average
13
12
11
10
9
8
7
6.7
6.5
6.3
5.9
5.4
5.0
4.9
4.7
4.5
4.2
4.1
4.2
4.1
4.1
4.0
3.9
3.7
3.5
3.4
LOAEL
90
95% UCL
14
12
11
10
9.3
8.5
7.6
6.9
6.7
6.5
6.1
5.6
5.2
5.0
4.9
4.6
4.3
4.2
4.3
4.2
4.2
4.1
4.0
3.8
3.6
3.5
NOAEL
90
Average
22
20
18
16
15
14
12
11
11
10
9.9
9.0
8.4
8.1
7.9
7.5
6.9
6.9
6.9
6.8
6.8
6.7
6.5
6.1
5.8
5.6
NOAEL
90
95% UCL
23
21
18
17
16
14
13
12
11
11
10
9.3
8.6
8.3
8.1
7.7
7.1
7.1
7.1
7.0
7.0
6.9
6.7
6.3
6.0
5.8
LOAEL
50
Average
13
11
10
9.0
8.2
7.5
6.8
6.1
5.7
5.5
5.2
4.8
4.5
4.2
4.0
3.9
3.6
3.5
3.5
3.4
3.4
3.3
3.2
3.1
3.0
2.9
LOAEL
50
95% UCL
13
12
10
9.3
8.4
7.7
7.0
6.3
5.9
5.6
5.3
4.9
4.6
4.3
4.1
4.0
3.7
3.6
3.6
3.5
3.5
3.4
3.3
3.2
3.1
2.9
NOAEL
50
Average
21
19
17
15
14
12
11
10
9.5
9.1
8.6
8.0
7.4
7.0
6.7
6.4
6.0
5.9
5.8
5.7
5.6
5.5
5.4
5.2
5.0
4.8
NOAEL
50
95% UCL
22
20
17
15
14
13
12
11
9.8
9.4
8.9
8.2
7.6
7.2
6.9
6.6
6.1
6.0
5.9
5.8
5.8
5.7
5.6
5.3
5.1
4.9
Bold values indicate exceedances
-------
TABLE 5-44: RATIO OF MODELED EGG CONCENTRATIONS BASED ON FISHRAND
FOR FEMALE BALD EAGLE USING TEQ FOR THE PERIOD 1993 • 2018
REVISED
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
3924
2770
2319
3010
2578
1767
1595
1555
1722
1596
1396
1066
1089
1240
1100
980
870
941
1093
987
1097
1031
876
792
777
797
NOAEL
152
95% UCL
3993
2812
2360
3050
2623
1776
1553
1516
1681
1549
1362
1036
1055
1207
1064
950
848
916
1064
951
1059
988
848
765
742
757
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
3046
2619
2176
2223
2072
1791
1553
1432
1491
1443
1298
1140
1026
1058
1034
976
862
858
930
885
914
867
821
774
704
680
NOAEL
113
95% UCL
3098
2664
2216
2258
2108
1822
1580
1457
1515
1467
1320
1159
1043
1075
1051
993
877
872
945
900
929
882
835
787
716
692
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
2338
2110
1862
1666
1560
1420
1266
1133
1081
1050
992
920
832
782
760
733
687
655
659
653
651
637
618
594
557
530
NOAEL
90
95% UCL
2376
2145
1894
1694
1585
1444
1287
1152
• 1099
1067
1009
935
846
795
773
745
699
666
671
664
663
648
629
604
566
539
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
2291
2016
1781
1588
1439
1308
1180
1067
982
929
881
825
760
705
669
639
598
572
562
551
546
537
526
508
485
464
NOAEL
50
95% UCL
2330
2050
1811
1614
1463
1330
1199
1085
998
944
895
839
773
716
680
649
608
581
572
561
555
546
535
517
493
471
Bold values indicate exceedances
-------
TABLE 5-45: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE BAT FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
Year
1993
1994
199S
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
2012
2013
2014
201S
2016
2017
2018
LOAEL
152
Average
3.6
33
3.2
3.2
3.0
2.9
2.7
2.8
2.7
2.6
2.5
23
23
2A
2.4
2.4
23
23
23
23
23
23,
2.1
2.2
23.
23,
LOAEL
152
95%UCL
33
33
3.4
34
33.
3.0
2.9
2.9
2.9
2.8
2.7
2.7
2.6
2.6
2.6
23
2.5
23
2.4
2.4
2.4
23
23
23
23
23
NOAEL
152
Average
17
16
15
15
14
13
13
13
13
12
12
12
12
11
11
11
11
11
11
10
10
10
10
10
10
10
NOAEL
152
95% UCL
18
16
16
16
15
14
14
14
14
13
13
12
12
12
12
12
12
12
11
11
11
11
11
11
11
11
LOAEL
113
Average
2.8
2.7
2.6
23
2A
23
23.
2.2
22
2.1
2.0
2.0
2.0
1.9
1.9
1.9
1.9
1.8
1.8
13
1.8
1.7
1.7
1.7
1.7
1.7
LOAEL
113
95% UCL
3.0
2JS
2.7
2.6
2.6
2.4
24
23
23
23.
2.1
2.1
2.1
2.1
2.1
2.0
2.0
2.0
2.0
1.9
1.9
1.9
1.8
1.8
13
1.9
NOAEL
113
Average
13
13
12
12
11
11
10
10
10
93
9A
9A
93
9.1
9.0
83
83
8.7
83
8.4
83
8.2
8.1
7.9
7.9
8.0
NOAEL
113
95% UCL
14
13
13
12
12
11
11
11
11
10
10
10
10
93
9.6
9.5
9.4
93
93.
9.0
8.9
83
8.6
83
83
8.7
LOAEL
90
Average
23
23.
2.1
2.0
1.9
13
1.8
1.7
1.7
1.6
1.6
13
13
13
1.5
13
IA
IA
1A
1A
1A
13
13
13
13
13
LOAEL
90
95% UCL
2.4
23
23.
2.1
2.0
2.0
1.9
13
13
1.7
1.7
1.6
1.6
1.6
1.6
1.6
13
13
13
13
13
IA
1.4
1.4
1A
1.4
NOAEL
90
Average
11
10
9.7
93
9.0
8.7
83
8.0
8.0
7.7
7.4
7J
7.2
7.0
6.9
63
6.7
6.7
6.6
63
6.4
63
63
6.1
6.0
5.9
NOAEL
90
95%UCL
11
11
10
93
93
9.2
83
83
8.4
8.2
7.9
7.7
7.7
73
7.4
73
7J
7.2
7.0
7.0
6.9
63
6.7
63
63
63
LOAEL
50
Average
17
16
15
15
14
14
13
13
13
12
12
11
11
11
11
11
11
11
10
10
10
10
9.9
9.7
9.6
9A
LOAEL
50
95% UCL
18
17
16
16
15
14
14
14
13
13
13
12
12
12
12
12
12
11
11
11
11
11
11
10
10
10
NOAEL
50
Average
73
73
7J
6.9
6.6
6.4
63
6.0
5.9
5.7
5.6
53
SA
53
S3,
5.1
5.0
5.0
4.9
43
43
4.7
4.6
4.6
43
4.4
NOAEL
50
95% UCL
83
8.0
7.6
73
7.0
63
6.6
6A
63.
6.0
5.9
5.7
5.7
5.6
53
53
5.4
SA
53
53.
5.1
5.0
5.0
4.9
43
43
Bold values indicate exceedances
TAMS/MCA
-------
TABLE 5-46: RATIO OF MODELED DIETARY DOSES TO TOXTCTTY BENCHMARKS
FOR FEMALE BAT ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
REVISED
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
58
54
51
52
49
46
44
45
44
43
41
40
40
39
39
38
38
37
37
36
36
35
35
35
35
35
LOAEL
152
9595. UCL
61
57
54
54
52
49
47
48
47
45
43
43
43
42
42
41
40
40
39
39
38
38
37
38
38
38
NOAEL
152
Average
581
536
514
515
490
463
442
449
443
427
405
402
397
394
388
383
377
372
368
362
356
351
346
348
348
350
NOAEL
152
95% UCL
612
565
543
544
518
491
470
476
470
453
432
431
426
422
415
411
404
398
394
388
382
376
372
376
377
379
LOAEL
113
Average
46
44
42
40
39
37
36
35
35
34
33
32
32
31
31
31
30
30
29
29
29
28
28
27
27
28
LOAEL
113
95% UCL
48
46
44
42
41
40
38
37
37
36
35
35
34
34
33
33
32
32
32
31
31
30
30
29
30
30
NOAEL
113
Average
459
435
416
401
391
373
361
352
349
338
326
325
321
315
311
307
302
299
294
291
287
283
278
274
274
277
NOAEL
113
95% UCL
484
459
439
424
413
395
383
374
370
359
347
347
344
337
333
330
324
321
316
312
307
303
299
295
295
299
LOAEL
90
Average
37
35
34
32
31
30
29
28
27
27
26
25
25
24
24
24
23
23
23
22
22
22
21
21
21
20
LOAEL
90
95% UCL
39
37
35
34
33
32
30
29
29
28
27
27
26
26
26
25
25
25
24
24
24
23
23
23
22
22
NOAEL
90
Average
367
351
336
322
310
299
288
278
275
267
257
249
248
243
240
236
233
231
227
224
221
218
214
211
208
204
NOAEL
90
95%UCL
387
370
354
340
328
316
305
294
291
283
274
265
265
260
257
252
250
247
243
240
237
234
230
226
223
219
LOAEL
50
Average
27
26
25
24
23
22
22
21
20
20
19
18
19
18
18
18
17
17
17
17
16
16
16
16
16
15
LOAEL
50
95% UCL
29
27
26
25
24
23
23
22
22
21
20
20
20
19
19
19
19
19
18
18
18
17
17
17
17
16
NOAEL
50
271
260
248
240
228
221
215
209
203
196
192
185
186
182
179
176
174
172
169
167
164
162
160
157
155
153
NOAEL
50
95% UCL
285
275
262
253
241
234
228
221
215
208
204
197
198
194
191
189
187
185
181
179
177
174
172
169
167
165
BOICJ values indicate exceedances
-------
TABLE 5-47: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE RACCOON FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
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
0.7
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
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
0.4
LOAEL
152
95% UCL
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
0.5
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
33
2.9
2.8
29
2.7
2.5
2.4
2.4
2.4
2.3
2.2
2.1
2.1
2.1
2.0
2.0
2.0
2.0
1.9
1.9
1.9
1.9
1.8
1.8
1.8
1.8
NOAEL
152
95% UCL
3.4
3.1
2.9
3.0
2.8
2.6
2.5
2.5
2.5
2.4
2.3
2.3
2.2
2.2
2.2
2.1
2.1
2.1
2.1
2.0
2.0
2.0
1.9
1.9
1.9
2.0
LOAEL
113
Average
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
0.3
0.3
0.3
0.3
0.3
0.3
LOAEL
113
95% UCL
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
0.4
0.4
0.4
0.4
0.4
0.3
0.3
0.3
0.3
0.3
0.3
NOAEL
113
Average
2.6
2.5
2.3
2.3
2.2
2.1
2.0
1.9
1.9
1.8
1.8
1.7
1.7
1.7
1.7
1.6
1.6
1.6
1.6
1.6
1.5
1.5
1.5
1.5
1.4
1.5
NOAEL
113
95% UCL
2.6
2.5
23
23
2.2
2.2
2.1
2.0
2.0
2.0
1.9
1.9
1.8
1.8
1.8
1.8
1.7
1.7
1.7
1.7
1.6
1.6
1.6
1.6
13
1.6
LOAEL
90
Average
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
0.3
0.3
0.3
0.2
0.2
0.2
0.2
0.2
LOAEL
90
95% UCL
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
0.3
0.3
0.3
0.3
0.3
0.3
0.2
NOAEL
90
Average
2.1
2.0
1.9
1.8
1.7
1.7
1.6
1.5
1.5
IS
1.4
13
1.3
13
13
13
1.2
1.2
1.2
12
1.2
1.2
1.1
1.1
1.1
1.1
NOAEL
90
95% UCL
2.2
2.1
2.0
1.9
1.8
1.7
1.7
1.6
1.6
15
1.5
1.4
1.4
1.4
1.4
13
13
13
13
13
13
1.2
1.2
1.2
1.2
1.2
LOAEL
50
Average
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
0.2
LOAEL
50
95% UCL
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
0.2
0.2
0.2
0.2
0.2
0.2
0.2
0.2
NOAEL
50
Average
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
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.9
0.9
0.8
0.8
0.8
NOAEL
50
95% UCL
1.7
1.6
1.5
1.4
1.4
1.3
1.3
1.2
1.2
1.1
1.1
1.1
1.1
1.0
1.0
1.0
1.0
1.0
1.0
1.0
0.9
0.9
0.9
0.9
0.9
0.9
Bold values indicate exceedances
-------
TABLE 5-48: RATIO OF MODELED DIETARY DOSES TO TOHCITY BENCHMARKS
FOR FEMALE RACCOON ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
REVISED
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
12
11
10
11
10
9.3
8.8
8.9
8.9
8.5
8.0
7.8
7.7
7.7
7.6
7.4
7.2
7.2
7.2
7.1
7.0
6.8
6.7
6.6
6.6
6.6
LOAEL
152
95% UCL
13
11
11
11
11
9.8
9.3
9.3
9.4
9.0
8.5
8.3
8.2
8.2
8.0
7.9
7.7
7.7
7.6
7.5
7.4
7.2
7.1
7.1
7.1
7.1
NOAEL
152
Average
121
109
103
107
100
93
88
89
89
85
80
78
77
77
76
74
72
72
72
71
70
68
67
66
66
66
NOAEL
152
95% UCL
127
114
109
112
105
98
93
93
94
90
85
83
82
82
80
79
77
77
76
75
74
72
71
71
71
71
LOAEL
113
Average
9.7
9.1
8.5
8.4
8.1
7.6
7.3
7.1
7.1
6.8
6.6
6.5
6.3
6.2
6.1
6.1
5.9
5.9
5.8
5.7
5.6
5.6
5.5
5.3
5.3
5.3
LOAEL
113
95% UCL
9.7
9.2
8.6
8.4
8.1
8.0
7.7
7.5
7.4
7.2
6.9
6.8
6.7
6.6
6.5
6.4
63
63.
63.
