oEPA
United States Office of Water EPA-842-R-24-002
Environmental Protection 4304T September 2024
Agency
FINAL
FRESHWATER AQUATIC LIFE AMBIENT WATER QUALITY
CRITERIA AND ACUTE SALTWATER AQUATIC LIFE
BENCHMARK
for
PERFLUOROOCTANOIC ACID
(PFOA)
September 2024
U.S. Environmental Protection Agency Office of Water, Office of Science and
Technology, Health and Ecological Criteria Division
Washington, D.C.
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Acknowledgements
Technical Analysis Leads:
James R. Justice, Office of Water, Office of Science and Technology, Health and Ecological
Criteria Division, Washington, DC
Amanda Jarvis, Office of Water, Office of Science and Technology, Health and Ecological
Criteria Division, Washington, DC
Brian Schnitker, Office of Water, Office of Science and Technology, Health and Ecological
Criteria Division, Washington, DC
Mike Elias, Office of Water, Office of Science and Technology, Health and Ecological Criteria
Division, Washington, DC
Reviewers:
Kathryn Gallagher, Colleen Flaherty and Elizabeth Behl, Office of Water, Office of Science and
Technology, Health and Ecological Criteria Division, Washington, DC
EPA Peer Reviewers (2020):
Jed Costanza, Office of Chemical Safety and Pollution Prevention, Office of Pollution
Prevention and Toxics, Existing Chemical Risk Assessment Division, Washington, DC
Alexis Wade, Office of General Counsel, Water Law Office, Washington, DC
Richard Henry, Office of Land and Emergency Management, Office of Superfund Remediation
and Technology Innovation, Edison, NJ
Kelly O'Neal, Office of Land and Emergency Management, Office of Superfund Remediation
and Technology Innovation, Washington, DC
Gerald Ankley, Lawrence Burkhard, Russ Erickson, Matthew Etterson, Russ Hockett, Dale Hoff,
Sarah Kadlec, Dave Mount, Carlie LaLone, and Dan Villeneuve, Office of Research and
Development, Center for Computational Toxicology and Exposure, Great Lakes Toxicology and
Ecology Division, Duluth, MN
Anthony Williams, Office of Research and Development, Center for Computational Toxicology
and Exposure, Chemical Characterization and Exposure Division, Durham, NC (Research
Triangle Park)
Colleen Elonen, Office of Research and Development, Center for Computational Toxicology and
Exposure, Scientific Computing and Data Curation Division, Duluth, MN
Robert Burgess, Office of Research and Development, Center for Environmental Measurement
and Modeling, Atlantic Coastal Environmental Sciences Division, Narragansett, RI
11
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Sandy Raimondo, Office of Research and Development, Center for Environmental Measurement
and Modeling, Gulf Ecosystem Measurement and Modeling Division, Gulf Breeze, FL
Susan Cormier, Office of Research and Development, Center for Environmental Measurement
and Modeling, Watershed and Ecosystem Characterization Division, Cincinnati, OH
Mace Barron, Office of Research and Development, Center for Environmental Solutions and
Emergency Response, Homeland Security and Materials Management Division, Gulf Breeze, FL
Cindy Roberts, Office of Research and Development, Office of Science Advisor, Policy, and
Engagement, Science Policy Division, Washington, DC
Karen Kesler and Lars Wilcut, Office of Water, Office of Science and Technology, Standards
and Health Protection Division, Washington, DC
Rebecca Christopher and Jan Pickrel, Office of Water, Office of Wastewater Management,
Water Permits Division, Washington, DC
Rosaura Conde and Danielle Grunzke, Office of Water, Office of Wetlands, Oceans, and
Watersheds, Watershed Restoration, Assessment, and Protection Division, Washington, DC
Dan Arsenault, Region 1, Water Division, Boston, MA
Brent Gaylord, Region 2, Water Division, New York, NY
Hunter Pates, Region 3, Water Division, Philadelphia, PA
Renea Hall, Joel Hansel, Lauren Petter, and Kathryn Snyder, Region 4, Water Division, Atlanta,
GA
Aaron Johnson and Sydney Weiss, Region 5, Water Division, Chicago, IL
Russell Nelson, Region 6, Water Division, Dallas, TX
Ann Lavaty, Region 7, Water Division, Lenexa, KS
Tonya Fish and Maggie Pierce, Region 8, Water Division, Denver, CO
Terrence Fleming, Region 9, Water Division, San Francisco, CA
Mark Jankowski, Region 10, Lab Services and Applied Sciences Divisions, Seattle, WA
EPA Peer Reviewers (2023):
Tyler Lloyd, Office of Chemical Safety and Pollution Prevention, Office of Pollution Prevention
and Toxics, Existing Chemicals Risk Management Division, Washington, DC
in
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Thomas Glazer, Office of General Counsel, Water Law Office, Washington, DC
Stiven Foster and Kathleen Raffaele, Office of Land and Emergency Management, Office of
Program Management, Washington, DC
Kelly O'Neal, Office of Land and Emergency Management, Office of Superfund Remediation
and Technology Innovation, Washington, DC
Glynis Hill and Sharon Cooperstein, Office of Policy, Office of Regulatory Policy and
Management, Policy and Regulatory Analysis Division, Washington, DC
Cindy Roberts and Emma Lavoie, Office of Research and Development, Office of Science
Advisor, Policy, and Engagement, Science Policy Division, Washington, DC
Kay Edly and Sydney Weiss, Region 5, Water Division, Chicago, IL
We would like to thank Russ Erickson, Dave Mount, and Russ Hockett, Office of Research and
Development, Center for Computational Toxicology and Exposure, Great Lakes Toxicology and
Ecology Division, Duluth, MN, for their technical support and contribution to this document.
We would like to thank Sandy Raimondo and Crystal Lilavois, Office of Research and
Development, Center for Environmental Measurement and Modeling, Gulf Ecosystem
Measuring and Modeling Division, Gulf Breeze, FL, for their work assisting the Office of Water
in developing the estuarine/marine benchmarks using Interspecies Correlation Estimates (ICE).
iv
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Table of Contents
Acknowledgements ii
Table of Contents v
List of Tables vii
List of Figures ix
List of Appendices xii
Acronyms xiii
Notices xvi
Foreword xvii
Executive Summary xix
1 INTRODUCTION AM) BACKGROUND 1
1.1 Previously Derived PFOA Toxicity Values and Thresholds 2
1.2 Overview of Per- and Polyfluorinated Substances (PFAS) 8
1.2.1 Physical and Chemical Properties of PFOA 12
2 PROBLEM FORMULATION 15
2.1 Overview of PFOA Sources 15
2.1.1 Manufacturing of PFOA 15
2.1.2 Sources of PFOA to Aquatic Environments 17
2.2 Environmental Fate and Transport of PFOA in the Aquatic Environment 18
2.2.1 Environmental Fate of PFOA in the Aquatic Environment 18
2.2.2 Environmental Transport of PFOA in the Aquatic Environment 19
2.3 Transformation and Degradation of PFOA Precursors in the Aquatic Environment... 20
2.3.1 Biodegradation of fluorotelomer-based precursors 21
2.3.2 Biodegradation of side-chain polymers 22
2.3.3 Biodegradation of other polyfluoroalkyl substances 23
2.3.4 Non-microbial biodegradation of other polyfluoroalkyl substances 25
2.4 Environmental Monitoring of PFOA in Abiotic Media 25
2.4.1 PFOA Occurrence and Detection in Ambient Surface Waters 26
2.5 Bioaccumulation and Biomagnification of PFOA in Aquatic Ecosystems 29
2.5.1 PFOA Bioaccumulation in Aquatic Life 29
2.5.2 Factors Influencing Potential for PFOA Bioaccumulation and Biomagnification in
Aquatic Ecosystems 31
2.5.3 Environmental Monitoring of PFOA in Biotic Media 33
2.6 Exposure Pathways of PFOA in Aquatic Environments 37
2.7 Effects of PFOA on Biota 37
2.7.1 Mechanisms of PFOA Toxicity 37
2.7.2 Potential Interactions with Other PFAS 39
2.8 Conceptual Model of PFOA in the Aquatic Environment and Effects 40
2.9 Assessment Endpoints 43
2.10 Measurement Endpoints 44
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2.10.1 Overview of Toxicity Data Requirements 44
2.10.2 Measure of PFOA Exposure Concentrations 45
2.10.3 Measures of Effect 50
2.11 Analysis Plan 53
2.11.1 Derivation of Water Column Criteria 53
2.11.2 Derivation of Tissue-Based Criteria 54
2.11.3 Translation of Chronic Water Column Criterion to Tissue Criteria 55
3 EFFECTS ANALYSIS FOR AQUATIC LIFE 58
3.1 Toxicity to Aquatic Life 58
3,1.1 Summary of PFOA Toxicity Studies Used to Derive the Aquatic Life Criteria 58
3.2 Derivation of the PFOA Aquatic Life Criteria 90
3.2.1 Derivation of Water Column-based Criteria 90
3.2.2 Derivation of Tissue-Based Criteria 97
3.3 Summary of Freshwater PFOA Aquatic Life Criteria and the Acute Estuarine/Marine
Benchmark 104
4 EFFECTS CHARACTERIZATION FOR AQUATIC LIFE 107
4.1 Influence of Using Non-North American Resident Species on PFOA Criteria 107
4.1.1 Freshwater Acute Water Criterion with Resident Organisms 107
4.1.2 Freshwater Chronic Water Criterion with North American Resident Organisms ..110
4.2 Consideration of Relatively Sensitive Qualitatively Acceptable Water Column-Based
Toxicity Data 112
4.2.1 Consideration of Qualitatively Acceptable Acute Data 113
4.2.2 Consideration of Qualitatively Acceptable Chronic Data 119
4.3 Acute to Chronic Ratios 129
4.4 Tissue-based Toxicity Studies Compared to the Chronic Tissue-based Criteria 131
4.5 Effects on Aquatic Plants 133
4.6 Protection of Threatened and Endangered Species 134
4.6.1 Quantitatively Acceptable Acute Toxicity Data for Listed Species 134
4.6.2 Quantitatively Acceptable Chronic Toxicity Data for Listed Species 135
4.6.3 Qualitatively Acceptable Toxicity Data for Listed Species 135
4.7 Summary of the PFOA Aquatic Life Criteria and the Supporting Information 135
5 REFERENCES 137
vi
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List of Tables
Table Ex-1. Recommended Perfluorooctanoic acid (PFOA) Ambient Water Quality Criteria
for the Protection of Aquatic Life in Freshwater xxi
Table Ex-2. Acute Perfluorooctanoic Acid (PFOA) Benchmark for the Protection of Aquatic
Life in Estuarine/Marine Waters xxi
Table 1-1. Previously Derived PFOA Toxicity Values and Thresholds 4
Table 1-2. Two Primary Categories of PFAS1 10
Table 1-3. Classification and Chemical Structure of Perfluoroalkyl Acids (PFAAs).1 11
Table 1-4. Chemical and Physical Properties of PFOA 13
Table 2-1. Summary of Assessment Endpoints and Measures of Effect Used in the Criteria
Derivation for PFOA 52
Table 2-2. Evaluation Criteria for Screening Bioaccumulation Factors (BAFs) in the Public
Literature 57
Table 3-1. Summary Table of Minimum Data Requirements per the 1985 Guidelines
Reflecting the Number of Acute and Chronic Genus and Species Level Mean Values
in the Freshwater and Saltwater Toxicity Datasets for PFOA 60
Table 3-2. The Four Most Sensitive Genera Used in Calculating the Acute Freshwater
Criterion (Sensitivity Rank 1-4) 61
Table 3-3. Ranked Freshwater Genus Mean Acute Values 69
Table 3-4. Estuarine/Marine Acute PFOA Genera 72
Table 3-5. Ranked Estuarine/Marine Genus Mean Acute Values 76
Table 3-6. The Most Sensitive Genera Used in Calculating the Chronic Freshwater Water
Column Criterion (Sensitivity Rank 1-4) 77
Table 3-7. Ranked Freshwater Genus Mean Chronic Values 87
Table 3-8. Ranked Estuarine/Marine Genus Mean Chronic Values 90
Table 3-9. Freshwater Final Acute Value and Criterion Maximum Concentration 91
Table 3-10. Freshwater Final Chronic Value and Criterion Continuous Concentration 93
Table 3-11. Summary Statistics for PFOA BAFs in Invertebrate Tissues and Various Fish
Tissues1 98
Table 3-12. Recommended Perfluorooctanoic acid (PFOA) Ambient Water Quality Criteria
for the Protection of Aquatic Life in Freshwater 105
Table 4-1. Ranked Freshwater Genus Mean Acute Values with North American Resident
Organisms 108
Table 4-2. Freshwater Exploratory Final Acute Value and Acute Water Column Concentration
with North American Resident Organisms (zebrafish included) 110
Table 4-3. Ranked Freshwater Genus Mean Chronic Values with Resident Organisms Ill
Table 4-4. Freshwater Exploratory Final Chronic Value and Chronic Water Column
Concentration with North American Resident Organisms 112
vii
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Table C-l. EC50 to EC10 ratios from all quantitatively acceptable chronic concentration-response
curves with species similar to H. azteca (i.e., small members of the subphylum
Crustacea) and with endpoints that were based on reproduction per female C-6
Table L-l. Surrogate Species Measured Values for PFOA and Corresponding Number of
ICE Models for Each Surrogate L-8
Table L-2. Comparison of ICE-predicted and measured values of PFOA for species using
both scaled values (entered as mg/L) and values potentially beyond the model
domain (entered as (J,g/L) L-10
Table L-3. All ICE Models Available in Web-ICE v3.3 for Saltwater Predicted Species
Based on Surrogates with Measured PFOA L-16
Table L-4. ICE-estimated Species Sensitivity to PFOA L-18
Table L-5. Ranked Estuarine/Marine Genus Mean Acute Values L-21
Table L-6. Estuarine/Marine Final Acute Value and Protective Aquatic Acute Benchmark. ...L-22
Table M-l. Measured Perfluorooctanoic acid (PFOA) Concentrations in Surface Waters
Across the United States M-l
Table N-l. Characteristics of adult fish sampled for the calculation of PFOA reproductive
tissue BAFs N-2
Table N-2. Summary Statistics for PFOA Freshwater BAFs in Additional Fish Tissues1 N-3
Table N-3. PFOA Concentrations for Additional Fish Tissue Values.1'2 N-4
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List of Figures
Figure 1-1. Chemical Structure of the Linear Isomer of Perfluorooctanoic acid (PFOA) 12
Figure 2-1. Synthesis of Perfluorooctanoic acid (PFOA) by Electrochemical Fluorination
(ECF) 16
Figure 2-2. Map Indicating Sampling Locations for Perfluorooctanoic acid (PFOA) Measured
in Surface Waters Across the United States (U.S.) Based on Data Reported in the
Publicly Available Literature 27
Figure 2-3. Distribution of the minimum and maximum concentrations (ng/L) of
Perfluorooctanoic acid (PFOA) measured in surface waters for each state or
waterbody (excluding the Great Lakes) with reported data in the publicly available
literature 28
Figure 2-4. Conceptual Model Diagram of Sources, Compartmental Partitioning, and Trophic
Transfer Pathways of Perfluorooctanoic acid (PFOA) in the Aquatic Environment
and its Bioaccumulation and Effects in Aquatic Life and Aquatic-dependent
Wildlife 42
Figure 3-1. Ranked Freshwater Acute PFOA GMAVs Fulfilling the Acute Family MDR 71
Figure 3-2. Acceptable Estuarine/Marine GMAVs 76
Figure 3-3. Freshwater Genus Mean Chronic Values for PFOA 88
Figure 3-4. Estuarine/Marine Genus Mean Chronic Values for PFOA 90
Figure 3-5. Ranked Freshwater Acute PFOA GMAVs and CMC 92
Figure 3-6. Freshwater Quantitative GMCVs and CCC 94
Figure L-l. Example ICE Model for Rainbow Trout (surrogate) and Atlantic Salmon
(predicted) L-4
Figure L-2. Ranked Estuarine/Marine Acute PFOA GMAVs Used for the Aquatic Life
Acute Benchmark Calculation L-22
Figure L-3. Americamysis bahia (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-27
Figure L-4. Americamysis bahia (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values L-27
Figure L-5. Americamysis bahia (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-28
Figure L-6. Americamysis bahia (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values L-28
Figure L-7. Danio rerio - embryo (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-29
Figure L-8. Danio rerio - embryo (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values L-29
Figure L-9. Danio rerio - embryo (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-30
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Figure L-10. Danio rerio - embryo (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values L-30
Figure L-l 1. Daphnia magna embryo (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values L-31
Figure L-12. Daphnia magna (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values L-31
Figure L-13. Daphnia magna (X-axis) and Lepomis macrochirus (Y-axis) regression model
used for ICE predicted values L-32
Figure L-14. Daphnia magna (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-32
Figure L-15. Daphnia magna (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-33
Figure L-16. Daphnia magna embryo (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values L-33
Figure L-17. Lampsilis siliquoidea (X-axis) and Daphnia magna (Y-axis) regression
model used for ICE predicted values L-34
Figure L-18. Lampsilis siliquoidea (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values L-34
Figure L-19. Lampsilis siliquoidea (X-axis) and Ligumia recta (Y-axis) regression model
used for ICE predicted values L-3 5
Figure L-20. Lampsilis siliquoidea (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-3 5
Figure L-21. Lampsilis siliquoidea (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values L-36
Figure L-22. Lepomis macrochirus (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values L-36
Figure L-23. Lepomis macrochirus (X-axis) and Daphnia magna (Y-axis) regression
model used for ICE predicted values L-37
Figure L-24. Lepomis macrochirus (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values L-37
Figure L-25. Lepomis macrochirus (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-3 8
Figure L-26. Lepomis macrochirus (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-3 8
Figure L-27. Lepomis macrochirus embryo (X-axis) and Pimephales promelas (Y-axis)
regression model used for ICE predicted values L-39
Figure L-28. Ligumia recta (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values L-39
Figure L-29. Lithobates catesbeianus (X-axis) and Daphnia magna (Y-axis) regression
model used for ICE predicted values L-40
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Figure L-30. Lithobates catesbeianus (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values L-40
Figure L-31. Lithobates catesbeianus (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-41
Figure L-32. Lithobates catesbeianus (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values L-41
Figure L-33. Oncorhynchus mykiss (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values L-42
Figure L-34. Oncorhynchus mykiss (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-42
Figure L-35. Oncorhynchus mykiss (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values L-43
Figure L-36. Oncorhynchus mykiss (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values L-43
Figure L-37. Oncorhynchus mykiss (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-44
Figure L-38. Oncorhynchus mykiss (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values L-44
Figure L-39. Pimephales promelas (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values L-45
Figure L-40. Pimephales promelas (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-45
Figure L-41. Pimephales promelas (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values L-46
Figure L-42. Pimephales promelas (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values L-46
Figure L-43. Pimephales promelas (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-47
Figure L-44. Pimephales promelas (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-47
Figure L-45. Pimephales promelas (X-axis) and Xenopus laevis (Y-axis) regression model
used for ICE predicted values L-48
Figure L-46. Xenopus laevis (X-axis) and Pimephales promelas (Y-axis) regression model
used for ICE predicted values L-48
Figure M-l. Distribution of the minimum and maximum concentrations (ng/L) of
Perfluorooctanoic acid (PFOA) measured in surface water samples collected from
the Great Lakes as reported in the publicly available literature M-8
Figure M-2. Comparison of relatively high (A; greater than 30 ng/L) and low (B; less than 30
ng/L) maximum Perfluorooctanoic acid (PFOA) concentrations (ng/L) measured in
XI
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surface water samples collected across the United States (U.S.) as reported in the
publicly available literature M-14
List of Appendices
Appendix A Acceptable Freshwater Acute PFOA Toxicity Studies A-l
Appendix B Acceptable Estuarine/Marine Acute PFOA Toxicity Studies B-l
Appendix C Acceptable Freshwater Chronic PFOA Toxicity Studies C-l
Appendix D Acceptable Estuarine/Marine Chronic PFOA Toxicity Studies D-l
Appendix E Acceptable Freshwater Plant PFOA Toxicity Studies E-l
Appendix F Acceptable Estuarine/Marine Plant PFOA Toxicity Studies F-l
Appendix G Other Freshwater PFOA Toxicity Studies G-l
Appendix H Other Estuarine/Marine PFOA Toxicity Studies H-l
Appendix I Acute-to-Chronic Ratios I-1
Appendix J Unused PFOA Toxicity Studies J-l
Appendix K EPA Methodology for Fitting Concentration-Response Data and Calculating
Effect Concentrations K-l
Appendix L Derivation of Acute Protective PFOA Benchmarks for Estuarine/Marine
Waters through a New Approach Method (NAM) L-l
Appendix M Occurrence of PFOA in Abiotic Media M-l
Appendix N Translation of The Chronic Water Column Criterion into Other Fish Tissue
Types N-l
Appendix O Bioaccumulation Factors (BAFs) Used to Calculate PFOA Tissue Values 0-1
Appendix P Example Data Evaluation Records (DERs) P-l
xii
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Acronyms
8:2 FTAC
fluorotelomer acrylate
8:2 FTMAC
fluorotelomer methacrylate
8:2 FTOH
fluorotelomer alcohol
8:2 FTS
fluorotelomer stearate
8:2 HMU
aliphatic diurethane ester
ACR
Acute-to-Chronic Ratio
ADP
Action Development Process
AFFF
aqueous film-forming foam
AIC
Akaike information criteria
ASTM
American Society for Testing and Materials
AWQC
National Recommended Ambient Water Quality Criteria
BAF
bioaccumulation factor
BMF
biomagnification factors
C8/C8-PFPIA
Bis(perfluorooctyl) phosphinic acid
C8-PFPA
Perfluorooctyl phosphonic acid
CAS
Chemical Abstracts Service
CASRN
Chemical Abstracts Service Registry Number
C-F
carbon fluorine
C-R
concentration-response
CC
Chronic Criterion
CCC
Criterion Continuous Concentration
CMC
Criterion Maximum Concentration
CWA
Clean Water Act
DER
Data Evaluation Record
dpf
days post fertilization
diPAPs
polyfluoroalkyl phosphoric acid diesters
drc
dose-response curve
dw
dry weight
ECF
Electrochemical fluorination
ECOTOX
ECOTOXicology database
ECx
Effect concentration at x percent level
EPA
U.S. Environmental Protection Agency
EU
European Union
FACR
Final Acute-to-Chronic Ratio
FAV
Final Acute Value
FCV
Final Chronic Value
FTEOs
fluorotelomer ethoxylates
FTCAs
fluorotelomer carboxylates
FTOH
fluorotelomer alcohol
GLI
Great Lakes Initiative
GMAV
genus mean acute value
GMCV
genus mean chronic value
GSD
genus sensitivity distribution
GSI
gonadal somatic index
hpf
hours post fertilization
Xlll
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HSI
ICE
ICx
Kd
Koc
Kow
LCx
LOECs
LOQ
MATC
MC
MDRs
NAMs
NOECs
NPDES
OECD
OCSPP
ORD
OW
ppt
PFAAs
PFAS
PFCAs
PFdiCAs
PFdiSAs
PFECAs
PFESAs
PFO
PFOA
PFOAAms
PFOS
PFOSI
PFPAs
PFPIAs
PFSAs
PFSIAs
pKa
QACs
SMACR
SMAV
SMCV
SOP
hepatic somatic index
Interspecies Correlation Estimation
Inhibitory concentration at x percent level
partitioning coefficients
Organic carbon water partitioning coefficient
n-octanol-water partition coefficient
Lethal concentration at x percent level
Lowest Observed Effect Concentrations
limit of quantification
Maximum Acceptable Toxicant Concentration
Maximum Criterion
minimum data requirements
New Approach Methods
No Observed Effect Concentrations
National Pollutant Discharge Elimination System
Organization for Economic Cooperation and Development
Office of Chemical Safety and Pollution Prevention
Office of Research and Development
Office of Water
parts per thousand
perfluoroalkyl acids
per- and polyfluorinated substances
Perfluorocarboxylic acids, perfluoroalkyl carboxylates or perfluoroalkyl
carboxylic acids
Perfluoroalkyl dicarboxylic acids
Perfluoroalkane disulfonic acids
Perfluoroalkylether carboxylic acids
Perfluoroalkylether sulfonic acids
Perfluorooctanoate
Perfluorooctanoic acid, pentadecafluoro-l-octanoic acid,
pentadecafluoro-n-octanoic acid, octanoic acid, pentadecafluoro-,
perfluorocaprylic acid, pentadecafluorooctanoic acid, perfluoroalkyl
carboxylic acid or perfluoroheptanecarboxylic acid
perfluorooctaneamido quaternary ammonium salt
Perfluorooctane sulfonic acid or perfluorooctane sulfonate
Perfluorooctane sulfinic acid
Perfluoroalkyl phosphonic acids
Perfluoroalkyl phosphinic acids
Perfluoroalkane sulfonic acids or perfluorokane sulfonates
Perfluoroalkyl sulfinic acids
acid dissociation constant
quaternary ammonium polyfluoroalkyl surfactants
species mean acute-to-chronic ratio
species mean acute value
species mean chronic value
standard operating procedure
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SSD
species sensitivity distribution
TMDLs
Total Maximum Daily Loads
TSCA
Toxic Substances Control Act
U.S.
United States
web-ICE
Web-based Interspecies Correlation Estimation
WQS
water quality standards
WW
wet weight
WWTPs
wastewater treatment plants
XV
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Notices
This document provides information that states and authorized Tribes may consider when
establishing water quality standards under the Clean Water Act (CWA) to protect aquatic life
from effects of Perfluorooctanoic acid (PFOA). Under the CWA, states and authorized Tribes
establish water quality criteria to protect designated uses. State and Tribal decision makers retain
the discretion to adopt approaches that are scientifically defensible that differ from these
recommended criteria or benchmarks, including to reflect site-specific conditions. While this
document contains the Environmental Protection Agency's (EPA) scientific recommendations
regarding ambient concentrations of PFOA that protect aquatic life, the PFOA Criteria
Document does not substitute for the Clean Water Act or the EPA's regulations; nor is this
document or the values it contains a regulation itself. This document does not establish or affect
legal rights or obligations, or impose legally binding requirements on the EPA, states, Tribes, or
the regulated community. It cannot be finally determinative of the issues addressed. This
document has been approved for publication by the Office of Science and Technology, Office of
Water, U.S. Environmental Protection Agency.
Mention of trade names or commercial products does not constitute endorsement or
recommendation for use. This document can be downloaded from:
https://www.epa.gov/wqc/aqiiatic4ife-criteria-perfluorooctarioic-acid-pfoa.
xvi
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Foreword
The Clean Water Act (CWA) Section 304(a)(1) (P.L. 95-217) directs the Administrator of
the EPA to develop and publish water quality criteria recommendations that accurately reflect
the latest scientific knowledge on the kind and extent of all identifiable effects on health and
welfare that might be expected from the presence of pollutants in any body of water, including
groundwater. This document includes EPA's recommended ambient water quality criteria
(AWQC) for the protection of aquatic life based upon consideration of all available information
relating to effects of perfluorooctanoic acid on aquatic organisms in freshwaters, as well as an
informational acute saltwater benchmark developed under CWA Section 304(a)(2).
Aquatic life benchmarks, developed by the EPA under 304(a)(2) of the CWA, are
informational values that EPA generates when there are limited high quality toxicity data
available and data gaps exist for several aquatic organism families. EPA develops aquatic life
benchmarks to provide information that states and Tribes may consider in their water quality
protection programs, including when developing water quality standards. In developing aquatic
life benchmarks, data gaps may be filled using new approach methods (NAMs), such as
computer-based toxicity estimation tools (e.g., EPA's Web-ICE) or other new approach methods
intended to reduce reliance on additional animal testing (https://www.epa.gov/chemical-
research/epa-new-approach-methods-work-plan-reducirm-use-verteb rate-animals-chemical).
including the use of read-across estimates based on other chemicals with similar structures. Like
criteria recommendations developed under Section 304(a)(1), the EPA's aquatic life benchmark
values are not regulatory, nor do they automatically become part of a state's water quality
standards.
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Under CWA Section 303, states or authorized Tribes adopt water quality standards and
submit them to EPA for review and approval. If approved by EPA as water quality standards,
they become the CWA water quality standards applicable in ambient waters within that state or
authorized Tribe. A state or authorized Tribe may, where appropriate, adopt water quality criteria
that have the same numerical values as recommended criteria or benchmarks developed by EPA
under CWA Section 304. States and authorized Tribes have discretion to adopt criteria that
modify EPA's recommended criteria to reflect site-specific conditions, such as the local water
chemistry or ecological conditions, or to develop criteria based on other scientifically defensible
methods that are protective of designated uses (40 C.F.R. 131.11 [b]). Guidelines to assist the
states and authorized Tribes in modifying the criteria presented in this document are contained in
the Water Quality Standards Handbook (see Chapter 3 titled "Water Quality Criteria"; U.S. EPA
2023).
Deborah G. Nagle
Director
Office of Science and Technology
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Executive Summary
The U.S. Environmental Protection Agency (EPA) developed the recommended
perfluorooctanoic acid (PFOA) freshwater aquatic life ambient water quality criteria and an acute
saltwater benchmark in accordance with the provisions of Section 304(a) of the Clean Water Act.
This document provides the EPA's basis for and derivation of the national PFOA ambient water
quality criteria recommendations to protect aquatic life. The EPA has derived the recommended
PFOA aquatic life criteria and benchmark to be consistent with methods described in the EPA's
"Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of
Aquatic Organisms and Their Uses'' (U.S. EPA 1985).
PFOA is an organic, human-made perfluorinated compound, consisting of a seven-carbon
backbone and a carboxylate functional group. PFOA (and other related chemicals in the
perfluorocarboxylic acids, PFCAs) is used primarily in specialized applications associated with
surface coatings in a variety of industrial and commercial products. This document provides a
critical review of toxicity data for aquatic life identified in the EPA's literature search for PFOA,
including the anionic form (CAS No. 45285-51-6), the acid form (CAS No. 335-67-1), and the
ammonium salt (CAS No. 3825-26-1). It also quantifies the toxicity of PFOA to aquatic life and
provides final criteria recommendations to protect aquatic life in freshwater from the acute and
chronic toxic effects of PFOA.
The Aquatic Life Ambient Water Quality Criteria for PFOA document includes water
column-based acute and water column-based chronic criteria, as well as chronic tissue-based
criteria for freshwaters. Quantitatively-acceptable estuarine/marine toxicity data only fulfilled
three of the eight minimum data requirements (MDRs) for deriving an acute estuarine/marine
criterion, and one of the eight MDRs for deriving a chronic estuarine/marine criterion per the
1985 Guidelines. The EPA did, however, include an acute aquatic life benchmark for
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estuarine/marine environments in Appendix L, using available estuarine/marine species toxicity
data and the New Approach Methods (NAMs) application of the EPA Office of Research and
Development's (ORD) peer-reviewed web-based Interspecies Correlation Estimate tool (Web-
ICE; Version 3.3; https://www.epa.gov/webice/). The estuarine/marine benchmarks are CWA
Section 304(a)(2) information provided for states and authorized Tribes to consider in their
state/tribal water quality protection programs. However, the acute estuarine/marine benchmark
magnitude is less certain than the freshwater criteria as the benchmark was based on both direct
laboratory-based and estimated PFOA acute toxicity data (Appendix L).
The final freshwater acute water column-based criterion magnitude is 3.1 mg/L, and the
final chronic water column-based chronic criterion magnitude is 0.10 mg/L. The final chronic
freshwater criterion also contains tissue-based criteria with magnitudes of 6.49 mg/kg wet weight
(ww) for fish whole-body, 0.133 mg/kg ww for fish muscle tissue, and 1.18 mg/kg ww for
invertebrate whole-body tissue. All criteria are intended to be equally protective against adverse
PFOA effects and are intended to be independently applicable. The three tissue criteria
magnitudes (for fish and invertebrate tissues) are translations of the chronic water column
criterion for freshwater using bioaccumulation factors (B AFs) derived from a robust national
dataset of BAFs (Burkhard 2021). The assessment of the available data for fish, invertebrates,
amphibians, and plants indicates these criteria recommendations are expected to protect the
freshwater aquatic community.
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Table Ex-1. Recommended Perfluorooctanoic acid (PFOA) Ambient Water Quality
Criteria for the Protection of Aquatic Life in Fres
lwater.
Chronic
Chronic
Acute \\aid-
Invertebrate
lisli
Co! ii m n
Chronic \\aler
Wliole-
Wliole-
Chronic Fish
Type/Media
(CMC)1-4
Column (C C C )1-5
liodv12
liodv12
Muscle12
Magnitude
3.1 mg/L
0.10 mg/L
1.18 mg/kg
WW
6.49 mg/kg
WW
0.133 mg/kg
WW
Duration
One-hour average
Four-day average
Instantaneous3
Frequency
Not to be
Not to be exceeded
exceeded more
more than once in
Not to be exceeded6
than once in three
three years on
years on average
average
1 All five of these water column and tissue criteria are intended to be independently applicable and no one criterion takes
primacy. All of the above recommended criteria (acute and chronic water column and tissue criteria) are intended to be
protective of aquatic life. These criteria are applicable throughout the year.
2 Tissue criteria derived from the chronic water column concentration (CCC) with the use of bioaccumulation factors and are
expressed as wet weight (ww) concentrations.
3 Tissue data provide instantaneous point measurements that reflect integrative accumulation of PFOA over time and space in
aquatic life population(s) at a given site.
4 Criterion Maximum Concentration; applicable throughout the water column.
5 Criterion Continuous Concentration; applicable throughout the water column.
6 PFOA chronic freshwater tissue-based criteria should not be exceeded, based on measured tissue concentrations representing the
central tendency of samples collected at a given site and time.
Table Ex-2. Acute Perfluorooctanoic Acid (PFOA) Benchmark for the Protection of
Aquatic Life in Estuarine/Marine Waters.
Type/Media
Acute W aler Column Benchmark
Magnitude
7.0 mg/L
Duration
One hour average
Frequency
Not to be exceeded more than once in three years on average
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1 INTRODUCTION AND BACKGROUND
National Recommended Ambient Water Quality Criteria (AWQC) are established by the
EPA under the CWA. Section 304(a)(1) states that aquatic life criteria serve as recommendations
to states and authorized Tribes by defining ambient water concentrations that are expected to
protect against unacceptable adverse ecological effects to aquatic life resulting from exposure to
pollutants found in water. States and authorized Tribes may adopt these criteria into their water
quality standards (WQS) to protect the designated uses of water bodies. States and authorized
Tribes may also modify these criteria to reflect site-specific conditions or use other scientifically
defensible methods to develop criteria before adopting these into standards. After adoption,
states/authorized Tribes submit new and revised WQS to the EPA for review and approval or
disapproval. When approved by the EPA, the state's/Tribe's WQS become the applicable WQS
for CWA purposes. Such purposes include identification of impaired waters and establishment of
Total Maximum Daily Loads (TMDLs) under CWA Section 303(d) and derivation of water
quality-based effluent limitations in permits issued under the CWA Section 402 National
Pollutant Discharge Elimination System (NPDES) program. The EPA recommends the adoption
of both the acute and chronic water column criteria as well as the chronic tissue-based criteria to
ensure the protection of aquatic life through all exposure pathways, including direct aqueous
exposure and bioaccumulation. Aquatic life benchmarks, developed by EPA under 304(a)(2) of
the CWA, are informational values that EPA generates when there are limited high quality
toxicity data available and data gaps exist for several aquatic organism families. EPA provided
an acute estuarine/marine benchmark in Appendix L as additional information on protective
values that states and tribes may consider in their water quality programs.
This document provides a critical review of toxicity data identified in the EPA's literature
search for PFOA, including the anionic form (CAS No. 45285-51-6), the acid form (CAS No.
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335-67-1), and the ammonium salt (CAS No. 3825-26-1). It also quantifies the toxicity of PFOA
to aquatic life and provides criteria to protect aquatic life in freshwater from the acute and
chronic toxic effects of PFOA.
The EPA derived the recommended criteria using the best available data to reflect the
latest scientific knowledge on the toxicological effects of PFOA on aquatic life. The EPA
developed the criteria following the general approach outlined in the EPA's "Guidelines for
Deriving Numerical Water Quality Criteria for the Protection of Aquatic Organisms and Their
Uses" (U.S. EPA 1985). The PFOA criteria, if adopted and implemented, are expected to be
protective of most aquatic organisms, including species listed as threated or endangered, in the
community (i.e., approximately 95 percent of tested aquatic organisms representing the aquatic
community) and are derived to be protective of aquatic life designated uses established by states
and authorized Tribes for freshwaters. The estuarine/marine benchmarks are also intended to be
protective of aquatic life designated uses, but they are based on fewer empirical PFOA toxicity
data and, therefore, have greater inherent uncertainty. The criteria recommendations presented
herein are the EPA's best estimate of the concentrations of PFOA, with associated frequency and
duration specifications, that would protect sensitive aquatic life from unacceptable acute and
chronic effects of PFOA.
1.1 Previously Derived PFOA Toxicity Values and Thresholds
Other jurisdictions (e.g., states, countries, etc.) have previously published PFOA acute
and chronic criteria, benchmarks, or thresholds, including values for both freshwater and marine
systems. These values focus exclusively on water column-based values only; no other
jurisdiction has previously derived tissue-based values. Within the United States, no states or
Tribes have CWA Section 303(c) approved PFOA water quality standards for the protection of
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aquatic life. However, several states have published draft/interim acute and chronic ecological
screening level values/benchmarks for the protection of aquatic life (e.g., Michigan, Minnesota,
Texas, Florida, California).
These publicly available freshwater acute values range from 4.47 mg/L in Texas (TCEQ
2021) to 20 mg/L in Florida (Stuchal and Roberts 2019) (Table 1-1). The EPA's freshwater acute
PFOA criterion (3.1 mg/L) is slightly lower than the range of state-derived values. No acute
estuarine/marine criteria, benchmarks, or protective values have been established for PFOA,
other than the benchmark provided herein.
Publicly available freshwater chronic values for other jurisdictions range from 0.22 mg/L
in Australia/New Zealand (95% species protection level; CRC CARE 2017; EPAV 2016; HEPA
2020; Table 1-1) to 2.27 mg/L in Texas (TCEQ 2021), which are all higher than the EPA's
freshwater chronic criterion of 0.10 mg/L, which includes more recently published data not
accounted for in other previously published values.
Previously published estuarine/marine chronic values are available for Australia/New
Zealand with a chronic protective value of 0.22 mg/L (95% species protection level) and
California with a chronic "interim final screening level value" of 0.54 mg/L (Table 1-1).
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Table 1-1. Previously Derived PFOA Toxicity Values and Thresholds.
State/Country
of
Applicability
Aquatic l.il'e Protective
Value (mg/F)
Criteria or Benchmark and Calculation Approach
Source
Freshwater Acute
Texas
4.47
Based on NOAELs, LOAELs, or similar values from specific toxicological
studies. Contact the TCEQ for more information. This is an acute surface water
benchmark and does not represent a CWA Section 303(c) approved water
quality standard for PFOA.
TCEQ 2021
Michigan
7.7
Calculated from a species sensitivity distribution (SSD) consisting of two
species-specific values. The Final Acute Value (FAV) was based on the lowest
ECso divided by a safety factor of 13 (following the U.S. EPA Great Lakes
Initiative [GLI; U.S. EPA 1995a]). This protective value is a translation of
narrative water quality criteria and does not represent a CWA Section 303(c)
approved water quality standard for PFOA.
EGLE 2010
Minnesota
15
Calculated from a species sensitivity distribution (SSD) consisting of three
species-specific values. The Maximum Criterion (MC) was based on the lowest
EC50 divided by a safety factor of 13 (following the U.S. EPA Great Lakes
Initiative [GLI; U.S. EPA 1995a]). This draft value is does not represent a
CWA Section 303(c) approved water quality standard for PFOA.
STS/MPCA
2007
Florida
20
Secondary Acute Value (SAV) calculated using the U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995a) Tier II Methodology. FAV calculated as the
lowest GMAV (unspecified) divided by a safety factor of 5.2. This value was
released in a White Paper sponsored by Florida Department of Environmental
Protection and is considered a draft eco-based surface water screening level, it
is not a CWA Section 303(c) approved water quality standard.
Stuchal and
Roberts
2019
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State/Country
of
Applicability
Aquatic Lil'e Protective
Value (mg/L)
Criteria or Benchmark and Calculation Approach
Source
Freshwater Chronic
Australia, New
Zealand
0.019
(99% species protection - high
conservation value systems)
Guidelines calculated from a species sensitivity distribution (SSD) consisting of
12 species-specific values for fish, insects, crustaceans, rotifers, algae, and
plants following the guidance of Warne et al. (2018) and Batley et al. (2014)
CRC CARE
2017,
EPAV
2016,
HEP A 2020
0.22
(95% species protection -
slightly to moderately disturbed
systems)
0.632
(90% species protection - highly
disturbed systems)
1.824
(80% species protection - highly
disturbed systems)
California
0.54
(99% species protection)
HC1 calculated from an acute and chronic NOEC-based SSD as reported in
DoD-SERDP Project ER18-1614 (DoD-SERDP 2019). Acute NOEC values
were converted to chronic values using mean acute-to-chronic ratios derived
from Giesy et al. (2010). This value represents an "Interim Final Environmental
Screening Level" and does not represent a CWA Section 303(c) approved water
quality Standard for PFOA.
San
Francisco
Bay
RWQCB
2020; DoD-
SERDP
2019
Michigan
0.88
Final Chronic Value (FCV) was calculated as the FAV Final Acute: Chronic
ratio (ACR) (following the GLI; U.S. EPA 1995a). This protective value is a
translation of narrative water quality criteria and does not represent a CWA
Section 303(c) approved water quality standard for PFOA.
EGLE 2010
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State/Country
of
Applicability
Aquatic Life Protective
Value (mg/L)
Criteria or Benchmark and Calculation Approach
Source
Florida
1.3
Secondary Chronic Value (SCV) calculated using the U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995a) Tier II Methodology with acute-to-chronic
ratio (ACR) of 15.3. SCV = SAV (20,000 |ig/L) - ACR (15.3) = 1,300 |ig/L or
1.3 mg/L. This value was released in a White Paper sponsored by Florida
Department of Environmental Protection and is considered a draft eco-based
surface water screening level, it is not a CWA Section 303(c) approved water
quality standard.
Stuchal and
Roberts
2019
Minnesota
1.7
Chronic Criterion (CC) calculated as the FAV ^ a generic ACR following
Minnesota Rules Chapter 7050. No species-specific ACRs were available at the
time to calculate the FACR. This draft value idoes not represent a CWA
Section 303(c) approved water quality standard for PFOA.
STS/MPCA
2007
Texas
2.77
Based on NOAELs, LOAELs, or similar values from specific toxicological
studies. Contact the TCEQ for more information. This is a chronic surface
water benchmark and does not represent a CWA Section 303(c) approved water
quality standard for PFOA.
TCEQ 2021
Marine Chronic
California
0.54 (99% species
protection)
HC1 calculated from an acute and chronic NOEC-based SSD as reported in
DoD-SERDP Project ER18-1614 (DoD-SERDP 2019). Acute NOEC values
were converted to chronic values using mean acute-to-chronic ratios derived
from Giesy et al. (2010). This value represents an "Interim Final Environmental
Screening Level" and does not represent a CWA Section 303(c) approved water
quality Standard for PFOA.
San
Francisco
Bay
RWQCB
2020; DoD-
SERDP
2019
Australia, New
Zealand
0.019
(99% species protection - high
conservation value systems)
Freshwater values are to be used on an interim basis until final marine guideline
values can be set using the nationally agreed process under the Australian and
New Zealand Guidelines for Fresh and Marine Water Quality.
HEP A 2020
0.22
(95% species protection -
slightly to moderately disturbed
systems)
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Stsitc/Coiinlrv
of
Applicability
Aqusilic Life Protective
Value (mg/L)
Crilerisi or lieiichmnrk niul (nlcuhilion Approach
Source
0.632
(90% species protection - highly
disturbed systems)
1.824
(80% species protection - highly
disturbed systems)
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1.2 Overview of Per- and Polyfluorinated Substances (PFAS)
Perfluorooctanoic acid (PFOA), and its salts, belong to the per- and polyfluorinated
substances (PFAS) group of chemicals. PFAS are a large group of structurally diverse
anthropogenic chemicals that include PFOA, PFOS, and thousands of other fully or partially
fluorinated chemicals. There are many families or subclasses of PFAS, and each contains many
individual structural homologues and can exist as either branched-chain or straight-chain isomers
(Buck et al. 2011; U.S. EPA 2021). These PFAS families can be divided into two primary
categories: non-polymers and polymers. The non-polymer PFAS include perfluoroalkyl and
polyfluoroalkyl substances. Polymer PFAS include fluoropolymers, perfluoropolyethers, and
side-chain fluorinated polymers (Table 1-2). Several U.S. federal, state, and industry
stakeholders as well as European entities have posited various definitions of what constitutes a
PFAS. OECD, an international organization comprised of 38 countries, recently published a
practical guidance regarding the terminology of PFAS (OECD 2021). The OECD-led
"Reconciling Terminology of the Universe of Per- and Polyfluoroalkyl Substances:
Recommendations and Practical Guidance" workgroup provided an updated definition of PFAS,
originally posited in part by Buck et al. (2011), as follows: "PFASs are defined as fluorinated
substances that contain at least one fully fluorinated methyl or methylene carbon atom (without
any H/Cl/Br/I atom attached to it), i.e. with a few noted exceptions, any chemical with at least a
perfluorinated methyl group (-CF3) or a perfluorinated methylene group (-CF2-) is a PFAS". It
is not within the scope of this framework to compare and contrast the various definitions, or the
nuances associated with defining or scoping PFAS; rather the reader of this document is referred
to OECD (2021) for review. Generally, the structural definition of PFAS includes chemicals that
contain at least one of the following three structures:
8
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R-(CF2)-CF(R')R", where both the CF2 and CF moieties are saturated carbons, and none
of the R groups can be hydrogen (TSCA draft definition);
R-CF2OCF2-R', where both the CF2 and CF moieties are saturated carbons, and none of
the R groups can be hydrogen; and
CF3C(CF3)R'R ", where both the CF2 and CF moieties are saturated carbons, and none of
the R groups can be hydrogen.
It should also be noted that what defines or constitutes a PFAS may change or evolve
over time and under different purviews (e.g., federal, state, international).
9
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Table 1-2. Two Primary Categories of PFAS1.
I'l-AS Non-polymers
Slriiclural Klemciils
Kxample I'I'AS lamilies
Perfluoroalkyl acids
Compounds in which all carbon-
hydrogen bonds, except those on
the functional group, are replaced
with carbon-fluorine bonds
Perfluoroalkyl carboxylic and
sulfonic acids (e.g., PFOA,
PFOS), perfluoroalkyl phosphonic
and phosphinic acids,
perfluoroalkylether carboxylic and
sulfonic acids
Polyfluoroalkyl acids
Compounds in which all carbon-
hydrogen bonds on at least one
carbon (but not all) are replaced
with carbon-fluorine bonds
polyfluoroalkyl carboxylic acids,
polyfluoroalkylether carboxylic
and sulfonic acids
I'I'AS Polymers
Slriiclural Klcmenls
Kxample PI' AS lamilies
Fluoropolymers
Carbon-only polymer backbone
with fluorines directly attached
polytetrafluoroethylene,
polyvinylidene fluoride,
fluorinated ethylene propylene,
perfluoroalkoxyl polymer
Polymeric
perfluoropolyethers
Carbon and oxygen polymer
backbone with fluorines directly
attached to carbon
F-(CmF2mO-)nCF3, where the
CmF2mO represents -CF20, -
CF2CF20, and/or -CF(CF3)CF20
distributed randomly along
polymer backbone
Side-chain fluorinated
polymers
Non-fluorinated polymer
backbone with fluorinated side
chains with variable composition
n:l or n:2 fluorotelomer-based
acrylates, urethanes, oxetanes, or
silicones; perfluoroalkanoyl
fluorides; perfluoroalkane sulfonyl
fluorides
1: Amalgamation of information from Figure 9 of OECD (2021) and Buck et al. (2011).
PFOA belongs to the perfluoroalkyl acids (PFAAs) of the non-polymer perfluoroalkyl
substances category of PFAS (Table 1-2). PFAAs are among the most researched PFAS (Wang
et al. 2017). The family PFAAs includes perfluoroalkyl carboxylic, sulfonic, sulfinic,
phosphonic, and phosphinic acids (Table 1-3). PFAAs are highly persistent and are frequently
found in the environment (Ahrens 2011; Wang et al. 2017). PFAAs may dissociate to their
anions in aqueous environmental media, soils, or sediments depending on their acid strength
(pKa value). Although the protonated and anionic forms may have different physiochemical
10
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properties, the anionic form is the dominant form in the aquatic environment, including in the
toxicity tests used to derive the PFOA criteria.
Table 1-3. Classification and Chemical Structure of Perfluoroalkyl Acids (PFAAs).1
Classification
I'linclional Croup
r.xamplc
Perfluoroalkyl carboxylic acids
(PFCAs)
Or
Perfluoroalkyl carboxylates (PFCAs)
-COOH
Perfluorooctanoic acid (PFOA)2
-COO"
Perfluorooctanoate (PFO)
Perfluoroalkane sulfonic acids (PFSAs)
Or
Perfluorokane sulfonates (PFSAs)
-so3h
Perfluorooctane sulfonic acid (PFOS)
-so3-
Perfluorooctane sulfonate (PFOS)
Perfluoroalkyl sulfinic acids (PFSIAs)
-so2h
Perfluorooctane sulfinic acid (PFOSI)
Perfluoroalkyl phosphonic acids
(PFPAs)
-P(=0)(OH)2
Perfluorooctyl phosphonic acid (C8-
PFPA)
Perfluoroalkyl phosphinic acids
(PFPIAs)
-P(=0)(OH)(CmF2m+l)
Bis(perfluorooctyl) phosphinic acid
(C8/C8-PFPIA)
Perfluoroalkylether carboxylic acids
(PFECAs)
CF3(OCF2)COO-
perfluoro (3,5,7-trioxaoctanoic) acid
Perfluoroalkylether sulfonic acids
(PFESAs)
CF3(0CF2)S03H
6:2 Cl-PFESA
Perfluoroalkyl dicarboxylic acids
(PFdiCAs)
hooc-cf2-cooh
9:3 Fluorotelomer betaine
Perfluoroalkane disulfonic acids
(PFdiSAs)
H03S-CF2-S03H
Perfluoro-l,4-disulfonic acid
1: Modified from Buck et al. (2011) and OECD (2021).
2: At most environmentally relevant pH conditions, PFOA occurs in the anionic form.
Perfluoroalkyl carboxylic acids (PFCAs), including PFOA, consist of a general chemical
structure of CnF2n+iCOOH. This chemical structure makes PFOA (see Figure 1-1) extremely
strong and stable, and resistant to hydrolysis, photolysis, microbial degradation, and metabolism
(Ahrens 2011; Beach et al. 2006; Buck et al. 2011).
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HO^^O
Figure 1-1. Chemical Structure of the Linear Isomer of Perfluorooctanoic acid (PFOA).
(Source: United States EPA Chemistry Dashboard; https://comptox.epa.gov/dashboard).
1.2.1 Physical and Chemical Properties of PFOA
Physical and chemical properties along with other reference information for PFOA are
provided in Table 1-4. These physical and chemical properties helped to define the
environmental fate and transport of PFOA in the aquatic environment. In the environment,
PFOA rapidly ionizes in water to its anionic form (perfluorooctanoate, PFO). PFOA is highly
stable and is resistant to hydrolysis, photolysis, volatilization, and biodegradation (UNEP 2015).
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Table 1-4. Chemical and Physical Properties of PFOA.
Property
PI-OA. acidic form1
Source
Chemical
Abstracts Service
Registry Number
(CASRN)
335-67-1
Chemical
Abstracts Index
Name
2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-
pentadecafluorooctanoic acid
Synonyms
PFOA; Pentadecafluoro-1-
octanoic acid; Pentadecafluoro-n-
octanoic acid; Octanoic acid,
pentadecafluoro-;
Perfluorocaprylic acid;
Pentadecafluorooctanoic acid;
Perfluoroheptanecarboxylic acid;
Chemical
Formula
C8HF15O2
Molecular Weight
(grams per mole
[g/moll)
414.07
PubChem Identifier (CID 9554) (URL:
https://pubchem.ncbi.nlm.nih.gov/compound/9554);
Lide(2007)
Color/Physical
State
White powder (ammonium salt)
PubChem Identifier (CID 9554) (URL:
httDs://Dubchem.ncbi.nlm.nih.gov/comDOund/9554)
Boiling Point
192.4 ฐC
HSDB (2012); Lide (2007); SRC (2016)
Melting Point
54.3 ฐC
HSDB (2012); Lide (2007); SRC (2016)
Vapor Pressure
0.525 mm Hg at 25 ฐC
(measured)
0.962 mm Hg at 59.25 ฐC
(measured)
Hekster et al. (2003); HSDB (2012); SRC (2016)
AT SDR (2015); Kaiser et al. (2005)
Kaw
0.00102 (experimentally
determined; equivalent to
Henry's Law Constant of
0.000028 Pa-m3/mol at 25 ฐC)
Li et al. (2007)
Kow
Not measurable
UNEP (2015)
Organic carbon
water partitioning
coefficient (Koc)
2.06
Higgins and Luthy (2006)
pKa
3.15 (mean measured)
Burns et al. (2008) and 3M Company (2003) as
reported in EPA Chemistry Dashboard (URL:
https://comptox.epa.gov/dashboard/dsstoxdb/results
?search=DTXSID8031865#DroDerties )
Solubility in
Water
9,500 mg/L (estimated);
3,300 mg/L at 25 ฐC (measured)
Hekster et al. (2003);
Inoue et al. (2012)
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Property
PI-OA. sicidic form1
Source
Half-Life in
Water
Stable
UNEP (2015)
Half-Life in Air
Stable
UNEP (2015)
1: PFOA is most commonly produced as an ammonium salt (CASRN 3825-26-1). Properties specific to the salt are
not included.
PFOA is water soluble, nonvolatile, and stable, with a low vapor pressure and is a solid at
room temperature (UNEP 2015). The EPA's chemistry dashboard reported a mean experimental
acid dissociation constant (pKa) for PFOA of 3.15 calculated from pKa values determined from
Burns et al. (2008) and 3M Company (2003). Burns et al. (2008) measured an acid dissociation
constant (pKa) for PFOA of 3.8 using a standard water-methanol mixed solvent approach, which
indicates PFOA is a moderate acid, while 3M Company (2003) reported a measured PFOA pKa
of 2.5.
Due to the surfactant properties of PFOA, it forms three layers when added to octanol and
water in a standard test system used to measure a n-octanol-water partition co-efficient (Kow),
thus preventing direct measurement (EFSA 2008; Giesy et al. 2010). Although a Kow cannot be
directly measured, a Kow for PFOA has been estimated from its individual water and octanol
solubilities (estimated PFOA Kow range = 2.69 - 6.3; UNEP 2015); however, the veracity of
such estimates is uncertain (UNEP 2015). Lacking a reliable Kow for PFOA precludes
application of Kow-based models commonly used to estimate various physiochemical properties
for organic compounds, including bioconcentration factors and soil adsorption coefficients.
Further, the unusual characteristics of PFOA would bring into question the use of Kow as a
predictor of environmental behavior; for example, bioaccumulation of PFOA is thought to be
mediated via binding to proteins rather than partitioning into lipids (EFSA 2008; Giesy et al.
2010), the latter being the theoretical basis for Kow-based prediction of bioaccumulation.
14
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2 PROBLEM FORMULATION
A problem formulation provides a strategic framework for water quality criteria
development under the CWA by focusing on the most relevant chemical properties and
endpoints. In the problem formulation, the purpose of the assessment is stated, the problem is
defined, and a plan for analyzing and characterizing risk is developed. The structure of this
problem formulation was consistent with the EPA's Guidelines for Ecological Risk Assessment
(U.S. EPA 1998).
2.1 Overview of PFOA Sources
2.1.1 Manufacturing of PFOA
PFOA is primarily produced through Electrochemical Fluorination (ECF) in which an
organic raw material, in the case of PFOA as octanoyl fluoride (C7H15COF), undergoes
electrolysis in anhydrous hydrogen fluoride solution. This electrolysis leads to the replacement
of all the hydrogen atoms by fluorine atoms and results in perfluorooctanoyl fluoride
(C7F15COF), which is the major raw material used to manufacture PFOA and PFOA salts (Figure
2-1; Buck et al. 2011). Electrochemical Fluorination typically results in a mixture of branched
and linear isomers.
15
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C7H15COCI
C7H15COF
(Octanoyl Fluoride)
S I g
ง I c
^ 1 o
ฎ I
HF, e | ง HF, e
C7F15COF
(Perfl uorooctanoy 1
Fluoride)
Treatment with. Strong
Acid and Base
C7F15CO2M"
(PFOA Salts)
Figure 2-1. Synthesis of Perfluorooctanoic acid (PFOA) by Electrochemical Fluorination
(ECF).
Modified from Buck et al. (2011).
Initial production of PFOA started in the 1940s and commercial production and use as
protective coatings starting in the mid-1950s. From 1951 - 2004 the total global historic industry
wide emission of PFCAs (including PFOA) from all sources (i.e., direct and indirect sources
such as manufacture, use, consumer products, and PFC A precursors) ranged from 3,200 tons to
7,300 tons (Prevedouros et al. 2006). In 2006, the EPA initiated the 2010/2015 PFOA
Stewardship Program, resulting in major PFOA producers committing to a 95% reduction in
PFOA facility emissions and product contents across the globe by 2010. The 2010/2015 PFOA
Stewardship Program further aimed to eliminate PFOA emissions and product content by 2015
(U.S. EPA 2006).
C7F15CO2H
(PFOA)
16
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2,1,2 Sources of PFOA to Aquatic Environments
PFCAs, including PFOA are primarily released into water (95% of PFCAs are emitted to
water; 3M Company 2000b) and can enter the aquatic environment from both industrial and
consumer products during manufacturing, along the supply chains, during product use and/or
disposal (Ahrens and Bundschuh 2014; Kannan 2011). Occurrence of PFOA in the aquatic
environment arises from both direct and indirect sources (Ahrens et al. 201 la). However, the
quantitative assessments of their production, direct and indirect emissions, and environmental
measurements are lacking (Ahrens and Bundschuh 2014; Prevedouros et al. 2006).
The direct sources of PFOA to the aquatic environment include both municipal and
industrial wastewater treatment plants (WWTPs), landfill leachate, and runoff from contaminated
biosolids (Renner 2009). WWTPs in particular are an important source of PFOA to aquatic
systems (Ahrens et al. 2009).
Indirect sources of PFOA to aquatic environments include dry and wet atmospheric
deposition, runoff from contaminated soils, and consumer product use and disposal (Kannan
2011). Identification of indirect sources of PFOA and understanding their relative contribution to
aquatic ecosystems is difficult. Overall, the presence of indirect sources of PFOA and their
contributions are dependent on the system and the nearby land uses. Overall PFAS
concentrations, including PFOA, in the environment are positively correlated with population
density. Overall, PFOA occurrence in aquatic environments is driven by legacy PFOA sources
because use of PFOA in the United States was largely phased out by 2010, and completely
phased out by 2015 in accordance with the EPA's 2010/2015 PFOA Stewardship Program.
In addition to direct discharge, environmental breakdown of precursor compounds
containing a seven-member perfluoro moiety can provide an additional source of PFOA.
Metabolic transformation of PFAS precursors, such as PAPs, FTCAs, FTUCAs, FTSAs, and
17
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FASAAs, and the degradation of volatile PFAS, such as FTOHs, FASAs, and FASEs, can be
potential sources of PFOA as these compounds can transform into more persistent PFAS,
including PFCAs and PFOA (Ahrens and Bundschuh 2014). For example, fluoroacrylate
polymers can breakdown in soil and release fluorotelomer alcohol (FTOH) which can further
degrade into PFOA (Russell et al. 2008). Similarly, polyfluoroalkyl phosphoric acid diesters
(diPAPs) are used in commercial applications, such as food packaging, and can be found in
WWTP sludge and contaminated biosolids. In environmental media, diPAPs can release FTOH
that further degrades into PFOA (Lee et al. 2010; Sinclair and Kannan 2006; Washington et al.
2009). Current understanding of these transformation processes remains limited, and additional
work is needed to fully understand these processes and their role in generation of sources of
PFOA to aquatic environments (Lau et al. 2007).
PFOA can also be re-emitted to the aquatic environment from ice melt and sediment
transport. Release of PFOA will continue into the future from the transformation of other PFAS
and the historical products still in use (e.g., consumer goods manufactured and/or obtained
before the PFOA discontinuation).
2.2 Environmental Fate and Transport of PFOA in the Aquatic Environment
2.2,1 Environmental Fate of PFOA in the Aquatic Environment
In natural waters near neutral pH, PFOA rapidly dissociates into ionic components. In
aquatic environments, PFOA has an affinity to remain in the water column rather than sediments,
but can also adsorb to sediments in the presence of organic carbon, with the partitioning
coefficients (Kd) increasing with salinity (Canada 2012; Hekster et al. 2003). Because of its
water solubility and preferential binding to proteins, once PFOA enters a waterbody it tends to
remain dissolved in the water column, where it is mobile, unless it adsorbs to organic particulate
matter or is assimilated by organisms.
18
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PFOA has low volatility in the ionized form but can adsorb to particles in air where it can
be transported globally, including remote locations (Del Vento et al. 2012; Shoeib et al. 2006).
PFOA is water soluble and has been found in surface water, ground water, and drinking water.
Because of the relatively low log Koc of PFOA, it does not easily adsorb to sediments and tends
to stay in the water column.
In the water column, and other environmental compartments, PFOA is stable and
resistant to hydrolysis, photolysis, volatilization, and biodegradation (Higgins and Luthy 2006;
UNEP 2015). The persistence of PFOA has been attributed to the strong carbon-fluorine (C-F)
bond. Additionally, there are limited indications that naturally occurring defluorinating enzymes
exist that can break a C-F bond. Consequently, no biodegradation or abiotic degradation
processes for PFOA are known. In aquatic environments, the only dissipation mechanisms for
PFOA are physical mechanisms, such as environmental dilution and sorption.
2.2.2 Environmental Transport of PFOA in the Aquatic Environment
The physiochemical properties discussed in Section 1.2.1 above enable PFOA to be
highly persistent in the aquatic environment. PFOA tends to be distributed in waters and in the
atmosphere (Ahrens 2011; Yamashita et al. 2008). PFOA concentrations in seawater are
typically greater than PFOS, which has been attributed to the relatively lower bioaccumulation
potential, lower sorption to sediments, and greater water solubility (Ahrens 2011; Ahrens et al.
2009).
Numerous uncertainties exist in the understanding of environmental transport of PFOA in
aquatic systems. Both point and non-point sources contribute PFOA to the aquatic environment.
PFOA can be transported from these sources into rivers, streams, lakes, and marine
environments. There is a general decrease in PFOA concentrations along a transport pathway
resulting from dilution in the water column. For example, measured PFOA concentrations in
19
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WWTP effluents were generally an order of magnitude greater than riverine concentrations, with
the upper end of riverine concentrations being similar to WWTP effluents. Although minimum
and maximum PFOA concentrations in coastal waters (ranging from hundreds of pg/L to several
ng/L) were below corresponding measurements in riverine systems, coastal PFOA
concentrations in general were largely similar to riverine concentrations. Open oceans contained
the lowest PFOA concentrations resulting from immense dilution. Overall, open ocean
concentrations of PFOA were roughly 2.5 orders of magnitude lower than those reported in
WWTP effluents (Ahrens 2011).
PFOA has been found in a diversity of environments, including in the arctic and
Antarctic, despite the limited number of manufacturing facilities and/or small population sizes
typically found in these areas (Del Vento et al. 2012; Shoeib et al. 2006). Although PFOA has
low volatility, particularly in the ionized form, it can absorb to air particles before being
deposited via atmospheric deposition to these remote regions. For example, Kim and Kannan
(2007) reported PFOA in snow in the United States ranging from below the limit of
quantification to 20 ng/L. Similarly, Young et al. (2007) reported mean PFOA concentrations in
snow in Canada ranging from 0.01 ng/L to 0.15 ng/L.
The continued presence of PFOA in open oceans and in remote polar areas may be due to
multiple exposure pathways, including those caused by direct production, use, and discharge of
PFOA itself, degradation and transformation of precursor compounds, and via long range
aqueous and atmospheric transport (Armitage et al. 2009).
2.3 Transformation and Degradation of PFOA Precursors in the Aquatic
Environment
Included among major sources of PFOA to the environment is from the abiotic and biotic
transformation and degradation of polyfluoroalkyl precursor substances (see Section 2.2.2
20
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above). Polyfluoroalkyl substances are one type of precursor substance that have the potential to
be transformed abiotically or biotically into PFCAs such as PFOA (Buck et al. 2011). On a
global scale, production volumes of polyfluoroalkyl substances, many of which are likely
polyfluoroalkyl precursor substances that ultimately degrade or transform to PFOA, greatly
exceed direct emissions of PFOA through its manufacture, use and disposal (Butt et al. 2014; Liu
and Mejia Avendano 2013). According to the OECD (2006), there were approximately one
thousand polyfluorylalkyl chemicals commercially produced at the time that could conceivably
degrade to PFCAs such as PFOA. For example, Buck et al. (2011) identified 42 families of
compounds and numerous individual PFAS detected in environmental and human matrices,
many of which have not been evaluated for their biodegradability (Liu and Mejia Avendano
2013). Any or all members of these PFAS have the ability to be transformed or degraded to
PFAAs such as PFCAs or perfluoroalkane sulfonic acids (PFSAs).
Critical reviews by Butt et al. (2014) and Liu and Mejia Avendano (2013) provided a
comprehensive summary of the qualitative and quantitative relationships between biodegradation
and transformation of polyfluoroalkyl precursors and generation of PFOA and other PFCAs. The
most well-studied polyfluoroalkyl precursor substances are fluorotelomer-based compounds,
which are produced through telomerization technology and are associated with PFOA as the final
product (Buck et al. 2011).
2.3,1 Biodegradation of fluorotelomer-based precursors
The aerobic biodegradation pathway of fluorotelomer alcohols (8:2 and 6:2 FTOH) have
been thoroughly studied. Dinglasan et al. (2004) was among the first to investigate the
biotransformation of 8:2 FTOH in a mixed microbial culture. Additional studies of the aerobic
microbial degradation of 8:2 FTOH by Liu et al. (2010) and Wang et al. (2005, 2009, and 2012)
have since confirmed the formation of PFOA via this pathway. The observed half-lives of 8:2
21
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FTOH ranged from <2 days to 30 days in these laboratory studies. Molar yield of PFOA ranged
anywhere from 0.5 to 40% depending on type of microbes or microcosm used in the study, with
Wang et al. (2009) observing relatively higher PFOA yield in aerobic soils relative to PFOA
yield in pure bacterial culture (Liu et al. 2010). Thus, aerobic microbial degradation of 8:2 FTOH
can be a significant source of PFOA in some environmental compartments. Anaerobic microbial
degradation of 8:2 FTOH, on the other hand, is inefficient and likely an insignificant source of
PFOA to the environment (Zhang et al. 2013b). Additional studies are needed, however, to
understand anaerobic biodegradability of FTOHs and related compounds in general (Liu and
Mejia Avendano 2013).
Aerobic biodegradation of several other types of fluorotelomer-based, polyfluoroalkyl
precursor substances definitively linked to PFOA formation include: fluorotelomer stearate (8:2
FTS) with observed half-life in aerobic soils of 5-28 days and molar yield of about 01.7-4%
(Dasu et al. 2012, 2013); fluorotelomer acrylate (8:2 FTAC) and fluorotelomer methacrylate (8:2
FTMAC) monomers with observed half-life in aerobic soils of 3-5 days and 15 days and molar
yields of 7.8 and 10%, respectively (Royer 2011); fluorotelomer ethoxylates (FTEOs) with
observed half-life in unfiltered WWTP effluent of approximately one day and molar yield of
about 0.3%) (Fromel and Knepper 2010); and fluorotelomer urethanes, specifically, aliphatic
diurethane ester (8:2 HMU), with an half-life in aerobic soils of >180 days and molar yield of
0.9% (Dasu 2011).
2.3.2 Biodegradation of side-chain polymers
At present, a crucial need exists to understand the potential degradation of side-chain
fluorinated polymers in natural environments because they currently represent a high percentage
of all commercial and industrial PFAS sales products (Liu and Mejia Avendano 2013). Side-
chain polymers are those with polyfluoroalkyl or perfluoroalkyl chains attached to non-
22
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fluorinated backbones (Buck et al. 2011). Russell et al. (2008) investigated the biodegradation of
a high molecular weight (-40,000 amu, 100-300 nm in diameter) polyacrylate polymer aqueous
dispersion product in four aerobic soils over two years. Two approaches (molar mass balance
and kinetic modeling) gave conflicting results. The molar mass balance approach indicated no
evidence of biodegradation because the PFOA generated was mostly accounted for by impurity
(residual non-polymerized PFAS) degradation. Conversely, using the kinetic modeling approach
half-lives of PFOA were estimated to be around 1,200-1,700 years among the four soils tested.
Upon further investigation using a low molecular weight (-3,500 amu) polyurethane polymer
product and a similar approach, Russell et al. (2010) clearly demonstrated biodegradability of the
low molecular weight polyurethane polymer product compared to the polyacrylate polymer, as
the levels of PFOA produced were several orders of magnitude greater than what the impurities
could account for. Applying a similar kinetic modeling approach, the half-lives of the
polyurethane polymer were estimated to range from 28 to 241 years among the four test soils.
Given the large disparity in half-life prediction between the two studies, however, additional
research is needed to clarify the contributions of polyfluoroalkyl polymers to PFOA formation
due to the high percentage of side-chain fluorinated polymers that exist in commercial and
industrial sales products.
2.3,3 Biodegradation of other polyfluoroalkyl substances
Recently, Mejia-Avendano et al. (2016) examined the formation of PFOA from aerobic
biotransformation of quaternary ammonium polyfluoroalkyl surfactants (QACs). Capitalizing on
several recent studies focused on identifying specific PFAS in major PFAS-based aqueous film-
forming foam (AFFF) formulations, all the newly identified PFAS were polyfluoroalkyl
compounds. These compounds have perfluoroalkyl carbon chain lengths varying from four to 12
and possess functionalities such as sulfonyl, thioether, tertiary amine, quaternary ammonium,
23
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carboxylate, sulfonate, amine oxide, andbetaine, etc. (Mejia-Avendano et al. 2016). Importantly,
the identified cationic PFAS in these studies contain either tertiary amine or quaternary
ammonium groups. In this first study of the fate of polyfluoroalkyl cationic surfactants used in
aqueous AFFF formulations, the biotransformation of perfluorooctaneamido quaternary
ammonium salt (PFOAAmS) was characterized by a DT50 value (time necessary to consume
half of the initial mass) of 142 days and significant generation of perfluoroalkyl carboxylic acid
(PFOA) at a yield of 30% (mol) by day 180. Three novel biotransformation intermediates were
identified for PFOAAmS, and it was demonstrated that despite overall high stability of QACs
and their biocide nature, the ones with perfluoroalkyl chains can be substantially biotransformed
into perfluoroalkyl acids in aerobic soil.
The above microbial biotransformation and degradation pathways are all dependent on
environmental conditions, degradation kinetics, and the chemical structures and properties of the
individual polyfluoroalkyl precursors (Buck et al. 2011; Butt et al. 2014; Liu and Mejia
Avendano 2013). Of particular importance is the environmental stability of key chemical
linkages (such as esters and ethers) as the stability of these chemical linkages determines the
stability of the overall PFAS (Liu and Mejia Avendano 2013). It is evident through these studies
that the biotransformation and biodegradability of polyfluoroalkyl precursor substances is due to
the breakdown of the non-fluorinated functionality of the precursor substances, which precedes
the breakdown of the perfluorinated carbons. In contrast, perfluoroalkyl chemicals in general
resist biotransformation and defluorination under natural conditions. Using 14C-labeled PFOA to
examine five different microbial communities, a range of electron donors for reductive
defluorination processes, and the possibility of co-metabolism during reductive dechlorination of
24
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trichloroethene, Liou et al. (2010) was able to confirm that PFOA is highly resistant to microbial
degradation in natural environments.
2.3.4 Non-microbial biodegradation of other polyfluoroalkyl substances
Butt et al. (2014) reviewed the current state of knowledge regarding the
biotransformation of fluorotelomer-based, polyfluoroalkyl precursor substances that degrade to
form PFCAs (PFOA) in microbial systems, rats, mice, and fish. Consistent with information
presented above, the majority of biotransformation studies thus far used 8:2 FTOH (a
fluorotelomer alcohol) as the substrate; only a few studies of non-FTOH biotransformation exist.
The biotransformation studies of 8:2 FTOH metabolism universally show the formation of
PFOA. As above, the overall yield of PFOA is low, presumably because of the multiple branches
in the biotransformation pathways, including conjugation reactions in animal systems which are
capable of phase II metabolism and results in the formation of conjugated metabolites such as
glucuronide, sulfate, and glutathione metabolites. Butt et al. (2014) also showed that
fluorotelomer carboxylates (FTCAs) appear to be more stable in animal models, whereas they
are relatively labile in microbial systems. In contrast, the unsaturated form of FTCAs - FTUCAs
appear to be readily degraded in animal models.
2.4 Environmental Monitoring of PFOA in Abiotic Media
PFOA has been detected in a variety of environmental abiotic matrices in aquatic
environments around the globe. These abiotic media include surface water, soils, sediments,
groundwater, air, and ice caps (Butt et al. 2010; Lau et al. 2007). Water is expected to be the
primary environmental media that PFOA is found (Lau et al. 2007). Occurrence and detection of
PFOA in other aquatic abiotic media found in the aquatic environment are summarized below.
25
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2.4,1 PFOA Occurrence and Detection in Ambient Surface Waters
PFOA is one of the dominant PFAS detected in ambient surface waters, along with PFOS
(Ahrens 2011; Benskin et al. 2012; Dinglasan-Panlilio et al. 2014; Nakayama et al. 2007;
Remucal 2019; Zareitalabad et al. 2013). Most of the current, published PFOA occurrence
studies have focused on a handful of broad geographic regions, many times targeting sites with
known manufacturing or industrial uses of PFAS, such as the Great Lakes, the Cape Fear River
and waterbodies near Decatur, Alabama (Figure 2-2; Boulanger et al. 2004; Cochran 2015;
Hansen et al. 2002; Konwick et al. 2008; Nakayama et al. 2007; 3M Company 2001).
26
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Figure 2-2. Map Indicating Sampling Locations for Perfluorooctanoic acid (PFOA)
Measured in Surface Waters Across the United States (U.S.) Based on Data Reported in the
Publicly Available Literature.
Colorado sampling coordinates were not available, these data are represented by the dash marks to
indicate measured PFOA surface water concentrations are available in Colorado.