6.1
6.0
59
5.8
5.7
5.7
5.7
NOAEL
113
Average
97
91
85
84
81
76
73
71
71
68
66
65
63
62
61
61
59
59
58
57
56
56
55
53
53
53
NOAEL
113
95% UCL
97
92
86
84
81
80
77
75
74
72
69
68
67
66
65
64
63
62
62
61
60
59
58
57
57
57
LOAEL
90
Average
7.7
7.3
6.9
6.6
6.4
6.1
5.8
5.6
5.5
5.4
5.2
5.0
4.9
4.8
4.7
4.6
4.6
4.5
4.4
4.4
4.3
4.3
4.2
4.1
4.0
4.0
LOAEL
90
95% UCL
8.1
7.7
7.3
6.9
6.7
6.4
6.1
5.9
5.8
5.6
5.4
5.3
5.2
5.1
5.0
4.9
4.8
4.8
4.7
4.7
4.6
4.5
4.5
4.4
4.3
4.2
NOAEL
90
Average
77
73
69
66
64
61
58
56
55
54
52
50
49
48
47
46
46
45
44
44
43
43
42
41
40
40
NOAEL
90
95% UCL
81
77
73
69
67
64
61
59
58
56
54
53
52
51
50
49
48
48
47
47
46
45
45
44
43
42
LOAEL
50
Average
5.9
5.6
5.3
5.1
4.8
4.6
4.4
4.3
4.1
4.0
3.9
3.7
3.7
3.6
3.6
3.5
3.4
3.4
3.3
3.3
3.2
3.2
3.1
3.1
3.0
3.0
LOAEL
50
95% UCL
6.2
5.9
5.5
5.3
5.0
4.8
4.7
4.5
4.3
4.2
4.1
3.9
3.9
3.8
3.8
3.7
3.6
3.6
3.5
3.5
3.4
3.4
3.3
3.3
3.2
3.2
NOAEL
50
Average
59
56
53
51
48
46
44
43
41
40
39
37
37
36
36
35
34
34
33
33
32
32
31
31
30
30
NOAEL
50
95% UCL
62
59
55
53
50
48
47
45
43
42
41
39
39
38
38
37
36
36
35
35
34
34
33
33
32
32
Bold values indicate exceedances
-------
TABLE 5-49: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE MINK FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
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
1.4
1.1
1.0
1.2
1.0
0.8
0.8
0.7
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
0.5
0.4
0.4
0.4
LOAEL
152
95% UCL
1.5
1.1
1.0
1.3
1.1
0.9
0.8
0.8
0.8
0.8
0.7
0.6
0.6
0.6
0.6
0.6
0.5
0.6
0.6
0.5
0.6
0.5
0.5
0.5
0.5
0.5
NOAEL
152
Average
46
35
32
39
33
27
25
24
26
24
22
18
18
20
18
17
15
17
17
17
18
16
15
14
14
14
NOAEL
152
95%UCL
48
37
34
41
35
28
27
25
27
25
23
19
19
21
19
18
16
18
18
18
19
17
16
15
15
15
LOAEL
113
Average
1.2
1.1
0.9
1.0
0.9
0.8
0.7
0.7
0.7
0.7
0.6
0.6
0.5
0.5
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
1.3
1.2
1.0
1.0
0.9
0.8
0.7
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
NOAEL
113
Average
40
36
30
32
29
26
23
23
23
22
20
18
18
18
17
16
15
15
16
16
16
15
15
14
13
13
NOAEL
113
95%UCL
41
37
32
33
31
27
24
24
24
23
21
19
18
18
18
17
16
16
17
16
16
16
15
14
14
14
LOAEL
90
Average
1.0
0.9
0.8
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
0.3
0.3
0.3
0.3
0.3
LOAEL
90
95%UCL
1.0
0.9
0.8
0.8
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
NOAEL
90
Average
31
29
26
24
22
21
19
17
17
16
16
15
14
13
13
13
12
12
12
12
12
11
11
11
10
9.9
NOAEL
90
95% UCL
33
30
27
25
23
22
20
18
18
17
16
15
14
14
14
13
12
12
12
12
12
12
12
11
11
10
LOAEL
50
Average
0.9
0.8
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
LOAEL
50
95% UCL
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
0.3
0.3
0.3
NOAEL
50
Average
29
26
23
21
19
18
16
15
14
14
13
12
12
11
11
10
9.8
9.6
9.5
93
9.2
9.1
8.9
8.6
83
8.1
NOAEL
50
95% UCL
30
27
24
22
20
19
17
16
15
14
14
13
12
12
11
11
10
10
9.9
9.7
9.6
9.5
9.3
9.0
8.7
8.4
Bold values indicate exceedances
-------
TABLE 5-50: RATIO OF MODELED DIETARY DOSE TO TOXICITY BENCHMARKS
FOR FEMALE OTTER FOR TRI+ CONGENERS FOR THE PERIOD 1993 - 2018
REVISED
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
73
6.1
7.9
6.8
5.4
4.6
4.4
4.9
4.5
4.0
3.0
3.0
3.5
3.0
2.8
2.5
2.6
3.1
2.6
3.0
2.7
2.4
2.2
2.0
1.9
LOAEL
152
95% UCL
10
7.4
6.2
8.0
6.9
5.5
4.7
4.5
5.0
4.6
4.0
3.1
3.1
3.5
3.1
2.8
2.5
2.7
3.1
2.7
3.1
2.7
2.5
2.2
2.0
2.0
NOAEL
152
Average
335
236
198
257
220
177
149
144
160
147
129
99
99
112
99
91
81
86
100
86
98
87
79
72
65
63
NOAEL
152
95%UCL
341
240
201
260
224
180
151
146
163
150
131
101
100
114
100
92
82
87
102
87
100
88
81
73
66
64
LOAEL
113
Average
8.0
6.9
5.7
5.8
5.4
4.7
4.1
3.8
3.9
3.8
3.4
3.0
2.7
2.8
2.7
2.6
23
23
2.4
2.3
2.4
2.3
2.2
2.0
1.8
1.8
LOAEL
113
95%UCL
8.1
7.0
5.8
5.9
5.5
4.8
4.1
3.8
4.0
3.9
3.5
3.0
2.7
2.8
2.8
2.6
23
23
2.5
2.4
2.4
23
2.2
2.1
1.9
1.8
NOAEL
113
Average
260
223
186
190
177
153
133
122
127
123
111
97
88
90
88
83
74
73
79
76
78
74
70
66
60
58
NOAEL
113
95% UCL
264
227
189
193
180
155
135
124
.129
125
113
99
89
92
90
85
75
74
81
77
79
75
71
67
61
59
LOAEL
90
Average
6.1
5.5
4.9
4.4
4.1
3.7
33
3.0
2.8
2.8
2.6
2.4
2.2
2.1
2.0
1.9
1.8
1.7
1.7
1.7
1.7
1.7
1.6
1.6
1.5
1.4
LOAEL
90
95% UCL
6.2
5.6
5.0
4.4
4.2
3.8
3.4
3.0
2.9
2.8
2.6
2.5
2.2
2.1
2.0
2.0
1.8
1.7
1.8
1.7
1.7
1.7
1.7
1.6
13
1.4
NOAEL
90
Average
199
180
159
142
133
121
108
97
92
90
85
79
71
67
65
63
59
56
56
56
56
54
53
51
48
45
NOAEL
90
95% UCL
203
183
162
145
135
123
110
98
94
91
86
80
72
68
66
64
60
57
57
57
57
55
54
52
48
46
LOAEL
50
Average
6.0
5.3
4.7
4.2
3.8
3.4
3.1
2.8
2.6
2.4
23
2.2
2.0
1.9
1.8
1.7
1.6
13
1.5
1.4
1.4
1.4
1.4
1.3
13
1.2
LOAEL
50
95% UCL
6.1
5.4
4.8
4.2
3.8
33
3.1
2.8
2.6
2.5
23
2.2
2.0
1.9
1.8
1.7
1.6
13
13
13
13
1.4
1.4
1.4
13
1.2
NOAEL
50
Average
195
172
152
135
123
112
101
91
84
79
75
70
65
60
57
55
51
49
48
47
47
46
45
43
41
40
NOAEL
50
95% UCL
199
175
155
138
125
113
102
93
85
81
76
72
66
61
58
55
52
50
49
48
47
47
46
44
42
40
Hold values indicate exceedances
-------
TABLE 5-51: RATIO OF MODELED DIETARY DOSES TO TOXICTTY BENCHMARKS
FOR FEMALE MINK ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
REVISED
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
265
204
185
226
191
154
146
135
150
137
123
105
105
114
103
98
87
98
99
97
101
93
85
79
79
81
NOAEL
152
95% UCL
277
212
192
236
199
160
152
141
156
143
129
110
110
120
108
102
92
103
104
101
106
98
90
83
83
85
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
230
209
175
184
169
149
134
131
131
125
117
105
101
102
99
94
87
88
91
90
89
87
83
78
75
75
NOAEL
113
95% UCL
239
216
182
191
176
155
139
135
136
129
121
109
105
106
103
98
90
91
95
93
93
91
87
81
78
78
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
181
165
148
138
129
120
109
100
98
95
90
83
79
77
75
72
68
67
68
67
66
65
63
60
58
57
NOAEL
90
95% UCL
188
171
154
143
134
124
113
104
101
98
93
87
82
80
78
75
71
70
70
69
69
68
66
63
61
59
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
167
149
134
122
111
103
95
87
82
79
75
70
67
64
61
59
56
55
54
53
53
52
51
49
48
46
NOAEL
50
95% UCL
173
155
139
126
116
107
98
91
85
82
78
73
69
66
64
62
58
58
56
56
55
54
53
51
50
48
Bold values indicate exceedances
-------
TABLE S-S2: RATIO OF MODELED DIETARY DOSES TO TOXICITY BENCHMARKS
FOR FEMALE OTTER ON A TEQ BASIS FOR THE PERIOD 1993 - 2018
REVISED
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
1954
1380
1155
1499
1284
881
795
775
858
795
696
532
543
618
549
488
434
469
545
492
547
514
437
395
388
398
NOAEL
152
95% UCL
1988
1400
1175
1519
1306
885
774
756
838
772
679
517
526
602
531
474
423
457
530
474
528
493
423
381
370
377
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
1517
1304
1084
1107
1032
892
774
714
743
719
647
568
511
527
515
486
430
428
463
441
456
432
409
386
351
339
NOAEL
113
95% UCL
1542
1327
1104
1124
1050
908
787
726
755
731
658
578
520
536
524
495
437
435
471
448
463
439
416
392
357
345
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
1164
1051
927
830
777
707
631
565
538
523
494
458
415
389
379
365
342
326
329
325
325
317
308
296
277
264
NOAEL
90
95% UCL
1183
1068
943
844
789
719
641
574
547
532
502
466
421
396
385
371
348
332
334
331
330
323
313
301
282
269
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
1141
1004
887
791
716
651
588
532
489
463
439
411
379
351
333
318
298
285
280
275
272
267
262
253
242
231
NOAEL
50
95% UCL
1160
1021
902
804
728
662
597
540
497
470
446
418
385
357
339
324
303
290
285
279
277
272
266
257
246
235
Bold values indicate exceedances
-------
RM 152
0.06
„ 0.05
y = -0.000138+ U R = 0.996
T1 ' ' ' I ' ' ' ' I ' ' ' ' I
0 0.01 0.02 0.03 0.04 0.05 0.06
Farley Model Output (USEPA, 1999)
RM 90
0.03
„ 0.025-
a
O,
g
2 0.02-
u
•o
o
>,0.015
•£ o.oi:
>
u
01 0.005
' y =-0.000209 + 1.01 x R = 1
0.005 0.01 0.015 0.02 0.025 0.03
Farley Model Output (USEPA, 1999)
0.035
0.03-
D.
50.025
2 0.02-
1 0.015
a,
-o
.2 0.01
0.005
0.041
RM 113
"y =-0.000186 + Ix R= 0.999
-rrjTT i i | i ,
0 0.005 0.01 0.015 0.02 0.025 0.03 0.035
Farley Model Output (USEPA, 1999)
RM 50
0 0.005 0.01 0.015 0.02 0.025 0.03 0.035 0.04
Farley Model Output (USEPA, 1999)
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000a
Notes:
1. The x-axis is the model data used in the Addendum to the Baseline Ecological Risk Assessment
for Future Risks in the Lower Hudson River (USEPA, 1999c).
2. The y-axis is the revised Farley fate and transport model with the Upper River Loads from the
RBMR (USEPA, 2000a).
3. Monthly arithmetic averages are shown.
Figure ni-1
Comparison for Tri+ PCBs in the Dissolved Phase of the Water Column Between the
Revised Model Output versus the Data Presented in the
Lower River Ecological Risk Assessment
TAMS/MCA
-------
RM 152
RM 113
0.07
0.06
0.04:
0.03:
.2 0.02
0.01
0.08
0.07:
0.06-
0.05-
0.04
0.03-
0.02:
0.01
0
" y = -0.000753 + 1.0 U R2= 0.994
0 0.01 0.02 0.03 0.04 0.05 0.06 0.07
Farley Model Output (USEPA, 1999)
RM 90
"y =-0.000735 + l.Olx R2=
TT-J I I njrr I I | i i
0 0.01 0.02 0.03 0.04 0.05 0.06 0.07 0.08
Farley Model Output (USEPA, 1999)
0.08
y =-0.000682+ l.Olx R2= 0.999
I | I I I I [ I
0 0.01 0.02 0.03 0.04 0.05 0.06 0.07 0.08
Farley Model Output (USEPA, 1999)
RM 50
0.06
~ 0.05-
3
O.
a
2 0.04
u
-o
o
>, 0.03"
u
n
i
"g 0.02:
>
4>
^ o.oi-
"y = -0.000686 + l.Olx R'= 1
•
0 0.01 0.02 0.03 0.04 0.05 0.06
Farley Model Output (USEPA, 1999)
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000
Notes:
1.
The x-axis is the model data used in the Addendum to the Baseline Ecological Risk Assessment
for Future Risks in the Lower Hudson River (USEPA, 1999c).
2. The y-axis is the revised Farley fate and transport model with the Upper River Loads from the
RBMR (USEPA, 2000a).
3. Annual arithmetic averages are compared.
Figure EI-2
Comparison for Total PCBs in the Water Column (Whole Water) Between the Revised
Model Output versus the Data Presented in the
Lower River Ecological Risk Assessment
TAMS/MCA
-------
RM 152
RM 113
O
u.
•a
y = -0.0141 + 1.02x R'= 0.998
0 0.5 1 1.5 2
Farley Model Output (USEPA, 1999)
RM 90
1.5
0.5
y = -0.0177 + 1.02x R = 1
0 0.5 1 1.5
Farley Model Output (USEPA, 1999)
T)
O
I
•o
0 0.5 1 1.5
Farley Model Output (USEPA, 1999)
RM 50
1.2
_ 08-
0.6-
0.4-
0.2
'y = -0.0137+ 1.03x R =
0 0.2 0.4 0.6 0.8 1 1.2
Farley Model Output (USEPA, 1999)
Notes:
I.
Sources: Farley et al., 1999, Hudson River Database Release 4.1 and USEPA, 2000a
The x-axis is the model data used in the Addendum to the Baseline Ecological Risk Assessment
for Future Risks in the Lower Hudson River (USEPA, 1999c).
4. The y-axis is the revised Farley fate and transport model with the Upper River Loads from the
RBMR (USEPA, 2000a).
5. Annual arithmetic averages are compared.
Figure m-3
Comparison for Total PCBs in the Sediment (0-2.5 cm) Between the Revised Model
Output versus the Data Presented in the Lower River Ecological Risk Assessment
TAMS/MCA
-------
Comments
-------
Federal
-------
4S. DEPARTMENT OF COMMERCE
h ational Oceanic and Atmospheric Administration
N itional Ocean Service
C fice of Response and Restoration
C >astal Protection and Restoration Division, 290 Broadway. Rm 1831
Nbw York, New York 10007
EF-1
Januarys, 2000
Alison Hess
U.S. EPA
Sediment Projects/Caribbean Team
290 Broadway, 19th Floor
New York, NY 10007
Dear Alison:
Thank you for the opportunity to review the December 1999 Phase 2 Report - Review Copy,
Further Site Characterization and Analysis, Volume 2E - Baseline Ecological Risk Assessment for
Future Risks in the Lower Hudson River, Hudson River PCBs Reassessment RI/FS. The
following comments are submitted by the National Oceanic and Atmospheric Administration
(NOAA).
Background
The primary objectives of the addendum to the baseline ecological risk assessment (ERA) are to
quantify future risks to selected biological receptors and communities exposed to releases of PCBs
in the lower tidal estuarine portion of the river between Federal Dam and the Battery in the absence
of remediation. The upper freshwater non-tidal portion of the river between Federal Dam and
Hudson Falls Hudson River was the primary focus of a August 1999 ERA although the Lower
Hudson was also evaluated.
Modeled concentrations of PCBs in sediment, water column, striped bass and white perch were
obtained from a model developed by Farley et al. 1999. Future concentrations of fish were also
derived from a modified FISHRAND model (EPA 1999,2000). White perch was the only species
whose concentration was estimated from both models. Risk evaluations were based on future
exposure from sediment, water and fish.
Summary
The Hudson River Superfund Site encompasses the 200 miles of the Hudson from the Verrazano
to Hudson Falls, encompassing freshwater, brackish and estuarine habitats. The ERA Addendum
focuses on Lower Hudson River: PCBs were examined as total PCBs (expressed as tri+ PCBs)
and toxic equivalents (TEQs).
Eight species of fish comprised of foragers, omnivores, semi- piscivores and piscivores were
evaluated. Measured PCB tissue contaminant levels were utilized for all species except the
federally endangered shortnose sturgeon for which body burdens were modeled. Five species
each of birds and four species of mammals were evaluated to represent various trophic positions.
Risks to benthic invertebrate communities were determined by comparing modeled concentrations
of sediment and water to existing guidelines, standards and criteria. Toxic reference values
-------
NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999 1/28/00
(TRVs) based on body burden or dietary dose were selected for survival, growth and reproductive
endpoints of fish, birds and mammals. '
Selected fish NOAELs ranged from 0.16 to 3.1 mg/kg wet weight PCBs. Fish LOAELs ranged
from 1.5 to 15 mg/kg wet weight PCBs. Fish egg NOAELs, which were reported as lipid-
normalized TEQ concentrations, ranged from 0.29 to 8 ug/kg lipid and LOAELs ranged from 0.6
to 103 ug/kg lipid. Based on whole body concentrations, the lowest TRVs were calculated for
pumpkinseed, yellow perch, white perch, largemouth bass, striped bass, brown bullhead and
shortnose sturgeon. Spottail shiner had the highest whole body. Brown bullhead had the highest
TEQ TRY based on NOAELs while spottail shiners had the highest based on LOAELs. All other
fish species had the lowest TRY for TEQs.
TRVs were developed for each bird species based on dietary dose and egg concentration for total
PCBs and TEQs from available field and laboratory studies. In general, TRVs from laboratory
studies were lower than those derived from field studies. Based on lab studies, NOAELs for total
dietary PCBs ranged from 0.01 to 0.26 mg/kg/day and the LOAELs ranged from 0.07 to 2.6
mg/kg/day. The NOAEL and LOAEL for total PCBs in eggs was 0.33 and 2.21 mg PCBs/kg egg,
respectively. TRVs derived from dietary TEQs were lower than those from egg TEQs. The
highest TRVs were associated with the tree swallow for field conducted diet and egg concentration
studies.
Mink and otter were more sensitive to dietary intake (lab studies) of PCBs (NOAEL=0 01 mg
PCBs/kg/day; LOAEL 0.07 mg PCBs/kg/day) than raccoon or little brown bat (NOAEL=0.032 mg
PCBs/kg/day; LOAEL 0.15 mg PCBs /kg/day). The NOAEL and LOAEL across species based on
laboratory dietary doses of TEQ was the same (NOAEL= 0.0001 ug TEQ/kg/day; LOAEL=0.001
ug TEQ/kg/day). The total PCB and TEQ NOAELs for mink and otter developed from field
studies was up to an order of magnitude lower than those based on laboratory toxicity studies.
A weight of evidence approach was followed to assess risks of adverse effects to receptors of
concern exposed to Hudson River PCBs. Assessment endpoints were evaluated against the
various lines of evidence available. The following conclusions were drawn about future exposure
to Lower Hudson River PCBs:
• risks to fish and wildlife are greatest in the upper reaches of the Lower Hudson River and
decrease downstream with concomitant decreasing PCB concentrations,
• many species are expected to be at risk at least through the year 2018 - the upper bound of the
forecasting exercise,
• modeled sediment and water concentrations generally exceed existing guidelines, standards and
criteria for the protection of aquatic health;
• animals using areas designated as significant habitats may be adversely affected by PCBs,
• PCBs may adversely affect survival, growth, and reproduction of fish, especially the higher
trophic levels;
• PCBs may adversely affect survival, growth, and reproduction of waterfowl, omnivorous, and
piscivorous birds while no risk is expected for insectivorous birds;
• PCBs may adversely affect survival, growth, and reproduction of insectivorous, omnivorous,
and piscivorous mammals; and
• threatened and endangered species are particularly susceptible.
-------
NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999
1/28/00
EF-1.1
EF-1.2
Comments
The Baseline Ecological Risk Assessment Addendum (ERA) is a companion document to the
August Hudson River ERA. The report determines the future risks in the Lower Hudson River.
Problem formulation including assessment and measurement endpoints, the exposure assessment
including modeled exposure concentrations and exposure pathways, th^ effects assessment
including development and selection of TRVs, risk characterization including an evaluation of the
assessment endpoints and an uncertainty analysis are described. Overall the document is well-
organized and clearly written. *
The fate and transport and bioaccumulation modeling presented in the Baseline Modeling Report
plus Farley's model for the Lower Hudson provides the primary exposure information for the ERA
Addendum. While the ERA Addendum describes the quantification of PCB fate and transport and
discusses the modeled exposure concentrations (sediment, water, benthos, fish), there is no
substantial discussion of the limitations of the models. Moreover, revisions to the BMR will be
released in a report at the end of January. Have these modifications been accounted for in the
predictions contained within the ERA Addendum? If not, how will these changes impact
predictions of risk?
The Farley model relies on HUDTOX load estimates at the Thompson Island Dam (TIP) to predict
sediment, water and fish (white perch, striped bass) PCBs in the Lower Hudson. This required
the conversion of tri+ PCB loads to homologue loads at the Federal Dam because dichloro through
hexachloro homologues are state variables in the Farley model of the Lower Hudson River. The
conversion process is not clearly explained. What are die uncertainties associated with this
conversion and the potential implications for the predicted sediment, water and fish tissue
concentrations.
There are a number of aspects of the Hudson River system that the fate and transport and
bioaccumulation models are not addressing. For example, as identified in our comments to the
May 1999 Baseline Modeling Report (7/1/99) and the August 1999 Ecological Risk Assessment
(9/7/99), potential effects of daily changes in water level on nearshore shallow-water PCB deposits EF-1.3
and non-scour related movement of PCB-contaminated sediment may result in significant
underestimation of resuspension of sediments and/or PCB loading to the river. This represents
major uncertainty in the exposure assessment for the risk assessment, since the future sediment,
water, and fish tissue PCB concentrations forecasted by these models employ HUDTOX loads at
Federal Dam to predict future risk in the Lower Hudson River. The implications of the uncertainty
resulting from the model inputs to risk assessment should be addressed.
The food chain modeling (FISHRAND) used a generic growth rate for lake trout as an input EF-1.4
parameter for the species of fish modeled rather than attempting to capture the difference in their
growth. Did Farley et al. use species-specific growth rates for white perch and striped bass? How
sensitive are the two models to growth rates and what implications does this have for predictions of
future PCBs in fish and risks to receptors?
Water column and sediment data used in the exposure assessment are averaged without regard for EF-1.5
habitat occupied by the receptors of concern or changes in physico-chemical conditions of the
river. In addition, TOC and lipid content were set to a single value for the entire Lower Hudson.
How sensitive is the model output to these assumptions and what implications does this have for
the derived risks?
The risk assessment did not provide clear criteria for selection of laboratory studies that are used to
define TRVs for fish species other than giving preference to studies on closely-related species. EF-1.6
Because of the importance of the TRV in the determination of risk and all of the uncertainty
associated with the selection of appropriate TRVs, relying on one or two laboratory studies to
determine the TRVs for effects in fish should be evaluated in the context of other studies,
particularly considering the limited number of studies available. For example, the selection of the
-------
NOAA commenis on Hudson River Ecological Risk Assessment Addendum, December 1999 1/28/00
toxicity reference values (TRY) for fish total PCS body burden relied to a great extent on a single
study (Bengtsson 1980) that used a commercial mixture (Clophen A50) which was not available in
the United States and is different from mixtures used in the Hudson River. Other studies (Hansen
et al. 1974, USAGE 1988) generated similar NOAELs and LOAELs and should have been
presented to support the laboratory- and field-based TRVs.
Field- and laboratory-based fish results were handled differently (Section B.2.1). When a
laboratory study could not be identified for a particular fish species or one in the same taxonomic EF-1.7
family, an interspecies uncertainty factor was applied to derive a TRY. For field-based studies,
NOAELs and LOAELs were only developed for the species studied or one within the same family.
An interspecies uncertainty factor was not applied if the receptor of interest belonged to a different
family. An explanation for the different data treatments should be provided.
Toxicity quotients were calculated using the field TRY as the denominator when both field and
laboratory TRVs were developed for the same fish species. An alternative approach would have
been to develop toxicity quotients from field and lab TRVs to bracket risk.
The TRVs developed for Hudson River total PCS body burdens in fish focus solely on growth,
survival and reproductive capacity. They may underestimate risk because they do not consider
other adverse effects such as immune suppression reported in the literature. The uncertainty kl1 -1.8
section should address these other adverse effects since they are not accounted for in the TRV
selection process. Additionally, modeling results tend to underestimate fish concentrations which
would result in lower toxicity quotients. This could also contribute to an underestimation of risk.
Specific Comments
Executive Summary
Pages ES-9 and ES-11: Statements about bald eagles appear to be inconsistent. Page ES-11 states EF-1.9
that "future PCB exposures (predicted from 1003 to 2018) are not expected to be of a sufficient
magnitude to prevent reproduction or recruitment" appears to contradict page ES-9's reference to
the lack of breeding success.
Chapter 3
Page 15 Para 3: The upstream boundary conditions used in the BMR assumes that flow-related EF-1.10
changes (increases) in loading during high flow events will not occur. The BMR model does not
address potential impact of high flow events on the Interim Cap on the Remnant Deposits or other
areas of high concentrations of PCBs that may remain between the plant sites and Rogers Island.
Data from the January 1999 high flow event suggest that setting the upstream boundary at 10 ng/1
could underestimate the loading. This uncertainty should be addressed.
Pages 13, 14, 16, 17 , 24 and 26: Farley et al. (1999) and FISHRAND model output parameters EF-1.11
are stated but the descriptions are inconsistent. For example, the Farley et al. (1999) model
generated sediment, water and fish (white perch) PCB concentrations (Section 3.0, Para 5), and
sediment, water and fish (striped bass, white perch) PCB concentrations (Section 3.1.1; Section
3.1.1.3, Para 2). In the case of FISHRAND model, PCB body burdens were calculated for all
fish (Section 3.0, Para 2; Section 3.1.1), and all fish receptors except striped bass (Section
3.1.1.2, Para 1; Section 3.1.2.3; Section 3.2.4).
Page 16, Estimation of striped bass body burdens: It is not clear whether this estimation was EF-1.12
based on wet weight or lipid-normalized values.
-------
NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999 1/28/00
Page 20 Para 2: It is suggested that the Farley et al. model overestimates loss of PCBs from the EF-1.13
water column during the summer. The importance of this finding relative to uptake of PCBs by
fish should be discussed.
Page 20 Para 3: According to Figure 3-7, empirical data at two locations fell outside the EF-1 14
boundaries of the modeled data. The text indicates only RM 47.
Pages 22-23: Modeled fish concentrations are compared against empirical data. Does the model , EF-1.15
capture the PCB concentrations reported by NYSDEC for 1998?
Page 25, Para 5: A TOC of 2.5% was used to estimate organic carbon-normalized PCB sediment r^ i i /:
concentrations. What effect does selection of a 2.5% TOC have on the outcome of the EJ< -l.io
comparisons as TOC concentrations in the Lower Hudson ranged from 0.35% to 4.9% for
individual samples? The sediment guidelines section provides PCB guidelines and standards.
Table 3-7 should include sediment guidelines developed by EPA (1993) for protection of fish, FF 1 17
birds and mammals based on TCDD sediment concentrations since PCBs are also evaluated as
TEQs.
Page 25, Para 7: Predicted benthic invertebrate PCB concentrations for the year 1993 could have EF-1.18
been compared to empirical data. Are predictions a reasonable estimate of actual measurements?
Page 27 Para 4: Elsewhere the report indicates that PCB concentrations in striped bass in food EF-1.19
web region 1 are predicted using a striped bass to largemouth bass ratio. The last sentence implies
the ratio was used for all of the Lower Hudson striped bass estimates.
Chapter 4
Table 4-1: The NOAA (1999) report on the effects of PCBs on fish reproduction and development EF-1.20
demonstrates that Hudson River fish contain PCBs at concentrations above levels shown to cause
reproductive and developmental effects and that these effect levels are, in some cases, below the
TRVs presented in Table 4-1.
Table 4-1: Other laboratory (Hansen et al. 1974) and field studies (Adams et al. 1989,1990, FF 1 21
1992, USAGE 1988) document fish effect levels similar to Bengtsson (1980) thereby providing
further weight of evidence to support selection of these TRVs. Hansen et al. (1974) is described
in Appendix B but USAGE (1988) was not. The NOAEL from the USAGE (1988) study on
fathead minnows are 5.6 mg/kg. Dividing by a factor of 10 for all species not in the Cyprinidae
family would result in an NOAEL of 0.56 mg/kg. Hence Hansen et al. (1974), Bengtsson (1980)
and USAGE (1988) yield similar TRVs. The field NOAEL developed for pumpkinseed and
largemouth bass from USAGE (1988) is also comparable to the field NOAEL in Adams et al.
(1989,1990,1992). The field-based NOAEL (0.5 mg/kg PCBs) reported in the ERA Addendum
was based on reproduction (Adams et al. 1989,1990,1992). A lower value (0.3 mg/kg RGBs)
should have been selected based on growth. It should be noted that Adams et al. (1989,1990,
1992) assessed endpoints besides those related to survival, growth and reproduction and these
effects were also observed at concentrations below 0.5 mg/kg PCBs. Adverse effects were
associated with DNA integrity, detoxification enzymes, lipid metabolism, community structure and
histological indices.
Page 36: The discussion on the use of TEFs to derive TEQs should have included an explanation EF-1.22
of data quality issues and the impacts on the TEQ calculations. For example, PCB 126 was
frequently classified as below detection and PCB 81 was not measured. In addition, the congeners
of primary importance (by weight and lexicologically) for the water column were mostly below
detection.
-------
NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999 1/28/00
Pag^s'section 5.1.1.1: Tables 3-2 and 3-3 are actually Tables 3-6 and 3-7. In 1993. mean EF-1.23
PCBs were 1.213 ppm at RM 152,0.828 ppm at RM 113,0.872 ppm at RM 90 and 0.806 ppm at
RM 50. Predicted sediment PCBs (Table 3-6) underestimate empirical means in 1993 by 0.07 to
0 42 ppm For organic carbon-normalized sediments, estimates also underpredict observed
concentrations at RM 152, RM 113 and RM 50 with the greatest difference again observed for RM
50. A TOC of 2.5% was used from Farley's model while average TOC for each of these RM
segments ranged from 2.5% to 3.6% and individual samples ranged from 0.35% to 5.3%.
Page 39 Top: Forecasted sediment concentrations exceeded NYSDEC benthic aquatic life chronic EF-1.24
toxicity criterion at RM 152 and RM 113 for the duration of the modeling period at the 95% UCL.
RM 90 only exceeded criterion until 2011.
Page 39 Eara 2 and Table 5-1: The organic carbon-normalized SEL (Persaud et al. 1993) should EF-1.25
read 13 rag/kg instead of 1.3 mg/kg.
Page 40 .Section 5.1.2.1; Page 44, Section 5.2.2.1; Page 47, Section 5.3.2.1; Page 49, Section FF 1 -,
5 42 1- Page 53, Section 5.5.2.1; Page 55, Section 5.6.2.1; Page 56 Section 5.7.2.1;, Page 59, ^*'*••«>
Section 5821; Page 61, Section 5.9.3.1; Page 63, Section 5.10.2.1: The NYSDEC surface
water standard for protection of wildlife is 1.2 x 10'4 ug/1 total PCBs (NYSDEC 1998). It replaced
the wildlife criterion of 0.001 ug/1 in 1998 (Stoner 2000).
Page 40, Last Para: The analysis on forage fish reproductive effects assumes that measurements of EF-1.27
young-o'f-year spottail shiner and age 1 pumpkinseed are equivalent to concentrations in mature
adults. According to Table 4-1, the TRY for spottail shiner on an NOAEL basis is 1.6 mg/kg not
15 mg/kg as stated in this paragraph. The NOAEL derived from laboratory studies, therefore,
resulted in a TRY for pumpkinseed that is an order of magnitude lower than for spottail shiner. If
comparisons are made between laboratory and field studies then the difference is reduced to 3 fold.
Page 41 Section 5.2.1.2: The authors assume that PCBs partition equally into the lipid phase of EF-1.28
eggs and into the lipid phase of adult fish "tissue". There is good justification for this assumption,
but it does not necessarily follow that it is appropriate to establish TRVs based on lipid-normalized
concentrations. What is the evidence to indicate that lipid-normalized concentrations are more
directly related to the reproductive effects than the wet weight concentrations in eggs?
Page 41, Section 5.2.1.3: RM 133 should read RM 113. This discussion should also note that EF-1.29
since the literature-derived TRVs were based on whole body concentration studies, the fish fillets
were converted to whole body for direct comparison.
Pages 41-43, 52-53,58, 61: Additional measurement endpoints should have included a
comparison of measured and modeled fish TEQ concentrations reported by EPA (1993) to pose a EF-1.30
risk to fish, avian and mammalian receptors. For example, low risk to piscivorous fish was
associated with fish concentration of 50 pg/g TCDD and high risk at 80 pg/g TCDD. Fish TCDD
concentrations of 6 pg/g and 60 pg/g were identified as posing a low to high risk to avian wildlife
respectively; where high risk is defined as causing 50-100% mortality in embryos and young of
sensitive species. For mammalian wildlife, fish TCDD concentrations of 0.7 pg/g pose a low risk
and 7 pg/g pose a high risk.
Page 42, Section 5.2.1.5: An NOAEL from field data was developed for white perch but not for EF-1.31
yellow perch since no field studies were identified that examined effects of PCBs on yellow perch
or on a species in the same genera or family. It is not clear why the values for white perch or other
species could not be used with the application of an uncertainty factor, as done for laboratory
studies. If this was done, how would the TRVs change and what impact would it have on the
toxicity quotients?
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NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999 1/28/00
Pages 48,50, and 54: Current trends in bird usage should be compared to historical usage, EF-1.32
especially prior to GE's use of PCBs at their two Hudson River facilities.
Chapter 6
Page 69: While the ERA attempts to address effects associated with congeners eliciting dioxin-like EF-1.33
behavior, it does not attempt to examine effects from congeners that have different mechanisms of '
action. This potentially further underestimates risk to receptors associated with releases from the
GE facilities and should be addressed within the uncertainty section.
Page 71 Top: While the model may be able to reproduce general trends, it was incapable of EF-1.34
picking up year to year changes in fish concentrations and generally underestimated average
surface sediment concentrations in 1993, the only year data are available for the Lower Hudson.
Page 71 Para 1: There is also uncertainty associated with changes in upstream boundary
conditions and potential releases from the remnant deposits.
Page 71-72, Section 6.3.2.2: The report discusses sensitivity analyses for avian and mammalian EF-1.36
receptors and refers to the BMR for a more detailed analysis of uncertainty and sensitivity in the
FISHRAND model. A sensitivity analysis should be conducted for fish toxicity values taking into
account exposure parameters (i.e., growth rates, lipid), TRVs (i.e., assuming 1:1 egg:tissue for
TEQs, fillet to whole body ratios) and exposure media concentrations (i.e., annually averaged-
sediment and summer averaged-water column concentrations).
Chapter 7 ' _ t .-
Page 75, Para 4: The report states that the uncertainty in sediment and water forecasts is u,r -i.J /
approximately a factor of two. The basis for this statement should be explained.
Page 76: The sensitivity of tree swallows to PCBs should be compared to other insectivorous EF-1.38
avian species since the lack of predicted risk may underestimate threats (impairment of survival,
growth, reproduction) to others in the same feeding guild.
Appendix A
Page A-6 Para 1: "However, the 2.4 percent summer-to-spring change in hexachloro homologue EF-1.39
ratio". The 2.4% change was between fall-winter and spring. Between summer and spring the
change was smaller (1.4%).
Appendix B
Page B-10 Para 1: See comment above on Page 41, Section 5.2.1.2.
Pages B-10 to B-ll, Section B.2.3.1, Para 1: Neither Hansen et al. (1971) nor Hansenetal. EF-1.41
(1974) focused on adult mortality. Hansen et al. (1971) evaluated toxicity of Aroclor 1254 to
juvenile spot and Hansen et al. (1974) examined the effect of Aroclor 1254 on the eggs of
sheepshead minnow. Their 1974 study represents a more sensitive endpoint than the (1971) study
and should have been considered in the development of TRVs.
Pages B-10 to B-ll, Section B.2.3.1, Para 2, and Sections 2.3.3-2.3.8: Hansen et al (1974)
established an NOAEL of 1.9 mg/kg PCBs and a LOAEL of 9.3 mg/kg PCBs based on early life
stage survival, where TRVs were based on adult female sheepshead minnow concentrations that
were directly associated with effects on their respective eggs.
The field study by USAGE (1988) with fathead minnows should also be described since the
toxicity endpoint was reproductive success and the NOAEL 5.6 mg/kg. For less closely related
species (non-Cyprinidae), the NOAEL be 0.56 mg/kg, upon employing an uncertainty factor of
10.
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NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999
1/28/00
Pages B-10 to B-l 1, Section B.2.3.1, Para 3, and Sections 2.3.3-2.3.8: NOAA does not support EF-1.42
the rationale for selecting Bengtsson (1980)w over Hansen et al. (1974)°°. The Hansen et al.
(1974) laboratory study should have been selected along with Bengtsson (1980) and the USAGE
(1988) field study along with Adams et al. (1989,1990,1992). For spottail shiner, the resultant
laboratory NOAELs and LOAELs would be 1.6ft) and 1.9*1 mg/kg PCBs and 9.3(b) and 15(1) mg/kg
PCBs, respectively, based on these two studies. All of the other seven fish receptors would have
NOAELs and LOAELs an order of magnitude lower (NOAEL: 0.16(1) and 0.19* mg/kg PCBs;
LOAEL: O^S^and 1.5W mg/kg PCBs). The field-based NOAELs (mg/kg PCBs) were 3.1 for
white perch and striped bass (Westin et al. 1983), 0.3 for pumpkinseed and largemouth bass
(Adams et al. 1989,1990,1992), and 5.6 for spottail shiner (USAGE 1988). Values are
summarized in the table below.
TRVs for PCBs
• (mg/kg PCBs wet WL)
White Perch, Largemouth
Striped Bass Bass,
Pumpkinseed
Spottail Yellow Perch,
Shiner Brown Bullhead,
Shorinose Sturgeon
References
Tissue Concentration
Lab-based NOAEL 0.16 0.16 1.6
Lab-based NOAEL 0.19 0.19 1.9
0.16 Bengtsson 1980
0.19 Hansen etaJ. 1974
Field-based
Field-based
Field-based
NOAEL
NOAEL
NOAEL
3.1
0.03
0.56
0.31
0.3
0.56
0.31
0.03
5.6
0.31
0.03
0.56
1983
Westin et al.
Adams etal.
1989,1990.1992
USACE 1988
Lab-based
Lab-based
LOAEL
LOAEL
0.93
1.5
0.93
1.5
9.3
15
0.93 Hansen etal. 01974
1.5 Bengtsson 1980
Pages B-ll, Section B.2.3.1, Para 3, and Sections 2.3.3, 2.3.4, 2.3.6, 2.3.8: Adams et al.
(1989,1990,1992) also documented reductions in length, weight and growth potential
(RNA/DNA ratio). The NOAEL for growth was 0.3 mg/kg PCBs while the NOAEL of 0.5 ppm
PCB was based on fecundity (clutch size). In addition, other adverse effects were reported at
concentrations lower than the NOAEL selected for lethality, growth and reproduction. These
included significant differences between reference and contaminated site fish for the following
parameters: DNA integrity (strand breaks), detoxification enzymes (P450, CBS, NADPH,
EROD), histological indices (liver and spleen parasites, macrophage aggregates in the liver,
necrotic liver parenchyma) and lipid metabolism (serum triglycerides, body triglyceridles,
phospholipids). Still, other effects (i.e., species richness, total body lipid, liver-somatic index)
were observed but results were either not significantly different from the reference or statistical
analyses were not presented.
Page B-17. The NOAEL and LOAEL are transposed. The LOAEL TRY for white perch should EF-1.43
read 0.6 ug TEQs/kg lipid. The NOAEL TRY for white perch should read 0.29 ug TEQs/kg lipid.
Tables B-5 and B-6 should include USACE (1988). EF-1.44
Table B-6: This table lists Adams et al. (1989,1990,1992) as providing EL-no effect and EL- EF-1.45
effect while the text in Appendix B indicates that an NOAEL and LOAEL was derived from these
studies. The Adams et al. (1989,1990,1992) field investigations evaluated redbreast sunfish
from four sites along the East Fork Poplar Greek. Contaminant concentrations and observed
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NOAA comments on Hudson River Ecological Risk Assessment Addendum, December 1999 1/28/00
effects encompass a range of values and should be considered representative of NOAEL/LOAELs
rather than EL-no effect or EL-effect.
Table B-7: It is not clear from the table whether lipid values were reported by the referenced study > FF-1 46
or estimated by the authors of the report. If estimated, the estimation procedure and the potential
implications should be discussed.
Figure B-2: Selected toxicity endpoims are shown for selected aroclors. Two of them, an NOAEL EF-1.47
of 11.6 mg/kg PCB and an LOAEL of 36 mg/kg PCB in fathead minnow cannot be found by
cross-referencing Table B-5. What is the reference? Why were these endpoints selected over
others presented in Table B-5? Why was the focus of the figure limited to laboratory-based
studies?