Concentrations of PFOA in surface waters vary widely (Figure 2-3), with observed
concentrations ranging over seven orders of magnitude and detected generally between pg and
ng per liter with some sites with reported concentrations in |ig/L (Zareitalabad et al. 2013). For
the purposes of this overview and comparison, all concentrations reported here are in ng per liter
(ng/L). Unlike other contaminants commonly found in aquatic ecosystems, PFAS are synthetic
compounds and therefore have no natural source. Thus, the occurrence of any PFAS in the
environment is an indication of anthropogenic sources, including consumer and industrial use,
27
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long-range transport, atmospheric deposition, surface water runoff, and general persistence in the
environment.
10000
1000
^ 100
<
ฃ 10
Oi
s
o
0.1
0.01
0.001
Figure 2-3. Distribution of the minimum and maximum concentrations (ng/L) of
Perfluorooctanoic acid (PFOA) measured in surface waters for each state or waterbody
(excluding the Great Lakes) with reported data in the publicly available literature.
The distribution is arranged alphabetically by state and waterbody.
PFOA concentrations in surface water tend to increase with levels of urbanization.
Across the Great Lakes region, PFOA was higher in the downstream lakes of Erie and Ontario
and lower in the upstream lakes of Superior, Michigan, and Huron (Remucal 2019). Similarly,
Zhang et al. (2016) observed measured PFOA concentrations in urban areas (urban average
PFOA concentration = 10.17 ng/L; n = 20) to be more than three time greater than concentrations
in rural areas (rural average PFOA concentration = 2.95 ng/L; n = 17) within New Jersey, New
York, and Rhode Island. Temporal variation of PFOA in surface waters remains largely
unknown due to data limitations. See Appendix M for further discussion of PFOA occurrence in
surface waters and other abiotic media such as aquatic sediments, groundwater, air, and ice.
ฆ
n |i
U ซ
~ 1
ฆ
AL CA CO DE FL GA LA MI MN NJ NM NY NC EI SC IN H WA
River
28
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2.5 Bioaccumulation and Biomagnification of PFOA in Aquatic Ecosystems
PFOA is found in aquatic ecosystems around the globe (e.g., Ankley et al. 2020; Giesy
and Kannan 2001; Houde et al. 2008). Although they were used predominantly in more
populated areas, these compounds are resistant to hydrolysis, photolysis, and biodegradation,
which facilitates their long-range transport to aquatic ecosystems in the remote arctic and mid-
oceanic islands (Haukas et al. 2007; Houde et al. 2006a). Several physical-chemical properties of
PFAS contribute to their bioaccumulation within aquatic and aquatic-dependent species once
they have entered an aquatic ecosystem.
2.5.1 PFOA Bioaccumulation in Aquatic Life
In contrast to many persistent organic pollutants which tend to partition to fats, PFOA
preferentially binds to proteins (Martin et al. 2003a, 2003b). Within the body, PFOA tends to
bioaccumulate within protein-rich tissues, such as the blood serum proteins, liver, kidney, and
gall bladder (De Silva et al. 2009; Jones et al. 2003; Martin et al. 2003a, 2003b). PFOA may also
bind to ovalbumin, and the transfer of PFOA to such albumin in eggs can be an important
mechanism for depuration in female oviparous species, as well as a mechanism for maternal
transfer (Jones et al. 2003; Kannan et al. 2005).
The stability of PFOA contributes to its bioaccumulation potential, as PFOA has not been
found to undergo biotransformation within the organism and is primarily depurated through
excretion in urine or across gill surfaces (De Silva et al. 2009; Martin et al. 2003b). Within an
organism, PFOA may undergo enterohepatic recirculation, in which PFOA is excreted from the
liver in bile to the small intestine, then reabsorbed and transported back to the liver (Goecke-
Flora and Reo 1996). Among PFAS, this process becomes increasingly more efficient the longer
the perfluorinated chain length, resulting in longer half-lives for chemicals like PFOA with a
relatively long chain length, as they are less readily excreted. PFAS with carboxylate head
29
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groups, such as PFOA, are less efficiently resorbed by the small intestine and transported back to
the liver than sulfonate PFAS, resulting in lower bioaccumulation levels (Hassell et al. 2020;
Martin et al. 2003a).
Sex differences in the elimination rate of PFOA chemicals have been observed in some
species. Lee and Schultz (2010) observed that the elimination rate of PFOA from blood plasma
was ten times faster in female fathead minnows compared to males. The faster elimination rate
may be related to sex hormones (i.e., androgen and estrogen) levels, as the elimination rate in
females decreased four-fold following exposure to the androgen trenbolone (Lee and Schultz
2010). This pattern has also been observed in rats, where the elimination of PFOA was 70 times
faster in females than males and was attributed to sex-related differences in the expression of
organic anion transporters in kidneys resulting in higher excretion rates (Kudo et al. 2002). The
mechanism for the higher elimination rate in female fathead minnows has not been determined,
and the degree to which gender-related differences in elimination rate apply to other fish species,
or other taxonomic groups, is unknown. However, it does suggest that the sex of the organism
should be considered when assessing ecosystem level bioaccumulation, and that there may be
another mechanism in addition to egg production that can result in lower concentration of PFAS
in females.
The structure of PFOA also contributes to its bioaccumulation potential, with linear
forms being more bioaccumulative than branched forms (De Silva et al. 2009; Hassell et al.
2020). The preferential accumulation of linear PFOA occurs because the elimination rate, of
branched isomers of PFOA is higher, particularly across gill surfaces (De Silva et al. 2009). This
pattern has also been observed in the field, as the proportion of branched isomers was higher in
30
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water and sediment compared to fish tissue in Taihu Lake, China (Fang et al. 2014) and Lake
Ontario (Houde et al. 2008).
2.5.2 Factors Influencing Potential for PFOA Bioaccumulation and Biomagnification in
Aquatic Ecosystems
PFOA binding to the surface of sediment organic matter and biofilms is influenced by
both hydrophobic and electrostatic effects, resulting from the hydrophobicity of the
perfluorinated chain and the hydrophilicity of the carboxylate head groups (Higgins and Luthy
2006). In a series of laboratory studies, Higgins and Luthy (2006) demonstrated that PFOA
sorption to sediments increased with increasing organic content, increasing calcium ions, and
decreasing pH. The strongest effect was observed in response to increasing organic content,
demonstrating the importance of hydrophobic effects, while the increased sorption in response to
calcium ions and decreasing pH demonstrated the role of electrostatic effects (Higgins and Luthy
2006). Across all PFAS, sorption to sediments increased with increasing perfluorinated chain
length, and for a given chain length, PFAS such as PFOS, had approximately 1.7 times the
sorption capacity as perfluoroalkyl carboxylic acids (PFCA) such as PFOA (Higgins and Luthy
2006). The capacity of PFOA to bind to particulate matter increases with increasing salinity.
Jeon et al. (2010) observed that water column PFOA partitioned more readily to particulate
organic matter as salinity increased from 10 to 34 ppt, resulting in increased uptake of PFOA in
Pacific oysters (Crassostrea gigas). In a recent review, Li et al. (2018a) found no single
parameter strongly predicted PFOA sorption to sediments and Li et al. (2019) reported that the
protein content of soil was a better predictor of sorption than organic carbon. Overall, these
results suggest that sorption to sediments should be an important mechanism for PFOA entry into
an aquatic ecosystem.
31
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Evidence of the PFOA sediment pathway in aquatic ecosystems, although mixed, overall
demonstrates the importance of bioaccumulation from sediments and biofilms via diet into
aquatic invertebrates. In laboratory studies, PFOA concentrations in sediment were positively
correlated to PFOA tissue concentrations in Lumbriculus variegatus (Lasier et al. 2011), but not
for Chironomusplumosus (Wen et al. 2016) or the amphipods Gammarus fossarum and G. pulex
(Bertin et al. 2016). In field studies PFOA concentrations were positively correlated between
sediments and biofilms and benthic feeding organisms (Lescord et al. 2015; Loi et al. 2011;
Martin et al. 2004; Penland et al. 2020). In addition, the distribution of PFAS in sediments was
more similar to their distribution in the tissues of benthic invertebrates (Lescord et al. 2015) and
benthic-feeding fish (Thompson et al. 2011) than they were to their distribution in pelagic
organisms.
PFOA can also enter aquatic organisms directly from the water column through
respiration. Because of its binding affinity to proteins, PFOA can enter the body of gill-breathing
organisms by binding to proteins in the blood at gill surfaces (De Silva et al. 2009; Jones et al.
2003; Martin et al. 2003a, 2003b).
The relative distribution of PFOA in tissues is related to the primary route of exposure
(dietary or respiratory). In rainbow trout, the rank order of PFOA concentrations following
aqueous exposure was blood>kidney>liver (Martin et al. 2003a). In contrast, their rank order
following dietary exposure was liver>blood>kidney (Goeritz et al. 2013). Hong et al. (2015)
observed the highest concentrations of PFAS in the intestines of green eel goby; soft tissues,
shell, and legs of shore crabs; and gills and intestines of oysters, suggesting bioaccumulation
through both dietary and aqueous uptake in invertebrates, and primarily dietary uptake in fish.
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Data from multiple field studies suggest trophic biomagnification potential of PFOA is
low, and is often not observed, particularly with respect to aquatic organisms. In a review of
PFOA and PFOS concentration data across major taxonomic groups, Ahrens and Bundschuh
(2014) found that maximum PFOA and PFOS concentrations were similar for invertebrates, but
that maximum PFOS concentrations in fish were nearly an order of magnitude greater than
PFOA, and several orders of magnitude greater for aquatic-dependent birds and mammals.
When individual aquatic species pairs were considered, biomagnification factors (BMF)
greater than one, indicating biomagnification, have been observed for PFOA (e.g., Fang et al.
2014; Penland et al. 2020; Tomy et al. 2009), suggesting trophic biomagnification. However,
when ecosystem-level biomagnification is assessed using trophic biomagnification factors
(TMF), which measures the change in the concentration of a chemical per trophic level within a
food web, PFOA is nearly always shown not to biomagnify (Loi et al. 2011; Martin et al. 2004;
Tomy et al. 2004; Xu et al. 2014; Zhou et al. 2012), unless aquatic-dependent species, such as
aquatic-dependent birds, are included in the food web model (Houde et al. 2006a; Kelly et al.
2009; Tomy et al. 2009). The overall lack of biomagnification in PFOA relative to PFOS is
attributed to its physical-chemical properties, including a shorter perfluorinated chain length and
the carboxylate head group, both of which are associated with less efficient assimilation into
tissues and faster excretion rates (e.g., Martin et al. 2003a, 2003b).
2.5,3 Environmental Monitoring of PFOA in Biotic Media
Generally, PFOA is one of the dominant PFAS detected in aquatic ecosystems, along
with PFOS (Ahrens 2011; Benskin et al. 2012; Dinglasan-Panlilio et al. 2014; Nakayama et al.
2007; Remucal 2019; Zareitalabad et al. 2013). PFAS were first detected in human serum
samples in the late 1960s, and subsequent studies across several continents demonstrated the
global distribution of PFAS in humans (Giesy and Kannan 2001; Houde et al. 2006b). Since
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then, the global distribution of PFAS in tissues of aquatic and aquatic-dependent species has
been demonstrated in studies conducted in freshwater and marine environments across every
continent, including remote regions far from direct sources, such as the high arctic, Antarctica,
and oceanic islands (Giesy and Kannan 2001; Houde et al. 2006b).
In lentic surface waters of the United States, one of the most comprehensive studies of
PFOA concentrations included fish muscle tissue data from 157 near shore sites across the Great
Lakes selected following a probabilistic design as part of the 2010 National Coastal Condition
Assessment (Stahl et al. 2014). In this study, PFOA was measured in fish collected at 12% of the
sites, with a 90th centile concentration of 0.16 ng/g wet weight (ww), and a maximum
concentration of 0.97 ng/g ww (Stahl et al. 2014). Lake trout (31% of samples), smallmouth bass
(14%>), and walleye (13%>) were the most commonly sampled species from the Great Lakes
samples.
Martin et al. (2004) measured PFOA in whole body samples of invertebrates and fish in
Lake Ontario, near the town of Niagara-on-the-Lake. PFOA concentrations were much higher in
the benthic amphipod Diporeia hoya (90 ng/g ww) than in the more pelagic My sis relicta (2.5
ng/g ww), suggesting sediments are an important source of PFOA in this area (Martin et al.
2004). Among the four fish species sampled, PFOA concentrations were highest in the slimy
sculpin (44 ng/g ww), which feeds on M. relicta and D. hoya. Although lake trout occupy the
highest trophic level at this site, their PFOA concentrations were the lowest of all sampled fish
species (1.0 ng/g ww) (Martin et al. 2004). PFOA concentrations were lower in lake trout than in
alewife (1.6 ng/g ww), which comprise 90%> of the lake trout diet, suggesting a lack of PFOA
biomagnification in this system (Martin et al. 2004).
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Guo et al. (2012) measured PFOA in lake trout muscle tissues in Canadian waters of
Lakes Ontario, Erie, Huron, and Superior, as well as Lake Nipigon, Ontario. The average PFOA
concentration across all sites was 0.045 ng/g ww and was not significantly different (P<0.1)
across the different lakes (Guo et al. 2012). Finally, Delinsky et al. (2010) sampled bluegill,
black crappie, and pumpkinseed muscle tissues in 59 lakes in Minnesota, including four lakes in
the Minneapolis-St. Paul metropolitan area, and did not detect PFOA in any of the samples (limit
of quantification = 0.77 ng/g ww; see Table 2 of Delinsky et al. 2009).
In flowing surface waters of the United States, one of the most comprehensive studies of
PFOA concentrations included fish muscle tissue data from 164 urban river sites (5th order or
higher) across the coterminous U.S. selected following a probabilistic design as part of the 2008-
2009 National Rivers and Streams Assessment and the National Coastal Condition Assessment
(Stahl et al. 2014). Largemouth bass (34% of samples), smallmouth bass (25%), and channel
catfish (11%) were the most commonly sampled species from the urban stream sites (Stahl et al.
2014). PFOA was not detected in any of the urban river sites (Stahl et al. 2014). The lack of
detection may have been related to the method detection limit of 2.37 ng/g ww, which was
higher than the highest PFOA concentration measured in the Great Lakes coastal survey
described above, which also followed a probabilistic sampling design (Stahl et al. 2014).
In 2005, Ye et al. (2008) detected average PFOA concentrations of 0.17 ng/g ww and 0.2
ng/g ww from whole body composite samples of multiple fish species from the Ohio River and
Mississippi River, respectively. PFOA was not detected (<1.0 ng/g ww) in whole body
composite fish samples collected from the Missouri River (Ye et al. 2008). Delinsky et al. (2010)
sampled PFOA in bluegill, black crappie, and pumpkinseed muscle tissues at eleven locations
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along the upper Mississippi River in 2007, and did not detect it at any location, including the
heavily impacted Pool Two site in the Minneapolis-St. Paul metropolitan area.
In a more recent study, Penland et al. (2020) measured PFAS concentrations in
invertebrates and vertebrates along the Yadkin - Pee Dee River, in North and South Carolina in
2015. PFOA was detected in whole body tissues of unionid mussels (7.41 ng/g ww) and aquatic
insects (10.68 ng/g ww), but was not detected in Asian clam, snails, or crayfish. PFOA was
measured in muscle tissue of two of the 11 sampled fish species, the channel catfish (21.19 ng/g
ww) and notchlip redhorse (45.66 ng/g ww). PFOA was not detected in the eggs of a robust
redhorse sample, which had the highest measured PFOS concentration (482.9 ng/g ww) of any
sample from the Penland et al. (2020) study.
Houde et al. (2006a) measured whole body PFOA in six fish species in Charleston
Harbor, South Carolina, and whole body PFOA of zooplankton and five fish species in Sarasota
Bay, Florida. Charleston Harbor was the more developed of the two sites and had higher overall
PFOA concentrations. PFOA was detected in four of the six fish species in Charleston Harbor
and ranged from 0.5 ng/g ww in spot to 1.8 ng/g ww in spotted seatrout. In Sarasota Bay, PFOA
concentrations averaged 0.3 ng/g ww in zooplankton and was not detected in any of the fish
species (Houde et al. 2006a).
Overall, these results illustrate the distribution of PFOA in biotic media collected from
invertebrate and fish samples. In contrast to PFOS, PFOA concentrations in biotic media are
often low, or below detection levels, highlighting the lower overall bioaccumulation potential for
this chemical, based on its physical-chemical properties, including a shorter perfluorinated chain
length, and a carboxylate head group. In addition, trophic biomagnification is rarely observed
with PFOA, as concentrations in invertebrates are often similar to concentrations in fish.
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2.6 Exposure Pathways of PFOA in Aquatic Environments
There are multiple potential exposure pathways of PFOA in the aquatic environment,
including: (1) direct aqueous (dermal and respiratory) exposure, (2) direct exposure from
contaminated sediment (for benthic organisms), (3) diet (e.g., bioaccumulation and
biomagnification), and (4) maternal transfer (Ankley et al. 2020). Exposure of PFOA through
water and sediment occurs through direct contact with the respective media, such as water
passing across the gills, or consumption of suspended and deposited sediments (Prosser et al.
2016). Elevated PFOA concentrations in eggs of fish and piscivorous birds suggests that PFOA
may maternally transfer to offspring. Given these exposure pathways, aquatic organisms, such as
fish and aquatic invertebrates, are exposed to PFOA when it is present in the environment. This
exposure occurs through multiple exposure routes including water, sediment, diet, and maternal
transfer.
2.7 Effects of PFOA on Biota
Currently, PFOA aquatic ecotoxicity data are primarily available for freshwater fish,
aquatic invertebrates, plants, and algae. Section 3 and Section 4 provide study summaries of
individual publicly available ecotoxicity studies, Appendix A through Appendix F summarize
the current quantitatively acceptable PFOA aquatic life ecotoxicity data, and Appendix G and
Appendix H list current qualitatively acceptable PFOA aquatic life ecotoxicity data.
2.7.1 Mechani sm s of PF OA Toxi city
The mechanisms underpinning the toxicity of PFOA to aquatic organisms, like other
PFAS, is an active and on-going area of research. Much work is still needed from a mechanistic
perspective to better understand how the different modes of action elicit specific biological
responses. Molecular disturbance at the cellular- and organ-level resulting in effects on
reproduction, growth and development at the individual-level are associated with the sex-related
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endocrine system, thyroid-related endocrine system, and neuronal-, lipid-, and carbohydrate-
metabolic systems (see Ankley et al. 2020 and Lee et al. 2020 for the latest reviews on the
subject). The underlying mechanisms of PFOA toxicity to aquatic animals, and fish in particular,
appear to be related to oxidative stress, apoptosis, thyroid disruption, and development-related
gene expression (Lee et al. 2020). The published research suggests that many of these molecular
pathways interact with each other and could be linked. For example, for several PFAS including
PFOA, oxidative stress appears correlated with effects on egg hatching and larval formation,
linking reproductive toxicity, oxidative stress, and developmental toxicity (Lee et al. 2020). The
actual mechanism(s) through which PFAS induce oxidative stress require additional study, but
increased B-oxidation of fatty acids and mitochondrial toxicity are proposed triggers (Ankley et
al. 2020).
Of particular importance to this document is that PFOA exposure-related disruption of
the sex-related endocrine system (e.g., androgen and estrogen) at the molecular, tissue and organ
levels appears to have adverse reproductive outcomes in fish and invertebrates, and likely in both
freshwater and saltwater and via multiple exposure routes, i.e., waterborne and dietary (Lee at al.
2020). The reproductive effects were observed in the Fo, Fi and F2 generations of zebrafish,
Danio rerio, in the multi-generational PFOA exposure reported by Lee et al. (2017).
It is clear that PFOA, and many other PFAS, cause a wide range of adverse effects in
aquatic organisms, including: reproductive failure, developmental toxicity; androgen, estrogen
and thyroid hormone disruption; immune system disruption; and, neuronal and developmental
damage. Study of the systematic interactions among the relevant biological pathways in fish is a
research need, as well as a better understanding of several knowledge gaps in non-fish aquatic
organisms where mechanistic-based investigations need to be prioritized.
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2,7,2 Potential Interactions with Other PFAS
PFAS may occur as mixtures in the environment. Occurrence studies document the
presence of complex mixtures of PFAS in surface waters in the U.S. and across the globe
(Ahrens 2011; Ahrens and Bundschuh 2014; Giesy and Kannan 2002; Houde et al. 2006a,
2006b; Keiter et al. 2012; Wang et al. 2017; see Section 2.4.1). Although the EPA's PFOA
criteria are based solely on single chemical exposure aquatic toxicity tests, it is recognized that
PFAS are often introduced into the environment as end-use formulations comprised of mixtures
of PFAS and or PFAS-precursors, the ecological effects of which are poorly understood (Ankley
et al. 2020). It is useful, therefore, to briefly summarize the types of interactions that might be
expected based on the few PFAS mixtures studies involving PFOA and one or more PFAS to
date. Note that for purposes of this document, the reader is referred to Ankley et al. (2020) and
elsewhere for more comprehensive reviews of PFAS mixtures in general, and the challenges they
are expected to present in ecological risk assessment. Beyond PFOA and PFOS, systematic
reviews of chemical mixture studies across various compound classes indicate that departures
from dose additivity are uncommon and rarely exceed minor deviations (~2-fold) from
predictions based on additivity (Martin et al. 2021).
Findings of the PF AS-specific studies described below are as reported by the study
authors without any additional interpretation or analysis of uncertainty. At both the organismal
and cellular levels, studies on zebrafish (Danio rerio; Ding et al. 2013), a water flea (Daphnia
magna; Yang et al. 2019), a bioluminescent cyanobacterium (Anabaena sp.\ Rodea-Palomares et
al. 2012), or with cultured hepatocytes of the cyprinid, Gobiocypris rarus (Wei et al. 2009),
demonstrate that the effects observed from in vivo and in vitro tests on PFAS mixtures vary.
PFAS mixture studies on zebrafish reported interactions for combinations of PFOA and PFOS,
but departures from additive models were also minor (Ding et al. 2013). Menger et al. (2020)
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reported zebrafish behavioral effects from a PFAS mixture that were less than individual PFAS,
however evaluation of chemical dose response and comparison to mixture models was not
conducted. Yang et al. (2019) exposed the water flea, Daphnia magna, to single and binary
mixtures of PFOA and PFOS. The authors reported synergism in acute and chronic toxic effects.
Conversely, Rodea-Palomares et al. (2012) showed binary PFOA and PFOS mixture as having
an antagonistic interaction at the whole range of effect levels tested using the bioluminescent
cyanobacterium, Anabaena.
In tests with cultured hepatocytes of the cyprinid G. rarus, co-exposure of PFOA with a
mixture of five other PFAS [PFNA, PFDA, PFDoA, PFOS, 8:2 FTOH] altered genes involved in
multiple biological functions and processes, including fatty acid metabolism and transport,
xenobiotic metabolism, immune response, and oxidative stress. Additionally, greater than 80%
of the altered genes in both the PFOA- and PFOS-dominant mixture groups were of the same
gene set (Wei et al. 2009). Finally, Conley et. al. (2022) observed PFOA and PFOS interacting in
an additive manner to reduce pup body weight, pup liver weight, and maternal liver weight in the
Sprague-Dawley rat.
2.8 Conceptual Model of PFOA in the Aquatic Environment and Effects
A conceptual model depicts the relationship between a chemical stressor and ecological
compartments, linking exposure characteristics to ecological endpoints. The conceptual model
provided in Figure 2-4 summarizes sources, potential pathways of PFOA exposure for aquatic
life and aquatic-dependent wildlife and possible toxicological effects.
PFOA initially enters the aquatic environment through direct discharge from wastewater
treatment facilities, atmospheric deposition, and runoff from contaminant surfaces such as PFAS
disposal sites or contaminated biosolids. PFOA enters the aquatic environment primarily in the
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dissolved form and to a lesser extent, particle-bound forms. Exposure pathways for the biological
receptors of concern (i.e., aquatic organisms) and potential effects (e.g., impaired survival,
growth, and reproduction) in those receptors are represented in the conceptual model (Figure
2-4). Both direct (i.e., exposure from the water column which is represented by *) and indirect
(i.e., bioconcentrated by producers and bioaccumulated by consumers in higher trophic levels
represented by **) pathways are represented in the conceptual model.
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6C
a
U
PFOA Source
Point Sources
(from municipal and industrial
dischargers from applications
such as surfactants, textile stain
and soil repellents )
PFOA in Water
Dissolved & Particle-Bound
Degradation &
Metabolism of
Other PFASs
ฆ4
Aquatic Life
PFOA in Sediment
PFOA Source
Nonpoint Sources
(from landfill leachate, land
application of biosolids)
Producers
1" Trophic Transfer
(from phytoplankton, periphyton, macrophytes; e.g., algae, cyanobacteria, waterweed/common eelgrass)
-
o
P.
u
tu
Pi
Consumers
2"d Tr0phjc Transfer
(to zooplankton, macroinvertebrates;
e.g., cladocerans/copepods &
mayflies/ribbed mussels)
Consumers
3rd Trophic Transfer
(to predatory fish:
e.g., longnose dace/
American shad)
>
Consumers
4th Trophic Transfer
(to predatory fish;
e.g., largemouth bass/
striped bass)
Figure 2-4. Conceptual Model Diagram of Sources, Compartmental Partitioning, and
Trophic Transfer Pathways of Perfluorooctanoic acid (PFOA) in the Aquatic Environment
and its Bioaccumulation and Effects in Aquatic Life and Aquatic-dependent Wildlife.
PFOA sources represented in ovals, compartments within the aquatic ecosystem represented by rectangles, and
effects (on trophic levels of aquatic-dependent wildlife, represented by shaded box) in pentagons. Examples of
organisms in each trophic transfer provided as freshwater/marine. Movement of PFOA from water to receptors
indicated by two separate pathways: direct exposure to all trophic levels within box (*) and bioconcentration by
producers (**). Relative proportion of PFOA transferred between each trophic level is dependent on life history
characteristics of each organism.
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2.9 Assessment Endpoints
Assessment endpoints are defined as the explicit expressions of the environmental values
to be protected and are comprised of both the ecological entity (e.g., a species, community, or
other entity) and the attributes or characteristics of the entity to be protected (U.S. EPA 1998).
Assessment endpoints may be identified at any level of organization (e.g., individual, population,
community). In context of the CWA, aquatic life criteria for toxic substances are typically
determined based on the results of toxicity tests with aquatic organisms, for which adverse
effects on growth, reproduction, or survival are measured. This information is typically compiled
into a sensitivity distribution based on genera and representing the impact on taxa across the
aquatic community. Criteria are based on the 5th percentile of genera and are, thus intended to be
protective of approximately 95 percent of aquatic genera to ensure aquatic communities are
protected. Assessment endpoints consistent with the criteria developed in this document are
summarized in Table 2-1.
The use of laboratory toxicity tests to protect bodies of water and resident aquatic species
was based on the theory that effects occurring to a species in appropriate laboratory tests will
generally occur to the same species in comparable field situations. Since aquatic ecosystems are
complex and diverse, the 1985 Guidelines recommend that acceptable data be available for at
least eight genera with a specified taxonomic diversity (the standard eight minimum data
requirements, or MDRs). The intent of the eight MDRs is to serve as a typical surrogate sample
community representative of the larger and generally much more diverse natural aquatic
community, not necessarily the most sensitive species in a given environment. The 1985
Guidelines note that since aquatic ecosystems can tolerate some stress and occasional adverse
effects, protection of all species at all times and places are not deemed necessary (the intent is to
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protect 95 percent of a group of diverse taxa, and any commercially and recreationally important
species).
2.10 Measurement Endpoints
2,10,1 Overview of Toxicity Data Requirements
To ensure the protection of various components of an aquatic ecosystem, the EPA
collects acute toxicity test data from a minimum of eight diverse taxonomic groups.
Acute freshwater criteria require data from the following eight taxonomic groups:
a) the family Salmonidae in the class Osteichthyes
b) a second family in the class Osteichthyes, preferably a commercially or
recreationally important warmwater species (e.g., bluegill, channel catfish)
c) a third family in the phylum Chordata (may be in the class Osteichthyes or may
be an amphibian)
d) a planktonic crustacean (e.g., cladoceran, copepod)
e) a benthic crustacean (e.g., ostracod, isopod, amphipod, crayfish)
f) an insect (e.g., mayfly, dragonfly, damselfly, stonefly, caddisfly, mosquito,
midge)
g) a family in a phylum other than Arthropoda or Chordata (e.g., Rotifera, Annelida,
Mollusca)
h) a family in any order of insect or any phylum not already represented
Acute estuarine/marine criteria require data from the following taxonomic groups:
a) two families in the phylum Chordata
b) a family in a phylum other than Arthropoda or Chordata
c) a family from either Mysidae or Penaeidae
d) three other families not in the phylum Chordata (may include Mysidae or
Penaeidae, whichever was not used above)
e) any other family
Additionally, to ensure the protection of various components of the aquatic ecosystem
from long term exposures, chronic toxicity test data are recommended for the same minimum of
eight diverse taxonomic groups that are recommended for freshwater acute criterion derivation.
If the eight diverse taxonomic groups are not available to support chronic criterion derivation
using a genus distribution approach, the chronic criterion may be derived using an acute-to-
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chronic ratio (ACR) approach. To apply an ACR approach to derive a chronic freshwater
criterion a minimum of three taxa are recommended, with at least one chronic test being from an
acutely sensitive species. Acute-to-chronic ratios (ACRs) can be calculated with data for aquatic
organisms.
Chronic aquatic life criteria require data from the following taxonomic groups:
a) At least one is a fish
b) At least one is an invertebrate
c) At least one is an acutely sensitive freshwater species, for freshwater chronic
criterion (the other two may be saltwater species)
d) At least one is acutely sensitive saltwater species for estuarine/marine chronic
criterion (the other two may be freshwater species)
The 1985 Guidelines also specified at least one quantitative test with a freshwater alga or
vascular plant. If plants are among the most sensitive aquatic organisms, toxicity test data from a
plant in another phylum should also be available. Aquatic plant toxicity data are examined to
determine whether aquatic plants are likely to be adversely affected by the concentration
expected to be protective for other aquatic organisms.
2,10,2 Measure of PFOA Exposure Concentrations
These PFOA ambient water quality criteria are for the protection of aquatic life. This
criteria document provides a critical review of all data identified in the EPA's literature search
for PFOA, including:
the anionic form (CAS No. 45285-51-6),
the acid form (CAS No. 335-67-1), and;
the ammonium salt (CAS No. 3825-26-1).
PFOA toxicity studies typically do not conduct an analysis or separation sufficient to
determine if the test compound is purely linear or branched. Data for possible inclusion in the
PFOA criteria were obtained from published literature reporting acute and chronic exposures of
PFOA that were associated with mortality, growth, and reproduction. This set of published
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literature was identified using the ECOTOXicology database (ECOTOX;
https://cfpub.epa.gov/ecotox/) as meeting data quality standards. ECOTOX is a source of high-
quality toxicity data for aquatic life, terrestrial plants, and wildlife. The database was created and
is maintained by the EPA, Office of Research and Development, Center for Computational
Toxicology and Exposure. The ECOTOX search generally begins with a comprehensive
chemical-specific literature search of the open literature conducted according to ECOTOX
Standard Operating Procedures (SOPs; Elonen 2020). The search terms are often comprised of
chemical terms, synonyms, degradates and verified Chemical Abstracts Service (CAS) numbers.
After developing the literature search strategy, ECOTOX curators conduct a series of searches,
identify potentially applicable studies based on title and abstract, acquire potentially applicable
studies, and then apply the applicability criteria for inclusion in ECOTOX. Applicability criteria
for inclusion into ECOTOX generally include:
1. The toxic effects are related to single chemical exposure (unless the study is being
considered as part of a mixture effects assessment);
2. There is a biological effect on live, whole organisms or in vitro preparation including
gene chips or omics data on adverse outcome pathways potentially of interest;
3. Chemical test concentrations are reported;
4. There is an explicit duration of exposure;
5. Toxicology information that is relevant to OW is reported for the chemical of concern;
6. The paper is published in the English language;
7. The paper is available as a full article (not an abstract);
8. The paper is publicly available;
9. The paper is the primary source of the data;
10. A calculated endpoint is reported or can be calculated using reported or available
information;
11. Treatment(s) are compared to an acceptable control;
12. The location of the study (e.g., laboratory vs. field) is reported; and
13. The tested species is reported (with recognized nomenclature).
Following inclusion in the ECOTOX database, toxicity studies were subsequently
evaluated by the Office of Water. All studies were evaluated for data quality as described by
U.S. EPA (1985), the EPA's Office of Chemical Safety and Pollution Prevention (OPP)'s
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Ecological Effects Test Guidelines (U.S. EPA 2016b), and the EPA OW's internal data quality
standard operating procedure (SOP), which is consistent with OPP's data quality review
approach (U.S. EPA 2016b). Office of Water completed a Data Evaluation Record (DER) for
each species by chemical combination from the PFOA studies identified by ECOTOX. This in-
depth review ensured the studies used to derive the criteria resulted in robust scientifically
defensible criteria. Example DERs are shown in Appendix P with the intent to convey the
meticulous level of evaluation, review, and documentation each PFOA study identified by
ECOTOX was subject to.
Studies that did not fully meet the data quality objectives outlined by the EPA SOP were
not considered for inclusion in the criteria derivation, including some studies with other PFAS
exposures, but were considered qualitatively as supporting information and are characterized in
the Effects Characterization. These studies are listed in Appendix G and Appendix H.
Furthermore, only single chemical toxicity tests with PFOA were considered for possible
inclusion in criteria derivation, studies that tested chemical mixtures, including mixtures with
PFAS were excluded from criteria derivation. Both controlled laboratory experiments and field
observations/studies were included.
The 1985 Guidelines recommend only toxicity tests focused on North American resident
species be considered. Due to the EPA's interest in using all available quality data, particularly
for a data-sparse chemical like PFOA (relative to chemicals such as cadmium or ammonia),
toxicity studies were considered for possible inclusion regardless of the test species residential
status in North America. Use of non-North American residential species is also consistent with
other published aquatic life criteria (U.S. EPA 2018). Non-North American resident species also
serve as taxonomically-related surrogate test organisms for the thousands of untested resident
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species. Supporting analyses to evaluate the influence of including non-resident species on the
freshwater criteria magnitudes were conducted by limiting toxicity datasets to North American
resident species with established populations in North America (see Section 4.1). These analyses
provided an additional line-of-evidence that supports inclusion of non-resident species in PFOA
criteria derivation.
Toxicity tests used in many previous aquatic life criteria documents are typically based
on measured chemical concentrations only. For PFOA, the EPA has examined the issue of
whether nominal (unmeasured) and measured concentrations are in close agreement with each
other (Jarvis et al. 2023). While measured PFOA toxicity tests are generally preferred, results of
Jarvis et al. (2023) demonstrated that experimental conditions had little influence on observed
discrepancies between nominal and measured concentrations for PFOA, with the exception of
freshwater studies that contained substrate. Nominal and measured concentrations in the analysis
generally displayed a high degree of linear correlation (>0.98 freshwater, >0.84 saltwater) and
relatively low median percent differences (Jarvis et al. 2023). In freshwater tests, when tests with
substrate were removed, 89% of the 527 PFOA and PFOS measured concentrations were within
20% of their nominal counterparts (Note: the EPA's OCSPP's Ecological Effects Test
Guidelines (2016b) consider tests acceptable when measured concentrations are within 20% of
nominal, and Rewerts et al. (2021) suggested that PFAS-specific toxicity tests may even be
acceptable if measured and nominal concentrations do not differ by up to 30%.). Of the observed
disparities between measured and nominal concentrations in the PFOA freshwater data sets,
those with substrate (McCarthy et al. 2021; Oakes et al. 2004) were concluded to be the primary
contributor to observed differences with measured PFOA concentrations systematically lower
than corresponding nominal concentrations, indicating that added PFOA may have sorbed to
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substrate, reducing PFOA in the water column (Jarvis et al. 2023). PFOA concentrations in test
waters are expected to remain relatively constant over the course of acute and chronic exposures
given its ability to resist breakdown and transformation (Ahrens 2011). Since toxicity tests with
substrate and nominal concentrations only were not used quantitatively in PFOA criteria
derivation, the EPA determined nominal test concentrations adequately represent actual PFOA
exposures in standard acute and chronic laboratory-based toxicity tests. Consequently, PFOA
toxicity tests were not excluded from quantitative use in criteria derivation on the basis of
unmeasured test concentrations alone.
Typically, per the 1985 Guidelines acute toxicity data from all measured flow-through
tests would be used to calculate species mean acute values (SMAVs), unless data from a
measured flow-through test were unavailable, in which case the acute criterion would be
calculated as the geometric mean of all the available acute values (i.e., results of unmeasured
flow-through tests and results of measured and unmeasured static and renewal tests). Chronic
unmeasured flow-through tests, as well as measured and unmeasured static and renewal tests are
not typically considered to calculate chronic values. In the case of the PFOA, static, renewal, and
flow-through experiments were considered for possible inclusion for both species mean acute
and chronic values regardless of whether PFOA concentrations were measured because PFOA is
a highly stable compound, resistant to hydrolysis, photolysis, volatilization, and biodegradation
(Section 1.2.1) and, therefore, expected to vary only minimally in the course of a toxicity test.
Additionally, chronic values were based on endpoints and exposure durations that were
appropriate to the species. Thus, both life- and partial life-cycle tests were utilized for the
derivation of the chronic criterion. However, it should be noted that the 1985 Guidelines specify
life-cycle chronic tests are typically used for invertebrates. The chronic studies used in the
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derivation of the chronic water column-based PFOA criterion followed taxon-specific exposure
duration requirements from various test guidelines (i.e., the EPA's 1985 Guidelines and the
EPA's OCSPP's Ecological Effects Test Guidelines) when available. For example, only chronic
daphnid studies of 21 days were considered in the chronic criterion derivation because the EPA
1985 Guidelines states daphnid tests should begin with young < 24-hours old and last at least 21
days. When taxon-specific exposure duration requirements were not available for a particular test
organism in the PFOA toxicity literature, both life- and partial life-cycle tests were considered in
the derivation of the chronic criterion.
PFOA toxicity in aquatic life is manifested as effects on survival, growth, and
reproduction. Measurements of fish tissue may be linked to the chronic adverse effects of PFOA,
since PFOA is highly persistent and potentially bioaccumulative.
2.10.3 Measures of Effect
Each assessment endpoint requires one or more "measures of ecological effect," which
are defined as changes in the attributes of an assessment endpoint itself or changes in a surrogate
entity or attribute in response to chemical exposure. Ecological effects data were used as
measures of direct and indirect effects to growth, reproduction, and survival of aquatic
organisms.
2.10.3.1 Acute Measures of Effect
The acute measures of effect on aquatic organisms are the lethal concentration (LCso),
effect concentration (ECso), or inhibitory concentration (IC50) estimated to produce a specific
effect in 50 percent of the test organisms. LC50 is the concentration of a chemical that is
estimated to kill 50 percent of the test organisms. EC50 is the concentration of a chemical that is
estimated to produce a specific effect in 50 percent of the test organisms. And the IC50 is the
concentration of a chemical that is estimated to inhibit some biological process (e.g., enzyme
50
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inhibition associated with an apical endpoint such as mortality) in 50 percent of the test
organisms.
2.10.3,2 Chronic Measures of Effect
The endpoint for chronic exposures is the effect concentration estimated to produce a
chronic effect on survival, growth, or reproduction in 10 percent of the test organisms (ECio).
The EPA selected an ECio to estimate a low level of effect that would be both different from
controls and not expected to be severe enough to cause effects at the population level for a
potentially bioaccumulative contaminant, such as PFOA. The use of the ECio, instead of an EC20,
is also consistent with the use of this metric for the bioaccumulative pollutant selenium in the
recent 2016 Selenium Freshwater Aquatic Life Criteria (U.S. EPA 2016a). Use of a 10% effect
concentration for deriving chronic criteria magnitudes is also consistent with the harmonized
guidelines from OECD and the generally preferred effect level for countries such as Canada,
Australia, and New Zealand (CCME 2007; Warne et al. 2018).
Regression analysis was used preferentially to characterize a concentration-effect
relationship and to estimate concentrations at which chronic effects are expected to occur (i.e.,
point estimate). Reported No Observed Effect Concentrations (NOECs) and Lowest Observed
Effect Concentrations (LOECs) were only used for the derivation of a chronic criterion when a
robust ECio could not be calculated for the genus. A NOEC is the highest test concentration at
which none of the observed effects are statistically different from the control. A LOEC is the
lowest test concentration at which the observed effects are statistically different from the control.
When LOECs and NOECs were used, a Maximum Acceptable Toxicant Concentration (MATC)
was calculated, which is the geometric mean of the NOEC and LOEC. For the calculation of a
chronic criterion, point estimates were selected for use as the measure of effect in favor of
MATCs, as MATCs are highly dependent on the concentrations tested. Point estimates also
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provide additional information that is difficult to determine with an MATC, such as a measure of
effect level across a range of tested concentrations. A decision rule was also applied to the PFOA
toxicity data when an author-reported NOEC or LOEC was used in conformity with the 2013
Ammonia Freshwater Aquatic Life Criteria (U.S. EPA 2013) such that "greater than" values for
concentrations of a relatively low magnitude compared to the other available toxicity data, and
"less than" values for concentrations of relatively high magnitude were considered to add little
significant information to the analyses and were not used quantitatively. Conversely, if data from
studies with relatively low "less than" values indicated a significant effect or studies with
relatively high "greater than" values only found an incomplete response for a chronic endpoint
(indicating low toxicity of the test material), those data significantly enhanced the understanding
of PFOA toxicity. Thus, the decision rule was applied as follows: "greater than" (>) high toxicity
values and "less than" (<) low toxicity values were used quantitatively to derive the chronic
water column-based PFOA criterion (U.S. EPA 2013). Data that met the quality objectives and
test requirements were utilized quantitatively in deriving freshwater criteria for aquatic life and
are presented in Table 3-3 and Table 3-7.
Table 2-1. Summary of Assessment Endpoints and Measures of Effect Used in the Criteria
Derivation for PFOA.
Assessment Kndpoinls lor (lie Aquatic
(ommunilY
Measures of K fleet
Aquatic Life: Survival, growth, and
reproduction of freshwater and
estuarine/marine aquatic life (i.e., fish,
amphibians, aquatic invertebrates)
For effects from acute exposure:
1. LCso, EC50, or IC50 concentrations in water
For effects from chronic exposure:
1. EC 10 concentrations in water
2. NOEC and LOEC concentrations in water;
Only used when an EC10 could not be calculated
for a genus.
NOEC = No observed effect concentration
LOEC = Lowest observed effect concentration
ECio = 10% Effect Concentration
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2.10.3.3 Summary of Independent Calculation of Toxicity Values
Toxicity values, including LC50 and EC 10 values, were independently calculated from the
data presented in the toxicity studies meeting the inclusion criteria described above when
adequate concentrations-response data were published in the study or could be obtained from
authors. When concentration-response data were not presented in toxicity studies, concentration-
response data were requested from study authors to independently calculate toxicity values. In
cases where study authors did not respond to the EPA's request for data or were unable to locate
concentration-response data, the toxicity values were not independently calculated by the EPA,
and the reported toxicity values were retained for criteria deviation. Where concentration-
response data were available, they were analyzed using the statistical software program R
(version 3.6.2) and the associated dose-response curve (drc) package. The R drc package has
various models available for modeling a concentration-response relationship for each toxicity
study. The specific model used to calculate toxicity values was selected following the details
provided in Appendix K, and the models performed well on most or all statistical metrics. The
independently calculated toxicity values used to derive the PFOA aquatic life criteria were
included in each study summary below and were used to derive criteria for aquatic life, where
available. Details relating to the independent verification of toxicity values for each toxicity
study used to derive the criteria were included in Appendix A.2 and Appendix C.2.
2.11 Analysis Plan
2.11.1 Derivation of Water Column Criteria
During CWA Section 304(a) criteria development, the EPA reviews and considers all
relevant toxicity test data. Information available for all relevant species and genera were
reviewed to identify: 1) data from acceptable tests that meet data quality standards, and 2)
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whether the acceptable data meet the MDRs as outlined in the EPA's 1985 Guidelines (U.S. EPA
1985). The taxa represented by the different MDR groups represent taxa with different
ecological, trophic, taxonomic and functional characteristics in aquatic ecosystems, and are
intended to be a representative subset of the diversity within a typical aquatic community. MDRs
for derivation of acute and chronic freshwater criteria were met for all the eight MDRs.
Acute and chronic MDRs for PFOA estuarine/marine criteria derivation were not met
and, consequently, acute and chronic estuarine/marine criteria were not derived. The EPA is,
however, including an acute aquatic life benchmark for estuarine/marine environments (see
Appendix L), using available estuarine/marine species toxicity data and application of ORD's
peer-reviewed web-ICE tool. A minimal number of tests from acceptable studies of aquatic algae
and vascular plants were also available for possible derivation of a Final Plant Value. However,
the relative sensitivity of freshwater plants to PFOA exposures indicated plants are less sensitive
than aquatic vertebrates and invertebrates so plant criteria were not developed.
2.11.2 Derivation of Tissue-Based Criteria
Chronic toxicity studies (both laboratory and field studies) were further screened to
ensure they contained the relevant chronic PFOA exposure conditions to aquatic organisms (i.e.,
dietary, or dietary and waterborne PFOA exposure), measurement of chronic effects, and
measurement of PFOA in tissue(s). The EPA considered deriving tissue-based criteria using
empirical toxicity tests with studies that exposed organisms to PFOA in water and/or diet and
reported exposure concentrations based on measured tissue concentrations. This approach would
also correspond with the 2016 Selenium Aquatic Life Freshwater Criterion, which is the only
304(a) aquatic life criterion with tissue-based criterion elements. However, the freshwater
chronic PFOA toxicity data with measured tissue concentrations were limited, with no
quantitatively acceptable tissue-based tests. Qualitatively acceptable tissue-based tests were
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reported for four species (three fish species and one amphibian) across five publications.
Therefore, there were insufficient data to derive tissue-based criteria using a GSD approach from
empirical tissue data from toxicity studies. The EPA thus developed protective tissue-based
criteria through a bioaccumulation factor approach (Burkhard 2021).
2.11.3 Translation of Chronic Water Column Criterion to Tissue Criteria
Because there were insufficient chronic toxicity data with measured tissue concentrations
to derive chronic PFOA tissue criteria using a GSD approach, the EPA derived PFOA chronic
tissue-based criteria by translating the chronic freshwater water column criterion (see Section
3.2.1.3) into tissue-based criteria magnitudes using bioaccumulation factors and the following
equation:
Tissue Criteria = Chronic Water Column Criterion x BAF (Equation 1)
The resulting tissue-based criteria magnitudes correspond to the tissue type from the BAF used
in the equation (see Section 2.11.3.1).
2,11,3.1 Aquatic Life Bioaccumulation Factors
A bioaccumulation factor (BAF) is determined from field measurements and is calculated
using the equation:
BAF = Cbl0ta (Equation 2)
Cwater
Where:
Cuota = PFOA concentration in the organismal tissue (s)
Cwater = PFOA concentration in water where the organism was collected
The EPA considered BAF data from field measurements to capture all PFOA exposure
routes, i.e., dietary, water, contact with sediments via dermal exposure and ingestion, and
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maternal transfer. Depending on the tissue residue measurement, BAFs can be based upon
residues in the whole organism, muscle, liver, or any other tissue.
Searching for literature reporting on PFOA was implemented by developing a series of
chemical-based search terms. These terms included chemical names and Chemical Abstracts
Service registry numbers (CASRN or CAS), synonyms, tradenames, and other relevant chemical
forms (i.e., related compounds). Databases searched were Current Contents, ProQuest CSA,
Dissertation Abstracts, Science Direct, Agricola, TOXNET, and UNIFY (database internal to the
U.S. EPA's ECOTOX database). The literature search yielded numerous citations and the
citation list was further refined by excluding citations on analytical methods, human health,
terrestrial organisms, bacteria, and where PFOA was not a chemical of study. The citations
meeting the search criteria were reviewed for reported BAFs and/or reported concentrations in
which BAFs could be calculated for freshwater and estuarine/marine species. BAFs from both
freshwater and estuarine/marine species were considered because: (1) inclusion of
estuarine/marine BAFs expanded the relatively limited PFOA BAF dataset, and (2) Burkhard
(2021) did not specifically observe notable differences in PFAS BAFs between freshwater and
estuarine/marine systems, instead stating additional research is needed to formulate conclusions.
Data from papers with appropriate BAF information were further screened for data
quality. Four factors were evaluated in the screening of the BAF literature: (1) number of water
samples, (2) number of organism samples, (3) water and organism temporal coordination in
sample collection, and (4) water and organism spatial coordination in sample collection.
Additionally, the general experimental design was evaluated. Table 2-2 below outlines the
screening criteria for study evaluation and ranking. Only BAFs of high and medium quality were
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used to derive the tissue criteria (Appendix O). For further details on BAFs compilation and
ranking, see Burkhard (2021).
Table 2-2. Evaluation Criteria for Screening Bioaccumulation Factors (BAFs) in the Public
Literature.
Screening l-'actor
High Quality
Medium Quality
Low Quality
Number of Water Samples
>3
2-3
1
Number of Organism Samples
>3
2-3
1
Temporal Coordination
Concurrent
collection
Within one year
Collection period >1 year
Spatial Coordination
Collocated
collection
Within 1-2 km
Significantly different
locations
(>2 km)
General Experimental Design
Mixed species tissues
samples
Modified from Burkhard (2021).
PFOA bioaccumulation potential in aquatic life is expected to be relatively low compared
to PFOS bioaccumulation, or PFOA bioaccumulation in aquatic-dependent birds and mammals
The high water solubility of PFOA may allow aquatic organisms to excrete PFOA through the
gills. Conversely, aquatic dependent birds and mammals lack such an excretion mechanism, as
the low vapor pressure of PFOA limits its ability to be transferred across the alveolar membrane
from lungs to air (Canada 2012; Kelly et al. 2004). The tissue criteria for fish and invertebrates
presented here protect aquatic life populations from PFOA exposure accumulating in tissues and
provide U.S. states and Tribes greater context to their fish-tissue monitoring programs, which are
actively measuring PFAS, including PFOA.
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3 EFFECTS ANALYSIS FOR AQUATIC LIFE
3.1 Toxicity to Aquatic Life
All available studies relating to the acute and chronic toxicological effects of PFOA on
aquatic life were considered in the derivation of these national recommended PFOA criteria.
Data for possible inclusion in these PFOA criteria were obtained from published literature
reporting acute and chronic exposures of PFOA that were associated with mortality, survival,
growth, and reproduction. The latest search was conducted through the March 2024 ECOTOX
database update. Acute and chronic data meeting the quality objectives and test requirements
were utilized quantitatively in deriving these criteria for aquatic life and are presented in
Appendix A: Acceptable Freshwater Acute PFOA Toxicity Studies; Appendix B: Acceptable
Estuarine/Marine Acute PFOA Toxicity Studies; Appendix C: Acceptable Freshwater Chronic
PFOA Toxicity Studies, and; Appendix D: Acceptable Estuarine/Marine Chronic PFOA Toxicity
Studies.
3.1.1 Summary of PFOA Toxicity Studies Used to Derive the Aquatic Life Criteria
Quantitatively acceptable acute PFOA toxicity data were available for 27 freshwater
species, representing 19 genera and 17 families in five phyla, and four estuarine/marine species,
representing four genera and three families in three phyla (Table 3-1). Quantitatively acceptable
chronic PFOA toxicity data were available for 13 freshwater species, representing 12 genera and
ten families in three phyla and two estuarine/marine species, representing two genera and two
families in two phyla. The following study summaries present the key acute and chronic
freshwater toxicity data with effect values that were used quantitatively to derive the acute and
chronic freshwater and estuarine/marine criteria to protect aquatic life. Study summaries for the
most sensitive taxa are presented below and are grouped by acute or chronic exposure and sorted
by sensitivity to PFOA.
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Acute and chronic values were presented as reported by the study authors for each
individual study, unless stated otherwise. The EPA independently calculated these toxicity
values if sufficient raw data were available for the EPA to conduct statistical analyses. The
EPA's independently-calculated toxicity values were used preferentially, where available.
Author-reported toxicity values and the EPA's independently calculated values (where available)
were included in each study summary and in appendices, as applicable. The results of all toxicity
values, such as LC values, EC values, NOECs, LOECs, and species- and genus-mean values, are
given to four significant figures to prevent round-off error in subsequent calculations, not to
reflect the precision of the value. The specific toxicity value utilized in the derivation of the
corresponding PFOA criteria is stated for each study at the end of the summaries below and in
the respective appendices.
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Table 3-1. Summary Table of Minimum Data Requirements per the 1985 Guidelines
Reflecting the Number of Acute and Chronic Genus and Species Level Mean Values in the
Freshwater and Saltwater Toxicity Dataset
ts for PFOA.
MDU11 h
l-'reshwater
GMAV
S.M.W
GMCV
SMCV
Tamil) Salniomdae in the cla^b (JileichlliN
1
1
1
1
Second family in the class Osteichthyes, preferably
a commercially or recreationally important
warmwater species
3
3
2
2
Third family in the phylum Chordata (may be in
the class Osteichthyes or may be an amphibian,
etc.)
5
10
2
2
Planktonic Crustacean
3
6
3
4
Benthic Crustacean
1
1
1
1
Insect
1
1
1
1
Family in a phylum other than Arthropoda or
Chordata (e.g., Rotifera, Annelida, orMollusca)
4
4
1
1
Family in any order of insect or any phylum not
already represented
1
1
1
1
Total
19
27
12
13
MDU
Saltwater1
GMAV
S.M.W
CMC V
SMCV
Family in the phylum Chordata
0
0
1
1
Family in the phylum Chordata
0
0
0
0
Either the Mysidae or Penaeidae family
2
2
0
0
Family in a phylum other than Arthropoda or
Chordata
1
1
0
0
Family in a phylum other than Chordata
1
1
1
1
Family in a phylum other than Chordata
0
0
Family in a phylum other than Chordata
0
0
Any other family
0
0
Total
4
4
2
2
a The 1985 Guidelines require that data from a minimum of eight families are needed to calculate an
estuarine/marine criterion. Insufficient data exist to fulfill all eight of the taxonomic MDR groups. Consequently,
the EPA cannot derive an estuarine/marine acute criterion, based on the 1985 Guidelines. However, the EPA has
developed estuarine/marine benchmarks through use of surrogate data to fill in missing MDRs using the EPA's
WeblCE tool and other New Approach Methods. These benchmarks are provided in Appendix L.
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3.1.1.1 Summary of Acute PFO A Toxicity Studies Used to Derive the Freshwater Aquatic Life
Criterion
The acute data set for PFO A contains 19 genera representing all eight taxonomic MDR
groups (Table 3-3). Quantitatively-acceptable data for acute PFOA toxicity were available for
four freshwater fish species, representing four genera and three families and fulfilled two of the
eight MDRs. Quantitatively acceptable data for acute PFOA toxicity were also available for 13
freshwater invertebrate species, representing ten genera and nine families, and fulfilled five of
the eight MDRs. Quantitatively acceptable acute data were available for 10 freshwater
amphibian species, representing five genera and five families fulfilling one of the MDRs.
Summaries of studies for the most sensitive acute genera are described below, with the four most
sensitive genera provided in Table 3-2.
Table 3-2. The Four Most Sensitive Genera Used in Calculating the Acute Freshwater
Criterion (Sensitivity Rank 1-4).
Ranked Below from Most to Least Sensitive
Uank
(ienus
(i.MAY (lllg/l.)
Species
1
Moina
8.885
Cladoceran
(Moina macrocropa)
Cladoceran
(Moina micrura)
2
Neocloeon
13.05
Mayfly
(Neocloeon triangulifer)
3
Chydorus
93.17
Cladoceran
(Chydorus sphaericus)
4
Daphnia
142.8
Cladoceran
(Daphnia carinataf
Cladoceran
(Daphnia magna)
Cladoceran
(Daphnia pulicaria)
a Not a North American resident but in same genus as resident species.
3.1.1.1.1 Most acutely sensitive genus: Moina (cladoceran)
Ji et al. (2008) performed a 48-hour static, unmeasured acute test of PFOA (CAS # 335-
67-1, purity unreported) with Moina macrocopa. Testing method followed U.S. EPA/600/4-
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90/027F (2002). The test involved four replicates of five daphnids (<24 hours old) each in five
unmeasured test concentrations (62.5, 125, 250, 500 and 1,000 mg/L) plus a negative control.
Dilution water was moderately hard reconstituted water. Survival of daphnids in the negative
control was not reported, although EPA/600/4-90/027F requires at least 90% survival for test
acceptability. The author-reported 48-hour EC so for the study was 199.51 mg/L. The EPA
performed C-R analysis for the test. The EPA-calculated EC so (166.3 mg/L PFOA) was
acceptable for quantitative use in deriving the acute freshwater PFOA criterion and served
directly as thq Moina macrocopa Species Mean Acute Value (SMAV).
Razak et al. (2023) tested the acute toxicity of perfluorooctanoic acid (PFOA, >98%
purity) to Moina micrura in a 48-hour static measured experiment. Testing methods followed
OECD 202 (OECD 2004) with nominal testing concentrations of 10, 25, 50, 75, 100, 250, 500,
750, 1,000, 2,500, 5,000, 7,500, and 10,000 |ig/L, plus a control, with four replicates per
treatment. Filtered surface lake water was used as test water. Measured concentrations were not
reported, but the authors noted they were 94.3ฑ6.1% of nominal on average. Each replicate
consisted of 10 neonates (<48 hours old) in 50 mL of solution in a 100 mL beaker, and
organisms were not fed during the study. Lethal effect concentrations (LC) were calculated using
Probit analysis, and the 48-hour LCso value of 474.7 |ig/L, or 0.4747 mg/L was determined to be
acceptable for quantitative use. The 48-hr LCso was acceptable for quantitative use in deriving
the acute freshwater PFOA criterion and served directly as thq Moina micrura SMAV.
Moina macrocopa (SMAV of 166.3 mg/L) is much more tolerant to acute freshwater
PFOA exposures thanM micrura (SMAV of 0.4747 mg/L) but both species were used to
determine the Moina Genus Mean Acute Value (GMAV) of 8.885 mg/L. If the EPA were to
exclude theM micrura SMAV on the basis of it being an overly sensitive outlier (relative toM
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macrocopa and the overall quantitatively acceptable acute data), that would result in the final
PFOA acute criterion potentially being underprotective of untested sensitive invertebrate species
or other taxa, considering that the available data serve as surrogate information for the thousands
of untested freshwater species. Conversely, excluding theM macrocopa SMAV on the basis of
it being a tolerant outlier (relative toM micrura) would result in the final PFOA acute criterion
being highly influenced by a single test/species with an LC50 that was uniquely sensitive (i.e., M
micrura SMAV = 0.4747 mg/L) compared to the overall acute data. Averaging theM micrura
and M. moina SMAVs resulted in a GMAV (8.885 mg/L) that was the most sensitive GMAV,
yet still in the general range of the data overall.
3.1.1.1.2 Second most acutely sensitive genus: Neocloeon (mayfly)
Soucek et al. (2023) exposed the parthenogenetic mayfly, Neocloeon triangulifer, to
PFOA (CAS # 335-67-1, 95% purity) in water in a 96-hour acute toxicity test. The test was
performed under static, non-renewal conditions beginning with <24-hour old larvae. Mayflies
were fed live diatom biofilm scraping beginning on day 0. Feeding only occurred on day 0, and
the authors indicated test organisms required food to survive the entire 96-hour exposure, with
previous studies demonstrating greater than 80% mortality at 48 hours with no food (Soucek and
Dickinson 2015). Percent survival in the control treatment after 96 hours was 100%. The EPA-
calculated acute LC50 (i.e., 13.045 mg/L) was similar to the author-reported LC50 of 13.451
mg/L. The EPA-calculated LC50 value (i.e., 13.045 mg/L) was acceptable for quantitative use in
deriving the acute freshwater PFOA criterion and served directly as the Neocloeon GMAV.
3.1.1.1.3 Third most acutely sensitive genus: Chydorus (cladoceran)
Le and Peijnenburg (2013) performed a 48-hour static unmeasured acute PFOA toxicity
test with the cladoceran, Chydorus sphaericus. The authors reported the 48-hour EC50 was 0.22
mM PFOA (91.10 mg/L). The EPA performed concentration-response (C-R) analysis for the test
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and calculated a LC50 of 93.17 mg/L PFOA that is acceptable for quantitative use. No other
quantitatively acceptable acute toxicity data were available for Chydorus sphaericus or other
members of the genus Chydorus. Therefore, the LC50 (i.e., 93.17 mg/L) from this test served
directly as the Chydorus sphaericus SMAV and the Chydorus GMAV.
3.1.1.1.4 Fourth most acutely sensitive genus: Daphnia (cladoceran)
Logeshwaran et al. (2021) conducted an acute PFOA test with the cladoceran, Daphnia
cannula, and PFOA (95% purity, purchased from Sigma-Aldrich Australia) following OECD
guidelines (2000) with slight modifications. Authors used nominal test concentrations (0, 0.5, 1,
2.5, 5, 10, 20, 30, 40, 50, 100, 150, 200 and 250 mg/L PFOA) with three replicates per treatment.
No mortality occurred in the controls. The author-reported 48-hour EC50 was 78.2 mg/L PFOA.
The EPA-calculated 48-hour EC50 value was 66.80 mg/L, which was acceptable for quantitative
use. No other quantitatively acceptable acute tests were available for this species and the EC50 of
66.80 mg/L from Logeshwaran et al. (2021) served directly as the Daphnia carinata SMAV.
Boudreau (2002) performed a 48-hour static PFOA (CAS # 335-67-1, >97% purity)
acute test with Daphniapulicaria, following ASTM E729-96 (1999). Five unmeasured test
concentrations plus a negative control were used with 3-4 replicates per treatment and 10
daphnids per replicate. Nominal concentrations were 0 (negative control), 26.3, 52.6, 105, 210
and 420 mg/L. Mortality of daphnids in the negative control was not reported, but the protocol
followed by the authors (i.e., ASTM E729-96) required > 90% survival in negative controls. The
48-hour D. pulicaria EC50 reported in the publication was 203.7 mg/L, which was acceptable for
quantitative use. No other quantitatively acceptable acute tests were available for this species and
the EC50 of 203.7 mg/L from Boudreau (2002) served directly as the I), pulicaria SMAV.
Boudreau (2002) also performed a 48-hour static PFOA (CAS # 335-67-1, >97% purity)
acute test with Daphnia magna following the same methods used in the D. pulicaria acute test.
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The 48-hour D. magna EC50 reported in the publication was 223.6 mg/L, which was acceptable
for quantitative use.
Colombo et al. (2008) conducted a 48-hour static PFOA (ammonium salt, CAS # 3825-
26-1, 99.7% purity) acute test on Daphnia magna. Authors stated the test followed OECD test
guideline 202 (1992). The nominal test concentrations included control, 100, 178, 316, 562 and
1,000 mg/L, with four replicates/treatment and five animals/replicate. No mortality was observed
in the controls, and the 48-hour EC50 reported in the study was 480 mg/L, which was acceptable
for quantitative use.
Ji et al. (2008) performed a 48-hour static acute test of PFOA (CAS # 335-67-1, purity
unreported; obtained from Sigma Aldrich, St. Louis, MO) on I), magna. Authors stated the test
followed U.S. EPA/600/4-90/027F (2002). The test involved four replicates of five daphnids
each in five unmeasured test concentrations plus a negative control. Nominal concentrations
were 0 (negative control), 62.5, 125, 250, 500 and 1,000 mg/L. Mortality of daphnids in the
negative control was not reported, although EPA/600/4-90/027F requires at least 90% survival
for test acceptability. The author-reported 48-hour EC50 for the study was 476.52 mg/L (95% C.I.
= 375.3 - 577.7 mg/L). The EPA performed C-R analysis for the test. The EPA-calculated EC50
was 542.5 mg/L PFOA and was acceptable for quantitative use.
Li (2009) conducted a 48-hour static PFOA (ammonium salt, >98%) purity) acute test
with Daphnia magna. Authors stated the test generally followed OECD 202 (1984). The test
employed five replicates of six daphnids each in five test concentrations (nominal range = 31-
250 mg/L) plus a negative control. No control daphnids were immobile at the end of the test. The
author-reported 48-hour EC50 for the study was 181 mg/L (95%> C.I.: 166-198 mg/L) which was
averaged across three tests. The EPA performed C-R analysis for each individual test. All three
65
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tests had acceptable curves with the EPA-calculated EC50 values of 220.8 mg/L, 157.9 mg/L, and
207.3 mg/L, which were acceptable for quantitative use.
Yang et al. (2014) conducted a 48-hour acute test of PFOA (CAS # 335-67-1, 99%
purity) with Daphnia magna, following ASTM E729 (1993). The test employed three replicates
of 10 daphnids each in six test concentrations plus a negative and solvent control. Nominal
concentrations were 0 (negative and solvent controls), 50, 80, 128, 204.8, 327.68 and 524.29
mg/L. Test concentrations were measured in low and high treatments only. Negative control and
solvent control mortality were 0% each. The author-reported 48-hour LC50 was 201.85 mg/L
(95% C.I. = 134.68 - 302.5 mg/L). The EPA performed C-R analysis for the test and fit an
acceptable curve with an EPA-cal culated LC50 of 222.0 mg/L PFOA, which was acceptable for
quantitative use.
Barmentlo et al. (2015) performed a 48-hour static, measured acute test of PFOA (CAS
# 335-67-1, >96% purity) with Daphnia magna following OECD 202 (2004) test guidelines. The
test involved four to six replicates of five daphnids each in five test concentrations plus a
negative control. Nominal concentrations were not provided, but PFOA was measured in the
control, lowest, and highest test concentrations. Based on these measurements, the authors
interpolated all test concentrations to be: 0.053 (negative control), 81, 128, 202, 318 and 503
mg/L. The author-reported 48-hour EC50 was 239 mg/L (95% C.I. = 190-287 mg/L). The EPA
performed C-R analysis for the test and fit an acceptable curve with an EPA-cal culated EC50 of
215.6 mg/L PFOA, which was acceptable for quantitative use.
Ding et al. (2012a) conducted a 48-hour static, partially measured acute test on PFOA
(CAS # 335-67-1; 96% purity from Sigma Aldrich) with I), magna. The test generally followed
OECD test guideline 202 (2004). Authors employed four replicates of five daphnids each in six
66
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test concentrations plus a negative control. Nominal concentrations were 0 (negative control),
144.9, 165.6, 186.3, 207.0, 227.7, and 248.4 mg/L. Concentrations of PFOA were confirmed in
the highest and lowest concentrations, though only nominal concentrations were reported. It was
stated that the verified concentrations were "well in line with nominal concentrations". The 48-
hour ECso was reported as 211.6 mg/L (95% C.I. = 184.7 - 255.5 mg/L). The EPA performed C-
R analysis for the test. The EPA-calculated EC50 was 216.1 mg/L PFOA, which was acceptable
for quantitative use.
Lu et al. (2016) evaluated the acute toxicity of PFOA (CAS# 335-67-1, 98% purity,
purchased from Tokyo Chemical Industry Co., Ltd., Tokyo, Japan) on Daphnia magna
immobilization. The test was conducted following modified OECD standard test procedure 202,
whereby five concentration treatments (3, 10, 30, 100 and 300 mg/L) plus a blank control were
employed with three replicates per treatment. Authors reported immobility/survival to be a more
sensitive endpoint than survival alone. The author-reported 48-hour EC50 for immobility/survival
was 110.7 mg/L and the EPA-calculated 48-hour EC50 was 114.6 mg/L, which was acceptable
for quantitative use.
Yang et al. (2019) evaluated the acute effects of PFOA (CAS# 335-67-1, purchased from
Sigma-Aldrich in St. Louis, MO) on Daphnia magna in a 48-hour unmeasured static exposure.
Authors stated the protocol for all testing followed OECD Guideline 202. Nominal acute test
concentrations included 0 (control), 66.67, 79.92, 96.06, 115.1, 138.3, and 166.0 mg/L PFOA,
with four replicates per treatment. Authors reported an LC50 of 120.9 mg/L PFOA. The EPA-
calculated 48-hour LC50 was 117.2 mg/L, which was acceptable for quantitative use.
Chen et al. (2022) tested the acute toxicity of ammonium perfluorooctanoic acid (APFO,
>98% purity) to Daphnia magna in a 48-hour unmeasured, static experiment. The acute APFO
67
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exposure generally followed test method OECD 202 (OECD 2004). For each replicate, 10
neonates were exposed to 50 mL of test solutions with nominal concentrations of 0, 10, 50, 100,
200, and 500 mg/L APFO, made by diluting the stock solution with culture medium. There were
three replicates per treatment and test organisms were not fed during the experiment. Control
mortality was less than 10% in the AFPO and reference toxicity tests. The 48-hour EC so for
immobility was 156.9 mg/L, which was acceptable for quantitative use.
Quantitatively acceptable D. magna acute values from Boudreau (2002; EC so = 223.6
mg/L), Colombo et al. (2008; ECso = 480 mg/L), Ji et al. (2008; ECso = 542.5 mg/L), Li (2009;
ECso = 220.8, 157.9, and 207.3 mg/L), Yang et al. (2014; LC50 = 222.0 mg/L), Barmentlo et al.
(2015; ECso = 215.6 mg/L), Ding et al. (2012a; ECso = 216.1 mg/L), Lu et al. (2016; ECso =
114.6 mg/L), Yang et al. (2019; LCso = 117.2 mg/L), and Chen et al. (2022; ECso = 156.9 mg/L)
were taken together as a geometric mean value to calculate the D. magna SMAV of 213.9 mg/L.
The D. carinata SMAV (i.e., 66.80 mg/L), I), pulicaria SMAV (i.e., 203.7 mg/L), and D. magna
SMAV (i.e., 213.9 mg/L) were used to determine the Daphnia GMAV of 142.8 mg/L.
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Table 3-3. Ranked Freshwater Genus Mean Acute Values.
Rank11
(/MAY
(ing/l. PI-OA)
MI)U
(roup1'
Genus
Species
SMAY1'
(ing/l. PI OA)
1
8.885
D
Moina
Cladoceran,
Moina macrocopa
166.3
Cladoceran,
Moina micrura
0.4747
2
13.05
F
Neocloeon
Mayfly,
Neocloeon triangulifer
13.05
3
93.17
D
Chydorus
Cladoceran,
Chydorus sphaericus
93.17
4
142.8
D
Daphnia
Cladoceran,
Daphnia carinata
66.80
Cladoceran,
Daphnia magna
213.9
Cladoceran,
Daphnia pulicaria
203.7
5
150.0
H
Brachionus
Rotifer,
Brachionus calyciflorus
150.0
6
161.0
G
Ligumia
Black sandshell,
Ligumia recta
161.0
7
164.4
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
164.4
8
377.0
C
Xenopus
Frog,
Xenopus sp.
377.0
9
383.6
H
Dugesia
Planaria,
Dugesia japonica
383.6
10
431.5
E
Neocaridina
Green neon shrimp,
Neocaridina denticulata
431.5
11
450.4
B
Danio
Zebrafish,
Danio rerio
450.4
12
593.6
B
Pimephales
Fathead minnow,
Pimephales promelas
593.6
13
646.2
C
Hyla
Gray treefrog,
Hyla versicolor
646.2
14
664.0
B
Lepomis
Bluegill,
Lepomis macrochirus
664.0
15
681.1
G
Physella
Bladder snail,
Physella acuta
681.1
16
689.4
C
Ambystoma
Jefferson salamander,
Ambystoma jeffersonianum
1,070
Small-mouthed salamander,
Ambystoma texanum
407.3
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Ksink"
Ci.MAV
(ing/l. PI OA)
MI)U
(roup1'
(it'll IIS
Species
SMAY1'
(ing/l. PI OA)
Eastern tiger salamander,
Ambystoma tigrinum
752.0
17
793.9
C
Anaxyrus
American toad,
Anaxyrus americanus
793.9
18
951.5
C
Lithobates
American bullfrog,
Lithobates catesbeiana
1,020
Green frog,
Lithobates clamitans
1,070
Northern leopard frog,
Lithobates pipiens
751.7
Wood frog,
Lithobates sylvatica
999
19
4,001
A
Oncorhynchus
Rainbow trout,
Oncorhynchus mykiss
4,001
a Ranked from the most sensitive to the most resistant based on Genus Mean Acute Value,
b From Appendix A: Acceptable Freshwater Acute PFOA Toxicity Studies,
c MDR Groups - Freshwater:
A. the family Salmonidae in the class Osteichthyes
B. a second family in the class Osteichthyes, preferably a commercially or recreationally important
warmwater species (e.g., bluegill, channel catfish, etc.)
C. a third family in the phylum Chordata (may be in the class Osteichthyes or may be an amphibian, etc.)
D. aplanktonic crustacean (e.g., cladoceran, copepod, etc.)
E. abenthic crustacean (e.g., ostracod, isopod, amphipod, crayfish, etc.)
F. an insect (e.g., mayfly, dragonfly, damselfly, stonefly, caddisfly, mosquito, midge, etc.)
G. a family in a phylum other than Arthropoda or Chordata (e.g., Rotifera, Annelida, Mollusca, etc.)
H. a family in any order of insect or any phylum not already represented.
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1.0 T
0.9 --
J ฐ-
0.7 -
u
a
C
4i
>
a 0.6
9
s
= 0.5
ง 0-4
C4
oป
3 0.3 +
a
-------
The following section provides information and summaries of studies for the sensitive
estuarine/marine taxa, based on the limited available data (Table 3-4). Study summaries for the 4
most sensitive genera are provided below.
Table 3-4. Estuarine/Marine Acute PFOA Genera.
Ranked Below from Most to Least Sensitive
Uank
Genus
GMAV
(nig/I,)
Species
1
Siriella
15.5
Mysid
(Siriella armataf
2
Mytilus
17.58
Mediterranean mussel
(Mytilus galloprovincialis)
3
Strongylocentrotus
20.63
Purple sea urchin
(Strongylocentrotus purpuratus)
4
Americamysis
24
Mysid
(Americamysis bahia)
a Not a North American resident species, but a member of the Siriella genus and serves as a surrogate for untested
Siriella species residing in North America (Heard et al. 2006).
3.1.1.2.1 Most sensitive estuarine/marine genus: Siriella (mysid)
Mhadhbi et al. (2012) performed a 96-hour static, unmeasured acute test with PFOA
(96% purity) on the mysid, Siriella armata. Mysids were exposed to one of ten nominal PFOA
treatments (0.1, 0.5, 1, 2, 5, 10, 20, 30, 40 and 80 mg/L). Neonates were fed 10-15 Artemia
salina nauplii daily and mortality was recorded after 96 hours. The 96-hour LCso reported in the
study was 15.5 mg/L PFOA and was acceptable for quantitative use. No other quantitatively
acceptable acute toxicity data were available for Siriella armata or other members of the genus
Siriella. Therefore, the LCso (i.e., 15.5 mg/L) from this test served directly as the Siriella
GMAV. Although S. armata is not a North American resident species, it is a member of the
Mysidae Family and serves as a surrogate for untested mysid species, including members of the
genus Siriella, residing in North America (Heard et al. 2006).