Figure B-3: Selected toxicity endpoints are shown for selected fish egg dioxin equivalencies. Why EF-1.48
were these endpoints selected over others presented in Table B-7? Why was the focus of the figure'
limited to laboratory-based studies?
Thank you for your continual efforts in keeping NOAA apprised of the progress at this site. Please
contact me at (212) 637-3259 or Jay Field at 206-526-6404 should you have any questions or
would like further assistance.
Sincerely,
Lisa Rosman
NOAA Coastal Resource Coordinator
References
Hansen, DJ., S.C, Schimmel, and J. Forester. 1974. Aroclor 1254 in eggs of sheepshead minnows: effect on
fertilization success and survival of embryos and fry. Proc. Southeastern Assoc. Came Fish. Comm. pp. 805-812.
NOAA 1999. Reproductive, Developmental and Immunotoxic Effects of PCBs in Fish: a Summary of Laboratory
and Field Studies. Prepared by Emily Monosson for Nauonal Oceanic Atmospheric Administration, March 1999.
NYSDEC 1998. Ambient Water Quality Standards and Guidance Values and Groundwater Effluent Limitations,
Memorandum, Division of Water Technical and Operational Guidance Series (1.1.1), New York State Department of
Environmental Protection, June 1998 (see page 54).
Stoner, Scott 2000. NYSDEC, Division of Water, Personal Communications, Jan 11,2000.
USAGE 1988. Relationship between PCB tissue residues and reproductive success of fathead minnows.
Environmental effects of dredging. Technical notes. EEDP-Ol-13. U.S. Army Corps of Engineers, Engineer
Waterways Experiment Station, Environmental Laboratory. Vicksburg, MS
USEPA 1993. Interim Report on Data and Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to
Aquatic Life and Associated Wildlife, EPA/600/R-93/055, Office of Research and Development, Washington, DC.,
March 1993.
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NOAA comments on Hudson River Ecological Risk Assessment Addendum. December 1999 1/28/00
cc: Mindy Pensak, DESA/HWSB
Gina Ferreira, ERRD/SPB
Roben Hargrove, DEPP/SPMM
Charles Merckel, USFWS
Anne Secord, USFWS
William Ports, NYSDEC
Ron Sloan, NYSDEC
Sharon Shutler, NOAA
10
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State
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New York State Department of Environmental Conservation
Division of Environmental Remediation
Bureau of Central Remedial Action, Room 228
50 Wolf Road. Albany, New York 12233-7010 TT SL1
Phone: (518} 457-1741 • FAX: (518) 457-7925 AJkJ J.
Website: www.dec.3tate.ny.ua
February 4,2000
Allison A. Hess
Project Manager
U.S. Environmental Protection Agency
Region 2
290 Broadway, 19th Floor
New York, New York 10007-1866
Dear Ms. Hess:
RE: Hudson River PCB Reassessment RI/FS
Site No. 5-46-031
The New York State Department of Environmental Conservation has completed its review of the
Phase 2 Report - Further Site Characterization and Analysis, Volume 2E - Baseline Ecological Risk
Assessment (BERA) for Future Risks in the Lower Hudson, Hudson River PCBs Reassessment RI/FS, dated
December 1999. Our comment on the Lower Hudson BERA is provided below.
On page 40, the reference to the "NYSDEC wildlife bioaccumulation criterion of 0.001 ug/L" for
PCB is an older number which has changed. The current number for PCB is 1.2x10-4 ug/L (ppb) for ES-1.1
protection of wildlife which is a promulgated New York State ambient water quality standard (see 6 NYCRR
Part 703).
i In general, we agree with EPA's conclusion that receptors in close contact with the Hudson River
are at an increased ecological risk as a result of exposure to PCBs in sediments, water, and/or prey.
If you have any questions regarding the comments please contact this office at 518-457-5637.
Sincerely,
7. «*
William T. Ports P.E.
Project Manager
Remedial Section A
Bureau of Central Remedial Action
Division of Environmental Remediation
cc: John Davis, NYSDOL
Robert Montione, NYSDOH
Jay Held. NOAA
LisaRosman,NOAA
Anne Secord, USF&WD
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Local
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SARATOGA COUNTY
ENVIRONMENTAL MANAGEMENT COUNCIL
PETER BALET «ORGE HODGSON
CHAIRMAN DIRECTOR
EL-1
January 26,2000
Alison A. Hess, CPG
USEPA, Region 2
290 Broadway, 19th Floor
New York, N.Y. 10007-1866
Dear Ms. Hess:
Enclosed you will find the Saratoga County Environmental Management Council's
(SCEMC's) comments on the Baseline Ecological Risk Assessment For Future Risks in the
Lower Hudson River and the Human Health Risk Assessment for the Mid-Hudson River
prepared by the Council's chief technical advisor, David Adams.
Many of the SCEMC's previous comments on the Hudson River Reassessment's Phase 2
Human Health Risk and Ecological Risk Assessment Reports transmitted to you on September 2,
1999 apply to these reports as well. The Council believes these latest Ecological and Human
Health Assessments also reflect an unrealistic and excessive degree of "scientific"
over-conservatism in calculating the human health and ecological risks.
In the enclosed comments, David Adams makes a number of appropriate and what we
feel are valid observations relating to the unavailability and inconsistencies of important
modeling information not being provided to the public for its review prior to its being used by
EPA in these reports. The unavailability of EPA's revised baseline modeling information and
EPA's lack of agency/peer review of the Farley model are important areas of methodological
concern as these tools are crucial in determining the magnitude of the Reassessment's risk
assessments. The SCEMC requests, at this time, a copy of EPA's revised modeling information
for our review and comment. This information should also be provided to all Reassessment
public information repositories.
Once again, it becomes apparent that EPA has not developed an adequate overall
methodological framework for the Reassessment when it relies on a model (Farley's) to assess
mid and lower river risks which requires PCB monitoring information on a homolog basis rather
than a congener basis which was the type of data collected during the Reassessment monitoring
period This lack of adequate pre-project planning now requires the need for data conversion
which introduces yet "another undefined level of uncertainty into the calculated risks". The
, Council also feels it is inappropriate to utilize a limited number of striped bass samples to draw
what we believe to be erroneous conclusions in regarding PCB concentrations found in
larffemouth bass populations. Again, the need for additional PCB Homolog sampling for
SO WEST HIGH STREET BALLSTON SPA. N.V. 12020 (6181884-4778
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representative fish species found in the mid and lower Hudson River should have been
anticipated and is indicative of the poor methodological planning inherent throughout EPA's
Hudson River PCB Reassessment process.
Peter M.Balet
Chairman
cc:" Doug Tomchuk,USEPA, Region 2
SCEMC Members
Darryl Decker, Chr., Government Liaison Committee, CD*
The Honorable John Sweeney
JobnWanska,USGAO
Dr. George Putman, Scientific & Technical Committee, CIP
William Ports, NYSDEC
Ned Sullivan, Scenic Hudson
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SARATOGA COUNTY
ENVIRONMENTAL MANAGEMENT COUNCIL
PETER BALET GEORGE HODGSON
CHAIRMAN DIRECTOR
COMMENTS ON PHASE 2 - VOLUME 2E
A BASELINE ECOLOGICAL RISK ASSESSMENT FOR FUTURE RISKS
IN THE LOWER HUDSON RIVER
AND ON VOLUME 2F
A HUMAN HEALTH RISK ASSESSMENT FOR THE MID-HUDSON RIVER
HUDSON RIVER PCB'S REASSESSMENT RI/FS
DECEMBER, 1999
Prepared By: David D. Adams, Member, Saratoga County EMC and Government Ualson
Committee, January 2,2000
General Comments
1 Both of these risk assessments and the revised EPA FISHRAND Model for the Upper Hudson River
are based on the revised EPA PCB Fate and Transport Model and the Farley, eL al. Model for the
Lower Hudson River. Reports describing these models and the model results were not made
available by EPA with the risk assessment reports. It Is Improper for EPA to present reports to the EL-1.1
public for review and comments when Information vital to the review is not available to the general
public. Before presenting these reports, EPA should have made the revised EPA model reports and
the Farley, et. al. Model report available In the designated PCB Reassessment repositories for review
along with the risk assessment reports. I was able to obtain a copy of the Fariey, et. al. Model
report through the courtesy of Alison Hess of EPA. Results of my review of the Fariey Model are
presented as appropriate In the comments on the Risk Assessment Reports. My review was
constrained, however, by not having the model revisions made after March, 1999. EPA Is
requested to forward Information on these revisions. I still await the revised EPA model reports
which have not yet been Issued.
2 In EPA's public presentation of the Risk Assessment Reports, EPA stated that EPA does not plan to
review the Farley Model. The reason given was that the Reassessment and subsequent remediation
decision being done by EPA Is for the Upper Hudson only. The logic of this position Is difficult.»
understand. If the risk assessments of the Mid and Lower Hudson are of no significance to EPA s,
study of the Upper Hudson, then why were the risk assessments done? If the results of the risk
assessments may have bearing on EPA's decision about remedial action In the Upper Hudson, then
EPA owes the public the assurance that the risk assessments have been done on a sound basis. This
assurance requires EPA's review of the Farley Model and also review by an appropriate Independent
review panel. EPA Is requested to respond as to the use of these risk assessments and based on that
response, as to whether the Farley Model will be reviewed. While overall the Fariey Model appears
SO west HIOH STREET BALLSTON SPA.tN r 12020 (6181884-4778
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to be a good and credible model, the following are some of my questions/concerns that arose from
my review of the report by Farley, et. al. which illustrate why review of the Farley Model Is needed:
a. The very sharp concentration gradient shown in Fig. 1-1 for di RGB's between RM159 and
RM144 is suspect as it Is not clear what could cause such a gradient. Also, there is no
explanation for the second bar graph at RM159. If this bar graph Is selected, the sharp gradient
for dl disappears. Is it possible there is something wrong with the data presented In the first bar
graph?
b. In many places, values of parameters are stated or assumed with little or no Justification.
Examples are the sediment thicknesses assigned to each model segment (p. 19); the use of the
1989 Mohawk River and Upper Hudson River flows as a constant yearly flow repeated annually
throughout the PCB simulations (P. 24); sedimentation rates, suspended solids concentrations,
settling velocity, suspended sediment loads from the Upper Hudson and Mohawk River during
high and low flow periods, sediment loads from the Lower Hudson Watershed and their
distribution In the model segments (P. 26); production rate of solids by phytoplankton, the
stolchlometrlc conversion factor, the decomposition percentage for phytoplankton, and average-
annual sedimentation rates (P. 27); fraction of organic carbon In sediments (P. 30); the values
for a (P. 56); use of Mohawk River PCB concentrations for Passalc, Hackensack, and Puritan
Rh/ers°(P. 40).
c. The specification rather than modeling of hydrodynamlc, organic carbon, and sediment EL-1.2c
transport (P. 18).
d. The lack of data to support model calculated values (see P. 28 8C Fig. 2-5 where data are
lacking above RM25 for low flow and RM12 for high flow and P. 55 fit Fig. 3-1 where data are
lacking below RM80).
e. The assignment of PCB Initial conditions for sediments for model segments missing sediment
cores. Based on the distribution of cores, It appears only 6 or 7 segments out of 26 segments In tLrl -2e
the model have core data (PP. 41 s 45).
f. There seems to be a very large number of parameter adjustments required to calibrate the bio- EL-1.2f
accumulation model (P. 54).
g. The rather poor fit In several Instances of the data to the model calculations for PCB homologue £L-1.2g
concentrations In surface sediments (P. 59 6C Fig. 3-5).
h. The apparent over prediction of total PCB's In perch (P. 75 fit Fig. 3-14). EL-1.2h
3. EPA also stated in Its public presentation that the only PCB source considered to the Lower Hudson
was the PCB's coming over the Troy Dam. While I could not find an explicit statement in the model
discussion in the Ecological Risk Assessment Report to this effect, the presentation In the Report
appears to be based on the Upper Hudson as the only source to the Lower Hudson. Farley, et. al.
state on P. 41 of their report that while the Upper Hudson dominated the loading to the Lower
Hudson in the early 1990's, the Upper Hudson loads continued to decrease In the 1990's and by EL-1.3
1997 are estimated to be'slightly less than one-half of the total PCB load to the Lower Hudson.
EPA Is requested to justify assuming all the PCB loading comes from the Upper Hudson In view of
the position stated by Farley, et. al. As a minimum, EPA should provide values for the risks
assuming that the Upper Hudson load is eliminated and 50% of the PCB load to the Lower Hudson
remains Into the future as no action to remove these loads appear to be underway. These risk values
would put into proper perspective the possible contribution of PCB loads from the Upper Hudson to
risks in the Lower Hudson.
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referenced as appropriate.
«: The need to convert EPA model Upper Hudson PCB Inputs to the Farley Model from ^congeners EL-1.5
comments on Appendix A.
Vol. 2E Baseline Ecological Risk Assessment Comments
Section 3.1.1.1; P. 15: Please Identify the "few changes" needed to make the Farley Model usable by EL-1.6
en A AIM "PPA h reauested to orovlde an evaluation of the potential effects of starting tne mo
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striped bass are the highest of any of the fish species ranging up to twice the PCB concentration in brown
bullheads which represent the major fraction of fish consumed (52% per Table 2-7 of Vol. 2F). Thus,
the product of the percent species in the diet times the PCB concentration makes striped bass as
significant as brown bullhead In contributing to the human health risk from eating fish. , , . _
EL-l./c
The situation for avian and mammal populations Is less clear. While many Include fish in their diet, in
most cases, but not all, the fish seem to be smaller than striped bass. Because EPA does not provide
definitive information, either In the August, 1999 or December, 1999 reports, It is not possible to
determine the fraction of the avian and mammal receptors diet that Is assumed to come from striped
bass but it Is likely striped bass contribute in EPA's analysis to at least some of the avian and mammal
receptors.
Because of the ma|or significance of striped bass to the risk assessments, it Is very important that proper
selection be made of the modeled PCB concentrations In-striped bass to be used In the risk assessments.
The trend for PCB concentration with decreasing river mile shows declining concentrations with
decreasing river mile until New York City Is reached. Review of Figure 3-18 for largemouth bass from
Vol. 2E (the species EPA uses to estimate striped bass PCB concentrations at RM150 and RM113)
Indicates this decline Is not linear but rather decreases from RM113 to RM90, and finally has a much
more gradual decline from RM90 to RM50. This trend Is Important because of how EPA calculates the
future yearly PCB concentrations In each fish species used In the human health risk assessment. While
not stated, (see comments on Sect. 2.3.1, P. 9 of Vol. 2F) It appears this average is calculated assuming
a linear variation with distance. This assumption would overestimate the PCB concentration In
largemouth bass and therefore striped bass. Use of a technique such as graphical Integration would seem
to be a more appropriate way to calculate the average concentration for these species. It Is also of note
that EPA provides curves vs. time for all fish species at each river mile except for striped bass. EPA Is
requested to provide the curve for striped bass. But of more consequence Is the fact that EPA has
chosen to use striped bass concentrations only at RM152 K 113 In both the ecological and human
health risk assessments, while using concentrations at RM152, RM113, RM90 and RM50 for all other
species in the ecological risk assessment and RM152, RM113, and RM90 In the human health risk
assessment This is done, despite the fact that Farley, et. al. do not even consider striped bass in this
region (Region 1) and the likely sharp drop-off In PCB concentration in striped bass from RM152 to
RM90.
The approach EPA has taken for striped bass is certainly overly conservative and likely Incorrect In
calculating the contribution of striped bass to the risk assessments. EPA should recalculate the risks using
a more accurate approach. It is recommended that EPA use striped bass concentrations at RM90 In the
human health risk assessment, and that the ecological risk to striped bass be evaluated at RM90 and
RM50 as was done for other flsh species. Whether the lack of striped bass PCB concentrations for these
river miles affects the ecological risk to other species at these locations Is unclear because EPA has not
Identified the amount of striped bass in the diets of receptors. In recalculating the PCB concentrations In
striped bass, EPA should also define and account for any size restrictions New York Imposes on catching
and retaining striped bass. Size is related to age and is important because PCB concentration in striped
bass decreases with age due to the migratory nature of striped bass as discussed in the Farley, et. al.
report on P. 78 and shown by Figs. 3-16 through 3-19 of the report. It is my understanding that NYS
limits keeping striped bass to flsh 18" or greater. Fish of this size would be expected to be older than 0-
2 yr. Age class which exhibits peak PCB concentrations. The excess conservatism In the EPA calculation
of PCB concentration in striped bass Is Illustrated by comparing Table 3-18 of EPA's Vol. 2E with Fig.
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3-16 of the Farley report. Table 3-18 shows median values for the years from 1993 to 1997 of 36 to
24 at RM152 and 5 to 3.5 for RM1.13. For fish born In 1987, Fig. 3-16 gives a mean of about 3 for
Food Region 2. Fig. 3-19 shows data points ranging from 1 to 2 (one year about 5) over this time*
period for flsh 6 to 17-years-old.
The use of largemouth bass, which are a non-migratory fish as a surrogate for striped bass, a migratory
flsh, Is In itself questionable. More uncertainty In the calculation for striped bass arises from the large
difference between the ratios of striped bass to largemouth bass PCB concentrations at RM152 (2.5),
and RM113 (.52) (see P.I 7). EPA Is requested to provide an explanation for this difference as there isf
no apparent reason for it. What are the ratios for RM90 and RM50? It Is also of interest that the ratios EL-l./d
(and also those for White Perch) have dropped considerably In recent years. Shouldn't any ratio, If used
to calculate striped bass concentrations, be based on the more recent data for future predictions?
Going back to P. 16, EPA Is requested to explain why the FISHRAND Model was used for all flsh
species except striped bass as again the reasons are not apparent Would using FISHRAND for striped „, . -
bass eliminate or reduce some of the concerns discussed above? Also, Farley, et. al. make a distinction' 1LLr1'
between ages of striped bass (2-6 yrs. and 6-16 yrs.). Does EPA modeling do this? If not, why not9
Section 3.1.1.3; PP.17s 18: Why Is there no discussion of the second part of Table 3-3, the period EL-1.8
from 4/91 to 2/96? Table 3-3 does not seem to agree with Fig. 3-2. Table 3-3 shows more penta
coming from HUDTOX but Fig. 3-2 shows the opposite. Also, Table 3-3 shows a delta of -18 kg for
hexa but Fig. 3-2 shows a delta of about -52 kg. Please explain these differences. It would be helpful If
EPA would sdck to one set of units as less arithmetic would be required.
Section 3.1.1.4;P.20: The comparison of measured striped bass body burdens to modeled values In EL-1.9
Fig. 3-9 Is for Region 2 only, whereas EPA uses only modeled values In Region 1 In Its health risk
assessment. EPA Is requested to show a plot of the EPA model results vs. data for Region 1 (RM152 SC
RM113) so the proper comparison can be made.