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3.1.1.2.2 Second most sensitive estuarine/marine genus: Mytilus (mussel)
The acute toxicity of PFOA (purity not provided) on the Mediterranean mussel, Mytilus
galloprovincialis, which occurs in California and other parts of the Pacific Northwest (Green
2014), was evaluated by Fabbri et al. (2014). The endpoint was the percent reduction of normal
D-larvae in each well. Authors noted that controls had >80% normal D-larvae across all tests,
meeting the >75% acceptability threshold outlined by ASTM (2004). PFOA was only measured
once in one treatment which was similar to the nominal concentration. The percentage of normal
D-larva decreased with increasing test concentrations. The NOEC and LOEC reported for the
study were 0.00001 and 0.0001 mg/L, respectively. Although authors report -27% effect at the
LOEC (i.e., 0.0001 mg/L), the test concentrations failed to elicit 50% malformations in the
highest test concentration, and an EC so was not determined. Therefore, the EC so for the study
was greater than the highest test concentration (1 mg/L). The 48-hour EC so based on
malformation of >1 mg/L was quantitatively acceptable.
Hayman et al. (2021) reported the results of a 48-hour static, measured acute PFOA
(CAS # 335-67-1, 95% purity, purchased from Sigma-Aldrich, St. Louis, MO) test on the
Mediterranean mussel, Mytilus galloprovincialis. Authors note that tests followed U.S. EPA
(1995b) and ASTM (2004) protocols. Six test solutions were made in 0.45 |im filtered seawater
(North San Diego Bay, CA) with PFOA dissolved in methanol. The highest concentration of
methanol was 0.02% (v/v) and each treatment solution contained five replicates. At test
termination (48 hours), larvae were enumerated for total number of larvae that were alive at the
end of the test (normally or abnormally developed) as well as number of normally-developed (in
the prodissoconch "D-shaped" stage) larvae. There were no significant differences between
solvent control and filtered seawater, suggesting no adverse effects of methanol. The author
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reported 48-hour EC50, based on normal larvae survival was 9.98 mg/L PFOA. The EPA-
calculated 48-hour EC50 value was 17.58 mg/L, which was acceptable for quantitative use.
Although the 48-hour EC50 based on malformation of >1 mg/L from Fabbri et el. (2014)
met the EPA's data quantitatively objectives, it was not used directly in the calculation of the M.
galloprovincialis SMAV because it was a "greater than LC50" value and a definitive LC50 was
available for the same species as reported by Hayman et al. (2021). The definitive LC50 value
from Hayman et al. (2021) of 17.58 mg/L served directly as the Mytilus galloprovincialis SMAV
and as the Mytilus GMAV.
3.1.1.2.3 Third most sensitive estuarine/marine genus: Strongylocentrotus (urchin)
Hayman et al. (2021) reported the results of a 96-hour static, measured PFOA (CAS #
335-67-1, 95% purity, purchased from Sigma-Aldrich, St. Louis, MO) test with the purple sea
urchin, Strongylocentrotuspurpuratus. Authors note that tests followed U.S. EPA (1995b) and
ASTM (2004) protocols. Six test solutions were made in 0.45 |im filtered seawater (North San
Diego Bay, CA) with PFOA dissolved in methanol and each treatment was replicated five times.
The highest concentration of methanol was 0.02% (v/v). At test termination (96 hours), the first
100 larvae were enumerated and observed for normal development (organisms distinguished as
being in the four arm pluteus stage). There were no significant differences between solvent
control and filtered seawater, suggesting no adverse effects of methanol. The author reported 96-
hour EC50, based on normal development, was 19 mg/L PFOA. The EPA-calculated 96-hour
EC50 value was 20.63 mg/L, which was acceptable for quantitative use. The EC50 value of 20.63
mg/L was the only acceptable acute value for Strongylocentrotus purpuratus or any members of
the genus Strongylocentrotus. Therefore, it served directly as the Strongylocentrotus purpuratus
SMAV and the Strongylocentrotus GMAV.
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3.1.1.2.4 Fourth most sensitive estuarine/marine genus: Americamysis (mysid)
Hayman et al. (2021) conducted a 96-hour static, measured test to assess effects of
PFOA (CAS # 335-67-1, 95% purity, purchased from Sigma-Aldrich, St. Louis, MO) on the
mysid, Americamysis bahia. Authors note that tests followed U.S. EPA (2002) protocols. Six test
solutions were made in 0.45 |im filtered seawater (North San Diego Bay, CA) with PFOA
dissolved in methanol. The highest concentration of methanol was 0.02% (v/v) and each test
solution was replicated six times with five mysids per replicate. There were no significant
differences between solvent control and filtered seawater, suggesting no adverse effects of
methanol. No organisms were found dead in the controls at test termination. The EPA was
unable to fit a concentration-response model with significant parameters and relied on the author-
reported 96-hour LCso of 24 mg/L PFOA as the quantitatively acceptable acute value. The LC50
value of 24 mg/L was the only acceptable acute value for Americamysis bahia or any members
of the genus Americamysis. Therefore, it served directly as the Americamysis bahia SMAV and
the Americamysis GMAV.
The estuarine/marine acute data set for PFOA contained four genera (Figure 3-2)
representing only three of the eight taxonomic MDR groups. The missing MDR groups included
two families in the phylum Chordata, two families in a phylum other than Chordata, and any
other family not already represented (Table 3-5). As noted above, the EPA used the available
acute toxicity data and ORD's peer-reviewed web-ICE tool to develop aquatic life benchmarks
for consideration by states and Tribes (see Appendix L).
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Tab
e 3-5. Ranked Estuarine/Marine Genus Mean Acute Values.
GMAV
MDR
SMAVb
Rank3
(mg/L PFOA)
Group0
Genus
Species
(mg/L PFOA)
1
15.5
C
Siriella
Mysid,
Siriella armata
15.5
2
17.58
D
Mytilus
Mediterranean mussel,
Mytilus galloprovincialis
17.58
3
20.63
F
Strongylocentrotus
Purple sea urchin,
Strongylocentrotus
purpuratus
20.63
4
24
C
Americamysis
Mysid,
Americamysis bahia
24
a Ranked from the most sensitive to the most resistant based on Genus Mean Acute Value.
B From Appendix B: Acceptable Estuarine/Marine Acute PFOA Toxicity Studies
c MDR Groups identified in Footnote C of Table 3-1.
1.0
0.9
a
.2 0.8
*->
u
S3
u
0.7 --
tj
ฃ
ซ 0.6
3
s
3 0.5
g 0.4
X
QJ
r= 0.3
a
ai
S 0.2
0.1
0.0
I Invertebrate (Other)
~ Invertebrate (Mollusk)
Americamysis
Strongylocentrotus
Mytilus
Siiiella
10
Genus Mean Acute Value (mg/L PFOA)
100
Figure 3-2. Acceptable Estuarine/Marine GMAVs.
76
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3.1,1.3 Summary of Chronic PFOA Toxicity Studies Used to Derive the Freshwater Aquatic
Life Criterion
Acceptable chronic PFOA toxicity data in freshwater were available for a total of 13
species representing 12 genera and ten families in three phyla. All eight of the required MDRs
were fulfilled. The following section provides information and summaries of studies for the
sensitive taxa with effect values used in the quantitative calculation of the chronic freshwater
PFOA criterion (Table 3-6). The chronic data set for PFOA based on quantitatively acceptable
data contains 12 genera representing all eight taxonomic MDR groups (Table 3-7).
Table 3-6. The Most Sensitive Genera Used in Calculating the Chronic Freshwater Water
Column Criterion (Sensitivity Rank 1-4).
Ranked Below from Most to Least Sensitive
Rank
(ienus
(;\l( V
(mg/l<)
Species
1
Hyalella
0.147
Amphipod
(Hyalella azteca)
2
Lithobates
0.288
American bullfrog
(Lithobates catesbeiana)
3
Daphnia
0.3695
Cladoceran
(Daphnia carinatdf
Cladoceran
(Daphnia magna)
4
Brachionus
0.7647
Rotifer
(Brachionus calyciflorus)
a Not a North American resident but in same genus as resident species.
3.1.1.3.1 Most chronically sensitive genus: Hyalella (amphipod)
Bartlett et al. (2021) evaluated the chronic effects of PFOA (CAS# 335-67-1, 96%
purity, solubility in water at 20,000 mg/L, purchased from Sigma-Aldrich) on Hyalella azteca
via a 42-day static-renewal, measured study. Methods for this study were adapted from
Borgmann et al. (2007), and organisms were two to nine days old at the test initiation. A 100
mg/L stock solution was prepared to yield measured test concentrations of 0 (control), 0.84, 3.3,
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8.9, 29 and 97 mg/L PFOA. Two separate tests were performed with five replicates per
concentration and 20 amphipods per replicate. At test termination (day 42), adults were sexed
and weighed, as well as their young counted. The 42-day author-reported LCio value for survival
was 23.2 mg/L PFOA. The author-reported ECio values for growth and reproduction were 0.160
mg/L and 0.0265 mg/L, respectively. The EPA only performed C-R analysis for the growth and
reproduction-based endpoints for this test, given the apparent tolerance of the survival-based
endpoint. The EPA calculated ECio values for the 42-day growth endpoint (i.e., control
normalized wet weight/amphipod) and the 42-day reproduction endpoint (i.e., number of
juveniles per female). The 42-day growth-based ECio of 0.488 mg/L was not selected as the
primary endpoint from this test because it was more tolerant than the reproduction-based ECio of
0.147 mg/L, which was acceptable for qualitative use. The ECio value of 0.147 mg/L was the
only acceptable chronic value for Hyalella azteca or any members of the genus Hyalella.
Therefore, it served directly as the Hyalella azteca Species Mean Chronic Value (SMCV) and
the Hyalella Genus Mean Chronic Value (GMCV).
3.1.1.3.2 Second most chronically sensitive genus: Lithobates (frog)
Flynn et al. (2019) evaluated the chronic effects of PFOA (CAS# 335-67-1, purchased
from Sigma-Aldrich) on the American bullfrog (Lithobates catesbeiana, formerly, Rana
catesbeiana) during a 72-day static-renewal unmeasured exposure. The authors tested a negative
control and two treatment concentrations (i.e., 0.144 and 0.288 mg/L), which were the only three
PFOA-only treatments within the larger factorially-designed experiment. Each treatment
contained 10 tadpoles (Gosner stage 25) and treatments were replicated four times. On day 72 of
the experiment, all tadpoles were euthanized and measured (snout vent length and mass). The
most sensitive chronic endpoint was growth (snout-vent length), with a 72-day NOEC and LOEC
of 0.144 mg/L and 0.288 mg/L, respectively. The EPA could not independently calculate an ECio
78
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value because there were minimal effects observed across the limited number of treatment
concentrations tested. Consequently, the EPA used the LOEC of 0.288 mg/L as the chronic value
from this chronic test. The LOEC was used preferentially to the MATC from this test because a
-7% reduction in snout-vent length relative to control responses was observed at the LOEC (i.e.,
0.288 mg/L), which is a similar effect level to the chronic 10% effect level (i.e., ECio) used
preferentially to derive the chronic criterion. The LOEC value of 0.288 mg/L was the only
acceptable chronic value for Lithobates catesbeiana or any members of the genus Lithobates.
Therefore, it served directly as the Lithobates catesbeiana SMC V and the Lithobates GMCV.
3.1.1.3.3 Third most chronically sensitive genus: Daphnia (cladoceran)
Logeshwaran et al. (2021) conducted a static-renewal unmeasured PFOA (95% purity,
purchased from Sigma-Aldrich Australia) chronic toxicity test with the cladoceran, Daphnia
carinata. Authors stated the chronic test protocol followed OECD guidelines (2012). Authors
tested a negative control and five PFOA concentrations (i.e., 0.001, 0.01, 0.1, 1.0 and 10 mg/L
PFOA). Each test treatment was replicated 10 times with one daphnid (six to 12 hours old) per
treatment. At test termination (21 days) test endpoints included survival, days to first brood,
average offspring in each brood and total live offspring. No mortality occurred in the controls or
lowest test concentration. Of the three endpoints measured, average offspring in each brood and
total live offspring were the more sensitive endpoints with 21-day NOEC and LOEC values of
0.01 and 0.1 mg/L PFOA, respectively. The EPA was unable to calculate statistically robust ECio
estimates from C-R models for these endpoints, largely because of the 10X dilution series across
five orders of magnitude. The LOECs for these endpoints were not selected as the chronic value
because the LOECs produced a 29.23% reduction in the average number of offspring per brood
relative to controls and a 39.89% reduction in the total living offspring relative to controls.
Therefore, the MATC (i.e., 0.03162 mg/L) was selected as the quantitatively acceptable chronic
79
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value from this test. The MATC value of 0.03162 mg/L was the only acceptable chronic value
for Daphnia carinata, and it served directly as the Daphnia carinata SMCV.
Centre International de Toxicologic (2003) and Colombo et al. (2008) conducted a 21-
day renewal measured chronic test on PFOA with the daphnid, Daphnia magna. Authors stated
that the toxicity test conducted followed OECD test guideline 211. Average number of live
young was the most sensitive endpoint reported by Colombo et al. (2008), with a NOEC of 20
mg/L. The author-reported NOEC for the average number of live young was 20 mg/L, the LOEC
was 44.2 mg/L and the MATC was 29.73 mg/L. The EPA performed C-R analysis for each
reported endpoint. The most sensitive endpoint with an acceptable C-R curve was the average
number of live young per starting adult, with an EPA-calculated ECio of 20.26 mg/L PFOA and
was acceptable for quantitative use.
Ji et al. (2008) conducted a chronic life-cycle test on the effects of PFOA with Daphnia
magna. Authors stated that the D. magna test followed OECD 211 (1998). The most sensitive
endpoint for D. magna reported in the publication was days to first brood with a 21-day NOEC
of 6.25 mg/L (LOEC = 12.5 mg/L; MATC = 8.839 mg/L); however, number of young per
starting female (an endpoint not reported in the publication, which only assessed number of
young per surviving female) was calculated by the EPA and considered to be a more sensitive
endpoint with an EPA-calculated ECio of 7.853 mg/L. Therefore, the EPA-calculated ECio of
7.853 mg/L PFOA for D. magna (number of young per starting female) was considered
quantitatively acceptable.
Li (2010) conducted an unmeasured chronic life cycle 21-day test on the effects of PFOA
on Daphnia magna. Authors stated the test followed OECD 211 (1998). The I), magna 21-day
NOEC (reproduction as number of young per female, broods per female, and mean brood size)
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was 10 mg/L (LOEC = 32 mg/L; calculated MATC = 17.89 mg/L). The EPA performed C-R
analysis for each reported endpoint. The EPA also revaluated all endpoints that were based on
number of surviving females to be based on the number of starting females. This recalculation
was done with the intent to account for starting females that were unable to contribute to the
population as reproduction/female due to mortality. The most sensitive endpoint with an
acceptable C-R curve was the number of young per starting female with an EPA-calculated ECio
of 12.89 mg/L PFOA and was acceptable for quantitative use.
Yang et al. (2014) evaluated the chronic 21-day renewal, measured test of PFOA with
Daphnia magna, following ASTM E729 (1993). The author-reported D. magna 21-day ECio for
reproduction (total number of spawning) was 7.02 mg/L. The EPA performed C-R analysis for
each reported endpoint. Both chronic survival and reproduction endpoints resulted in acceptable
C-R curves. The EPA-calculated EC io for reproduction as total number of spawning events was
6.922 mg/L, similar to the ECio reported by the authors (i.e., 7.02 mg/L). Chronic survival was
more sensitive than reproduction, with an EPA-calculated ECio of 5.458 mg/L PFOA. Therefore,
the survival based ECio calculated by the EPA (i.e., 5.458 mg/L) was acceptable for quantitative
use.
Lu et al. (2016) evaluated the chronic toxicity of PFOA (CAS# 335-67-1, 98% purity,
purchased from Tokyo Chemical Industry Co., Ltd., Tokyo, Japan) on Daphnia magna
immobilization, growth and reproduction in a 21-day semi-static test with unmeasured treatment
solutions. Authors stated the test protocol followed OECD Test Method 211. Neonates (<24 h
old) were exposed to PFOA in one of six PFOA treatments (i.e., 0 [control] 0.032, 0.16, 0.8, 4
and 20 mg/L), with 20 replicates for each treatment. The 21-day growth and reproductive NOEC
and LOEC values were 0.032 and 0.16 mg/L PFOA, respectively. The EPA was unable to fit a
81
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C-R model with significant parameters to the chronic data associated with reproduction from this
test. The EPA-calculated ECio values for mean intrinsic rate of increase (r; population-level
endpoint that accounts for births and deaths over time) and growth (as length) were 0.0173 mg/L
and 0.0124 mg/L, respectively. Both ECio values were nearly two times lower than the NOEC
for both endpoints (i.e., 0.032 mg/L) and four times lower than the LOEC values (i.e., 0.16
mg/L) for both endpoints, where a 15.2% reduction in intrinsic rate of natural increase ฎ and an
11.9% reduction in length were observed. As a result, the MATC of 0.0716 mg/L for growth and
reproduction was selected as the most appropriate chronic value for quantitative use to in
deriving the chronic water column-based criterion.
Yang et al. (2019) evaluated the chronic effects of PFOA (CAS# 335-67-1, purchased
from Sigma-Aldrich in St. Louis, MO) on Daphnia magna via a 21-day unmeasured, static-
renewal test that assessed reproductive effects. The protocol for testing followed OECD
Guideline 211. Authors tested a negative control and four PFOA treatment concentrations
(6.708, 10.10, 15.11, and 22.61 mg/L), with each treatment replicated lOtimesand each replicate
containing one neonate (12-24 hours old) in a 100 mL glass beaker. The reproductive NOEC and
LOEC values were 6.708 and 10.10 mg/L PFOA, respectively. The EPA performed C-R analysis
for the test. The EPA-calculated ECio based on mean offspring at 21-days as a proportion of the
control response was 8.084 mg/L and was used quantitatively to derive the chronic water column
criterion.
The chronic values from Centre International de Toxicologie/Colombo et al. (2003/2008;
ECio = 20.26 mg/L; endpoint = average number of live young), Ji et al. (2008; ECio = 7.853
mg/L; endpoint = number of live young per starting female), Li (2010; ECio =12.89 mg/L;
endpoint = number of young per starting female), Yang et al. (2014; ECio = 5.458 mg/L;
82
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endpoint = survival), Lu et al. (2016; MATC = 0.0716 mg/L; endpoint = length and rate of
natural increase), and Yang et al. (2019; ECio = 8.084 mg/L; endpoint = mean offspring as a
proportion of control response) were taken together as a geometric mean to serve as the Daphnia
magna SMCV (i.e., 4.317 mg/L).
The/), carinata SMCV (i.e., 0.03162 mg/L) and the/), magna SMCV (i.e., 4.317 mg/L)
were used to calculate the Daphnia GMCV of 0.3695 mg/L. Exclusion of the D. carinata SMCV
under the basis of it being an overly sensitive outlier (relative to D. magna and the overall
quantitatively acceptable chronic data) would have resulted in the Daphnia GMCV being
potentially underprotective. Conversely, excluding the I), magna SMCV under the basis of being
a tolerant outlier (relative to D. carinata) would result in the Daphnia GMCV being highly
influenced by a single test/species with a relatively sensitive chronic value. The/), carinata
chronic value was also an MATC, calculated as the geometric mean of the NOEC (0.01 mg/L)
and LOEC (0.1 mg/L) from a 10X dilution series, meaning the MATC was influenced by a
relatively low NOEC and broadly spaced dosing. However, the corresponding LOEC (i.e., 0.1
mg/L) remains relatively sensitive and produced effects of 29.23% to 39.89%, depending on the
specific offspring-based endpoint. Using both the D. magna and I), carinata SMAVs resulted in
a Daphnia GMCV (0.3695 mg/L) that was among the four most sensitive genera and in the
general range of the data overall.
3.1.1.3.4 Fourth most chronically sensitive genus: Brachionus (rotifer)
Zhang et al. (2013a) conducted a chronic life-cycle renewal test of PFOA (CAS # 335-
67-1, 96% purity) with Brachionus calyciflorus. The test consisted of a negative control and four
PFOA concentrations (0.25, 0.5, 1.0, 2.0 mg/L PFOA). For each treatment level, fifteen amictic
rotifers were placed individually into culture plate wells containing two mL of test solution that
was renewed daily. Numbers of eggs produced and starting rotifer lifetimes were recorded for
83
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every individual, and the test was conducted until every starting rotifer from every treatment
level died, which occurred around 200 hours after test initiation. Data from this test were used to
construct survivorship and fertility tables using conventional life-history techniques, which were
used to calculate net reproductive rate, generation time, and intrinsic rate of natural increase. The
EPA calculated EC 10 values from C-R. data reported in the publication, and the most sensitive
endpoint with an acceptable C-R curve was the intrinsic rate of natural increase, with an ECio of
0.5015 mg/LPFOA.
The intrinsic rate of natural increase (d"1) is a population-level endpoint that accounts for
births and deaths over time. In Zhang et al. (2013a), the intrinsic rate of natural increase was
calculated as the natural log of the lifetime net reproductive rate for all individuals within a
population (defined here as a PFOA treatment level) divided by the average generation time of
those individuals.
The ECio calculated for the intrinsic rate of natural increase was similar to the ECio value
for average net reproductive rate (0.514 mg/L). Zhang et al. (2013a) also reported significant
reductions in egg size, with an EPA-calculated ECio = 0.193 mg/L. However, this endpoint
displayed a relatively poor concentration response relationship and may not be relevant for
assessing population level effects. For these reasons, it was not selected as the primary effect
concentration from this study. Zhang et al. (2013a) also reported effects to average juvenile
period, which was a relatively tolerant endpoint. Juvenile period decreased with increasing
exposure concentration, with the average juvenile period being about 16% faster than the control
responses in the highest treatment concentration (2.0 mg/L). Effects to chronic apical endpoints
in this publication and Zhang et al. (2014b) generally appear as a threshold effect from 0.25
mg/L to 1.0 mg/L, providing further support for selection of the ECio value (i.e., 0.5015 mg/L)
84
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based on rate of natural increase as the primary chronic value for quantitative use from Zhang et
al. (2013a).
In addition to the life cycle exposure, Zhang et al. (2013a) also conducted a second multi-
generational 28-day study to measure effects of PFOA on growth patterns, population density,
and population dynamics. The 28-day test consisted of a negative control and two PFOA
concentrations (0.25, and 2.0 mg/L PFOA). Population densities were lower than controls at both
PFOA treatment levels; however, because this study was limited to two treatment levels, it was
considered to be of secondary importance compared to the life-cycle test.
Zhang et al. (2014b) describes the results of three experiments involving Brachionus
calyciflorus exposures to PFOA (CAS # 335-67-1, 96% purity). The effects of PFOA
concentration on mictic ratios of B. calyciflorus was examined by placing individual neonates in
culture plate wells with two mL of medium containing of two PFOA concentrations (0.25 mg/L
and 2.0 mg/L) plus a control. Each treatment level, as well as the control, was replicated three
times. All eggs produced by these exposed individuals were individually incubated in culture
wells with one mL control medium. The mature Fi offspring were subsequently identified as
producing mictic or amictic eggs, and these data were used to calculate mictic ratios. The
proportions of mictic eggs increased with increasing PFOA concentration (0.56 - control, 0.72 -
0.25 mg/L, 0.75 - 2.0 mg/L), and the results were statistically significant (p<0.05). In contrast,
mictic ratios were not affected by PFOA concentrations in Zhang et al. (2013a). Because of the
inconsistent result in the mictic ratio endpoint between Zhang et al. (2013a) and Zhang et al.
(2014b), it was not selected as the representative endpoint from either publication.
The effects of PFOA concentration on resting egg production of B. Calyciflorus was
examined by exposing rotifers to one of five PFOA concentrations (plus control) in the dark for
85
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six days. Resting eggs were collected on the sixth day and then hatched in control medium 30
days later. Resting egg production decreased with increasing PFOA concentration. The EPA-
calculated ECio calculated from C-R data reported in Figure 1 of Zhang et al. (2014b) was 0.076
mg/L. Because there was only one replicate (as implied by lack of error bars in Figure 1 of the
publication, no clear description of replicates in the methods section, and no author-reported
statistical analysis of this endpoint), resting egg production from this study was not considered
quantitatively acceptable but was retained for qualitative use. In a second resting egg exposure
study, resting eggs were produced under control conditions, then allowed to hatch while exposed
to one of five PFOA concentrations plus a control. In this study, the effects of PFOA exposure
on resting egg hatching rate were not statistically significant.
Finally, the effects of PFOA concentration on B. calyciflorus population growth was
examined during a four-day study in which 10 neonates were placed into chambers with 10 mL
of medium containing one of eight PFOA concentrations, plus a control. Each treatment level, as
well as the control, was replicated at least six times. After four days, the total numbers of rotifers
in each chamber were counted, and these data were used to calculate the intrinsic rate of natural
increase (d"1), the most sensitive acceptable endpoint from this study, with an EPA-calculated
ECio of 1.166 mg/L. Beyond Zhang et al. (2013a) and Zhang et al. (2014b), no other
quantitatively acceptable chronic tests were available for Brachionus calyciflorus. The EPA-
calculated ECio values from Zhang et al. (2013a) (i.e., 0.5015 mg/L; endpoint = rate of natural
increase) and Zhang et al. (2014b) (i.e., 1.166 mg/L; endpoint = rate of natural increase) were
taken together as a geometric mean to serve as the Brachionus calyciflorus SMCV (i.e., 0.7647
mg/L). No other quantitatively acceptable chronic toxicity data were available for other members
86
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of the genus Brachionus and the Brachionus calyciflorus SMCV (i.e., 0.7647 mg/L) served
directly as the Brachionus GMCV.
Table 3-7. Ranked Freshwater Genus Mean Chronic Values.
Uank11
(;\K vh
(ing/l. PI OA)
MI)U
(roup1'
Genus
Species
SMCV1'
(ing/l. PI-OA)
1
0.147
E
Hyalella
Amphipod,
Hyalella azteca
0.147
2
0.288
C
Lithobates
American bullfrog,
Lithobates catesbeiana
0.288
3
0.3695
D
Daphnia
Cladoceran,
Daphnia carinata
0.03162
Cladoceran,
Daphnia magna
4.317
4
0.7647
G
Brachionus
Rotifer,
Brachionus calyciflorus
0.7647
5
2.194
D
Moina
Cladoceran,
Moina macrocopa
2.194
6
>3.085
H
Neocloeon
Mayfly,
Neocloeon triangulifer
>3.085
7
9.487
C
Oryzias
Medaka,
Oryzias latipes
9.487
8
23.56
D
Ceriodaphnia
Cladoceran,
Ceriodaphnia dubia
23.56
9
>30
B
Gobiocypris
Rare minnow,
Gobiocypris rarus
>30
10
>40
A
Oncorhynchus
Rainbow trout,
Oncorhynchus mykiss
>40
11
>76
B
Pimephales
Fathead minnow,
Pimephales promelas
>76
12
88.32
F
Chironomus
Midge,
Chironomus dilutus
88.32
a Ranked from the most sensitive to the most resistant based on Genus Mean Chronic Value,
b From Appendix C: Acceptable Freshwater Chronic PFOA Toxicity Studies,
c MDR Groups identified in Footnote C of Table 3-3.
87
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1.0
0.9 -
a
.2 0.8 4
U
C3
"-1 0.7 4
&H
o.i 4
o.o
ฆ Invertebrate
a Insect
Fish
~ Amphibian
Chironomus a
Piniephales (non-definitive, greater than value)
Oncorhynchus (non-definitive, greater than value)
Gobiocypris (non-definitive, gr eater than value)
Ceriodaphnia ฆ
Oryzias
A Neocloeon (non-definitive,
greater than value)
ฆ Moina
Brachionus
ฆ Daphnia
~ Lithobates
Hyalella
0.01
0.1 1 10
Genus Mean Chronic Value (mg/L PFOA)
100
Figure 3-3. Freshwater Genus Mean Chronic Values for PFOA.
3.1.1.4 Summary of Chronic PFOA Toxicity Studies Used to Derive the Estuarine/Marine
Aquatic Life Criterion
Quantitatively-acceptable empirical data for chronic PFOA toxicity were available for
two estuarine/marine species, fulfilling two of the eight MDRs (Table 3-8).
3.1.1.4.1 Most chronically sensitive estuarine marine species: Oryzias (fish)
Oh et al. (2013) evaluated the chronic toxicity of PFOA on the Japanese medaka,
Oryzias latipes, in a 30-day static exposure. Prior to test initiation, the fish were acclimated to a
seawater environment. Fish were exposed for 30 days to one PFOA concentration (1.0 mg/L), a
0.22 |im filtered seawater control, and a DMSO carrier solvent control. The 30-day condition
factor NOEC of 1.0 mg/L PFOA was selected as the primary endpoint from this study. The
methods section of Oh. et al. (2013) also stated, "In our preliminary study, fish mortality was
altered 30 days after perfluorinated compound exposure, suggesting that repeated exposure to
88
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PFCs for 30 days at 1 |ag/mL [1 mg/L] causes adverse effect on 0. latipesThe statement about
mortality-based effects in their preliminary test is in direct conflict with the condition-based
results from the primary test that was the focus of the publication. Few details are provided about
the preliminary test. Results of the final test (i.e., condition factor NOEC of 1.0 mg/L reported in
Table 1 of Oh et al. 2013) were retained as quantitatively acceptable because they provide
chronic estuarine/marine data that were otherwise limited.
3.1.1.4.2 Second most chronically sensitive estuarine/marine species: Paracentrotus (sea
urchin)
Savoca et al. (2022) tested chronic toxicity of perfluorooctanoic acid (PFOA, >98%
purity) to sea urchins (Paracentrotus lividus) in a 28-day measured, static experiment. Four sea
urchins were placed into one of nine aquariums containing 15 L of seawater at nominal
concentrations (0 [control], 10, and 100 mg/L), with three replicates per concentration treatment.
The average recovery of PFOA was stable and close to nominal across all treatments throughout
the experiment, and PFOA in control water was less than 10 ng/L. Coelomic fluid was sampled
from all individuals weekly to measure PFOA uptake. During the exposure period, sea urchins at
the highest test concentration showed sublethal signs of toxicity, including reduced spine
mobility, spine loss, and reduced ability to remain anchored to the bottom of the test vessels.
Mortality was only observed at the highest test concentration. Adult mortality, adult PFOA
accumulation in coelomic fluid, percentage of normal and abnormal embryos, and gene
expression in embryos were measured. The NOEC and LOEC for adult mortality were 10 mg/L
and 100 mg/L, respectively, and the MATC of 31.62 mg/L was determined to be acceptable for
quantitative use.
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Table 3-8. Ranked Estuarine/Marine Genus Mean Chronic Values.
GMCVb
MDR
SMCVb
Rank3
(mg/L PFOA)
Group0
Genus
Species
(mg/L PFOA)
1
>10
A
Oiyzias
Japanese medaka,
Oryzias latipes
>10
2
31.62
E
Paracentrotus
Purple sea urchin,
Paracentrotus lividus
31.62
a Ranked from the most sensitive to the most resistant based on Genus Mean Chronic Value,
b From Appendix D: Acceptable Estuarine/Marine Chronic PFOA Toxicity Studies,
c MDR Groups identified in Footnote C of Table 3-3.
A Fish
ฆ Invertebrate
ฆ Paracentrotus
A Oryzias (non-definitive, greater than value)
1 10 100
Genus Mean Chronic Value (mg/L PFOA)
Figure 3-4. Estuarine/Marine Genus Mean Chronic Values for PFOA.
3.2 Derivation of the PFOA Aquatic Life Criteria
3.2.1 Derivation of Water Column-based Criteria
3.2.1,1 Derivation of Acute Water Criterion for Freshwater
The acute data set for PFOA contains 19 genera representing all eight taxonomic MDR
groups. GMAVs for the four most sensitive genera were within a factor of 16.1 of each other.
0.9 -
fi
2 0.8 -
ซ 0.6 -
3
s
3 0.5 -
2 0.4 -
Pt
SS 03 "
a
PLh
0.1 -
0.0
0.1
90
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The freshwater Final Acute Value (FAV) (i.e., the 5th percentile of the genus sensitivity
distribution, intended to address 95 percent of the genera) for PFOA is 6.237 mg/L, calculated
using the procedures described in the 1985 Guidelines (Table 3-9). The FAV is lower than all the
GMAVs for the tested species. The FAV was then divided by two to obtain a concentration
yielding a minimal effect acute criterion value. Based on the above, the FAV/2, which is the
freshwater acute criterion water column magnitude (criterion maximum concentration, CMC), is
3.1 mg/L PFOA (rounded to two significant figures) and is expected to be protective of 95% of
freshwater genera exposed to PFOA under short-term conditions of one-hour of duration, if the
one-hour average magnitude is not exceeded more than once in three years (Figure 3-5).
Table 3-9. Freshwater Final Acute Value and Criterion Maximum Concentration.
Calculated iTeshualer I'.
\V based on 4 lowest \allies
olal Number of(i\l.\Ys in Data Set N
(i.MAV
Kniik
(ienus
(mg/l.)
ln((>MAY)
Ih((;ma\ )2
P=U/(N+I)
st| rl( P)
1
Mount
8.885
2.18
4.77
I) o^o
0.224
2
Neocloeon
13.05
2.57
6.60
0.100
0.316
3
Chydorus
93.17
4.53
20.56
0.150
0.387
4
Daphnia
142.8
4.96
24.61
0.200
0.447
ฃ (Sum):
14.25
56.54
0.50
1.37
S2 =
208.28
S = slope
L =
-1.396
L = X-axis intercept
A =
1.831
A = InFAV
FAV =
6.237
P = cumulative probability
CMC =
3.1 mg/L PFOA (rounded to two significant figures)
91
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1.0 T
0.9 --
J ฐ-8
V
2
C 0.7
0)
ซ 0.6
3
= 0.5
S 0.4
a
at
P*
0.2
0.1 --
0.0
Oncorhynchus
~ Litliobates
~ Anaxyrus
~ Ambystoma
ฆ Physella
~ Lepomis
~ Hyla
Piniephales
Danio
~ Neocaridiiia
~ Dugesia
~ Xenopus
Lampsilis
ฆ Ligumia
~ Brachionus
~ Daplinia
~ Chydoms
~ Neocloeon
~ Moina
ฆ
Invertebrate (Mollusk)
~
Invertebrate (Other)
~
Insect
Fish
~
Amphibian
-CMC
10 100 1,000
Genus Mean Acute Value (mg/L PFOA)
10,000
Figure 3-5. Ranked Freshwater Acute PFOA GMAVs and CMC.
3.2.1.2 Derivation of Acute Water Criterion for Estuarine/Marine Water
The 1985 Guidelines state that data from a minimum of eight families are needed to
calculate an estuarine/marine FAV. Insufficient data exist to fulfill all eight of the taxonomic
MDR groups. Notably, no acceptable test data on fish species were available. Since data were
available for only three families, an estuarine/marine FAV could not be derived, and
consequently, the EPA cannot derive an estuarine/marine acute criterion. The EPA has, however,
developed an estuarine/marine acute benchmark value using available empirical data and the
EPA/ORD's web-ICE tool to estimate missing data. The acute estuarine/marine benchmark is
provided in Appendix L.
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3.2.1.3 Derivation of Chronic Water Criterion for Freshwater
The freshwater Final Chronic Value (FCV) (i.e., the 5th percentile of the genus sensitivity
distribution, intended to address 95 percent of the genera) for PFOA is 0.1041 mg/L, calculated
using the procedures described in the 1985 Guidelines (Table 3-10). The freshwater chronic
criterion water column magnitude (CCC) is the FCV rounded to two significant figures, or 0.10
mg/L PFOA, and is expected to be protective of 95% of freshwater genera potentially exposed to
PFOA if the four-day average magnitude is not exceeded more than once in three years.
Table 3-10. Freshwater Final Chronic Value and Criterion Continuous Concentration.
Calculated iTcshualci' 1 CV based on 4 lowest \allies
olal Number of(i\ICVs in Data Set 12
(i.MCV
Uank
(ion us
(mg/l.)
in((;M( \)
in((;M( \ )2
P=U/(N+I)
sqrl(P)
1
Hyalella
0.147
-1.917
3.676
0.077
0.277
2
Lithobates
0.288
-1.245
1.550
0.154
0.392
3
Daphnia
0.3695
-0.996
0.991
0.231
0.480
4
Brachionus
0.7647
-0.268
0.072
0.308
0.555
ฃ (Sum):
-4.43
6.29
0.77
1.70
S2 =
32.54
S = slope
L =
-3.538
L = X-axis intercept
A =
-2.262
A = InFCV
FCV =
0.1041
P = cumulative probability
CCC =
0.10 mg/L PFOA (rounded to two significant figures)
93
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1.0
0.9 --
a
.2 0.8 4
u
C3
"-1 0.7 4
&H
o.i 4
o.o
ฆ
Invertebrate
A
Insect
Fish
~
Amphibian
-CCC
o.oi
Chironomus a
Pimephales (non-definitive, greater than value)
Oncorhynchus (non-definitive, greater than value)
Gobiocypris (non-definitive, greater than value)
Ceriodaphnia ฆ
Oryzias
^ Neocloeon (non-definitive,
greater than value)
ฆ Moina
Brachionus
ฆ Daphnia
~ Lithobates
Hyalella
-+-
-+-
0.1 1 10
Genus Mean Chronic Value (mg/L PFOA)
100
Figure 3-6. Freshwater Quantitative GMCVs and CCC.
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3,2,1,4 Deriving A Protective Duration Component of the Water Column-Based Chronic
Criterion
The EPA 1985 Guidelines set the standard chronic duration at four days for two primary
reasons. The 1985 Guidelines state, "An averaging period offour days seems appropriate for use
with the CCC for two reasons. First, it is substantially shorter than the 20 to 30 days that is
obviously unacceptable. Second, for some species it appears that the results of chronic tests are
due to the existence of a sensitive life stage at some time during the test
Among tests with chronically sensitive genera, Bartlett et al. (2021) measured effects of
PFOA on H. azteca (the most chronically sensitive GMCV) survival at 7, 14, 21, 28, 35, and 42
days and concluded, "Toxicity increased approximately two-fold over the duration of exposure,
with LCsos of 110 mg/L after seven days and 51 mg/L after 42 days." Based on Table S6 of
Bartlett et al. (2021), LC50 values decreased from seven to 21-days and remained generally stable
from 28 to 42 days, suggesting the chronic PFOA LC50 value became time-independent between
21 and 28 days. Although//, azteca survival was tolerant at seven days, clear time-dependent
toxicity may not occur for more sensitive endpoints such as reproduction.
Bartlett et al. (2021) determined effects to reproduction to be more sensitive than long-
term survival, but only measured effects to reproduction after 42 days of exposure, which is
substantially longer than the four-day chronic duration. Bartlett et al. (2021) noted, effects to
amphipod reproduction are typically the result of effects to growth under the premise, "larger
amphipods have a greater reproductive output" and "reduced growth delays sexual maturity."
However, results observed in Bartlett et al. (2021), suggested "the effects on reproduction may
also have occurred independently of growth." Therefore, the reproductive-specific effects
observed by Bartlett et al. (2021) may not have been caused by the long-term effects of reduced
growth but were possibly the result of a sexually-developing and uniquely-sensitive life stage
95
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that existed during a relatively brief duration within the longer 42-day test. Such instances are
among the two primary reasons why the 1985 Guidelines prescribed the standard four-day
chronic duration.
PFOA effects observed for other chronically sensitive and short-lived species further
suggests a four-day chronic duration was appropriate. For example, the SMCV for Brachionus
calyciflorus and the Brachionus GMCV (fourth most sensitive genus) are both the geometric
mean of a full life-cycle test (i.e., up to 200 hours) test by Zhang et al. (2013a) and a four-day
test by Zhang et al. (2014b). Chronic values for both studies correspond to the effect on
population intrinsic growth rate. The full life-cycle test lasted up to 200 hours and yielded an
EPA-calculated ECio of 0.5015 mg/L. The four-day test yielded an EPA-calculated ECio of
1.166 mg/L, suggesting chronic PFOA effects to short-lived species may occur after four days
and increase with exposure duration.
Similarly, the SMCV for Moina macrocopa and the Moina GMCV (fifth most sensitive
genus) are based on a seven-day life-cycle test by Ji et al. (2008), which also suggested
reproductive effects (endpoint of number of young per starting female) occur in as little as seven
days.
Overall, no chronic PFOA toxicity tests systematically evaluated time-to-effect, reported
effect data at time intervals at a high enough resolution to model the speed of toxic action,
assessed time variable PFOA exposures, or assessed the potential for latent toxicity. However,
chronic tests, including life cycle tests with relatively short-lived species suggest chronic effects
may occur at durations shorter than those of standard toxicity tests (e.g., 28 days) and a chronic
four-day duration component of the water column criterion was considered protective for these
species/genera. Therefore, the EPA has set the duration component of the PFOA chronic water
96
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column criterion at four days to reflect the chronic criterion duration recommended in the 1985
Guidelines. This four-day duration component of the chronic water column criterion is also
consistent with U.S. EPA (1991), which considered the default four-day chronic averaging
period as "the shortest duration in which chronic effects are sometimes observed for certain
species and toxicants," and concludes that four-day averaging "should be fully protective even
for the fastest acting toxicants."
3,2,1.5 Derivation of Chronic Water Criterion for Estuarine/Marine Water
The 1985 Guidelines recommend that data from a minimum of eight families are needed
to calculate an estuarine/marine FCV or sufficient data to transform an estuarine/marine FAV to
a CCC through the use of an ACR. Insufficient data exist to fulfill all eight of the taxonomic
MDR groups. There are only two quantitatively acceptable GMCVs for estuarine/marine genera
and no estuarine/marine ACRs. Consequently, the EPA could not derive an estuarine/marine
chronic criterion (see Appendix L for derivation of acute estuarine/marine benchmarks).
3,2,2 Derivation of Tissue-Based Criteria
Chronic PFOA toxicity data with measured tissue concentrations were limited. There
were no aquatic life tissue-based toxicity studies considered acceptable for quantitative use.
Therefore, there were not sufficient data to derive chronic tissue criteria using a sensitivity
distribution approach. Instead, the water column chronic criterion was transformed into
corresponding tissue-based criteria through a BAF approach, as outlined in Section 2.11.3. The
chronic PFOA tissue-based criteria were derived by translating the chronic freshwater column
criterion (i.e., 0.10 mg/L; see equation 1 in Section 3.2.1.3) into corresponding tissue-based
criteria. The resulting tissue criterion corresponded to the tissue type from the 20th percentile
BAF used in the equation (see Section 2.11.3). The 20th centile BAF was used to derive tissue-
97
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based criteria as a relatively conservative BAF estimate in order to protect species across taxa
and across water bodies with variable bioaccumulation conditions. That is, use of the 20th centile
BAF protects species and conditions where the bioaccumulation of PFOA and resultant tissue-
based exposures is relatively low as well as those conditions with the bioaccumulation potential
of PFOA is relatively high.
3,2,2,1 PFOA Bioaccumulation Factors (BAFs)
Section 2.11.3.1 above summarizes the literature search, calculation, and evaluation of
the PFOA BAFs for aquatic life. These BAFs were compiled by and can be found in Burkhard
(2021). BAFs used in the derivation of the PFOA tissue-based criteria consisted of two or more
water and organism samples each and were collected within one year and 2 km distance of one
another. In order to derive more protective tissue criteria across and within water bodies, the
distributions of BAFs used to derive tissue criteria were based on the lowest species-level BAF
reported at a site. When more than one BAF was available for the same species within the same
waterbody, the species-level BAF was calculated as the geometric mean of all BAFs for that
species at that site. Summary statistics for the PFOA BAFs used in derivation of the tissue-based
criteria are presented in Table 3-11 and individual BAFs are provided in Appendix O.
Table 3-11. Summary Statistics for PFOA BAFs in Invertebrate Tissues and Various Fish
Tissues1.
20"'
(ieomelric
Median
Cenlile
Mean liAl
UAI-"
HA I-"
Mini in ii in
Maximum
(1 ./kg-w el
(1 ./kg-w el
(1 ./kg-w el
(1 ./kg-we!
(1./kg-w el
Category
n
weight)
weigh!)
weight)
weigh!)
weigh!)
Invertebrates
21
105.3
84.8
11.76
0.985
9,680
Fish Muscle
17
7.152
7.94
1.331
0.292
656
Fish Whole Body
25
198.6
219
64.93
1
16,273
1 Based on the lowest species-level BAF measured at a site (i.e., when two or more BAFs were available for the
same species at the same site, the species-level geometric mean BAF was calculated, and the lowest species-level
BAF was used to represent that site).
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3,2,2.2 Deriving Tissue-Based Criteria from the Chronic Water Column Criterion
Invertebrate whole-body, fish whole-body, and fish muscle tissue criteria were derived
separately by multiplying the freshwater chronic water column criterion by the respective 20th
centile of the distribution of BAFs described in Section 3.2.2.1, using Equation 1 from Section
2.11.3. The use of a 20th centile BAF results in more protective tissue criteria than those derived
from a BAF based on a central tendency measure (e.g., geometric mean or median), which would
only be protective on average or approximately 50% of the time.
The invertebrate whole-body tissue chronic criterion was calculated by multiplying the
20th centile BAF of 11.76 L/kg wet weight and the PFOA freshwater chronic water criterion of
0.10 mg/L, which resulted in a chronic invertebrate whole-body tissue criterion of 1.18 mg/kg
wet weight. The fish muscle tissue chronic criterion was calculated by multiplying the 20th
centile BAF of 1.331 L/kg wet weight and the PFOA freshwater chronic water criterion of 0.10
mg/L, which resulted in a chronic fish muscle-based chronic criterion of 0.133 mg/kg wet
weight. The fish whole-body tissue chronic criterion was calculated by multiplying the 20th
centile BAF of 64.93 L/kg wet weight and the PFOA freshwater chronic water criterion of 0.10
mg/L, which resulted in a chronic fish whole-body tissue criterion of 6.49 mg/kg wet weight.
The chronic tissue-based criteria are expected to be protective of 95% of freshwater genera
potentially exposed to PFOA under long-term exposures if the tissue-based criteria magnitudes
are not exceeded.
The EPA acknowledges that there is uncertainty in deriving protective tissue criteria
magnitudes by transforming the chronic water column criterion (which was based on tests that
only added PFOA to the water column) into tissue concentrations through field-measured
bioaccumulation data of paired water and tissue concentrations in waterbodies. Nevertheless, the
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chronic water column criterion is based on chronic toxicity tests that fed test organisms. In these
tests, PFOA can directly affect species based on direct water column exposure and/or sorb to
added food that is consumed by test organisms before eliciting chronic effects from dietary
exposure. Therefore, the chronic water column criterion magnitude accounts for water column-
based and, to a possible lesser extent, dietary-based effects, while the field-based BAFs account
for water column-based and dietary-based PFOA exposure in tissues.
The tissue criteria will provide information to states, Tribes, and stakeholders on potential
effects to aquatic organisms based on aquatic tissue monitoring data. No available quantitatively
acceptable data on the effects of dietary exposures to aquatic species were available, thus the
EPA elected to develop protective values for aquatic organism tissues based on the observed
relationship between water column concentrations and tissue concentrations and observed PFOA
toxicity in chronic tests where PFOA was only added directly to the water column.
3.2.2,3 Deriving Protective Duration and Exceedance Frequencies for the Tissue-based
Chronic Criteria
3.2.2.3.1 Duration: Chronic Tissue-Based Criteria
PFOA concentrations in tissues are generally expected to change only gradually over
time in response to environmental fluctuations. The chronic tissue-based criteria averaging
periods, or duration components, were therefore specified as instantaneous, because tissue data
provide point, or instantaneous, measurements that reflect integrative accumulation of PFOA
over time and space in population(s) at a given site.
3.2.2.3.2 Frequency: Chronic Tissue-Based Criteria
The PFOA tissue-based criteria frequencies are set as "not to be exceeded" to ensure
protection of aquatic life populations. This frequency accounts for the many variables
influencing ecological recovery and the uncertainty in PFAS-specific ecological recovery time
due to the lack of case studies on recovery rates following elevated PFOA concentrations in
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aquatic biota. Ecological recovery times following chemical disturbances are context-dependent,
being largely dependent on: (1) biological variables such as the presence of nearby source
populations or generational time of taxa affected, (2) physical variables such as lentic and lotic
habitat considerations where recovery rates in lentic systems may be slower than lotic systems
where the pollutant may be quickly flushed downstream, and (3) chemical variables such as the
persistence of a chemical and potential for residual effects.
PFOA-specific case studies are unavailable to directly inform rates of ecological recovery
following elevated concentrations in fish and aquatic invertebrates. Metals and other chemical
pollutants may be retained in the sediment and biota, where they can result in residual effects
over time that further delay recovery. Few studies are available concerning PFOA elimination or
depuration half-life in aquatic animals, however the data that exist indicate a short half-life. For
example, the elimination half-life for PFOA in adult rainbow trout exposed to PFOS for 28 days
via the diet followed by 28 days depuration was estimated to be seven days in muscle tissue
(Falk et al. 2015), while the terminal half-life in rainbow trout receiving a one-time intra-arterial
injection of PFOA was 12.6 days (Consoer et al. 2014). Additionally, the depuration half-life in
northern leopard frog tadpoles via 40-day aqueous exposure to 0.10 mg/L PFOA was estimated
to be only 2.6 days (Hoover et al. 2017). It's unclear whether PFOA half-life in aquatic organism
tissues is the mechanistic result of rapid depuration via gills or an artifact of these measurements
that are taken during relatively short testing times (e.g., 28 days) where a steady state condition
between PFOA and water and tissues has not occurred. Long-term uptake and subsequent
excretion rates of PFOA have been extensively studied in humans relative to aquatic life. For
example, Li et al. (2017) reported a median PFOA half-life of 2.7 years in human serum
101
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following exposure to PFOA in drinking water, which authors stated was in the range of
previously published estimates.
Due to chemical retention in both the environment and tissues, ecosystems impacted by
discharges of bioaccumulative pollutants (such as selenium) recover from chemical disturbances
at relatively slow rates. For example, Lemly (1997) concluded that although water quality in
Belews Lake, North Carolina, (a freshwater reservoir) had recovered significantly in the decade
since selenium discharges were halted in 1985, the threat to fish had not been eliminated. The
effects of selenium that led to severe reproductive failure and deformities in fish, were still
measurable (as fish deformities) in 1992 (seven years later) and in 1996 (ten years later). Lemly
(1997, pg. 280) estimated based on these data that"the timeframe necessary for complete
recovery from selenium contamination from freshwater reservoirs can be on the order of
decades
Beyond bioaccumulation, chemical-specific considerations such as degradation vs.
persistence may also provide a mechanism influencing ecological recovery rates. The persistence
of PFOA has been attributed to the strong C-F bond, with no known biodegradation or abiotic
degradation processes for PFOA (see Section 2.3). Similarly, metals do not degrade and may
persist in aquatic systems following elevated discharge. The persistence of metals may explain
why metals had the second longest median recovery time of any disturbance described in a
systematic review of aquatic ecosystem recovery (Gergs et al. 2016). Gergs et al. (2016) showed
recovery times following metal disturbances ranged from roughly six months to eight years
(median recovery time = one year; 75th centile ~ three years; n = 20). Unlike metals, however,
PFOA is not naturally-occurring and aquatic organisms possess no evolved detoxification
mechanisms to aid in recovery at the individual level. Furthermore, the degradation of other
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PFAS into PFOA may represent a continued source of PFOA in aquatic systems that further
delays recovery.
The persistence and bioaccumulative/human-made nature of PFOA in aquatic systems, in
combination with the documented recovery times of pollutants with similar chemical attributes
(Lemly 1997; Gergs et al. 2016; Mebane 2022), suggests aquatic systems may recover from
PFOA tissue criteria exceedances on the order of five to 10 years. However, recovery times
could be longer if the source of PFOA and other PFAS that degrade into PFOA is not removed.
There is a large amount of uncertainty in specifying a time interval associated with ecological
recovery from PFOA tissue criteria exceedances given the lack of PFOA-specific examples of
ecological recovery and the many situational-specific factors influencing recovery (Mebane et al.
2020). For example, the lack of PFOA degradation in the environment, and the fact that other
PFAS in the environment can degrade into PFOA could act as ongoing PFOA sources that
further delay recovery. Given these uncertainties, the PFOA tissue-based criteria frequencies are
set as "not to be exceeded" to ensure protection of aquatic life populations. Moreover, if tissue-
based criteria were exceeded, then PFOA has likely built up through the food web and PFOA
source reservoirs are likely to exist, representing a broad level of PFOA contamination
throughout the aquatic ecosystem.
The "not to exceed" frequency components of the tissue-based criteria do not suggest
aquatic ecosystems could never recover from PFOA tissue-based criteria exceedance, under the
right conditions. For example, ecological recovery from such an exceedance could begin once
PFOA sources are decreased or eliminated, PFOA source reservoirs existing within the
ecosystem (including other PFAS that degrade into PFOA), became permanently isolated from
possible uptake by the ecosystem (e.g., long-term burial with no benthic disturbances), and
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unaffected organisms were able to repopulate the system through immigration and/or
reproductive events producing generations that are no longer exposed to PFOA.
Evaluation of PFOA concentrations in tissues would likely include evaluating the central
tendency of tissue concentrations of samples for a given species, collected at a specific site and
time. Considering a measure of central tendency to assess tissue-based exposures in the field
relative to the criteria is appropriate because the criteria are intended to protect aquatic life
populations.
3.3 Summary of Freshwater PFOA Aquatic Life Criteria and the Acute
Estuarine/Marine Benchmark
This Aquatic Life Ambient Water Quality Criteria for PFOA document includes water
column based acute and chronic criteria and tissue-based criteria for fresh waters. Acute and
chronic water column criteria magnitudes for estuarine/marine waters could not be derived at this
time due to data limitations; however, an acute estuarine/marine benchmark is provided for
states/authorized tribal consideration (see Appendix L). The freshwater acute water column-
based criterion magnitude is 3.1 mg/L, and the chronic water column-based chronic criterion
magnitude is 0.10 mg/L. The fish whole-body tissue criterion magnitude is 6.49 mg/kg wet
weight, the fish muscle tissue criterion magnitude is 0.133 mg/kg wet weight and the invertebrate
whole-body tissue criterion magnitude is 1.18 mg/kg wet weight (Table 3-12). The assessment of
the available data for fish, invertebrates, amphibians, and plants indicates these criteria will
protect the freshwater aquatic community.
The freshwater chronic water column criterion is more strongly supported than the
chronic tissue-based criteria because the water column-based chronic criterion was derived
directly from the results of empirical toxicity tests. The chronic tissue-based criteria are
relatively less certain because they were derived by transforming the chronic water column
104
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criterion into tissue concentrations through BAFs, with any uncertainty and variability in the
underlying BAFs then propagating into the resultant tissue-based criteria magnitudes.
Table 3-12. Recommended Perfluorooctanoic acid (PFOA) Ambient Water Quality
Criteria
'or the Protection of Aquatic Life in Freshwater.
Typo/Modi a
Acute \Ysiler
Column ((M( )'4
Chronic \\ aler
Column (( ( ( )'5
Inverlebrale
Wliole-
liodv12
lisli
\\ liolo-
liodv12
Fish Muscle12
Magnitude
3.1 mg/L
0.10 mg/L
1 1S mu kg
WW
(> -N mu kg
WW
ii 133 mu kg
WW
Duration
One-hour average
Four-day average
Instantaneous3
Frequency
Not to be exceeded
more than once in
three years on
average
Not to be exceeded
more than once in
three years on
average
Not to be exceeded6
1 All five of these water column and tissue criteria are intended to be independently applicable and no one criterion takes
primacy. All of the above recommended criteria (acute and chronic water column and tissue criteria) are intended to be
protective of aquatic life. These criteria are applicable throughout the year.
2 Tissue criteria derived from the chronic water column concentration (CCC) with the use of bioaccumulation factors and are
expressed as wet weight (ww) concentrations.
3 Tissue data provide instantaneous point measurements that reflect integrative accumulation of PFOA over time and space in
aquatic life population(s) at a given site.
4 Criterion Maximum Concentration; applicable throughout the water column.
5 Criterion Continuous Concentration; applicable throughout the water column.
6 PFOA chronic freshwater tissue-based criteria should not be exceeded, based on measured tissue concentrations representing
the central tendency of samples collected at a given site and time.
This Aquatic Life Ambient Water Quality Criteria and Acute Saltwater Benchmark for
PFOA document includes a water column based acute benchmark for estuarine/marine waters.
The derivation of this benchmark is described in detail in Appendix L. The saltwater acute
benchmark 7.0 mg/L (magnitude component), expressed as a one-hour average (duration
component), that is not to be exceeded more than once in three years on average (Table 3-13).
Aquatic life benchmarks, developed under 304(a)(2) of the CWA, are informational
values that the EPA generates when there are limited high quality toxicity data available and data
gaps exist for several aquatic organism families. The EPA develops aquatic life benchmarks to
provide information that states and Tribes may consider in their water quality protection
programs. In developing aquatic life benchmarks, data gaps may be filled using new approach
105
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methods (NAMs), such as computer-based toxicity estimation tools (e.g., the EPA's Web-ICE)
or other new approach methods intended to reduce reliance on additional animal testing
(https://www.epa.gov/chemical-research/epa-new-approach-methods-work-plan-reducing-use-
vertebrate-ani mal s-chemi cal). including the use of read-across estimates based on other
chemicals with similar structures. The EPA's aquatic life benchmark values are not regulatory,
nor do they automatically become part of a state's water quality standards.
Table 3-13. Acute Perfluorooctanoic Acid (PFOA) Benchmark for the Protection of
Aquatic Life in Estuarine/Marine Waters.
Type/Media
Acute Wnlor Column Benchmark
Magnitude
7.0 mg/L
Duration
One hour average
Frequency
Not to be exceeded more than once in three years on average
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4 EFFECTS CHARACTERIZATION FOR AQUATIC LIFE
This section describes supporting information for the derivation of these PFOA aquatic
life criteria. Specifically, this chapter: (1) assesses the influence of including non-North
American resident species in criteria derivation (Section 4.1), (2) considers relatively sensitive
toxicity data from qualitatively acceptable studies that were used as supporting information
(Section 4.2), (3) describes the available PFOA ACRs (Section 4.3), (4) compares the tissue-
based criteria magnitudes to the empirical tissue-based effect concentrations available (Section
4.4), (5) evaluates aquatic plant tolerance to PFOA exposures (Section 4.5), and (6) evaluates the
tolerance of threatened and endangered species to PFOA exposures (Section 4.6).
4.1 Influence of Using Non-North American Resident Species on PFOA
Criteria
The EPA conducted an additional analysis of the water column-based criteria by limiting
the toxicity datasets to organisms that are residents to the conterminous United States that have
established populations to evaluate the influence of including non-North American resident
species in criteria derivation.
4.1.1 Freshwater Acute Water Criterion with Resident Organisms
Three species, the green neon shrimp (Neocaridina denticulata), the cladoceran (Daphnia
carinata) and the planarian (Dugesia japonica), are not resident or reproducing in the
conterminous United States, while it remains uncertain if there are established resident zebrafish
(Danio rerio) populations in the conterminous United States (U.S. FWS 2018). Nevertheless,
zebrafish are common ecotoxicity test organisms that serve as taxonomic surrogates for untested
fish species and are also considered in effects assessments conducted under the Toxic Substances
Control Act (TSCA) and the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA).
Removal of the green neon shrimp, Daphnia carinata, and Dugesia japonica, while retaining
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zebrafish truncated the freshwater acute dataset to 24 species representing 17 genera (Table 4-1).
The freshwater acute dataset truncated to North American resident species only was missing one
MDR group (a benthic crustacean). The Daphnia carinata SMAV was the third most sensitive;
however, its removal along with the removal of the other non-North American resident species
had limited impact on the exploratory FAV and subsequent acute water column concentration
(Table 4-2). The acute water column concentration based on North American resident species
only, including zebrafish, was 2.6 mg/L PFOA, which was slightly lower than the chronic
criteria CMC (CMC of 3.1 mg/L) based on both North American and non-North American
species. Had zebrafish also been removed from the exploratory FAV and acute water column
concentration based on North American resident species only, the four most sensitive genera
would have remained the same, but the number of genera in the dataset would have decreased by
one, which causes the resulting exploratory FAV (i.e., 4.793 mg/L) and acute water column
concentration (i.e., 2.4 mg/L) to decrease slightly. The exploratory FAV and CMC based on
North American resident species only with zebrafish excluded were both similar to the FAV and
CMC described in Section 3.2.1.1.
Table 4-1. Ranked Freshwater Genus Mean Acute Values with North American Resident
Organisms.
Uank11
(/MAY
(ing/l. PI OA)
MI)U
Croup1
(ienus
Species
SMAV1'
(mg/L PI OA)
1
8.885
D
Moina
Cladoceran,
Moina macrocopa
166.3
Cladoceran,
Moina micrura
0.4747
2
13.05
F
Neocloeon
Mayfly,
Neocloeon triangulifer
13.05
3
93.17
D
Chydorus
Cladoceran,
Chydorus sphaericus
93.17
4
150.0
H
Brachionus
Rotifer,
Brachionus calyciflorus
150.0
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Ksiuk"
Ci.MAV
(ing/l. PI OA)
MI)U
(roup1'
(CI1IIS
Species
SMAY1'
(ing/l. PI OA)
5
161.0
G
Ligumia
Black sandshell,
Ligumia recta
161.0
6
164.4
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
164.4
7
208.7
D
Daphnia
Cladoceran,
Daphnia magna
213.9
Cladoceran,
Daphnia pulicaria
203.7
8
377.0
C
Xenopus
Frog,
Xenopus sp.
377.0
9
450.4
B
Danio
Zebrafish,
Danio rerio
450.4
10
593.6
B
Pimephales
Fathead minnow,
Pimephales promelas
593.6
11
646.2
C
Hyla
Gray treefrog,
Hyla versicolor
646.2
12
664.0
B
Lepomis
Bluegill,
Lepomis macrochirus
664.0
13
681.1
G
Physella
Bladder snail,
Physella acuta
681.1
14
689.4
C
Ambystoma
Jefferson salamander,
Ambystoma jeffersonianum
1,070
Small-mouthed salamander,
Ambystoma texanum
407.3
Eastern tiger salamander,
Ambystoma tigrinum
752.0
15
793.9
C
Anaxyrus
American toad,
Anaxyrus americanus
793.9
16
951.5
C
Lithobates
American bullfrog,
Lithobates catesbeiana
1,020
Green frog,
Lithobates clamitans
1,070
Northern leopard frog,
Lithobates pipiens
751.7
Wood frog,
Lithobates sylvatica
999
17
4,001
A
Oncorhynchus
Rainbow trout,
Oncorhynchus mykiss
4,001
a Ranked from the most sensitive to the most tolerant based on Genus Mean Acute Value,
b From Appendix A: Acceptable Freshwater Acute PFOA Toxicity Studies,
c MDR Groups identified in Footnote C of Table 3-3.
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Table 4-2. Freshwater Exploratory Final Acute Value and Acute Water Column
Concentration with North American Resident Organisms (zebrafish included).
Calculated iTeshualer 1 AY based on 4 lowest \allies
olal Number of(i\l AYs in Dala Set 17
(/MAY
Uank
(ienus
(mg/l.)
ln((;.\l.\Y)
Ih((;ma\ )2
P=U/(N+I)
s(|ii(l*)
1
Muina
8.885
2.18
4.77
U.U59
U.243
2
Neocloeon
13.05
2.57
6.60
0.118
0.343
3
Chydorus
93.17
4.53
20.56
0.176
0.420
4
Brachionus
150.0
5.01
25.11
0.235
0.485
ฃ (Sum):
14.30
57.04
0.59
1.49
S2 =
181.32
S = slope
L =
-1.444
L = X-axis intercept
A =
1.567
A = InFAV
FAV =
4.793
P = cumulative probability
CMC =
2.4 mg/L PFOA (rounded to two significant figures)
4,1,2 Freshwater Chronic Water Criterion with North American Resident Organisms
Three species, the cladoceran (Daphnia carinata), the rare minnow (Gobiocypris rarus)
and the medaka (Oryzias latipes), are not resident or reproducing in the conterminous United
States. Removal of these species truncated the freshwater chronic dataset to ten species
representing ten genera (Table 4-3). The freshwater chronic dataset truncated to North American
resident species fulfilled all MDR groups. Calculating the FCV based on the chronic GSD
comprised of North American resident species only resulted in an exploratory FCV of 0.05857
mg/L and a chronic water column criterion of 0.059 mg/L (rounded to two significant figures;
Table 4-4). The exploratory chronic water column concentration CCC (0.059 mg/L PFOA) based
on North American species only was about half of the CCC based on North American and non-
North American species (0.10 mg/L PFOA) and is lower than all of the quantitatively-acceptable
GMCVs (Table 3-7). The reduction in the exploratory FCV based on North American resident
species only was primarily an artifact of the FCV calculation procedure rather than inclusion of
more sensitive toxicity data. That is, the reduced "//" used in the exploratory criterion calculation
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and the increase in the Daphnid GMCV (which was the result of excluding I), carinata SMCV),
increased the slope of the GSD which decreased the extrapolated FCV. The EPA retained the
chronic water column criterion which includes North American and non-North American
species, with a magnitude of 0.10 mg/L, to ensure the fullest, high-quality dataset available is
used to represent the thousands of untested aquatic taxa present in U.S. ecosystems when
deriving the chronic criterion for PFOA.
Table \
-3. Ranked Freshwater Genus Mean Chronic Values with Resident C
)rganisms.
Unnk11
OK Vh
(ing/l. PI OA)
MI)U
Croup1'1
(ienus
Species
SMCV1'
(ing/l. PI OA)
1
0.147
E
Hyalella
Amphipod,
Hyalella azteca
0.147
2
0.288
C
Lithobates
American bullfrog,
Lithobates catesbeiana
0.288
3
0.7647
G
Brachionus
Rotifer,
Brachionus calyciflorus
0.7647
4
2.194
D
Moina
Cladoceran,
Moina macrocopa
2.194
5
>3.085
H
Neocloeon
Mayfly,
Neocloeon triangulifer
>3.085
6
4.317
D
Daphnia
Cladoceran,
Daphnia magna
4.317
7
23.56
D
Ceriodaphnia
Cladoceran,
Ceriodaphnia dubia
23.56
8
>40
A
Oncorhynchus
Rainbow trout,
Oncorhynchus mykiss
>40
9
>76
B
Pimephales
Fathead minnow,
Pimephales promelas
>76
10
88.32
F
Chironomus
Midge,
Chironomus dilutus
88.32
a Ranked from the most sensitive to the most resistant based on Genus Mean Chronic Value,
b From Appendix C: Acceptable Freshwater Chronic PFOA Toxicity Studies,
c MDR Groups identified in Footnote C of Table 3-3.
Ill
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Table 4-4. Freshwater Exploratory Final Chronic Value and Chronic Water Column
Concentration with North American Resident Organisms.
Calculated I'lvshwalcr 1 CV Ixiscd on 4
ouesl \allies
Tolal \ ii in her of (iMCVs in
Data Set l<)
ok v
Kniik
(iCIlllS
(mg/l.)
ln((;\l( V)
in((;M( \ )2
P=U/(N+I)
sqrKP)
1
llyulcllu
U. 147
-1.917
3.676
U.U91
U.3U2
2
Lithobates
0.288
-1.245
1.550
0.182
0.426
3
Brachionus
0.7647
-0.268
0.072
0.273
0.522
4
Moina
2.194
0.786
0.617
0.364
0.603
ฃ (Sum):
-2.64
5.91
0.91
1.85
S2 =
82.45
S = slope
L =
-4.868
L = X-axis intercept
A =
-2.838
A = InFCV
FCV =
0.05857
P = cumulative probability
CCC =
0.059 mg/L PFOA (rounded to two significant figures)
4.2 Consideration of Relatively Sensitive Qualitatively Acceptable Water
Column-Based Toxicity Data
A multitude of studies were identified as not meeting the EPA's data quality guidelines
for inclusion in the criteria derivation; however, these studies were used qualitatively as
supporting information to the PFOA criteria derived to protect aquatic life and provide additional
evidence of the observed toxicity and effects of PFOA, including the relative sensitivities. Most
of these studies produced relatively tolerant effect concentrations relative to the criteria and are
listed in Appendix G. The key studies used qualitatively in the derivation of the PFOA criteria
were identified as being from either relatively sensitive genera or relatively sensitive tests and
are described below. That is, qualitatively acceptable data for tests with species among the four
most sensitive genera that were used to derive the criteria magnitudes are discussed.
Qualitatively acceptable tests with relevant exposure durations and apical effects that also
observed effect concentrations less than or similar to (e.g., factor of two) the corresponding
criteria magnitude are discussed. Qualitatively acceptable studies described below were
separated by acute (Section 4.2.1) and chronic (Section 4.2.2) data and only included those
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studies that reported apical endpoints. The toxicity values summarized as part of this Effects
Characterization were not used in any quantitative analyses or in the numerical derivation of the
PFOA aquatic life criteria.
4.2.1 Consideration of Qualitatively Acceptable Acute Data
4,2,1.1 Qualitatively Acceptable Acute Data for Species Among the Four Most Sensitive
Genera Used to Derive the Acute Water Column Criterion
4.2.1.1.1 Most acutely sensitive genus, Moina
There were no qualitatively acceptable acute tests for species within the genus, Moina.
4.2.1.1.2 Second most acutely sensitive genus, Neocloeon
There were no qualitatively acceptable acute tests for species within the genus,
Neocloeon.
4.2.1.1.3 Third most acutely sensitive genus, Chydorus
There were no qualitatively acceptable acute tests for species within the genus, Chydorus.
4.2.1.1.4 Fourth most acutely sensitive genus, Daphnia
3M Co. (2000a) exposed D. magna to PFOA (CAS # 335-67-1) in a 48-hour static,
unmeasured acute toxicity test that followed USEPA-TSCA Guideline 797.1300. The toxicant
was part of the 3M production lot number 269 and was characterized as mixture of PFOA (96.5-
100% of the compound) and C6, C7 and C9 perfluoro homologue compounds (0-3.5% of the
compound). The 48-hour reported EC50, based on death/immobility, was 360 mg/L PFOA. This
test was not acceptable for quantitative use because of possible mixture effects from other
perfluoro homologue compounds in the test substance, the amount of isopropanol and missing
exposure details, but was retained for qualitative use.
3M Co. (2000a) summarized four 48-hour static, unmeasured APFO (CAS # 3825-26-1)
acute toxicity tests with the cladoceran, Daphnia magna and APFO. The toxicant was part of the
3M production lot number 37 and was characterized as a mixture of APFO (96.5-100%) of the
compound) and C6, C7 and C9 perfluoro analogue compounds (0-3.5% of the compound). The
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48-hour EC50 determined from tests conducted in May 1982, based on death/immobility, was
>1,000 mg/L APFO, while the EC50 for a subsequent test in June 1982 was reported to be 126
mg/L APFO. Possible mixture effects of other perfluoro analogue compounds, and the authors
indication that the tests may have included food, did not make these tests acceptable for
quantitative use and they were retained for qualitative use.
3M Co. (2000a) summarized a 48-hour static, unmeasured acute toxicity test with the
cladoceran, Daphnia magna and APFO (CAS # 3825-26-1). The toxicant was part of the 3M
production lot number 390 and was characterized as a mixture of APFO (78-93% of the
compound) and C5, C6 and C7 perfluoro analogue compounds (7-22% of the compound). The
author-reported 48-hour EC50, based on mortality, was 221 mg/L APFO. The possible mixture
effects of APFO with other perfluoro analogue compounds in the test material and possible low
(<80%) test substance purity did not make this test acceptable for quantitative use. This test was
retained for qualitative use only.
3M Co. (2000a) summarized a 48-hour static, unmeasured acute toxicity test with the
cladoceran, Daphnia magna and APFO (CAS # 3825-26-1). The acute test followed USEPA-
TSCA Guideline 797.1300 protocol. The toxicant was part of the 3M production lot number
HOGE 205 and was not sufficiently characterized but was considered a mixture of APFO (30%
of the compound) and water (80% of the compound). The author-reported 48-hour EC50, based
on mortality, was 1,200 mg/L test substance. The authors reported that the test substance was
considered a mixture of APFO (with unreported purity) and other impurities, so the EC50 does
not accurately reflect the toxicity of APFO and therefore the value was not acceptable for
quantitative use but was retained for qualitative use.
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3M Co. (2000a) summarized a 48-hour static, unmeasured acute toxicity test with the
cladoceran, Daphnia magna and APFO (CAS # 3825-26-1). The acute test followed test
guidance from OECD 202. The toxicant was part of the 3M production lot number 2327 and was
characterized as a mixture of APFO (<45% of the compound), water (50% of the compound),
inert perfluorinated compound (<3% of test substance), and Cs and C7 perfluoro analogue
compounds (1-2% of the compound). The author-reported 48-hour EC50, based on
death/immobility, was 584 mg/L test substance. Because of the possible mixture effects of the
inert perfluorinated compounds, other perfluoro analogue compounds and unreported test
substance purity, the test was not acceptable for quantitative use but was retained for qualitative
use.
3M Co. (2000a) summarized a 21-day static-renewal, unmeasured chronic toxicity test
with the cladoceran, Daphnia magna, and APFO (CAS # 3825-26-1) and also briefly described a
corresponding acute test with a reported 48-hour EC50 of 266 mg/L APFO. Very few details were
provided about the acute test methodology. The test compound was assumed to be that of the
chronic test, where the toxicant was part of the 3M production lot number 37 and was
characterized as a mixture of APFO (96.5-100%) of the compound) and C6, C7 and C9 perfluoro
analogue compounds (0-3.5% of the compound). The 48-hour EC50 from this test was not used
quantitatively because of missing study details and the possible presence of additional PFAS, but
the study was retained for qualitative use.
Overall, these D. magna acute effect concentrations were all greater than the FAV (i.e.,
6.237 mg/L) and the acute criterion magnitude (i.e., 3.1 mg/L). These additional data suggest the
Daphnia GMAV (i.e., 142.8 mg/L) used to derive the acute criterion was sufficiently protective.
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4,2,1,2 Consideration of Relatively Sensitive Freshwater Tests based on Qualitatively
Acceptable Data
This section focuses on qualitatively acceptable tests that were most relevant to informing
the appropriateness of the acute freshwater criterion. Specifically, those tests used to
qualitatively inform the acute freshwater criterion magnitude were identified as most relevant if
they met all parameters listed below:
1. reported effect concentrations that were less than or similar to (e.g., factor of two) the
acute criterion magnitude;
2. evaluated an animal species;
3. conducted the test for a relevant acute exposure duration (e.g., -48 hours to -96 hours);
4. evaluated apical effects (i.e., acute mortality/inhibition), and;
5. not already discussed in the previous section (i.e., not a species discussed among the four
most sensitive genera).
The toxicity values summarized below were not used quantitatively to derive the acute PFOA
criterion. Results of each individual study (as well as the rationale why a study was not
quantitatively acceptable) were considered relative to the acute criterion magnitude to ensure the
acute PFOA criterion was not underprotective and to provide additional supporting evidence of
the potential toxicity of PFOA to aquatic organisms.
4.2.1.2.1 Genus: Danio (zebrafish)
Truong et al. (2014) evaluated the sub-chronic effects of 1,060 compounds (U.S. EPA
ToxCast phase 1 and 2) on zebrafish, Danio rerio, through the use of high-throughput
characterization of multidimensional in vivo effects. The effects of APFO and PFOA on
mortality, growth, behavior, morphology, histology and physiology were observed until 120
hours post fertilization (hpf) (114-hour exposure duration) with the water quality conditions not
reported. The most sensitive endpoint was mortality with a reported LOEC of 0.02759 mg/L
APFO. There were no effects of PFOA on mortality for zebrafish embryos with a reported
NOEC of 26.50 mg/L PFOA. This test was not used quantitatively and retained for qualitative
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use only because the exposure durations were too long for an acute test and too short for a
chronic test.
Han et al. (2021) evaluated the bioconcentration factors and phenotypic toxicities of 74
PFAS in zebrafish after five days of exposure. Toxicity tests followed OECD test guidance with
minor modifications. Embryos (64-cell stage) were exposed to one measured PFAS
concentration or DMSO solvent controls with solutions renewed daily. At test termination no
mortality effects were observed at 3.94 |iM PFOA (1.631 mg/L) and 3.67 |iM APFO (1.582
mg/L). This test was not used quantitatively and retained for qualitative use only because the
exposure durations were too long for an acute test and too short for a chronic test with no effects
observed.
Rericha et al. (2021) conducted a 114-hour static unmeasured PFOA toxicity test with
the zebrafish, Danio rerio. The authors reported a 114-hour mortality NOEC of 0.60 |iM, or
0.2484 mg/L PFOA, based on a molecular weight of 414.07 g/mol. This test was not acceptable
for quantitative use as the exposure duration was too long to be an acute test and too short to be a
chronic test. Additionally, the authors only tested one exposure concentration that produced a
relatively low NOEC. Although relatively low, the NOEC of 0.2484 mg/L does not suggest/).
rerio early life stages are sensitive to sub-chronic PFOA exposures.
Haimbaugh et al. (2022) evaluated the acute toxic effects of low-level (<700 ng/L)
PFOA on zebrafish from zero to five days post fertilization. Unmeasured test concentrations (7,
70, or 700 ng/L PFOA) were renewed daily. At test termination, the highest test concentration
(700 ng/L or 0.0007 mg/L) had no effects on mortality or abnormal development. This test was
not used quantitatively and retained for qualitative use only because the exposure durations were
too long for an acute test and too short for a chronic test with no effects observed.
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Yu et al. (2022) evaluated the sub-chronic effects of PFOA on embryonic Danio rerio.
Exposure to PFOA concentrations up to 0.100 mg/L had no effect on growth/length,
mortality/hatch, or survival of larvae. The 70-hour growth, mortality, and hatch NOECs of 0.100
mg/L PFOA were not acceptable for quantitative use because of the short exposure duration,
especially in the context of other available tests for this species that were of a sufficient exposure
duration. Yu et al. (2022) also reported that heartbeat (at 48 hpf) and locomotor behavior (at 120
hpf) were significantly decreased by PFOA at 0.025 mg/L; however, these effects did not
translate to effects on growth, mortality, or hatch at nominal test concentrations that were four
times greater.
Liu et al. (2023b) evaluated the acute effects of PFOA on zebrafish embryos (5 hpf)
exposed to one of five test concentrations (1, 2, 5, 10, and 20 mg/L). Test concentrations were
renewed daily and test concentrations were measured with hatching rate, mortality and deformity
recorded at test termination (96 hpf). The author reported that the 91-hour LCso was
approximately 2 mg/L PFOA. This test was not acceptable for quantitative use because the
exposure duration was not a 96-hour exposure duration and other quantitatively acceptable data
suggests the relatively-sensitive LCso reported here (~2 mg/L) is an outlier, with the Danio
SMAV being 450.4 mg/L.
Wang et al. (2023) evaluated the acute effects of PFOA on zebrafish embryos (2 hpf)
exposed to 5 or 500 |ig/L up to 120 hours post fertilization. Test concentrations were renewed
daily and PFOA concentrations were measured using high-performance liquid chromatography -
tandem mass spectrometry. At 96 hpf, there were no effects observed on mortality or abnormal
development at 491.5 |ig/L (0.4915 mg/L PFOA). Since this exposure was shorter than the
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recommended 96 hour exposure (only 94 hours) and no effects were observed, the test was only
used qualitatively.
Quantitatively acceptable acute tests used to calculate the SMAV (Ye et al. 2007;
Hagenaars et al. 2011; Zhao et al. 2016; Godfrey et al. 2017a; Stengel et al. 2017, 2018)
suggested D. rerio is relatively tolerant to acute PFOA exposures with a I). rerio SMAV of
450.4 mg/L. Another quantitatively acceptable acute test with the zebrafish was conducted by
Corrales et al. (2017). However, the LCso value from this test was excluded from the SMAV
calculation because a comparative assessment between this LCso value and the other seven
quantitatively-acceptable values available indicated the LCso was an outlier, about 5 times lower
than the other lowest LCso value. It is expected that/), rerio will be tolerant to acute PFOA
exposures.
4.2,2 Consideration of Qualitatively Acceptable Chronic Data
4,2,2.1 Qualitatively Acceptable Chronic Data for Species Among the Four Most Sensitive
Genera Used to Derive the Chronic Water Column Criterion
4.2.2.1.1 Most chronically sensitive genus, Hyalella
Kadlec et al. (2024) tested the sub-chronic toxicity of PFOA (96-99% purity) to Hyalella
azteca for seven days in a measured, static-renewal experiment. Testing protocols followed
ASTM methodologies (ASTM 2002) with a shortened exposure period. The test was conducted
with a 0.5x dilution series of measured PFOA concentrations, with four replicates of each
concentration, and -10 organisms per replicate (note: some replicates were overstocked and
included 11 test organisms). Mean test concentrations were 0.53 (control), 5.3, 11, 21, 41 and 83
mg/L. Amphipods were placed in 100 mL glass beakers with two holes (covered by stainless-
steel mesh) on opposite sides and filed with 60 mL of test solution. Beakers also included 2.5 mL
of silica sand substrate to fully cover the bottoms of each chamber and each test chamber was
renewed daily. Little to no mortality was observed in any treatment with survival ranging from
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98-100% at test termination. The author reported EC20 for biomass was 5.0 mg/L. The EPA did
not recalculate an EC10 from this study since the exposure duration was too short for a chronic
exposure. However, this value is much larger than the 42-day exposure (Bartlett et al. 2021) with
an EPA-calculated EC10, based on reproduction, of 0.147 mg/L PFOA.
4.2.2.1.2 Second most chronically sensitive genus, Lithobates
Hoover et al. (2017) tested PFOA (96% purity) toxicity on the northern leopard frog,
Lithobatespipiens (formerly, Ranapipiens) in a chronic renewal toxicity test using measured
PFOA treatment concentrations. The 40-day NOEC was >1.0 mg/L PFOA based on Gosner
stage reached at test termination and snout-vent length. The test used water renewals rather than
the required flow-through design for chronic ALC development; however, leopard frogs
commonly do not tolerate flow-through test systems and the use of renewal system was
appropriate for this study organism. Also, PFOA was detected in the control organisms at
concentrations three orders of magnitude lower than any PFOA treatment groups, indicating the
trace contamination in controls may not be considered a significant issue. The 40-day NOEC of
>1.0 mg/L was classified as acceptable for quantitative use based on meeting data quality
objectives; however, it was not used to derive the chronic criterion because the study showed no
adverse effects at the highest treatment concentrations (i.e., 1.0 mg/L). Because the highest
treatment group that showed no effects was a relatively low treatment concentration, including
this NOEC value in the criterion calculation would have resulted in the criterion magnitude being
influenced by the relatively low-test concentration selected by study investigators (that did not
produce an adverse response), rather than a concentration-response relationship. Therefore, this
test was not used quantitatively and was considered qualitatively acceptable for use in criterion
derivation.
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Flynn et al. (2021) evaluated the chronic effects of PFOA (CAS # 335-67-1, >96%
purity, purchased from Sigma-Aldrich) on northern leopard frogs, Lithobatespipiens (formerly
Ranapipiens), via a 30-day sediment-spiked, static outdoor mesocosm study. At test termination
(30 days) there was no effect on survival or growth (snout-vent length and weight). The 30-day
NOEC, based on survival and growth, was 0.066 mg/L. However, on test-day five and at test
termination all frogs in the spiked sediment mesocosm were less developed, based on Gosner
stage, than the control mesocosms. The study was not acceptable for quantitative use because the
test design was an outdoor spiked sediment mesocosm exposure with algal and zooplankton
communities present and because of the relatively low NOEC value that did not quantitatively
inform criteria derivation based on an exposure-response effect.
Flynn et al. (2022) evaluated the chronic effects of PFOA on the northern leopard frog
[.Lithobates (formerly Rana)pipiens] during a 30-day measured renewal exposure. Animals were
monitored daily for mortality and abnormalities and sampled to quantify snout-vent length
(SVL), body mass, and developmental stage. The SVL and body mass were used to calculate the
scaled mass index (SMI). The NOEC for SVL was >1.376 mg/L and not used quantitatively
because of a lack of a clear dose-response relationship and high variability in organismal
responses despite a 10X dilution series. The NOEC and LOEC for growth (as mass) were 0.1251
and 1.376 mg/L resulting in a MATC of 0.4149 mg/L, which was also not used quantitatively for
the same reasons. The effects of PFOA on SMI and Gosner stage were also not quantitatively
acceptable because, while the highest test concentration was over 100X greater than the lowest,
the response in the highest test concentration did not differ from control responses. Results of
this test were retained for qualitative use.
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Overall, these three studies showed minimal effects to the northern leopard frog at the
concentrations tested, while the indoor laboratory test by Flynn et al. (2019; used to derive the
Lithobates catesbeiana SMCV) showed a -7% reduction in SVL after 72-day exposures at 0.288
mg/L. Although Hoover et al. (2017) reported a NOEC of 1.0 mg/L, the tests only consisted of a
40-day exposure, which may not have been long enough to elicit the chronic effects to SVL
observed by Flynn et al. (2019) after 72 days. For example, Flynn et al. (2019) reported effects
of PFOS on Lithobates catesbeiana tadpole mass after 21, 42, 56, 63, 70, and 72 days, with
PFOS dose-dependent effects only becoming apparent at 56 days. Results of Flynn et al. (2021)
could not meaningfully inform the appropriateness of the chronic criterion or the Lithobates
catesbeiana SMCV (0.288 mg/L) because Flynn et al. (2021) did not observe effects of PFOA
on Lithobates pipiens, with a relatively low NOEC of 0.066 mg/L. Similarly, results of Flynn et
al. (2022) did not suggest the Lithobates GMAV value was underprotective, as Flynn et al.
(2022) did not observe obvious PFOA effects with increasing exposure concentrations.
4.2.2.1.3 Third most chronically sensitive genus, Daphnia
3M Co. (2000a) summarized a 21-day static-renewal, unmeasured chronic toxicity test
with the cladoceran, Daphnia magna, and APFO (CAS # 3825-26-1). The toxicant was part of
the 3M production lot number 37 and was characterized as a mixture of APFO (96.5-100% of the
compound) and C6, C7 and C9 perfluoro analogue compounds (0-3.5% of the compound). The
test followed U.S. EPA (1982) and OECD (1997) test protocols. The 21-day NOEC and LOEC,
based on reproduction and survival were 22 and 36 mg/L APFO, respectively with a
corresponding MATC of 28.14 mg/L. This test was acceptable for qualitative use only because
of the possible mixture effects of other perfluoro analogue compounds and missing exposure
details but does not suggest D. magna will be chronically sensitive relative to the chronic
freshwater criterion.
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Seyoum et al. (2020) evaluated the chronic effects of PFOA (CAS# 335-67-1, >99%,
purchased from Sigma) on Daphnia magna neonates via a 21-day unmeasured, static-renewal
study. The study authors did not report following any specific protocol. The 21-day reproductive
(fecundity) LOEC of 0.4141 mg/L PFOA was reported by the study authors, where a -38.25%
reduction in mean number of daphnids relative to the control was observed. The EPA was unable
to fit a model with significant parameters to the reproduction-based concentration-response data
due to a lack of clear concentration-dependent effects beyond the LOEC. The reproduction-based
LOEC (i.e., 0.4141 mg/L) was selected as the chronic value from this test; however, it was not
considered acceptable for quantitative use because chronic responses in this test did not display
concentration-dependent effects beyond the LOEC despite a 25X increase in treatment
concentrations. Moreover, additional ECio values from other, quantitatively acceptable tests,
were available to inform the chronic sensitivity of Daphnia magna in criteria derivation.
4.2.2.1.4 Fourth most chronically sensitive genus, Brachionus
The qualitatively acceptable chronic value for Brachionus from Zhang et al. (2014b)
was discussed in greater detail in Section 3.1.1.3.4. Briefly, Zhang et al. (2014b) conducted a full
life-cycle test using renewal conditions for approximately four days on the rotifer, Brachionus
calyciflorus. Zhang et al. (2014b) reported several endpoints, including intrinsic rate of natural
increase, which was selected as the primary endpoint for criterion derivation. Zhang et al.
(2014b) also observed effects to resting egg production. Resting egg production is an
ecologically important endpoint for this species because it represents the final result of sexual
reproduction. NOEC and LOEC values were not reported for resting egg production, but 0.25
mg/L PFOA produced more than a 50% reduction in resting egg production. Based on the
authors description of results in the text, it was assumed the B. calyciflorus four-day NOEC for
resting egg production was 0.125 mg/L and the LOEC was 0.25 mg/L and the calculated MATC
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was 0.1768 mg/L, suggesting resting egg production may be a relatively sensitive endpoint.
Because there was only one replicate (as implied by lack of error bars in Figure 1 of the
publication, no clear description of replicates in the methods section, and no author-reported
statistical analysis of this endpoint), resting egg production from this study was not considered
quantitatively acceptable and was instead considered in a qualitative manner. Overall, effects to
chronic apical endpoints for this genus, reported in this publication and Zhang et al. (2013a),
generally appear as a threshold effect from 0.25 mg/L to 1.0 mg/L, providing support for the
endpoint and effect level selected for quantitative use in criterion derivation (i.e., intrinsic rate of
natural increase), and further suggests the chronic criterion magnitude is adequately protective of
the genus, Brachionus.
4,2,2.2 Consideration of Relatively Sensitive Freshwater Tests based on Qualitatively
Acceptable Data
This section focuses on qualitatively acceptable chronic tests that were most relevant to
informing the appropriateness of the chronic freshwater criterion. Specifically, those tests used to
qualitatively inform the chronic freshwater criterion magnitude were identified as most relevant
if they met all parameters listed below:
1. reported effect concentrations that were less than or similar to (e.g., factor of two) the
chronic freshwater criterion magnitude;
2. evaluated an animal species;
3. conducted the test for a non-acute exposure duration (e.g., greater than seven days);
4. evaluated apical effects (i.e., long-term survival, growth, and/or reproduction), and;
5. not already discussed in the previous section (i.e., not a species discussed among the four
most sensitive genera).
The toxicity values summarized below were not used quantitatively to derive the chronic
PFOA freshwater criterion. Results of each individual study (as well as the rationale why a study
was not quantitatively acceptable) were considered relative to the chronic criterion magnitude to
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ensure the chronic PFOA criterion was not underprotective and to provide additional supporting
evidence of the potential toxicity of PFOA to aquatic organisms.
4.2.2.2.1 Genus: Chironomus (midge)
Stefani et al. (2014) conducted a chronic (10 generation) test of PFOA with a midge,
Chironomus riparius. The 10 generations (each approximately 20 to 28 days) were tested under
static conditions. The NOEC for the study, based on effect on emergence, reproduction or sex
ratio at the only concentration tested, was 0.0089 mg/L PFOA. Marziali et al. (2019) provides
further analysis of the same chronic test conducted by Stefani et al. (2014; reported in Table G.l)
by reporting measurements of alterative endpoints/responses. The LOEC based on
developmental time, adult weight was 0.0098 mg/L (time-weighted average; NOEC <0.0098
mg/L). While Stefani et al. (2014) reported no effects across the chronic test, Marziali et al.
(2019) reported effects to select generations. Overall, however, effects were sporadic with
reductions in growth observed in several generations. There were no effects on "survival,
development, or reproduction" and Marziali et al. (2019) concluded "no effects at population
level (population growth rate) were proved, thus a toxicity risk in real ecosystems at the tested
concentrations seems unlikely." The results from these studies were deemed not acceptable for
quantitative use because of limited test concentrations assessed, and uncertainty pertaining to
sediment characteristics, as well as poor control survival in four of the 10 generations.
Zhai et al. (2016) exposed Chironomusplumosus larvae to PFOA (perfluorooctanoic
acid, obtained from Acros Organics, Morris Plains, NJ, 96% purity) spiked in sediment for 10.3
days. The sediment was collected from the upstream region of the Yongding River in Beijing,
China. Sediments were physically mixed with a steel blender at 60 rpm for 24 hours in the
darkness at room temperature. Then, 1 mL of PFOA methanol solution (20 mg/L) was added to
obtain a concentration of 100 ng/g PFOA, and the sample was thoroughly mixed in a fume hood.
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When the methanol had evaporated, 300 mL of distilled water was added to each beaker. Finally,
the beakers were sealed, kept in the darkness at 25ฐC, and aged for 60 days before test initiation.
The midge larvae used in this study were collected from the uncontaminated upstream area of the
Yangliuqing River in the outer suburbs of Tianjin, China. Male larvae in the third or fourth
instars of similar size were chosen and placed in pure water to purify their guts for more than 48
hours in the laboratory. There were 20 larvae in each 500 mL beaker containing sediment and
300 mL of solution and three replicates per concentration (9.8 ng PFOA/g sediment) and control
(pure water). The static measured exposure lasted 10 days without food supply and involved a
16:8 hour (light:dark) photoperiod at 25 ฑ 1ฐC. At the end of the experiment, the surviving larvae
were counted. The 10-day mortality NOEC was 9.8 ng/g, the only concentration tested. The test
was not acceptable for quantitative use because: (1) exposures were spiked sediments rather than
water only, (2) test duration (-10 days) was sub-chronic, too long for an acute test and too short
for a chronic test, and (3) no effects were observed at the highest concentration tested which
produced a relatively low "> NOEC" value that did not provide information on the sensitivity of
this species to PFOA at concentrations consistent with criteria magnitudes. Results of this test
were retained for qualitative use.
Quantitatively acceptable midge data (McCarthy et al. 2021) were available to derive the
chronic PFOA freshwater criterion and these chronic data further suggest Chironomus is
relatively tolerant to chronic PFOA exposures.
4.2.2.2.2 Genus: Salmo (Atlantic salmon)
Spachmo and Arukwe (2012) exposed Atlantic salmon {Salmo salaf) to PFOA (95%
purity) in a 52-day flow-through test with unmeasured treatment concentrations. The 52-day
growth NOEC and LOEC were 0.10 and >0.10 mg/L PFOA, respectively. These data were
acceptable for quantitative use based on meeting data quality objectives; however, were not used
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in deriving the chronic criterion because the study only included one treatment group that
showed no adverse effects. Because the one treatment group that showed no effects was a
relatively low treatment concentration, including this NOEC value in the criterion calculation
would have resulted in the criterion magnitude being strongly influenced by the low test
concentration selected by study investigators (that did not produce an adverse response), rather
than a concentration-response relationship.
Arukwe et al. (2013) exposed Atlantic salmon (Salmo salar) embryos to PFOA at 100
|ig/L for 49 days. This test was not used quantitatively because of the lack of effects in the one
treatment group tested. The 49-day growth-based NOEC of 0.1 mg/L PFOS was selected as the
most sensitive and relevant endpoint. However, given there was only one treatment group and
that the test concentration was a relatively low concentration (at 0.1 mg/L) compared to the other
PFOA data, resulting in a > low value, this test was not informative enough for quantitative use
in the derivation of the PFOA aquatic life criteria. Results of this test were retained for
qualitative consideration.
4.2.2.2.3 Genus: Cyprinus (carp)
Petre et al. (2023) evaluated the bioconcentration and biotransformation of PFOA in
common carp (Cyprinus carpio) through a 14-week exposure and 3-week depuration period.
Carp (30-40 g) were exposed to one of two PFOA treatments (10 and 100 |ig/L) in a static
renewal measured experiment with solutions renewed every 48 hours. After 14 weeks, the fish
were visually analyzed, measured, and weighed. No mortalities or visible physiological were
observed in any treatment. Therefore, the 14-week NOEC, based on mortality, weight and
length, was >92.1 |ig/L (>0.0921 mg/L). The test not acceptable for quantitative use because the
chronic value was a "greater than" NOEC which would have influenced the criterion magnitude
despite no effects observed at the chronic value.
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4.2.2.2.4 Genus: Danio (zebrafish)
Jantzen et al. (2017a) evaluated the effects of PFOA on the morphometries behavioral
and gene expression in D. rerio exposed via five-day static unmeasured exposures. Zebrafish
embryos were exposed at 3-hpf to PFOA for 120 hours in what is equivalent to a rapid early-life
stage test. The observation period in clean water was extended beyond the exposure time points
from 120 hpf to 14 days post fertilization (dpf) to assess possible latent effects. The five-day
(plus nine days for observation) chronic value for growth-based endpoints, including body
length, was an MATC of 0.6325 mg/L (NOEC = 0.2 mg/L; LOEC = 2.0 mg/L), but the MATC
for swimming activity, a non-apical endpoint, was reported as 0.06325 mg/L (NOEC = 0.02
mg/L; LOEC = 0.2 mg/L). The reported chronic values based on growth and swimming activity
were not considered quantitatively acceptable because of the relatively brief chronic (i.e., 5-day)
exposure duration compared to other acceptable acute exposures that indicated D. rerio was
tolerant to brief (i.e., 96-hours) PFOA exposures.
4.2.2.2.5 Genus: Oryzias (medaka)
Ji et al. (2008) evaluated the chronic toxicity of PFOA (CAS # 335-67-1, purity not
provided) to the Japanese medaka, Oryzias latipes, via renewal unmeasured exposures. For the
F0 fish exposure study, breeding medaka pairs were exposed to PFOA for 14 days. Eggs were
counted every day, and the eggs spawned on the seventh day were saved for the F1 generation
exposure study. For the F1 fish exposure study, fertilized eggs collected from F0 fish were
exposed until all living embryos had hatched. Newly hatched larvae were then randomly
transferred to 100 mL beakers and observed daily for swim-up success and survival for an
additional two weeks. Larvae were fed Artemia nauplii ad libitum twice daily. After 14 days,
replicates from each treatment group were transferred to beakers without PFOA for observation
through 100 days post hatch. The F0 (parental generation) adult survival, condition factor and
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adult male and female GSI and HSI 14-day LOECs were all >10 mg/L PFOA. For the F1
(progeny generation), the LOEC for larval survival was 0.1 mg/L, while the corresponding
NOEC was considered <0.1 because effects were observed in the lowest concentration tested.
This test was not used quantitatively because uncertainties associated with the responses across
the range of concentrations tested. In many instances, the authors did not report increasing
chronic effects with increasing concentrations that differed by an order of magnitude.
Additionally, endpoints associated with longer term effects to juveniles were also rejected
because of pseudoreplication resulting from a lack of replicates in the hatching stage. Since this
test is a static-renewal unmeasured test, the EPA chose to rely exclusively on the test by Lee et
al. (2017) to derive the SMCV for this species since Lee et al. (2017) was a flow-through
measured test with fewer concerns pertaining to test design (i.e., no pseudoreplication) and
results (lack of increasing effects despite a 10-fold increase in exposure concentrations).
4.3 Acute to Chronic Ratios
When sufficient empirical data are not available, the 1985 Guidelines allow the use of a
Final Acute-to-Chronic Ratio (FACR) to convert the FAV to a FCV as an alternative approach to
derive the chronic criterion (U.S.EPA 1985). An ACR approach was not used for the derivation
of the chronic freshwater PFOA criterion, which was derived from empirical chronic data with
all of the eight MDRs met. For illustrative purposes only, 11 individual ACRs for five
invertebrate species, and one amphibian could be calculated from the quantitatively acceptable
acute and chronic toxicity data (Appendix A and Appendix C). Appendix I includes the ACRs
for freshwater aquatic species with quantitatively acceptable acute values for which comparable
quantitatively acceptable chronic values were reported from the same study or same investigator
and laboratory combination. For each species where more than a single ACR was calculated,
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Species Mean Acute-Chronic Ratios (SMACRs) were also calculated as the geometric mean
value of individual ACRs. In the case of a single ACR within a species, that ACR was the
SMACR.
Individual ACRs ranged from <4.229 to 3,493 across all species and SMACRs ranged
from <4.299 to 3,493. Except for D. magna, all SMACRs consisted of a single ACR. Lithobates
catesbeiana had the largest ACR (i.e., 3,493). The denominator of the L. catesbeiana ACR was a
LOEC where authors reported a significant effect to snout vent length (SVL) despite a reduction
of only -7% relative to control responses (Flynn et al. 2019). The -7% decrease to SVL
observed at the LOEC was a relatively mild effect level compared to the denominator (i.e.,
chronic value) of most other PFOA ACRs, with chronic effect levels typically being ECio values
or MATCs that had corresponding LOECs that produced a >10% effect. Consequently, the
relatively mild effect to SVL observed by Flynn et al. (2019) may have contributed to an
artificially high ACR relative to the other PFOA ACRs available.
Daphnia carinata had the second highest ACR (2,113) and the denominator was based on
a MATC, with the corresponding NOEC and LOEC (which reduced reproduction by -40%) that
differed by a factor of 10 (Logeshwaren et al. 2021). The 10X difference between the D.
carinata NOEC and LOEC published by Logeshwaran et al. (2021) likely contributed to an
artificially low MATC that, in turn, produced an artificially high ACR.
The mayfly, Neocloeon triangulifer, had the lowest non-definitive ACR (<4.229). The
96-hour LCso of 13.05 mg/L was divided by the highest chronic test concentration from Soucek
et al. (2023) that did not result in any adverse survival, weight or emergence effects at test
termination (i.e., NOEC > 3.085 mg/L). The relatively high acute sensitivity of this species
produced a low ACR compared to other test organisms.
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Four out of the five D. magna ACRs ranged from 14.50 to 69.08, with the remaining
ACR from Lu et al. (2016) being 1,602. The D. magna ACR from Lu et al. (2016) was removed
from the SMACR calculation because it was an order of magnitude greater than other ACRs for
the species. Overall, the range of SMACRs was greater than a factor of 100 and there was not an
apparent relationship between SMACRs and SMAVs. The 1985 Guidelines do not provide for
calculation of a FACR under these circumstances.
4.4 Tissue-based Toxicity Studies Compared to the Chronic Tissue-based
Criteria
Tissue-based PFOA toxicity data were reported for four species (three fish and one frog
species) across five publications, all of which were classified as qualitatively acceptable. Feng et
al. (2015) conducted a 96-hour study with juvenile goldfish (Carassius auratus) and observed no
effects of PFOA on mortality or antioxidant enzyme activity in the highest aqueous PFOA
treatment concentration (4.931 mg/L, measured), which corresponded to liver, gill, and muscle
PFOA concentrations of 17.11, 35.13, and 6.07 mg/kg dry weight, respectively.
Giari et al. (2016) measured PFOA in several tissues of two-year-old common carp
{Cyprinus carpio) exposed to nominal PFOA water concentrations of 0.0002 mg/L and 2 mg/L,
plus a control, for 56 days. Corresponding tissue PFOA average concentrations in blood, liver,
and muscle in fish from the 2 mg/L treatment after 56 days were 0.0649, 0.0281, and 0.0075
mg/kg wet weight, respectively. No effects of mortality, condition factor, hepatic somatic index
(HSI) or gonadal somatic index (GSI) were observed. Manera et al. (2017) performed a separate
study that replicated the study design of Giari et al. (2016), in which two-year-old common carp
{Cyprinus carpio) were exposed to nominal PFOA water concentrations of 0.0002 mg/L and 2
mg/L, plus a control, for 56 days. PFOA was measured in liver, and average concentrations in
fish exposed to 2 mg/L for 56 days was 0.0284 mg/kg wet weight, similar to Giari et al. (2016).
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PFOA in the livers of fish exposed to 0.0002 mg/L and the control were below detection
(<0.0002 mg/kg ww). No apical endpoints were reported; however, evidence of degenerative
liver morphology in PFOA exposed fish was observed.
Hagenaars et al. (2013) exposed adult zebrafish (D. rerio) to aqueous PFOA for 28 days.
Several reproductive and biochemical endpoints were measured. Whole-body PFOA
concentrations in the highest concentration (1 mg/L PFOA, nominal) after 28 days averaged
0.550 mg/kg wet weight in males and 0.301 mg/kg wet weight in females. No statistically
significant differences were observed in reproductive endpoints (total egg production, fertilized
egg production, and hatching rate) for any treatment levels compared to controls. Statistically
significant effects were observed among non-apical endpoints. Decreased whole body glycogen
content was lower in male and female fish across all exposure treatments, and liver
mitochondrial electron transport activity was lower in males exposed to the highest PFOA
concentration. Differences in several liver proteins of PFOA exposed males and females were
also observed.
Hoover et al. (2017) exposed juvenile (Gosner stage 26) northern leopard frogs
(Lithobatespipiens, formerly, Ranapipiens) to three PFOA concentrations (0.011, 0.109, and
1.11 mg/L PFOA, respectively) for 40 days. Survival, growth (snout-vent length), and
developmental time were measured. Whole body PFOA concentrations in frogs exposed to the
highest aqueous treatment level averaged 3.61 mg/kg dry weight after 40 days. Tadpole moisture
content was not reported. In order to convert the reported dry weight concentrations to wet
weight concentrations, so that they would be more directly comparable to the whole-body fish
tissue criterion, a whole-body moisture content of 72.1% was applied, calculated as the average
for all fish collected as part of the USGS National Contaminant Biomonitoring Program (NCBP
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Fish Database). The resulting whole-body concentration at the highest treatment level after 40
days was 1.01 mg/kg wet weight. No effects of PFOA on mortality, growth, or development time
were reported.
Tissue PFOA concentrations reported in all but one of the qualitative studies were lower
than the chronic freshwater tissue-based criteria (i.e., invertebrate whole-body criterion =1.18
mg/kg ww; fish whole-body criterion = 6.49 mg/kg ww; fish muscle criterion = 0.133 mg/kg
ww). The exception was reported by Feng et al. (2015), where the 96-hour muscle-based NOEC
(i.e., 6.07 mg/kg dw) was greater than the muscle tissue criterion magnitude of 0.133 mg/kg.
However, the liver- and muscle-based NOECs reported by Feng et al. (2015) were from an acute
duration (96-hour exposure), whereas the tissue-based values were derived to protect species
from longer-term chronic exposures, where effects to sensitive species at concentrations lower
than the whole body-based NOEC reported by Feng et al. (2015) may occur. In addition, the
tissue concentrations were reported by Feng et al. (2015) as dry weight, as tissue specific percent
moisture contents were not available. Had the values in Feng et al. (2015) been expressed as wet
weight, the magnitude of the differences would have been smaller.
Although all other tissue-based concentrations were lower than the corresponding tissue-
based criteria, no statistically significant effects of apical endpoints were observed in any of
these studies. Results of these studies do not provide any evidence that the aquatic community
will experience unacceptable chronic effects at tissue-based criteria magnitudes.
4.5 Effects on Aquatic Plants
Available data for aquatic plants and algae were reviewed to determine if aquatic plants
were likely to be adversely affected by PFOA and if they were likely to be more sensitive to
PFOA than aquatic animals (see Section 4 and Appendix E: Acceptable Freshwater Plant PFOA
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Toxicity Studies). Toxicity values for freshwater plants reported in Appendix E were all greater
than the chronic freshwater criterion (i.e., 0.10 mg/L PFOA). The lowest effect concentrations
reported were all NOECs for the green alga (Chlorellapyrenoidosa). All of these were relatively
low NOEC values, which do not provide meaningful information about the sensitivity of this
alga relative to the chronic water column criterion. However, a definitive EC50 for the same
species of 190.99 mg/L (Xu et al. 2013) suggests the PFOA criteria magnitudes are protective of
C. pyrenoidosa. Excluding the relatively low NOECs, effect concentrations for freshwater plants
and algae ranged from 5 to 745.7 mg/L relative to animal chronic values of 0.03162 to 88.32
mg/L (Appendix C). Therefore, it was not necessary to develop a criterion based on the toxicity
of PFOA to aquatic plants and the PFOA freshwater criteria are expected to be protective of
freshwater plants.
4.6 Protection of Threatened and Endangered Species
The PFOA acute and chronic datasets include some data representing species that are
listed as threatened or endangered by the U.S. Fish and Wildlife Service and/or National Oceanic
and Atmospheric Administration (NOAA) Fisheries. Summaries are provided here describing the
available PFOA toxicity data for listed species indicating that the PFOA criteria are protective of
these listed species, based on available scientific data.
4.6.1 Quantitatively Acceptable Acute Toxicity Data for Listed Species
Quantitatively acceptable acute toxicity test data evaluating the effects of PFOA on
threatened and endangered freshwater species were available for rainbow trout (Oncorhynchus
mykiss) with a SMAV of 4,001 mg/L PFOA (DuPont Haskill Laboratory 2000). The rainbow
trout SMAV is more than 1,290 times greater than the recommended acute criterion (CMC) of
3.1 mg/L, indicating the acute criterion is protective of rainbow trout and expected to be
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protective of other listed salmonid species. There were no acceptable acute toxicity data for
endangered or threatened estuarine/marine aquatic species.
4.6.2 Quantitatively Acceptable Chronic Toxicity Data for Listed Species
Quantitatively acceptable chronic toxicity test data evaluating the effects of PFOA on
threatened and endangered freshwater species were available for rainbow trout (Oncorhynchus
mykiss) with a SMCV of >40 mg/L PFOA (Centre International de Toxicologie 2004; Colombo
et al. 2008). The rainbow trout SMCV is 400 times greater than the recommended chronic
criterion (CCC) of 0.10 mg/L, indicating the chronic criteria are protective of rainbow trout and
other listed salmonid species. There are no acceptable chronic toxicity data for endangered or
threatened estuarine/marine aquatic species.
4.6.3 Qualitatively Acceptable Toxicity Data for Listed Species
Focusing on qualitatively acceptable tests with apical endpoints and water column
exposures, there available data for the listed Atlantic salmon (Salmo salaf). Spachmo and
Arukwe (2012) and Arukwe et al. (2013) observed no adverse effects for growth at the highest
treatment concentration (0.1 mg/L PFOA) following 49-day and 53-day exposures, respectively.
For both exposures, the qualitatively-acceptable NOECs were greater than the recommended
acute criterion (CCC) of 0.10 mg/L, which further indicates that the chronic criteria are
protective of listed salmonid species. There were no qualitative acute or chronic toxicity data for
endangered or threatened estuarine/marine aquatic species.
4.7 Summary of the PFOA Aquatic Life Criteria and the Supporting
Information
The PFOA aquatic life criteria were developed to protect aquatic life against adverse
effects, such as mortality, altered growth, and reproductive impairments associated with acute
and chronic exposure to PFOA. This Aquatic Life Ambient Water Quality Criteria for
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Perfluorooctanoic acid (PFOA) document includes water column based acute and chronic criteria
and tissue-based criteria for fresh waters. Acute and chronic water column criteria magnitudes
for estuarine/marine waters could not be derived at this time due to data limitations; however
acute estuarine/marine benchmarks are provided in Appendix L. The freshwater acute water
column-based criterion magnitude is 3.1 mg/L, and the chronic water column-based chronic
criterion magnitude is 0.10 mg/L. The fish whole-body tissue criterion magnitude is 6.49 mg/kg
wet weight, the fish muscle tissue criterion magnitude is 0.133 mg/kg wet weight and the
invertebrate whole-body tissue criterion magnitude is 1.18 mg/kg wet weight (Table 3-12).
Although empirical PFOA toxicity data for estuarine/marine species were not available to fulfill
the eight MDRs directly, the EPA included an acute aquatic life benchmark for estuarine/marine
environments in Appendix L, using available estuarine/marine species toxicity data and a NAM
application of ORD's peer-reviewed web ICE tool. The estuarine/marine acute water column-
based benchmark magnitude is 7.0 mg/L; this value provides information on a concentration that
should be protective of aquatic estuarine/marine life from acute PFOA exposures. As noted
earlier, the benchmark value has greater uncertainty than the freshwater PFOA criteria, due to the
paucity of empirical data of PFOA effects on estuarine/marine organisms.
The EPA evaluated the influence of including non-North American resident species on the
acute and chronic criteria magnitudes and concluded their inclusion did not substantiality affect
the criteria magnitudes. These PFOA aquatic life criteria are expected to be protective of aquatic
life, such as fish and aquatic invertebrates, on a national basis.
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5 REFERENCES
3M Company. 2000a. Information on perfluorooctanoic acid and supplemental information on
perfluorooctane sulfonates and related compounds. EPA/OTS Doc. No. #FYI-OTS-0500-1378:
4297 pp.
3M Company. 2000b. Voluntary Use and Exposure Information Profile Perfluorooctanic Acid
and Salts; U.S. EPA Administrative Record AR226-0595; U.S. Environmental Protection
Agency: Washington, DC, 2000.
3M Company. 2001. Environmental Monitoring-Multi-City Study Water, Sludge, Sediment,
POTW Effluent and Landfill Leachate Samples. 3M Environmental Laboratory.
3M Company. 2003. Perfluorooctanoic Acid Physiochemical Properties and Environmental Fate
Data. 3M Center, St. Paul, MN. EPA. Office of Pollution Prevention and Toxics (OPPT); OPPT-
2003-00 12.
Adedara, I.A., T.P. Souza, J. Canzian, A.A. Olabiyi, J.V. Borba, E. Biasuz, G.R. Sabadin, F.L.
Goncalves, F.V. Costa, M.R.C. Schetinger, E.O. Farombi and D.B. Rosemberg. 2022. Induction
of aggression and anxiety-like responses by perfluorooctanoic acid is accompanied by
modulation of cholinergic- and purinergic signaling-related parameters in adult zebrafish.
Ecotoxicol. Environ. Saf. 239:10.
Ahrens, L. 2011. Polyfluoroalkyl compounds in the aquatic environment: a review of their
occurrence and fate. J. Environ. Monit. 13(1): 20-31.
Ahrens, L. and M. Bundschuh. 2014. Fate and effects of poly- and perfluoroalkyl substances in
the aquatic environment: a review. Environ. Toxicol. Chem. 33(9): 1921-1929.
Ahrens, L., S. Felizeter, R. Sturm, Z. Xie and R. Ebinghaus. 2009. Polyfluorinated compounds in
waste water treatment plant effluents and surface waters along the River Elbe, Germany. Mar.
Pollut. Bull. 58: 1326-1333.
Ahrens, L., S. Taniyasu, L.W.Y. Yeung, N. Yamashita, P.K.S. Lam, R. Ebinghaus. 2010.
Distribution of polyfluoroalkyl compounds in water, suspended particulate matter and sediment
from Tokyo Bay, Japan. Chemosphere 79(3): 266-272.
Ahrens, L., M. Shoeib, S. Del Venlo, G. Codling and C. Halsall. 2011a. Polyfluoroalkyl
compounds in the Canadian Arctic atmosphere. Environ. Chem. 8: 399-406.
Ahrens, L., M. Shoeib, T. Harner, S.C. Lee, R. Guo and E. J. Reiner. 201 lb. Wastewater
treatment plant and landfills as sources of polyfluoroalkyl compounds to the atmosphere.
Environ. Sci. Technol. 45: 8098-8105.
137
-------
Anderson, R. H., G. C. Long, R. C. Porter and J. K. Anderson. 2016. Occurrence of select
perfluoroalkyl substances at U.S. Air Force aqueous film-forming foam release sites other than
fire-training areas: Field-validation of critical fate and transport properties. Chemosphere. 150:
678-685.
Ankley, G.T., P. Cureton, R. A. Hoke. M. Houde, A. Kumar, J. Kurias, R. Lanno, C. McCarthy,
J. Newsted, C. J. Salice, B. E. Sample, M. S. Sepulveda, J. Steevens and S. Valsecchi. 2020.
Assessing the ecological risks of per- and polyfluoroalkyl substances: Current state-of-the
science and a proposed path forward. Environ. Toxicol. Chem. 40(3): 564-605.
Annunziato, K.M. 2018. Low molecular weight PFAS alternatives (C-6) result in fewer cellular
and behavioral alterations than long chain (C-8/C-9) PFASS in larval zebrafish. Ph.D. Thesis,
Rutgers, The State University of New Jersey, New Brunswick, NJ:188 pp.
Armitage, J. M., U. Schenker, M. Scheringer, J. W. Martin, M. MacLeod and I. T. Cousins.
2009. Modeling the global fate and transport of perfluorooctane sulfonate (PFOS) and precursor
compounds in relation to temporal trends in wildlife exposure. Environ. Sci. Technol. 43(24):
9274-9280.
Arukwe, A. and A.S. Mortensen. 2011. Lipid peroxidation and oxidative stress responses of
salmon fed a diet containing perfluorooctane sulfonic- or perfluorooctane carboxylic acids.
Comp. Biochem. Physiol. Part C. 154: 288-295.
Arukwe, A., M.V. Cangialosi, R.J. Letcher, E. Rocha and A.S. Mortensen. 2013. Changes in
morphometry and association between whole-body fatty acids and steroid hormone profiles in
relation to bioaccumulation patterns in salmon larvae exposed to perfluorooctane sulfonic or
perfluorooctane carboxyl. Aquat. Toxicol. 130/131: 219-230.
ASTM (American Society for Testing and Materials). 1981. ASTM-E-35.23. Proposed standard
practice for conducting toxicity tests with freshwater and saltwater algae. Draft No. 2.
ASTM (American Society for Testing and Materials). 1993. Standard guide for conduction acute
toxicity tests with fishes, macroinvertebrates, and amphibians. Annual Book of ASTM
Standards. Pp. 88-729.
ASTM (American Society for Testing and Materials). 1999. Standard guide for conducting acute
toxicity tests on test materials with fishes, macroinvertebrates, and amphibians: Designation E
729-96. In: Annual book of ASTM standards: water and environmental technology. 11th edition.
Pp. 224-244.
ASTM (American Society for Testing and Materials). 2002. Standard guide for conducting three-
brood, renewal toxicity tests with Ceriodaphnia dubia. E1295-01.
ASTM (American Society for Testing and Materials). 2004. Standard guide for conducting
laboratory toxicity tests with bioluminescent dinoflagellates. Method E 1924-97. (Reapproved
2004). ASTM. West Conshohocken, PA
138
-------
ASTM (American Society for Testing and Materials). 2005. Standard test method for measuring
the toxicity of sediment-associated contaminants with freshwater invertebrates. El706-05.
Annex 7 - Guidance for conducting a life-cycle test for measuring effects of sediment-associated
contaminants on Chironomus tentans. Philadelphia, PA, USA.
ASTM (American Society for Testing and Materials). 2006. Standard guide for conducting
laboratory toxicity tests with freshwater mussels. E2455-06. In Annual Book of ASTM
Standards. Philadelphia, PA.
ASTM (American Society for Testing and Materials). 2017. Standard E729-96: Guide for
conducting acute toxicity tests on test materials with fishes, macroinvertebrates, and amphibians.
West Conshohocken (PA). 22 p. https://www.astm.org/Standards/E729.htm
ATSDR (Agency for Toxic Substances and Disease Registry). 2015. Toxicological Profile for
Perfluoroalkyls. Draft for Public Comment. Agency for Toxic Substances and Disease Registry,
Public Health Service, United States Department of Health and Human Services, Atlanta, GA.
Accessed May 2016. http://www.atsdr.cdc.gov/ToxProfiles/tp200.pdf.
Awkerman, J.A., S. Raimondo, C.R. Jackson andM.G. Barron. 2014. Augmenting species
sensitivity distributions with interspecies toxicity estimation models. Environ. Toxicol. Chem.
33: 688-695.
Barmentlo, S.H., J.M. Stel, M. van Doom, C. Eschauzier, P. de Voogt and M.H.S. Kraak. 2015.
Acute and chronic toxicity of short chained perfluoroalkyl substances to Daphnia magna.
Environ. Pollut. 198: 47-53.
Bartlett, A.J., A.O. De Silva, D.M. Schissler, A.M. Hedges, L.R. Brown, K. Shires, J. Miller, C.
Sullivan and C. Spencer. 2021. Lethal and sublethal toxicity of perfluorooctanoic acid (PFOA) in
chronic tests with Hyalella azteca (amphipod) and early-life stage tests with Pimephales
promelas (fathead minnow). Ecotoxicol. Environ. Saf. 207: 10.
Batley, G.E., R.A. van Dam, M.St.J. Warne, J.C. Chapman, D.R. Fox, C.W. Hickey and J.L.
Stauber. 2014. Technical rationale for changes to the method for deriving Australian and New
Zealand water quality guideline values for toxicants. Prepared for the Council of Australian
Government's Standing Council on Environment and Water (SCEW). 40 pp.
Beach, S. A., J. L. Newsted, K. Coady and J. P. Giesy. 2006. Ecotoxicological evaluation of
perfluorooctanesulfonate (PFOS). Rev. Environ. Contam. Toxicol. 186: 133-174.
Bejarano, A.C. and J.R. Wheeler. 2020. Scientific basis for expanding the use of interspecies
correlation estimation models. Integr. Environ. Assess Manage. 16(4): 525-530.
Bejarano, A.C., S. Raimondo and M.G. Barron. 2017. Framework for optimizing selection of
interspecies correlation estimation models to address species diversity gaps in an aquatic
database. Environ. Sci. Technol. 51: 8158-8165.
139
-------
Benninghoff, A.D., W.H. Bisson, D.C. Koch, D.J. Ehresman, S.K. Kolluri and D.E. William.
2011. Estrogen-like activity of perfluoroalkyl acids in vivo and interaction with human and
rainbow trout estrogen receptors in vitro. Toxicol. Sci. 120(1): 42-58.
Benninghoff, A.D., G.A. Orner, C.H. Buchner, J.D. Hendricks, A.M. Duffy and D.E. Williams.
2012. Promotion of hepatocarcinogenesis by perfluoroalkyl acids in rainbow trout. Toxicol. Sci.
125(1): 69-78.
Benskin, J.P., D.C. Muir, B.F. Scott, C. Spencer, A.O. De Silva, H. Kylin, J.W. Martin, A.
Morris, R. Lohmann, G. Tomy, B. Rosenberg, S. Taniyasu andN. Yamashita. 2012.
Perfluoroalkyl acids in the Atlantic and Canadian Arctic Oceans. Environ. Sci. Technol. 46(11):
5815-5823.
Bernardini, I., V. Matozzo, S. Valsecchi, L. Peruzza, G.D. Rovere, S. Polesello, S. Iori, M.G.
Marin, J. Fabrello and M. Ci. 2021. The new PFAS C604 and its effects on marine invertebrates:
first evidence of transcriptional and microbiota changes in the Manila Clam Ruditapes
philippinarum. Environ. Int. 152: 17 p.
Berry, J.P., M. Gantar, P.D. Gibbs and M.C. Schmale. 2007. The zebrafish (Danio rerio) embryo
as a model system for identification and characterization of developmental toxins from marine
and freshwater microalgae. Comp Biochem. Physiol. C Toxicol. 145(1): 61-72.
Bertin, D., P. Labadie, B.J.D. Ferrari, A. Sapin, J. Garric, O. Geffard, H. Budzinski and M.
Babut. 2016. Potential exposure routes and accumulation kinetics for poly- and perfluorinated
alkyl compounds for a freshwater amphipod: Gammarus spp. (Crustacea). Chemosphere. 155:
380-387.
Borgmann, U., D.T. Bennie, A.L. Ball and V. Palabrica. 2007. Effect of a mixture of seven
pharmaceuticals on Hyalella azteca over multiple generations. Chemosphere. 66(7): 1278-1283.
Boudreau, T.M. 2002. Toxicity of perfluorinated organic acids to selected freshwater organisms
under laboratory and field conditions. Chapter 3: Toxicology of perfluoroalkyl carboxylic acids
(PFCAs) in relation to carbon-chain length. Masters of Science Thesis, University of Guelph,
Ontario, Canada. December 17, 2002.
Boudreau, T. M., P. K. Sibley, S. A. Mabury, D. G. C. Muir and K. R. Solomon. 2003.
Laboratory evaluation of the toxicity of perfluorooctane sulfonate (PFOS) on Selanastrum
140erfluorinate, Chlorella vulgaris, Lemna gibba, Daphnia magna, and Daphnia pulicaria. Arch
Environ Contam Toxicol. 44: 307-313.
Boulanger, B., J. Vargo, J.L. Schnoor and K.C. Hornbuckle. 2004. Detection of perfluorooctane
surfactants in Great Lakes water. Environ. Sci. Technol. 38(15): 4064-4070.
140
-------
Boulanger B, A.M. Peck, J.L. Schnoor, K.C. Hornbuckle. 2005. Mass budget of perfluorooctane
surfactants in Lake Ontario. Environ Sci Technol. 39(l):74-79. Erratum in: Environ Sci Technol.
2005 39(6): 1920.
Bouwmeester, M.C., S. Ruiter, T. Lommelaars, J. Sippel, H.M. Hodemaekers, E.J. van den
Brandhof, J.L.A. Pennings, J.H. Kamstra, J. Jelinek, J.P.J. Issa, J. Legler and L.T.M. van der
Ven. 2016. Zebrafish embryos as a screen for DNA methylation modifications after compound
exposure. Toxicol. Applied Pharmacol. 291: 84-96.
Brill, J.L., S.E. Belanger, J. Chaney, S.D. Dyer, S. Raimondo, M.G. Barron and C.A. Pittinger.
2016. Development of algal inter-species correlation estimation (ICE) models for chemical
hazard assessment. Environ. Toxicol. Chem. 35: 2368-2378.
Buck, R.C., J. Franklin, U. Berger, J.M. Conder, I.T. Cousins, P. de Voogt, A.A. Jensen, K.
Kannan, S.A. Mabury and S.P.J. vanLeeuwen. 2011. Perfluoroalkyl and polyfluoroalkyl
substances in the environment: Terminology, classification, and origins. Integrat. Environ.
Assess. Manag. 7(4): 513-541.
Burkhard, L. P. 2021. Evaluation of published bioconcentration factor (BCF) and
bioaccumulation factor (BAF) data for per-and polyfluoroalkyl substances across aquatic
species. Environ. Toxicol. Chem. 40(6): 1530-1543.
Burns, D. C., D.A. Ellis, H. Li, C.J. McMurdo and E. Webster. 2008. Experimental pKa
determination for perfluorooctanoic acid (PFOA) and the potential impact of pKa concentration
dependence on laboratory-measured partitioning phenomena and environmental modeling.
Environ. Sci. Technol. 42(24): 9283-9288.
Butt, C.M., U. Berger, R. Bossi and G.T. Tomy. 2010. Levels and trends of poly- and
perfluorinated compounds in the arctic environment. Sci. Total. Environ. 48: 2936-2965.
Butt, C.M., D.C.G. Muir, and S.A. Mabury. 2014. Biotransformation pathways of fluorotelomer-
based polyfluoroalkuyl substances: A review. Environ. Toxicol. Chem. 33(2): 243-267.
Canada (Department of the Environment, Department of Health). 2012. Screening Assessment
Report. Perfluorooctanoic Acid, its Salts, and its Precursors. August 2012.
https://www.ee. gc.ca/ese-ees/default.asp?lang=En&n=370AB 133-1
CCME (Canadian Council of Ministers of the Environment). 2007. Protocol for the derivation of
water quality guidelines for the protection of aquatic life. Winnipeg, Manitoba: Canadian
Council of Ministers of the Environment, https://ccme.ca/en/res/protocol-for-the-derivation-of-
water-qualitv-guidelines-for-the-protection-of-aquatic-life-2007-en.pdf
Centre International de Toxicologie. 2003. Ammonium perfluorooctanoate (APFO): Daphnia
magna reproduction test. Study No. 22658 ECD. Evreux (FR): 52 pp.
141
-------
Centre International de Toxicologie. 2004. Ammonium perfluorooctanoate (APFO): Early-life
stage toxicity test in rainbow trout under flow-through conditions. Study No. 22659 ECP. Evreux
(FR): OPPT Administrative Record AR226-1803: 112 pp.
Chen, C.C., Y. Shi, Y. Zhu, J. Zeng, W. Qian, S. Zhou, J. Ma, K. Pan, Y. Jiang, Y. Tao, and X.
Zhu. 2022. Combined Toxicity of Polystyrene Microplastics and Ammonium Pert]uorooctanoate
to Daphnia magna: Mediation of Intestinal Blockage. Water Res.219:10 p
Chen, P., Q. Wang, M. Chen, J. Yang, R. Wang, W. Zhong, L. Zhu and L. Yang. 2018.
Antagonistic estrogenic effects displayed by Bisphenol AF and perfluorooctanoic acid on
zebrafish (Danio rerio) at an Early developmental stage. Environ. Sci. Technol. Lett. 5(11): 655-
661.
Cochran, R.S. 2015. Evaluation of perfluorinated compounds in sediment, water, and passive
samplers collected from the Barksdale Air Force Base. MS Thesis. Texas Tech University,
Lubbock, TX United States. https://ttu-ir.tdl.org/bitstream/handle/2346/63633/COCHRAN-
THESIS-2015.pdf?sequence=l&isAllowed=v
Colombo, I., W. de Wolf, R.S. Thompson, D.G. Farrar, R.A. Hoke and J. L'Haridon. 2008.
Acute and chronic aquatic toxicity of ammonium perfluorooctanoate (APFO) to freshwater
organisms. Ecotoxicol. Environ. Saf. 71: 749-756.
Colorado Department of Public Health and the Environment. 2020. PFAS sampling project
results: surface water.
https://cohealthviz.dphe.state.co.us/#/site/EnvironmentalEpidemiologyPublic/views/PF AS result
s DRAFT/Samplelocationsdash?%3AshowAppBanner=false%26%3Adisplav count
Conley, J.M., C.S. Lambright, N. Evans, E. Medlock-Kakaley, A. Dixon, D. Hill, J. McCord,
M.J. Strynar, J. Ford and L.E. Gray. 2022. Cumulative maternal and neonatal effects of
combined exposure to a mixture of perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic
acid (PFOS) during pregnancy in the Sprague-Dawley rat. Environ. Int. 170:107631.
Consoer, D.M. 2017. A mechanistic investigation of perfluoroalkyl acid kinetics in rainbow trout
{Oncorhynchus mykiss). A dissertation submitted to the faculty of the University of Minnesota.
Consoer, D.M., A.D. Hoffman, P.N. Fitzsimmons, P.A. Kosian and J.W. Nichols. 2014.
Toxicokinetics of perfluorooctanoate (PFOA) in rainbow trout (Oncorhynchus mykiss). Aquat.
Toxicol. 156: 65-73.
Corrales, J., L.A. Kristofco, W.B Steele, G.N. Saari, J. Kostal, E.S. Williams, M. Mills, E.P
Gallagher, T.J. Kavanagh, N. Simcox, L.Q. Shen, F. Melnikov, J.B. Zimmerman, A.M.
Voutchkova-Kostal, P.T. Anastas and B.W. Brooks. 2017. Toward the design of less hazardous
chemicals: Exploring comparative oxidative stress in two common animal models. Chem. Res.
Toxicol. 30: 893-904.
142
-------
CRC CARE (Cooperative Research Centre for Contamination Assessment and Remediation of
the Environment). 2017. Assessment, management and remediation for PFOS and PFOA Part 4:
application of HSLs and ESLs. Technical Report No. 38.
Cui, Y., W. Liu, W. Xie, W. Yu, C. Wang and H. Chen. 2015. Investigation of the effects of
perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) on apoptosis and cell
cycle in a zebrafish (Danio rerio) liver cell line. Int. J. Environ. Res. Public Health. 12(12):
15673-15682.
Dasgupta, S., A. Reddam, Z. Liu, J. Liu and D.C. Volz. 2020. High-content screening in
zebrafish identifies perfluorooctanesulfonamide as a potent developmental toxicant. Environ.
Pollut. 256: 113550.
Dasu, K. 2011. Evaluating the biotransformation potential of commercial model fluorotelomer
monomers in soils. PhD Thesis, Purdue University.
Dasu, K., J. Liu and L.S. Lee. 2012. Aerobic soil biodegradation of 8:2 fluorotelomer stearate
monoester. Environ. Sci. Technol. 46: 3831-3836.
Dasu, K., L.S. Lee, R.F. Turco and L.F. Nies. 2013. Aerobic biodegradation of 8:2 fluorotelomer
stearate monoester and 8:2 fluorotelomer citrate triester in forest soil. Chemosphere. 91: 399-
405.
De Silva, A.O., P.J. Tseng and S.A. Mabury. 2009. Toxicokinetics of perfluorocarboxylate
Isomers in rainbow trout. Environ. Toxicol. Chem. 28: 330-337.
De Silva, A.O., C. Spencer, B.F. Scott, S. Backus and D.C. Muir. 2011. Detection of a cyclic
perfluorinated acid, perfluoroethylcyclohexane sulfonate, in the Great Lakes of North America.
Environ. Sci. Technol. 45(19): 8060-8066.
Del Vento, S., C. Halsall, R. Gioia, K. Jones and J. Dachs. 2012. Volatile per- and
polyfluoroalkyl compounds in the remote atmosphere of the western Antarctic Peninsula: an
indirect source of perfluoroalkyl acids to Antarctic waters? Atmos. Poll. Res. 3(4): 450-455.
Delinsky, A.D., M.J. Strynar, S.F. Nakayama, J.L. Varns, X.B. Ye, P.J. McCann and A.B.
Lindstrom. 2009. Determination of ten perfluorinated compounds in bluegill sunfish (Lepomis
macrochirus) fillets. Environ. Res. 109(8): 975-984.
Delinsky, A.D., M.J. Strynar, P.J. McCann, J.L. Varns, L. McMillan, S.F. Nakayama and A.B.
Lindstrom. 2010. Geographical distribution of perfluorinated compounds in fish from Minnesota
lakes and rivers. Environ. Sci. Technol. 44: 2549-2554.
Department of Defense (DoD), Strategic Environmental Research and Development Program
(SERDP). 2019. Guidance for Assessing the Ecological Risks of PFASs to Threatened and
Endangered Species at Aqueous Film Forming Foam-Impacted Sites SERDP Project ER18-
1614, July 2019.
143
-------
Ding, G.H., T. Fromel, E.J. van den Brandhof, R. Baerselman and W.J.G.M. Peijnenburg. 2012a.
Acute toxicity of poly- and perfluorinated compounds to two cladocerans, Daphnia magna and
Chydorus sphaericus. Environ. Toxicol. Chem. 31(3): 605-610.
Ding, G., M. Wouterse, R. Baerselman and W.J.G.M. Peijnenburg. 2012b. Toxicity of
polyfluorinated and perfluorinated compounds to lettuce {Lactuca sativa) and green algae
(Pseudokirchneriella subcapitata). Arch. Environ. Contam. Toxicol. 62: 49-55.
Ding, G., J. Zhang, M. Wang, Y. Chen, G. Luo and D. Xiong. 2012c. Evaluation and prediction
of mixture toxicity of PFOS and PFOA to zebrafish (Danio rerio) embryos. Adv. Mater. Res.
485: 297-300.
Ding, G., J. Zhang, Y. Chen, L. Wang, M. Wang, D. Xiong and Y. Sun. 2013. Combined effects
of PFOS and PFOA on zebrafish {Danio rerio) embryos. Arch. Environ. Contam. Toxicol. 64(4):
668-675.
Dinglasan, M.J.A., Y. Ye, E.A. Edwards and S.A. Mabury. 2004. Fluorotelomer alcohol
biodegradation yields poly- and perfluorinated acids. Environ. Sci. Technol. 38: 2857-2864.
Dinglasan-Panlilio, J. M., S.S. Prakash and J.E. Baker. 2014. Perfluorinated compounds in the
surface waters of Puget Sound, Washington and Clayoquot and Barkley Sounds, British
Columbia. Mar. Pollut. Bull. 78(1-2): 173-180.
Dong, H., G. Lu, X. Wang, P. Zhang, H. Yang, Z. Yan, J. Liu, and R. Jiang. 2023. Tissue-
specific accumulation, depuration, and effects of perfluorooctanoic acid on fish: Influences of
aqueous pH and sex. Sci. Total Environ. 861: 10 p.
Dragojevic, J., P. Marie, J. Loncar, M. Popovic, I. Mihaljevic, and T. Smital. 2020.
Environmental contaminants modulate transport activity of zebrafish organic anion transporters
Oatl and Oat3. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 231: 8 p.
Du, G., H. Huang, J. Hu, Y. Qin, D. Wu, L. Song, Y. Xia and X. Wang. 2013. Endocrine-related
effects of perfluorooctanoic acid (PFOA) in zebrafish, H295R steroidenesis and receptor reporter
gene assays. Chemosphere 91: 1099-1106.
DuPont Haskell Laboratory. 2000. Summaries of studies conducted at DuPont Haskell
Laboratory with ammonium perfluorooctanoate and perfluorononanoate (with cover letter dated
05-25-2000).
Dyer, S.D., D.J. Versteeg, S.E. Belanger, J.G. Chaney andF.L. Mayer. 2006. Interspecies
correlation estimates predict protective environmental concentrations. Environ. Sci. Technol. 40:
3102-3111.
144
-------
Dyer, S.D., D.J. Versteeg, S.E. Belanger, J.G. Chaney, S. Raimondo andM.G. Barron. 2008.
Comparison of species sensitivity distributions derived from interspecies correlation models to
distributions used to derive water quality criteria. Environ. Sci. Technol. 42(8): 3076-3083.
EFSA (European Food Safety Authority). 2008. Opinion of the scientific panel on contaminants
in the food chain on perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) and
their salts. EFSA Journal. 653: 1-131.
EGLE (Michigan Department of Environment, Great Lakes, and Energy). 2010. Rule 57 water
quality values. Surface Water Assessment Section.
Elnabarawy, M.T. 1980. Aquatic Toxicity Testing: FC-143 (Lot-37) L.R. 5626 S. Report No.
037; OPPT Administrative Record AR226-0504:3780-3786.
Elnabarawy, M.T. 1981. Multi-phase exposure/recovery algal assay test method. Report No. 006;
OPPT Administrative Record AR226-0506:3791-3810.
Elonen, C. 2020. ECOTOXicology Knowledgebase System User Guide Version 5.3. U.S.
Environmental Protection Agency, Cincinnati, OH, EPA/600/R-20/087.
EPAV (Environment Protection Authority Victoria). 2016. Commonwealth Environmental
Management Guidance on Perfluorooctane Sulfonic Acid (PFOS) and Perfluorooctanoic Acid
(PFOA). Australian Government, Department of the Environment and Energy.
Fabbri, R., M. Montagna, T. Balbi, E. Raffo, F. Palumbo and L. Canesi. 2014. Adaptation of the
bivalve embryotoxicity assay for the high throughput screening of emerging contaminants in
Mytilus galloprovincialis. Mar. Environ. Res. 99: 1-8.
Fabrello, J., M. Ciscato, L. Masiero, L. Finos, S. Valsecchi, S. Polesello, I. Bernardini, G. Dalla
Rovere, L. Bargellon. 2021. New compounds, old problems. The case of C604 - a substitute of
PFOA - and its effects to the clam Ruditapesphilippinarum. J. Hazard. Mater. 420: 11.
Fairbrother, A. 2008. Risk Management Safety Factor. In Encyclopedia of Ecology Vol. 4,
Ecotoxicology; S.R. J0rgensen andB.D. Fath, Eds; Elsevier: Oxford.
Falk, S., K. Failing, S. Georgii, H. Brunn and T. Stahl. 2015. Tissue specific uptake and
elimination of perfluoroalkyl acids (PFAAs) in adult rainbow trout (Oncorhynchus mykiss) after
dietary exposure. Chemosphere. 129: 150-156.
Fang, S., X. Chen, S. Zhao, Y. Zhang, W. Jiang, L. Yang and L. Zhu. 2014. Trophic
magnification and isomer fractionation of perfluoroalkyl substances in the food web of Taihu
Lake, China. Environ. Sci. Technol. 48: 2173-2182.
Fang, X., Y. Wei, Y. Liu, J. Wang and J. Dai. 2010. The identification of apolipoprotein genes in
rare minnow (Gobiocypris rarus) and their expression following perfluorooctanoic acid
exposure. Comp. Biochem. Physiol. PartC. 151: 152-159.
145
-------
Feng, C.L., F. Wu, S.D. Dyer, H. Change and X.L. Zhao. 2013. Derivation of freshwater quality
criteria for zinc using interspecies correlation estimation models to protect aquatic life in China.
Chemosphere. 90: 1177-1183.
Feng, M., Q. He, L. Meng, X. Zhang, P. Sun and Z. Wang. 2015. Evaluation of single and joint
toxicity of perfluorooctane sulfonate, perfluorooctanoic acid, and copper to Carassius auratus
using oxidative stress biomarkers. Aquat. Toxicol. 161: 108-116.
Fernandez-Sanjuan, M., M. Faria, S. Lacorte and C. Barata. 2013. Bioaccumulation and effects
of perfluorinated compounds (PFCs) in zebra mussels (Dreissenapolymorpha). Environ. Sci.
Pollut. Res. 20: 2661-2669.
Flynn, R.W., M.F. Chislock, M.E. Gannon, S.J. Bauer, B.J. Tornabene, J.T. Hoverman and M.
Sepulveda. 2019. Acute and chronic effects of perfluoroalkyl substance mixtures on larval
American bullfrogs (Rana catesbeiana). Chemosphere. 236: 7 p.
Flynn, R.W., M. Iacchetta, C. De Perre, L. Lee, M.S. Sepulveda and J.T. Hoverman. 2021.
Chronic per-/polyfluoroalkyl substance exposure under environmentally relevant conditions
delays development in northern leopard frog {Ranapipiens) larvae. Environ. Toxicol. Chem.
40(3): 711-716.
Flynn, R.W., G. Hoover, M. Iacchetta, S. Guffey, C. De Perre, B. Huerta, W.M. Li, J.T.
Hoverman, L. Lee and M.S. Sepulv. 2022. Comparative toxicity of aquatic per- and
polyfluoroalkyl substance exposure in three species of amphibians. Environ. Toxicol. Chem.
41(6): 1407-1415.
Fojut, T.L., A.J. Palumbo and R.S. Tjeerdema. 2012a. Aquatic life water criteria derived via the
UC Davis Method: II. Pyrethroid Insecticides. In. Tjeerdema, R.S., Ed., Springer, NY, NY. Rev.
Environ. Contam. Toxicol. 216: 51-104.
Fojut, T.L., A.J. Palumbo and R.S. Tjeerdema. 2012b. Aquatic life water criteria derived via the
UC Davis Method: III. Diuron. In. Tjeerdema, R.S., Ed., Springer, NY, NY. Rev. Environ.
Contam. Toxicol. 216:105-142.
Fromel, T. and T.P. Knepper. 2010. Fluorotelomer ethoxylates: Sources of highly fluorinated
environmental contaminants Part I: Biotransformation. Chemosphere. 80: 1387-92.
Furdui, V.I., P.W. Crozier, E.J. Reiner and S.A. Mabury. 2008. Trace level determination of
perfluorinated compounds in water by direct injection. Chemosphere. 73(1 Suppl): S24-30.
Garoche, C., A. Boulahtouf, M. Grimaldi, B. Chiavarina, L. Toporova, M.J. Den Broeder, J.
Legler, W. Bourguet and P. Ba. 2021. Interspecies differences in activation of peroxisome
proliferator-activated receptor gamma by pharmaceutical and environmental chemicals. Environ.
Sci. Technol. 55(24): 16489-16501.
146
-------
Gebreab, K.Y., M.N.H. Eeza, T. Bai, Z. Zuberi, J. Matysik, K.E. O'Shea, A. Alia and J.P. Berry.
2020. Comparative toxicometabolomics of perfluorooctanoic acid (PFOA) and next-generation
perfluoroalkyl substances. Environ. Pollut. 265: 14.
Gebreab, K.Y., D. Benetti, M. Grosell, J.D. Stieglitz and J.P. Berry. 2022. Toxicity of
perfluoroalkyl substances (PFAS) toward embryonic stages of mahi-mahi (Coryphaena
hippurus). Ecotox. 31(7): 1057-1067.
Geis, S.W., K.L. Fleming, E.T. Korthals, G. Searle, L. Reynolds and D.A. Karner. 2000.
Modifications to the algal growth inhibition test for use as a regulatory assay. Environ. Toxicol.
Chem. 19: 36-41.
Geng, Q., M. Guo, H. Wu, J. Peng, G. Zheng, X. Liu, Y. Zhai and Z. Tan. 2021. Effects of single
and combined exposure to BDE-47 and PFOA on distribution, bioaccumulation, and toxicity in
blue mussel (Mytilus galloprovincialis). Ecotoxicol. Environ. Saf. 228: 1-9.
Gergs, A., S. Classen, T. Strauss, R. Ottermanns, T. Brock, H. Ratte, U. Hommen and T. Preuss.
2016. Ecological recovery potential of freshwater organisms: Consequences for environmental
risk assessment of chemicals. Rev. Environ. Con. Toxicol. 236: 259-294.
Gewurtz, S B., S.M. Backus, A.O. De Silva, L. Ahrens, A. Armellin, M. Evans, S. Fraser, M.
Gledhill, P. Guerra, T. Harner, P.A. Helm, H. Hung, N. Khera, M.G. Kim, M. King, S.C. Lee,
R.J. Letcher, P. Martin, C. Marvin, D.J. McGoldrick, A.L. Myers, M. Pelletier, J. Pomeroy, E. J.
Reiner, M. Rondeau, M.C. Sauve, M. Sekela, M. Shoeib, D.W. Smith, S.A. Smyth, J. Struger, D.
Spry, J. Syrgiannis and J. Waltho. 2013. Perfluoroalkyl acids in the Canadian environment:
multi-media assessment of current status and trends. Environ. Int. 59: 183-200.
Giari, L., F. Vincenzi, S. Badini, C. Guerranti, B.S. Dezfuli, E.A. Fana and G. Castaldelli. 2016.
Common carp Cyprinus carpio responses to sub-chronic exposure to perfluorooctanoic acid.
Environ. Sci. Pollut. Res. 23: 15321-15330.
Giesy, J. P. and K. Kannan. 2001. Global distribution of perfluorooctane sulfonate in wildlife.
Environ. Sci. Tech. 35(7): 1339-1342.
Giesy, J. P. and K. Kannan. 2002. Perfluorochemical surfactants in the evironment. Envrion Sci
Technol. 36(7): 146-152.
Giesy, J. P., J.E. Naile, J.S. Khim, P.D. Jones and J.L. Newsted. 2010. Aquatic toxicology of
perfluorinated chemicals. Rev. Environ. Contam. Toxicol. 202: 1-52.
Godfrey, A.E. 2017. Endocrine disrupting effects of halogenated chemicals on fish. Chapter 4:
Sex specific endocrine disrupting effects of halogenated chemicals in Japanese medaka. Ph.D.
Thesis Dissertation, Purdue University, Department of Forestry and Natural Resources, West
Lafayette, IN.
147
-------
Godfrey, A., A. Abdel-moneim and M.S. Sepulveda. 2017a. Acute mixture toxicity of
halogenated chemicals and their next generation counterparts on zebrafish embryos.
Chemosphere 181: 710-712.
Godfrey, A., B. Hooser, A. Abdelmoneim, K.A. Horzmann, J.L. Freeman and M.S. Sepulveda.
2017b. Thyroid disrupting effects of halogenated and next generation chemicals on the swim
bladder development of zebrafish. Aquat. Toxicol. 193: 228-235.
Godfrey, A., B. Hooser, A. Abdelmoneim and M.S. Sepulveda. 2019. Sex-specific endocrine-
disrupting effects of three halogenated chemicals in Japanese medaka. J. Appl. Toxicol. 39(8):
1215-1223.
Goecke-Flora, C M. and N. V. Reo. 1996. Influence of carbon chain length on the hepatic effects
of peril uorinated fatty acids. A 19F- and 31P-NMR investigation. Chem. Res. Toxicol. 9(4):
689-95.
Goeritz, I., S. Falk, T. Stahl, C. Schafers and C. Schlechtriem. 2013. Biomagnification and
Tissue Distribution of Perfluoroalkyl Substances (PFASs) in Market-Size Rainbow Trout
(Oncorhynchus mykiss). Environ. Toxicol. Chem. 32(9): 2078-2088.
Gonzalez-Naranjo, V. and K. Boltes. 2014. Toxicity of ibuprofen and perfluorooctanoic acid for
risk assessment of mixtures in aquatic and terrestrial environments. Int. J. Environ. Sci. Technol.
11: 1743-1750.
Gorrochategui, E., S. Lacorte, R. Tucker and F.L. Martin. 2016. Perfluoroalkylated substance
effects in Xenopus laevis A6 kidney epithelial cells determined by ATR-FTIR spectroscopy and
chemometric analysis. Chem. Res. Toxicol. 29: 924-932.
Green, A. 2014. Invasive species report Mediterranean mussel. University of Washington,
Seattle, WA. Accessed January 2020: https://depts.washington.edu/oldenlab/wordpress/wp-
content/uploads/2015/09/Mytilus galloprovincialis Green 2014.pdf
Guo, R., E.J. Reiner, S.P. Bhavsar, P.A. Helm, S.A. Mabury, E. Braekevelt and S.A. Tittlemier.
2012. Determination of polyfluoroalkyl phosphoric acid diesters, perfluoroalkyl phosphonic
acids, perfluoroalkyl phosphinic acids, perfluoroalkyl carboxylic acids, and perfluoroalkane
sulfonic acids in lake trout from the Great Lakes Region. Anal. Bioanal. Chem. 404: 2699-2709.
Hagenaars, A., L. Vergauwen, W. De Coen and D. Knapen. 2011. Structure-activity relationship
assessment of four perfluorinated chemicals using a prolonged zebrafish early life stage test.
Chemosphere 82: 764-772.
Hagenaars, A., L. Vergauwen, D. Benoot, K. Laukens and D. Knapen. 2013. Mechanistic
toxicity study of perfluorooctanoic acid in zebrafish suggests mitochondrial dysfunction to play a
key role in PFOA toxicity. Chemosphere 91: 844-856.
148
-------
Haimbaugh, A., C. Wu, C. Akemann, D.N. Meyer, M. Connell, M. Abdi, A. Khalaf, D. Johnson,
and T.R. Baker. 2022. Multi- and transgenerational effects of developmental exposure to
environmental levels of PFAS and PFAS mixture in zebrafish (Danio rerio). Toxics 10(6): 334-
358.
Han, J., W. Gu, H. Barrett, D. Yang, S. Tang, J. Sun, J. Liu, H.M. Krause, K.A. Houck, and H.
Peng. 2021. A roadmap to the structure-related metabolism pathways of per- and polyfluoroalkyl
substances in the early life stages of zebrafish (Danio rerio). Environ. Health Perspect. 129(7):
15 p.
Han, Z.X., M. Zhang and C.X. Lv. 2011. Toxicokinetic behaviors and modes of perfluorooctane
sulfonate (PFOS) and perfluorooctane acid (PFOA) on tilapia (Oreochromis niloticus). Afr. J.
Biotechnol. 10(60): 12943-12950.