Section 3.1.1.5;P.21: Referring to Fig. 3-10, would It make more sense to plot the average of EL-1.10
FISHRAND values In Region 1 to compare to the Farley Model as It uses averages for Region 1 ?
Section 3.1.I.6;P.21: EPA Is requested to supply a comparison similar to Fig. 3-12 for striped bass. EL-1.11
Why are striped bass often omitted from data comparisons?
Section 3.1.2.2;P.23: Please explain what all the "x's represent on Figs. 3-16 « 3-17. It Is also noted EL-1.12
Fig. 3-17 shows results only for Region 2 despite the tide on the figure.
Section 3.1.2.3;P.24: Comparing Fig. 3-16 to Fig. 3-19, It appears the average value for Region 1 EL-1.13
from Fig. 3-19 Is about 50% higher for the year 2020 than the value from Fig. 3-16, but for Region 2
it appears Fig. 3-16 gives a somewhat higher value. Please explain why this changeover should occur.
Would using the Farley Model throughout give more Internally consistent results and thus be preferred
over FISHRAND? Again, why is there no forecast for striped bass?
Section 3.2, P.25: The selection of a river mile towards the upper end of each range to represent the EL-1.14
range Is another example of the excessive conservatism In the EPA assessments. Given the known drop
-------
off of PCB body burden with decreasing river mile, using the body burden at the selected river miles
instead of an appropriate average over the river mile segment introduces unnecessary extra conservatism.
Section 3.2.4;P.26: The use of brown bullhead results to represent short-nosed sturgeon makes the riskEL-1 15
assessment for the sturgeon very uncertain and of dubious value because of the unknown uncertainty.'
Also the need to extrapolate the fish PCB concentration data from standard fillets basis to whole body
wet weight basis produces more uncertainty of unknown magnitude Into the risk assessment again
decreasing the value of the calculated risks.
Section 3.3;PP.27-30: These sections are very similar to those In the August, 1999 Risk Assessment EL-1.16
"Reports. The comments previously submitted on these Items apply to this report as well and will not be
repeated here.
Section 4; PP.31-36: These sections are very similar to those In the August, 1999 Risk Assessment pT 117
Reports. The comments previously submitted on these items apply to this report as well and will not be KlLi~Lmi'
repeated here. Additional comments come from PP. B-10 at B-l I of Appendix B. The presentation In
Section B.2.3.1 on P. B-10 answers the question asked In the EMC's comments to the August, 1999
Risk Assessment Reports as to the amount of chlorine In chlophen compared to PCB's. However, no
Information Is given to justify that the behavior In fish of the chlorine In chlophen duplicates that of
PCB's. Page B-l 1 says "Hatchabillty was significantly reduced In fish with an average total PCB
concentration of 170 mg/kg....w I thought Bengtsson's testing was done with chlophen A50 and not
PCB's. This sentence should be corrected to state what was actually tested. The discussion here
introduces another factor of about 10 conservatism in the results by not using the 170 mg/kg and
15mg/kg data from Bentgsson study but rather the 15 mg/kg and 1.6 mg/kg data. This further adds to
the total excessive conservatism in the EPA risk assessments (also applies to other fish species In Section
B.2.3 of Appendix B). Does this new conservatism mean that EPA now considers the ecological risk
evaluation of these fish species in the August, 1999 risk assessment to be wrong?
Section 5.;P.37-55: Comments previously made on the August 1999 ERA regarding the over EL-1.18
conservatism In EPA's risk characterization apply to the report as well and will not be repeated here.
Section 5.2.1.9;P.43: As previously questioned, EPA is requested to explain why EPA reports EL-1 19
Measurement Endpolnts for striped bass only for RM152 and 113 and why these river miles should be
considered at all for striped bass.
Section 5.2.4.1 ;P.45«46: In view of the unquantified uncertainty In the calculation of body burdens in FT i -jn
the shortnosed sturgeon and the positive statements about the health of the shortnosed sturgeon in the LlLj~L'*v
last paragraph on this page, why does EPA Insist on putting forth a negative risk evaluation for the
shortnosed sturgeon? This question also applies to white perch as the discussion on P. 46 again indicates
a healthy situation and the discussion at the end of the paragraph represents speculation based on only
extremely conservative calculations and Is inconsistent with the facts shown by the field studies.
Section 5.4.3;P.50,Section 5.5.3.l;PP.53s54,Sectlon 5.3.3.l;PP.47«48A: EPA Is requested to EL-1.21
provide Information on what trends were seen in the Christmas bird counts. This Information would be
helpful In assessing what Is happening to the health of birds In the region.
-------
Section 573.1;P.57: The discussion in this paragraph leads to the conclusion that not enough „
raccoons would be affected by the PCB's in the Hudson to have an Impact on the raccoon popubtlon so ^ *«"
why Is EPA Insisting on singling out the potential risk to those few raccoons that might be affected?
Section A 2-P A-2: It is not clear what Is meant by the phrase "duplicate samples are equivalent.'' EL-1.23
DcS mis nfe'an *e PCB data from the duplicate samples are exactly equal? If not the case, why weren'd
the duplicate GE samples averaged as were the EPA duplicates?
-P.A-3- EPA Is requested to provide some discussion of what factors could effect the EL-1.24
mca procies and why tSese factors are not expected to change to Justify the assumption mad^
he dKon of the steps taken Is confusing In thar, it appears the first step described applies i*
Factor 2 and the second step to Factor 1 . Is this correct?
Section A3-P.A-3 and Flgs.A-1toA-5: The EPA mean values shown on these figures for the TID EL-1.25
?p^mabi; from years prior to 1 996) agree more with GE means (see Fig. A-9) for post 1 996 data
Sd^oTat alUrtui GE means for prior 1996 data. Since the GE data set for the TID Is much larger
225 amples Jrtor to 1996 and 293 samples after 1996) than the EPA data set of 4 to 12 samples, |
d? f«3 & HA data at the TID to calculate the ratio for homologues at Waterford (or £• > Troy
nable. Shouldn't the GE data be used to calculate the factors In Table A-2? EPA Is
this issue regarding the calculation EPA used to get Input to the Farley Model.
Section A 3-P A-4- EPA Is requested to provide the citation of the data used as the basis for the EL-1.26
SSment'm'at there Is Itale evidence of decline In PCB loads at the TID post-1995. Is this still true<
based on 1999 data?
Section A 3;P.A-4: See comment above on A-3 and Fig. A-l - A-5 questioning validity of factors givenf EL-1.27
In Table A-2. Also, why should these factors stay constant for 40 years?
Section A 5 P.A-7: The basis for the statement at the top of the page about releases from Baker Falls Is EL-1.28
undea" Weren't the malor releases from Baker Falls post 1990? If so, EPA Is requested to clarify wh*
the post 1 990 releases are not of concern.
Vol. 2F - Human Health Risk Assessment Comments
Section 2;PP.5-21: Comments previously submitted on Section 2 of the August, 1999 Risk Assessment
apply to this report as well and will not be repeated here.
Section 3-PP 23*24: Comments prevtously submitted on the August 1999 risk assessment regarding
non^cancer toxlclty values and cancer toxlclty apply to this report and will not be repeated here.
Section 2 3 1;P.8: Comments given above on the Ecological Risk Assessment regarding the EPV
approach to calculating PCB concentrations In striped bass apply here also.
Section 2 3 1 -P.9: The comment on Section 3.2, P. 25 of the Ecological Risk Assessment applies ; here
atow me selection of river miles to represent sections of the river as do comments about selecting a
more appropriate way to average values than straight linear averages.
-------
General Eleclric
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Feb-04-00 16:16 From- T-312 P ZB/79 F-180
EG-1
COMMENTS OF GENERAL ELECTRIC COMPANY ON
Hudson River
PCBs Reassessment RI/FS
Phase 2 Baseline Ecological
Risk Assessment for Future Risks
in the Lower Hudson River
February 4,2000
General Electric Company LWB Environmental Services, Inc.
Corporate Environmental Programs 105 Wesley Lane
320 Great Oaks Office Park, Suite 323 Oak Ridge, TN 37830
Albany, NY 12203
Quantitative Environmental Analysis, Inc
305 West Grand Avenue
Montvale.NJ 07645
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F«b-(U-00 IS:IB From- T-312 P 29/79 F-l
TABLE OF CONTENTS
1.0 EXECUTIVE SUMMARY AND INTRODUCTION
2.0 THE FUTURE RISK ERA DOES NOT PROVIDE THE INFORMATION
NECESSARY TO SUPPORT REMEDIAL ACTION DECISIONS ___________________ 6
3.0 EPA HAS REPEATED CRITICAL FLAWS IDENTIFIED IN GE'S AND
OTHERS' REVIEW OF THE BASELINE ERA ______________________________________________ 7
3.1 INADEQUATE CONSIDERATION OF POPULATION vs. INDIVDUAL-LEVEL EFFECTS ..... 7
3.2 IGNORING OR DISMISSING SITE-SPECIFIC DATA [[[ Zs
3.3 USE OF EXCESSIVELY CONSERVATIVE ASSUMPTIONS CONCERNING EXPOSURES ........
AND EFFECTS [[[ g
3 .4 INTERPRETATION OF EXCEEDENCES OF SEDIMENT EFFECTS CONCENTRATIONS' ........
AND OTHER SEDIMENT QUALITY GUIDELINES AS ACTUAL MEASURES OF EFFECTS 1 0
3.5 INAPPROPRIATE USE OF THETEQ APPROACH [[[ '" JQ
3.6 FAILURE TO CITE THE EXPERT REVIEW OF PCB EFFECTS ON PISH PREPARED FOR
NOAA [[[ n
4.0 THE ERA FOR FUTURE RISKS DOES NOT CONFORM TO BEST
SCIENTIFIC PRACTICE
5.0 THE MODELS USED TO PROJECT FUTURE PCB CONCENTRATIONS
IN WATER, SEDIMENT, AND BIOTA HAVE BEEN INADEQUATELY
REVIEWED AND ARE SERIOUSLY DEFICIENT __________________ ......................... 14
5. 1 EPA UPPER HUDSON RIVER MODEL (HUDTOX) USED TO PREDICT PCB LOADS
TO THE LOWER HUDSON RIVER [[[ 14
5.2 FARLEY ETAL. LOWERHUDSON RIVERMODEL USED TO PREDICT LOWER .............
HUDSON RIVER WATER AND SEDIMENT PCB CONCENTRATIONS ............ 15
5.3 MODELS USED TO PREDICT PCB CONCENTRATIONS IN LOWER HUDSON RIVER .....
PISH (FISHRAND AND FARLEY ETAL.) [[[ 16
5.3.1 food web structure ............................. ............. ,7
5.3.2 Calibration .............................................. ZZZZZZZZ .................... 75
5.4 OTHER MODEL DEVELOPMENT ISSUES ................................. ...."...."."..."...".. ............. 19
6.0 AVAILABLE DATA ON ECOLOGICAL RESOURCES OF THE LOWER
HUDSON DIRECTLY CONTRADICT EPA'S CONCLUSIONS _____________________ 20
6.1 BENTHIC MACROINVERTEBRATES .................... ™
6.2 FISH [[[ [[[ ,:
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TABLE OF CONTENTS (Com.)
7.0 EPA'S APPROACH TO EFFECTS ASSESSMENT FOR FISH AND
WILDLIFE IS EXCESSIVELY CONSERVATIVE, RELIES ON A SMALL
SUBSET OF THE AVAILABLE DATA, AND IGNORES OR
IMPROPERLY INTERPRETS KEY STUDIES -------------------------------------- 28
7.1 BENTHIC COMMUNITY STRUCTURE [[[ 28
7.2 FISH [[[ 29
7.3 BIRDS [[[ "!!!!." 30
7.3.1 TreeSwallo* [[[ j;
7.3.2 Mallard [[[ "^
7.3.3 Great Blue Heron [[[ 32
7.3.4 Belted Kingfisher [[[ 33
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LIST OF TABLES AND FIGURES
Table 1. Comparison of Lower Hudson River Future Risk ERA and Clinch River
ERA
Table 1. Compulation of PCB Levels in Fish - Future Risk ERA
Figure 1. Total PCB Concentration and Young-of-the-Year Production for Striped
Bass in the Lower Hudson River
m
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1.0 Executive Summary and Introduction
General Electric Company (GE) submits these comments on the Hudson River PCBs EG-1.1
Reassessment Rf/FS Phase 2 Baseline Ecological Risk Assessment for Future Risks in the
Lower Hudson River (Future Risk ERA), issued by The U.S. Environmental Protection
Agency (EPA) on December 29,1999.
PCBs have been present in the Hudson River environment for SO years, and at
significantly higher levels than are found today. For the last 25 years, PCB concentrations
in fish and wildlife in the Hudson River have been declining. During this period, other
pollutants in this river have generally declined and the management of wild populations,
particularly fish, has materially improved. EPA has studied and analyzed Hudson River
PCBs for the last 10 years and, even before this reassessment began, the Agency was
fully familiar with the river's aquatic resources through its involvement in the issuance of
the first water discharge permits to power plants on the Lower Hudson River in the
1970s.
As a result of public, scientific and regulatory interest in the environmental health of the
Hudson River, volumes of data on fish, wildlife, sediment and water quality have been
collected over the last 25 years. The data documenting conditions in the Lower Hudson
for this period are particularly rich for fish.
When it began its ecological risk assessment for the Lower Hudson, EPA had at its
disposal the entire record of a living river laboratory, a quarter century in length. These
data, collected at a time when PCB levels were higher, provided an unusual opportunity
to explore relationships between PCB levels and the sustainability of populations offish,
birds, and mammals. For a number of animal populations, there was sufficient data for
EPA to examine the potential for impacts due to PCBs and to determine whether at lower
future levels it is reasonable to suggest that animal populations would be affected.
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EPA could have built on this extensive historical record to produce a first-class
FC-1 2
ecological risk assessment. Unfortunately, the Agency did nothing to collect data on EjVJ"
wildlife or biotic populations in the Lower Hudson over the past 10 years and disregarded
the mine of data which it examined in the 1970s power plant cases and which has grown
larger with new data in each year since. EPA likewise ignored the extensive work of the
U.S. Fish and Wildlife Service and the National Marine fisheries Service in addressing
the most obvious, large-scale, Hudson-related biological emergency of the last 25 years -
the late '70s-early '80s crash of the coastal striped bass population, to which the Hudson
stock contributes, an event for which PCBs were considered, but rejected, as a cause,
before the real cause, overfishing, was established (Atlantic States Marine Fisheries
Commission [ASMFC], 1990).
What EPA produced is superficial, theoretical speculation that implies future risks to
wildlife populations without providing evidence of past effects and while ignoring clear
evidence that key wildlife populations are, in general, healthy and the communities
diverse. For many of the fish and wildlife species evaluated by EPA, the facts clearly
contradict EPA's conclusions. For example, the facts demonstrate that:
• The white perch population of the Lower Hudson River is relatively stable and
that the striped bass and shortnose sturgeon populations have increased
dramatically since the 1970s. The upward trend in striped bass is especially
important because EPA has concluded that risks to this species are especially
high.
• Although EPA predicts mat PCB levels in kingfishers range from 4 to 280 times
the level EPA says may pose a risk, a kingfisher population is documented by
EPA as successfully reproducing in the Lower Hudson.
According to reports from various sources, including the New York State Department of
Environmental Conservation (NYSDEC), the U.S. Fish and Wildlife Service (USFWS),
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The Audubon Society and others, the populaiions of other species are present and
growing, including bald eagles, which have returned to the Hudson after an absence of
more than 100 years and, contrary to EPA statements, are successfully reproducing in the
Lower Hudson River; mallard ducks, whose population is characterized as "demonstrably
secure," great blue herons, and raccoons. In some cases, EPA's report does not even
acknowledge these facts, and where it does, it discounts the data for no legitimate reason.
EPA's approach, including selective use of data, discounting information in a manner that
is inconsistent with the Agency's guidance and scientifically defensible practices, and
uncorroborated speculation about risks for which no site-specific evidence exists, is
highly misleading to the public and fails to provide regulators with a risk assessment that
is useful for choosing the most appropriate, scientifically defensible management options
for the Upper Hudson River. There is no sound basis to accept EPA's analytical approach
as plausible when it at dramatic variance with the facts.
The objective of the risk assessment should be to provide data and analysis on which to
base remedial decisionmaking for the Upper Hudson River. To the extent that an
examination of risks in the lower river is appropriate, the assessment must be useful to
the remedial manager as:
• A sound and reliable description of the effects of current and future PCS
exposures emanating solely from the Upper Hudson on biota in the
Hudson River Valley.
• A foundation for projecting the responses of those biota to alternative
remedies taking into account the effects of chemicals other than PCBs and
PCBs whose source is not the Upper Hudson River.
• A sound technical underpinning for comparing the ecological benefits
gained through remediation to the ecological costs of implementing
remedial actions.
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Like EPA's Baseline Ecological Risk Assessment (BERA), the future Risk ERA is
simply a screening-level assessment. As such, it does not reflect acceptable scientific
practice, is excessively conservative, and Is insufficient for use in determining the effect
of a remedy or selecting an appropriate remedy.
The Future Risk ERA repeats critical flaws identified by GE and oners in The BERA
including*
• Inadequate consideration of population vs. individual-level effects.
• Ignoring or dismissing site-specific data.
• Failure to use a weight-of-evidence approach correctly.
• Use of excessively conservative assumptions concerning exposures and effects.
• Interpretation of exceedances of Sediment Effects Concentrations and other
sediment quality guidelines as measures of actual effects.
• Inappropriate use of the TEQ approach.
• Failure to evaluate the usefulness of or even cite the expert review of PCB effects
on fish prepared for NOAA.
• Mathematical errors.
Rather than altering the assessment procedures to minimize or eliminate the identified
flaws, EPA used exactly the same approach in the Future Risk ERA. Consequently, this
assessment suffers from the same flaws as the BERA.
In the following sections, GE provides comments on EPA's Future Risk ERA,
specifically addressing:
» The Future Risk ERA does not provide the information necessary to support
remedial action decisions.
• EPA has repeated critical flaws identified in previous reviews of the BERA.
• The Future Risk ERA does not conform to best scientific practice.
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• The models used 10 project furore PCB concentrations in media have been
inadequately reviewed and are seriously deficient.
• Available data on ecological resources of the Lower Hudson River were not used
and directly contradict EPA's conclusions.
• EPA's approach lo effects assessment for fish and wildlife is excessively
conservative, relies on a small subset of the available data, and ignores or
improperly interprets key studies.
By concluding that PCBs may or may not pose risks to wildlife populations and offering
no evidence of past effects from PCBs, EPA failed to abide by the most fundamental
tenet of its own internal guidance - it did not quantify impacts on wildlife populations.
The Agency failed to use realistic exposure scenarios, failed to consider effects that might
be attributable to contaminants other than PCBs, and failed to distinguish PCBs from the
Upper Hudson and those originating in the mid-Hudson or elsewhere. This final poim is
most important. EPA is preparing to make a remedial decision for the Upper Hudson
River. If it intends to assert that its decision would benefit lower parts of the river as well
as the Upper Hudson, it must be able to show thai it has the ability to distinguish between
one PCB source and another. There is no indication in this report or any report that the
agency has thus far produced for this project, that EPA can do that with any scientific
certainty.