Hansen, K. J., H. Johnson, J. Eldridge, J. Butenhoff and L. Dick. 2002. Quantitative
characterization of trace levels of PFOS and PFOA in the Tennessee River. Environ. Sci.
Technol. 36(8): 1681-1685.
Hanson, M.L., J. Small, P.K. Sibley, T.M. Boudreau, R.A. Brain, S.A. Mabury and K.R.
Solomon. 2005. Microcosm evaluation of the fate, toxicity, and risk to aquatic macrophytes from
perfluorooctanoic acid (PFOA). Arch. Environ. Contam. Toxicol. 49: 307-316.
Hassell, K.L., T.L. Coggan, T. Cresswell, A. Kolobaric, K. Berry, N.D. Crosbie, J. Blackbeard,
V.J. Pettigrove and B.O. Clarke. 2020. Dietary uptake and depuration kinetics of perfluorooctane
sulfonate, perfluorooctanoic acid, and hexafluoropropylene oxide dimer acid (GenX) in a benthic
fish. Environ. Toxicol. Chem. 39(3): 595-603.
Haukas, M., U. Berger, H. Hop, B. Gulliksen and G. W. Gabrielsen. 2007. Bioaccumulation of
per- and polyfluorinated alkyl substances (PFAS) in selected species from the Barents Sea food
web. Environ. Pollut. 148(1): 360-371.
Hawkey, A.B., M. Mead, S. Natarajan, A. Gondal, O. Jarrett, and E.D. Levin. 2023. Embryonic
exposure to PFAS causes long-term, compound-specific behavioral alterations in zebrafish.
Neurotoxicol. Teratol. 97: 12 p.
Hayman, N.T., G. Rosen, M.A. Colvin, J. Conder and J.A. Arblaster. 2021. Aquatic toxicity
evaluations of PFOS and PFOA for five standard marine endpoints. Chemosphere 273: 7 p.
Hazelton, P.D. 2013. Emerging methods for emerging contaminants: novel approaches to
freshwater mussel toxicity testing. University of Georgia.
Hazelton, P.D., W.G. Cope, T.J. Pandolfo, S. Mosher, M.J. Strynar, M.C. Barnhart and R.B.
Bringolf. 2012. Partial life-cycle and acute toxicity of perfluoroalkyl acids to freshwater mussels.
Environ. Toxicol. Chem. 31: 1611-1620.
149
-------
Heard, R. W., W. W. Price, D. M. Knott, R. A. King D. M. Allen. 2006. A taxonomic guide to
the mysids of the South Atlantic Bight. NOAA Professional Paper NMFS 4, 37 p.
(https://spo.nmfs.noaa.gov/sites/default/files/pp4.pdf)
Hekster, F.M., R.W. Laane and P. de Voogt. 2003. Environmental and toxicity effects of
perfluoroalkylated substances. Rev. Environ. Contam. Toxicol. 179: 99-121.
HEPA (Heads of EPAs Australia and New Zealand). 2020. PFAS National Environmental
Management Plan Version 2.0. National Chemicals Working Group of the Heads of EPAs
Australia and New Zealand, (https://www.awe.gov.au/sites/default/files/documents/pfas-nemp-
2.pdf)
Higgins C. and R. Luthy. 2006. Sorption of perfluorinated surfactants on sediments. Environ.
Sci. Technol. 40(23): 7251-7256.
Holth, T.F., M. Yazdani, A. Lenderink and K. Hyllan. 2012. Effects of fluoranthene and
perfluorooctanoic acid (PFOA) on immune functions in Atlantic cod (Gadus morhua). Abstracts
Comp. Biochem. Physiol. Part A. 163: S39-S42.
Hong, S., J.S. Khim, T. Wang, J.E. Naile, J. Park, B.O. Kwon, S.J. Song, J. Ryu, G. Codling,
P.D. Jones, Y. Lu and J.P. Giesy. 2015. Bioaccumulation characteristics of perfluoroalkyl acids
(PFAAs) in coastal organisms from the west coast of South Korea. Chemosphere. 129:157-63.
Hoover, G.M. 2018. Effects of per/polyfluoroalkyl substance exposure on larval amphibians.
Ph.D. Thesis, Purdue University, West Lafayette, IN: 119 p.
Hoover, G.M., M.F. Chislock, B.J. Tornabene, S.C. Guffey, Y.J. Choi, C.D. Perre, J.T.
Hoverman, L.S. Lee and M.S. Sepulveda. 2017. Uptake and depuration of four
per/polyfluoroalkyl substances (PFASS) in Northern leopard frog Ranapipiens tadpoles.
Environ. Sci. Technol. Letters. 4(10): 399-403.
Hoover, G., S. Kar, S. Guffey, J. Leszczynski, and M.S. Sepulveda. 2019. In vitro and in silico
modeling of perfluoroalkyl substances mixture toxicity in an amphibian fibroblast cell line.
Chemosphere 233: 25-33.
Houde, M., J.W. Martin, R.J. Letcher, K.R. Solomon and D.C. Muir. 2006a. Biological
monitoring of polyfluoroalkyl substances: A review. Environ. Sci. Technol. 40(11): 3463-3473.
Houde, M., A. Trevor, D. Bujas, J. Small, R.S. Wells, P.A. Fair, G.D. Bossart, K.R. Solomon
and D. Muir. 2006b. Biomagnification of perfluoroalkyl compounds in the bottlenose dolphin
(Tursiops truncatus) food web. Environ. Sci. Technol. 40(13): 4138-4144.
Houde, M., G. Czub, J.M. Small, S. Backus, X. Wang, M. Alaee and D. C. Muir. 2008.
Fractionation and bioaccumulation of perfluorooctane sulfonate (PFOS) isomers in a Lake
Ontario food web. Environ. Sci. Technol. 42(24): 9397-9403.
150
-------
HSDB (Hazardous Substances Data Bank). 2012. Perfluorooctanoic acid. Accessed May 2016.
Hu, C., Q. Luo and Q. Huang. 2014. Ecotoxicological effects of perfluorooctanoic acid on
freshwater microalgae Chlamydomonas reinhardtii and Scenedesmus obliquus. Environ. Toxicol.
Chem. 33(5): 1129-1134.
Hu, J., D. Wang, N. Zhang, K. Tang, Y. Bai, Y. Tian, Y. Li, and X. Zhang. 2023. Effects of
perfluorooctanoic acid on Microcystis aeruginosa: Stress and self-adaptation mechanisms. J.
Hazard. Mater. 445: 12 p.
Hu, Y., F.L. Meng, Y.Y. Hu, N. Habibul and G.P. Sheng. 2020. Concentration- and nutrient-
dependent cellular responses of microalgae Chlorellapyrenoidosa to perfluorooctanoic acid.
Water Res. 185:116248-116248.
Huset, C.A., A.C. Chiaia, D.F. Barofsky, N. Jonkers, H. P.E. Kohler, C. Ort, W. Giger and J. A.
Field. 2008. Occurrence and mass flows of fluorochemicals in the Glatt Valley watershed,
Switzerland. Environ. Sci. Technol. 42(17): 6369-6377.
Inoue, Y., N. Hashizume, N. Yakata, H. Murakami, Y. Suzuki, E. Kikushima and M. Otsuka.
2012. Unique physicochemical properties of perfluorinated compounds and their
bioconcentration in common carp Cyprinus carpio. Arch. Environ. Contam. Toxicol. 62(4): 672-
680.
Jantzen, C.E., K.M. Annunziato and K.R. Cooper. 2016. Behavioral, morphometric, and gene
expression effects in adult zebrafish (Danio rerio) embryonically exposed to PFOA, PFOS, and
PFNA. Aquat. Toxicol. 180: 123-130.
Jantzen, C.E., K.A. Annunziato, S.M. Bugel and K.R. Cooper. 2017a. PFOS, PFNA and PFOA
sub-lethal exposure to embryonic zebrafish have different toxicity profiles in terms of
morphometries behavioral and gene expression. Aquat. Toxicol. 175: 168-170.
Jantzen, C.E., F. Toor, K.A. Annunziato and K.R. Cooper. 2017b. Effects of chronic
perfluorooctanoic acid (PFOA) at low concentration on morphometries, gene expression, and
fecundity in zebrafish (Danio rerio). Reproduct. Toxicol. 69: 34-42.
Jarvis, A.L., J.R. Justice, B. Schnitker and K. Gallagher. 2023. Meta-analysis comparing nominal
and measured concentrations of perfluorooctanoic acid and perfluorooctane sulfonate in aquatic
toxicity studies across various experimental conditions. Environ. Toxicol. Chem. 42(11): 2289-
2301.
Jeon, J., K. Kannan, H.K. Lim, H.B. Moon, J.S. Ra and S.D. Kim. 2010. Bioaccumulation of
perfluorochemicals in Pacific oyster under different salinity gradients. Environ. Sci. Technol.
44(7): 2695-2701.
151
-------
Ji, K., Y. Kim, S. Oh, B. Ahn, H. Jo and K. Choi. 2008. Toxicity of perfluorooctane sulfonic acid
and perfluorooctanoic acid on freshwater macroinvertebrates (Daphnia magna and Moina
macrocopa) and fish (Oryzias latipes). Environ. Toxicol. Chem. 27(10): 2159-2168.
Jones, P.D., W. Hu, W. De Coen, J.L. Newsted and J.P. Giesy. 2003. Binding of perfluorinated
fatty acids to serum proteins. Environ. Toxicol. Chem.: An International Journal. 22(11): 2639-
2649.
Johnson, B. R., P. C. Weaver, C. T. Nietch, J. M. Lazorchak, K. A. Struewing and D. H. Funk.
2015. Elevated major ion concentrations inhibit larval mayfly growth and development. Environ.
Toxicol. Chem. 34: 167-172.
Kadlec, S.M., W.J. Backe, R.J. Erickson, R. Hockett, S.E. Howe, I D. Mundy, E. Piasecki, H.
Sluka, L.K. Votava, and D.R. 2024. Sublethal toxicity of 17 per- and polyfluoroalkyl substances
with diverse structures to Ceriodaphnia dubia, Hyalella azteca, and Chironomus dilutus.
Environ. Toxicol. Chem. 43(2): 359-373.
Kaiser, M.A., B.S. Larsen, C-P.C. Kao and R.C. Buck. 2005. Vapor pressures of
perfluorooctanoic, -nonanoic, -decanoic, undecanoic, and dodecanoic acids. Jo. Chem. Eng.
Data. 50(6): 1841-1843.
Kalasekar, S.M., E. Zacharia, N. Kessler, N.A. Ducharme, J. Gustafsson, I.A. Kakadiaris and M.
Bondesson. 2015. Identification of environmental chemicals that induce yolk malabsorption in
zebrafish using automated image segmentation. Reproduct. Toxicol. 55: 20-29.
Kang, J.S., T.G. Ahn and J.W. Park. 2019. Perfluorooctanoic acid (PFOA) and perfluooctane
Sulfonate (PFOS) induce different modes of action in reproduction to Japanese medaka {Oryzias
latipes). J. Hazard. Mater. 368: 97-103.
Kannan, K. 2011. Perfluoroalkyl and polyfluoroalkyl substances: current and future perspectives.
Environ. Chem. 8(4): 333-338.
Kannan, K., L. Tao, E. Sinclair, S.D. Pastva, D.J. Jude and J.P. Giesy. 2005. Perfluorinated
compounds in aquatic organisms at various trophic levels in a Great Lakes food chain. Arch.
Environ. Contam. Toxicol. 48(4): 559-566.
Keiter, S., L. Baumann, H. Farber, H. Holbech, D. Skutlarek, M. Engwall and T. Braunbeck.
2012. Long-term effects of a binary mixture of perfluorooctane sulfonate (PFOS) and bisphenol
A (BPA) in zebrafish (Danio rerio). Aquat. Toxicol. 118/119: 116-129.
Kelly B.C., F.A. Gobas and M.S. McLachlan. 2004. Intestinal absorption and biomagnification
of organic contaminants in fish, wildlife and humans. Environ. Toxicol. Chem. 23(10): 2324-
2336.
152
-------
Kelly, B.C., M.G. Ikonomou, J.D. Blair, B. Surridge, D. Hoover, R. Grace and F.A. Gobas.
2009. Perfluoroalkyl contaminants in an arctic marine food web: Trophic magnification and
wildlife exposure. Environ. Sci. Technol. (43): 4037-4043.
Khan, E.A., X. Zhang, E.M. Hanna, F. Yadetie, I. Jonassen, A. Goksoyr and A. Arukwe. 2021.
Application of quantitative transcriptomics in evaluating the ex vivo effects of per- and
polyfluoroalkyl substances on Atlantic cod (Gadus morhua) ovarian physiology. Sci. Total
Environ. 755(1): lip.
Kim, J., G. Lee, V.M. Lee, K.D. Zoh and K. Choi. 2021. Thyroid disrupting effects of
perfluoroundecanoic acid and perfluorotridecanoic acid in zebrafish (Danio rerio) and rat
pituitary (GH3) cell line. Chemosphere. 262: 8.
Kim, K.S., D.H. Funk and D.B. Buchwalter. 2012. Dietary (periphyton) and aqueous Zn
bioaccumulation dynamics in the mayfly Centroptilum triangulifer. Ecotoxicol. 21: 2288-2296.
Kim, M., J. Son, M.S. Park, Y. Ji, S. Chae, C. Jun, J.S. Bae, T.K. Kwon, Y.S. Choo, H. Yoon, D.
Yoon, J. Ryoo, S.H. Kim, M.J. Park and H.S. Lee. 2013. In vivo evaluation and comparison of
developmental toxicity and teratogenicity of perfluoroalkyl compounds using Xenopus embryos.
Chemosphere. 93: 1153-1160.
Kim, S.-K. and K. Kannan. 2007. Perfluorinated acids in air, rain, snow, surface runoff, and
lakes: relative importance of pathways to contamination of urban lakes. Environ. Sci. Technol.
41(24): 8328-8334.
Kim, W.K., S.K. Lee and J. Jung. 2010. Integrated assessment of biomarker responses in
common carp (Cyprinus carpio) exposed to perfluorinated organic compounds. J. Haz. Mat. 180:
395-400.
Kimmel, C.B., WW. Ballard, S.R. Kimmel, B. Ullmann and T.F. Schilling. 1995. Stages of
embryonic development of the zebrafish. Dev. Dyn. 203: 253-310.
Konwick, B.J., G.T. Tomy, N. Ismail, J.T. Peterson, R.J. Fauver, D. Higginbotham and A.T.
Fisk. 2008. Concentrations and patterns of perfluoroalkyl acids in Georgia, USA surface waters
near and distant to a major use source. Environ. Toxicol. Chem. 27(10): 2011-2018.
Kudo, N., M. Katakura, Y. Sato and Y. Kawashima. 2002. Sex hormone-regulated renal
transport of perfluorooctanoic acid. Chem. Biol. Interact. 139(3): 301-16.
Kwadijk, C., P. Korytar and A. Koelmans. 2010. Distribution of perfluorinated compounds in
aquatic systems in the Netherlands. Environ. Sci. Technol. 44(10): 3746-3751.
Labine, L.M., E.A. Oliveira Pereira, S. Kleywegt, K.J. Jobst, A.J. Simpson, and M.J. Simpson.
2022. Comparison of sub-lethal metabolic perturbations of select legacy and novel perfluorinated
alkyl substances (PFAS) in Daphnia magna. Environ. Res. 212: 12 p.
153
-------
Lasier, P.J., J.W. Washington, S.M. Hassan and T.M. Jenkins. 2011. Perfluorinated chemicals in
surface waters and sediments from northwest Georgia, USA, and their bioaccumulation in
Lumbriculus variegatus. Environ. Toxicol. Chem. 30(10): 2194-2201.
Latala, A., M. Nedzi and P. Stepnowski. 2009. Acute toxicity assessment of perfluorinated
carboxylic acids towards the Baltic microalgae. Environ. Toxicol. Pharmacol. 28: 167-171.
Lau, C., K. Anitole, C. Hodes, D. Lai, A. Pfahles-Hutchens and J. Seed. 2007. Perfluoroalkyl
acids: a review of monitoring and toxicological findings. Toxicol. Sci. 99(2): 366-394.
Lee, H., J. D'eon and S.A. Mabury. 2010. Biodegradation of polyfluoroalkyl phosphates as a
source of perfluorinated acids to the environment. Environ. Sci. Technol. 44(9): 3305-3310.
Lee, J.J. and I.R. Schultz. 2010. Sex differences in the uptake and disposition of
perfluorooctanoic acid in fathead minnows after oral dosing. Environ. Sci. Technol. 44(44): 491-
496.
Lee, J.W., J.W. Lee, K. Kim, Y.J. Shin, J. Kim, S. Kim, H. Kim, P. Kim and K. Park. 2017.
PFOA-induced metabolism disturbance and multi-generational reproductive toxicity in Oryzias
latipes. J. Haz. Mat. 340: 231-240.
Lee, J.W., P.K. Seong, S.D. Yu and P. Kim. 2020. Adverse effects of perfluoroalkyl acids on fish
and other aquatic organisms: A review. Sci. Total Environ. 707: 135334.
Lee, W. and Y. Kagami. 2010. Effects of perfluorooctanoic acid and perfluorooctane sulfonate
on gene expression profiles in medaka (Oryzias latipes). Abstracts. Toxicol. Letters. 196S: S37-
S351.
Lemly, A. D. 1997. Ecosystem recovery following selenium contamination in a freshwater
reservoir. Ecotox. And Environ. Safety. 36(3): 275-281.
Lescord, G.L., K.A. Kidd, A.O. De Silva, M. Williamson, C. Spencer, X.W. Wang and D.C.G.
Muir. 2015. Perfluorinated and polyfluorinated compounds in lake food webs from the Canadian
High Arctic. Environ. Sci. Technol. 49: 2694-2702.
Li, F., X.L. Fang, Z.M. Zhou, X.B. Liao, J. Zou, B.L. Yuan and W.J. Sun. 2019. Adsorption of
perfluorinated acids onto soils: Kinetics, isotherms, and influences of soil properties. Sci. Tot.
Environ. 649: 504-514.
Li, F., Y. Yu, M. Guo, Y. Lin, Y. Jiang, M. Qu, X. Sun, Z. Li, Y. Zhai and Z. Tan. 2021a.
Integrated analysis of physiological, transcriptomics and metabolomics provides insights into
detoxication disruption of PFOA exposure in Mytilus edulis. Ecotoxicol. Environ. Saf. 214: 11
pp.
154
-------
Li, H., D. Ellis and D. Mackay. 2007. Measurement of low air-water partition coefficients of
organic acids by evaporation from a water surface. Jour. Chem. Engineer. Data. 52(5): 1580-
1584.
Li, M.H. 2008. Effects of nonionic and ionic surfactants on survival, oxidative stress, and
cholinesterase activity of planarian. Chemosphere. 70(10): 1796-1803.
Li, M.H. 2009. Toxicity of perfluorooctane sulfonate and perfluorooctanoic acid to plants and
aquatic invertebrates. Environ. Toxicol. 24(1): 95-101.
Li, M.H. 2010. Chronic effects of perfluorooctane sulfonate and ammonia perfluorooctanoate on
biochemical parameters, survival and reproduction of Daphnia magna. J. Health Sci. 56(1): 104-
111.
Li, M.H. 2011. Changes of cholinesterase and carboxylesterase activities in male guppies,
Poecilia reticulate, after exposure to ammonium perfluorooctanoate, but not to perfluorooctane
sulfonate. Fresenius Environ. Bull. 20(8a): 2065-2070.
Li, Y., T. Fletcher, D. Mucs, K. Scott. C.H. Lindh, P. Tallving and K. Jakobsson. 2017. Half-
lives of PFOS, PFHxS and PFOA after end of exposure to contaminated drinking water. Occup.
Environ. Med. 75: 46-51.
Li, Y.S., D.P. Oliver and R.S. Kookana. 2018a. A critical analysis of published data to discern
the role of soil and sediment properties in determining sorption of per and polyfluoroalkyl
substances (PFASs). Sci. Tot. Environ. 628-629: 110-120.
Li, Y., J. Wang, M. Zheng, Y. Zhang and S. Ru. 2018b. Development of ELISAs for the
detection of vitellogenin in three marine fish from coastal areas of China. Mar. Pollut. Bull.
133:415-422.
Li, Y., X. Liu, X. Zheng, M. Yang, X. Gao, J. Huang, L. Zhang and Z. Fan. 2021b. Toxic effects
and mechanisms of PFOA and its substitute GenX on the photosynthesis of Chlorella
pyrenoidosa. Sci. Tot. Environ.765: 144431.
Liang, X. and J. Zha. 2016. Toxicogenomic applications of Chinese rare minnow (Gobiocypris
rarus) in aquatic toxicology. Comp. Biochem. Physiol. PartD. 19: 174-180.
Lide, D.R. 2007. CRC Handbook of Chemistry and Physics. 88th ed. CRC Press, Taylor &
Francis, Boca Raton, FL. pp. 3-412.
Lim, J. 2022. Broad toxicological effects of per-/poly- fluoroalkyl substances (PFAS) on the
unicellular eukaryote, Tetrahymenapyriformis. Environ. Toxicol. Pharmacol. 95: 7 p.
Lin, H., Y. Feng, Y. Zheng, Y. Han, X. Yuan, P. Gao, H. Zhang, Y. Zhong, and Z. Liu. 2022a.
Transcriptomic analysis reveals the hepatotoxicity of perfluorooctanoic acid in black-spotted
frogs (Rana nigromaculata). Diversity 14(11): 10 p.
155
-------
Lin, H., Z. Liu, H. Yang, L. Lu, R. Chen, X. Zhang, Y. Zhong, and H. Zhang. 2022b. Per- and
polyfluoroalkyl substances (PFASs) impair lipid metabolism in Rana nigromaculata: A field
investigation and laboratory study. Environ. Sci. Technol. 56(18): 13222-13232.
Lin, H., H. Wu, F. Liu, H. Yang, L. Shen, J. Chen, X. Zhang, Y. Zhong, H. Zhang, and Z. Liu.
2022c. Assessing the hepatotoxicity of PFOA, PFOS, and 6:2 Cl-PFESA in black-spotted frogs
{Rana nigromaculata) and elucidating potential association with gut microbiota. Environ. Pollut.
312: 11 p.
Lindqvist, D. and E. Wincent. 2022. Kinetics and toxicity of an environmentally relevant mixture
of halogenated organic compounds in zebrafish embryo. Aquat. Toxicol. 252: 9 p.
Liou, J.S.C., B. Szostek, C.M. DeRito and E.L. Madsen. 2010. Investigating the biodegradability
of perfluorooctanoic acid. Chemosphere. 80: 176-83.
Liu, C. and K.Y.H Gin. 2018. Immunotoxicity in green mussels under perfluoroalkyl substance
(PFAS) exposure: Reversible response and response model development. Environ. Toxicol.
Chem. 37(4): 1138-1145.
Liu, C., Y. Du and B. Zhou. 2007a. Evaluation of estrogenic activities and mechanism of action
of perfluorinated chemicals determined by vitellogenin induction in primary cultured tilapia
hepatocytes. Aquat. Toxicol. 85: 267-277.
Liu, C., K. Yu, X. Shi, J. Wang, P.K.S. Lam, R.S.S. Wu and B. Zhou. 2007b. Induction of
oxidative stress and apoptosis by PFOS and PFOA in primary cultured hepatocytes of freshwater
tilapia {Oreochromis niloticus). Aquat. Toxicol. 82: 135-143.
Liu, C., V.W. Chang and K.Y. Gin. 2013. Environmental toxicity of PFCs: An enhanced
integrated biomarker assessment and structure-activity analysis. Environ. Toxicol. Chem. 32(10):
2226-2233.
Liu, C., V.W.C. Chang, K.Y.H. Gin and V.T. Nguyen. 2014a. Genotoxicity of perfluorinated
chemicals (PFCs) to the green mussel {Perna viridis). Sci. Tot. Environ. 487: 117-122.
Liu, C., V.W.C Chang and K.Y.H. Gin. 2014b. Oxidative toxicity of perfluorinated chemicals in
green mussel and bioaccumulation factor dependent quantitative structure-activity relationship.
Environ. Toxicol. Chem. 33(10): 2323-2332.
Liu, C., K.Y.H Gin and V.W.C. Chang. 2014c. Multi-biomarker responses in green mussels
exposed to PFCs: Effects at molecular, cellular, and physiological levels. Environ. Sci. Pollut.
Res. 21: 2785-2794.
Liu, G., X. Yan, C. Li, S. Hu, J. Yan, and B. Yan. 2023a. Unraveling the joint toxicity of
transition-metal dichalcogenides and per- and polyfluoroalkyl substances in aqueous mediums by
experimentation, machine learning and molecular dynamics. J. Hazard. Mater. 443: 13 p.
156
-------
Liu, H., Y. Chen, W. Hu, Y. Luo, P. Zhu, S. You, Y. Li, Z. Jiang, X. Wu, and X. Li. 2023b.
Impacts of PFOAC8, GenXC6, and their mixtures on zebrafish developmental toxicity and gene
expression provide insight about tumor-related disease. Sci. Total Environ. 858: 13 p.
Liu, J. and S. Mejia Avendano. 2013. Microbial degradation of polyfluoroalkyl chemicals in the
environment: A review. Environ. Intern. 61: 98-114.
Liu, J., N. Wang, B. Szostek, R.C. Buck, P.K. Panciroli, P.W. Folsom, L.M. Sulecki and C.A.
Bellin. 2010. 6-2 Fluorotelomer alcohol aerobic biodegradation in soil and mixed bacterial
culture. Chemosphere. 78: 437-444.
Liu, S., L. Yan, Y. Zhang, M. Junaid, and J. Wang. 2022. Toxicological effects of polystyrene
nanoplastics and perfluorooctanoic acid to Gambusia affinis. Fish Shellfish Immunol. 127: 1100-
1112.
Liu, W., S. Chen, X. Quan and Y.H. Jin. 2008a. Toxic effect of serial perfluorosulfonic and
perfluorocarboxylic acids on the membrane system of a freshwater alga measured by flow
cytometry. Environ. Toxicol. Chem. 27(7): 1597-1604.
Liu, Y., J. Wang, Y. Wei, H. Zhang, Y. Liu and J. Dai. 2008b. Molecular characterization of
cytochrome P450 1A and 3A and the effects of perfluorooctanoic acid on their mRNA levels in
rare minnow {Gobiocypris rarus). Aquat. Toxicol. 88: 183-190.
Liu, Y., J. Wang, Y. Liu, H. Zhang, M. Xu, and J. Dai. 2009. Expression of a novel cytochrome
9450 4T gene in rare minnow {Gobiocypris rarus). Comp. Biochem. Physiol. Part C. 150: 57-64.
Liu, Z., H. Lin, Y. Zheng, Y. Feng, C. Shi, R. Zhu, X. Shen, Y. Han, H. Zhang, and Y. Zhong.
2023c. Perfluorooctanoic acid and perfluorooctanesulfonic acid induce immunotoxicity through
the NF-kappaB pathway in black-spotted frog (Rana nigromaculata). Chemosphere 313: 8 p.
Logeshwaran, P., A.K. Sivaram, A. Surapaneni, K. Kannan, R. Naidu and M. Megharaj. 2021.
Exposure to perfluorooctanesulfonate (PFOS) but not perfluorooctanoic acid (PFOA) at ppb
concentration induces chronic toxicity in Daphnia carinata. Sci. Tot. Environ.769: 8 p.
Loi, E.I., L.W. Yeung, S. Taniyasu, P.K. Lam, K. Kannan andN. Yamashita. 2011. Trophic
magnification of poly- and perfluorinated compounds in a subtropical food web. Environ. Sci.
Technol. 45: 5506-5513.
Loos, R., J. Wollgast, T. Huber and G. Hanke. 2007. Polar herbicides, pharmaceutical products,
perfluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA), and nonylphenol and its
carboxylates and ethoxylates in surface and tap waters around Lake Maggiore in Northern Italy.
Anal. Bioanal. Chem. 387(4): 1469-1478.
157
-------
Loos, R., B.M. Gawlik, G. Locoro, E. Rimaviciute, S. Contini and G. Bidoglio. 2009. EU-wide
survey of polar organic persistent pollutants in European river waters. Environ. Pollut. 157(2):
561-568.
Lu, G.H., B.H. Ma, S. Li and L.S. Sun. 2016. Toxicological Effects of Perfluorooctanoic Acid
(PFOA) on Daphnia magna. Mat. Sci. Environ. Eng. Pp. 559-564.
MacDonald, M.M., A.L. Warne, N.L. Stock, S.A. Mabury, K.R. Solomon and P.K. Sibley. 2004.
Toxicity of perfluorooctane sulfonic acid and perfluorooctanoic acid to Chironomus tentans.
Environ. Toxicol. Chem. 23(9): 2116-2123.
Mahapatra, C.T., N.P. Damayanti, S.C. Guffey, J.S. Serafin, J. Irudayaraj and M.S. Sepulveda.
2017. Comparative in vitro toxicity assessment of perfluorinated carboxylic acids. J. Applied
Toxicol. 37: 699-708.
Manera, M., L. Giari, F. Vincenzi, C. Guerranti, J.A. DePasquale and G. Castaldelli. 2017.
Texture analysis in liver of common carp (Cyprinus carpio) sub-chronically exposed to
perfluorooctanoic acid. Ecol. Indicators. 81: 54-64.
Manera, M., G. Castaldelli, and L. Giari. 2022a. Perfluorooctanoic acid affects thyroid follicles
in common carp (Cyprinus carpio). Int. J. Environ. Res. Public Health 19(15): 9049-9060.
Manera, M., G. Castaldelli, C. Guerranti and L. Giari. 2022b. Effect of waterborne exposure to
perfluorooctanoic acid on nephron and renal hemopoietic tissue of common carp Cyprinus
carpio. Ecotoxicol. Environ. Saf. 234: 10.
Mao, W., M. Li, X. Xue, W. Cao, X. Wang, F. Xu, and W. Jiang. 2023. Bioaccumulation and
toxicity of perfluorooctanoic acid and perfluorooctanesulfonate in marine algae Chlorella sp.
Sci. Total Environ. 870: 10 p.
Marchetto, F., M. Roverso, D. Righetti, S. Bogialli, F. Filippini, E. Bergantino and E. Sforza.
2021. Bioremediation of per- and poly-fluoroalkyl substances (PFAS) by Synechocystis sp. PCC
6803: A chassis for a synthetic biology approach. Life (Basel). 11(12): 18.
Martin, J.W., S.A. Mabury, K.R. Solomon and D.C.G. Muir. 2003a. Bioconcentration and tissue
distribution of perfluorinated acids in rainbow trout (Oncorhynchus mykiss). Environ. Toxicol.
Chem. 22: 196-204.
Martin, J.W., S.A. Mabury, K.R. Solomon and D.C. Muir. 2003b. Dietary accumulation of
perfluorinated acids in juvenile rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem.
22(1): 189-195.
Martin, J.W., M.M. Smithwick, B.M. Braune, P.F. Hoekstra, D.C. Muir and S.A. Mabury. 2004.
Identification of long-chain perfluorinated acids in biota from the Canadian Arctic. Environ. Sci.
Tech. 38(2): 373-380.
158
-------
Martin, J.W., J.A. Brian, S. Beeson, J.D. Benskin and M.S. Ross. 2010. PFOS of PreFOS? Are
perfluorooctane sulfonate precursors (PreFOS) important determinants of human and
environmental perfluorooctoane sulfonate (PFOS) exposure? J. Envrion. Monit. 12: 1979-2004.
Martin, J.W., S.A. Mabury, K.R. Solomon and D.C.G. Muir. 2013. Progress toward
understanding the bioaccumulation of perfluorinated alkyl acids. Environ. Toxicol. Chem.
32(11): 2421-2423.
Martin, O., M. Scholze, S. Ermler, J. McPhie, S.K. Bopp, A. Kienzler, N. Parissis and A.
Kortenkamp. 2021. Ten years of research on synergisms and antagonisms in chemical mixtures:
a systematic review and quantitative reappraisal of mixture studies. Environ. Intern. 146:
106206.
Marziali, L., F. Rosignoli, S. Valsecchi, S. Polesello and F. Stefani. 2019. Effects of
perfluoralkyl substances (PFASs) on a multigenerational scale: A case study with Chironomus
riparius (Diptera, Chironomidae). Environ. Toxicol. Chem. 38(5): 988-999.
McCarthy, C.J., S.A. Roark, D. Wright, K. O'Neal, B. Muckey, M. Stanaway, J. Rewerts, J.
Field, T. Anderson and C.J. Sa. 2021. Toxicological response of Chironomus dilutus in single
chemical and binary mixture exposure experiments with 6 perfluoralkyl substances. Environ.
Toxicol. Chem. 40(8): 2319-2333.
Mebane, C. A. 2022. The capacity of freshwater ecosystems to recover from exceedences of
aquatic life criteria. Environ. Toxicol. Chem. 41(12): 2887-2910.
Mebane, C. A., T. S. Schmidt, J. L. Miller and L. S. Balistrieri. 2020. Bioaccumulation and
toxicity of cadmium, copper, nickel and zinc and their mixtures to aquatic insect communities.
Environ. Toxicol. Chem. 39: 812-833.
Mejia-Avendano, S., S. Vo Duy, S. Sauve and J. Liu. 2016. Generation of perfluoroalkyl acids
from aerobic biotransformation of quaternary ammonium polyfluoroalkyl surfactants. Environ.
Sci. Technol. 50(18): 9923-9932.
Menger, F., J. Pohl, L. Ahrens, G. Carlsson and S. Orn. 2020. Behavioural effects and
bioconcentration of per- and polyfluoroalkyl substances (PFASs) in zebrafish (Danio rerio)
embryos. Chemosphere. 245:11.
Mhadhbi, L., D. Rial, S. Perez and R. Beiras. 2012. Ecological risk assessment of
perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) in marine environment
using Isochrysis galbana, Paracentrotus lividus, Siriella armata and Psetta maxima. J. Environ.
Monit. 14(5): 1375-1382.
Miranda, A.F., C. Trestrail, S. Lekamge and D. Nugegoda. 2020. Effects of perfluorooctanoic
acid (PFOA) on the thyroid status, vitellogenin, and oxidant-antioxidant balance in the Murray
River rainbowfish. Ecotox. 29(2): 163-174.
159
-------
MPCA (Minnesota Pollution Control Agency). 2008. PCFs in Minnesota's Ambient
Environment: 2008 Progress Report.
Moody, C.A., J.W. Martin, W.C. Kwan, D.C. Muir and S.A. Mabury. 2002. Monitoring
perfluorinated surfactants in biota and surface water samples following an accidental release of
fire-fighting foam into Etobicoke Creek. Environ. Sci. Technol. 36(4): 545-551.
Moody, C.A., G.N. Hebert, S.H. Strauss and J.A. Field. 2003. Occurrence and persistence of
perfluorooctanesulfonate and other perfluorinated surfactants in groundwater at a fire-training
area at Wurtsmith Air Force Base, Michigan, USA. J. Environ. Monit. 5(2): 341-345.
Mortensen, A.S., R.J. Letcher, M.V. Cangialosi, S. Chu and A. Arukwe. 2011. Tissue
bioaccumulation patterns, xenobioticbiotransformation and steroid hormone levels in Atlantic
salmon (Salmo salaf) fed a diet containing perfluoroactane sulfonic or perfluorooctane
carboxylic acids. Chemosphere. 83: 1035-1044.
Mount, D.R., R.J. Erickson, T.L. Highland, J.R. Hockett, D.J. Hoff, C.T. Jenson, T.J. Norberg-
King, K.N. Peterson and S. Wisniewski. 2016. The acute toxicity of major ion salts to
Ceriodaphnia dubia. I. The influence of background water chemistry. Environ. Toxicol. Chem.
35: 3039-3057.
Myosho, T., A. Ishibashi, S. Fujimoto, S. Miyagawa, T. Iguchi, and T. Kobayashi. 2022. Preself-
feeding medaka fry provides a suitable screening system for in vivo assessment of thyroid
hormone-disrupting potential. Environ. Sci. Technol. 56(10): 6479-6490.
Nagel, R. 2002. DarT: the embryo test with the zebrafish Danio rerio - a general model in
ecotoxicology and toxicology. ALTEX 19(Suppll): 38-48.
Nakata, H., K. Kannan, T. Nasu, H.-S. Cho, E. Sinclair and A. Takemura. 2006. Perfluorinated
contaminants in sediments and aquatic organisms collected from shallow water and tidal flat
areas of the Ariake Sea, Japan: environmental fate of perfluorooctane sulfonate in aquatic
ecosystems. Environ. Sci. Technol. 40(16): 4916-4921.
Nakayama, S., M.J. Strynar, L. Helfant, P. Egeghy, X. Ye and A. B. Lindstrom. 2007.
Perfluorinated compounds in the Cape Fear drainage basin in North Carolina. Environ. Sci.
Technol. 41(15): 5271-5276.
Newsted, J.L., R. Holem, G. Hohenstein, C. Lange, M. Ellefson, W. Reagen and S. Wolf. 2017.
Spatial and temporal trends of poly- and perfluoroalkyl substances in fish fillets and water
collected from pool 2 of the Upper Mississippi River. Environ. Toxicol. Chem. 36(11): 3138-
3147.
Nguyen, V. T., M. Reinhard and G. Y. Karina. 2011. Occurrence and source characterization of
perfluorochemicals in an urban watershed. Chemosphere. 82(9): 1277-1285.
160
-------
NJDEP. (New Jersey Department of Environmental Protection). 2019. Investigation of Levels of
Perfluorinated Compounds in New Jersey Fish, Surface Water, and Sediment. New Jersey
Department of Environmental Protection Division of Science, R., and Environmental
Health,SRI 5-010. Pp. 1-46.
https://www.ni.gov/dep/dsr/publications/Investigation%20of%20Levels%20of%20Perfluorinated
%20Compounds%20in%20New%20Jersev%20Fish.%20Surface%20Water.%20and%20Sedime
nt.pdf
NMED (New Mexico Environment Department). 2021. PFAS. Data. Accessed January 2021.
Available online at: https://www.env.nm.gov/pfas/data/.
Norwegian Institute for Air Research. 2007a. ISBN 978-82-425-1962-7.
Norwegian Institute for Air Research. 2007b. ISBN 978-82-425-1984-9.
NRC (National Research Council). 2013. Assessing risks to endangered and threatened species
from pesticides; National Academies Press, National Research Council. Washington, DC. 142
pp.
Oakes, K.D, P.K. Sibley, K.R. Solomon, S.A. Mabury and G.J. Van Der Kraak. 2004. Impact of
perfluorooctanoic acid on fathead minnow (Pimephalespromelas) fatty acyl-CoA oxidase
activity, circulating steroids, and reproduction in outdoor microcosms. Environ. Toxicol. Chem.
23(8): 1912-1919.
OECD. 1984. Guideline for testing of chemicals 202 -Daphnia sp., Acute Immobilization Test
and Reproduction Test.
OECD. 1992. Test No. 203: Fish, Acute Toxicity Test. OECD Guidelines for the Testing of
Chemicals, Section 2, OECD Publishing, Paris, https://doi.org/10.1787/9789264069961-en.
OECD. 1997. Guideline 211: Daphnia magna reproduction test.
OECD. 1998. Daphnia magna reproduction test. OECD Guideline 211. Paris, France.
OECD. 2000. Guideline 202: Daphnia magna, Acute Immobilisation Test, Updated Guideline,
October 2000. Guidelines for the testing of chemicals.
OECD. 2004. Guideline for Testing of Chemicals - Daphnia sp., Acute Immobilization Test
202.
OECD. 2006. Lists of PFOS, PFAS, PFOA, PFCA, related compounds and chemicals that may
degrade to PFCA. Report ENV/JM/MONO (2006). Organisation for Economic Co-operation and
Development; 2007. 15 pp.
161
-------
OECD. 2011. OECD Guidelines for the testing of chemicals: freshwater alga and cyanobacteria,
growth inhibition test. OECD Guidelines for the Testing of Chemicals, Section 2: Effects on
Biotic Systems. OECD Publishing, Paris.
OECD. 2012. Test no. 211: Daphnia magna reproduction test. OECD Guidelines for the Testing
Chemicals, Section 2: Effect on Biotic Systems. OECD Publishing, Paris.
OECD. 2013. Test No. 236: Fish Embryo Acute Toxicity (FET) Test. OECD Guidelines for the
Testing Chemicals, Section 2: Effect on Biotic Systems. OECD Publishing, Paris. 22 pp.
OECD. 2021. Reconciling Terminology of the Universe of Per- and Polyfluoroalkyl Substances:
Recommendations and Practical Guidance, OECD Series on Risk Management, No. 61, OECD
Publishing, Paris
Oh, J.H., H.B. Moon and E.S. Choe. 2013. Alterations in differentially expressed genes after
repeated exposure to perfluorooctanoate and perfluorooctanesulfonate in liver of Oryzias latipes.
Arch. Environ. Contam. Toxicol. 64(3): 475-483.
Oliaei, F., D. Kriens and R. Weber. 2013. PFOS and PFC releases and associated pollution from
a PFC production plant in Minnesota (USA). Environ. Sci. Pollut. Res. 20: 1977-1992.
Otero-Sabio, C., M. Giacomello, C. Centelleghe, F. Caicci, M. Bonato, A. Venerando, J.M.
Graic, S. Mazzariol, L. Finos. 2022. Cell cycle alterations due to perfluoroalkyl substances
PFOS, PFOA, PFBS, PFBA and the new PFAS C604 on bottlenose dolphin (Tursiops
truncatus) skin cell. Ecotoxicol. Environ. Saf. 244: 10 p.
Pacchini, S., E. Piva, S. Schumann, P. Irato, D. Pellegrino, and G. Santovito. 2023. An
experimental study on antioxidant enzyme gene expression in Trematomus newnesi (Boulenger,
1902) experimentally exposed to perfluoro-octanoic acid. Antioxidants (Basel 12(352): 10 p.
Padilla, S., D. Corum, B. Padros, D.L. Hunter, A. Beam, K.A. Houck, N. Sipes, N. Kleinstreuer,
T. Knudsen, D.J. Nix and D.M. Reif. 2012. Zebrafish developmental screening of the ToxCast
Phase I chemical library. Reprod. Toxicol. 33: 174-187.
Palumbo, A.J., P.L. TenBrook, T.L. Fojut, I.R. Faria and R.S. Tjeerdema. 2012. Aquatic life
water criteria derived via the UC Davis Method: I. Organophosphate Insecticides. In. Tjeerdema,
R.S., ed., Springer, NY, NY. Rev. Environ. Contam. Toxicol. 216. Pp. 1-50.
Pan, Y., H. Zhang, Q. Cui, N. Sheng, L. W. Y. Yeung, Y. Sun, Y. Guo and J. Dai. 2018.
Worldwide distribution of novel perfluoroether carboxylic and sulfonic acids in surface water.
Environ. Sci. Technol. 52(14): 7621-7629.
Pecquet, A.M., A. Maier, S. Kasper, S. Sumanas and J. Yadav. 2020. Exposure to
perfluorooctanoic acid (PFOA) decreases neutrophil migration response to injury in zebrafish
embryos. BMC Res. Notes 13(1): 6 pp.
162
-------
Penland, T.N., W.G. Cope, T.J. Kwak, M.J. Strynar, C.A. Grieshaber, R.J. Heise and F.W.
Sessions. 2020. Trophodynamics of per- and polyfluoroalkyl substances in the food web of a
large Atlantic Slope river. Environ. Sci. Technol. 54(11): 6800-6811.
Petre, V.A., F.L. Chiriac, I.E. Lucaciu, I. Paun, F. Pirvu, V.I. Iancu, L. Novae, and S. Gheorghe.
2023. Tissue bioconcentration pattern and biotransformation of per-fluorooctanoic acid (PFOA)
in Cyprinus carpio (European carp) - An extensive in vivo study. Foods 12(7): 1423-1444.
Phelps, D.W., A.I. Palekar, H.E. Conley, G. Ferrero, J.H. Driggers, K.E. Linder, S.W. Kullman,
D.M. Reif, M.K. Sheats, J. 2023. Legacy and emerging per- and polyfluoroalkyl substances
suppress the neutrophil respiratory burst. J. Immunotoxicol. 20(1): 16 p.
Plumlee, M.H., J. Larabee and M. Reinhard. 2008. Perfluorochemicals in water reuse.
Chemosphere. 72(10): 1541-1547.
Popovic, M., R. Zaja, K. Fent and T. Smital. 2014. Interaction of environmental contaminants
with zebrafish organic anion transporting polypeptide, Oatpldl (Slcoldl). Toxicol. Appl.
Pharmacol. 280(1): 149-158.
Prevedouros, K., I.T. Cousins, R.C. Buck and S.H. Korzeniowski. 2006. Sources, fate and
transport of perfluorocarboxylates. Environ. Sci. Technol. 40(1): 32-44.
Prosser, R.S., K. Mahon, P.K. Sibley, D. Poirier and T. Watson-Leung. 2016. Bioaccumulation
of perfluorinated carboxylates and sulfonates and polychlorinated biphenyls in laboratory-
cultured Hexagenia spp., Lumbriculus variegatus and Pimephalespromelas from field-collected
sediments. Sci. Tot. Environ. 543: 715-726.
Raby, M., M. Nowierski, D. Perlov, X. Zhao, C. Hao, D. G. Poirier and P. K. Sibley. 2018.
Acute toxicity of 6 neonicotinoid insecticides to freshwater invertebrates. Environ. Toxicol.
Chem. 37: 1430-1445.
Raimondo, S. and M.G. Barron. 2020. Application of Interspecies Correlation Estimation (ICE)
models in estimating species sensitivity to pesticides. SAR QSAR Environ. Res. 31: 1-18.
Raimondo, S., C.R. Jackson, and M.G. Barron. 2010. Influence of taxonomic relatedness and
chemical mode of action in acute interspecies estimation models for aquatic species. Environ.
Sci. Technol. 44: 7711-7716.
Raimondo, S., C.R. Jackson, and M.G. Barron. 2015. Web-based Interspecies Correlation
Estimation (Web-ICE) for Acute Toxicity: User Manual. Version 3.3, EPA/600/R-15/192, U. S.
Environmental Protection Agency, Office of Research and Development, Gulf Ecology Division.
Gulf Breeze, FL.
Raimondo, S., C. Lilavois, and S. A. Nelson. 2024. Uncertainty analysis and updated user
guidance for interspecies correlation estimation models and low toxicity compounds. Integr
Environ Assess Manag. https://doi.org/10.1002/ieam.4884.
163
-------
Raine, J.C., S. Su, E. Lin, Z.L. Yang, J.P. Giesy and P.D. Jones. 2021. Prefertilization exposure
of rainbow trout eggs to per- and polyfluoroalkyl substances to simulate accumulation during
oogenesis. Environ. Toxicol. Chem. 40(11): 3159-3165.
Rainieri, S., N. Conlledo, T. Langerholc, E. Madorran, M. Sala and A. Barranco. 2017. Toxic
effects of perfluorinated compounds at human cellular level and on a model vertebrate. Food
Chem. Toxicol. 104: 14-25.
Razak, M.R., A.Z. Aris, A.H. Zainuddin, F.M. Yusoff, Z.N.B. Yusof, S.D. Kim, and K.W. Kim.
2023. Acute toxicity and risk assessment of perfluorooctanoic acid (PFOA) and
perfluorooctanesulfonate (PFOS) in tropical cladocerans Moina micrura. Chemosphere 313: 9 p.
Remucal, C. K. 2019. Spatial and temporal variability of perfluoroalkyl substances in the
Laurentian Great Lakes. Environ. Sci. Process. Impacts. 21(11): 1816-1834.
Renner, R. 2009. EPA finds record PFOS, PFOA levels in Alabama grazing fields. Environ. Sci.
Technol. 43(3): 1245-1246.
Rericha, Y., D. Cao, L. Truong, M. Simonich, J.A. Field and R.L. Tanguay. 2021. Behavior
effects of structurally diverse per- and polyfluoroalkyl substances in zebrafish. Chem. Res.
Toxicol. 34(6): 1409-1416.
Rewerts, J.N., EC. Christie, A.E. Rob el, T.A. Anderson, C. McCarthy, C.J. Sal ice and J.A.
Field. 2021. Key considerations for accurate exposures in ecotoxicological assessments of
perfluorinated carboxylates and sulfonates. Environ. Toxicol. Chem. 40: 677-688.
Rodea-Palomares, I., F. Leganesa, R. Rosal andF. Fernandez-Pinas. 2012. Toxicological
interactions of perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA) with
selected pollutants. J. Hazard. Mater. 201-202: 209-218.
Rodea-Palomares, I., M. Makowski, S. Gonzalo, M. Gonzalez-Pleiter, F. Leganes and F.
Fernandez-Pinas. 2015. Effect of PFOA/PFOS pre-exposure on the toxicity of the herbicides 2,4-
D, atrazine, diuron and paraquat to a model aquatic photosynthetic microorganism.
Chemosphere. 139: 65-72.
Rosal, R., I. Rodea-Palomares, K. Boltes, F. Fernandez-Pinas, F. Leganes and A. Petre. 2010.
Ecotoxicological assessment of surfactants in the aquatic environment: combined toxicity of
docusate sodium with chlorinated pollutants. Chemosphere. 81: 288-293.
Rotondo, J.C., L. Giari, C. Guerranti, M. Tognon, G. Castaldelli, E.A. Fano and F. Martini. 2018.
Environmental doses of perfluorooctanoic acid change the expression of genes in target tissues
of common carp. Environ. Toxicol. Chem. 37(3): 942-948.
Royer, L.A. 2011. An investigation of the biodegradation potential of 8:2 fluorotelomer esters in
environmentally relevant systems. PhD Thesis, Purdue University, Purdue, IN.
164
-------
Russell, M.H., W.R. Berti, B. Szostek and R.C. Buck. 2008. Investigation of the biodegradation
potential of a fluoroacrylate polymer product in aerobic soils. Environ. Sci. Technol. 42: 800-
807.
Russell, M.H., W.R. Berti, B. Szostek and R.C. Buck. 2010. Evaluation of PFO formation from
the biodegradation of a fluorotelomer-based urethane polymer product in aerobic soils. Polym.
Degrad. Stab. 95: 79-85.
Saez M., D.V. Moreno, J. Begona. And S. Van Leeuwen. 2008. Uncommon PFC-profile in arctic
ice samples from Russia. Organohalogen Compd. 70: 1870-1873.
Saito, N., K. Sasaki, K. Nakatome, K. Harada, T. Yoshinaga and A. Koizumi. 2003.
Perfluorooctane sulfonate concentrations in surface water in Japan. Arch. Environ. Contam.
Toxicol. 45(2): 149-158.
San Francisco Bay Regional Water Quality Control Board (RWQCB). Transmittal
Memorandum: Transmittal of Interim Final Environmental Screening Levels (ESLs) for Two
Per- and Polyfluoroalkyl Substances (PFAS): Perfluorooctane Sulfonate (PFOS) and
Perfluorooctanoate (PFOA); May 27, 2020; Alec Naugle, Chief, Toxics Cleanup Division.
Sanderson, H., T.M. Boudreau, S.A. Mabury and K.R. Solomon. 2003. Impact of
perfluorooctanoic acid on the structure of the zooplankton community in indoor microcosms.
Aquat. Toxicol. 62: 227-234.
Satbhai, K., C. Vogs, and J. Crago. 2022. Comparative toxicokinetics and toxicity of PFOA and
its replacement GenX in the early stages of zebrafish. Chemosphere 308: 9 p.
Savoca, D., A. Pace, V. Arizza, M. Arculeo, and R. Melfi. 2022. Controlled uptake of PFOA in
adult specimens of Paracentrotus lividus and evaluation of gene expression in their gonads and
embryos. Environ. Sci. Pollut. Res.: 26094-26106.
Scott, B.F., C. Spencer, E. Lopez and D. C. Muir. 2009. Perfluorinated alkyl acid concentrations
in Canadian rivers and creeks. Water Qual. Res. Jour.. 44(3): 263-277.
Scott, B.F., A.O. De Silva, C. Spencer, E. Lopez, S.M. Backus and D.C.G. Muir. 2010.
Perfluoroalkyl acids in Lake Superior water: Trends and sources. J. Great Lakes Res. 36(2): 277-
284.
Sedlak, M.D., J.P. Benskin, A. Wong, R. Grace and D. J. Greig. 2017. Per- and polyfluoroalkyl
substances (PFASs) in San Francisco Bay wildlife: Temporal trends, exposure pathways, and
notable presence of precursor compounds. Chemosphere. 185: 1217-1226.
Seyoum, A., A. Pradhan, J. Jass and P.E. Olsson. 2020. Perfluorinated alkyl substances impede
growth, reproduction, lipid metabolism and lifespan in Daphnia magna. Sci. Tot. Environ. 737:
12 pp.
165
-------
Shoeib, M., T. Harner and P. Vlahos. 2006. Perfluorinated chemicals in the Arctic atmosphere.
Environ. Sci. Technol. 40: 7577-7583.
Simcik, M.F. and K.J. Dorweiler. 2005. Ratio of perfluorochemical concentrations as a tracer of
atmospheric deposition to surface waters. Environ. Sci. Technol. 39(22): 8678-8683.
Sinclair, E. and K. Kannan. 2006. Mass loading and fate of perfluoroalkyl surfactants in
wastewater treatment plants. Environ. Sci. Technol. 40(5): 1408-1414.
Sinclair, E., D.T. Mayack, K. Roblee, N. Yamashita, K. Kannan. 2006. Occurrence of
perfluoralkyl surfactants in water, fish, and birds from New York State. Arch. Environ. Contam.
Toxicol. 50: 398-410.
Solan, M.E., M.E. Franco, and R. Lavado. 2022. Effects of perfluoroalkyl substances (PFASs)
and benzo[a]pyrene (BaP) co-exposure on phase I biotransformation in rainbow trout
(Oncorhynchus mykiss). Fish Physiol. Biochem. 48(4): 925-935.
Soucek, D.J. and A. Dickinson. 2015. Full-life chronic toxicity of sodium salts to the mayfly
Neocloeon triangulifer in tests with laboratory cultured food. Environ. Toxicol. Chem. 34(9):
2126-2137.
Soucek, D. J. A. Dickinson, C. Schlekat, E. Van Genderen E. J. Hammer. 2020. Acute and
chronic toxicity of nickel and zinc to a laboratory cultured mayfly, Neocloeon triangulifer, in
aqueous but fed exposures. Environ. Toxicol. Chem. 39: 1196-1206.
Soucek, D. J., R. A. Dorman, E. L. Pulster, B. G. Perrotta, D. M. Walters and J. A. Steevens.
2023. Perfluorooctanesulfonate adversely affects a mayfly (Neocloeon triangulifer) at
environmentally realistic concentrations. Environmental Science & Technology Letters. DOI:
10.1021/acs.estlett.3c00056 .
Spachmo, B. and A. Arukwe. 2012. Endocrine and developmental effects in Atlantic salmon
(Salmo salaf) exposed to perfluorooctane sulfonic or perfluorooctane carboxylic acids. Aquat.
Toxicol. 108: 112-124.
Spence, R., W.C. Jordan and C. Smith. 2006. Genetic analysis of male reproductive success in
relation to density in the zebrafish, Danio rerio. Front. Zool. 3: 5. https://doi.org/10.1186/1742-
9994-3-5
SRC (Syracuse Research Corporation). 2016. PHYSPROP Database. Accessed May 2016.
http://www.srcinc.com/what-we-do/environmental/scientific-databases.html.
Stahl, L.L., B.D. Snyder, A.R. Olsen, T.M. Kincaid, J.B. Wathen and H.B. McCarty. 2014.
Perfluorinated compounds in fish from U.S. urban rivers and the Great Lakes. Sci. Tot. Environ.
499: 185-195.
166
-------
Stefani, F., M. Rusconi, S. Valsecchi and L. Marziali. 2014. Evolutionary ecotoxicology of
perfluoralkyl substances (PFASs)inferred from multigenerational exposure: A case study with
Chironomus riparius (Diptera, Chironomidae). Aquat. Toxicol. 156: 41-51.
Stengel, D., S. Wahby and T. Braunbeck. 2017. In search of a comprehensible set of endpoints
for the routine monitoring of neurotoxicity in vertebrates: Sensory perception and nerve
transmission in zebrafish (Danio rerio) embryos. Environ. Sci. Pollut. Res. Int. 12: 19 pp.
Stengel, D., S. Wahby and T. Braunbeck. 2018. In search of a comprehensible set of endpoints
for the routine monitoring of neurotoxicity in vertebrates: sensory perception and nerve
transmission in zebrafish {Danio rerio) embryos. Environ. Sci. Pollut. Res. Int. 25(5): 4066-
4084.
Stevenson, C.N., L.A. MacManus-Spencer, T. Luckenbach, R.G. Luthy and D. Epel. 2006. New
perspectives on pefluorochemical ecotoxicology: inhibition and induction of an efflux transporter
in marine mussel, Mytilus californianus. Environ. Sci. Technol. 40: 5580-5585.
Stinckens, E., L. Vergauwen, G.T. Ankley, R. Blust, V.M. Darras, D.L. Villeneuve, H. Witters,
D.C. Volz and D. Knapen. 2018. An AOP-based alternative testing strategy to predict the impact
of thyroid hormone disruption on swim bladder inflation in zebrafish. Aquat. Toxicol. 200:1-12.
Stock, N.L., V.I. Furdui, D.C.G. Muir and S.A. Mabury. 2007. Perfluoroalkyl contaminants in
the Canadian Arctic: Evidence of atmospheric transport and local contamination. Environ. Sci.
Technol. 41: 3529-3536.
STS Consultants, Ltd. 2007. Surface water quality criterion for perfluorooctanoic acid. Prepared
for Minnesota Pollution Control Agency. St. Paul, Minnesota.
https://www.pca.state.mn.us/sites/default/files/pfoa-report.pdf
Stuchal, L. and S. Roberts. 2019. PFAS- Provisional Cleanup Target Levels and Screening
Levels. University of Florida, Center for Environmental and Human Toxicology, Contaminated
Media Forum. September, 2019.
Sun, W.Q., X.M. Zhang, Y. Qiao, N. Griffin, H.X. Zhang, L. Wang, and H. Liu. 2023a.
Exposure to PFOA and its novel analogs disrupts lipid metabolism in zebrafish. Ecotoxicol.
Environ. Saf. 259: 13 p.
Sun, X., Y. Xie, X. Zhang, J. Song, and Y. Wu. 2023b. Estimation of per- and polyfluorinated
alkyl substance induction equivalency factors for humpback dolphins by transactivation
potencies of peroxisome proliferator-activated receptors. Environ. Sci. Technol. 57(9): 3713-
3721.
Tang, J., X. Jia, N. Gao, Y. Wu, Z. Liu, X. Lu, Q. Du, J. He, N. Li, B. Chen, J. Jiang, W. Liu, Y.
Ding, W. Zhu and H. Zhang. 2018. Role of the Nrf2-ARE pathway in perfluorooctanoic acid
(PFOA)-induced hepatotoxicity in Rana nigromaculata. Environ. Pollut. 238: 1035-1043.
167
-------
Tang, L., W. Qiu, S. Zhang, J. Wang, X. Yang, B. Xu, J.T. Magnuson, E.G. Xu, M. Wu, and C.
Zheng. 2023. Poly- and perfluoroalkyl substances induce immunotoxicity via the TLR pathway
in zebrafish: Links to carbon chain length. Environ. Sci. Technol. 57(15): 6139-6149.
TCEQ (Texas Commission on Environmental Quality). 2021. 2021 Ecological Screening
Benchmarks, available at: https://www.tceq.texas.gov/remediation/eco (2021 Benchmarks),
accessed 01.13.22.
Thienpont, B., A. Tingaud-Sequeira, E. Prats, C. Barata, P.J. Babin and D. Raldua. 2011.
Zebrafish eleutheroembryos provide a suitable vertebrate model for screening chemicals that
impair thyroid hormone synthesis. Environ. Sci. Technol. 45(17): 7525-7532.
Thomas, D.G., H. Shankaran, L. Truong, R.L. Tanguay andK.M. Waters. 2019. Time-dependent
behavioral data from zebrafish reveals novel signatures of chemical toxicity using point of
departure analysis. Comput. Toxicol. 9: 50-60.
Thompson, J., A. Roach, G. Eaglesham, M. E. Bartkow, K. Edge and J. F. Mueller. 2011.
Perfluorinated alkyl acids in water, sediment and wildlife from Sydney Harbour and
surroundings. Mar. Poll. Bull. 62: 2869-2875.
Tilton, S.C, G.A. Orner, A.D. Benninghoff, H.M. Carpenter, J.D. Hendricks, C.B. Pereira and
D.E. Williams. 2008. Genomic profiling reveals an alternate mechanism for hepatic tumor
promotion by perfluorooctanoic acid in rainbow trout. Environ. Health Perspect. 116(8): 1047-
1055.
Tomy, G.T., S.A. Tittlemier, V.P. Palace, W.R. Budakowski, E. Braekevelt, L. Brinkworth and
K. Friesen. 2004. Biotransformation of A-ethyl perfluorooctanesulfonamide by rainbow trout
(Oncorhynchus mykiss) liver microsomes. Environ. Sci. Technol. 38: 758-762.
Tomy, G. T., K. Pleskach, S. H. Ferguson, J. Hare, G. Stern, G. Macinnis, C. H. Marvin and L.
Losefo. 2009. Trophodynamics of some PFCs and BFRs in a western Canadian Arctic marine
food web. Environ. Sci. Technol. 43: 4076-4081.
Tornabene, B.J., M.F. Chislock, M.E. Gannon, M.S. Sepulveda and J.T. Hoverman. 2021.
Relative acute toxicity of three per- and polyfluoroalkyl substances on nine species of larval
amphibians. Integr. Environ. Assess. Manag. 17(4): 684-689.
Truong, L., D.M. Reif, L. St Mary, M.C. Geier, H.D. Truong and R.L. Tanguay. 2014.
Multidimensional in vivo hazard assessment using zebrafish. Toxicol. Sci. 137(1): 212-233.
Truong, L., Y. Rericha, P. Thunga, S. Marvel, D. Wallis, M.T. Simonich, J.A. Field, D. Cao,
D.M. Reif, and R.L. Tanguay. 2022. Systematic developmental toxicity assessment of a
structurally diverse library of PFAS in zebrafish. J. Hazard. Mater. 431: 10 p.
Ulhaq, M., S. Orn, G. Carlsson, J. Tallkvist and L. Norrgren. 2012. Perfluorooctanoic acid
toxicity in zebrafish (Danio rerio). Abstracts Toxicol. Letters. 21 IS: S43-S216.
168
-------
Ulhaq, M., G. Carlsson, S. Orn and L. Norrgren. 2013. Comparison of developmental toxicity of
seven perfluroalkyl acids to zebrafish embryos. Environ. Toxicol. Pharmacol. 36: 423-426.
UNEP (United Nations Environmental Program). 2015. Proposal to list pentadecafluorooctanoic
acid (CAS No: 335-67-1, PFOA, perfluorooctanoic acid), its salts and PFOA-related compounds
in Annexes A, B and/or C to the Stockholm Convention on Persistent Organic Pollutants.
U.S. EPA (United States Environmental Protection Agency). 1978. The Selenastrum
capricornutum Printz algal assay bottle test: Experimental design, application, and data
interpretation protocol. U.S. Environmental Protection Agency, Washington, D.C., EPA/600/9-
78/018 (NTIS PB286950).
U.S. EPA (United States Environmental Protection Agency). 1982. Standard evaluation
procedure, Daphnia magna Life-cycle (21-day renewal) chronic toxicity test EPA 540/9-86-141.
Environmental Protection Agency, Office of Pesticide Programs. Washington D.C.
U.S. EPA (United States Environmental Protection Agency). 1985. Guidelines for deriving
numerical national water quality criteria for the protection of aquatic organisms and their uses.
National Technical Information Service No. PB85-227049.
U.S. EPA (United States Environmental Protection Agency). 1991. Technical support document
for water quality-based toxics control. U.S. EPA, Office of Water, Washington, DC. EPA/505/2-
90-001, PB91-127415.
U.S. EPA (United States Environmental Protection Agency). 1995a. Final water quality
guidance for the Great Lakes system. 60 Federal Register 15366-15425 (March 23, 1995)
(40 CFR Parts 9, 122, 123, 131, and 132).
U.S. EPA (United States Environmental Protection Agency). 1995b. Short-term methods
for estimating the chronic toxicity of effluents and receiving waters to West Coast marine
and estuarine organisms. Environmental Monitoring and Support Laboratory, Cincinnati,
OH. EPA/600/R-95/136.
U.S. EPA (United States Environmental Protection Agency). 1996. Ecological effects test
guidelines. OPPTS 850.4200: seed germination/root elongation toxicity test. EPA 712-C-96-154.
Washington, DC
U.S. EPA (United States Environmental Protection Agency). 1998. Guidelines for ecological
risk assessment. EPA/630/R-95/002F. Risk Assessment Forum. Office of Research and
Development, Washington, D.C.
U.S. EPA (U.S. Environmental Protection Agency). 1999. 1999 Update of ambient water quality
criteria for ammonia. EPA 822-R-99-014. Office of Water, Washington, DC.
169
-------
U.S. EPA (U.S. Environmental Protection Agency). 2000. Methods for measuring the toxicity
and bioaccumulation of sediment-associated contaminants with freshwater invertebrates, 2nd ed.
EPA/600/R-99/064, most recent update to the EPA 600/R-94/024. Office of Research and
Development, Mid-Continent Ecology Division, Duluth, MN; Office of Science and Technology,
Office of Water, Washington, DC.
U.S. EPA (United States Environmental Protection Agency). 2002. Methods for measuring the
acute toxicity of effluents and receiving waters to freshwater and marine organisms. Fifth
Edition. C. Weber (Ed). U.S. Environmental Protection Agency, Washington, D.C., EPA/600/4-
90/027F (NTIS PB94114733). 275 pp.
U.S. EPA (United States Environmental Protection Agency). 2006. PFOA Stewardship
Program (EPA-HQ-OPPT-2006-0621). Office of Prevention, Pesticides, and Toxic Substances.
Washington, D.C.
U.S. EPA (United States Environmental Protection Agency). 201 1. A field-based aquatic life
benchmark for conductivity in central Appalachian streams - final report (EPA/600/R-10/023F).
Washington, DC.
U.S. EPA. (U.S. Environmental Protection Agency). 2012. Ecological effects test guidelines.
OCSPP 850.4500: Algal Toxicity. Office of Chemical Safety and Pollution Prevention.
Washington, DC. EPA 712-C-006-18-002.
U.S. EPA. (U.S. Environmental Protection Agency). 2013. Aquatic life ambient water quality
criteria for ammonia - freshwater. Office of Water, Office of Science and Technology.
Washington, DC. EPA 822-R-18-002.
U.S. EPA (United States Environmental Protection Agency). 2016a. Recommended
aquatic life ambient water quality criterion for selenium in freshwater. 81 Federal
Register 45285-45287 (July 13, 2016).
U.S. EPA (U.S. Environmental Protection Agency). 2016b. Series 850 - Ecological effects test
guidelines. Office of Chemical Safety and Pollution Prevention, Washington, DC. Accessed
March 2021. https://www.epa.gov/test-guidelines-pesticides-and-toxic-substances/series-850-
ecological-effects-test-guidelines.
U.S. EPA (United States Environmental Protection Agency). 2018. Final aquatic life ambient
water quality criteria for aluminum - 2018 (EPA-822-R-18-001). Office of Water. Washington,
DC.
U.S. EPA (U.S. Environmental Protection Agency). 2021. Assessing and Managing Chemicals
under TSCA: Fact Sheet PFOA Stewardship Program, https://www.epa.gov/assessing-and-
managing-chemi cal s-under-tsca/ fact-sheet-20102015-pfoa-stewardship-program.
170
-------
U.S. EPA (United States Environmental Protection Agency). 2023. Water quality standards
handbook. EPA-820-B-14-008. Office of Water, Washington, DC. Available online at:
https://www.epa.gov/wqs-tech/water-qualitv-standards-handbook.
U.S. FWS. (U.S. Fish and Wildlife Service). 2018. Zebrafish (Danio rerio) ecological risk
screening summary, https://www.fws.gov/story/ecological-risk-screening-summaries
Vedagiri, U.K., R.H. Anderson, H.M. Loso and C.M. Schwach. 2018. Ambient levels of PFOS
and PFOA in multiple environmental media. Rem. J. 28(2): 9-51.
Villeneuve, D.L., B.R. Blackwell, J.E. Cavallin, J. Collins, J.X. Hoang, R.N. Hofer, K.A. Houck,
K.M. Jensen, M.D. Kahl. 2023. Verification of in vivo estrogenic activity for four per- and
polyfluoroalkyl substances (PFAS) identified as estrogen receptor agonists via new approach
methodologies. Environ. Sci. Technol. 57(9): 3794-3803.
Vogs, C., G. Johanson, M. Naslund, S. Wulff, M. Sjodin, M. Hellstrandh, J. Lindberg and E.
Wincent. 2019. Toxicokinetics of perfluorinated alkyl acids influences their toxic potency in the
zebrafish embryo (Danio rerio). Environ. Sci. Technol. 53(7): 3898-3907.
Wang, N., B. Szostek, R.C. Buck, P.W. Folsom, L.M. Sulecki, V. Capka, W.R. Berti and J.T.
Gannon. 2005. Fluorotelomer alcohol biodegradation-direct evidence that perfluorinated carbon
chains breakdown. Environ. Sci. Technol. 39: 7516-7518.
Wang, N., B. Szostek, R.C. Buck, P.W. Folsom, L.M. Sulecki and J.T. Gannon. 2009. 8-2
fluorotelomer alcohol aerobic soil biodegradation: pathways, metabolites, and metabolite yields.
Chemosphere. 75: 1089-1096.
Wang, N., R.C. Buck, B. Szostek, L.M. Sulecki and B.W. Wolstenholme. 2012. 5:3
Polyfluorinated acid aerobic biotransformation in activated sludge via novel "one-carbon
removal pathways". Chemosphere. 87: 527-534.
Wang, W.J., C. DeWitt, C.P. Higgins and I.T. Cousins. 2017. A never-ending story of per- and
polyfluoroalkyl substances (PFASs)? Environ. Sci. Technol. 5(51): 2508-2518.
Wang, X., B. Fan, M. Fan, S. Belanger, J. Li, J. Chen, X. Gao and Z. Liu. 2020. Development
and use of interspecies correlation estimation models in China for potential application in water
quality criteria. Chemosphere. 240: 124848
Wang, Y.H., S.N. Jiang, B.B. Wang, X. Chen, and G.H. Lu. 2023. Comparison of developmental
toxicity induced by PFOA, HFPO-DA, and HFPO-TA in zebrafish embryos. Chemosphere 311:
8 p.
171
-------
Warne, M.St.J., G.E. Batley, R.A. van Dam, J.C. Chapman, D.R. Fox, C.W. Hickey and J.L.
Stauber. 2018. Revised method for deriving Australian and New Zealand water quality guideline
values for toxicants - update of 2015 version. Prepared for the revision of the Australian and
New Zealand Guidelines for Fresh and Marine Water Quality. Australian and New Zealand
Governments and Australian state and territory governments, Canberra. 48 pp.
Wasel, O., K.M. Thompson, Y. Gao, A.E. Godfrey, J. Gao, C.T. Mahapatra, L.S. Lee, M.S.
Sepulveda and J.L. Freeman. 2020. Comparison of zebrafish in vitro and in vivo developmental
toxicity assessments of perfluoroalkyl acids (PFAAs). J. Toxicol. Environ. Health Part A.
84:125-136.
Wasel, O., K.M. Thompson, and J.L. Freeman. 2022. Assessment of unique behavioral,
morphological, and molecular alterations in the comparative developmental toxicity profiles of
PFOA, PFHxA, and PFBA using the zebrafish model system. Environ. Int. 170: 14 p.
Washington, J. W., J.J. Ellington, T.M. Jenkins, J.J. Evans, H. Yoo and S.C. Hafner. 2009.
Degradability of an acrylate-linked, fluorotelomer polymer in soil. Environ. Sci. Technol.
43(17): 6617-6623
Wei, Y., J. Dai, M. Liu, J. Wang, M. Xu, J. Zha and Z. Wang. 2007. Estrogen-like properties of
perfluorooctanoic acid as revealed by expressing hepatic estrogen-responsive genes in rare
minnows {Gobiocypris rarus). Environ. Toxicol. Chem. 26(11): 2440-2447.
Wei, Y., L.L. Chan, D. Wang, H. Zhang, J. Wang and J. Dai. 2008a. Proteomic analysis of
hepatic protein profiles in rare minnow (Gobiocypris rarus) exposed to perfluorooctanoic acid. J.
Proteome Res. 7: 1729-1739.
Wei, Y., Y. Liu, J. Wang, Y. Tao and J. Dai. 2008b. Toxicogenomic analysis of the hepatic
effects of perfluorooctanoic acid on rare minnows (Gobiocypris rarus). Toxicol. Appl.
Pharmacol. 226: 285-297.
Wei, Y., X. Shi, H. Zhang, J. Wang, B. Zhou and J. Dai. 2009. Combined effects of
polyfluorinated and perfluorinated compounds on primary cultured hepatocytes from rare
minnow {Gobiocypris rarus) using toxicogenomic analysis. Aquat. Toxicol. 95: 27-36
(Supplemental Journal Materials).
Weiss-Errico, M.J., J.P. Berry and K.E. O'Shea. 2017. Beta-cyclodextrin attenuates
perfluorooctanoic acid toxicity in the zebrafish embryo model. Toxics. 5(4): 10.
Wen, W., X. Xia, X. Chen, H. Wang, B. Zhu, H. Li and Y. Li. 2016. Bioconcentration of
perfluoroalkyl substances by Chironomusplumosus larvae in water with different types of
dissolved organic matters. Environ. Pollut. 213: 299-307.
Wesner, J. S., J. M. Kraus, T. S. Schmidt, D. M. Walters and W. H. Clements. 2014.
Metamorphosis enhances the effects of metal exposure on the mayfly, Centroptilum triangulifer.
Environ. Sci. Technol. 48: 10415-10422.
172
-------
Williams, T.D., A. Diab, F. Ortega, V.S. Sabine, R.E. Godfrey, F. Falciani, J.K. Chipman and
S.G. George. 2008. Transcriptomic responses of European flounder (Platichthys flesus) to model
toxicants. Aquat. Toxicol. 90(2): 83-91.
Willming, M. M., C. R. Lilavois, M.G. Barron and S. Raimondo. 2016. Acute toxicity prediction
to threatened and endangered species using Interspecies Correlation Estimation (ICE) models.
Environ. Sci. Technol. 50: 10700-10707.
Wixon, J. 2000. Featured organism: Danio rerio, the zebrafish. Yeast 17: 225-231.
Wu, J., Z. Liu, Z. Yan and X. Yi. 2015. Derivation of water quality criteria of phenanthrene
using interspecies correlation estimation models for aquatic life in China. Environ. Sci. Pollut.
Res. 22: 9457-9463.
Wu, J., Z. Yan, X. Yi, Y. Lin, J. Ni, X. Goa., Z. Liu and X. Shi. 2016. Comparison of species
sensitivity distributions constructed with predicted acute toxicity data from interspecies
correlation estimation models and measured acute data for Benzo[a]pyrene. Chemosphere. 144:
2183-2188.
Wu, Y., Q. Xiong, B. Wang, Y. Liu, P. Zhou, L. Hu, F. Liu, and G. Ying. 2023. Screening of
structural and functional alterations in duckweed (Lemna minor) induced by per- and
polyfluoroalkyl substances (PFASs) with FTIR spectroscopy. Environ. Pollut. 317: 10 p.
Xia, X., X. Chen, X. Zhao, H. Chen and M. Shen. 2012. Effects of carbon nanotubes, chars, and
ash on bioaccumulation of perfluorochemicals by Chironomusplumosus larvae in sediment.
Environ. Sci. Technol. 46: 12467-12475.
Xia, X., A.H. Rabearisoa, X. Jiang and Z. Dai. 2013. Bioaccumulation of perfluoroalkyl
substances by Daphnia magna in water with different types and concentrations of protein.
Environ. Sci. Technol. 47: 10955-10963.
Xia, X., Z. Dai, A.H. Rabearisoa, P. Zhao and X. Jiang. 2015a. Comparing humic substance and
protein compound effects on the bioaccumulation of perfluoroalkyl substances by Daphnia
magna in water. Chemosphere. 119: 978-986.
Xia, X., A.H. Rabaerisoa, Z. Dai, X. Jiang, P. Zhao and H. Wang. 2015b. Inhibition effect of
Na+ and Ca2+ on the bioaccumulation of perfluoroalkyl substances by Daphnia magna in the
presence of protein. Environ. Toxicol. Chem. 34(2): 429-436.
Xia, X., R. Yu, M. Li, L. Liu, K. Zhang, Y. Wang, B. Li, L. Zhang, G. Song, X. Zheng and X.H.
Bai. 2018. Molecular cloning and characterization of two genes encoding peroxiredoxins from
freshwater bivalve Anodonta woodiana: Antioxidative effect and immune defense. Fish Shellfish
Immunol. 82: 476-491.
173
-------
Xiang, J., Y. Mi, B. Luo, S. Gong, Y. Zhou and T. Ma. 2021. Evaluating the potential of KOH-
modified composite biochar amendment to alleviate the ecotoxicity of perfluorooctanoic acid-
contaminated sediment on Bellamya aeruginosa. Ecotoxicol. Environ. Saf. 219: 9.
Xiao, F. 2017. Emerging poly- and perfluoroalkyl substances in the aquatic environment: A
review of current literature. Water Res. 124: 482-495.
Xiao, F., M.F. Simcik, T.R. Halbach and J.S. Gulliver. 2015. Perfluorooctane sulfonate (PFOS)
and perfluorooctanoate (PFOA) in soils and groundwater of a U.S. metropolitan area: Migration
and implications for human exposure. Water Res. 72: 64-74.
Xu, D., C. Li, H. Chen and B. Shao. 2013. Cellular response of freshwater green algae to
perfluorooctanoic acid toxicity. Ecotoxicol. Environ. Saf. 88: 103-107.
Xu, J., C.S. Guo, Y. Zhang and W. Meng. 2014. Bioaccumulation and trophic transfer of
perfluorinated compounds in a eutrophic freshwater food web. Environ. Pollut. 184: 254-261.
Yamashita, N., S. Taniyasu, G. Petrick, S. Wei, T. Gamo, P.K. Lam and K. Kannan. 2008.
Perfluorinated acids as novel chemical tracers of global circulation of ocean waters.
Chemosphere. 70(7): 1247-1255.
Yang, H. B., Z. Ya-Zhou, Y. Tang, G. Hui-Qin, F. Guo, S. Wei-Hua, L. Shu-Shen, H. Tan and F.
Chen. 2019. Antioxidant defence system is responsible for the toxicological interactions of
mixtures: A case study on PFOS and PFOA in Daphnia magna. Sci. Total Environ. 667: 435-443
(Supplemental Journal Materials).
Yang, J. 2010. Perfluorooctanoic acid induces peroxisomal fatty acid oxidation and cytokine
expression in the liver of male Japanese medaka (Oryzias latipes). Chemosphere. 81: 548-552.
Yang, S., F. Xu, F. Wu, S. Wang and B. Zheng. 2014. Development of PFOS and PFOA criteria
for the protection of freshwater aquatic life in China. Sci. Total Environ. 470-471: 677-683.
Yang, Z., L. Fu, M. Cao, F. Li, J. Li, Z. Chen, A. Guo, H. Zhong, W. Li, Y. Liang, and Q. Luo.
2023. PFAS-induced lipidomic dysregulations and their associations with developmental toxicity
in zebrafish embryos. Sci. Total Environ. 861: 9 p.
Ye, L., L.L. Wu, C.J. Zhang, L. Chen, Y. Wang, S.C. Li, P. Huang, Y.H. Yang, Y. An and X.Y.
Sun. 2007. Aquatic toxicity of perfluorooctane acid and perfluorooctyl sulfonates to zebrafish
embryos. In: 2007 International Symposium on Environ. Sci. Technol. Pp. 134-137.
Ye, X., M. J. Strynar, S. F. Nakayama, J. Varns, L. Helfant, J. Lazorchak and A. D. Lindstrom.
2008. Perfluorinated compounds in whole fish homogenates from the Ohio, Missouri, and Upper
Mississippi Rivers, USA. Environ. Pollut. 156: 1227-1232.
Young, C.J., V.I. Furdui, J. Franklin, R.M. Koerner, D.C.G. Muir and S.A. Mabury. 2007.
Perfluorinated acids in arctic snow: new evidence for atmospheric formation. Environ. Sci.
Technol. 41(10): 3455-3461.
174
-------
Yu, J., W. Cheng, M. Jia, L. Chen, C. Gu, H.Q. Ren and B. Wu. 2022. Toxicity of
perfluorooctanoic acid on zebrafish early embryonic development determined by single-cell
RNA sequencing. J. Hazard. Mater. 427: 1-9.
Yu, T., G. Zhou, Z. Cai, W. Liang, Y. Du and W. Wang. 2021. Behavioral effects of early-life
exposure to perfluorooctanoic acid might synthetically link to multiple aspects of dopaminergic
neuron development and dopamine functions in zebrafish larvae. Aquat. Toxicol. 238: 1-7.
Yuan, Z., J. Zhang, Y. Zhang, H. Zhen and Y. Sun. 2015. The effect of perfluorooctanoic acid on
the planarian Dugesia japonica. Pol. J. Environ. Stud. 24(2): 801-807.
Yuan, Z., J. Zhang, C. Tu, Z. Wang and W. Xin. 2016a. The protective effect of blueberry
anthocyanins against perfluorooctanoic acid-induced disturbance in planarian (Dugesia
japonica). Ecotoxicol. Environ. Saf. 127: 170-174.
Yuan, Z., J. Zhang, B. Zhao, Z. Miao and X. Wu. 2016b. Effects of perfluorooctanoic acid on
neural genes expression and neuronal morphology in the planarian Dugesia japonica. Chem.
Ecol. 32(6): 575-582.
Yuan, Z., Z. Miao, X. Gong, B. Zhao, Y. Zhang, H. Ma, J. Zhang and B. Zhao. 2017. Changes
on lipid peroxidation, enzymatic activities and gene expression in planarian (Dugesia japonica)
following exposure to perfluorooctanoic acid. Ecotoxicol. Environ. Saf. 145: 564-568.
Zareitalabad, P., J. Siemens, M. Hamer and W. Amelung. 2013. Perfluorooctanoic acid (PFOA)
and perfluorooctanesulfonic acid (PFOS) in surface waters, sediments, soils and wastewater - A
review on concentrations and distribution coefficients. Chemosphere. 91(6): 725-732.
Zhai, Y., X. Xia, X. Zhao, H. Dong, B. Zhu, N. Xia and J. Dong. 2016. Role of ingestion route in
the perfluoroalkyl substance bioaccumulation by Chironomusplumosus larvae in sediments
amended with carbonaceous materials. J. Hazard. Mater. 302: 404-414.
Zhang, H., W. Fang, D. Wang, N. Gao, Y. Ding and C. Chen. 2014a. The role of interleukin
family in perfluorooctanoic acid (PFOA)-induced immunotoxicity. J. Hazard. Materials. 280:
552-560.
Zhang, H., L. Shen, W. Fang, X. Zhang and Y. Zhong. 2021. Perfluorooctanoic acid-induced
immunotoxicity via NF-Kappa B pathway in zebrafish (Danio rerio) kidney. Fish Shellfish
Immunol. 113: 9-19.
Zhang, J., X. Shao, B. Zhao, L. Zhai, N. Liu, F. Gong, X. Ma, X. Pan, B. Zhao, Z. Yuan and X.
Zhang. 2020. Neurotoxicity of perfluorooctanoic acid and post-exposure recovery due to
blueberry anthocyanins in the planarians Dugesia japonica. Environ. Pollut. 263:
175
-------
Zhang, J., N. Sun, J. Sun, B. Wang, X. Chen, J. Liu, B. Zhao, and Z. Yuan. 2022. The combined
effects of wood vinegar and perfluorooctanoic acid on enzymatic activities, DNA damage and
gene transcription in Dugesiajaponica. Chem. Ecol. 38(6): 527-543.
Zhang, L., J. Niu, Y. Li, Y. Wang and D. Sun. 2013a. Evaluating the sub-lethal toxicity of PFOS
and PFOA using rotifer Brachionus calyciflorus. Environ. Pollut. 180: 34-40.
Zhang, L., J. Niu, Y. Wang, J. Shi and Q. Huang. 2014b. Chronic effects of PFOA and PFOS on
sexual reproduction of freshwater rotifer Brachionus calyciflorus. Chemosphere. 114: 114-120.
Zhang, S., B. Szostek, P.K. McCausland, B.W. Wolstenholme, X. Lu, N. Wang and R.C. Buck.
2013b. 6:2 and 8:2 fluorotelomer alcohol anaerobic biotransformation in digester sludge from a
WWTP under methanogenic conditions. Environ. Sci. Technol. 47: 4227-4235.
Zhang, S. L. Wang, Z. Want, D. Fan, L. Shi and J. Lui. 2017. Derivation of freshwater water
quality criteria for dibutyltin dilaurate from measured data and data predicted using interspecies
correlation estimation models. Chemosphere. 171: 142-148.
Zhang, X., R. Lohmann, C. Dassuncao, X. C. Hu, A. K. Weber, C. D. Vecitis and E. M.
Sunderland. 2016. Source attribution of poly- and perfluoroalkyl substances (PFASs) in surface
waters from Rhode Island and the New York Metropolitan Area. Environ Sci Technol Lett. 3(9):
316-321.
Zhao, W.C., Q. Li, F. Xiao, G.L. Song, Y. Lu, H.B. Yang and C.X. Liao. 2016. The study of
acute toxicity of perfluorooctanoic acid on zebrafish. Architectural, Energy and Information
Engineering: Proceedings of the 2015 International Conference on Architectural, Energy and
Information Engineering (AEIE 2015), Xiamen, China. Pp.139-141.
Zhao, Z.L., X.Y. Zheng, Z.S. Han, S.S. Yang, H.J. Zhang, T. Lin, and C. Zhou. 2023. Response
mechanisms of Chlorella sorokiniana to microplastics and PFOA stress: Photosynthesis,
oxidative stress, extracellular polymeric substances and antioxidant system. Chemosphere 323:
10 p.
Zheng, X.M., H.L. Liu, W. Shi, S. Wei, J.P. Giesy and H.X. Yu. 2012. Effects of perfluorinated
compounds on development of zebrafish embryos. Environ. Sci. Pollut. Res. 19(7): 2498-2505.
Zhou, Y. and Z.S. Zhang. 1989. The method of aquatic toxicity test. Agricultural Press, Beijing.
Zhou, Z., Y. Shi and W. Li. 2012. Perfluorinated compounds in surface water and organisms
from Baiyangdian Lake in north China: Source profiles, bioaccumulation and potential risk. Bull.
Environ. Contam. Toxicol. 89: 519-524.
176
-------
Appendix A Acceptable Freshwater Acute PFOA Toxicity Studies
A.l Summary Table of Acceptable Quantitative Freshwater Acute PFOA Toxicity Studies
Species (lilVsliiiic'l
Method'
Tesl
Dui'iilioii
( hcmic;il /
PuriU
Pll
Temp.
(ฐC)
I'.ITeel
Author
Kcporlcri
I'.ITeel
Cone,
(niii/l.)
I-'. PA
Ciileuliiled
I'.ITeel
(ttne.
(niii/l.)
l-iiiiil I.ITccl
(one.
(niii/l.)'
Speeies
Menu
Aeule
Value
(111 ป/l.)
Reference
Planaria (0.9 cm),
Dugesia japonica
S,U
96 hours
PFOA
>98%
-
25
LC50
458
-
458
-
Li 2008
Planaria (0.9 cm),
Dugesia japonica
s,u
96 hours
PFOA
>98%
-
25
LC50
337h
321.8
321.8
-
Li 2009
Planaria (0.9 cm),
Dugesia japonica
s,u
96 hours
PFOA
>98%
-
25
LC50
337h
383.0
383.0
383.6
Li 2009
Fatmucket
(glochidia, <24 hours),
Lampsilis siliquoidea
S,M
24 hours
PFOA
96%
8.46
20
EC50
(viability)
164.4
-
164.4
-
Hazelton et al.
2012, Hazelton
2013
Fatmucket (juvenile, 4-6 weeks),
Lampsilis siliquoidea
R, M
96 hours
PFOA
96%
8.46
20
LC50
>500
-
>500f
164.4
Hazelton et al.
2012, Hazelton
2013
Black sandshell
(glochidia, <24 hours),
Ligumia recta
S,M
24 hours
PFOA
96%
8.46
20
EC50
(viability)
161.0
-
161.0
-
Hazelton et al.
2012, Hazelton
2013
Black sandshell
(juvenile, 4-6 weeks),
Ligumia recta
R, M
96 hours
PFOA
96%
8.46
20
LC50
>500
-
>500f
161.0
Hazelton et al.
2012, Hazelton
2013
Pewter Physa (mixed age),
Physella acuta
(formerly, Physa acuta)
S,U
96 hours
PFOA
>98%
-
25
LC50
672h
762.0
762.0
-
Li 2009
Pewter Physa (mixed age),
Physella acuta
S,U
96 hours
PFOA
>98%
-
25
LC50
672h
659.9
659.9
-
Li 2009
Pewter Physa (mixed age),
Physella acuta
S,U
96 hours
PFOA
>98%
-
25
LC50
672h
628.3
628.3
681.1
Li 2009
Rotifer (<2-hour old neonates),
Brachionus calyciflorus
s,ub
24 hours
PFOA
96%
-
20
LC50
150.0
-
150.0
150.0
Zhang et al. 2013a
Cladoceran (<24 hours),
Chydorus sphaericus
s,u
48 hours
PFOA
Unreported
-
20
EC50
(death/immobility)
91.10ฐ
93.17ฐ
93.17ฐ
93.17
Le and
Peijnenburg 2013
A-l
-------
Species (lilVsliiiic)
Method'
1 esl
Dui'iilioii
( licniic;il /
PuriU
Pll
Temp.
(ฐC)
r.iTcci
Author
Kcporlcri
I'.ITecl
Cone,
(niii/l.)
r.PA
Ciileuliiled
I'.ITecl
(ttne.
(111 ป/l.)
l-iiiiil r.lTccl
(one.
(inii/l.)'
Species
Mean
Acme
Value
(111 ป/l.)
Reference
Cladoceran (6-12 hours),
Daphnia carinata
S,U
48 hours
PFOA
95%
-
21
ECso
(death/immobility)
78.2
66.80
66.80
66.80
Logeshwaran et
al. 2021
Cladoceran (<24 hours),
Daphnia magna
s,u
48 hours
PFOA
>97%
-
21
ECso
(immobility)
223.6
-
223.6
-
Boudreau 2002
Cladoceran
(STRAUS-clone 5; 6-24 hours),
Daphnia magna
s,u
48 hours
APFO
99.7%
-
18-22
ECso
480d
-
480d
-
Colombo et al.
2008
Cladoceran (<24 hours),
Daphnia magna
s,u
48 hours
PFOA
Unreported
-
21
ECso
(immobility)
476.52
542.5
542.5
-
Ji et al. 2008
Cladoceran (<24 hours),
Daphnia magna
s,u
48 hours
PFOA
>98%
7.82-
7.91
25
LC50
18 lh
220.8
220.8
-
Li 2009
Cladoceran (<24 hours),
Daphnia magna
s,u
48 hours
PFOA
>98%
7.82-
7.91
25
LC50
18 lh
157.9
157.9
-
Li 2009
Cladoceran (<24 hours),
Daphnia magna
s,u
48 hours
PFOA
>98%
7.82-
7.91
25
LC50
18 lh
207.3
207.3
-
Li 2009
Cladoceran (<24 hours),
Daphnia magna
S,M
48 hours
PFOA
96%
-
20
EC50
(death/immobility)
211.6ฐ
216.1ฐ
216.1c
-
Ding et al. 2012a
Cladoceran (<24 hours),
Daphnia magna
S,M
48 hours
PFOA
99%
7
22
LC50
201.85
222.0
222.0
-
Yang et al. 2014
Cladoceran (<24 hours),
Daphnia magna
S,M
48 hours
PFOA
>96%
7.0-
7.82
20
EC50
(immobility)
239
215.6
215.6
-
Barmentlo et al.
2015
Cladoceran (<24 hours),
Daphnia magna
S,U
48 hours
PFOA
98%
-
20
EC50
(death/immobility)
110.7
114.6
114.6
-
Lu et al. 2016
Cladoceran (12-24 hours),
Daphnia magna
S,U
48 hours
PFOA
Unreported
6-8.5
20
LC50
120.9ฐ
117.2ฐ
117.2C
-
Yang et al. 2019
Cladoceran (<24 hours),
Daphnia magna
S,U
48 hours
APFO
>98%
-
20
EC50
(death/immobility)
156.9
-
156.9
213.9
Chen et al. 2022
Cladoceran (<24 hours),
Daphnia pulicaria
S,U
48 hours
PFOA
>97%
-
21
EC50
(immobility)
203.7
-
203.7
203.7
Boudreau 2002
Cladoceran (<24 hours),
Moina macrocopa
s,u
48 hours
PFOA
Unreported
-
25
EC50
(immobility)
199.51
166.3
166.3
166.3
Ji et al. 2008
Cladoceran (<48 hours),
Moina micrura
s,u
48 hours
PFOA
>98%
-
27
LC50
0.4747
-
0.4747
0.4747
Razak et al. 2023
A-2
-------
Species (lilVsliiiic)
Method'
1 esl
Dui'iilioii
( licinic;il /
PuriU
Pll
Temp.
(ฐC)
r.iTcci
Author
Kcpurlcri
I'.ITecl
Cone,
(niii/l.)
I'. PA
Ciileuliiled
r.lTeel
(ttne.
(mii/l.)
l-'iiiiil l-'.ITccl
(one.
(mii/l.)'
Speeies
Menu
Aeule
Value
(111 ป/l.)
Reference
Green neon shrimp,
Neocaridina denticulata
S,U
96 hours
PFOA
>98%
25
LC50
454h
499.7
499.7
-
Li 2009
Green neon shrimp,
Neocaridina denticulata
s,u
96 hours
PFOA
>98%
-
25
LC50
454h
428.1
428.1
-
Li 2009
Green neon shrimp,
Neocaridina denticulata
s,u
96 hours
PFOA
>98%
-
25
LC50
454h
375.5
375.5
431.5
Li 2009
Mayfly (<24 hr larva),
Neocloeon triangulifer
S,M
96 hours
PFOA
95%
-
23
LC50
13.45
13.05
13.05
13.05
Soucek et al. 2023
Rainbow trout (2.8 cm, 0.21 g),
Oncorhynchus mykiss
S,M
96 hours
APFO
99.4%
7.1-
7.2
11.8
LC50
4,001
-
4,001
4,001
DuPont Haskell
Laboratory 2000
Zebrafish (embryo),
Danio rerio
R,U
96 hours
PFOA
Unreported
7-8.5
26
LC50
499
-
499
-
Ye et al. 2007
Zebrafish (embryo),
Danio rerio
S,U
96 hours
PFOA
>97%
7.2-
7.5
26
LC50
>500
-
>500
-
Hagenaars et al.
2011
Zebrafish (3 mo),
Danio rerio
S,M
96 hours
PFOA
98%
-
23
LC50
118.82
122.5
122.5
-
Zhao et al. 2016
Zebrafish (embryo),
Danio rerio
S,U
96 hours
PFOA
Unreported
7.5
26-28
LC50
24.6
22.77
22.778
-
Corrales et al.
2017
Zebrafish (embryo, 4 hpf),
Danio rerio
R, U
96 hours
PFOA
Unreported
7-7.5
28
LC50
473h
548.00
548.0
-
Godfrey et al.
2017a
Zebrafish (embryo, 4 hpf),
Danio rerio
R, U
96 hours
PFOA
Unreported
7-7.5
28
LC50
473h
508.5
508.5
-
Godfrey et al.
2017a
Zebrafish (embryo, 4 hpf),
Danio rerio
R, U
96 hours
PFOA
Unreported
7-7.5
28
LC50
473h
547.0
547.0
-
Godfrey et al.
2017a
Zebrafish (embryo),
Danio rerio
R, U
96 hours
PFOA
Unreported
-
26
LC50
759
806.6
806.6
450.4
Stengel et al.
2017, 2018
Fathead minnow,
Pimephales promelas
S,U
96 hours
PFOA
95-98%
7.5-
7.7
19-20
LC50
843
852.7
852.7
-
3M Co. 2000a
Fathead minnow (larva),
Pimephales promelas
S,U
96 hours
PFOA
Unreported
7.5
25
LC50
413.2
-
413.2
593.6
Corrales et al.
2017
Bluegill (2.1 cm, 0.228 g),
Lepomis macrochirus
S,U
96 hours
APFO
99%
6.9-
7.4
21.4-
22.1
LC50
634
664.0
664.0
664.0
DuPont Haskell
Laboratory 2000
A-3
-------
Species (lilVsliiiic)
Method'
1 esl
Dui'iilioii
( licinic;il /
PuriU
Pll
Temp.
(ฐC)
r.iTcci
Author
Reported
I'.ITecl
Cone,
(niii/l.)
r.PA
Ciileuliiled
r.lTeel
(ttne.
(mii/l.)
l-'iiiiil l-'.ITccl
(one.
(mii/l.)'
Speeies
Menu
Aciilc
Value
(111 ป/l.)
Reference
American toad
(larva, Gosner stage 26),
Anaxyrus americanus
S,U
96 hours
PFOA
Unreported
21
LC50
711h
7814
781.4
-
Tornabene et al.
2021
American toad
(larva, Gosner stage 41),
Anaxyrus americanus
s,u
96 hours
PFOA
Unreported
-
21
LC50
711h
806.6
806.6
793.9
Tornabene et al.
2021
Gray treefrog
(larva, Gosner stage 26),
Hyla versicolor
s,u
96 hours
PFOA
Unreported
-
21
LC50
557
646.2
646.2
646.2
Tornabene et al.
2021
American bullfrog
(tadpole, Gosner stage 25),
Lithobates catesbeiana
(formerly, Rana catesbeiana)
s,u
96 hours
PFOA
Unreported
-
21
LC50
1,004
1,006
1,006
-
Flynnetal. 2019
American bullfrog
(larva, Gosner stage 26),
Lithobates catesbeiana
s,u
96 hours
PFOA
Unreported
-
21
LC50
1,060
1,035
1,035
1,020
Tornabene et al.
2021
Green frog
(larva, Gosner stage 26),
Lithobates clamitans
(formerly, Rana clamitans)
s,u
96 hours
PFOA
Unreported
-
21
LC50
1,070
-
1,070
1,070
Tornabene et al.
2021
Northern leopard frog
(larva, Gosner stage 26),
Lithobates pipiens
(formerly, Rana pipiens)
s,u
96 hours
PFOA
Unreported
-
21
LC50
752
751.7
751.7
751.7
Tornabene et al.
2021
Wood frog
(larva, Gosner stage 26),
Lithobates sylvatica
(formerly, Rana sylvatica)
s,u
96 hours
PFOA
Unreported
-
21
LC50
999
-
999
999
Tornabene et al.
2021
Frog (embryo stage 8.5),
Xenopus sp.
R,U
96 hours
PFOA
Unreported
-
23
LC50
377ฐ
-
377c
377
Kim et al. 2013
A-4
-------
Species (lilVsliiiic'l
Method'
1 esl
Dui'iilioii
( hcmic;il /
Piiriu
Pll
Temp.
(ฐC)
r.lTecl
Author
Reported
t'.ITecl
Cone,
(niii/l.)
r.PA
Ciileuliiled
t'.ITecl
(one.
(111 ป/l.)
l-'iiiiil I'.ITeel
(one.
dii"/!.)'
Speeies
Menu
Aenle
Value
(111 ป/l.)
Reference
Jefferson salamander
(larva, Harrison stage 40),
Ambystoma jeffersonianum
S,U
96 hours
PFOA
Unreported
21
LC50
1,070
-
1,070
1,070
Tornabene et al.
2021
Small-mouthed salamander
(larva, Harrison stage 40),
Ambystoma texanum
s,u
96 hours
PFOA
Unreported
-
21
LC50
474
407.3
407.3
-
Tornabene et al.
2021
Small-mouthed salamander
(larva, Harrison stage 45),
Ambystoma texanum
s,u
96 hours
PFOA
Unreported
-
21
LC50
1,000
-
l,000f
407.3
Tornabene et al.
2021
Eastern tiger salamander
(larva, Harrison stage 40),
Ambystoma tigrinum
s,u
96 hours
PFOA
Unreported
-
21
LC50
752
-
752
752
Tornabene et al.
2021
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b Chemical concentrations made in a side-test representative of exposure and verified stability of concentrations of PFOA in the range of concentrations tested under similar
conditions. Daily renewal of test solutions,
c Reported in moles converted to milligram based on a molecular weight of 414.07 mg/mmol.
d Concentration of APFO in publication indirectly determined through quantification of the anion (PFO-).
e Values in bold used in the SMAV calculation,
f Only the most sensitive life-stage used in the SMAV calculation,
g Value is considered an outlier and not used in SMAV calculation,
h Author pooled test of life stages.
A-5
-------
A.2 Detailed PFOA Acute Toxicity Study Summaries and Corresponding
Concentration-Response Curves (when calculated)
The purpose of this section is to present detailed study summaries for tests that were
considered quantitatively acceptable for criterion derivation, with summaries grouped and
ordered by genus sensitivity. C-R models developed by the EPA that were used to determine
acute toxicity values used for criterion derivation are also presented. C-R models included here
with study summaries were those for the four most sensitive genera. In many cases, authors did
not report concentration-response data in the publication/supplemental materials and/or did not
provide concentration-response data upon request by the EPA. In such cases, the EPA did not
independently calculate toxicity values and the author-reported effect concentrations were used
to derive the criterion.
A.2.1 Most acutely sensitive genus -Moina
Ji et al. (2008) performed a 48-hour static, unmeasured acute test of PFOA (CAS # 335-
67-1, purity unreported; obtained from Sigma Aldrich, St. Louis, MO) with Moina macrocopa.
Authors stated the test followed U.S. EPA/600/4-90/027F (2002). M. macrocopa used for testing
were obtained from brood stock cultured at the Environmental Toxicology Laboratory at Seoul
National University, Korea. Test organisms were less than 24-hours old at test initiation. Dilution
water was moderately hard reconstituted water (total hardness typically 80-100 mg/L as CaC03).
Experiments were conducted in glass jars of unspecified size and fill volume. Photoperiod was
assumed as 16-hour: 8-hour, light:dark, the same conditions as the daphnid cultures. Preparation
of test solutions was not described. The test involved four replicates of five daphnids each in five
unmeasured test concentrations plus a negative control. Nominal concentrations were 0 (negative
control), 62.5, 125, 250, 500 and 1,000 mg/L. Test temperature was maintained at 25 ฑ 1ฐC.
Authors noted water quality parameters (pH, temperature, conductivity, and dissolved oxygen)
A-6
-------
were measured 48-hours after exposure, but the information was not reported. Survival of
daphnids in the negative control was not reported, although EPA/600/4-90/027F requires at least
90% survival for test acceptability. The author-reported 48-hour EC50 for the study was 199.51
mg/L (95% C.I. = 163.9 - 245.1). The EPA performed C-R analysis for the test. The EPA-
calculated EC50 was 166.3 mg/L PFOA (95% C.I. = 138.6 - 194.1 mg/L) and was acceptable for
quantitative use.
Razak et al. (2023) tested the acute toxicity of perfluorooctanoic acid (PFOA) to Moina
micrura in a 48 hour static measured experiment. In addition to acute toxicity, the effects of
PFOA on heart rate, heart size, and individual size were also examined. PFOA (>98% purity)
analytical standards were purchased from Dr. Ehrenstorfer GmbH (Augsburg, Germany), and
PFOA and solvents for making test solutions were purchased from Fisher Scientific (New Jersey,
USA). Organisms were obtained from the Aquatic Animal Health and Therapeutics Laboratory
(Aquahealth) at the Institute of Bioscience, Universiti Putra Malaysia. Culturing procedures
followed International Organisation for Standardization (ISO) procedure 6341:2012. Cultures
were kept under a 12:12 light:dark cycle at 27ฑ1ฐC. Culture water was renewed every two
weeks, and culture organisms were fed green algae (Chlorella vulgaris) three times weekly. Both
culture and test water was filtered (0.2 |im) surface lake water. A stock solution of 100 mg/L
PFOA with filtered surface lake water was made just before testing began. Testing methods
followed OECD 202 (OECD 2004) with nominal testing concentrations of 10, 25, 50, 75, 100,
250, 500, 750, 1,000, 2,500, 5,000, 7,500, and 10,000 |ig/L, plus a control, with four replicates
per treatment. Measured concentrations were not reported, but the authors noted they were
94.3ฑ6.1% of nominal on average. Each replicate consisted of 10 neonates (<48 hours old) in 50
mL of solution in a 100 mL beaker, and organisms were not fed during the study. Nonparametric
A-7
-------
Kruskal-Wallace tests followed by post-hoc tests were used to calculate significant (P<0.05)
differences between controls and treatment concentrations for all endpoints. The lethal effect
concentrations (LCio, LC50, LC75, LC90) were calculated using Probit analysis, and the 48-hour
LC50 value of 474.7 |ig/L, or 0.4747 mg/L was determined to be acceptable for quantitative use.
C-R data could not be obtained for this test (beyond the visual presentation in the Razak et al.
(2023), so the EPA was unable to perform independent C-R analysis.
A.2.2 Second most acutely sensitive genus - Neocloeon
Soucek et al. (2023) exposed the parthenogenetic mayfly, Neocloeon triangulifer, to
PFOA (CAS # 335-67-1, 95% purity) in a 96-hour acute toxicity test. The test was performed
under static, non-renewal conditions beginning with <24-hour old larvae. Exposures consisted of
five mayfly nymphs per 30 mL polypropylene cup filled with 20 mL test solution. The control
and each of six PFOA test concentrations were replicated five times for a total of 25 test
organisms per treatment. Nominal test concentrations were 0.0 (control), 1.25, 2.5, 5.0, 10, and
20 mg/L PFOA. Mean measured PFOA concentrations (EPA Analytical Method 1633; LC-
MC/MS) were 0.0003 (control), 1.765, 3.896, 8.137, 13.39 and 30.40 mg/L PFOA, respectively.
Mayflies were exposed at 23 ฑ 1ฐC under a 16:8-hour light:dark cycle and fed live diatom
biofilm scraping beginning on day 0. Feeding only occurred on day 0, and the authors indicated
that test organisms required food to survive the entire 96-hour exposure, with previous studies
demonstrating greater than 80% mortality at 48 hours with no food (Soucek and Dickinson
2015). Percent survival in the control treatment after 96 hours was 100%. A steep concentration-
response relationship between percent survival and increasing PFOA test concentration was
observed. Survival decreased from 100% at 8.137 mg/L PFOA to 38.5% at 13.39 mg/L, and to
0% at 30.40 mg/L. The EPA-calculated acute LC50 (i.e., 13.045 mg/L) was similar to the author-
A-8
-------
reported LC50 of 13.451 mg/L. The EPA-calculated LC50 value (i.e., 13.045 mg/L) was
acceptable for quantitative use in deriving the acute freshwater PFOA criterion.
Recent research has led to the development of successful culturing methods for mayflies
that are used in laboratory-based toxicity studies (Soucek and Dickinson 2015; Soucek et al.
2023). These resulting toxicity studies have shown that mayflies (including Neocloeon
triangulifer) are commonly among the most sensitive taxa to different chemical exposures,
including PFOA, PFOS, major geochemical ions, pesticides, and heavy metals (Johnson et al.
2015; Kim et al. 2012; Raby et al. 2018; Soucek and Dickinson 2015; Soucek et al. 2020; Soucek
et al. 2023; Wesner et al. 2014). The high sensitivity of mayflies to contaminant exposures has
also been observed In mesocosm-based experiments (Mebane et al. 2020) and field-based
surveys (U.S. EPA 2011). Many of these laboratory-based toxicity tests used mayflies capable of
adapting to laboratory settings. It is hypothesized that mayfly species unable to survive in
laboratory settings may also be more sensitive to contaminant exposures than the relatively hardy
mayfly species (e.g., N. triangulifer) commonly used for toxicity testing.
Publication: Soucek et al. (2023)
Species: Mayfly (Neocloeon triangulifer)
Genus: Neocloeon
EPA-Calculated LCso: 13.045 mg/L (95% C.I. = 12.46 - 13.63 mg/L)
Concentration-Response Model Estimates:
Parameter
Kslimale
Sul. Krror
1-slal
p-value
b
10.8557
6.1850
1.7552
0.0792
e
13.4934
0.3271
41.248
<2.0e"16
A-9
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Concentration-Response Model Fit:
Soucek et al. 2023
Neocleon triangulifer
Weibull type 1, 2 para
0.01 I 10 100
PFOA ( mg/L)
A.2.3 Third most acutely sensitive genus - Chydorus
Le and Peijnenburg (2013) performed a 48-hour static unmeasured test on PFOA
(unreported purity) with the cladoceran, Chydorus sphaericus. Authors stated the test followed
the protocol of the "Chydotox toxicity test" developed by the National Institute for Public Health
and the Environment, The Netherlands. In-house cultures of neonates (<24 hours) were exposed
to 250 ML of test solutions in 2 ML vials of unreported material. Each vial contained five
neonates and each test concentration was replicated four times. No solvent was used in the test
solutions with 18-20 test concentrations. C. sphaericus was cultured at 20 ฑ 1ฐC and a cycle of
16-hour: 8-hour light:dark without the addition of food. At test termination vials were shaken
slightly and the mobility of the neonates was determined. The author-reported 48-hour EC so was
0.22 MM PFOA (91.10 mg/L). The EPA performed concentration-response (C-R) analysis for
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the test and calculated a LC50 of 93.17 mg/L PFOA (95% C.I. = 82.52 - 103.8 mg/L) that was
acceptable for quantitative use.
Publication: Le and Peijnenburg (2013)
Species: Cladoceran {Chydorus sphaericus)
Genus: Chydorus
EPA-Calculated LCso: 93.17 mg/L (95% C.I. = 82.52 - 103.8 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std.Error
t-stat
p-value
b
3.007
0.4165
7.2181
5.27e"13
d
0.9560
0.0248
38.5566
<2.2e"16
e
5.9057
5.9057
15.7767
<2.2e"16
Concentration-Response Model Fit:
Le and Peijnenberg 2013
Chydorus sphaericus
Log Logistic type 1, 3 para
PFOA ( mg'L )
A.2.4 Fourth most acutely sensitive genus -Dayhnia
Logeshwaran et al. (2021) conducted acute and chronic toxicity tests with the
cladoceran, Daphnia carinata, and PFOA (95% purity, purchased from Sigma-Aldrich
Australia). In-house cultures of daphnids were maintained in 2 L glass bottles with 30% natural
spring water in deionized water, 21ฐC and a 16-hour:8-hour light:dark photoperiod. The acute
test protocol followed OECD guidelines (2000) with slight modifications. A PFOA stock
A-ll
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solution (100 mg/L) was prepared in deionized water. Cladoceran culture medium was used to
prepare the PFOA stock and test solutions. Ten daphnids (six to 12 hours old) were transferred to
polypropylene containers containing one of 14 nominal test concentrations (0, 0.5, 1, 2.5, 5, 10,
20, 30, 40, 50, 100, 150, 200 and 250 mg/L PFOA). Each test treatment was replicated three
times and held under the same conditions as culturing. At test termination (48 hours) immobility
was determined after 15 seconds of gentle stirring. No mortality occurred in the controls. The
author-reported 48-hour EC50 was 78.2 mg/L PFOA. The EPA-calculated 48-hour EC50 value
was 66.80 mg/L (95% C.I. = 57.10 - 76.50 mg/L), which was acceptable for quantitative use.
Publication: Logeshwaran et al. (2021)
Species: Cladoceran {Daphnia carinata)
Genus: Daphnia
EPA-Calculated LCso: 66.80 mg/L (95% C.I. = 57.10 - 76.50 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.6249
0.1565
10.3860
<2.2e"16
e
83.6974
5.9263
14.1230
<2.2e"16
Concentration-Response Model Fit:
Logeshwaran et al. 2021
Daphnia carinata
Weibull type 1, 2 para
PFOA ( mg/L )
A-12
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Boudreau (2002) performed a 48-hour static unmeasured test on PFOA (CAS # 335-67-
1, >97% purity) with Daphnia magna and Daphniapulicaria as part of a Master's thesis at the
University of Guelph, Ontario, Canada. The results were subsequently published in the open
literature (Boudreau et al. 2003). Authors stated the test followed ASTM E729-96 (1999).
Daphnids used for testing were less than 24-hours old at test initiation. I). magna were obtained
from a brood stock (Dm99-23) at ESG International (Guelph, ON, Canada). D. pulicaria were
acquired from a brood stock maintained in the Department of Zoology at the University of
Guelph. Dilution water was clean well water obtained from ESG International. Total hardness
was softened by addition of distilled deionized water to achieve a range of 200-225 mg/L of
CaC03. Photoperiod was 16-hours of illumination under cool-white, fluorescent light between
380 and 480 lux. Laboratory-grade distilled water was used for all solutions with maximum
concentrations derived from stock solutions no greater than 450 mg/L. Test vessels consisted of
225 mL polypropylene disposable containers containing 150 mL of test solution. All toxicity
testing involved three to four replicates of 10 daphnids each in five unmeasured test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 26.3,
52.6, 105, 210 and 420 mg/L. Experiments were conducted in environmental chambers at a test
temperature of 21 ฑ 1ฐC. Authors note that temperature and pH were measured at the beginning
and end of the study, but this information is not reported. Mortality of daphnids in the negative
control was also not reported, although ASTM E729-96 requires at least 90% survival for test
acceptability. The 48-hour D. magna ECso reported in the publication was 223.6 mg/L. The 48-
hour D. pulicaria EC so reported in the publication was 203.7 mg/L.
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Publication: Boudreau (2002)
Species: Cladoceran {Daphnia pulicaria)
Genus: Daphnia
EPA-Calculated LCso: Not calculable, concentration-response data not available
Concentration-Response Model Fit: Not Applicable
Publication: Boudreau (2002)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: Not calculable, concentration-response data not available
Concentration-Response Model Fit: Not Applicable
Colombo et al. (2008) conducted a 48-hour static unmeasured acute test on ammonium
perfluorooctanoate (APFO, CAS # 3825-26-1, 99.7% purity) with the daphnid, Daphnia magna.
The authors stated that the toxicity test was conducted following OECD test guideline 202
(1992). Neonates, six to 24-hours old, were acclimated to test conditions for six-hours before test
initiation with test solutions made in reconstituted M4 media. There were four replicates for each
test treatment containing five animals each. Exposure vessel material and size were not reported.
Based on loading, exposure vessels contained at least 100 mL test solution. Nominal test
concentrations were used based on the known stability of the test substance in water. The
nominal test concentrations included control, 100, 178, 316, 562 and 1,000 mg/L. Dissolved
oxygen was >60% saturation and temperature were maintained between 18-22ฐC. Illumination
involved 16-hours of light with an unreported intensity. No mortality was observed in the
controls. C-R data were available for this acute test; however, the EPA was unable to fit a model
with significant parameters and relied on the 48-hour EC so reported in the study of 480 mg/L,
which was acceptable for quantitative use. The authors note the contribution of ammonia from
APFO exposure indicates that un-ionized ammonia could be a potential contributor to the
observed acute toxicity of APFO. The EPA believes ammonia does not contribute substantively
to the acute toxicity to D. magna in this test based on the following rationale. U.S. EPA (2013)
A-14
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derived a SMAV of 157.7 mg N/L at pH=7 and temperature of 20ฐC for I), magna. Using the
EPA equations, this translates to approximately 30.36 mg N/L at the authors' assumed acute test
pH=8.1 (from their Table 7) and their midrange test temperature of 20ฐC, which in turn translates
to an un-ionized ammonia concentration of approximately 1.76 mg un-ionized ammonia/L at this
pH and temperature. The authors' Table 7 lists the un-ionized ammonia concentration at their
APFO ECso as 0.73 mg un-ionized ammonia/L, which is 2.4-fold lower than the EPA's SMAV
for D. magna re-expressed as an un-ionized ammonia concentration for the test condition.
Therefore, it is unlikely that ammonia contributes substantively to the acute toxicity of APFO to
D. magna in this test, and therefore, the 48-hour EC50 is used in the calculation of the SMAV for
this species.
Publication: Colombo et al. (2008)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LC50: Not calculable, unable to fit a model with significant parameters
Concentration-Response Model Fit: Not Applicable
Ji et al. (2008) also performed a 48-hour static, unmeasured acute test of PFOA (CAS #
335-67-1, purity unreported; obtained from Sigma Aldrich, St. Louis, MO) with D. magna.
Authors stated that the test followed U.S. EPA/600/4-90/027F (2002). I). magna used for testing
were obtained from brood stock cultured at the Environmental Toxicology Laboratory at Seoul
National University, Korea. Test organisms were less than 24-hours old at test initiation. Dilution
water was moderately hard reconstituted water (total hardness typically 80-100 mg/L as CaC03).
Experiments were conducted in glass jars of unspecified size and fill volume. Photoperiod was
assumed as 16-hour: 8-hour, light:dark, the same conditions as the daphnid cultures. Preparation
of test solutions was not described. The test involved four replicates of five daphnids each in five
unmeasured test concentrations plus a negative control. Nominal concentrations were 0 (negative
A-15
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control), 62.5, 125, 250, 500 and 1,000 mg/L. Test temperature was maintained at 21 ฑ 1ฐC. The
authors noted that water quality parameters (Ph, temperature, conductivity, and dissolved
oxygen) were measured 48-hours after exposure, but the information was not reported. Mortality
of daphnids in the negative control was not reported, although EPA/600/4-90/027F requires at
least 90% survival for test acceptability. The author-reported 48-hour EC so for the study was
476.52 mg/L (95% C.I. = 375.3 - 577.7 mg/L). The EPA performed C-R analysis for the test.
The EPA-calculated ECso was 542.5 mg/L PFOA (95% C.I. = 461.1 - 623.8 mg/L), which was
acceptable for quantitative use.
A-16
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Publication: Ji et al. (2008)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 542.5 mg/L (95% C.I. = 461.1 - 623.8 mg/L),
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
3.7248
1.8230
2.0432
0.0410
d
0.8985
0.0393
22.8879
< 2.0 e16
e
598.5588
66.9972
8.9341
< 2.0 e16
Concentration-Response Model Fit:
Ji et al. 2008
Daphnia magna
WeibiiH type 1, 3 para
PFOA ( nig'! )
Li (2009) conducted a 48-hour static unmeasured acute test on PFOA (ammonium salt,
>98% purity) with Daphnia magna. The authors stated that the test followed OECD 202 (1984)
with slight modifications. D. magna used for the test were less than 24-hours old at test
initiation. Dilution water was dechlorinated tap water. The photoperiod consisted of 12-hours of
illumination at an unreported light intensity. A primary stock solution was prepared in dilution
water and did not exceed 400 mg/L. Exposure vessels were polypropylene of unreported
dimensions and 50 Ml fill volume. The test employed five replicates of six daphnids each in five
A-17
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test concentrations plus a negative control. Based on water solubility of test chemicals and
preliminary toxicity results, nominal test concentrations were in the range of 31-250 mg/L for
PFOA. The test was conducted in a temperature incubator at 25 ฑ 2ฐC. Water quality parameters
including water Ph, conductivity, and dissolved oxygen were measured at the beginning and at
the end of each test. Initial values of Ph were 7.82 ฑ 0.12 and 7.91 ฑ 0.03 after 48-hours. At the
start of the bioassays, dissolved oxygen and specific conductivity were 67.7 ฑ 6.8% saturation
and 101.8 ฑ 6.8 |iS/cm, respectively. After the 48-hour testing period, dissolved oxygen and
specific conductivity were 55.6 ฑ 1.26% saturation (implying 4.56 mg/L) and 109.1 ฑ 3.5 |iS/cm,
respectively. None of the control animals became immobile at the end of the test. The author-
reported 48-hour ECso for the study was 181 mg/L (95% C.I.: 166-198 mg/L) which was
averaged across three tests. The EPA performed C-R analysis for each individual test. All three
tests had acceptable curves with the EPA-calculated EC50 values of 220.8 mg/L (95% C.I. =
191.8 - 250.0 mg/L), 157.9 mg/L (95% C.I. = 135.9 - 180.0 mg/L), and 207.3 mg/L (95% C.I. =
176.1 - 238.5 mg/L), which were acceptable for quantitative use.
A-18
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Publication: Li (2009)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 220.8 mg/L (95% C.I. = 191.8 - 250.0 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std.Error
t-stat
p-value
b
3.2035
0.6834
4.6875
2.766 e6
e
247.6075
17.6535
14.0260
< 2.2 e16
Concentration-Response Model Fit:
Li 2009
Daphnia magna
Weibull type 1,2 para
PFOA ( mg'L )
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Publication: Li (2009)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 157.9 mg/L (95% C.I. = 135.9 - 180.0 mg/L)
Concentration-Response Model Estimates:
Parameter
Kslimale
Sid. Krror
1-slal
p-value
b
2.8137
0.4368
6.4423
1.177 e10
e
179.9061
12.3471
14.5707
< 2.2 e"16
Concentration-Response Model Fit:
Li 2009
Daphnia magna
WeibiiD type 1, 2 para
PFOA ( mg/L )
A-20
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Publication: Li (2009)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 207.3 mg/L (95% C.I. = 176.1 - 238.5 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
2.5732
0.4477
5.7479
9.036 e9
e
239.0336
19.0886
12.5223
< 2.2 e"16
Concentration-Response Model Fit:
Li 2009
Daphnia magna
Weibull type 1,2 para
PFOA ( mg/L )
Yang et al. (2014) conducted a 48-hour measured acute test of PFOA (CAS # 335-67-1,
99% purity) with Daphnia magna, following ASTM E729 (1993). Although the authors termed
the test conditions "static", they also mentioned PFOA measurements before and after renewal;
based on this distinction the test was assumed to be renewed at least once. Daphnids used for the
test were donated by the Chinese Research Academy of Environmental Sciences. The daphnids
were less than 24-hours old at test initiation. Dilution water was dechlorinated tap water (pH, 7.0
ฑ0.5; dissolved oxygen, 7.0 ฑ 0.5 mg/L; total organic carbon, 0.02 mg/L; and total hardness,
190.0 ฑ0.1 mg/L as CaC03). The photoperiod consisted of 12-hours of illumination at an
A-21
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unreported intensity. A primary stock solution was prepared by dissolving PFOA in deionized
water and solvent, DMSO, and proportionally diluted with dilution water to prepare the test
concentrations. Exposure vessels were 200 mL beakers of unreported material type containing
100 mL of test solution. The test employed three replicates of 10 daphnids each in six test
concentrations (measured in low and high treatments only) plus a negative and solvent control.
Nominal concentrations were 0 (negative and solvent controls), 50, 80, 128, 204.8, 327.68 and
524.29 mg/L. The authors provided mean measured concentrations before and after renewal:
49.62 and 43.93 mg/L (lowest concentration) and 526.9 and 476.41 mg/L (highest
concentration). Analyses of test solutions were performed using HPLC/MS and negative
electrospray ionization. The concentration of PFOA was calculated from standard curves (linear
in the concentration range of 1-800 ng/mL), and the average extraction efficiency was in the
range of 70-83%. The concentrations and chromatographic peak areas exhibited a significant
positive correlation (r = 0.9987, p < 0.01), and the water sample-spiked recovery was 99%. The
temperature, D.O., and pH were reported as having been measured every day during the acute
test, but results were not reported. Negative control and solvent control mortality were 0% each.
The author-reported 48-hour LCso for the study was 201.85 mg/L (95% C.I. = 134.7 - 302.5
mg/L). The EPA performed C-R analysis for the test and had an acceptable curve with an EPA-
calculated LCso of 222.0 mg/L PFOA (95% C.I. = 190.5 - 253.5 mg/L). The acute value was
acceptable for quantitative use.
A-22
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Publication: Yang et al. (2014)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 222.0 mg/L (95% C.I. = 190.5 - 253.5 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.1031
0.1773
6.2226
4.89 e10
e
309.4319
36.3820
8.5051
< 2.2 e"16
Concentration-Response Model Fit:
Yang et. al. 2014
Daphnia magna
Weibull type 1, 2 para
PFOA ( mg'L )
Barmentlo et al. (2015) performed a 48-hour static, measured acute test of PFOA (CAS
# 335-67-1, >96%) with Daphnia magna. The authors stated the test followed OECD 202 (2004)
guidelines for testing. D. magna used for testing were obtained from Grontmij, Amsterdam, and
cultured in M4 media according to OECD 211 (2012). Test organisms were less than 24-hours
old at test initiation. Dilution water was ISO medium. Experiments were conducted in 50 Ml
polypropylene tubes with 20 Ml of test solution. The photoperiod consisted of 16-hours of
illumination at an unreported intensity. PFOA stock was made with demineralized water. The
test involved four to six replicates of five daphnids each in five test concentrations plus a
A-23
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negative control. Nominal concentrations were not provided, but PFOA was measured in the
control, lowest and highest test concentrations. Based on these measurements, the authors
interpolated all test concentrations, 0.053 (negative control), 81, 128, 202, 318 and 503 mg/L.
Test temperature was maintained at 20 ฑ 1ฐC, Ph ranged from 7.00-7.82, and D.O. ranged from
8.54-9.42 mg/L. Mortality of daphnids in the negative control was not reported. The author-
reported 48-hour ECso for the study was 239 mg/L (95% C.I.: 190 - 287 mg/L). The EPA
performed C-R analysis for the test and had an acceptable curve with an EPA-calculated EC50 of
215.6 mg/L PFOA (95% C.I. = 181.7 - 249.5 mg/L). The acute value was acceptable for
quantitative use.
A-24
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Publication: Barmentlo et al. (2015)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 215.6 mg/L PFOA (95% C.I. = 181.7 - 249.5 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
2.3893
0.4288
5.5728
2.507 e8
e
251.3332
19.6475
12.7921
< 2.2 e"16
Concentration-Response Model Fit:
Barmentlo et al. 2015
Daphnia magna
WabuH type lt 2 para
PFOA ( mg/L )
Ding et al. (2012a) conducted a 48-hour static, partially measured acute test on PFOA
(CAS # 335-67-1; 96% purity from Sigma Aldrich) with I), magna. The test was performed
following OECD test guideline 202 (2004) with slight modifications. D. magna used for testing
were purchased from local suppliers and cultured for two months prior to use. Test organisms
were less than 24-hours old at test initiation. Dilution water was M4 solution prepared following
the OECD test guideline. The photoperiod consisted of a 16 hour:8-hour light:dark cycle at an
unreported light intensity. A primary stock solution was prepared in dilution (reconstituted M4)
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water. Exposure vessels were 50 Ml polypropylene disposable tubes containing 20 Ml of test
solution. The test involved four replicates of five daphnids each in six test concentrations plus a
negative control. Nominal concentrations were 0 (negative control), 0.35, 0.4, 0.45, 0.5, 0.55 and
0.6 Mm PFOA, or 0, 144.9, 165.6, 186.3, 207.0, 227.7 and 248.4 mg/L, respectively, after
conversion by multiplying the Mm concentration by a molecular weight of 414.07 g/mol for
PFOA. The subsequent concentrations are reported in the converted units of mg/L.
Concentrations of PFOA were confirmed in the highest and lowest concentrations, though only
nominal concentrations were reported. It was stated that the verified concentration was "well in
line with nominal concentrations". Test temperature was maintained at 20 ฑ 1ฐC. Observations
were made at 24-hours and 48-hours after test initiation. EC so values were reported for both
observational time periods. The 48-hour ECso was reported as 211.6 mg/L with the 95%
confidence levels of 184.7 - 255.5 mg/L and a NOEC of 207.0 mg/L. The EPA performed C-R
analysis for the test. The EPA-calculated ECso was 216.1 mg/L PFOA (95% C.I. = 206.1 - 225.9
mg/L) for D. magna, which was acceptable for quantitative use.
(It should be noted that these authors also conducted and reported on a similar acute
toxicity test with Chydorus sphaericus. However, this test was not used qualitatively or
quantitatively because personal correspondence with the authors indicated high control mortality.
Raw data provided by the authors did not clearly show 24- and 48-hour EC so data reported in the
publication. Data provided by the authors contained many PFOA C-R datasets for C. sphaericus.
Of these, the C-R dataset with nominal concentrations and 24- and 48-hour ECso values that best
matched those in the publication indicated high mortality in control responses. Therefore, the
results of this test were not included in the PFOA aquatic life criteria document.)
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Publication: Ding et al. (2012a)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 216.1 mg/L PFOA (95% C.I. = 206.1 - 225.9 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std.Error
t-stat
p-value
b
7.5008
1.3478
5.5650
2.621 e8
e
0.5478
0.0128
42.6800
< 2.2 e16
Concentration-Response Model Fit:
Ding et al. 2012
Daphnia magna
Weibull type 1,2 para
le-04 le-03 le-OQ le-01
PFOA ( mM )
Lu et al. (2016) evaluated the acute toxicity of PFOA (CAS# 335-67-1, 98% purity,
purchased from Tokyo Chemical Industry Co., Ltd., Tokyo, Japan) on Daphnia magna
immobilization. Reconstituted daphnia culture media was used for both culturing and test
solution preparation as described in OECD Test Guideline 202. D. magna cultures (originally
obtained from the Chinese Center for Disease Control and Prevention (Beijing, China) were fed
with the green algae Scenedesmus obliquits daily, maintained at 20ฐC and a light/dark
photoperiod of 16-hour/8-hour and the medium renewed three times weekly. The 48-hour static
unmeasured acute test was conducted via a modified OECD standard test procedure 202,
whereby five concentration treatments (3, 10, 30, 100 and 300 mg/L) plus a blank control were
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employed. Ten neonates (<24-hours old) from a designated brood were placed in a 100 Ml glass
beaker containing 45 Ml test solution for each test concentration and control. Test daphnids were
not fed during the testing period and each treatment was replicated three times. The status of
immobilization and mortality was checked at 48 hours (daphnids unable to swim within 15
seconds after gentle agitation of the test container are considered to be immobile and those
animals whose heartbeats have stopped are considered dead). The authors reported immobility/
survival to be a more sensitive endpoint than survival alone. The author-reported 48-hour EC so
for immobility/survival was 110.7 mg/L and the EPA-calculated 48-hour ECso was 114.595
mg/L (95% C.I. = 93.71 - 135.5 mg/L).
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Publication: Lu et al. (2016)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 114.595 mg/L (95% C.I. = 93.71 - 135.5 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std.Error
t-stat
p-value
b
0.5649
0.1050
5.3783
7.517 e8
e
219.2463
68.2427
3.2127
0.0013
Concentration-Response Model Fit:
Lu et al. 2016
Daphnia magna
Weibull type 1, 2 para
PFOA (mg/L )
Yang et al. (2019) evaluated the acute effects of PFOA (CAS# 335-67-1, purchased from
Sigma-Aldrich in St. Louis, MO) on Daphnia magna via a 48-hour unmeasured static exposure.
D. magna cultures were originally obtained from the Institute of Hydrobiology of Chinese
Academy of Science in Wuhan, China. Organisms were cultured in Daphnia Culture Medium
according to the parameters specified in OECD Guideline 202. Protocol for all testing followed
OECD Guideline 202. Cladocerans were cultured in artificial freshwater maintained at 20 ฑ 1ฐC
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under a 16-hour:8-hour light:dark photoperiod and a light intensity of 1,000-1,500 lux at the
surface of the water. Cultures were fed Scenedesmus obliquus daily and the water was changed
twice weekly. Reported water quality parameters include total hardness of 140-250 mg/L as
CaCC>3 and Ph of 6-8.5. Acute test concentrations included 0 (control), 0.000161, 0.000193,
0.000232, 0.000278, 0.000334 and 0.000401 mol/L (or 0 (control), 66.67, 79.92, 96.06, 115.1,
138.3 and 166.0 mg/L, respectively, given the molecular weight of the form of PFOA used in the
study, CAS # 335-67-1, of 414.07 g/mol). Five neonates (12-24 hours old) were placed randomly
in 100 Ml glass beakers filled with 60 Ml test solution, with four replicates per concentration.
Organisms were observed for mortality at 48 hours, and the authors reported a LCso of 0.000292
mol/L, or 120.9 mg/L PFOA. The EPA-calculated 48-hour LCso was 117.192 mg/L (95% C.I. =
112.2-122.2 mg/L).
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Publication: Yang et al. (2019)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated LCso: 117.192 mg/L (95% C.I. = 112.2 - 122.2 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
6.0229
1.0808
5.5724
2.512 e8
d
0.9998
0.0103
96.6600
< 2.2 e16
e
117.1921
4.8691
24.0685
< 2.2 e16
Concentration-Response Model Fit:
Yang et al. 2019
Daphnia magna
Log Logistic type 1, 3 para
PFOA ( mg/L )
Chen et al. (2022) tested the acute toxicity of ammonium perfluorooctanoic acid (APFO)
to Daphnia magna in a 48-hour unmeasured, static experiment. In addition, the authors tested the
acute effects of pristine and aged polystyrene particles, and the combined effects of these
microplastics and AFPO, but only the results of the AFPO exposure are summarized here. AFPO
(>98% purity) was purchased from Shanghai Aladdin Biochemical Technology Co., Ltd., China.
Organisms were purchased from Everbright Algae Species Co., China. They were housed in
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culture medium described in method OECD 202 (OECD 2004) at 20ฑ1ฐC under a 12:12
light:dark cycle at 2000ฑ200 lux. Culture medium was renewed twice a week, and organisms
were fed green algae daily. The acute APFO exposure generally followed test method OECD
202 (OECD 2004). Stock solution of 50 g/L APFO was prepared by mixing the chemical with
ultrapure laboratory water. For each replicate, 10 neonates were exposed to 50 Ml of test
solutions with nominal concentrations of 0, 10, 50, 100, 200, and 500 mg/L APFO, made by
diluting the stock solution with culture medium. There were three replicates per treatment, and
test vessels were housed on a bath oscillator rotating at 150 rpm. Test organisms were not fed
during the experiment. A reference toxicity test was performed with K2Cr07. Control mortality
was less than 10% in the AFPO and reference toxicity tests. The LCso and EC so values were
calculated using a logistic model in GraphPad Prism, version 9.2.0, and the 48-hour EC so for
immobility was 156.9 mg/L, which was determined to be acceptable for quantitative use.
A.2.5 Fifth most acutely sensitive genus - Brachionus
Zhang et al. (2013a) performed a 24-hour static test of PFOA (CAS # 335-67-1, 96%
purity) with Brachionus calyciflorus. Organisms were neonates less than two-hours old at test
initiation. All animals were parthenogenetically-produced offspring of one individual from a
single resting egg collected from a natural lake in Houhai Park (Beijing, China). The rotifers
were cultured in an artificial inorganic medium at 20ฐC (16-hour:8-hour, light:dark; 3,000 lux)
for more than six months before toxicity testing to acclimate to the experimental conditions. All
toxicity tests were carried out in the same medium and under the same conditions as during
culture (i.e., Ph, temperature, illumination). Solvent-free stock solutions of PFOA (1,000 mg/L)
were prepared by dissolving the solid in deionized water via sonication. After mixing, the
primary stock was proportionally diluted with dilution water to prepare the test concentrations.
Exposures were in 15 Ml, six-well cell culture plates (assumed plastic) each containing at total of
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10 Ml of test solution. The test employed seven measured test concentrations plus a negative
control. Each treatment consisted of one replicate plate of 10 rotifers each in individual cells and
repeated six times. Nominal concentrations were 0 (negative control), 60, 80, 100, 120, 140, 160,
and 180 mg/L. PFOA concentrations were not measured in the rotifer exposures, but rather, in a
side experiment using HPLC/MS. The side experiment showed that the concentration of PFOA
measured every eight-hours over a 24-hour period in rotifer medium with green algae incurs
minimal change in the concentration range from 0.25 to 2.0 mg/L. The acute test was conducted
without green algae added to the exposure medium. Although this rotifer species has a short life
span, a 24-hour unfed test is not expected to cause starvation and 0% mortality was observed at
test termination in the negative control. The study reported 24-hour LCso was 150.0 mg/L. The
acute value was acceptable for quantitative use.
Publication: Zhang et al. (2013a)
Species: Rotifer (Brachionus calyciflorus)
Genus: Brachionus
EPA-Calculated LCso: Not calculable, concentration-response data not available
Concentration-Response Model Fit: Not Applicable
A.2.6 Sixth and seventh most acutely sensitive genera - Ligumia and Lamysilis (mussels)
Hazelton (2013) and Hazelton et al. (2012) evaluated the acute effects of PFOA (96%
purity) on two freshwater mussels: Lampsilis siliquoidea and Ligumia recta. Acute toxicity was
observed under static conditions over a 24-hour period (<24-hour old glochidia) or renewal
conditions over a 96-hour period (four to six-week-old juveniles). The authors stated that the
tests followed ASTM E2455-06 (2006). Dilution water was hard reconstituted water.
Photoperiod and light intensity were not reported. No details were provided regarding primary
stock solution and test solution preparation. Experiments were conducted in 3.8 L glass jars of
unspecified fill volume. The test employed three replicates of 150 glochidia or seven juvenile
mussels each in six measured test concentrations plus a negative control (10 juveniles for the
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control treatment). Nominal concentrations were 0 (negative control), 0.005, 0.05, 0.5, 5, 50 and
500 mg/L, while corresponding mean measured concentrations were less than the limit of
quantification (LOQ, specifics not provided), 0.0051, 0.0484, 0.490, 4.8, 51 and 476 mg/L
PFOA, respectively. Analyses of test solutions were performed at the U.S. EPA National
Exposure Research Laboratory in Research Triangle Park, NC using HPLC/MS. Measured test
concentrations of PFOA were within 10% of target in water from acute tests. Recovery of PFOA
standards ranged from 91.2-108%. For all acute tests, alkalinity ranged from 97 to 110 mg/L as
CaC03 with a mean of 104.4 mg/L; total hardness ranged from 132 to 162 mg/L as CaC03 with
a mean of 149.6 mg/L; conductivity ranged from 514 to 643 |iS/cm with a mean of 556.5 |iS/cm;
Ph ranged from 8.05 to 8.56 with a mean of 8.46; and dissolved oxygen ranged from 8.16 to 9.46
mg/L with a mean of 8.62 mg/L (n = 12 for alkalinity and total hardness, n = 55 for all other
parameters). Exposures were conducted at 20ฐC. Mortality of mussels in the negative control
was <10% in all exposures. The 24-hour EC so reported for glochidia of L. siliquoidea was 164.4
mg/L (95% C.I.: 116.0 - 232.8 mg/L) and fori, recta, 161.0 mg/L (95% C.I.: 135.0 - 192.7
mg/L). The 96-hour LCso values for the juvenile L. siliquoidea and L. recta were greater than the
highest test concentration (500 mg/L). The study reported 24-hour ECsos for L. siliquoidea and
for L. recta represent acute values acceptable for quantitative use for the two mussel species. The
juvenile life stage is less sensitive, such that its LCsos were not used quantitatively in SMAVs.
A.2.7 Eighth most acutely sensitive genus - Xenoyus
Kim et al. (2013) conducted a 96-hour renewal unmeasured assay with perfluorooctanoic
acid (PFOA) using the frog embryo teratogenesis assay -Xenopus (FETAX). PFOA stock
solutions were prepared by dissolving PFOA in dimethyl sulfoxide (DMSO), and then diluting in
FETAX medium for exposure solutions (DMSO did not exceed 0.15%). Adult Xenopus were
purchased from Nasco (Fort Atkinson, WI) and housed in clear plastic aquariums with
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dechlorinated tap water at 18 ฑ 2ฐC with a 12-hour light cycle and fed three times a week.
Ovulation was induced by injecting 1,000 IU of human chorionic gonadotropin just under the
skin of a female in the evening. The next day, females laid eggs in 60 mm plastic dishes. The
eggs were immediately fertilized in 0. IX modified Barth solution (MBS) (Xenopus testes were
obtained from sacrificed males). Following successful fertilization, the jelly coat was removed
by swirling the embryos in a 2% L-cysteine solution. The embryos were then transferred to IX
MBS containing 3% Ficoll 400. Unfertilized eggs and dead embryos were removed and
maintained at 22 ฑ 0.5ฐC. Finely cleaved embryos in the blastula stage (stage 8.5) were selected,
with 20-25 embryos used per concentration (nominal concentrations of 100, 500, 750, 1000 and
1,250 |iM PFOA, or 41.4, 207.0, 310.6, 414.1 and 517.6 mg/L PFOA, respectively). DMSO
alone (0.1%) and FETAX medium alone were used as controls. Embryos were incubated at 23ฐC
until the end of the assay. The media were changed every day, and dead embryos were removed.
At the end of the experiments, embryo mortality was recorded and surviving embryos were fixed
in 4% formaldehyde to check for malformation. Head-tail lengths and malformations were
analyzed to measure growth inhibition. The authors reported a 96-hour LCso of 377 mg/L PFOA
and the value was acceptable for quantitative use.
A.2.8 Ninth most acutely sensitive genus - Dugesia
Li (2008) conducted a 96-hour static, unmeasured acute toxicity test on PFOA (CAS #
3825-26-1, >98% purity) with the planarian, Dugesia japonica (a non-North American species).
The test organisms were originally collected from Nan-shi stream located in Wu-lai, Taipei
County, Taiwan in 2004 and maintained in the laboratory in dechlorinated tap water. The
planarians had a body length of 0.9 ฑ0.1 cm at test initiation. The dilution water was
dechlorinated tap water and a primary stock solution of PFOA was prepared in the same dilution
water. The photoperiod consisted of 12-hours of illumination at an unreported intensity.
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Exposure vessels were polypropylene beakers of unreported dimensions with a 50 Ml fill
volume. The test employed five replicates of five planarians each in at least five test
concentrations plus a negative control. Nominal test concentrations were in the range of 100-750
mg/L PFOA. The test temperature was maintained at 25 ฑ 1ฐC. No other water quality
parameters were reported for test solutions. Survival of negative control animals was not
reported. The study reported a 96-hour LCso was 458 mg/L (95% C.I. = 427 - 491 mg/L). The
acute value was acceptable for quantitative use.
Li (2009) conducted a second 96-hour static, unmeasured acute test of PFOA
(ammonium salt, >98% purity) with Dugesia japonica. Again, the tested individuals were
originally collected from Nan-shi stream located in Wu-lai, Taipei County, Taiwan in 2004 and
maintained in the laboratory in dechlorinated tap water. The planarians had a body length of 0.9
ฑ 0.1 cm at test initiation. The dilution water was dechlorinated tap water and a primary stock
solution of PFOA was prepared in the same dilution water. The photoperiod consisted of 12-
hours of illumination at an unreported intensity. Exposure vessels were made of polyethylene
with unreported dimensions and 50 Ml fill volume. The test employed three replicates of 10
planarians each in at least five test concentrations plus a negative control. The test was repeated
three times with different test concentrations. Nominal test concentrations were in the range of
150-750 mg/L PFOA. The test temperature was maintained at 25 ฑ 2ฐC. Water quality
parameters including Ph, conductivity, and D.O. were reported as having been measured at the
beginning and end of each test, but the information was not provided. Organisms were not fed,
and no mortality was observed in the control groups in any of the three tests. The author-reported
96-hour LCso was 337 mg/L (95% C.I. = 318-357 mg/L) which was averaged across the three
tests. The EPA performed C-R analysis for each individual test. Two of the tests had acceptable
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curves with the EPA-calculated LC50 values of 321.8 mg/L PFOA (95% C.I. = 290.6 - 353.1
mg/L) and 383.0 mg/L PFOA (95% C.I. = 347.8 - 418.2 mg/L) which were acceptable for
quantitative use. The third curve had a poor concentration response and the LC50 (427.7 mg/L;
95% C.I. = 251.4 - 604.1 mg/L) was, therefore, not used quantitatively but considered for
qualitative use only.
A.2.9 Tenth most acutely sensitive genus - Neocaridina
Li (2009) conducted a 96-hour acute test on PFOA (ammonium salt, >98%) purity) with
the freshwater shrimp species, Neocaridina denticulata (a non-North American species). Test
conditions were static (no solution renewal), and test concentrations were unmeasured. Test
organisms were obtained from an unspecified local supplier and acclimated in the laboratory for
at least seven days prior to the experiments. N. denticulata of unspecified age were used at test
initiation and were reported to be 1.3 ฑ 0.2 cm long. Dilution water was dechlorinated tap water.
The photoperiod consisted of 12-hours of illumination at an unreported light intensity. A primary
stock solution was prepared in dilution water. Exposure vessels were polypropylene beakers of
unreported dimensions and 1 L fill volume. The test employed five replicates of six organisms
each in at least five test concentrations plus a negative control. Each treatment was tested three
different times. Nominal test concentrations were in the range of 50-1,000 mg/L PFOA. The test
temperature was maintained at 25 ฑ 2ฐC. Water quality parameters including Ph, conductivity,
and D.O. were reported as having been measured at the beginning and end of each test, but the
information is not reported. Mortality of negative control animals was 10%> for one treatment, but
0%> in others. The author-reported 96-hour LC50 reported in the study was 454 mg/L (95%> C.I.:
418-494 mg/L) which was averaged across three tests. The EPA performed C-R analysis for each
individual test. All three tests had acceptable curves with the EPA-calculated LC50S of 499.7
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mg/L (95% C.I. = 457.4 - 542.1 mg/L), 428.1 mg/L (95% C.I. = 396.3 - 459.9 mg/L), and 375.5
mg/L (95%) C.I. = 296.5 - 454.4 mg/L), which were acceptable for quantitative use.
A.2.10 Eleventh most acutely sensitive genus -Danio
The acute effects of PFOA on the zebrafish, Danio rerio., have been reported by
numerous researchers. Ye et al. (2007) evaluated the acute effects of perfluorooctanoic acid
(PFOA, purity not reported) on D. rerio in a 96-hour static-renewal test with unmeasured
treatment concentrations. The PFOA stock solution of 1,500 mg/L was maintained at a Ph of 8.2
with phosphate buffer and test substances were agitated in the reconstituted water by
ultrasonification. The solutions were stored at 4 ฑ 1ฐC. No added solvents were used to
formulate test concentrations from the stock solution. Test fish (AB strain) were obtained from
the School of Life Sciences at Fandan University, Shanghai, China. Breeding fish (1.5 years old)
were fed live brine shrimp twice daily and kept with a light:dark period of 14:10 in aquaria
containing aerated natural water. The Ph ranged from 7 to 8.5 and water temperature was
maintained at 26 ฑ1ฐC in test solutions. Embryos were obtained from spawning adults, usually
five male and three female. For each toxicant concentration, 48 embryos were randomly
distributed into each well of 24-well polystyrene multi-well plates, with four eggs per well. Each
well was filled with 2.5 mL of test solution, which was totally renewed daily by transferring
embryos in newly cleaned wells. Nominal exposure concentrations were 0 (control), 1.5, 45, 100,
400, 800 and 1,500 mg/L PFOA. The multi-well plates were kept at 26 ฑ 1ฐC, with a
photoperiod of 16:8-hour light:dark. The observations of embryos were made at distinct stages,
which represent important steps of zebrafish development. The author reported 96-hour LCso
was 499 mg/L PFOA. The EPA could not calculate an LCso value because sufficient C-R data
were not reported by Ye et al. (2007). The author-reported LCso value (i.e., 499 mg/L) was
acceptable for quantitative use in deriving the freshwater acute PFOA criterion.
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Hagenaars et al. (2011) exposed/), rerio embryos to PFOA (CAS #335-67-1, purity
>97%) under static unmeasured conditions for 120 hours. The PFOA was dissolved in medium-
hard reconstituted laboratory water, which was aerated and kept at 26ฐC until use (no solvent).
Adult wildtype zebrafish (breeding stock) were obtained from a commercial supplier (Aqua
hobby, Heist-op-den-berg, Belgium) and kept in aerated and biologically filtered medium-hard
reconstituted freshwater. Four males and four females were used for egg production, with
fertilized eggs collected in egg traps within 30 minutes of spawning. Eggs were transferred to the
test solutions within 60 minutes after spawning. Eggs with anomalies or damaged membranes
were discarded, and fertilized eggs were separated from the non-fertilized eggs using a
stereomicroscope. Twenty normally shaped fertilized eggs per exposure concentration were
divided over a 24-well plastic plate and each embryo was placed individually in 2 mL of the test
solution. The remaining four wells were filled with clean water and used for the control eggs.
Two replicate plates were used for each exposure concentration resulting in 40 embryos per
exposure condition at the beginning of the experiment. The 24-well plates were covered with a
self-adhesive foil, placed in an incubation chamber at 26 ฑ 0.3ฐC, Ph 7.2-7.5 and subjected to a
14-hour: 10-hour, light:dark cycle. A test was considered valid if more than 90% of the controls
successfully hatched and showed neither sublethal nor lethal effects. The authors reported a 96-
hour LCso of >500 mg/L PFOA and was classified as quantitative.