Therefore, this report should be given no weight in the Agency's deliberations over the
appropriate remedial strategy for the Upper Hudson River.
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2.0 The Future Risk ERA does not provide the information necessary to
support remedial action decisions
As we have previously explained, it is inappropriate for EPA 10 base a remedial decision
EG-1.3
for sediments in the Upper Hudson on risk reduction TO biota in the Lower Hudson.
Should EPA nevertheless persist in examining risks in the Lower Hudson, it is clear that,
like the BERA, the Future Risk ERA in its present form will not provide useful
information for the risk manager.
To support remedial action decisions for the Upper Hudson River, the Future Risk ERA
must be based on an objective evaluation of all available information concerning the risks
to ecological resources posed by present and future exposures to PCBs. As described in
the following sections of GE1; comments, this information should include:
• Site-specific data concerning PCB and other chemical exposures and effects on
populations and communities based on a variety of independent lines of evidence.
• Estimates of concentrations of PCBs in sediment, water, and biota based on
properly calibrated and verified models.
• A thorough review of all available data.
The Future Risk ERA fails to include any of the above information. It is based on
inadequately verified models, excessively conservative Toxicity Quotients (TQs) based
on a limited evaluation of literature-derived test data, a focus on individual organisms,
and a failure to consider important and relevant site-specific data. Therefore, the Future
Risk ERA cannot support scientifically sound decisions about remedial actions on the
Hudson River.
1 See. Nov. 6,1997 letter from Angus Macbeth to Richard Caspe; May 5,1998 letter from Angus Macbeth
to Douglas Fischer
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3.0 EPA has repeated critical flaws identified in GE's and others' review of
the Baseline ERA
GE's and other's comments on the BERA identified a number of critical flaws, which EG-1.4
render the document inadequate for supporting remedial decisionmaking for the Hudson
River. In the Future Risk ERA, EPA has not addressed any of these flaws.
3.1 Inadequate consideration of population vs. individual-level effects
As noted in GE's comments on the BERA, decisions concerning remedial action needs
for the Hudson River must consider:
(1) Whether the sustainabiliry of exposed biological populations and
communities is being threatened by the presence of PCBs in Upper Hudson
River sediment.
(2) Whether the positive effects of a particular remedy will be greater than any
negative ecological effects of carrying out the remedy. EPA's Risk
Management Guidance clearly states that populations are the appropriate level
of ecological organization for assessment. (EPA I999a, Ecological Risk
Assessment and Risk Management Principles for Superfund Sites. USEPA
Office of Solid Waste and Emergency Response, Washington, D.C., Directive
9285.7-28P).
A focus on populations rather than individuals is necessary because compensatory
mechanisms that operate in all biological populations permit these populations to sustain
themselves in spite of the death or impairment of some individuals that occurs due to
natural and anthropogenic stressors. Even if statistically significant reductions in
survival, growth and reproduction of some individuals are observed, such data alone
cannot be used directly to estimate adverse effects to populations, communities, or
ecosystems (Forbes and Calow, 1999). Survival, growth, and reproductive rates are
interrelated in complex ways, and apparent adverse changes in one of these factors (e.g.,
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Feb-04-00 16:19 From- T.312 p 39/79 F.,80
a reduction in fecundity) are often offset by compensatory changes in others (e.g.,
increased growth and survival of young).
In the Future Risk ERA, EPA indicates that it considers population-level effects by
comparing the magnitudes of TQs over the 25-year modeling period TO the life spans of
the receptor species (p. 9). EPA asserts that population-level effects are more likely if the
TQ exceeds 1 for the life spaa of a species. This approach does not consider
compensatory processes and is not supported by any published studies. In fact, £PA did
not even implement the approach described on page 9. The risk characterization in
Section 5 does not even discuss the life spans of the various receptor species, much less
compare them to the duration of the modeling period.
3.2 Ignoring or dismissing site-specific data
GE's comments on the BERA noted that EPA had not examined or incorporated site- EG_L5
specific data such as biological surveys, whole-media toxiciry tests, or reproductive
effects studies. According to Suter (1999), site-specific ecotoxicological studies "can
provide a firm basis for decision making, often resulting in savings in remedial costs far
beyond the cost of performing the studies." This is particularly true where, as in the
Lower Hudson, PCB concentrations in biota have been declining over a long period of
time. GE's previous comments included a comparison between the data used by EPA
and the data collected by the Department of Energy for the Clinch River ecological
assessment. Table 1 presents a similar comparison between the Future Risk ERA and the
Clinch River ERA. Whereas the BERA included limited site-specific data concerning the
effects of PCBs on Hudson River biota, the Future Risk ERA includes no data specific to
the Lower Hudson River.
Like the BERA, die Future Risk ERA ignores or discounts existing site-specific data. For
the Lower Hudson, extensive data on the condition of ecological resources are available,
especially for fish. As in the BERA, EPA explicitly discounts these data for risk
assessment, arguing on page 45 that reproduction and recruitment of fish might be
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impaired by exposure to PCfis, even though populations are increasing. The implication
is that only comparisons between measured or modeled exposures and Toxicity
Reference Values (TRVs) are relevant. This conflicts with established principles of
ecological risk assessment (e.g., Suter, 1993) and with EPA's own Superfund guidance
(EPA, 1997a).
3.3 Use of excessively conservative assumptions concerning exposures
and effects
EG-1.6
In its comments on the BERA, GE noted that, even accepting the proposition that the TQ
approach provides useful information for an assessment, EPA's application of TQs in the
BERA provides highly inflated risk estimates that are not useful in remedial
decisionmaking. Both the exposure assessment and the effects assessment used by EPA
employed data, models, and assumptions that are inappropriate for site-specific
assessments.
Like the BERA, the Future Risk ERA employs water and sediment-quality guidelines
designed to be protective such that exposure concentrations belo\v the criteria can be
confidently presumed to be safe. Site-specific studies of the type EPA chose not to
perform (such as those used in the Clinch River ERA) are required to determine whether
exposures that exceed the guidelines are actually causing any adverse effects. Similarly,
in selecting TRVs for use in assessing effects on fish and wildlife, EPA consistently
chose the lowest value from the range of available test results, and often adjusted those
values even lower with lOx uncertainty factors. The resulting TRVs are generally lower
Than any exposure concentrations at which effects have been observed in any test system.
We may be confident that exposures thai are lower than the TRVs will have no adverse
effects, but additional information - again, information that EPA chose not to collect - is
required to determine whether adverse effects will occur at the exposure levels actually
seen in the lower Hudson.
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3.4 Interpretation of exceedences of Sediment Effects Concentrations
and other sediment quality guidelines as actual measures of effects
G£'s comments on the BERA included an extensive discussion of the lack of validity of EG-1 7
NOAA's Sediment Effects Concentrations (SECs) as measures of actual effects on
benthic invertebrate communities. GE provided a thorough review of the inherent
limitations of the SECs and other generic sediment quality guidelines, including
statements from the developers of the guidelines themselves that these values are intended
as screening values, not as measures of effects. In the Future Risk ERA, EPA continues
to use generic sediment-quality criteria as the primary measure of risks to benthic
invertebrates.
3.5 Inappropriate use of the TEQ approach
GE previously noted that the toxicity equivalency (TEQ) approach, in its current state of EG-1.8
development, is a screening approach rather than a primary assessment approach. The
developers of the approach themselves have expressed caution concerning improper use
of the TEQs. EPA has inappropriately handled non-deiect readings of PCS congeners by
using full detection limits for non-detect values, even though standard risk assessment
practice typically involves using one-half of the detection limit for non-deiects and in the
human health risk assessment a value of 0 was used for non-detect. As noted by GE in
comments on me BERA, EPA has assumed that nondetects of B2#126 are present at the
detection limit. This results in the TEQ-based risk assessments being driven by a
chemical not even detected (non-quantified concentrations of B2#126).
In the case offish, ihe review performed for NOAA of the TEQ approach concluded that,
because of insufficient understanding of inter-species variations in sensitivity to dioxin-
like compounds, the approach should not be applied to Hudson River fish species
(NOAA, 1999).
In these circumstances, the Future Risk ERA should not employ the TEQ approach.
10
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3.6 Failure to cite the expert review of PCB effects on fish prepared for
NOAA EG-1.9
In its previous comments, GE noted that NOAA commissioned a review by Dr. Emily
Monosson of effects of PCBs on fish, with specific reference to Hudson River fish
populations (NOAA, 1999). The review concluded that adverse effects on early life
stages of Hudson River fish species might occur at tissue concentrations exceeding 5 ppm
(whole body, wet weight), and that physiological effects on adult fish might occur at
tissue concentrations exceeding 12.5 ppm (whole body, wet weight). One might question
these values in light of the site-specific data, but in any event, they are far higher than the
TRVs used by EPA in both the BERA and the Future Risk ERA.
This review was published by the same NOAA office that published the report on
Sediment Effects Concentrations that £PA used in its assessment of risks to benthic
invertebrates. Both reports were issued in March, 1999. There is no indication that EPA
evaluated the applicability of the Monosson study. EPA's failure to examine the
Monosson review violates common sense and the Agency's own guidelines, which
require the EPA to consider all relevant evidence when performing its risk assessments.
Will EPA choose the results that give the lowest possible acceptable PCB levels
regardless of the quality of the data? This is scientifically indefensible.
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4.0 The ERA for Future Risks does not conform to best scientific practice
EG-1.10
Like the BERA, the Furore Risk ERA relies almost exclusively on Toxichy Quotients"
(TQs), i.e., comparisons between measured or modeled exposure concentrations and
concentrations believed to be potentially harmful to organisms. Such screening-level data
and models, as applied by EPA, are deliberately designed to be conservative, i.e., to
minimine the possibility that any potential adverse effects will be missed. They
necessarily overstate the actual effects of most chemicals at most sites. The Ecological
Risk Assessment Guidance for Superfund (EPA, 1997) explicitly states that decisions to
require remedial action based solely on the screening-level calculations performed by
EPA "would not be technically defensible." As noted by GE in comments on the BERA,
a scientifically defensible ecological risk assessment should use a variety of independent
techniques for measuring and characterizing ecological risks, e.g.:
• Measurements of the abundance, diversity, and other characteristics of
exposed invertebrate, fish, and wildlife communities.
• Measurements of reproductive success in fish, birds, and mammals.
, whole-media, and dietary toxicity tests using selected receptors
or appropriate surrogate species.
These techniques are described in EPA's Guidelines for Ecological Risk Assessment
(EPA, 1998) and Ecological Risk Assessment Guidance for Superfund (EPA, 1997).
Each type of measurement typically requires knowledge of and data relevant to the
population dynamics of the species for appropriate use in assessing risks to wild
populations. Measures of effects on individual organisms must be interpreted in the
context of the distribution, abundance, and temporal dynamics of the exposed
populations.
As noted in GE's comments on the BERA, these techniques have been successfully
applied at other large Superfund sites such as the Clark Fork River (Canfield et al., 1 994)
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and the Clinch River Study Area, Tennessee (Cook et al, 1999). Table 1 contrasts the
assessment performed for the Clinch River Study Area to the EPA's Future Risk ERA.
In addition to the TQ approach used by EPA, the Clinch River assessment used site-
specific toxicity tests, histopathological studies, avian reproduction studies, a mink
dietary toxicity test, and local/regional fish and benthic macroinvenebrate surveys. In
contrast with the deterministic TQs used in the Hudson River assessment, Monte Carlo
analyses and other probabilistic approaches were used in the Clinch River risk assessment
to characterize the likelihood mat adverse effects might occur as a result of exposure to
PCBs and other chemicals.
Data collection to support the Clinch River assessment began in 1989, the same year EPA
initiated its reassessment of PCBs in the Hudson River. EPA had ample time to perform
similar studies for the Hudson River, but chose not to do so.
EPA's approach to evaluating the small amount of field data that were discussed in the
Future Risk ERA also fails to meet accepted standards of scientific inference. In the
Clinch River assessment, all of the lines of evidence were considered together in making
determinations concerning the existence and magnitude of risks. Lack of concordance
between different types of evidence relevant to a given endpoint was taken to indicate
that the risk assessment was inconclusive. In the Future Risk ERA, EPA discounted all
lines of evidence other than TQs, arguing that the failure of field data to support the TQs
simply showed that other factors were masking the adverse effects caused by exposure 10
PCBs. Such an approach is scientifically indefensible.
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5.0 The models used to project future PCB concentrations in water,
sediment, and biota have been inadequately reviewed and are
seriously deficient
EG-1.11
All three of the models used by EPA in the exposure assessment component of the Future
Risk ERA have deficiencies that compromise their value for projecting future PCB
concentrations in sediment, water, and biota. Two of these models - EPA's HUDTOX
and FISHRAND models - were recently revised, and it is the modified models that were
used in the risk assessments. Our comments are based on oral presentations of the
modified models to the peer reviewers of EPA's Baseline Modeling Report (BMR), and
we reserve our right to supplement these comments after further review of the revised
BMR, which EPA just released in late January 2000.
5.1 EPA Upper Hudson River model (HUDTOX) used to predict PCB
loads to the Lower Hudson River
EG-1.12a
The use of the EPA Upper Hudson River model (HUDTOX) to predict PCB load passing
Troy to the Lower Hudson River relies on the presumption that this model accurately
predicts the rime vends of PCB concentrations at Troy. As detailed in GE's Comments
on the BMR (GE, 1999), GE has concerns that HUDTOX has not been properly and fully
developed and is inadequate for predicting future PCB concentrations. One of the most
significant of these concerns relates to the model's ability to describe PCB fate
downstream of the Thompson Island Dam (TTD). The equations and coefficients
describing sediment transport in the 34 miles between The TID and Troy are inconsistent
with the equations and coefficients used in the Thompson Island Pool and inaccurately
represent the processes critical to PCB fate in the river (GE, 1999).
The inaccuracy of the HUDTOX-predicted PCB load to the Lower Hudson River is
exacerbated by the necessity to convert the HUDTOX PCB metric (PCBs with 3 or more
chlorine atoms; tri-i-) ro the homolog characterization of PCBs used in the Farley et al.
(1999) Lower Hudson River model. This conversion was made using factors that may
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not be generally applicable because they were developed from 1993 HD and Waterford
data thai were influenced by the 1991-1993 elevated upstream source.
The ratio of each PCB homolog to tri-t- was calculated in two steps. The first step was to
calculate the seasonal averages of these ratios for all of the measurements made at the GE
TID West sampling station between 1991 and 1998. The second step was to convert
these ratios to equivalent ratios at Waterford. This step was accomplished using the
differences in PCB composition between the TID and Waterford observed in the 1993
EPA Phase 2 sampling program. This assumes that the differences observed in 1993
apply over all times, a presumption that was never tested. There are several reasons why
the presumption may be invalid, first, the 1993 EPA Phase 2 TID station was located
along the west shoreline 200 feet upstream of the GE TID West station. Both stations
provide poor representations of the overall PCB flux passing TID and they are not
replicate locations. Second, the 1993 EPA Phase 2 data reflect a period in which PCB
load from the vicinity of Hudson Falls was a significant component of the PCBs passing
the TID. This condition is not representative of the entire 1991 to 1998 period; a period
over which conditions have transiiioned from one in which the Hudson Falls source
dominates to one in which sediment sources dominate. Thus, a ratio developed from a
snapshot in time may not be applicable to the full historical period or to the future.
5.2 Farley at al. Lower Hudson River model used to predict Lower
Hudson River water and sediment PCB concentrations EG-1.13
EPA has used the Farley et al. (1999) Lower Hudson River model without having
conducted a critical review to determine its validity and accuracy. EPA has not
developed an understanding of the veracity of the predicted water and sediment PCB
concentrations and the relationship of those concentrations to the various PCB sources.
Because the predictions are the basis for the risk calculations, the lack of understanding
of model veracity undermines the utility of the risk assessment.
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Concerns about model veracity are pertinent in view of apparent deviations between the
model and site data. These deviations raise questions about the ability of the model to
accurately describe the relative contributions of external and sediment PCS sources and
to accurately predict time trends.
The model is biased toward lower chlorinated PCfis relative to the observed PCB
composition. For example, data indicate that dicblorobiphenyl constitutes about 20
percent of the sum of di- through pentachlorobiphenyl present at river mile 125, whereas
the model computes thai it constitutes about 40 percent. (See Figure 3-2 of Farley et al.
1999). Dichlorobiphenyl is a reasonable tracer of the Upper Hudson River source and the
upward bias of the model may indicate underestimation of the rate at which the Upper
Hudson River source declines as water moves downstream.
The water column and sediment model-data comparisons were limited to a single year
(1993), an inadequate duration to test the model's ability to predict time trends
accurately. Water column data for comparison to the model were available for only 3
locations over the more than 150 miles of river. The model predicts PCB levels that
compare poorly with these data. The model's predictions are significantly lower than the
summer data and do not predict the extent of concentration decline from Troy to the mid-
river in April (Figure 3-5 of the Future Risk ERA report). These differences suggest that
the model underestimates sources within the lower river (probably local sediments) and
under estimates the loss rate of Upper Hudson River PCBs. The comparison of model
and surface sediment data (Figure 3-7 of the Future Risk ERA report) excludes important
data (i.e., the USEPA Phase 2 high resolution cores) that indicate that the model under
predicts 1993 surface sediment PCB levels.
5.3 Models used to predict PCB concentrations in Lower Hudson River
fish (FISHRAND and Farley et al.)
PCB concentrations in fish in the Lower Hudson River were computed using two models,
FISHRAND (EPA, 1999b) and Farley et al., (1999). Each model was used to predict
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PCB concentrations in selected species in the Lower Hudson River (Table 2). These
models are similar, in that they are mechanistic bioenergetic-based simulation models of
bioaccumulation in aquatic organisms. However, they differ in some of the formulations
used 10 describe the key processes, and the impacts of these differences have not been
evaluated. In addition, as mentioned above, EPA has used the Farley, et al., (1999)
Lower Hudson River model without having conducted a critical review to determine its
validity and accuracy. Thus, the validity of the predicted fish PCB concentrations has not
been fully evaluated, undermining the utility of the risk assessment.
A preliminary review of Farley et al. (1999) and FISHRAND (EPA, 1999b) has revealed
several weaknesses in parameterization and calibration of the models. These are divided
into three categories: food web structure, calibration, and other issues associated with
model development.
5.3.1 Food web structure
fish can accumulate PCBs from both the surface sediments and the water column. PCB
concentrations in the sediments and water column may exhibit different rates of natural
recovery and different responses to remedial activities. Thus, the realism of the projected
fish concentrations is affected by the accuracy of the presumed food web. The two
bioaccumulation models of the Lower Hudson River are inconsistent in their descriptions
of contaminant sources to the food web. FISHRAND includes both sediment- and water
column-associated food webs for the resident fish and the striped bass, based on the fact
that the striped bass concentrations are computed from the largemouth bass
concentrations, and the statement thai the parameterization of FISHRAND is the same as
in the Upper Hudson River. In contrast, Farley includes only a water column source to
the food web of the striped bass. To develop reliable projections, this inconsistency must
be reconciled, and the final food web structure must be considered in light of the
available information.