Zhao et al. (2016) performed a 96-hour static acute PFOA (98% purity; obtained from
J&K Scientific) test on zebrafish, D. rerio. Authors stated that the test followed OECD Guideline
203. Zebrafish were obtained from Wuhan Institute of Hydrobiology at the Chinese Academy of
Science. Test organisms were three months old and 2.5-3.0 cm in length at test initiation. The
dilution water used for testing was not provided. Experiments were conducted in 5-liter glass fish
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tanks filled with three liters of solution. The photoperiod and preparation of test solutions was
not described. The test involved 10 fish in five unmeasured test concentrations plus a negative
control. Nominal concentrations were 0 (negative control), 72.3, 86.8, 104.2, 125.0 and 150.0
mg/L. However, nominal concentrations of 50, 100, 200 mg/L PFOA were confirmed by HPLC
in a side study without organisms. Measured concentrations in the side study were 80-120
percent of nominal. Test temperature was maintained at 23 ฑ 2ฐC, but other water quality
parameters were not reported. There was no mortality in the negative control at the end of the
exposure period. The author-reported 96-hour LCso was 118.82 mg/L (95% C.I. = 107.46 -
134.33 mg/L). The EPA performed C-R analysis for the test and calculated an LCso of 122.5
mg/L PFOA (95% C.I. = 119.1 - 125.8 mg/L), which was acceptable for quantitative use in
deriving the acute PFOA freshwater criterion. D. rerio embryos (4 hours post-fertilization; hpf)
were also subjected to PFOA by Godfrey et al. (2017a) in a 96-hour acute toxicity test using
static renewal exposures that were not analytically confirmed. Stock solutions were prepared by
dissolving PFOA in 1 L of reverse osmosis (RO) water containing 12.5 |iL Replenish (Seachem
Laboratories Inc.) and then adjusted to neutral Ph (7-7.5). Adult zebrafish, AB wild-type, were
maintained at a water temperature of 28 ฑ 1ฐC and a photoperiod of 14-h L: 10-h D. Fish were
fed twice daily, Artemia nauplii in the morning and Tetramin in the afternoon, and genders were
kept separate overnight at a ratio of 2 males: 1 female. Randomly collected embryos (20 per
concentration, gastrula stage, 4-hpf) were placed in plastic petri dishes containing 25 Ml of
exposure solution for 96-hours at 28ฐC. Each test consisted of a minimum of two replicates per
dose and test solutions were renewed daily (nominal exposure solutions ranged from 250-1,000
mg/L PFOA). The author-reported 96-hour LCso was 473.0 mg/L PFOA which was averaged
across four tests. The EPA performed C-R analysis for each individual test. Three tests had
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acceptable curves with the EPA-calculated LCsos of 548.0 mg/L (95% C.I. = 530.6 - 565.5
mg/L), 508.5 mg/L (95% C.I. = 471.4 - 545.6 mg/L), and 547.0 mg/L (95% C.I. = 516.0 - 578.0
mg/L), which were acceptable for quantitative use. The fourth test had an unacceptable curve and
therefore the EPA-calculated LCso of 560.1 mg/L PFOA (95% C.I. = 556.4 - 563.8 mg/L) was
not used.
Stengel et al. (2017, 2018) exposed/), rerio embryos to PFOA for 96-hours using
renewal unmeasured procedures as specified in OECD (2013) guidelines. PFOA stock and
exposure solutions were prepared in reconstituted laboratory water. All adult zebrafish used for
breeding were wild-type descendants of the "Westaquarium" strain and obtained from the
Aquatic Ecology and Toxicology breeding facilities at the University of Heidelberg. Details of
zebrafish maintenance, egg production and embryo rearing are provided as described previously
(Kimmel et al. 1995, 1988; Nagel 2002; Spence et al. 2006; Wixon 2000) and have been updated
for the purpose of the zebrafish embryo toxicity test by Lammer et al. (2009). Embryos were
exposed at the latest from 1 hpf in glass vessels, which had been preincubated (saturated) for at
least 24 hours, to a series of nominal PFOA dilutions (0, 400, 512, 640, 800 and 1,000 mg/L).
After verifying fertilization success, embryos were individually transferred to the wells of 24-
well plates, which had been pre-incubated with 2 mL of the test solution per well for 24 hours
prior to the test start and kept in an incubator at 26.0 ฑ 1,0ฐC under a 14-hour: 10-hour light:dark
regime. In order to prevent evaporation or cross-contamination between the wells, the plates
were sealed with self-adhesive foil. Embryo tests were classified as valid if the mortality in the
negative control was <10%, and if the positive control (3,4-dichloroaniline) showed mortalities
between 20% and 80%. All fish embryo tests were run in three independent replicates. The
author-reported 96-hour LCso was 759 mg/L PFOA. The EPA performed C-R analysis for the
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test and had an acceptable curve with an EPA-calculated LC50 of 806.6 mg/L (95% C.I. = 773.6
- 839.6 mg/L) and was determined to be quantitatively acceptable for criterion derivation.
Corrales et al. (2017) exposed/), rerio embryos to PFOA for 96-hours employing static
unmeasured procedures (U.S. EPA 2002, OECD 2013). Tropical 5D wild type adult zebrafish
were kept at a density of less than four fish per liter in a z-mod recirculating system with water
(Ph 7.0, 260 ppm Instant Ocean) maintained at 26-28ฐC and a 16-hour:8-hour light/dark cycle.
Zebrafish were fed twice daily with brine shrimp (Artemia sp. Nauplii) and once per day with
TetraMin Tropical Flakes. Sexually mature fish were bred to produce embryos for toxicity
studies. Glass beakers were used as experimental units; 15 zebrafish embryos in 100 Ml beakers
containing 30 ml test solution. Before the start of each experiment, all solutions were titrated to
Ph 7.5 following standard methods. General water chemistry measures (e.g., alkalinity, total
hardness, dissolved oxygen, and temperature) were also routinely monitored (assume same
culture and test physico-chemical test conditions). The author-reported 96-hour LC50 was 24.6
mg/L PFOA. The EPA-calculated LC50 was 22.77 mg/L PFOA (95% C.I. = 13.30 - 32.20
mg/L), which was acceptable for quantitative use. This LC50 value, however, was excluded from
derivation of the acute criterion because a comparative assessment between this LC50 value and
the other seven quantitatively-acceptable zebrafish LC50 values available (Ye et al. 2007; Zhao et
al. 2016; Godfrey et al. 2017a; Hagenaars et al. 2011; Stengel et al. 2017, 2018), indicated that
the LC50 reported by Corrales et al. (2017) was an outlier, falling out more than five times lower
than the next lowest D. rerio LC50 value (Zhao et al. 2016) and nearly 20 times lower than the
final D. rerio SMAV.
A.2.11 Twelfth most acutely sensitive genus - Pimeyhales
The 3M Company (2000a) conducted a 96-hour static, unmeasured acute toxicity test
with the fathead minnow, Pimephalespromelas, and PFOA (CAS # 335-67-1). The toxicant was
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part of the 3M production and was characterized as a mixture of PFOA (95-98% of the
compound) and perfluorochemical inert compounds (1-5% of the compound). No specific test
protocol was identified. A stock of PFOA was made by dissolving the test substance with NaOH
and diluting the stock with carbon-filtered well water to make five test concentrations (690, 750,
810, 870 and 930 mg/L) plus a control (well water only). Exposures were conducted in glass
beakers with 5 L of test solution and five fish per beaker (0.5 g/L fish loading). There were two
replicates for each treatment. Test conditions throughout the experiment varied little (D.O.: 6.1-
7.7 mg/L; Ph: 7.5-7.7; 19-20ฐC). No mortality occurred in the control and PFOA treatments
<750 mg/L. The author-reported 96-hour LCso was 843 mg/L. The EPA performed C-R analysis
for the test. The EPA-calculated LCso was 852.7 mg/L PFOA (95% C.I. = 834.4 - 870.9 mg/L),
which was acceptable for quantitative use and was used for criterion derivation.
Corrales et al. (2017) also evaluated the acute toxicity of PFOA to the fathead minnow
(Pimephalespromelas). Embryos were exposed to PFOA for 96-hours employing static
unmeasured procedures (U.S. EPA 2002, OECD 2013). Fish were housed in a flow-through
system supplied with aged, dechlorinated tap water at a constant temperature of 25 ฑ1ฐC under a
16-hour: 8-hour light/dark photoperiod. They were fed twice daily with brine shrimp (Artemia sp.
Nauplii) and TetraMin Tropical Flakes. Individuals were aged to at least 120 days before
breeding at which time they were placed in tanks in a 1:4-5 male to female ratio. Embryos were
collected, and within 24 hours post hatched larvae were used for toxicity studies. Glass beakers
were used as experimental units; 10 fathead minnow larvae were placed in each 500 Ml beaker
containing 200 ml test solution. Before the start of each experiment, all solutions were titrated to
Ph 7.5 following standard methods. General water chemistry measures (e.g., alkalinity, total
hardness, dissolved oxygen, and temperature) were also routinely monitored (assume same
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culture and test physico-chemical test conditions). The reported 96-hour LC50 was 413.2 mg/L
PFOA and was determined to be quantitatively acceptable for criterion derivation.
A.2.12 Thirteenth most acutely sensitive genus - Hyla
Tornabene et al. (2021) conducted an acute PFOA (purchased from Sigma Aldrich,
Catalog # 171468-25G; purity not provided) toxicity test with the gray treefrog, Hyla versicolor.
The acute test followed standard 96-hour acute toxicity test guidance (U.S. EPA 2002; ASTM
2017). Frog egg masses were collected from the field in the wetlands of Indiana near the campus
of Purdue University. Collected egg masses were raised outdoors in 200 L polyethylene tanks
filled with well water. Experiments began when frogs reached Gosner stage 26, defined as when
larvae are free swimming and feeding. Before test initiation larvae were acclimated to test
conditions (21ฐC and 12-hour: 12-hour light:dark photoperiod) for 24 hours. A stock solution of
PFOA (2,000 mg/L) was made in UV-filtered well water and diluted with well water to reach
test concentrations (ranged from 0-2,000 mg/L PFOA). Test concentrations were not measured in
test solutions based on previous research that demonstrated limited degradation under similar
conditions. Larva were transferred individually to 250 Ml plastic cups with 200 Ml of test
solution and were not fed during the exposure period. There were nine to 10 replicates for each
treatment and no mortality occurred in the controls. The author reported 96-hour LC50 was 557
mg/L and the EPA-calculated 96-hour LC50 values was 646.2 mg/L (95% C.I. = 588.0 - 704.4
mg/L), which was acceptable for quantitative use. Note, the authors also reported a qualitatively
acceptable test for the same species (Gosner stage 40) that is described in Table G. 1.
A.2.13 Fourteenth most acutely sensitive genus -Lepomis
The DuPont Haskell Laboratory (2000) evaluated the acute toxicity of ammonium
perfluorooctanoate (APFO, 99% purity) to the bluegill sunfish, Lepomis macrochirus. The static
unmeasured GLP study exposed 2.1 cm fish for 96 hours (dilution water not identified). Fish
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used in this study were not fed approximately 24-hours prior to and during the test. Bluegill
sunfish were assigned to the test chambers using random numbers. Nominal APFO
concentrations were 262, 328, 410, 512, 640, 800 and 1,000 mg/L. Glass aquaria (20L)
containing 10 L of test solution were employed. Positions of test chambers in the water bath used
to maintain constant temperature were assigned using random numbers. Ten fish were added to
each replicate using random numbers (2 replicates per concentration; total 20 fish per
concentration). A photoperiod of 16 hours light (312-344 lux) versus eight hours darkness was
employed with 25 minutes of transitional light (<2.15 lux) preceding and following the 16-hour
light interval. Observations for mortality and behavioral effects were made daily. All chemical
and physical parameters were within expected ranges. Total alkalinity and EDTA total hardness
of the dilution water control were 79 mg/L CaCC>3 and 76 mg/L CaCC>3, respectively. During the
test, dissolved oxygen concentrations ranged from 6.7-8.5 mg/L, Ph ranged from 6.9-7.4, and
temperature ranged from 21.4-22.1ฐC. No fish died in the controls. The authors reported a 96-
hour LCso of 634 mg/L APFO. The EPA performed C-R analysis for the test and had an
acceptable curve with an EPA-calculated LC50 of 664.0 mg/L (95% C.I. = 631.4 - 696.7 mg/L),
which was determined to be quantitatively acceptable for criterion derivation.
A.2.14 Fifteenth most acutely sensitive genus -Physella
Li (2009) conducted a 96-hour static unmeasured acute test on PFOA (ammonium salt,
>98% purity) with the snail species, Physella acuta (Note: formerly called Physa acuta). The test
organisms were collected from a ditch located in Shilin of Taipei City in June 2005. Snails were
fed with lettuce and half of the culture medium was changed with dechlorinated water every two
weeks, implying a holding time of greater than two weeks. Snails of mixed ages (shell length 0.6
ฑ 0.2 cm) were used at test initiation. The dilution water was dechlorinated tap water, and a
primary stock solution of PFOA was prepared in the same dilution water. The photoperiod
A-45
-------
consisted of 12-hours of illumination at an unreported intensity. Exposure vessels were made of
polyethylene with unreported dimensions and 1 L fill volume. The test employed five replicates
of six snails each in at least five test concentrations plus a negative control. Nominal test
concentrations were in the range of 100-1,000 mg/L PFOA. The test temperature was maintained
at 25 ฑ 2ฐC. Water quality parameters including pH, conductivity, and D.O. were reported as
having been measured at the beginning and end of each test, but the information is not reported.
Organisms were not fed, and no animals died in the control groups. The author-reported 96-hour
LCso was 672 mg/L (95% C.I.: 635-711 mg/L) which was averaged across three tests. The EPA
performed C-R analysis for each individual test. All three tests had acceptable curves with the
EPA-calculated LCsos of 762.0 mg/L (95% C.I. = 706.1 - 817.9 mg/L), 659.9 mg/L (95% C.I. =
607.9 - 711.8 mg/L), and 628.3 mg/L (95% C.I. = 582.9 - 673.7 mg/L), which were acceptable
for quantitative use.
A.2.15 Sixteenth most acutely sensitive genus - Ambystoma
Tornabene et al. (2021) conducted acute toxicity tests with three species of salamanders
in the genus Ambystoma and PFOA (purchased from Sigma Aldrich, Catalog # 171468-25G;
purity not provided). Acute tests followed standard 96-hour acute toxicity test guidance (U.S.
EPA 2002; ASTM 2017). The three test species (Jefferson salamander, Ambystoma
jeffersonianum\ small-mouthed salamander, A. texanum\ eastern tiger salamander, A. tigrinum)
were collected from the field in the wetlands of Indiana near the campus of Purdue University.
Collected egg masses were raised outdoors in 200 L polyethylene tanks filled with well water.
Experiments began when salamanders reached Harrison stage 40, defined as when larvae are free
swimming and feeding. Before test initiation larvae were acclimated to test conditions (21ฐC and
12-hour: 12-hour light:dark photoperiod) for 24 hours. An additional acute test with Harrison
stage 45 small-mouthed salamanders was run to determine if toxicity varied between life stages.
A-46
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A stock solution of PFOA (2,000 mg/L) was made in UV-filtered well water and diluted with
well water to reach test concentrations (ranged from 0-2,000 mg/L PFOA). Test concentrations
were not measured in test solutions based on previous research that demonstrated limited
degradation under similar conditions. Larva were transferred individually to 250 Ml plastic cups
with 200 Ml of test solution and were not fed during the exposure period. The number of
replicates varied by species, life stage and treatment; five replicates per treatment for Jefferson
salamander and Harrison stage 45 small-mouthed salamander, five to seven replicates per
treatment for Harrison stage 40 small-mouthed salamander, and 20 replicates in the control and
10 replicates in each treatment for eastern tiger salamander. Only one salamander larva died in
the controls across all tests (eastern tiger salamander test), acute values from the four tests
include:
Jefferson salamander: The author-reported 96-hour LCso was 1,070 mg/L. The EPA
was unable to fit a C-R model with significant parameters and relied on the author
reported value as quantitatively acceptable.
Harrison stage 40 small-mouthed salamander: The author-reported 96-hour LCso
was 474 mg/L. The EPA-calculated LCso was 407.3 mg/L (95% C.I. = 303.7 - 0.510.9
mg/L), which was acceptable for quantitative use.
Harrison stage 45 small-mouthed salamander: The author-reported 96-hour LCso
was 1,000 mg/L. The EPA was unable to fit a C-R model with significant parameters
and relied on the author-reported value as quantitatively acceptable; however, the LCso
from this test was more than two times greater than the Harrison stage 40 small-
mouthed salamander indicating the Harrison stage 45 was a relatively tolerant life
stage. As a result, the LCso from this test was not used in the SMAV calculation for A.
texanum.
Eastern tiger salamander: The author-reported 96-hour LCso was 752 mg/L.
Concentration-response data from this test lacked partial effects and the EPA was
A-47
-------
unable to fit a C-R model with significant parameters and relied on the author reported
value as quantitatively acceptable.
A.2.16 Seventeenth most acutely sensitive genus - Anaxyrus
Tornabene et al. (2021) conducted acute PFOA (purchased from Sigma Aldrich, Catalog
# 171468-25G; purity not provided) toxicity tests with the American toad, Anaxyrus americanus.
The acute tests followed standard 96-hour acute toxicity test guidance (U.S. EPA 2002; ASTM
2017). The toad egg masses were collected from the field in the wetlands of Indiana near the
campus of Purdue University. Collected egg masses were raised outdoors in 200 L polyethylene
tanks filled with well water. Experiments began when frogs reached Gosner stage 26, defined as
when larvae are free swimming and feeding. An additional acute test with Gosner stage 41 was
conducted to determine if toxicity varied between life stages. Before test initiation larvae were
acclimated to test conditions (21ฐC and 12-hour: 12-hour light:dark photoperiod) for 24 hours. A
stock solution of PFOA (2,000 mg/L) was made in UV-filtered well water and diluted with well
water to reach test concentrations (ranged from 0-2,000 mg/L PFOA). Test concentrations were
not measured in test solutions based on previous research that demonstrated limited degradation
under similar conditions. Larva were transferred individually to 250 Ml plastic cups with 200 Ml
of test solution and were not fed during the exposure period. The number of replicates varied by
treatment for both tests; 10 replicates for each treatment except the 1,750 mg/L PFOA exposure
which had nine replicates. No mortality occurred in any of the control groups. The authors did
not find a significant difference between the life stages of the American toad, so results of the
two tests were pooled to determine the 96-hour author-reported LCso of 711 mg/L. The EPA-
calculated 96-hour LCso value was 781.4 mg/L (95% C.I. = 748.3 - 814.4 mg/L) for the Gosner
stage 26 test and was 806.6 (95% C.I. = 760.6 - 852.6 mg/L) mg/L for the Gosner stage 41 test,
both of which were quantitatively acceptable for use.
A-48
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A.2.17 Eighteenth most acutely sensitive genus - Lithobates
Flynn et al. (2019) evaluated the acute effects of PFOA (CAS# 335-67-1, purchased
from Sigma-Aldrich) on the American bullfrog (Lithobates ca/esbeiana, formerly, Rana
catesbeiana) during a 96-hour static unmeasured study. Testing followed Purdue University's
Institutional Animal Care and Use Committee Guidelines Protocol #16010013551. American
bullfrog eggs were taken from a permanent pond in the Martell Forest outside of West Lafayette,
Indiana. The eggs from a single egg mass were acclimated in 100 L outdoor tanks filled with 70
L of aged well water and covered with a 70% shade cloth. Once hatched, tadpoles were fed
rabbit chow and TetraMin ad libitum and were acclimated to laboratory conditions for 24 hours
before testing (21ฐC and a 12-hour: 12-hour light:dark photoperiod). A 2,000 mg/L PFOA stock
solution was prepared with reverse osmosis water to produce 12 nominal test concentrations of
PFOA [0 (control), 10, 100, 250, 500, 750, 1,000, 1,250, 1,500, 1,750, 2,000 and 2,500 mg/L],
Each test treatment contained 10 replicates with one Gosner Stage 25 tadpole in each 250 Ml
plastic tub. Mortality was monitored twice daily. The author reported a 96-hour LCso value of
1,004 mg/L PFOA. The EPA performed C-R analysis for the test and the EPA-calculated 96-
hour LCso was 1,006 mg/L (95% C.I. = 992.8 - 1,018 mg/L).
Tornabene et al. (2021) conducted acute PFOA (purchased from Sigma Aldrich, Catalog
# 171468-25G; purity not provided) toxicity tests with four species of frogs in the genus
Lithobates (formerly, Rana). Acute tests followed standard 96-hour guidance (U.S. EPA 2002;
ASTM 2017). The four test species (American bullfrog, Lithobates catesbeiana; green frog, L.
clamitans; northern leopard frog, L. pipiens; wood frog, L. sylvatica) were collected from a field
in the wetlands of Indiana near the campus of Purdue University. Collected egg masses were
raised outdoors in 200 L polyethylene tanks filled with well water. Experiments began when
frogs reached Gosner stage 26, defined as when larvae are free swimming and feeding. Before
A-49
-------
test initiation larvae were acclimated to test conditions (21ฐC and 12-hour: 12-hour light:dark
photoperiod) for 24 hours. A stock solution of PFOA (2,000 mg/L) was made in UV-filtered well
water and diluted with well water to reach test concentrations (ranged from 0-2,000 mg/L
PFOA). Test concentrations were not measured in test solutions based on previous research that
demonstrated limited degradation under similar conditions. Larva were transferred individually
to 250 mL plastic cups with 200 mL of test solution and were not fed during the exposure period.
The number of replicates varied by species and treatment; 30 replicates in the control and five to
20 replicates in each treatment for American bullfrog, 10 replicates for each treatment for green
frog, northern leopard frog and wood frog. No mortality occurred in any of the control groups.
Acute values from the four tests include:
American bullfrog: The author-reported 96-hour LCso was 1,060 mg/L. The EPA-
calculated LCso was 1,035 mg/L (95% C.I. = 1,020 - 1,049 mg/L), which was
acceptable for quantitative use.
Green frog: The author-reported 96-hour LCso was 1,070 mg/L. The EPA was unable to
fit a C-R model with significant parameters and relied on the author reported value as
quantitatively acceptable.
Northern leopard frog: The author-reported 96-hour LCso was 752 mg/L. The EPA-
calculated LCso was 751.7 mg/L (95% C.I. = 713.0 - 790.5 mg/L), which was acceptable
for quantitative use.
Wood frog: The author-reported 96-hour LCso was 999 mg/L. The EPA was unable to
fit a C-R model with significant parameters and relied on the author reported value as
quantitatively acceptable.
A.2.18 Nineteenth most acutely sensitive genus - Oncorhynchus
The acute effects of ammonium perfluorooctanoate (APFO, 99.4% purity) to
Oncorhynchus mykiss were investigated by researchers at the DuPont Haskell Laboratory
(2000). The static measured GLP study exposed 2.8 cm fish for 96-hours (dilution water not
A-50
-------
identified). Rainbow trout used in this study were not fed approximately 29-hours prior to and
during the test. Rainbow trout were assigned to the test chambers using random numbers. The
addition of fish to the test solutions was initiated approximately 41 minutes after test solution
mixing was completed. Mean measured concentrations of ammonium perfluorooctanoate were
554, 1,090, 2,280, 4,560 and 9,360 for the 625, 1,250, 2,500, 5,000 and 10,000 mg/L nominal
dose levels, respectively (measured directly by high performance liquid chromatography/tandem
mass spectrometry). Control solutions showed no detectable concentrations of ammonium
perfluorooctanoate. All test substance solutions were clear and colorless with no insoluble test
substance present during the test. Test chambers were stainless steel aquaria that held
approximately 9 L of test solution. Two replicate test chambers were used per test concentration
with 10 fish in each chamber (total of 20 fish per concentration). Each chamber was covered
with a glass plate to prevent fish from escaping. Mortality and behavioral observations were
made at test start, every 24-hours thereafter, and at test end. All chemical and physical
parameters for the definitive test were within expected ranges. Total alkalinity and EDTA total
hardness of the dilution water control were 49 mg/L CaCC>3 and 122 mg/L CaCC>3, respectively.
During the test, dissolved oxygen concentrations ranged from 7.5-11.2 mg/L, pH ranged from
7.1-7.2, and mean temperature was 11.8ฐC (range 11.6-12.1ฐC). A photoperiod of 16 hours light
(approximately 199-450 lux) and eight hours darkness was employed, which included 30
minutes of transitional light (11-157 lux) preceding and following the 16-hour light interval. The
authors reported a 96-hour LCso of 4,001 mg/L APFO and no mortality or sublethal effects were
observed at concentrations <2,500 mg/L APFO. Unlike Colombo et al. (2008), DuPont Haskell
Laboratory (2000) did not observe any ammonia toxicity resulting from the APFO test salt,
perhaps a result of the relatively lower pH and temperature used by DuPont Haskell Laboratory
A-51
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(2000), which decreased the proportion of un-ionized ammonia. Results of this study were
considered quantitatively acceptable for use in criterion derivation.
A-52
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Appendix B Acceptable Estuarine/Marine Acute PFOA Toxicity Studies
B.l Summary Table of Acceptable Quantitative Estuarine/Marine Acute PFOA Toxicity Studies
Species (lilVsliiiic)
Method'1
1 CM
Dui'iilion
( hcmiciil /
PuriU
pll
Temp.
(ฐC)
Siiliniit
(DDll
I'.ITecl
Author
Kcporlcd
I'.ITecl
Cone,
(inii/l.)
l-'.IW
( iilculiiled
I'.ITecl
(one.
tlll!ป/l.)
l-iiiiil I'-ITccl
Cone.
(mซ/l.)'
Species
Mesin
Acute
Value
(lliu/l->
Reference
Purple sea urchin (embryo),
Strongylocentrotus purpuratus
S,M
96 hours
PFOA
95%
15
30
EC50
(normal
development)
19
20.63
20.63
20.63
Hayman et
al. 2021
Mediterranean mussel (larva),
Mytilus galloprovincialis
S,U
48 hours
PFOA
Unreported
7.9-
8.1
16
36
EC50
(malformation)
>1
-
>lc
-
Fabbri et al.
2014
Mediterranean mussel
(embryo),
Mytilus galloprovincialis
S,M
48 hours
PFOA
95%
-
15
30
EC50
(normal and
surviving)
9.98
17.58
17.58
17.58
Hayman et
al. 2021
Mysid (3-days old),
Americamysis bahia
S,M
96 hours
PFOA
95%
-
20
30
LC50
24
-
24
24
Hayman et
al. 2021
Mysid (neonate, <24 hours),
Siriella armata
S,U
96 hours
PFOA
96%
-
20
-
LC50
15.5
-
15.5
15.5
Mhadhbi et
al. 2012
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b Values in bold used in the SMAV calculation
B-l
-------
B.2 Detailed Study Summaries of Acute Saltwater PFOA Toxicity Studies
Considered for Use in Saltwater Criterion Derivation
The purpose of this section is to present detailed study summaries for acute
estuarine/marine tests that were considered quantitatively acceptable for criterion derivation,
with summaries grouped and ordered by genus sensitivity. Unlike Appendix A.2 and Appendix
C.2, the EPA-calculated C-R models were not presented below for the four most sensitive
estuarine/marine genera because an estuarine/marine criterion was not developed exclusively
based on these empirical data. Rather, an estuarine/marine benchmark was derived using a NAM,
which is further described in Appendix L.
B.2.1 Most acutely sensitive estuarine/marine genera - Siriella (mysid)
Mhadhbi et al. (2012) performed a 96-hour static, unmeasured acute test with PFOA
(96% purity) on the mysid, Siriella armata. A stock solution of PFOA was made either with
filtered sea water from the Ria of Vigo (Iberian Peninsula) for low exposure concentrations, or
with DMSO for high PFOA concentrations (a final maximum DMSO concentration of 0.01%
(v/v) in the test medium). However, the authors do not indicate what is considered a high-test
concentration, so it is unclear which test concentrations actually used DMSO as a solvent. If
DMSO was used, a solvent control was also included. Mysids were exposed to one of ten
nominal PFOA treatments (0.1, 0.5, 1, 2, 5, 10, 20, 30, 40 and 80 mg/L). Mysids were also
collected from the same source as the dilution water and quarantined before use in 100 L plastic
tanks with circulating sand-filtered seawater. The adult stock was fed daily and maintained at
laboratory conditions (17-18ฐC, salinity between 34.4-35.9 ppt, and oxygen 6 mg/L). Twenty
neonates (<24-hours old) were used per each treatment. To prevent cannibalism, a single
individual was added to each glass vial with 2-4 Ml of test solution. Vials were incubated at
20ฐC with a 16-hour light period. Neonates were fed 10-15 Artemia salina nauplii daily and
B-2
-------
mortality was recorded after 96 hours. The 96-hour LC50 reported in the study was 15.5 mg/L
PFOA and was acceptable for quantitative use.
B.2.2 Second most acutely sensitive estuarine/marine genus -Mytilus (mussel)
The acute toxicity of perfluorooctanoic acid (PFOA, purity not provided) on the
Mediterranean mussel, Mytilus galloprovincialis, which occurs in California and other parts of
the Pacific Northwest (Green 2014), was evaluated by Fabbri et al. (2014). Sexually mature
mussels were purchased from an aquaculture farm in the Ligurian Sea (La Spezia, Italy) and held
for two days for gamete collection. Gametes were held in artificial sea water (ASW) made of
analytical grade salts and at a constant temperature of 16 ฑ 1ฐC. It is assumed that the gametes
were held at the same environmental conditions as the adults, so test salinity was assumed to be
36 ppt with a Ph of 7.9-8.1. Embryos were transferred to 96-well microplates with a minimum of
40 embryos/well. Each treatment had six replicates. Embryos were incubated with a 16-hour:8-
hour light:dark photoperiod and exposed to one of six nominal PFOA concentrations (0.00001,
0.0001, 0.001, 0.01, 0.1 and 1 mg/L) or controls. The PFOA stock was made with ethanol, and
ASW control samples run in parallel. This included ethanol at the maximal final concentration of
0.01%. Each experiment was repeated four times. At test termination (48 hours), the endpoint
was the percentage of normal D-larvae in each well, including malformed larvae and pre-D
stages. The acceptability of test results was based on controls for a percentage of normal D-shell
stage larvae of >75% (ASTM 2004). Authors noted that controls had >80% normal D-larvae
across all tests. PFOA was only measured once in one treatment which was similar to the
nominal concentration, 0.000081 mg/L versus the nominal concentration of 0.0001 mg/L. PFOA
was below the limit of detection in the control ASW (0.04 ng/L). The percentage of normal D-
larva decreased with increasing test concentrations. The NOEC and LOEC reported for the study
were 0.00001 and 0.0001 mg/L, respectively. However, the test concentrations failed to elicit
B-3
-------
50% malformations in the highest test concentration, and an EC50 was not determined. Therefore,
the EC50 for the study was greater than the highest test concentration (1 mg/L). The 48-hour EC50
based on malformation of >1 mg/L was acceptable for quantitative use.
Hayman et al. (2021) report the results of a 48-hour static, measured test on the effects
of PFOA (CAS # 335-67-1, 95% purity, purchased from Sigma-Aldrich, St. Louis, MO) on the
Mediterranean mussel, Mytilus galloprovincialis. The authors note that the tests followed U.S.
EPA (1995b) and ASTM (2004) protocols. Mussels were collected in the field (Sand Diego Bay,
CA) and conditioned in a flow-through system at 15ฐC. Mussels were induced to spawn by heat-
shock and approximately 250 embryos (2-cell stage) were added to 20 Ml borosilicate glass
scintillation vials with 10 Ml of test solution. There were five replicates per test concentration.
Test conditions were 30 ppt, 15ฐC and a 16-hour:8-hour light:dark photoperiod. Six test solutions
were made in 0.45 |im filtered seawater (North San Diego Bay, CA) with PFOA dissolved in
methanol. The highest concentration of methanol was 0.02% (v/v). Measured test concentrations
ranged from 1.5-52 mg/L. Controls were made in the same seawater and the acute test also
included a solvent control. At test termination (48 hours), larvae were enumerated for total
number of larvae that were alive at the end of the test (normally or abnormally developed) as
well as number of normally-developed (in the prodissoconch "D-shaped" stage) larvae. There
were no significant differences between solvent control and filtered seawater, suggesting no
adverse effects of methanol. The author reported 48-hour EC50, based on normal survival larvae,
was 9.98 mg/L PFOA. The EPA-calculated 48-hour EC50 value was 17.58 mg/L (95% C.I. =
13.73 -21.43 mg/L), which was acceptable for quantitative use.
B.2.3 Third most acutely sensitive estuarine/marine genus - Stronsvlocentrotus (sea urchin)
Hayman et al. (2021) report the results of a 96-hour static, measured test on the effects
of PFOA (CAS # 335-67-1, 95% purity, purchased from Sigma-Aldrich, St. Louis, MO) on the
B-4
-------
purple sea urchin, Strongylocentrotuspurpuratus. The authors noted that tests followed U.S.
EPA (1995b) and ASTM (2004) protocols. Sea urchins were collected in the field (Sand Diego
Bay, CA) and conditioned in a flow-through system at 15ฐC. They were induced to spawn by
KC1 injection and approximately 250 embryos (2-cell stage) were added to 20 Ml borosilicate
glass scintillation vials with 10 Ml of test solution. There were five replicates per test
concentration. Test conditions were 30 ppt, 15ฐC and a 16-hour:8-hour light:dark photoperiod.
Six test solutions were made in 0.45 |im filtered seawater (North San Diego Bay, CA) with
PFOA dissolved in methanol. The highest concentration of methanol was 0.02% (v/v). Measured
test concentrations ranged from 1.5-52 mg/L. Controls were made in the same seawater and the
acute test also included a solvent control. At test termination (96 hours), the first 100 larvae were
enumerated and observed for normal development (four-arm pluteus stage). There were no
significant differences between solvent control and filtered seawater, suggesting no adverse
effects of methanol. The author reported 96-hour ECso, based on normal development, was 19
mg/L PFOA. The EPA-calculated 96-hour ECso value was 20.63 mg/L (95% C.I. = 19.74 -
21.52 mg/L), which was acceptable for quantitative use.
B.2.4 Fourth most acutely sensitive estuarine/marine genus - Americamysis (mysid)
Hayman et al. (2021) conducted a 96-hour static, measured test to assess effects of
PFOA (CAS # 335-67-1, 95% purity, purchased from Sigma-Aldrich, St. Louis, MO) on the
mysid, Americamysis bahia. The authors noted that tests followed U.S. EPA (2002) protocols.
Mysids were purchased from a commercial supplier (Aquatic Research Organisms, Hampton,
NH) and acclimated to test conditions (30 ppt, 20ฐC and a 16-hour: 8-hour light:dark
photoperiod). Six test solutions were made in 0.45 |im filtered seawater (North San Diego Bay,
CA) with PFOA dissolved in methanol. The highest concentration of methanol was 0.02% (v/v).
Measured test concentrations ranged from 1.1-29 mg/L. The highest test concentration (61.7
B-5
-------
mg/L) was reported as nominal only because the sample was mistakenly not sent to the lab for
verification. Controls were made in the same seawater and the acute test also included a solvent
control. Five mysids (three-days old) were added to 120 Ml polypropylene cups and 100 Ml of
test solution with six replicates per treatment. Living mysids were counted and dead organisms
were removed daily. There were no significant differences between solvent control and filtered
seawater, suggesting no adverse effects of methanol. No organisms were found dead in the
controls at test termination. The EPA was unable to fit a concentration-response model with
significant parameters and relied on the author-reported 96-hour LC50 of 24 mg/L PFOA as the
quantitatively acceptable acute value.
B-6
-------
Appendix C Acceptable Freshwater Chronic PFOA Toxicity Studies
C.l Summary Table of Acceptable Quantitative Freshwater Chronic PFOA Toxicity Studies
Species (lil'cs(;iปc)
Mclhori'1
Tesl
Diimlion
Chcmic;il /
Pu ii( \
Pll
Temp.
(ฐC)
Chronic Y;iluc
l.nripoinl
Aii I hซir
Reported
Chronic
Value
(inii/l.)
r.PA
( iilcnliilcd
Chronic
Value
(mป/l.)
liiiiil
Chronic
Value
(niii/l.)'
Species
Mean
Chronic
\ iilne
(mป/l.)
Reference
Rotifer
(<2-hours old neonates),
Brachionus calyciflorus
R,Ub
Up to
200
hours
PFOA
96%
20
ECio
(intrinsic rate of natural
increase)
0.3536
0.5015
0.5015
-
Zhang et al.
2013a
Rotifer
(<2-hours old neonates),
Brachionus calyciflorus
R,Ub
4 days
PFOA
96%
-
20
ECio
(intrinsic rate of natural
increase)
2.828
1.166
1.166
0.7647
Zhang et al.
2014b
Cladoceran (<8 hours),
Ceriodaphnia dubia
R, M
7 days
PFOA
99.5%
7.72
(7.6-
7.78)
24.7
(24.0-
25.2)
ECio
(neonates/female)
26.2
20.42
20.42
-
Kadlec et al.
2024
Cladoceran (<8 hours),
Ceriodaphnia dubia
R, M
7 days
PFOA
99.5%
7.58
(7.47-
7.67)
24.5
(24.0-
25.2)
ECio
(neonates/female)
25.1
21.69
21.69
-
Kadlec et al.
2024
Cladoceran (<8 hours),
Ceriodaphnia dubia
R, M
7 days
PFOA
99.5%
7.64
(7.59-
7.69)
24.9
(23.9-
25.8)
ECio
(neonates/female)
33.8
29.54
29.54
23.56
Kadlec et al.
2024
Cladoceran (6-12 hours old),
Daphnia carinata
R, U
21 days
PFOA
95%
-
21
MATC
(average # of offspring per
brood and total # of living
offspring)
0.03162
-
0.03162
0.03162
Logeshwaran
et al. 2021
Cladoceran (STRAUS-clone
5; 6-24 hours old),
Daphnia magna
R, M
21 days
APFO
99.7%
7.56-
8.26
18-22
ECio
(average # offspring per
starting adult)
29.73
20.26
20.26d
-
Centre
International
de Toxicologie
2003;
Colombo et al.
2008
Cladoceran,
Daphnia magna
R, U
21 days
PFOA
Unreported
-
21
ECio
(# young/starting female)
17.68
7.853
7.853
-
Ji et al. 2008
Cladoceran (<24 hours old),
Daphnia magna
R, U
21 days
APFO
>98%
-
20
ECio
(# young/starting female)
17.89
12.89
12.89
-
Li 2010
Cladoceran (<24 hours old),
Daphnia magna
R, M
21 days
PFOA
99%
7
22
ECio
(survival)
7.028
5.458
5.458
-
Yang et al.
2014
Cladoceran (<24 hours old),
Daphnia magna
S,U
21 days
PFOA
98%
-
20
MATC
(growth and reproduction)
0.07155
-
0.07155
-
Lu et al. 2016
C-l
-------
Species (lil'cs(;iปc)
Mclhori'1
Tesl
Diimlion
Chcmic;il /
Pu ii( \
Pll
Temp.
(ฐC)
Chronic Y;iluc
l.nripoinl
Aii I hซir
Reported
Chronic
Value
(inii/l.)
r.PA
( iilcnliilcd
Chronic
Value
(mป/l.)
liiiiil
Chronic
Value
(niii/l.)'
Species
Mean
Chronic
\ iilne
(mป/l.)
Reference
( ladoccran (12-24 hours old),
Daphnia magna
R, U
21 days
PFOA
Unreported
6-8.5
20
ECio
(# of offspring)
8.23 lf
8.084f
8.084f
4.317
Yang et al.
2019
Cladoceran (<24 hours old),
Moina macrocopa
R, U
7 days
PFOA
Unreported
-
25
ECio
(# young/starting female)
4.419
2.194
2.194
2.194
Ji et al. 2008
Amphipod (2-9 days old),
Hyalella azteca
R, M
42 days
PFOA
96%
8.1
25
ECio
(# of juveniles/female)
0.0265
0.147
0.147
0.147
Bartlett et al.
2021
Midge (2-day old larvae),
Chironomus dilutus
R, M
19 days
PFOA
97%
00
20.0-
24.0
ECio
(survival)
89.8
88.32
88.32
88.32
McCarthy et
al. 2021
Mayfly (<24 hr larva),
Neocloeon triangulifer
R, M
23 days
PFOA
95%
-
23
NOEC
(survival, weight, emergence)
>3.085
-
>3.085
>3.085
Soucek et al.
2023
Rainbow trout
(embryo-larval-juvenile),
Oncorhynchus mykiss
F, M
85 days
(ELS)
APFO
99.7%
6.0-
8.5
11.1-
14.4
LOEC
(growth and mortality)
>40
-
>40d
>40
Centre
International
de Toxicologie
2004;
Colombo et al.
2008
Rare minnow (adult),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
25
LOEC
(survival)
>30
-
>30
>30
Wei et al. 2007
Fathead minnow (<18 hpf),
Pimephales promelas
R, M
21 days
PFOA
96%
7.4-
7.8
25
LOEC
(mortality and growth)
>76
-
>76
>76
Bartlett et al.
2021
Medaka
(adult-FO, embryo-Fl, F2),
Oryzias latipes
F, M
259 days6
PFOA
Unreported
7.5
25
MATC
(F2: sac-fry survival; FO, Fl,
F2: fecundity)
9.487
-
9.487
9.487
Lee et al. 2017
American bullfrog
(tadpole, Gosner stage 25),
Lithobates catesbeiana
(formerly, Rana catesbeiana)
R, U
72 days
PFOA
Unreported
-
21
LOEC
(snout vent length)
0.288
-
0.288
0.288
Flynn et al.
2019
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b Chemical concentrations made in a side-test representative of exposure and verified stability of concentrations of PFOA in the range of concentrations tested under similar
conditions. Daily renewal of test solutions.
C-2
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0 Values in bold used in SMCV calculation.
d Concentration of APFO in publication indirectly determined through quantification of the anion (PFO-).
6 Total exposure period across FO, Fl, and F2 generations.
f
Reported in moles, converted to grams based on a molecular weight of 414.07 g/mol.
g Value represents an ECio based on reproduction.
C-3
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.2 Detailed PFOA Chronic Toxicity Study Summaries and Corresponding
Concentration-Response Curves (when calculated)
The purpose of this section is to present detailed study summaries for tests that were
considered quantitatively acceptable for criteria derivation, with summaries grouped and ordered
by genus sensitivity. C-R models developed by the EPA that were used to determine chronic
toxicity values used for criterion derivation are also presented. C-R models included here with
study summaries were those for the four most sensitive genera. In many cases, authors did not
report concentration-response data in the publication/supplemental materials and/or did not
provide concentration-response data upon request by the EPA. In such cases, the EPA did not
independently calculate toxicity values and the author-reported effect concentrations were used
to derive the criterion.
C.2.1 Most chronically sensitive genus - Hyalella
Bartlett et al. (2021) evaluated the chronic effects of PFOA (CAS# 335-67-1, 96%
purity, solubility in water at 20,000 mg/L, purchased from Sigma-Aldrich) on Hyalella azteca
via a 42-day static-renewal, measured study. Methods for this study were adapted from
Borgmann et al. (2007), and organisms were two to nine days old at the test initiation.
Experiments were conducted in standard artificial media with water quality characteristics of 52
to 60 mg/L alkalinity as CaCC>3, average specific conductivity of 0.41 Ms/cm, dissolved oxygen
of 5.2 to 8.8 mg/L, total hardness of 120 to 140 mg/L CaCCb, average Ph of 8.1 and average
temperature of 25ฐC. A 100 mg/L stock solution was prepared to yield measured test
concentrations of 0 (control), 0.84, 3.3, 8.9, 29 and 97 mg/L PFOA. Two separate tests were
performed with five replicates per concentration and 20 amphipods per replicate in 2-L HDPE
containers filled with 1 L of testing solution, 2.5 mg of ground TetraMin and one piece of 5x5
cm cotton gauze. Test organisms were fed 2.5 mg TetraMin three times a week during weeks one
C-4
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and two, 5 mg TetraMin three times a week during weeks three and four, and 5 mg TetraMin five
times a week during weeks five and six. At test termination (day 42), adults were sexed and
weighed, as well as their young counted. The 42-day author-reported LCio value for survival was
23.2 mg/L PFOA. The author-reported ECio values for growth and reproduction were 0.160
mg/L and 0.0265 mg/L, respectively. The EPA only performed C-R analysis for the growth and
reproduction-based endpoints for this test, given the apparent tolerance of the survival-based
endpoint. The EPA calculated ECio values for the 42-day growth endpoint (i.e., control
normalized wet weight/amphipod) and the 42-day reproduction endpoint (i.e., number of
juveniles per female). The 42-day growth-based ECio of 0.488 mg/L (95% C.I. = 0.319 - 0.657
mg/L) was not selected as the primary endpoint from this test because it was more tolerant than
the reproduction-based ECio of 0.147 mg/L (95% C.I. = 0.147 - 0.147 mg/L). The EPA-
calculated ECio was 0.147 mg/L with a corresponding ECso of 0.911 mg/L. While the ECio was
relatively uncertain due to a lack of partial effects around the 10% effect level, the EC so estimate
remained relatively certain given the 47% effect observed in the lowest treatment concentration.
ECso to ECio ratios from all quantitatively acceptable chronic concentration-response curves with
similar species (i.e., small members of the subphylum Crustacea) and endpoints (i.e.,
offspring/female) were evaluated to understand the variability in the EC50:EC 10 ratios and
provide further context to the reasonableness of the H. azteca ECio estimate. Overall, three
quantitatively acceptable chronic concentration-response curves with similar species/endpoints
were available. See Table C-l below for a description of the individual C-R curves and resultant
ECsoiECio ratios from each C-R curve.
C-5
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Table C-l. ECso to ECio ratios from all quantitatively acceptable chronic concentration-
response curves with species similar to H. azteca (i.e., small members of the subphylum
Crustacea) and with endpoints that were based on reproduction per female.
i:iปa-
i:iปa-
(itlculitlecl
(itlculitlecl
IX 5:i:c ...
C'iliilion
Species
Knilpoinl
KC'ซii (nig/I.)
IX'in (nig/I.)
Ritlio
Ji et al. 2008
Daphnia
magna
(# young/starting
female)
61.67
7.853
7.853
Li 2008
Daphnia
magna
(# young/starting
female)
40.75
12.89
3.161
Ji et al. 2008
Moina
macrocopa
(# young/starting
female)
12.77
2.194
5.819
ECsoiECio ratios from the three tests with similar species/endpoints ranged from 3.161 to
7.852 with a geometric mean ration of 5.247. Dividing the H. azteca reproduction-based ECso
(i.e., 0.911 mg/L) by the geometric mean ECso:ECio ratio (i.e., 5.247) produced an estimated H.
azteca ECio of 0.174 mg/L, which was similar to ECio value calculated directly from the H.
azteca C-R curve (i.e., 0.147 mg/L). The ECio value calculated directly from the H. azteca C-R
dataset was, therefore, hypothesized to provide a robust estimate of a 10% reproductive-based
effect concentration despite the lack of partial low-level effects observed along the C-R curve.
The 42-day average number of young per female ECio value of 0.147 mg/L (95% C.I. = 0.147 -
0.147 mg/L) calculated from C-R data reported by Bartlett et al. 2021 was retained for
quantitative use.
C-6
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Publication: Bartlett et al. (2021)
Species: Hyalella azteca
Genus: Hyalella
EPA-Calculated ECio: 0.147 mg/L (95% C.I. = 0.147 - 0.147 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std.Error
t-stat
p-value
b
1.0325
2.1536 e6
479432
1.328 e6
d
1.7400
1.3029 e7
1335511
4.767 e7
e
1.2996
2.5393 e6
511775
1.244 e6
Concentration-Response Model Fit:
Bartlett et al. 2021
Hyalella azteca
Weibull type 1, 3 para
le-03 le-02 le-01 le+00 lefOl
PFOA ( mg/L )
C.2.2 Second most chronically sensitive genus - Lithobates
Flynn et al. (2019) evaluated the chronic effects of PFOA (CAS# 335-67-1, purchased
from Sigma-Aldrich) on the American bullfrog (.Lithobates catesbeiana, formerly, Rana
catesbeiana) during a 72-day static-renewal unmeasured exposure. Testing followed Purdue
University's Institutional Animal Care and Use Committee Guidelines Protocol #16010013551.
American bullfrog eggs were taken from a permanent pond in the Martell Forest outside of West
C-7
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Lafayette, Indiana. The eggs from a single egg mass were acclimated in 100 L outdoor tanks
filled with 70 L of aged well water and covered with a 70% shade cloth. Once hatched, tadpoles
were fed rabbit chow and TetraMin ad libitum and were acclimated to laboratory conditions for
24 hours before testing (21ฐC and a 12-hour: 12-hour light:dark photoperiod). A 2,000 mg/L
PFOA stock solution was prepared with RO water to produce three concentrations for the
chronic test (0, 0.144 and 0.288 mg/L). Each chronic test treatment contained 10 tadpoles
(Gosner stage 25), replicated four times, in 15-L plastic tubs filled with 10 L of aged UV-
irradiated, filtered well water. Complete water changes were performed every three to four days,
at which time chemical treatments were reapplied. Each experimental unit was fed daily at a
constant rate (10% per capita) based on tadpole wet biomass in the control treatment to assure
that food was not limiting. On day 72 of the experiment, all tadpoles were euthanized, measured
(snout vent length and mass) and staged. The most sensitive chronic endpoint was growth (snout-
vent length), with a 72-day NOEC and LOEC of 0.144 mg/L and 0.288 mg/L, respectively. The
EPA could not independently calculate an ECio value because there were minimal effects
observed across the limited number of treatment concentrations tested. Consequently, the EPA
used the LOEC of 0.288 mg/L as the chronic value from this chronic test. The LOEC was used
preferentially to the MATC from this test because a -7% reduction in snout-vent length relative
to control responses was observed at the LOEC (i.e., 0.288 mg/L).
Publication: Flynn et al. (2019)
Species: American bullfrog (Lithobates catesbeiana)
Genus: Lithobates
EPA-Calculated ECio: Not calculable, unable to fit a model with significant parameters
Concentration-Response Model Fit: Not Applicable
C-8
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C.2.3 Third most chronically sensitive genus - Daphnia
Logeshwaran et al. (2021) conducted a PFOA (95% purity, purchased from Sigma-
Aldrich Australia) chronic toxicity test with the cladoceran, Daphnia carinata. In-house cultures
of daphnids were maintained in 2 L glass bottles with 30% natural spring water in deionized
water, 21ฐC and a 16-hour: 8-hour light:dark photoperiod. The chronic test protocol followed
OECD guidelines (2012). A PFOA stock solution (100 mg/L) was prepared in deionized water.
Cladoceran culture medium was used to prepare the PFOA stock and test solutions. One daphnid
(6-12 hours old) was transferred to each 100 Ml polypropylene container containing 50 Ml of the
nominal test solution (0, 0.001, 0.01, 0.1, 1.0 and 10 mg/L PFOA). Each test treatment was
replicated 10 times with test solutions renewed and daphnids fed daily. At test termination (21
days) test endpoints included survival, days to first brood, average offspring in each brood and
total live offspring. No mortality occurred in the controls or lowest test concentration. Of the
three endpoints measured, average offspring in each brood and total live offspring were the more
sensitive endpoints with 21-day NOEC and LOEC values of 0.01 and 0.1 mg/L PFOA,
respectively. The EPA was unable to calculate statistically robust ECio estimates from C-R
models for these endpoints, largely because of the 10X dilution series across five orders of
magnitude. The LOECs for these endpoints were not selected as the chronic value because the
LOECs produced a 29.23% reduction in the average number of offspring per brood relative to
controls and a 39.89% reduction in the total living offspring relative to controls. Therefore, the
MATC (i.e., 0.03162 mg/L) was selected as the quantitatively acceptable chronic value form this
test.
Publication: Logeshwaran et al. (2021)
Species: Cladoceran (Daphnia carinata)
Genus: Daphnia
EPA-Calculated ECio: Not used, unable to fit a statistically robust model
Concentration-Response Model Fit: Not Applicable
C-9
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Ji et al. (2008) conducted a chronic life-cycle test on the effects of PFOA (CAS # 335-
67-1, purity was not reported; obtained from Sigma Aldrich, St. Louis, MO, USA) with Daphnia
magna. The test was done under renewal conditions over a 21-day period and test solutions were
not analytically confirmed. Authors stated that the D. magna test followed OECD 211 (1998). I),
magna used for testing were obtained from brood stock cultured at the Environmental
Toxicology Laboratory at Seoul National University, Korea. Test organisms were less than 24-
hours old at test initiation. Dilution water was moderately hard reconstituted water (total
hardness typically 80-100 mg/L as CaCCb). Experiments were conducted in glass jars of
unspecified size and fill volume. Photoperiod was assumed to be 16-hours of illumination, the
same conditions as the daphnid cultures used as the source of the experimental organisms.
Preparation of test solutions was not described. The test involved 10 replicates of one daphnid
each in five nominal test concentrations plus a negative control. Nominal concentrations were 0
(negative control), 3.125, 6.25, 12.5, 25 and 50 mg/L and test solutions were renewed three times
per week. Test temperature was 21 ฑ 1ฐC for the D. magna test. Authors note that the water
quality parameters (pH, temperature, conductivity, and dissolved oxygen) were measured after
changing the medium, but the information is not reported. Survival of daphnids in the negative
control was 100%. The most sensitive endpoint for D. magna reported in the publication was
days to first brood with a 21-day NOEC of 6.25 mg/L (LOEC = 12.5 mg/L; MATC = 8.839
mg/L); however, number of young per starting female (an endpoint not reported in the
publication, which only assessed number of young per surviving female) was calculated by the
EPA and considered to be a more sensitive endpoint with an EPA-calculated ECio of 7.853 mg/L
C-10
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(95% C.I. = 4.253 - 11.45 mg/L). Therefore, the EPA-calculated ECio of 7.853 mg/L PFOA for
D. magna (number of young per starting female) was considered quantitatively acceptable.
Publication: Ji et al. (2008)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated ECio: 7.853 mg/L (95% C.I. = 4.253 - 11.45 mg/L)
Concentration-Response Model Estimates:
Parameter
Kslimale
Sul. Krror
1-slal
p-value
b
-0.4609
0.0633
-7.2762
0.0054
d
83.0500
3.3807
24.5662
0.0001
e
37.5761
6.5642
5.7244
0.0106
Concentration-Response Model Fit:
Ji et al. 200S
Daphnia magna
WeibiiB type 2, 3 para
0
3 7(S'
1
2 ง0'
m
c
3
Gs
งปฆ
2
40'
\.
\
S
\
\
\
\
0.! 1.0 10.0
PFOA ( mg/L )
Li (2010) conducted an unmeasured chronic life cycle 21-day test on the effects of PFOA
(ammonium salt, >98%) purity) on Daphnia magna. The authors stated that the test followed
OECD 211 (1998). D. magna used for the test were maintained in the laboratory for more than
one year and were less than 24-hours old at test initiation. Dilution water was distilled water with
C-ll
-------
ASTM medium salts added (0.12 g/L CaS04.2H20, 0.12 g/L MgS04, 0.192 g/L NaHC03, and
0.008 g/L KC1). The calculated total hardness was 169 mg/L as CaCCb. The photoperiod had 16-
hours of illumination with an unreported light intensity. A primary stock solution (1,000 mg/L)
was prepared in ASTM medium. Exposure vessels were 50 Ml polypropylene culture tubes with
50 Ml fill volume. The test involved 10 replicates of one daphnid each in five nominal test
concentrations plus a negative control and each test was repeated three times. Nominal
concentrations were 0 (negative control), 1, 3.2, 10, 32 and 100 mg/L. Test temperature was
maintained at 20 ฑ 1ฐC. Water quality parameters measured in test solutions were not reported.
Survival of daphnids in the negative control was 96.7% across all three tests. The D. magna 21-
day NOEC (reproduction as number of young per female, broods per female, and mean brood
size) was 10 mg/L (LOEC = 32 mg/L; calculated MATC = 17.89 mg/L). The EPA performed C-
R analysis for each reported endpoint. The EPA also revaluated all endpoints that were based on
number of surviving females to be based on the number of starting females. This recalculation
was done with the intent to account for starting females that were unable to contribute to the
population as reproduction/female due to mortality. The most sensitive endpoint with an
acceptable C-R curve was the number of young per starting female with an EPA-calculated ECio
of 12.89 mg/L PFOA (95% C.I. = 8.292 - 17.49 mg/L) and was acceptable for quantitative use.
C-12
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Publication: Li (2010)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated ECio: 12.89 mg/L (95% C.I. = 8.292 - 17.49 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.2765
0.2577
4.9540
0.0158
d
145.8089
3.0147
48.3655
1.946 e5
e
57.4122
8.7260
6.5795
0.0071
Concentration-Response Model Fit:
Li 2010
Daphnia magna
Weibutl type 1, 3 para
PFOA (mg''L )
Yang et al. (2014) evaluated the chronic 21-day renewal, measured test of PFOA (CAS #
335-67-1, 99% purity) with Daphnia magna, following ASTM E729 (1993). Daphnids used for
the test were donated by the Chinese Research Academy of Environmental Sciences. The
daphnids were less than 24-hours old at test initiation. Dilution water was dechlorinated tap
water (Ph, 7.0 ฑ 0.5; dissolved oxygen, 7.0 ฑ 0.5 mg/L; total organic carbon, 0.02 mg/L; and
total hardness, 190.0 ฑ0.1 mg/L as CaCCb). The photoperiod consisted of 12-hours of
illumination at an unreported intensity. A primary stock solution was prepared by dissolving
C-13
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PFOA in deionized water and DMSO solvent, and proportionally diluted with dilution water to
prepare the test concentrations. Exposure vessels were 200 Ml beakers of unreported material
type containing 100 Ml of test solution. The test employed ten replicates of one daphnid each in
six test concentrations plus a negative and solvent control. Nominal concentrations were 0
(negative and solvent controls), 5, 7.5, 11.25, 16.88, 25.31 and 37.97 mg/L and were renewed at
48-hour intervals. Test concentrations were measured in low and high treatments only. The
authors provided mean measured concentrations before and after renewal: 4.96 and 4.49 mg/L
(lowest concentration) and 37.66 and 32.88 mg/L (highest concentration). Analyses of test
solutions were performed using HPLC/MS and negative electrospray ionization. The
concentration of PFOA was calculated from standard curves (linear in the concentration range of
1-800 ng/Ml), and the average extraction efficiency was in the range of 70-83%. The
concentrations and chromatographic peak areas exhibited a significant positive correlation (r =
0.9987, p < 0.01), and the water sample-spiked recovery was 99%. The temperature, DO, and Ph
were reported as having been measured every day during the test, but results are not provided.
Negative control and solvent control survival were 90% and 100%, respectively. The author-
reported D. magna 21-day ECio for reproduction (total number of spawning) was 7.02 mg/L. The
EPA performed C-R analysis for each reported endpoint. Both chronic survival and reproduction
endpoints resulted in acceptable C-R curves. The EPA-calculated ECio for reproduction as total
number of spawning events was 6.922 mg/L (95% C.I. = 4.865 - 8.979 mg/L), similar to the
ECio reported by the authors (i.e., 7.02 mg/L). Chronic survival was more sensitive than
reproduction, with an EPA-calculated ECio of 5.458 mg/L PFOA (95% C.I. = 3.172 - 7.743
mg/L). Therefore, the survival based ECio calculated by the EPA (i.e., 5.458 mg/L) was
acceptable for quantitative use.
C-14
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Publication: Yang et al. (2014)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated ECio: 5.458 mg/L (95% C.I. = 3.172 - 7.743 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.2765
0.2577
4.9540
0.0158
d
145.8089
3.0147
48.3655
1.946 e5
0.0158
57.4122
8.7260
6.5795
0.0071
Concentration-Response Model Fit:
Yang et al. 2014
Daphnia magna
Weibull type 2, 3 para
PFOA ( mgfL )
Centre International de Toxicologic (2003) and Colombo et al. (2008) conducted a 21-
day renewal measured chronic test on ammonium perfluorooctanoate (APFO CAS # 3825-26-1,
99.7% purity) with the daphnid, Daphnia magna. The authors stated that the toxicity test was
conducted followed OECD test guideline 211. There were 10 replicates for each test treatment
containing one neonate, six to 24-hours old, each. Exposure vessel material and size were not
reported but filled with 50 Ml of test solution. Stock solutions of APFO were prepared by
dissolving the test substance directly in M4 media. The stock was diluted with M4 media to
C-15
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make the nominal test concentrations: control, 6.25, 12.5, 25, 50 and 100 mg/L. Test solutions
were analyzed by ion chromatography with electrochemical detection. Measured concentrations
were 60% saturation and
temperature were maintained between 18-22ฐC. Illumination included 16-hours of light with an
unreported intensity. Test solutions were typically renewed every three days and daphnids were
fed daily. Control survival met the minimum survival guidance (80%). Average number of live
young was the most sensitive endpoint reported by Colombo et al. (2008), with a NOEC of 20
mg/L. Based on the author-reported EC so for the average number of live young, the LOEC was
44.2 mg/L and the MATC was 29.73 mg/L. The EPA performed C-R analysis for each reported
endpoint. The most sensitive endpoint with an acceptable C-R curve was the average number of
live young, with an EPA-calculated ECio of 20.26 mg/L PFOA (95% C.I. = -75.48 - 116.0
mg/L), which was acceptable for quantitative use. Although the D. magna C-R curve from
Colombo et al. (2008) displayed relatively wide 95% confidence bands, the C-R curve was
retained for use because the ECio is just beyond the NOEC, where effects quickly increase from
0% to nearly 100%. Although there is a lack of partial effects, there appears to be a threshold
effect that occurs above the NOEC. Note, Colombo et al. (2008) reported treatment-mean C-R
data for the number of offspring per starting adult. In contrast, Centre International de
Toxicologie (2004) reported replicate-level C-R data for number of offspring per surviving adult
(producing an EPA-calculated ECio of 21.25 mg/L; Model = Weibull Type 1, 3 parameters).
Basing the number of offspring on starting adult (as reported by Colombo et al. [2008]) rather
than per surviving adult (as reported by Centre International de Toxicologie [2003]), produces a
more sensitive endpoint that accounts for those parents that could not contribute to the overall
offspring count as a result of their own mortality. Because Colombo et al. 2008 and Centre
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International de Toxicologic (2003) both reported only treatment-mean C-R survivorship data,
the number of offspring per starting adult was based on treatment mean data rather than
replicate-level data.
The authors note that the contribution of ammonia from APFO exposure in this test
indicates that un-ionized ammonia could be a potential contributor to the observed chronic
toxicity of APFO. The EPA believes ammonia does not contribute substantively to the chronic
toxicity to D. magna in this test based on the following rationale. US EPA (2013) derived a
GMCV of 41.46 mg N/L at pH=7 and temperature of 20ฐC for D. magna. Using the EPA
equations, this translates to 12.6 mg N/L at the authors' assumed acute test pH=8.2 (from their
Table 7) and the midrange reported test temperature (20ฐC). This in turn translates to an un-
ionized ammonia concentration of approximately 0.91 mg un-ionized ammonia/L at this pH and
temperature. Table 7 of Colombo et al. (2008) lists un-ionized ammonia concentration at the
APFO 21-day reproductive NOEC as 0.048 mg un-ionized ammonia/L, which is almost twenty
times lower than the EPA's SMAV for D. magna re-expressed as an un-ionized ammonia
concentration for the test condition. Therefore, it is highly unlikely that ammonia contributed to
the chronic toxicity of APFO to D. Magna in this test, and therefore, the EPA-calculated ECio of
20.26 mg/L APFO is used in the calculation of the GMCV for this species.
Publication: Colombo et al. (2008)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated ECio: 20.26 mg/L (95% C.I. = -75.48 - 116.0 mg/L)
Concentration-Response Model Estimates:
Parameter
Kslimale
Sul. Krror
1-slal
p-value
b
2.79764
0.39868
7.0172
0.005944
d
65.85398
1.79847
36.6167
4.48 e5
e
45.28323
4.01408
11.2811
0.001494
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Concentration-Response Model Fit:
Colombo et al. 2008
Daphnia magna
Weibull type 1, 3 para
so-
licit 1 10 100
PFOA(mg'X)
Lu et al. (2016) evaluated the chronic toxicity of PFOA (CAS# 335-67-1, 98% purity,
purchased from Tokyo Chemical Industry Co., Ltd., Tokyo, Japan) on Daphnia magna
immobilization, growth and reproduction. Reconstituted daphnia culture media was used for both
culturing and test solution preparation as described in OECD Test Guideline 202. D. magna
cultures (originally obtained from the Chinese Center for Disease Control and Prevention
(Beijing, China) were fed with the green algae Scenedesmus obliquus daily, maintained at 20ฐC
and a light/dark photoperiod of 16-hours/8-hours and the medium renewed three times weekly.
The 21-day chronic test endpoints were assessed by a semi-static unmeasured test according to
OECD Test Method 211. Neonates (<24-hours old) were exposed to six concentrations of PFOA
[0 (control), 0.032, 0.16, 0.8, 4 and 20 mg/L] maintained at 20 ฑ 1ฐC. One test organism was
exposed in a 100 mL glass beaker filled with 45 mL of test solution, and there were 20 replicates
for each exposure concentration. The daphnids were fed 1 x 106 cells of Scenedesmus obliquus per
animal per day, and the test solution was renewed every other day. Survival, growth and
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reproduction (fecundity) was determined during the 21-day exposure. The 21-day growth and
reproductive NOEC and LOEC were 0.032 and 0.16 mg/L PFOA, respectively. The EPA was
unable to fit a C-R model with significant parameters to the chronic data associated with
reproduction from this test. The EPA-calculated ECio values for mean intrinsic rate of increase ฎ
and growth (as length) were 0.0173 mg/L (95% C.I. = 0.0170 - 0.0177 mg/L) and 0.0124 mg/L
(95% C.I. = 0.0048 - 0.0200 mg/L), respectively. Both ECio values were nearly two times lower
than the NOEC of 0.032 mg/L and four times lower than the LOEC value (i.e., 0.16 mg/L) where
only 15.2%) and 11.9% reductions in intrinsic rate of natural increase (r) and length were
observed, respectively. As a result, the MATC of 0.07155 mg/L for growth and reproduction was
selected as the most appropriate chronic value for quantitative use to in deriving the chronic
water column-based criterion.
Publication: Lu et al. (2016)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated ECio: Not used, unable to fit a statistically robust model
Concentration-Response Model Fit: Not Applicable
Yang et al. (2019) evaluated the chronic effects of PFOA (CAS# 335-67-1, purchased
from Sigma-Aldrich in St. Louis, MO) on Daphnia magna via a 21-day unmeasured, static-
renewal test that assessed reproductive effects. D. magna cultures were originally obtained from
the Institute of Hydrobiology of Chinese Academy of Science in Wuhan, China. Organisms were
cultured in Daphnia Culture Medium according to the parameters specified in OECD Guideline
202. Protocol for all testing followed OECD Guideline 211. Cladocerans were cultured in
artificial freshwater maintained at 20 ฑ 1ฐC under a 16-hour: 8-hour light:dark photoperiod and a
light intensity of 1,000-1,500 lux at the surface of the water. Cultures were fed Scenedesmus
obliquus daily and the water was changed twice weekly. Reported water quality parameters
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include total hardness of 140-250 mg/L as CaCCb and pH of 6-8.5. The 21-day chronic study had
nominal concentrations of 0 (control), 0.0000162, 0.0000244, 0.0000365 and 0.0000546 mol/L
(or 0 (control), 6.708, 10.10, 15.11 and 22.61 mg/L, respectively, given the molecular weight of
the form of PFOA used in the study, CAS # 335-67-1, of 414.07 g/mol). One neonate (12-24
hours old) was placed in A 100 mL glass beaker, replicated 10 times, and each container filled
wiTH 80 mL of test solution maintained at 20 ฑ 1ฐC and a 16-hour: 8-hour light: dark photoperiod
with a light intensity of 1,000-1,500 lux. D. magna were fed S. obliquus and test solutions were
renewed every 72 hours. Test organisms were counted daily, with any young removed. The
reproductive NOEC and LOEC were 0.0000162 and 0.0000244 mol/L, or 6.708 and 10.10 mg/L
PFOA, respectively. The EPA performed C-R analysis for the test. The EPA-calculated ECio
based on mean offspring at 21-days as a proportion of the control response was 8.084 mg/L
(95% C.I. = 7.830 - 8.334 mg/L) and was used quantitatively to derive the chronic water column
criterion.
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Publication: Yang et al. (2019)
Species: Cladoceran {Daphnia magna)
Genus: Daphnia
EPA-Calculated ECio: 8.084 mg/L (95% C.I. = 7.830 - 8.334 mg/L)
Concentration-Response Model Estimates:
Parameter
Kslimale
Sul. Krror
1-slal
p-value
b
-0.9632
0.1065
-9.0420
0.0029
e
19.2161
0.9269
20.7320
0.0002
Concentration-Response Model Fit:
Yang et al. 2019
Daphnia magna
Weibull type 2, 2 para
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were cultured in an artificial inorganic medium at 20ฐC (16-hours:8-hours, light:dark; 3,000 lux)
for more than six months before toxicity testing to acclimate to the experimental conditions.
Culture medium was an artificial inorganic medium and all toxicity tests were carried out in the
same culture medium and under the same conditions as during culture (i.e., pH, temperature,
illumination). Solvent-free stock solutions of PFOA (1,000 mg/L) were prepared by dissolving
the solid in deionized water via sonication. After mixing, the primary stock was proportionally
diluted with dilution water to prepare the test concentrations. Exposures were carried out in 24-
well cell culture plates (assumed plastic) containing 2 mL of test solution per cell. The test
employed four measured test concentrations plus a negative control. Each treatment consisted of
one replicate plate of 15 rotifers, with one rotifer per cell. Treatments were repeated six times.
Nominal concentrations were 0 (negative control), 0.25, 0.5, 1.0 and 2.0 mg/L. PFOA
concentrations were not measured in the rotifer exposures, but rather, in a side experiment using
HPLC/MS. The side experiment showed that the concentration of PFOA measured every 8 hours
over a 24-hour period in rotifer medium with green algae incurs minimal change in the
concentration range from 0.25 to 2.0 mg/L. One hundred percent survival was observed at 24
hours in the negative control in the corresponding acute test, but survival information is not
provided for the life-cycle test. Zhang et al. (2013 a) demonstrated rotifer body size and mictic
ratio after 28-days were relatively tolerant endpoints with reported NOECs of > 1.0 mg/L and 2.0
mg/L, respectively. The EPA performed C-R analysis for the remaining reported endpoints from
C-R data reported in the publication. The most sensitive endpoint with an acceptable C-R curve
was the intrinsic rate of natural increase with an EPA-calculated ECio of 0.5015 mg/L PFOA
(95% C.I. = 0.1458 - 0.8572 mg/L), which was acceptable for quantitative use. The intrinsic rate
of natural increase (d"1) is a population level endpoint that accounts for births and deaths over
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time. In this study, the intrinsic rate of natural increase was defined as the natural log of the
lifetime net reproductive rate for all individuals within a population (defined here as a PFOA
treatment level) divided by the average generation time of those individuals. The effect
associated with intrinsic rate of natural increase is similar to other chronic apical effects reported
by Zhang et al. (2013a). For example, Zhang et al. (2013a) also reported net reproductive rate
and juvenile period which produced an EPA-calculated ECio value of 0.514 mg/L (95% C.I. =
0.1958 - 0.8329 mg/L). Zhang et al. (2013a) also reported effects to average juvenile period,
which was a relatively tolerant endpoint. Juvenile period decreased with increasing exposure
concentration, with the average juvenile period being about 16% faster than the control responses
in the highest treatment concentration (2.0 mg/L; the EPA was unable to fit a statistically-robust
C-R model for this endpoint). Zhang et al. (2013a) reported significant reductions in egg size
with an EPA-calculated ECio = 0.193 (95% C.I. = -0.1606 - 0.5466 mg/L); however, this
endpoint displayed a relatively poor concentration response relationship and may not be relevant
for assessing population level effects and was, therefore, not selected as the primary effect
concentration from this test. Effects to chronic apical endpoints in this publication and Zhang et
al. (2014b) generally appear as a threshold effect from 0.25 mg/L to 1.0 mg/L, providing further
support for the endpoint and effect level selected for quantitative use from Zhang et al. (2013a).
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Publication: Zhang et al. (2013a)
Species: Rotifer (Brachionus calyciflorus)
Genus: Brachionus
EPA-Calculated ECio: 0.5015 mg/L (95% C.I. = 0.1458 - 0.8572 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
-0.6515
0.1231
-5.2930
0.0339
d
0.5058
0.0144
35.0478
0.0008
e
1.8042
0.2240
8.0546
0.0151
Concentration-Response Model Fit:
Zhang et al. 2013
Brachionus calyciflorus
Weibull type 2, 3 para
PFOA ( mg'L )
Zhang et al. (2014b) reports the results of a similar chronic life-cycle test of PFOA
(CAS # 335-67-1, 96% purity) with Brachionus calyciflorus. The full life-cycle test used renewal
conditions for approximately four days. B. calyciflorus used for the test were less than two-hours
old at test initiation. All animals were parthenogenetically-produced offspring of one individual
from a single resting egg collected from a natural lake in Houhai Park (Beijing, China). The
rotifers were cultured in an artificial inorganic medium at 20ฐC (16-hours:8-hours, light:dark;
3000 lux) for more than six months before toxicity testing to acclimate to the experimental
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conditions. Culture medium was an artificial inorganic medium and all toxicity tests were carried
out in the same culture medium and under the same conditions as during culture (i.e., Ph,
temperature, illumination). Solvent-free stock solutions of PFOA (1,000 mg/L) were prepared by
dissolving the solid in deionized water via sonication. After mixing, the primary stock was
proportionally diluted with dilution water to prepare the test concentrations. Exposure vessels
and size were not reported for the four-day reproductive assay but were likely 6-well cell culture
plates (assumed plastic) each containing at total of 10 Ml of test solution. The test employed
eight test concentrations plus a negative control. Each treatment consisted of six replicates of 10
rotifers each in individual cells. The numbers of living rotifers were counted after four days for
each treatment level. Nominal concentrations were 0 (negative control), 0.125, 0.25, 0.50, 1.0,
2.0, 4.0, 8.0 and 16.0 mg/L. PFOA concentrations were not measured in the rotifer exposures,
but rather in a side experiment using HPLC/MS. The side experiment showed that the
concentration of PFOA measured every eight-hours over a 24-hour period in rotifer medium with
green algae incurs minimal change in the concentration, ranging from 0.25 to 2.0 mg/L. Negative
control survival was not provided for the life-cycle test.
Resting egg production is an ecologically important endpoint for this species because it
represents the final result of sexual reproduction. Based on authors description of results in the
text, "PFOA exposure significantly reduced resting egg production of B. calyciflorus females
during the three-day period." NOEC and LOEC values were not reported, but 0.25 mg/L PFOA
produced more than a 50% reduction in resting egg production. Therefore, it is assumed The B.
calyciflorus four-day NOEC for resting egg production was 0.125 mg/L and the LOEC was 0.25
mg/L, with a calculated MATC is 0.1768 mg/L. Concentration response data from Figure 1 of
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Zhang et al. (2014b) were estimated (WebPlotDigitizer1) and used to derive an EPA-calculated
ECio of 0.076 mg/L (95% C.I. = 0.054 - 0.098 mg/L), further suggesting resting egg production
may be a relatively sensitive endpoint. Because there was only one replicate (as implied by lack
of error bars in Figure 1 of the publication, no clear description of replicates in the methods
section, and no author-reported statistical analysis of this endpoint), resting egg production from
this study was not considered quantitatively acceptable but was retained for qualitative use.
Beyond resting egg production, PFOA did not clearly affect hatching rate of resting eggs when
exposed to PFOA during the formation or hatching period, enhanced hatching rate relative to
controls in most treatments (nominal test concentration range = 0-2.0 mg/L; see figures 3 and 4
of Zhang et al. 2014b) and displayed no clear concentration-response relationship, suggesting
rotifer hatching rate was a relatively tolerant endpoint from this publication. In contrast to Zhang
et al. (2013a), which observed no effect of PFOA on mictic ratio after 28 days at a nominal
concentration as high as 2.0 mg/L PFOA, Zhang et al. (2014b) stated PFOA significantly
increased the F1 mictic ratio from 0.56 in the control treatment to 0.75 and 0.72 in nominal
PFOA test concentrations of 0.25 mg/L and 2.0 mg/L, respectively. Given conflicting results of
PFOA on rotifer mictic ratio, it was not selected as the primary endpoint from Zhang et al.