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Striped bass migration patterns are described inaccurately.
Largemouth bass is a resident fish, while sniped bass is migratory. Because predicted
largemouth bass PCB concentrations are used to estimate sniped bass concentrations, the
contribution to the striped bass of PCBs originating south of Region 1, that is, in the
estuary, is underestimated in the ERA. Projected concentrations in the striped bass are
determined by the changes in the loads from the various PCB sources in the Lower
Hudson River. Migratory striped bass migrate between the coastal ocean, the river and
the Harbor and are therefore exposed to PCBs from many sources. Inaccurate description
of the relative contributions of each source can therefore lead to inaccurate projections.
5.3.2 Calibration
Farlev does not compute realistic temporal trends in striped bass PCB levels.
Computed total PCB concentrations in striped bass ages 6-16 years are consistently lower
than the data prior to 1992 and generally greater than the data after 1992 (Figure 3-9 of EG-1 .15a
the Future Risk ERA report). This is important because it indicates that the rate of
natural recovery is not being accurately modeled. It may be due to inaccuracies in the
food web structure, in particular the contribution of sediment and water column PCBs, or
to inaccurate temporal trends in water column PCBs computed by the fate model.
Resonse of model fish at RM 1 S2 to the events of 1 991 is unrealistic.
At river mile (RM) 152, lipid-based PCB concentrations in largemouth bass, white perch,
brown bullhead and yellow perch increased in 1992 following the Allen Mill event and _~, .
decreased thereafter (Figure 3-12a of the Future Risk ERA report). In contrast, model
calculations for these fish exhibit no response to these events. This suggests that
exposure concentrations and food web structure may be inaccurate.
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FISHRAND computations on a wet weight and lioid basis are inconsistent.
For lareemouth bass and white perch at RM 152, wet weight-based concentrations
FP 1 15c
computed by the model run through the error bars and exhibit limited bias with respect to ^
the data. In contrast, lipid-based levels are generally lower than the data (Figure 3-12a of
the Future Risk ERA report). This suggests that the lipid contents are not representative
of the fish for which PCB data are available.
5.4 Other model development issues
Size of fish modeled mav not reflect consumption patterns bv ecological recepiors.
To develop a relationship between largemouth bass and striped bass concentrations, EPA
compared concentrations in fish greater than 25 centimeters (cm) in length, because those £G-
are consumed by anglers. It is unclear what size classes are used in the model
calculations. Size classes consumed by wildlife should be used.
Fish growth rates are not site-specific.
fish growth rates can control the computed PCB concentrations. For example, if growth
rates are unrealistically high, then the predicted degree of bioaccumulation is likely to be
unrealistically low. To calibrate a model with less bioaccumulation, the exposure
concentrations must be increased. This is done, for example, by increasing the VQ.I iru
contribution to the food web from more contaminated sources. Thus, realistic growth
rates are needed to characterize the contaminant sources to the food web as accurately as
possible. It is our understanding that FISHRAND employed generic growth rates; site-
specific data should be used when available.
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6.0 Available data on ecological resources of the Lower Hudson directly
contradict EPA's conclusions
Substantial data are available concerning the condition of the ecological resources of the
Lower Hudson River. Information concerning long-term trends in the abundance of
various fish species, including three of the receptor species considered in the Future Risk
ERA, are especially complete. This information directly contradicts EPA's conclusions
concerning the risks posed by future exposures to PCBs.
6.1 Benthic macroinvertebrates
Based on the comparison of modeled Lower Hudson River PCB surface water and
sediment concentrations with screening criteria and guidelines, EPA contends that there
is the potential for adverse effects on benthic organisms. As noted in GE's comments on
the BERA, NYSDEC (1993) found that the abundance of pollution-intolerant filter-
feeding macroinvertebrates has increased throughout the Hudson River as a result of
improved water quality since 1972. Hudson River macroinvertebrate communities are
comparable in structure to those in other New York rivers, and currently considered
slightly impacted based on the type of species present in the river (Plafkin et al., 1989;
NYSDEC, 1993).
In addition to improvements at several sites in the Upper Hudson River, NYSDEC (1993)
noted improvements in macroinvertebrate populations in the Lower Hudson River over
the last two decades. The number of pollution-sensitive species increased below Troy
Dam at Castleton and Saugerties between 1973 and 1983. Numbers declined from 1983
to 1991, but 1991 values were still higher than those of the early 1970s. These data
demonstrate that: (1) the benthic community improved even in the presence of PCB
concentrations greater than levels currently exhibited; and (2) changes in species
composition appear to occur independent of changes in PCB levels.
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There is more evidence that the improvements in macroinvenebraie communities of the
Hudson River noted by NYSDEC (1993) are likely independent of any changes in PCB
concentrations. Exponent (1998a,b) found thai the macroinvenebraie communities of the
Upper Hudson River had abundant populations and high species richness (i.e., total
number of taxa), in areas with higher PCB concentrations. These results together with
the results of macroinvertebrate surveys conducted by EPA (as reported in The BERA)
suggest that PCBs currently have no major impact on macroinvenebrate communities of
the Hudson River. Because it is highly unlikely that PCB concentrations in the Lower
Hudson River reach the high concentrations in study area sediments sampled by
Exponent (1998a,b), it can be concluded that there is no apparent risk, present or future,
from GE-associated PCBs to macroinvertebrates of the Lower Hudson River.
6.2 Fish
The Hudson River utility companies recently completed a comprehensive assessment of
the impacts of power plants on the biological resources of the Hudson River (Central
Hudson Gas & Electric Corporation et al., 1999) as part of a Draft Environmental Impact
Statement (DEIS). The assessment summarizes 25 years of data on the distribution and
abundance of the major fish populations inhabiting the Lower Hudson. Trends in the
abundance of 16 fish species were evaluated, including striped bass, white perch, and
shormose sturgeon. The major conclusions from the DEIS are summarized below.
6.2.1 Striped bass
Information on the abundance of striped bass life stages in the Lower Hudson is available
from sampling programs conducted both by the utility companies and by NYSDEC.
These data include a river wide ichthyoplankton sampling program, two beach seine
surveys, a trawl survey, and a mark-recapture program. NYSDEC also samples striped
bass in 7 bays around western Long Island Sound, conducts a haul seine survey to obtain
information on the length, age, sex distribution, and mortality rates for the adult
population, and monitors the striped bass bycatch in the American shad fishery. The data
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derived from these programs represent one of the most extensive data sets available for
any esruarine fish species.
As documented in the DEIS, Urge year classes of striped bass, as measured by the utility
and NYSDEC beach seine surveys, were produced in 1977,1978,1983, and 1984. When
these fish reached reproductive age in the mid and late 1980s, numbers of striped bass
larvae collected in the utilities' river wide ichthyoplankton survey increased dramatically.
Correspondingly strong year classes, as measured in the beach seine surveys, were
produced in four consecutive years, from 1987 through 1990. The abundance of adult
striped bass increased steadily from 1980 through the mid-1990s. According to the DEIS,
the Hudson River striped bass population may now have reached its carrying capacity.
Sniped bass are, according 10 the DEIS, now a dominant predator in the estuary,
controlling the abundance of many other fish species.
In addition to the utility-sponsored studies, research on the migratory behavior of sniped
bass has shown that adult striped bass collected immediately below Troy Dam (RM 152)
appear to be a cohort of nonmigratory male fish that have resided in fresh water for their
entire lifetimes (Secor, 1999). These fish, which frequently have higher PCB body
burdens, are unrepresentative of the population as a whole. Fish that migrate annually
between marine and fresh water, and probably dominate the spawning stock, have much
lower body burdens. The adult females sampled by NYSDEC in April and May, in the
mid and lower estuary, provide the most relevant data concerning PCB concentrations in
spawning female striped bass and are the only data that should be used for risk
assessment.
Figure 1 compares time trends in PCB concentrations in adult female striped bass,
collected during the spawning season in the mid and lower Hudson, to trends in the
NYSDEC striped bass juvenile index. This index, which is a measure of the density of
juvenile striped bass present in the Hudson River estuary during the late summer and
early fall, has been accepted by the Atlantic States Marine Fisheries Commission
(ASMFC) as a valid indicator of year-class production in the Hudson River stnped bass
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T-312 P.54/78 F-180
population and is used in the ASMFC's annual striped bass stock assessments. From
1976 through 1997, the annual production of young striped bass from the Hudson has
fluctuated without trend; PCB concentrations in the spawning females that produced
these fish have declined steadily over the same period. The ASMFC concluded that
"[gjiven the very healthy status of the Hudson River stock, which is well documented to
have relatively high tissue concentrations of PCBs, it would appear that such levels...
may not pose a threat to sniped bass from a population biology perspective" (ASMFC,
1990). Clearly, there is no ewtence that high maternal PCB concentrations in the late
1970s adversely affected striped bass recruitment. The obvious implication of this result
is that future, lower maternal concentrations will similarly have no effect on striped bass
recruitment.
Young-of-ihe-Year Abundance
and I «TB Concentrations in Striped Bass
so
60
a
•3
Q
a
20
0
1970
20
o
15 ~
10
a
£
*
1975 1980
1985
1990 1965
2000
... .YOYwveybyNYSDEC_____Tou]|>CBConcenirnioa
TOUJ Kf Qntcaunttloa: Artngt V- 2SE for female anped t>ou (>}000g, Afril/MqJtUa)
Sou fee: JutdeTS.dtf, NYSDEC **a*ut
Abundance: GeemOrie ml** nutnker ptr 200' iOM ktulfor 6 Mr* sampling penal ->/- 2 S£
Source: Qrqfl Envirorumtrual Imfoa Siatmtm, Dteimtur 1999
Figure 1. Total PCB Concentration and Y oung-of-thc-Year Production for Striped
Bass in the Lower Hudson River
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6.2.2 White perch
White perch are sampled in many of the same programs lhat sample striped bass. The
abundance of white perch larvae and juveniles increased rapidly in the late 1970s, bur has
fluctuated and generally declined since the mid-1980s. A variety of factors may have
contributed to the decline; however, The DEIS concluded that competition with young
striped bass and predanon by older sniped bass are the most likely cause (Central Hudson
Gas & Electric Corporation et al., 1999). In addition, the re-growth of large beds of
water chesmui in The upper estuary following cessation of herbicide treatments in 1976 is
believed to have reduced the quality of the habitat for juvenile fish and may also have
contributed to the recent decline (Central Hudson Gas & Electric Corporation et al.,
1999).
6.2.3 Shortnose sturgeon
Published mark-recapture studies discussed in GE's comments on the BERA show a
large increase in the abundance of shormose sturgeon in the Lower Hudson between the
1970s and the 1990s. These studies indicate thai the size of the spawning stock of
shormose sturgeon in the Hudson has increased fourfold, from approximately 14,000 fish
to 60,000 fish during That interval. These studies are supported by data on the abundance
of yearling shormose from the utilities' monitoring program. The utilities' data show a
substantial increase in abundance of young sturgeon since 1990. In light of these data,
NMFS has recommended that the status of the population be changed from "endangered"
to "threatened."
6.2.4 Atlantic Tomcod
The Atlantic torncod is relevant to the Future Risk ERA because studies performed in the
in the 1970s found liver tumors in 80% of the adult torncod examined (Klauda et al.,
1981). Exposure to PCBs was suggested as a possible cause; however elevated levels of
PAH-sensitive biomarkers in Hudson River tomcod suggest increased exposure to
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polycyclic aromatic hydrocarbons (PAHs), consistent with previous studies (Wirgin et
al., 1994). Thermal stress to tomcod during wanner months and the potential occurrence
of a genetically distinct population of tomcod in the Hudson River thai is predisposed to
neoplasia may also contribute to the prevalence of tumors (El-Zahr, et al., 1993; Schultz
et al.. 1993; Wirgin et al., 1991). Despite the tumors, population trends in this species
have been relatively stable, with abundance increasing somewhat from 1983-1989 and
decreasing somewhat from 1989 through 1997. The DEIS concludes that improved
sewage treatment in the lower estuary, resulting in reduced food availability and
increased competition, may be responsible for the recent decline. Data collected during
the 1995-1996 spawning season indicate that the incidence of liver tumors has dropped to
less than 2%.
6.2.5 Summary of Risks to Fish Community of the Lower Hudson River
Changes in the fish community as a whole, measured by the number of species present,
appear to have been determined by three factors based on analyses performed by experts
in fisheries biology (Central Hudson Gas & Electric Corporation et al., 1999):
(1) Improved water quality in the Lower Hudson, which increased the number of
marine species entering the lower estuary.
(2) Increased abundance of sniped bass, which reduced the abundance of many
species throughout the lower estuary.
(3) Increased abundance of water chestnut, which has reduced the availability of
habitat for freshwater fish in the upper estuary.
PCS exposures, which have declined steadily over the entire period covered in the DEIS,
do not explain any of the observed changes. The observation of increasing, i.e.,
recovering, populations of fish occurring in previous periods of relative high PCB
concentrations suggests that PCBs are unlikely to have a significant impact on population
dynamics in the future when PCB levels are expected to decline.
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6.3 Birds and Mammals
As noted by G£ in comments on the BERA, data demonstrating the health of bird and
mammal populations throughout the Hudson Valley are available from a variety of
sources. For example, data show that mallards are "demonstrably secure" throughout the
New York Bight watershed and are "widespread, abundant and secure in the state of New
York" (USFWS, 1997). NYSDEC (1997) reports that, on the basis of breeding surveys,
the mallard population using the Hudson River estuary is "stable to increasing." Mid-
winter counts of waterfowl show generally increasing numbers of mallards and other
species with a peak in 1995 of more than 16,000 birds (NYSDEC, 1997). North
American Breeding Bird Survey data (analyzed in Sauer et al., 1997) indicate that
populations of mallard ducks have significantly increased at a rate of 5.7 percent per year
within the region that includes the Hudson River (i.e., the Ridge and Valley Province)
since 1966.
The Future Risk ERA itself acknowledges that Audubon Society Christmas bird counts
and other sources of local information on the bird species present in the Lower Hudson
Valley show that:
(1) Tree swallows are present throughout the Lower Hudson Valley.
(2) Waterfowl are extremely abundant.
(3) Belted kingfishers and great blue herons are breeding throughout the Lower
Hudson.
(4) Bald eagles are returning.
EPA's statement that the eagles have not successfully reproduced is incorrect. In fact, the
Hudson River bald eagle population has become reestablished in recent years. The first
bald eagle nesting attempt on the Hudson River in over 100 years occurred in 1992 along
the Lower Hudson River, but no fledglings were successfully produced at this nest until
1997 (Nye 1999, pers. comm.). Since then, three bald eagle territories have been active
on the Lower Hudson River. Four eaglets were fledged from these territories in 1998,
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including three from a single nest in Columbia County. Four eaglets were also fledged in
1999, including three from a single nest in Green County.
The Future Risk ERA also acknowledges that raccoons are abundant throughout the
Lower Hudson Valley, and that mink and river otter are present. EPA discounts the
significance of ihe occurrence of raccoon populations on the grounds that raccoons likely
obtain food from sources other than the Hudson River. In the 1960s, the Hudson River
Valley Commission (HRVC, 1966) reported that the raccoon, cottontail rabbit, gray
squirrel, muskrat, skunk, and beaver were plentiful along the Hudson River. Numerous
localized studies of biota in wetland and riparian areas along the Lower Hudson River
reported the presence of mammalian species that are common throughout the eastern
U.S., including raccoon, muskrat, beaver, and white-tailed deer (Kiviat, 1986, 1997;
Kiviat and Tashiro, 1987; Kiviat and Stapleton, 1987).
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7.0 EPA's approach to effects assessment for fish and wildlife is
excessively conservative, relies on a small subset of the available
data, and ignores or improperly interprets key studies.
All of GE's comments on The TRVs used in the BERA apply equally TO the Future Risk
ERA, because in almost all cases the same TRVs are used in both documents. The only
exception is the study of fiengsston (1980), for which EPA apparently lowered the
NOAEL and LOAEL in response to comments from NOAA on the BERA. In addition to £Q_ 1 ,ga
its previous comments, G£ believes it is important to emphasize that the effects
assessment component of the Future Risk ERA is based on a mere handful of studies that
are treated in an excessively conservative manner. Therefore, not only does EPA make
inappropriate use of an overly conservative screening-level approach, its approach is
further compromised by a biased treatment of the available literature-derived
lexicological data.
7.1 Benthic Community Structure
EPA states that the assessment endpoint to be used for evaluation of risks to the benthic
community is benthic community structure,2 but the measurement endpoints selected
were (1) comparison of modeled water column chemical concentrations to water quality
criteria and (2) comparison of modeled sediment chemical concentrations to guideline
values. Neither of these endpoints that were actually used is directly representative of EG-l.Sb
benthic community structure. These methods are suitable only for screening assessments.
The Furore Risk ERA should rely on direct measurement of the abundance, diversity, and
other characteristics of invertebrate communities. Data on benthic community structure
are available from EPA (1993) (reported as part of the BERA), Exponent (1998a,b), and
NYSDEC (1993).
2 The text of the Future Risk ERA uses the ambiguous phrase "benihic community structure as a food
source"—whether this b intended to mean community structure or biomass is unclear, bui in either case,
the measurement endpoints used are inappropriate.
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Water and sediment quality criteria (or guideline values) are inappropriate measurement
endpoints for assessment of benihic community structure. Criteria values are derived
from toxiciry tests on individuals, and do not represent community-level effects.
EG-U9
7.2 Fish
The following studies provided all of the TRVs for me eight fish species evaluated:
• Bengtsson (1980), effects of exposure to Clophen ASO on the minnow Phoximu
phoxinus.
• Walker et al. (1994), effects of dioxin on lake trout eggs and fry.
• Adams et al. (1989,1990,1992) study of redbreast sunfish (Lepomis aurirus)
exposed to multiple chemicals in the field.
• Olivieri and Cooper (1997), study of effects of dioxin on the fathead minnow
(Pimephales promelas).
• Elonen et al. (1998), study of the effects of dioxin on channel catfish (laalwus
punciana).
• Westin et al. (1993), study of effects of PCBs on larval striped bass (Morone
saxarilis).
The study by Bengtsson (1980) was the source of laboratory-derived TRVs for 7 of the 8
fish species. The TRVs for 6 of these species were derived by applying lOx uncertainty
factors to the NOAEL and LOAEL calculated in the paper. The study by Walker et al.
(1994) was the source of TEQ-based TRVs for 6 of the 8 species. No uncertainty factors
were applied to results from this study; however, because salmonids appear to be
uniquely sensitive to dioxin compared to other tested taxonomic groups, the relevance of
the study to Hudson River fish species is questionable. The NOAELs derived from the
two field studies used by EPA (Adams et al., Westin et al.) are unbounded NOAELs,
meaning that no effects on survival, growth, or reproduction attributable to PCBs were
actually observed.
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The review performed by Monosson (NOAA, 1999) for NOAA, which evaluated all
available literature on the toxiciry of Aroclor 1254 to fish, concluded that adverse effects
could be expected at exposure concentrations of approximately 25 ppra in the livers of
adult fish (equivalent to approximately 12.S ppm in fillets of Hudson River fish) or
approximately 5 ppm (whole body) in larvae. The NOAA value for adult fish is nearly an
order ot magnitude higher than the LOAEL TRVs EPA used for pumpkinseed, brown
bullhead, yellow perch, white perch, largemouth bass, striped bass, and shortnose
sturgeon. As noted in Section 2 of these comments, EPA ignored the report's conclusion
that the TEQ approach should not be applied to Hudson River fish species.
As noted in GE's comments on the BERA, the values developed in the Monosson report
are still conservative: a review by Niimi (1996) concluded That even higher exposures
may be required before actual reductions in survival or reproduction are observed in
typical fish species. Thus, EPA's approach to evaluating the toxiciry of PCBs to fish is
highly selective and superficial and the effects predicted by EPA's TQs have not been
observed in the exposed populations themselves.
EG-1.20
7.3 Birds
For birds, the following laboratory studies on gallinaceous birds (e.g., chickens and
pheasants) provided a large traction of the TRVs used by EPA:
• Scon (1977), effects of PCBs on the chicken.
• Nosek et al. (1992), effects of dioxin on the pheasant.
• Powell ei al. (1996), effects of PCB congeners on the chicken.
EPA acknowledges that gallinaceous birds, such as chickens and pheasants, are
extremely sensitive to PCBs. The use of TRVs derived from these studies is therefore
expected to significantly overstate the actual risks of PCBs to wild birds. Alternative
data sources more relevant to avian receptors at the Hudson River which avoid this
overprediction are discussed in the following sections.
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As GE explained in its comments on the BERA, EPA's use of the lowest available
NOAEL when multiple studies were available is inappropriate. Because a NOAEL can
be considerably lower Than an effects threshold, selection of the highest NOAEL for the
species of interest or a surrogate will minimize the gap between the NOAEL and the
actual threshold for observable effects.
The derivation of TRVs in the Future Risk ERA also follows an outdated "margin-of-
safery" method in applying uncertainly factors which introduces unnecessary
conservatism into the risk assessment. Rather than using default uncertainty factors of
10, human health risk assessors (Douison et al., 1996) use a method that considers values
from 1 to 10 where appropriate, depending on the availability of data for the chemical in
question. Ecological risk assessors seem to be following suit, particularly with regard to
interspecies extrapolations (e.g., EPA Region 10, 1997 [EPA, 1997b]; Hoff and
Henningsen, 1998). EPA's ERA guidelines (EPA, 1998) note thai "uncertainty factors
can be misused, especially when used in an overly conservative fashion, as when chains
of factors are multiplied together without sufficient justification."
in several instances, EPA considers a 10-week exposure period to be subchronic, and a
subchronic-to-chronic uncertainty factor of 10 is applied to the NOAEL. This is the case
for the tree swallow, mallard, great blue heron, bald eagle, and belted kingfisher's dietary
TEQ-based TRV. However, according TO Sample et al. (1996), 10 weeks is considered
the transition point from a subchronic to a chronic exposure duration for avian species,
rendering such a large uncertainty factor unnecessary.
7.3.1 Tree Swallow
The field studies conducted by the U.S. Fish and Wildlife Service which addressed
effects of PCBs at concentrations higher than likely to be found in the Lower Hudson
make it irrelevant to predict PCB-related effects on the basis of extrapolations of data
from laboratory studies. Ample field data have been collected from areas adjacent to the
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Hudson River (Secord and McCarty, 1997; McCany and Secord, 1999 a,b). These data
indicate that the reproductive success of tree swallows is not being affected by PCBs in
the Hudson River. EPA's statements regarding these studies are misleading. McCany
and Secord have been unable to illustrate a dose-response relationship between tree
swallow reproduction and PCB contamination. The differences in reproductive
parameters between the Ithaca and Hudson River tree swallow populations fall within the
natural variation observed elsewhere in tree swallow populations. Likewise, the
behavioral data referred to by EPA do not correlate with reproductive parameters.
7.3.2 Mallard
Our of the three studies that have examined PCB toxicity in mallards, EPA selected the
study with the lowest NOAEL for TRY development. As shown above, this approach is
erroneous. The NOAEL found by Risebrough and Anderson (1975), based on a dietary
Aroclor 1254 dose of 40 ppm, is recommended as the TRY. Risebrough and Anderson
(1975) did not measure PCB concentrations in eggs associated with this level of
exposure. However, Heath et al. (1972) established a NOAEL for Aroclor 1254 at a
slightly lower dose (25 ppm), and measured a corresponding egg concentration of 45
ppm. Additionally, because these TWO studies used exposure durations of 150 and 511
days (Risebrough and Anderson, 1975; Heath et al., 1972, respectively), should not apply
a subchronic-to-chronic uncertainly factor as it did for the Custer and Heinz (1980) study.
7.3.3 Great Blue Heron
The studies selected by EPA for TRY development for the great blue heron were less
appropriate than other available studies and were incorrectly interpreted. Speich et al.
(1992) examined potential effects of environmental concentrations of PCBs, from both
pristine and industrialized areas, on great blue heron reproduction in western Washington
State. The authors noted that they were unable to detect any PCB-related effects on egg
mortality that would have been predicted on the basis of chicken studies. Therefore, the
egg concentration of 16 ppm (wet weight), representing the highest reported mean egg
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concentration in a reproductively healthy colony, could be considered an .unbounded
NOAEL. This concentration is 48-fold higher than The TRY (0.33 mg/kg egg) derived by
EPA on the basis of effects in chickens.
Field data in Sanderson et al. (1994) are used to derive a TEQ-based TRY in great blue
heron eggs. However, the authors reported an improvement in the reproductive success
of the colony with the highest measured TEQ concentrations. Though EPA used an egg
concentration of 0.5 ug TEQ/kg egg as a LOAEL based on a reduction in body weight,
Sanderson et al. (1994) did not find reduced body weights in the birds.
7.3.4 Belted Kingfisher
Species-specific studies are not available for the kingfisher, however, the studies selected
by EPA for TRV development were less appropriate than other available studies for
species similar to the kingfisher. As indicated above, there are available studies for
species with similar feeding habits to those of the kingfisher (e.g., great blue heron)
which would provide more representative TRVs than those derived using gallinaceous
bird studies.
7.3.5 Bald Eagle
The TRV for total PCS concentrations in bald eagle eggs - 3.0 mg/kg - is based on a field
study of population productivity and egg contaminant concentrations for a large number
of sites (Wiemeyer et al.. 1993). This value is inappropriate for two reasons:
(1) Wiemeyer et al. (1993) report that productivity was not statistically different
in eggs in three concentration ranges: <3.0,3.0 - <5.6,5.6-<13 (Wiemeyer ei
al., 1993 Table 10). Productivity was significantly reduced for PCB
concentrations >13 mg/kg. Thus, based upon these data, a NOAEL of 13
mg/kg is more appropriate.
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(2) Wiemeyer et al. (1993) could not demonsrraie impacts of PCBs on
productivity because of the strong correlation between PCB and DDE levels.
Thus, a LOAEL cannot be determined, and the degree of conservatism in the
NOAEL of 13 mg/kg is unknown.
DDE concentrations in fish collected recently near Catskill, New York average
approximately 0.27 ppm whole body (NYSDEC database: HUDORG.dbf). Using an
egg/fish DDE ratio of 22 (Oiesy et al., 1995), an egg level of approximately 6 mg/kg is
estimated. This is greater than the NOAEL of 3.6 mg/kg estimated by Wiemeyer et al.
(1993) for DDE in bald eagles. This suggests that DDE may be having an impact on bald
eagle productivity in the Lower Hudson River.
EPA also ignored or discounted two other field studies on potential effects of PCBs on
bald eagles. Elliot et al. (1996) evaluated hatching success and morphological,
physiological, and histological parameters in bald eagle eggs collected near pulp mills in
British Columbia. Laboratory hatching success did not differ between eggs from pulp
mill sites and from reference locations, though Elliot et al. (1996) did find positive
associations between PCB exposure and biochemical and morphological responses. The
unbounded NOAEL for hatching success based on this data is >400 pg/g TEQ (wet
weight) in eggs. Additionally, Donaldson et al. (1999) studied reproductive success of
breeding bald eagles along Lake Erie in Canada from 1980 to 1996. The author
concluded that the reproductive success of the colony was not impaired, and found an
unbounded NOAEL of >26.4 mg/kg total PCBs (wet weight) in eggs based on nest
reproductive success. Both of these NOAELs are significantly higher than those selected
by EPA.
EG-1.21
7.4 Mammals
As noted in GE's comments on the BERA, the TRVs for little brown bat and raccoon are
based on laboratory studies of rats (Murray et al. 1979; Linder et al. 1974). The study by
Murray et al. (1979) was also used to derive TEQ-based dietary TRVs for ™™lc and river
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oner. EPA calculated TRVs by applying 1 Ox uncertainty factor to the LOAELs and
NOAELs from these studies.
The very limited available data concerning effects of PCBs on mammalian species other
man rodents and mink indicate that EPA should be very cautious about basing remedial
decisions on TQs calculated for these species. Data sources and approaches that EPA
could use to more appropriately assess potential effects of PCBs on mink and river oner
are described below.
7.4.1 Mink
EPA used a field study by Tillett et al. (1996) to derive both a NOAEL and a LOAEL for
TEQs in the diet of mink at Lake Michigan. However, the method used to administer
PCBs to the test animals did not exclude other environmental toxicants known to be
present in Great Lakes fish (Giesy, et al. 1994), the srudy is inappropriate for use in
deriving a LOAEL. On page 34 of the Future Risk ERA, EPA states that "because of the
potential contribution of other contaminants (e.g., metals, pesticides, etc.) to observed
effects in field studies, [this] ERA and ERA Addendum use field studies to establish
NOAEL TRVs, but not LOAEL TRVs." According to EPA's own selection criteria, this
study should not have been used to derive a LOAEL TRV.
Mink laboratory studies thai investigate the reproductive effects of Aroclor 1254
resulting from chronic dietary exposure are typically considered relevant and
scientifically sound for the development of protective mink NOAEL and LOAEL values
for PCBs. EPA's choice of the srudy by Aulerich and Ringer (1977) is consistent with
Sample et al. (1996); however, it should be used similarly to derive a TRV. While EPA
applies a subchronic-to-chronic uncertainly factor of 10 to the NOAEL and LOAEL,
Sample et al. (1996) states that because the treatment period extended before and
throughout the reproductive stage, the study should be considered chronic in duration.
As a result, the NOAEL and LOAEL should not be conservatively adjusted to account for
the exposure duration.
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An alternative approach to TRV development based on dietary levels of PCBs is the
determination of critical body residues of PCBs developed from dose-response
relationships. A study by Leonards et al. (1995) evaluated dose-response relationships
for PCB body burdens and mink reproductive parameters from nine feeding studies.
Leonards et al. (1995) proposed critical body residues of 1.2 ug/g total PCBs (wet
weight) and 160 pg/g TEQ (wet weight) based on effects on mink litter siie. Because
PCB whole-body concentrations in mink were more closely correlated with reproductive
effects than PCB concentrations in food, these critical whole-body residue levels should
serve as PCB TRVs. EPA should use the results of ongoing residue studies for rurbearers
by NYSDEC in conjunction with these TRVs.
7.4.2 River Otter
EPA selected TRVs for the river otter using NOAEL and LOAEL TRVs for mink, based
on the assumption that because the two species are in the same phylogenetic family, they
must be similarly sensitive to PCBs. Recent data examining reproductive health in
mustelids found thai river otters were not as susceptible to PCB-induced effects as mink
(Harding et al., 1999). The Agency should take account of this information.
7.5 General Limitations of TRVs and the TQ Approach EG-1 22
As previously indicated, the TQ approach, which incorporates the TRVs, is a highly
conservative screening-level approach that is inappropriate for use in an ecological risk
assessment of the scale of the Hudson River assessments. Since this approach focuses on
potential risks to individuals, it is not sufficient to demonstrate a significant risk at the
population, community, or ecosystem level. EPA's selective treatment of the available
scientific literature and overly conservative application of uncertainty factors in deriving
TRVs further negates any use this approach has on decisions regarding remedial actions.
36
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CBB-U4-UU ID:*; rrom- n\t r tt/in r-iou
8.0 Conclusions
In its Baseline Ecological Risk Assessment for Future Risks in the Loner Hudson River. EG-1.23
EPA relied exclusively on models and ignored site-specific data demonstrating that PCBs
have not adversely affected ecological resources of the Lower Hudson River in the past,
and will not do so in the future. The models used by EPA to predict future concentrations
of PCBs in water, sediment, and fish tissue contain many deficiencies and have been
inadequately reviewed to date. The Toxiciry Reference Values used by EPA to estimate
risks to fish and wildlife are conservative, screening-level values selectively derived from
the scientific literature. EPA's conclusions, which are that important fish and wildlife
species in the lower Hudson are presently at risk and will in the future continue to be at
risk, are unambiguously contradicted by a wealth of data on the past and present status of
those species. Data that were available to EPA show that:
• The reproductive success of the Hudson River siriped bass population, as measured
by the number of juvenile fish produced each year, was as high in the 1970s, when
PCB concentrations in adult female striped bass were at their highest measured levels,
as in recent years, when concentrations are much lower. The abundance of adult
striped bass has increased dramatically over that same period, as has the abundance of
shormose sturgeon.
• The Lower Hudson River Valley supports healthy, reproducing populations of the
wildlife populations addressed by EPA. These include piscivorous birds such as the
kingfisher, for which EPA predicted that reproductive effects would occur as a result
of PCB exposures.
• Bald eagles are now successfully reproducing in the Lower Hudson River Valley, for
the first time in 100 years.
EPA's failure to properly consider these facts in the Future Risk ERA is inconsistent with
best scientific practice in ecological risk assessment and with the agency's own
guidelines.
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This assessment does noi provide a sound and reliable description of the effects of current
and risks of future PCB exposures on biota in the Hudson River Valley. It does not
provide a scientifically valid foundation for either estimating the responses of the biota of
the Lower Hudson River to alternative remedies that would reduce inputs of PCBs from
the upper Hudson or for comparing the ecological benefits gained through remedial
actions to the ecological costs of implementing remedial actions.
The report should not be used by EPA in making decisions regarding remedial actions in
the upper Hudson River.
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Table 1. Comparison of Lower Hudson River Future Risk ERA and Clinch River
ERA
Hudson River ERA
Clinch River ERA
Problem Formulation
Assessment endpoints:
Maintenance of benthic community structure;
protection and maintenance of local fish,
insectivorous birds, waterfowl, piscivorous
birds, and wildlife; protection of threatened
and endangered species; protection of
significant habitats
Measurement endpoints:
Water and sediment-quality criteria, Chronic
TRVs (reproduction endpoint) for fish, birds,
and mammals
Assessment endpoints:
Reductions in benthic community richness or
abundance; reductions in fish species richness
or abundance; increased frequency of gross
pathologies in fish communities; reduced
abundance or production of piscivorous and
insectivorous wildlife
Measurement endpoints:
Near-field and far-field biological survey data
(fish and benthic invertebrates), whole-
sediment toxicity tests; whole-water toxicity
tests, fish histopathology, water and sediment-
quality criteria; chronic TRVs for fish, birds,
and mammals, blue heron reproductive
success, mink dietary toxicity studies
Exposure Assessment
Modeled concentrations of PCBs (iri+) and
TEQsinfish
Modeled oral doses (tr+ and TEQs) to avian
and mammalian receptors using conservative
exposure assumptions; modeled egg
concentrations in birds
Measured concentrations of Aroclors in fish
(whole body), water, and sediment
Measured concentrations of Aroclors in great
blue heron eggs and chicks
Modeled oral doses to avian and mammalian
receptors (by sub-area), using (1) conservative
exposure assumptions, and (2) Monte Carlo
analysis of all exposure parameters
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T-3IZ P 77/79 F-190
Effects Assessment
Hudson River ERA
Clinch River ERA
TRVs for PCB and TEQ concentrations in fish
tissue
Field-derived (tree swallow and bald eagle) or
literature-derived (other species) TRVs for
fish, birds, mammals
TRY for PCB concentrations in fish tissue
(whole body, adult)
Literature-derived TRVs for birds and
mammals
Site-specific assessment offish hisiopaihology
and reproductive condition
Whole-sediment toxiciry tests
Whole-water toxicity tests
Analysis offish and benthic community
composition at local and regional scales
Site-specific mink dietary toxicity study
Site-specific study of great blue heron
reproductive success
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Feb-04-00 16:30 From-
T-312 P 78/79 F-1BO
Risk Characterization
Hudson River ERA
Clinch River ERA
All assessment Midpoints: Comparison of
water and sediment concentrations to water
and sediment-quality criteria
Fish: Comparison of trif and TEQ
concentrations in fish tissue to literature-
derived TRVs
Overview of population trends for selected
species
Birds: Comparison of modeled oral doses and
egg concentrations (urr and TEQs) to field-
derived (uee swallow and bald eagle) or
literature-derived (ether species) TRVs.
Qualitative overview of occurrence data for
various species
Mammals: Comparison of modeled doses (tri-t-
and TEQs) to literature-derived TRVs
Benthic Invertebrates: Comparison of
maximum sediment concentration to sediment-
quality criteria; comparison of empirical
distribution functions for sediment toxicity to
cumulative distribution of measured sediment
concentrations
Whole-sediment toxicity tests
Fish: Comparison of observed concentration in
fish tissue to TRVs
Whole-water toxicity lest results
Comparison of frequencies of
histopathological and reproductive condition
indicators in study area to observed values in
unexposed upstream reservoir
Canonical discriminant analysis offish
community composition (reservoir scale);
analysis of species richness (reservoir scale
and local scale)
Birds: Comparisons of modeled dose
distributions (cumulative frequencies from
Monte Carlo analysis) to TRVs
Comparison of blue heron reproductive
success in on-site and off-site rookeries;
comparison of osprey reductive success in
nests adjacent to site to observed range of
North American values
Mammals: Comparisons of modeled dose
distributions (cumulative frequencies tram
Monte Carlo analysis) to TRVs
Comparison of toxicity observed in mink
dietary study to toxicity predicted from
exposure model and literature-derived TRVs
47
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T-312 P.79/T9 F-l
Table 2. Computation of PCB Levels in Fish - Future Risk ERA
Largemouth bass,
White perch,
brown bullhead,
pumpkinseed,
yellow perch,
sponail shiner
While perch
113,152
Region 1 (60-152)
FISHRAND
FARLEY
White perch
Region 2 (12-60)
FARLEY
Striped bass
113
FARLEY
Striped bass
152
Largemouih bass from
FISHRAND multiplied
by a data-based
STB/LMB ratio
48
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