(2014b). The most sensitive quantitatively acceptable endpoint was the intrinsic rate of natural
increase. The intrinsic rate of natural increase (d"1) is a population level endpoint that accounts
for births and deaths over time. In this study, the intrinsic rate of natural increase was defined as
the natural log of the net increase in the number of rotifers (surviving parents and offspring) for
each PFOA treatment level over a four-day exposure period. This endpoint was conceptually
1 WebPlotDigitizer is an online application used to convert values shown in figures to numerical values. This
application was used to obtain numerical concentration-response data when they were only reported in figures. The
application is free and available online (WebPlotDigitizer - Extract data from plots, images, and maps
(automeris.io')').
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equivalent to the intrinsic rate of natural increase endpoint calculated by Zhang et al. (2013a) but
was a simplification of the calculations performed in Zhang et al. (2013b), in that it only applied
to the four-day observational period, whereas the intrinsic rate of natural increase calculated in
Zhang et al. (2013a) represented the full lifetimes of all individuals within each population (i.e.,
exposure concentration). The EPA-calculated ECio for this endpoint was 1.166 mg/L (95% C.I. =
0.7720- 1.559 mg/L).
Publication: Zhang et al. (2014b)
Species: Rotifer (Brachionus calyciflorus)
Genus: Brachionus
EPA-Calculated ECio: 1.166 mg/L (95% C.I. = 0.7720 - 1.559 mg/L)
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.1913
0.2118
5.6236
0.0014
d
0.2253
0.0081
27.9366
1.392 e7
e
7.7080
0.8191
9.4102
8.183 e5
Concentration-Response Model Fit:
Zhang et al. 2014b
Brachionus calyciflorus
Weibull type 1, 3 para
PFOA ( mg'L )
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C.2.5 Fifth most chronically sensitive genus -Moina
Ji et al. (2008) conducted a chronic life-cycle test on the effects of PFOA (CAS # 335-
67-1, purity unreported; obtained from Sigma Aldrich, St. Louis, MO, USA) with Moina
macrocopa. Tests were done under renewal conditions over a seven-day period and test solutions
were not analytically confirmed. Authors stated that theM macrocopa test followed a protocol
developed and reported by S.R. Oh (2007) (Master's thesis, Seoul National University, Seoul,
Korea), which is similar to OECD 211 (1998), but with slight modification (i.e., shorter test
duration, exposure temperature and different feeding regime: 100 |iL yeast:cerophyll:Tetramin
mixture and 200 |iL algae suspension per day). M. macrocopa used for testing were obtained
from brood stock cultured at the Environmental Toxicology Laboratory at Seoul National
University, Korea. Test organisms were less than 24 hours old at test initiation. Dilution water
was moderately hard reconstituted water (total hardness typically 80-100 mg/L as CaC03).
Experiments were conducted in glass jars of unspecified size and fill volume. Photoperiod for the
test was not reported but was assumed to be 16-hours of illumination, the same conditions as the
daphnid cultures reported in this same publication. Preparation of test solutions was not
described. The test involved ten replicates of one individual each in five nominal test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 3.125,
6.25, 12.5, 25 and 50 mg/L and test solutions were renewed three times per week. Test
temperature was 25 ฑ 1ฐC forM macrocopa. Authors note that the water quality parameters (Ph,
temperature, conductivity, and dissolved oxygen) were measured after changing the medium, but
the information was not reported. Survival of daphnids in the negative control was 100%. TheM
macrocopa seven-day NOEC (reproduction: number of young per adult) was 3.125 mg/L, the
LOEC was 6.25 mg/L and the MATC is 4.419 mg/L. The EPA performed C-R analysis for this
study. The most sensitive endpoint with an acceptable C-R curve was the number of young per
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starting female. The EPA-calculated ECio was 2.194 mg/L PFOA (95% C.I. = -0.7120 - 5.010
mg/L) for M. macrocopa. The lowest treatment concentration produced a greater than a 10%
effect which forced the ECio calculation to extrapolate beyond the lowest treatment
concentration (i.e., not the control, but the nominal treatment of 3.25 mg/L). However, the
resultant ECio value (i.e., 2.194 mg/L) was considered acceptable for quantitative use because it
was largely in agreement with the 14.3% effect observed at the test concentration of 3.125 mg/L.
C.2.6 Sixth most chronically sensitive genus - Neocloeon
Soucek et al. (2023) conducted a chronic life-cycle test to determine the effects of PFOA
(CAS # 335-67-1, 95% purity) on the parthenogenetic mayfly, Neocloeon triangulifer. The test
was performed under renewal conditions over 28 days beginning with <24-hour old nymphs.
Single mayfly exposures were static without renewal for the first four days due to the small size
of starting organisms and renewed three times per week thereafter by transferring organisms to
new exposure chambers. From day 0 to 14, mayflies were exposed in 30 Ml polypropylene cups
with 20 Ml exposure water. Organisms were transferred after 14 days into 250 Ml glass beakers
with 100 or 150 Ml of test water (or control) and to 300 Ml tall form beakers for emergence. At
test initiation, there were 16 replicates per test concentration and control. Nominal test
concentrations were 0.0 (control), 0.047, 0.094, 0.375, 0.750, 1.500 and 3.000 mg/L PFOA.
Mean measured PFOA concentrations (EPA Analytical Method 1633; LC-MC/MS) were
0.000176 (control), 0.043, 0.078, 0.419, 0.672, 1.739 and 3.085 mg/L PFOA, respectively.
Mayflies were exposed at 23 ฑ 1ฐC under a 16:8-hour light:dark cycle and fed 0.2 Ml diatom
slurry plus small scraping on test days 0 and 4 followed by live diatom biofilm scraping after day
4 on solution renewal days. Percent survival in the control after 14 days was 94%. Percent
survival of mayflies after 14 days in the remaining seven test concentrations ranged from 75 to
100%). On test day 14, replicates one through eight were destructively sampled to evaluate 14-
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day weight. Replicates 9-18 were maintained under test conditions until test day 23. No clear
effects of PFOA on 14-day weight or survival were observed in any of the treatments. Similarly,
after 28-days of exposure, no clear effects of PFOA were observed on survival to pre-emergent
nymph (PEN) stage, survival to emergence, number of days to emergence, imago live weight, or
survival to subimago. The relatively high >NOEC (i.e., >3.085 mg/L) suggests TV. triangulifer is
relatively tolerant to chronic PFOA exposures and does not occur among the four most sensitive
genera. The >NOEC (i.e., >3.085 mg/L) was acceptable for quantitative use in deriving the
chronic freshwater PFOA criterion.
C.2.7 Seventh most chronically sensitive genus - Oryzias
Lee et al. (2017) conducted a multiple generation exposure to determine the effects of
PFOA (CAS # 335-67-1, purity was not reported) on the reproductive toxicity and metabolic
disturbances to Oryzias latipes. Fish were originally received from the Department of Risk
Assessment of the National Institute of Environmental Research (NIER; South Korea) and
maintained according to the following conditions: dissolved oxygen 7-8 mg/L, Ph 7.5 ฑ 0.2,
water temperature 25 ฑ 1ฐC, 16-hour light, 8-hour dark photoperiod and total hardness 55-57
mg/L (as CaC03). O. latipes were fed Artemia salina once daily, based on the OECD test
guideline 240 feeding schedule. Adults (about 13 weeks old, when genders could be visually
differentiated) through the F2 generation were exposed to three nominal concentrations of PFOA
(0.3, 3 and 30 mg/L PFOA) for a total exposure period of 259 days. At test initiation, four pairs
of both genders were introduced into the test chambers (8 L glass tank) of a flow-through
exposure system. PFOA solutions were replenished five times daily to keep the same water
quality as fish maintaining condition. PFOA exposure continued for three weeks, during which
eggs produced by mating of F0 fish were removed from test chamber and counted daily for
fecundity. During test week four, (spawning period), the F1 generation eggs (n = 192) were
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obtained per concentration. Right after the spawning period was finished, FO fish were used to
evaluate metabolism disturbance. The F1 eggs were pooled and redistributed into an incubation
chamber containing PFOA solution. After the hatching was completed, the test organisms were
returned into test chambers and raised under flow-through PFOA exposure conditions until they
reached adult stage (at about 13 weeks old), during which sac-fry survival rate, hatching rate, and
abnormality of F1 were analyzed. When F1 fish reached the adult stage, sex ratio of total F1 fish
was determined and 32 individuals of F1 fish were used to analyze gonadosomatic (GSI),
hepatosomatic (HSI), condition factor (K), VTG expression, and histological alterations. In
addition, other F1 fish (32 male and female fish) were used for obtaining the F2 generation eggs
same as the FO generation. The exposure conditions to F2 fish were carried out in the same
manner as F1 fish. Consequently, FO fish were exposed to PFOA for four weeks and F1 and F2
fish were exposed to PFOA across all life cycle stages without exposure pause. The exposure
regime was applied equally in all test groups. The 259-day MATC of 9.487 mg/L PFOA was
reported for F2 sac-fry survival and fecundity for the FO, F1 and F2 generations and represented
the most sensitive endpoints from the study. Reproductive responses reported by Lee et al.
(2017) appear to be control normalized; however, use of control normalized data in this study
does not alter conclusions from hypothesis-based testing (i.e., use of aNOEC, LOEC, or
MATC). Beyond F2 survival (which had control mortality), the EPA attempted C-R analysis for
all endpoints reported by Lee et al. (2017). Given the large dilution factor between PFOA
treatments, C-R models could either not be fit, or when models could be fit, they performed
poorly on statistical metrics and were not used. Therefore, the 259-day MATC of 9.487 mg/L
PFOA was considered to be quantitatively acceptable for criterion derivation. The large dilution
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factor from this test does not support concentration response modeling, consequently EPA relied
on an MATC (i.e., 9.487 mg/L) as the chronic effect level from this test.
C.2.8 Eighth most chronically sensitive genus - Ceriodayhnia
Kadlec et al. (2024) tested the chronic toxicity of pefluorooctanoic acid (PFOA) to
Ceriodaphnia dubia for seven days in three measured, renewal experiments. Similar chronic tests
were also performed with Chironomus dilutus and Hyallela azteca, but this summary is limited
to the results of the C. dubia tests. Test chemicals were obtained from Sigma, Alfa Aesar,
Synquest, and Toronto Research Chemical (purity 96-99%). Test organisms were obtained from
in-house cultures maintained following ASTM and EPA protocols. Test water was UV-treated
and sand-filtered Lake Superior water supplemented with Na2SC>4, NaCl, KC1, CaCh x 2H2O,
and MgChx 6H2O. Testing protocols followed species-specific ASTM methodologies (ASTM
2002). Three separate tests were conducted, each with a 0.5x dilution series of measured PFOA
concentrations, with ten replicates of each concentration, and one organism per replicate. Test 1
mean concentrations were 0.019 (control), 0.19, 0.31, 0.62, 1.3, 2.7, 5.5, 11, 22, 44 and 91 mg/L.
Test 2 mean concentrations were 0.67 (control), 6.7, 14, 28, 55, and 109 mg/L. Test 3 mean
concentrations were 0.35 (control), 3.5, 6.8, 14, 30 and 57 mg/L. C. dubia neonates (<8-hour)
were placed in one ounce polystyrene cups filled with 15 mL of test solution. Dissolved oxygen
and pH were measured twice in each exposure per treatment and twice in the stock solutions.
Test chambers were placed in a water bath to maintain a steady temperature under a 16:8
light:dark cycle. Study authors reported average water quality measurements of 24.7ฐC, 8.6 mg/L
DO, 7.8 pH, 52 mg/L as CaC03 hardness, 41 mg/L as CaC03 alkalinity, and 145 |imhos/cm
conductivity. Testing solutions were renewed daily, and organisms were fed daily with 100 |ig/L
YCT and algae. Control survival was 96.8% with a mean reproduction of 26.8 offspring per
surviving female. EC20S and EC50S for survival and young per surviving female were calculated
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following methods described in Mount et al. (2016), using custom software written with Intel
Visual Fortran Compiler Xe (Intel Corporation) and Winteracter 13.0 (Interactive Software
Services). The author reported EC20S for young per surviving female for tests 1, 2, and 3 were
26.2, 25.1, and 33.8 mg/L, respectively. Concentration-response data were reported for these
tests, allowing the EPA to independently model concentration-response curves using the dose-
response curve package in R. The EPA-calculated EC10 values for tests 1, 2, and 3 were 20.42,
21.69, and 29.54 mg/L, respectively, which were determined to be acceptable for quantitative
use.
C.2.9 Ninth most chronically sensitive genus - Gobiocyyris
The chronic toxicity of PFOA (98% purity) on the rare minnow (Gobiocypris rarus; not
North American resident species) was investigated by Wei et al. (2007) using flow-through
unmeasured exposure conditions. Two hundred and forty mature male and female rare minnows
(about nine months old, 1.4 ฑ 0.4 g, 47.7 ฑ3.6 mm) were obtained from a laboratory hatchery
and randomly assigned to eight 20 L glass tanks (30 individuals per tank). Fish were supplied
with dechlorinated tap water under continuous flow-through conditions at 25 ฑ 2ฐC and a
photoperiod of 16-hours:8-hours light:dark. During the 28-day exposure period, fish were fed a
commercial granular food (Tetra) at a daily rate of 0.1% body weight. Waste and uneaten food
were removed daily. After a one-week acclimation period, 30 randomly selected male and 30
female rare minnows (gender determined by observing the shape of the abdomen and the
distance between the abdomen fin and the stern fin) were assigned to one of the four nominal
PFOA exposures (0, 3, 10 or 30 mg/L PFOA). Each treatment was performed in duplicate tanks.
The flow rate of the test solution was 8 L/hour, and actual PFOA concentrations in the tanks
were not verified by chemical analysis. During the exposure period, there were separate inputs
for water and PFOA and the mixer helped mix PFOA and water before flowing into the tanks.
C-33
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The concentration of mixed solution flowing out from the mixer were kept at 3, 10 or 30 mg/L
PFOA by adjusting the input flow rate of concentrated PFOA and water, respectively. After 14-
day and 28-day exposure periods, fish were anesthetized on ice, and liver samples were taken
and immediately frozen in liquid nitrogen and stored at -80ฐC until analyzed. No mortality was
observed in any treatments. The 28-day LOEC (survival) was >30 mg/L PFOA and was
acceptable for use as a high-unbounded value from a high-quality study which provides relevant
sensitivity information for this fish species.
C.2.10 Tenth most chronically sensitive genus - Oncorhynchus
Centre International de Toxicologic (2004) and Colombo et al. (2008) evaluated the
chronic effects of ammonium perfluorooctanoate (APFO, CAS #3825-26-1, 99.7% purity) to
embryos of the rainbow trout, Oncorhynchus mykiss. Stock solutions of APFO were prepared by
dissolving the test substance directly in the test media or dilution water and then diluting the
stock solution to provide a geometric series of test concentrations (nominal concentrations of
3.13, 6.25, 12.5, 25 and 50 mg/L APFO). The early-life-stage (ELS) test was performed under
flow-through conditions and in compliance with OECD test guideline 210. Unfertilized trout
eggs and sperm were received from a commercial supplier and the eggs were fertilized in the
laboratory. One hundred and eighty newly fertilized eggs were randomly selected and allocated,
60 eggs per replicate, to the three replicate test vessels for each control and test concentration.
Authors stated that the number of surviving fish was reduced randomly to 30 per replicate just
after the end of the hatching period (day 26) in the control. The number of surviving fish was
again reduced randomly to 15 per replicate when swim-up and feeding began on day 50.
Actively feeding juveniles were fed trout chow two to four times per day, corresponding to
approximately 4% of their body weight per day, from day 50 to the end of the 85-day test. Test
solutions were continuously renewed during the study by pumping the stock solutions into
C-34
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flowing dilution water with a peristaltic pump system at a replacement rate of 5.76 times the test
vessel volume per day. Dilution water pH was 6.0-8.5, total hardness was 150 mg/L as CaCCb,
and water temperature was kept between 11.1 and 12.5ฐC for embryos and between 11.6 and
14.4ฐC for larvae and juvenile fish. The dissolved oxygen concentration was greater than 60% air
saturation, the light/dark cycle was maintained at constant darkness until seven days after
hatching, then 16 hours light and eight hours dark through test end. Observations were made
daily as follows: eggs-marked loss of translucency and change in coloration, white opaque
appearance; embryos-absence of body movement or heartbeat; larvae and juvenile fish-
immobility, absence of respiratory movement or heartbeat, white opaque coloration of the central
nervous system, lack of reaction to mechanical stimulus, and abnormalities. The reported 85-day
growth and mortality NOEC was 40 mg/L PFO"; however, the authors' note that the contribution
of ammonia from APFO exposure indicates that un-ionized ammonia could be a potential
contributor to the observed toxicity of APFO. Although the authors cite U.S. EPA (1999) for un-
ionized ammonia toxicity values, that document (and the subsequent U.S. EPA [2013] criteria
document) expressed toxicity in terms of a relationship between total ammonia nitrogen and pH
and temperature. For rainbow trout, U.S. EPA (1999) declined to specify a chronic value, due to
inconsistencies between tests. However, U.S. EPA (2013) set the rainbow trout chronic value at
6.66 mg TAN/L (Total Ammonia Nitrogen/L) at pH = 7. Using the normalization equations in
U.S. EPA (2013), the rainbow trout chronic value translated to 3.60 mg N/L at the authors'
assumed chronic test pH of 7.8 (see table 7 of Colombo et al. [2008]), which in turn translated to
a rainbow trout chronic value as un-ionized ammonia of 0.064 mg un-ionized ammonia/L at pH
7.8 and a reported test of temperature (13ฐC). Table 7 of Colombo et al. (2008) listed the un-
ionized ammonia concentration at their APFO NOEC as 0.013 mg un-ionized ammonia/L, which
C-35
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is 4.9-fold lower than the EPA's chronic value for rainbow trout re-expressed as an un-ionized
ammonia concentration for the test condition. Therefore, the EPA does not believe ammonia was
a confounding factor in this test and the study was determined to be quantitatively acceptable for
criterion derivation.
C.2.11 Eleventh most chronically sensitive genus - Pimeyhales
Bartlett et al. (2021) also evaluated the chronic effects of PFOA (CAS# 335-67-1, 96%
purity, solubility in water at 20,000 mg/L, purchased from Sigma-Aldrich) on fathead minnows
(Pimephalespromelas) via a 21-day early-life stage static-renewal, measured study. The authors
followed OECD Test Guideline 210, except that the test ended at 16 days post-hatch (dph)
compared to 28 dph for the standard OECD test. Test water (i.e., stock solutions, exposure
solutions and controls) was charcoal-filtered UV-sterilized Burlington City water from Lake
Ontario (total hardness 120-130 mg/L, alkalinity 89-93 mg/L, Ph 7.4-7.8), and was maintained
in a header tank prior to use in testing. Fathead minnow eggs (<18-hour post fertilization) were
purchased from Aquatox Labs (Guelph, ON) and exposed to nine nominal PFOA concentrations:
0.01, 0.032, 0.1, 0.32, 1, 3.2, 10, 32 and 100 mg/L. The tests were divided into low concentration
(0.01-10 mg/L) and high concentration (32-100 mg/L) tests, with five days in the egg stage and
16 days in the larval fish stage. Tests were initiated with eggs from five to ten egg batches (from
different fathead minnow breeding groups) to maximize genetic diversity and variability. There
were 20 eggs per beaker, with eight replicates of controls and four replicates of each PFOA
concentration in each of the two tests. Embryos and larvae were held in glass, Nitex mesh
bottomed (mesh size 500 |im) egg cups within 800-M1 HDPE beakers filled to 700 Ml with test
solution. Beakers containing fathead minnow eggs/larvae were aerated, loosely covered, and held
in a 25ฐC incubator with a photoperiod of 16 hours light and 8 hours dark. Larvae were fed 10
|iL/fish (0-9 dph) and 20 |iL/fish (9-16 dph) of newly hatched brine shrimp slurry per day. The
C-36
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first feeding (half of the daily aliquot) was two hours prior to the daily solution changeover (to
remove excess food and waste), and the second feeding (the other half of the daily aliquot) was
after solution changeover, so that food was available at all times during the tests. Endpoints
evaluated were survival to hatch, time to hatch, hatching success, deformities at hatch, uninflated
swim bladder, survival from the egg until nine and 16 dph, and weight, length, tail length, and
condition factor of larvae at nine and 16 dph. The reported 21-day NOEC for mortality, weight,
length, and condition factor was 76 mg/L PFOA. The EPA could not independently calculate an
ECio value because no effects were observed across the range of concentrations tests. Because
the NOEC of 76 mg/L was a relatively tolerant NOEC value it was considered quantitatively
acceptable for criteria derivation.
C.2.12 Twelfth most chronically sensitive genus - Chironomus
McCarthy et al. (2021) conducted a 19-day "abbreviated full life cycle" PFOA (97%
purity, purchased from Sigma-Aldrich) toxicity test on the midge, Chironomus dilutus. The
PFOA stock solution was dissolved in reconstituted moderately hard water without the use of a
solvent and stored in polyethylene at room temperature until use. Authors reported that they
followed standard protocols (ASTM 2005; U.S. EPA 2000) with slight modifications. Exposure
vessels for both experiments were 1 L high-density polyethylene beakers containing natural
field-collected sediment with 60 mL of sediment and 105 mL of test solution. PFOA test
solutions were added via pipette to the beakers with the tip just above the sediment substrate.
Nominal test concentrations were 0, 26, 87, 149, 210 and 272 mg/L PFOA, respectively. Test
concentrations were based on the results of a 10-day range finding test conducted by McCarthy
et al. (2021). Egg cases were obtained from Aquatic Biosystems or USGS Columbia
Environmental Research Center and held as free-swimming hatched embryos (<24 hour after
hatch) before testing. Each beaker held 12 organisms with five replicates per exposure treatment.
C-37
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Solutions were renewed every 48 hours. PFOA treatment concentrations were measured on days
10, 15 and 20 in the 20-day exposure. Mean measured PFOA concentrations in the 20-day
exposure were 0 (control), 19.9, 59.4, 145, 172 and 227 mg/L PFOA. Percent survival in the
control treatment was 82%. The most sensitive endpoint was survival with an author reported 19-
day ECio of 89.8 mg/L PFOA. The EPA-calculated survival-based ECio was 88.32 mg/L (95%
C.I. = 15.40 - 161.3 mg/L), which was acceptable for quantitative use.
C-38
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Appendix D Acceptable Estuarine/Marine Chronic PFOA Toxicity Studies
D,1 Summary Table of Acceptable Quantitative Estuarine/Marine Chronic PFOA Toxicity Studies
Spocios (lil'cshitic)
Method'
Tesl
Dui'iilioii
( hcniiciil /
PuriU
Dll
Temp.
(ฐC)
S;ilinil>
(DDI)
Chronic Yiiluc
1. ii(l point
Author
Reported
Chronic
\ iilue
(mii/l.)
l-'.PA
( iilculiilcd
(hidiiic
\;ilne
(lliu/l - >
liiiiil
(hidiiic
Yiiluc
(niii/l.)1'
Species
Mciin
(hidiiic
\ illllC
Reference
Purple sea urchin
(adult),
Paracentrotus
lividus
S,M
28 days
PFOA
>98%
17.00
35
MATC
(survival)
31.62
-
31.62
31.62
Savoca et al.
2022
Japanese medaka,
Oryzias latipes
S,U
30 days
Perfluoro-
octanoate
96%
7.72
24.82
34.52
NOEC
(growth - condition
factor)
1.0
-
>1.0
>1.0
Oh et al. 2013
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b Values in bold used in the SMCV calculation
D-l
-------
D.2 Detailed Study Summaries of Chronic Saltwater PFOA Toxicity Studies
Considered for Use in Saltwater Criterion Derivation
The purpose of this section is to present detailed study summaries for chronic
estuarine/marine tests that were considered quantitatively acceptable for criterion derivation,
with summaries grouped and ordered by genus sensitivity. Unlike Appendix A.2 and Appendix
C.2, the EPA-calculated C-R models were not presented below for the four most sensitive
estuarine/marine genera because a chronic estuarine/marine criterion or benchmark was not
developed.
D.2.1 Most chronically sensitive estuarine/marine genera - Oryzias
Oh et al. (2013) evaluated the chronic toxicity of PFOA on the Japanese medaka,
Oryzias latipes, in a 30-day static exposure. PFOA (96% pure, CAS No. 335-67-1) was dissolved
in filtered seawater with a minimal concentration (<0.001%) of DMSO used as a vehicle to
prevent cellular damage to the fish. Only nominal concentrations were used throughout the
study. Prior to test initiation, the fish were acclimated to a seawater environment. The fish were
maintained at 25ฐC under a constant photoperiod of 16:8 hour (light:dark) and water quality was
monitored by measuring the pH, dissolved oxygen, and temperature. Fish from the 3rd generation
of 0. latipes (n = 7/group) that had acclimated to seawater for over one month were used in the
exposure experiments. Fish were exposed for 30 days to one PFOA concentration (1.0 mg/L), a
0.22 |im filtered seawater control, and a DMSO carrier solvent control. Test conditions were
maintained at an average temperature of 24.82ฐC, pH of 7.72, dissolved oxygen of 6.04 mg/L,
and salinity of 34.52 practical salinity units, with fish fed daily. The 30-day condition factor
NOEC of 1.0 mg/L PFOA was selected as the primary endpoint from this study. In the methods
section, Oh et al. (2013) also stated, "In our preliminary study, fish mortality was altered 30 days
after perfluorinated compound exposure, suggesting that repeated exposure to PFCs for 30 days
D-2
-------
at 1 |ig/mL [1 mg/L] causes adverse effect on 0. latipesThe statement about mortality-based
effects in their preliminary test is in direct conflict with the condition factor-based results from
the primary test described in the publication. Few details are provided about the preliminary test.
Results of the final test (i.e., condition factor NOEC of 1.0 mg/L; reported in Table 1 of Oh et al.
2013) were retained as quantitatively acceptable because they provide chronic estuarine/marine
data that were otherwise limited.
D.2.2 Second most chronically sensitive estuarine/marine genera - Paracentrotus
Savoca et al. (2022) tested chronic toxicity of perfluorooctanoic acid (PFOA) to sea
urchins (Paracentrotus lividus) in a 28-day measured, static experiment. Analytical grade PFOA
(>98% purity) was obtained from Sigma-Aldrich. Thirty-six adult sea urchins were collected off
the coast near Capo Zafferano, Italy, a site where little to no PFOA exposure was expected, and
were acclimated in the laboratory for eight days in 200 L of filtered (30 jam) seawater. Four sea
urchins were placed into one of nine aquariums containing 15 L of seawater at nominal
concentrations (0 (control), 10, and 100 mg/L), with three replicates per concentration treatment.
Organisms were kept at 17ฑ1ฐC with a 35ฑ1% salinity under a 12:12 light:dark cycle. To avoid
PFOA uptake, organisms were not fed during the experiment and tanks did not have filtration.
Instead, continuous aeration was employed, and PFOA in test water was monitored weekly in
each tank. The average recovery of PFOA was stable and close to nominal across all treatments
throughout the experiment, and PFOA in control water was less than 10 ng/L. Coelomic fluid
was sampled from all individuals weekly to measure PFOA uptake. During the exposure period,
sea urchins at the highest test concentration showed sublethal signs of toxicity, including reduced
spine mobility, spine loss, and reduced ability to remain anchored to the bottom of the test
vessels. Mortality was only observed at the highest test concentration. After the exposure, or in
the case of urchins exposed to the highest PFOA concentrations, after the onset of adverse
D-3
-------
effects, egg samples were dissected from the gonads of each organism, fertilized, and held at
18ฑ1ฐC for 48 hours. Adult mortality, adult PFOA accumulation in coelomic fluid, percentage of
normal and abnormal embryos, and gene expression in embryos were measured. Adverse effects
of PFOA were observed for all endpoints. ANOVA or Kruskal-Wallis tests were used to assess
significant differences in PFOA treatments compared to controls. The NOEC and LOEC for
adult mortality were 10 mg/L and 100 mg/L, respectively, and the MATC of 31.62 mg/L was
determined to be acceptable for quantitative use.
D-4
-------
Appendix E Acceptable Freshwater Plant PFOA Toxicity Studies
E.l Summary Table of Acceptable Quantitative Freshwater Plant PFOA Toxicity Studies
Species
Method'1
Tesl
Dui'iilioii
( hcinic;il
/ Piiriu
Pll
Temp.
<ฐC)
r.iTcci
Kcpurlcri
HITccl
( onccnlmlion
(inii/l.)
Reference
Cyanobacteria,
Microcystis aeruginosa
S,U
12 days
PFOA
>95%
25
MATC
(population abundance - cell
density)
15.81
Hu et al. 2023
Green alga,
Chlamydomonas reinhardtii
s,u
96 hours
PFOA
>96%
6.8
25
ECso
(growth)
51.9
Huetal. 2014
Green alga,
Chlamydomonas reinhardtii
s,u
8 days
PFOA
>96%
6.8
25
MATC
(cell number)
28.28
Hu et al. 2014
Green alga,
Chlamydomonas sorokiniana
s,u
96 hours
PFOA
>96%
7.1
25
ECso
(population abundance inhibition
rate)
140.59
Zhao et al. 2023
Green alga (7.0 x 105 cells/mL),
Chlorella pyrenoidosa
S,M
96 hours
PFOA
>98%
-
25
ECso
(growth)
190.99
Xu et al. 2013
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.73
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.95
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.95
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.86
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.84
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.97
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.94
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.95
Hu et al. 2020
Green alga,
Chlorella pyrenoidosa
S,M
10 days
PFOA
>98%
7.1
23
NOEC
(biomass)
0.83
Hu et al. 2020
Green alga (9 x 105 cells/mL),
Chlorella pyrenoidosa
S,U
96 hours
PFOA
>95%
-
25
NOEC
(growth)
0.1
Li et al. 2021b
E-l
-------
Species
Mclhori'1
Tesl
Dui'iilioii
( hciiiiciil
/ Piiriu
Pll
Temp.
<ฐC)
HITccl
Kcpuricri
HITccl
( onccnlmlion
(inii/l.)
Reference
Green alga (1.5 x 104 cells/mL),
Chlorella vulgaris
S,U
96 hours
PFOA
95%
23
IC50
(cell density)
115.5
Boudreau 2002
Green alga,
Raphidocelis subcapitata
(formerly, Pseudokirchneriella subcapitata
and Selenastrum capricornutum)
s,u
14 days
APFO
96.5-100%
-
23
ECio
(cell count)
5
3M Co. 2000a;
Elnabarawy
1981
Green alga (1.5 x 104 cells/mL),
Raphidocelis subcapitata
s,u
96 hours
PFOA
95%
-
23
IC50
(cell density)
123.4
Boudreau 2002
Green alga (log phase growth),
Raphidocelis subcapitata
S,M
96 hours
APFO
99.7%
-
21-25
MATC
(biomass and growth rate)
16.07
Colombo et al.
2008
Green alga (7.0 x 105 cells/mL),
Raphidocelis subcapitata
S,M
96 hours
PFOA
>98%
-
25
ECso
(growth)
207.46
Xu et al. 2013
Green alga,
Scenedesmus obliquus
S,U
96 hours
PFOA
>96%
6.8
25
ECso
(growth)
44.0
Hu et al. 2014
Green alga,
Scenedesmus obliquus
S,U
8 days
PFOA
>96%
6.8
25
NOEC
(cell number)
40
Hu et al. 2014
Green alga,
Scenedesmus quadricauda
S,M
96 hours
PFOA
99%
7
22
ECso
(growth inhibition rate)
269.63
Yang et al. 2014
Duckweed,
Lemna minor
S,M
96 hours
PFOA
>97.6%
6.5
25
NOEC
(population growth rate)
9.478
Wu et al. 2023
Water milfoil (4 cm apical shoots),
Myriophyllum sibiricum
S,M
14 days
PFOA
Unreported
8.3-8.7
17.8-22.0
ECio
(dry weight)
8.7
Hanson et al.
2005
Water milfoil (4 cm apical shoots),
Myriophyllum sibiricum
S,M
21 days
PFOA
Unreported
8.3-8.7
17.8-22.0
ECio
(dry weight)
7.9
Hanson et al.
2005
Water milfoil (4 cm apical shoots),
Myriophyllum sibiricum
S,M
35 days
PFOA
Unreported
8.3-8.7
17.8-22.0
ECio
(wet weight)
21.6
Hanson et al.
2005
Water milfoil (4 cm apical shoots),
Myriophyllum spicatum
S,M
14 days
PFOA
Unreported
8.3-8.7
17.8-22.0
ECio
(dry weight)
18.1
Hanson et al.
2005
Water milfoil (4 cm apical shoots),
Myriophyllum spicatum
S,M
21 days
PFOA
Unreported
8.3-8.7
17.8-22.0
ECio
(plant length)
5.7
Hanson et al.
2005
Water milfoil (4 cm apical shoots),
Myriophyllum spicatum
S,M
35 days
PFOA
Unreported
8.3-8.7
17.8-22.0
ECio
(dry weight)
19.7
Hanson et al.
2005
E-2
-------
Species
Method'
Tesl
Dui'iilioii
( hemic;il
/ Piiriu
Dll
Temp.
(ฐC)
r.iToci
KcpurU'ri
I'-ITocl
( onceii 1 r;i 1 ion
(inii/l.)
KoI'oiviico
Lettuce (seed),
Lactuca sativa
S,U
5 days
PFOA
96%
-
EC50
(root elongation)
745.7b
Ding et al.
2012b
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, NR=not reported
b Reported in moles converted to milligram based on a molecular weight of 414.07 mg/mmol.
E-3
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E.2 Summary of Quantitatively Acceptable Plant PFOA Toxicity Studies
E.2.1 Cvanobacteria. Microcystis aeruginosa
Hu et al. (2023) tested perfluorooctanoic acid (PFOA) on Microcystis aeruginosa for 12
days in a static, unmeasured experiment. PFOA (>95% purity) was purchased from Sigma-
Aldrich Corp. (St. Louis, MO, USA). Test organisms were obtained from the Freshwater Algae
Culture Collection at the Institute of Hydrobiology in Wuhan, China, and were cultured in BG-
11 medium within Erlenmeyer flasks at 25ฐC and 30|imol m"V photosynthetic photon flux
density. M. aeruginosa in the exponential growth phase were added to 250 mL Erlenmeyer flasks
containing 100 mL of test solution at a density of approximately 1.45xl06 cells/mL. The
experimental design consisted of nominal concentrations of 0 (control), 1,10, 25, and 50 mg/L
PFOA, diluted from a PFOA stock solution of 1,000 mg/L prepared in BG-11 medium. All
concentrations were repeated in triplicate, and each Erlenmeyer flask was manually shaken twice
daily during the 12-day experiment. One mL of algal-test chemical solution was sampled from
each flask daily to measure cell density. Chlorophyll a, photosynthesis, respiration, cellular ROS
production, growth inhibition, enzyme activity, emission spectra, carbohydrate content, and total
RNA were also measured. Statistical differences between treatments and controls were assessed
using ANOVA and Dunnett's tests (P<0.05). Algal cells began showing signs of growth
inhibition after two days of exposure, and decreased cell density in the highest PFOA
concentration after four days. Starting on day five, cell density was significantly lower than
controls in the 25 and 50 mg/L treatments. The 12-day NOEC and LOEC for cell density were
10 and 25 mg/L, and the MATC of 15.81 was determined to be acceptable for quantitative use.
E.2.2 Green alga. Chlamydomonas reinhardtii
Hu et al. (2014) evaluated the growth inhibition of PFOA (>96% purity) with
Chlamydomonas reinhardtii in 96-hour and eight-day static exposures. Authors stated that the
E-4
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tests followed OECD test guidance 201 (OECD 2006). Chlamydomonas reinhardtii were
supplied by UTEX Culture Collection of Algae, University of Texas at Austin. Dilution medium
was described as modified high-salt media at a pH of 6.8. Algae in exponential growth phase
were exposed to nominal concentrations of 0 (negative control), 1,3.16, 10, 31.6, 100, 316 and
1,000 mg/L in the 96-hour exposure, and 0, 5, 10, 20 and 40 mg/L in the eight-day exposure.
Experiments were initiated by inoculating equal cell numbers of lxlO4 cells/mL in the 96-hour
exposure and 5xl06 cells/mL in the eight-day exposure into 250 mL flasks containing a total
volume of 100 mL of algal cell suspension per flask. There were five replicates per treatment in
the 96-hour exposure and three replicates in the eight-day exposure. Algae were incubated at
25ฐC under cool-white fluorescence lights at 85-90 |imol photons/[m2 x s] irradiance with a 16-
hour:8-hour light:dark cycle. The 96-hour growth EC so (inhibition based on optical density) was
51.9 mg/L. The 8-day MATC based on cell number was 28.28 mg/L (NOEC and LOEC are 20
and 40 mg/L, respectively). The plant values from the study were acceptable for quantitative use.
E.2.3 Green alga. Chlamydomonas sorokiniana
Zhao et al. (2023) tested perfluorooctanoic acid (PFOA) on Chlorella sorokiniana for 96
hours in a 96-hour, static, unmeasured experiment. The study also investigated the effects of
polystyrene microplastics on C. sorokiniana, both independently and in combination with PFOA,
but only the independent effects of PFOA are summarized here. PFOA (>96% purity) was
purchased from Aladdin (Shanghai, China). Algae were purchased from the Freshwater Algae
Culture Collection at the Institute of Hydrobiology in Wuhan, China. Before the experiment, the
algae were cultured in BG-11 media in an illumination incubator at 25ฑ1ฐC on a 12:12 hour
light-dark cycle at a light intensity of 2,000ฑ50 lux. Algal cultures were shaken three times daily.
C. sorokiniana in the exponential growth phase were added to test vessels containing test
solution at a density of approximately l.OxlO6 cells/mL. The experimental design consisted of
E-5
-------
nominal concentrations of 0 (control), 0.05, 0.5, and 5.0 mg/L PFOA, diluted from a PFOA stock
solution of 500 mg/L prepared in BG-11 medium, with three replicates per treatment. Cell
density and chlorophyll a were measured every 24 hours, and population growth inhibition rate
was calculated after 96 hours. Several additional endpoints related to morphology and oxidative
stress were also measured after 96 hours. Statistically significant (p<0.05) differences between
treatment were assessed using ANOVA and least significant different tests. The EC so for
population growth inhibition rate was calculated using SPSS 19 software. The 96-hour EC so
value of 140.59 mg/L for population growth inhibition rate was determined to be acceptable for
quantitative use.
E.2.4 Green alga. Chlorella yyrenoidosa
Xu et al. (2013) performed a 96-hour static, measured algal growth inhibition test on
PFOA (>98% purity) with Chlorellapyrenoidosa. Algae were obtained from the Aquatic
Organism Research Institute of the Chinese Academy of Science and precultured for three
generations prior to initiating the test. Dilution medium consisted of number one culture medium
supplemented with aquatic number four nutrient solution (Zhou and Zhang 1989). Algae in
logarithmic growth phase (7.0 x 105 cells/mL) were inoculated in medium containing PFOA at 0
(negative control), 30, 60, 90, 120, 150, 180, 240, 300 and 360 mg/L. Tests were conducted in
100 mL conical flasks with 50 mL of solution with each concentration replicated three times.
Exposure concentrations were verified via UHPLC-MS/MS using the Agilent 1290 Infinity
UHPLC system interfaced with an Agilent 646-0 Triple Quadrupole mass. Algae were exposed
under a 12-hour: 12-hour light:dark cycle at 3,000-4,000 lux and 25ฐC. Chlorophyll concentration
and permeability of cell membranes was determined after 96-hours of exposure. The reported 96-
hour growth ECso (inhibition based on optical density) was 190.99 mg/L and was considered to
be acceptable for quantitative use.
E-6
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Hu et al. (2020) also evaluated the toxic effects of PFOA (>98% purity) on the green
alga, Chlorellapyrenoidosa, in both nutrient rich and nutrient limited media. PFOA was
dissolved in deionized water to make stock solutions. Green algae were purchased from the
Chinese Academy of Sciences and cultivated in sterile blue-green medium (BG11) with a 14-
hour photoperiod and pH of 7.1, at 23 ฑ 1ฐC. Algae in the cultures were in the logarithmic
growth phase before they were used for testing. Nominal PFOA test concentrations in standard
growth media (BG11) were 0.05, 0.2, 0.5 and 1.0 mg/L. In the eight nutrient-limited, -enriched,
or -starved tests only one PFOA test concentration (1.0 mg/L) was examined. All experiments
included controls, lasted for 10 days and were replicated three times. The initial cell density for
each treatment was 1.0 x 106 cells/mL and followed OECD test guidelines (OECD 2011). PFOA
was measured in all test treatments at test termination. After exposure day 10, cell density, dry
biomass, chlorophyll content, malondialdehyde content, catalase/peroxidase activities, and cell
membrane property of algae were measured. In all tests there was no effect on plant biomass
compared to the controls at the highest or only test concentration. The 10-day biomass NOECs
for the study were 0.73, 0.95, 0.95, 0.86, 0.84, 0.97, 0.94, 0.95 and 0.83 mg/L PFOA for the
BG11 media, N,P-limited, N-limited, P-limited, N,P-enriched, N,P-starved, N-starved, P-starved
and N,P-enriched tests, respectively. All these NOECs were considered acceptable for
quantitative use. Although these NOEC values are relatively low compared to the other plants
and animals tested, they represent NOEC values and, therefore, do not suggest effects are
expected to occur at these relatively low concentrations.
Li et al. (2021b) conducted a 12-day static, unmeasured toxicity test with PFOA (>95%
purity, purchased from Sigma-Aldrich) on the green alga, Chlorella pyrenoidosa. The FACHB-9
strain of the green alga was purchased from the Institute of Hydrobiology, Chinese Academy of
E-7
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Sciences. The alga was cultured in BG-11 medium at 25ฐC under a 12-hour: 12-hour light:dark
photoperiod (2,000 lux) and shaken manually every 12 hours. Two PFOA solutions (0.100 and
100 |ig/L) were prepared in sterile water and control solutions were sterile water only. Test
solutions were added to flasks containing an initial density of 9 x 105 cells/mL growing in BG-11
medium, with three flasks for each treatment. The variation in algal density was observed every
day over the 12-day exposure period and chlorophyll pigment content and photosynthetic activity
was observed on days 3, 6, 9 and 12. During the first half of the test there was no significant
difference in growth of PFOA treatments and the control. On day 12, the growth was reduced by
6.76 and 14.4% relative to the control, in the 0.1 and 100 |ig/L PFOA treatments, respectively.
Later time points (i.e., >4-12 days of exposure) were not used quantitatively use because the
exposure duration was too long (U.S. EPA 2012). The four-day cell density-based endpoint with
a NOEC of 0.1 mg/L was acceptable for quantitative use.
E.2.5 Green alga. Chlorella vulgaris
Boudreau (2002) performed a 96-hour static algal growth inhibition test on PFOA (acid
form, CAS # 335-67-1, 95% purity) with Chlorella vulgaris as part of a Master's thesis at the
University of Guelph, Ontario, Canada. The author stated that the algal growth inhibition tests
followed protocols found in ASTM E 1218-97a (ASTM 1999) and Geis et al. (2000). Chlorella
vulgaris (UTCC 266 strain) used for testing were obtained as slants from the University of
Toronto Culture Collection (UTCC; Toronto, Canada). Stock concentrations were prepared in
laboratory-grade distilled water with a maximum concentration that did not exceed the critical
micelle concentration for PFOA of 450 mg/L. Dilution medium was Bristol's algal growing
media. Toxicity testing consisted of a range-finder test and at least two definitive tests. Nominal
test concentrations were 0 (negative control), 6.7, 12.5, 25, 50, 100, 200 and 400 mg/L. Tests
were conducted in 60 x 15 mm polyethylene disposable Petri dishes containing 20 mL of test
E-8
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solution. Each Petri dish was inoculated with 1.5 x 104 cells/mL at initiation and replicated four
times per test concentration. Tests were continuously illuminated with cool-white, fluorescent
light between 3,800 and 4,200 lux and incubated at 23 ฑ 1ฐC. Replicate Petri dishes were
manually shaken twice a day during testing. Toxicity test endpoints included cell density and
chlorophyll-c/ content. The reported ICio, IC25 and IC50 based on growth inhibition (measured as
either chlorophyll-a or cell density) were 0.014 M (95% Confidence Interval, C.I.: 0.013-0.016),
0.034 M (95% C.I.: 0.032-0.040) and 0.279 M (95% C.I.: 0.249-0.320). Note that the lex's for
PFOA were reported in molar (M) units, but the EPA judged the units were misreported and
were actually millimolar (mM) units. This judgement was based on the reported test
concentrations in Table 3.1 of the publication and the reported effect concentrations (ICx) would
not fall within this range unless the values were in mM units. Accordingly, the lex reported as
mM were converted to mg/L by multiplying the mM concentration by a molecular weight of
414.07 g/mol for PFOA. The calculated 96-hour IC10, IC25 and IC50 expressed as mg/L from the
study were 5.797, 14.07 and 115.5, respectively and acceptable for quantitative use.
E.2.6 Green alga. Rayhidocelis subcayitata
(formerly known as Selenastrum capricornutum and Pseudokirchneriella subcapitata)
Elnabarawy (1981) and 3M Company (2000a) provide the results of four separate
toxicity tests completed with the green alga, Raphidocelis subcapitata (formerly Selenastrum
capricornutum), and APFO (CAS # 3825-26-1). The toxicant was part of the 3M production lot
number 37 and was characterized as a mixture of APFO (96.5-100%) of the compound) and C6,
C7 and C9 perfluoro analogue compounds (0-3.5%> of the compound). The toxicity tests followed
a protocol modified from U.S. EPA-600/9-78-018 (1978) and ASTM-E-35.23 (1981). There
were four separate exposure regimes: 1) a four-day exposure + 10-day recovery period, 2) a
seven-day exposure + seven-day recovery period, 3) a 10-day exposure + four-day recovery
E-9
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period, and 4) a 14-day continuous exposure. Only the results of the continuous exposure are
presented here due to the confusion if reported effect concentrations are reported before or after
the recovery period. A bacteria-free culture of the alga was obtained from the USEPA (Corvallis,
OR) and stored in the dark until testing. Seven-day old stock cultures with an initial density of 1
x 104 cells/mL were placed in 250 mL flasks with 50 mL of test solution. There were three
replicates for each of the six nominal test concentrations (100, 180, 320, 560, 1000 and 1800
mg/L) and control. Nutrient medium was used as the dilution media for all test treatments and
were not renewed during the exposure. Algae were grown at 23 ฐC and continuously shaken at
100 rpm. The author-reported ECio, based on cell counts, was 5 mg/L for the 14-day exposure
and is acceptable for quantitative use.
Boudreau (2002) also performed a 96-hour static algal growth inhibition test on PFOA
(acid form, CAS # 335-67-1, 95% purity) with Raphidocelis subcapitata as part of the Master's
thesis at the University of Guelph, Ontario, Canada. The author stated that the algal growth
inhibition test with R. subcapitata similarly followed protocols found in ASTM E 1218-97a
(ASTM 1999) and Geis et al. (2000). R. subcapitata (UTCC 37 strain) used for testing were
obtained as slants from the University of Toronto Culture Collection (UTCC; Toronto, Canada).
Stock concentrations were prepared in laboratory-grade distilled water with a maximum
concentration that did not exceed the critical micelle concentration for PFOA of 450 mg/L.
Dilution medium was Bristol's algal growing media. Toxicity testing consisted of a range-finder
test and at least two definitive tests. Nominal test concentrations were 0 (negative control), 6.7,
12.5, 25, 50, 100, 200 and 400 mg/L. Tests were conducted in 60 x 15 mm polyethylene
disposable Petri dishes with 20 mL of test solution. Each Petri dish was inoculated with 1.5 x 104
cells/mL at initiation and replicated four times per test concentration. Tests were continuously
E-10
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illuminated with cool-white, fluorescent light between 3,800 and 4,200 lux and incubated at 23
ฑ1ฐC. Replicate Petri dishes were manually shaken twice a day during testing. Toxicity test
endpoints included cell density and chlorophyll-a content. The reported ICio, IC25 and IC50 based
on growth inhibition (measured as either chlorophyll-a or cell density) were 0.130 M (95% C.I.:
0.020-0.162), 0.197 M (95% C.I.: 0.166-0.231) and 0.298 M (95% C.I.: 0.274-0.317). As noted
above, although the lex for PFOA were reported in molar (M) units in the thesis, the EPA judged
the units were misreported and were actually millimolar (mM). This judgement was based on the
reported test concentrations in Table 3.1 of the publication and the reported effect concentrations
(ICx) would not fall within this range unless the values were in mM units. Accordingly, the lex
reported as mM were converted to mg/L by multiplying the mM concentration by a molecular
weight of 414.07 g/mol. The calculated 96-hour IC10, IC25 and IC50 expressed as mg/L from the
study were 53.83, 81.57 and 123.4, respectively and acceptable for quantitative use.
More recently, Xu et al. (2013) conducted a 96-hour static, measured algal growth
inhibition test on PFOA (acid form, >98% purity) with Raphidocelis subcapitata. Algae were
obtained from the Aquatic Organism Research Institute of the Chinese Academy of Science and
precultured for three generations prior to initiating the test. Dilution medium consisted of number
one culture medium supplemented with aquatic number four nutrient solution (Zhou and Zhang
1989). Algae in logarithmic growth phase (7.0 x 105 cells/mL) were inoculated in medium
containing nominal concentrations of PFOA at 0 (negative control), 30, 60, 90, 120, 150, 180,
240, 300 and 360 mg/L. Tests were conducted in 100 mL conical flasks with 50 mL of solution.
Each test concentration and the control were replicated three times. Exposure concentrations
were verified via UHPLC-MS/MS using the Agilent 1290 Infinity UHPLC system interfaced
with an Agilent 646-0 Triple Quadrupole mass. Algae were exposed under a 12-hour: 12-hour
E-ll
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light:dark cycle at 3,000-4,000 lux and 25ฐC. Chlorophyll concentration and permeability of cell
membranes was determined after 96 hours of exposure. The reported 96-hour growth EC50
(inhibition based on optical density) was 207.46 mg/L and was considered to be acceptable for
quantitative use.
Colombo et al. (2008) evaluated growth inhibition with Raphidocelis subcapitata on
ammonium perfluorooctanoate (APFO, the ammonium salt of PFOA, CAS # 3825-26-1, 99.7%
purity). Authors stated that the 96-hour algal growth inhibition test followed OECD test guidance
201 and European Commission directive 92/69/EEC. The source of R. subcapitata used for
testing was not reported, but presumably from an in-house culture as the medium reported to be
used for both culturing and testing was reconstituted water recommended via the French algae
test guideline (AFNOR T 90-304). The media differs slightly from the OECD recommended
media with regard to concentrations of P, N, and chelators. Stock solutions of APFO were
prepared by dissolving the test substance directly in the test media and diluting to provide a
geometric series of test concentrations. A range-finding and two definitive tests were conducted.
Definitive tests included six negative control replicates and three replicates at each PFOA
concentration. Tests were initiated via inoculation with 1 x 104 cells/mL from an algal culture in
log phase growth and carried out under continuous illumination with approximately 2,000 lux
and at 21-25ฐC. Test solutions were agitated to keep algae in suspension during the 96-hour
exposure and growth was determined at 24-hour intervals by counting an aliquot of test solution
from each replicate test chamber. Test concentrations measured in the second definitive algal test
and were 0 (negative control), 5.76, 11.37, 22.70, 46.33, 95.87, 180.67 and 369.67 mg/L. APFO
was determined as PFOA from a calibration curve of peak area against APFO concentrations in
standard solutions. The limit of quantification (LOQ) of the analytical method was 1 mg/L.
E-12
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Linearity was checked with a resulting coefficient of determination for the calibration curve of
greater than 0.999 in the range of 1-100 mg/L. Accuracy and precision were demonstrated by
analyzing six solutions containing nominal concentrations of 2.03 and 50.7 mg/L APFO in Milli-
Q water. The mean measured concentrations were 2.02 and 53.7 mg/L, respectively, with
calculated precision of 6% and 2% and accuracy of 99% and 106%, respectively. The reported
96-hour NOEC, based on biomass and growth rate, was 11.37 mg/L. The reported 96-hour
LOEC was 22.70 mg/L. The calculated MATC was 16.07 mg/L and was considered to be
acceptable for quantitative use.
E.2.7 Green alga. Scenedesmus obliguus
Hu et al. (2014) evaluated algal growth inhibition of PFOA (>96% purity) with
Scenedesmus obliquus in both a 96-hour and eight-day static unmeasured exposures. Authors
stated that the tests followed OECD test guidance 201 (OECD 2006). S. obliquus were supplied
by UTEX Culture Collection of Algae, University of Texas at Austin. Dilution medium was HB-
4 media adjusted to a pH of 6.8. Algae in exponential growth phase were exposed to nominal
concentrations of 0 (negative control), 1, 3.16, 10, 31.6, 100, 316 and 1,000 mg/L in the 96-hour
exposure, and 0, 5, 10, 20 and 40 mg/L in the eight-day exposure. Experiments were initiated by
inoculating equal cell numbers of 5 x 103 cells/mL in the 96-hour exposure and 5 x 106 cells/mL
in the eight-day exposure into 250 mL flasks containing a total volume of 100 mL of algal cell
suspension per flask. There were five replicates per treatment in the 96-hour exposure and three
replicates in the eight-day exposure. Algae were incubated at 25ฐC under cool-white
fluorescence lights at 85-90 |imol photons/[m2 x s] irradiance with a 16-hour:8-hour light:dark
cycle. The 96-hour growth EC so (inhibition based on optical density) was 44.0 mg/L. The eight-
day NOEC based on cell number was 40 mg/L (the highest test concentration). The plant values
from the study were acceptable for quantitative use.
E-13
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E.2.8 Green alga. Scenedesmus quadricauda
Yang et al. (2014) conducted a 96-hour renewal, measured test on the growth effects of
PFOA (acid form, CAS #335-67-1, 99% purity) with the green alga, Scenedesmus quadricauda.
Algae were obtained from in-house cultures originally supplied by the Chinese Research
Academy of Environmental Sciences. The algae used for testing were inoculated at a cell density
equal to 2.0 x 104 cells/mL in 50 mL beakers. PFOA was dissolved in deionized water and
DMSO (amount not provided) and then diluted with M4 medium. Algae in logarithmic growth
phase were exposed to 0 (solvent control), 80.00, 128.00, 204.80, 327.68, 524.29 and 838.86
mg/L. Each treatment was replicated three times. Measured concentrations ranged from 75.68
mg/L (before renewal) to 78.8 mg/L (after renewal) in the lowest treatment, and from 764.13
(before renewal) to 831.45 mg/L (after renewal) in the highest treatment. The experiments were
conducted at 22 ฑ 2ฐC with a 12-hour: 12-hour light: dark cycle. The initial pH of the test solution
was 7.0 ฑ 0.5, total hardness was 190 ฑ0.1 mg/L as CaCCb, and total organic carbon was 0.02
mg/L. Algae concentrations in the beakers were measured daily with a microscope. The 96-hour
growth inhibition EC so was reported as 269.63 mg/L and was acceptable for quantitative use.
E.2.9 Duckweed. Lemna minor
Wu et al. (2023) tested perfluorooctanoic acid (PFOA) on duckweed, Lemna minor, for
96 hours in a static, measured experiment. PFOA (>97.6% purity) was obtained from Dr.
Ehrenstorfer GmbH (Augsburg, Germany). DMSO solvent (> 99.9% purity) was obtained from
Merck (Germany). Duckweed was obtained from in house cultures that had been grown in a
modified Swedish Standards Institute (SIS) medium. Duckweed cultures were housed in 150 mm
petri dishes with 100 mL SIS medium that was changed every two weeks. SIS medium pH was
adjusted to 6.5 by using NaOH or HC1. Plants were cultured at 25ฑ1ฐC and 60% humidity under
a 12:12 light dark cycle at 2,000 lux. Duckweed was precultured for 1 week in clean SIS media
E-14
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for seven days prior to testing. Duckweed experiments were conducted in 6-well poly-propylene
plates to avoid PFOA sorption to container walls. Each replicate well contained 10 mL of test
material and two colonies approximately the same size of a 3-frond L. minor. Nominal test
concentrations were 0 (control), 0.001, 0.1, and 10 mg/L PFOA. All test chambers included
DMSO solvent, and each treatment had three duplicates. PFOA concentrations measured on day
0 were 0.93ฑ0.04, 97.2ฑ0.89, and 9,478ฑ4.74 |ig/L. PFOA in solvent controls were not reported
for day 0 but was below detection levels when measured after 96 hours. The number of fronds in
each treatment well were counted after 48 and 96 hours. In addition, Fourier-transform infrared
spectroscopy (FTIR) was performed on a subset of fronds from each treatment at the end of the
96-hour exposure to examine responses to PFOA at the biochemical level. Statistically
significant (p<0.05) differences between treatment groups were assessed with one-way ANOVA
followed by Dunnett's tests using SPSS Statistics 26. No statistically significant differences in
frond number were observed. However, FTIR analysis revealed structural and functional
alterations in response to PFOA at the biochemical level. The reported NOEC of 9.478 mg/L for
population growth rate was determined to be acceptable for quantitative use.
E.2.10 Watennilfoil. Myrioyhyllum sy.
Hanson et al. (2005) conducted a 35-day microcosm study on PFOA (sodium salt
donated by 3M Co., purity not provided) with the submerged watermilfoils, Myriophyllum
spicatum and M. sibiricum. The study was conducted in 12,000 L outdoor microcosms at the
University of Guelph Microcosm Facility located in Ontario, Canada using in-house cultures of
Myriophyllum spp. Each microcosm was below ground and was flush with the surface. Plastic
trays filled with sediment (1:1:1 mixture of sand, loam and organic matter, mostly manure) were
placed in the bottom of each microcosm. The total carbon content of the sediment was 16.3%.
Ten apical shoots, 4 cm in length, from in-house cultures using the same sediment were
E-15
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transferred to each microcosm, with three separate microcosms used for each treatment (nominal
concentrations 0, 0.3, 10, 30 and 100 mg/L). Endpoints of toxicity that were monitored on days
14, 21 and 35 of the study included growth in plant length, root number, root length, longest root,
node number, wet mass, dry mass and chlorophyll-a and -b content. PFOA treatments were
dissolved in the same water (well water) used to supply the microcosms. Results showed that
measured concentrations remained similar to nominal concentrations throughout the entire
exposure period and did not change appreciably over the course of the study. The time-weighted
average measured concentrations were 0 (negative control), 0.27, 0.65, 23.9 and 74.1 mg/L.
Water quality over the length of the 35-day microcosm experiment was: dissolved oxygen: 7.3-
8.5 mg/L; temperature: 17.8-22.0ฐC; pH: 8.3-8.7; total hardness: 217.5 mg/L as CaCC>3. The
light:dark cycle was outdoor ambient cycles beginning June 13, 2000 (Guelph, Ontario). The
watermilfoil species were equally sensitive to PFOA. The 35-day ECio (based on weight) was
21.6 mg/L for M. sibiricum and 19.7 mg/L for M. spicatum. The plant values were acceptable for
quantitative use.
E.2.11 Lettuce. Lactuca sativa
Ding et al. (2012b) conducted a microcosm study where water lettuce, Lactuca sativa,
was exposed to PFOA (CAS# 335-67-1, 96% purity) for 5 days. Authors stated the test protocols
followed U.S. EPA (1996). Test vessels were plastic and test solutions were static and
unmeasured. The test employed six exposure concentrations and a negative control, with each
treatment being replicated three times. Petri dishes containing lettuce seeds were placed in a
plant test chamber with a constant room temperature of 18ฐC ฑ2ฐC and a photoperiod or 16-
hours light and 8-hours dark. After 5 days, the number of germinated seeds was counted, and the
length of the roots was measured with a ruler to the closest millimeter. The author reported EC so
E-16
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(endpoint = root elongation) was 1.801 mM, which was converted to 745.7 mg/L and was
retained for quantitative use.
E-17
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Appendix F Acceptable Estuarine/Marine Plant PFOA Toxicity Studies
Species
Method'
losl
Dui'iilioii
( homiciil
/ Piiriu
pll
Temp.
(ฐC)
Siiliniit
(|)|)D
IIIWl
Uc'INirk'd r.lTwl
( oiicoiilriilioii
(mป/l.)
Reference
Green alga,
Chlorella sp.
s,u
96 hours
PFOA
>98%
23
30b
ECso
(population abundance)
127.35
Mao et al. 2023
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, NR=not reported
b Salinity of Erdschreiber's medium
F-l
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F.l Summary of Quantitatively Acceptable Plant PFOA Toxicity Studies
F. 1.1 Green alga. Chlorella sp.
Mao et al. (2023) tested perfluorooctanoic acid (PFOA) on an estuarine/marine Chlorella
sp. For seven days in a static, unmeasured experiment. PFOA (>98% purity) was obtained from
Tokyo Chemical Industry Co. Ltd. Algae were obtained from the Freshwater Algae Culture
Collection at the Institute of Hydrobiology in Wuhan, China. Algae were obtained from the
Algae Culture Collection at the Institute of Hydrobiology in Wuhan, China. Algae were cultured
in Erdschreiber medium within a conical flask inside an illumination incubator at 23ฑ1ฐC under
a 12:12 light:dark cycle at 5,000 lux. Algae were shaken three times per day to prevent sticking
to the sides of the flask, and inoculated once every two weeks to maintain optimal growth.
Chlorella sp. In the exponential growth phase were added to test vessels containing test solution
at a density of approximately 5.0xl04 cells/mL. The experimental design consisted of nominal
concentrations of 0 (control), 5, 10, 20, 40, 80, 160, and 320 mg/L PFOA. Testing protocols
followed OECD guidelines, and all experiments were performed in triplicate. Statistical analysis
included one-way ANOVA, using SPSS version 26 software. Algal cell density and size were
measured daily, and algal growth inhibition was calculated using the equation provided in the
OECD guidelines. Chlorophyll a, maximum quantal yield, cell membrane integrity, esterase
activity relative to control, relative electron transfer rate, and reactive oxygen species were
reported after 1, 3, 5, and 7 days, respectively. Algae exhibited maximum growth at 20 mg/L, but
growth significantly declined at 80 mg/L and higher concentrations. Increasing PFOA
concentrations also inhibited chlorophyll a, and increased oxidative stress. The 96-hour EC so for
algal growth inhibition was 127.35 mg/L, and was determined to be acceptable for quantitative
use.
F-2
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Appendix G Other Freshwater PFOA Toxicity Studies
G.l Summary Table of Qualitative Freshwater PFOA Toxicity Studies
Species (lil'cs(;iปc)
Method"
lost Dui'iiiion
( homiciil /
Pn ril
pll
Temp.
(ฐC)
r.ricci
Chronic
limits
(NOI'.C-
i.or.ci
(111 Si/I.)
Reported
I'.ITecl
(one.
(mti/l.)
Deficiencies
Reference
Cyanobacteria,
Anabaena sp.
S,M
24 hours
PFOA
96%
-
ECso
(bioluminescence
inhibition)
-
19.81
Duration too
short for a plant
test, missing
some exposure
details, non-
apical endpoint
Rodea-Palomares
et al. 2012
Cyanobacteria,
Anabaena sp.
s,u
24 hours
PFOA
96%
7.8
28
ECso
(bioluminescence
inhibition)
-
78.88
Duration too
short for a plant
test, missing
some exposure
details, non-
apical endpoint
Rodea-Palomares
et al. 2015
Blue green alga,
Synechocystis sp.
S,M
4 days
PFOA
96%
7.5
30
NOEC
(population abundance)
1->1
1
Duration, focus
of study was
PFOA removal
Marchetto et al.
2021
Blue green alga,
Synechocystis sp.
F, M
12-15 days
PFOA
96%
7.5
30
NOEC
(biomass)
1->1
1
Duration, focus
of study was
PFOA removal
Marchetto et al.
2021
Green alga,
Raphidocelis
subcapitata
(formerly, Selenastrum
capricornutum)
S,U
96 hours
PFOA
96.5-100%
2.3-
10.3
-
ECso
(cell density and growth
rate)
-
90
Possible mixture
effects of other
perfluoro
homologue
compounds and
the amount of
isopropanol, wide
pFl range
3M Company
2000a
Green alga,
Raphidocelis
subcapitata
S,U
96 hours
(+ 10-day
recovery period)
APFO
96.5-100%
-
23
ECio
(cell count)
-
5.3
Unclear if
chronic value
determined
before or after
recovery period
3M Company
2000a/Elnabarawy
1981
Green alga,
Raphidocelis
subcapitata
S,U
7 days
(+ 7-day
recovery period)
APFO
96.5-100%
-
23
ECio
(cell count)
-
3.3
Unclear if
chronic value
determined
before or after
recovery period
3M Company
2000a/Elnabarawy
1981
Green alga,
Raphidocelis
subcapitata
S,U
10 days
(+ 4-day
recovery period)
APFO
96.5-100%
-
23
ECio
(cell count)
-
2.9
Unclear if
chronic value
determined
before or after
recovery period.
3M Company
2000a/
Elnabarawy 1981
G-l
-------
( hrunic
l.imils
Reported
(NOI'.C-
HITccl
( hem ic;il /
Temp.
i.or.C)
(one.
Species (lil'cs(;iiic)
Method'
lost Dui'iiiion
Pn ril
pll
(ฐC)
r.nvci
(111 Si/I.)
(iiiii/l.)
Deficiencies
Reference
Test substance
was not
characterized but
considered a
mixture of APFO
and other PFAS
impurities; test
substance purity
unknown and not
used
Green alga,
Raphidocelis
subcapitata
APFO
Unknown
ECso
(cell count)
quantitatively
3M Company.
2000a
S,U
96 hours
1,980
because other
tests from 3M
Company
(2000a) indicated
test purity as low
as 78%. 3M
Company 2000a
also indicated
other tests with
green alga lacked
information on
test substance
purity.
Duration too
Green alga,
PFOA
96%
ECso
(growth)
short for a plant
Raphidocelis
S,M
72 hours
-
21-24
-
96.2
test, missing
Rosal et al. 2010
subcapitata
some exposure
details
Duration too
Green alga,
Raphidocelis
subcapitata
S,U
4.5 hours
PFOA
96%
-
-
ECso
(photosynthetic
efficiency)
-
748.2ฐ
short for a plant
test, missing
some exposure
details, non-
apical endpoint
Ding et al. 2012b
Green alga
(104 cells/mL),
Scenedesmus obliquus
S,U
72 hours
PFOA
Unreported
7.5
22
NOEC
(growth rate)
-
828.1ฐ
Duration too
short for a plant
test
Liu et al. 2008a
Culture water not
Duckweed,
Lemna gibba
S,U
7 days
PFOA
95%
-
-
IC50
(wet weight)
-
79.92
characterized,
missing some
Boudreau 2002
exposure details
Protozoa,
S,U
2 hours
PFOA
7.2
25
EC50
713.9ฐ
Single cell
Lim 2022
Tetrahymena pyriformis
Unreported
(population abundance)
organism
G-2
-------
Species (lil'es(;iปc)
Method'
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
l-llccl
(one.
(inti/l.)
Deficiencies
Reference
Protozoa,
Tetrahymena pyriformis
S,U
96 hours
PFOA
Unreported
7.2
25
EC50
(population abundance)
-
65.1
Single cell
organism
Lim 2022
Tubificid worm
(0.03 g, 0.8 cm),
Limnodrilus
hoffmeisteri
S,M
96 hours
PFOA
99%
7
22
LC50
-
568.20
Atypical source
of organisms
Yang et al. 2014
Planaria (0.9 cm),
Dugesia japonica
S,U
96 hours
PFOA
>98%
-
25
LC50
-
427.7
Poor
concentration-
response curve
Li 2009
Planaria (10-12 mm),
Dugesia japonica
R,U
96 hours
PFOA
96%
-
20
LC50
-
39.35
Atypical source
of the test
organisms
Yuanetal. 2015
Planarian,
Dugesia japonica
R,U
10 days
PFOA
96%
-
20
LOEC
(decrease mRNA
expression levels of
neural genes DjFoxD,
DjotxA and DjotxB)
<0.5-0.5
0.5
Duration too long
for an acute test
and too short for
a chronic test,
non-apical
endpoint
Yuanetal. 2016b
Planarian,
Dugesia japonica
S,U
10 days
PFOA
96%
-
20
LOEC
(elevated lipid
peroxidation; increased
mRNA expression
levels of HSP 40 and
HSP 70)
<0.5-0.5
0.5
Duration too long
for an acute test
and too short for
a chronic test,
non-apical
endpoint
Yuanetal. 2017
Planarian,
Dugesia japonica
R, U
10 days
PFOA
Unreported
-
25
LOEC
(growth, behavioral,
genetic and
biochemistry changes)
-
15
Duration, effects
no longer seen
after 8-day
recovery period
Zhang et al. 2020
Planarian,
Dugesia japonica
R, U
10 days
PFOA
Unreported
-
-
LOEC
(genetic and
biochemistry changes)
-
15
Duration too long
for an acute test
and too short for
an acute test,
atypical
endpoints
Zhang et al. 2022
Chinese pond mussel
(1 year),
Sinanodonta woodiana
(formerly, Anodonta
woodiana)
S,U
48 hours
PFOA
Unreported
7.0
24
LC50
-
192.08
Missing some
exposure details,
duration
Xia et al. 2018
G-3
-------
Species (lil'cs(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
( limnic
limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITcct
(one.
(iiiii/l.)
Deficiencies
Reference
Snail (adult, 1 yr),
Sinotaia aeruginosa
(formerly, Bellamya
aeruginosa)
Sediment
21 days
PFOA
>96%
7.98-
8.02
24
LOEC
(biochemistry and
enzyme changes)
-
0.00276 -
0.00286
Sediment
exposure,
missing exposure
details
Xiang et al. 2021
Mud snail (4.0 g, 2.0
cm)
Cipangopaludina
cathayensis
S, M
96 hours
PFOA
99%
7
22
LC50
-
740.07
Atypical source
of organisms
Yang et al. 2014
Rotifer
(<2-hour old neonates),
Brachionus calyciflorus
R, Ud
4 days
PFOA
96%
-
20
EC10
(resting egg production)
0.125-
0.25
0.1768
(EPA-
Calculated
ec10:
0.07758)
Only one
replicate
Zhang et al. 2014b
Cladoceran,
Daphnia magna
s,u
48 hours
PFOA
96.5-100%
7.5-
8.4
19.4-
20.2
EC50
(death/immobility)
-
360
Possible mixture
effects of other
perfluoro
homologue
compounds and
the amount of
isopropanol;
missing exposure
details
3M Company
2000a
Cladoceran,
Daphnia magna
s,u
48 hours
APFO
96.5-100%
-
-
EC50
(death/immobility)
-
>1,000
Possible mixture
effects of other
perfluoro
analogue
compounds;
<50% effect in
highest test
concentration;
Authors indicate
test may have
included food
3M Company
2000a
Cladoceran,
Daphnia magna
s,u
48 hours
APFO
96.5-100%
-
-
EC50
(death/immobility)
-
126
Possible mixture
effects of other
perfluoro
analogue
compounds;
authors indicate
test may have
included food
3M Company
2000a
G-4
-------
Chronic
limits
Reported
(NOI'.C-
r.lTeel
( hem ic;il /
Temp.
i.or.C)
(one.
Species (lilVsliiuo)
Method"
lost Dui'iilion
Pn ril
pll
(ฐC)
r.nvci
(111 Si/I.)
(iiiii/l.)
Deficiencies
Reference
Possible mixture
Cladoceran (<24 hours
old),
Daphnia magna
s,u
48 hours
APFO
78-93%
8.0-
8.1
21
ECso
(death/immobility)
-
221
effects of other
perfluoro
analogue
compounds; low
test substance
purity (i.e., 78%).
3M Company
2000a
Test substance is
considered a
mixture of APFO
and other
Cladoceran (<24 hours
old),
Daphnia magna
s,u
48 hours
APFO
Unreported
OO 00
'u> ~
20.3-
20.8
LC50
-
1,200
impurities; test
substance purity
unknown and not
used
quantitatively
because other
tests from 3M
Company
(2000a) indicated
test purity as low
as 78%.
3M Company
2000a
Possible mixture
effects of the
inert
Cladoceran (<24 hours
old),
Daphnia magna
s,u
48 hours
APFO
Unreported
00 00
r
19.5-
20.1
EC50
(death/immobility)
-
584
perfluorinated
compounds and
other perfluoro
analogue
compounds; test
substance purity
unknown and not
used
quantitatively
because other
tests from 3M
Company
(2000a) indicated
test purity as low
as 78%.
3M Company
2000a
Cladoceran,
Daphnia magna
R,U
48 hours
APFO
96.5-100%
-
-
EC50
(death/immobility)
-
266
Missing test
details
3M Company
2000a
Cladoceran,
Daphnia magna
R,U
21 days
APFO
96.5-100%
-
-
MATC
(survival and
reproduction)
22-36
28.14
Missing test
details
3M Company
2000a
G-5
-------
Species (lil'cs(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.ricci
Chronic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Cladoceran (<24 hours
old),
Daphnia magna
R,U
21 days
PFOA
>99%
7.5
23
LOEC
(fecundity)
<0.4141-
0.4141
0.4141ฐ
Chronic
responses in this
test did not
display
concentration-
dependent effects
beyond the
LOEC despite a
25X increase in
treatment
concentrations
Seyoum et al.
2020
Cladoceran
(adult, ~14 days),
Daphnia magna
S,M
48 hours
APFO
98%
-
21
MATC
(arginine, lysine, uridine
G-6mountG-6ation)
27.88-
46.33
35.94
Non-apical
endpoint
Labine et al. 2022
Amphipod (7 days),
Hyalella azteca
R, M
7 days
PFOA
99.5%
7.31
(6.85-
7.76)
22.6
(21.9-
23.2)
EC20
(growth - biomass)
-
5.0
Duration too
short for a
chronic exposure
Kadlec et al. 2024
Oriental river prawn
(0.30 g, 4.0 cm),
Macrobrachium
nipponense
S,M
96 hours
PFOA
99%
7
22
LC50
-
366.66
Atypical source
of organisms
Yang et al. 2014
Midge (larva, 10 days
old),
Chironomus dilutus
R, U
10 days
PFOA
>97%
-
23
NOEC
(survival and growth)
-
100
Lack of effects in
all concentrations
tested; Range-
finding
experiment only
with few details
reported in the
publication;
Primary focus of
the publication
was on PFOS
MacDonald et al.
2004
Midge (larva, 10 days
old),
Chironomus dilutus
S,M
10 days
PFOA
97%
-
-
MATC
(mortality)
26-272
84.10
Range-finding
experiment only;
results of the
definitive 19-day
test by McCarthy
et al. 2021 were
used
preferentially
over this test.
McCarthy et al.
2021
Midge (Instar, 3 days),
Chironomus dilutus
R, M
7 days
PFOA
99.5%
6.96
(6.77-
7.19)
22.7
(22.2-
23.0)
EC20
(growth - biomass)
-
70.6
Duration too
short for a
chronic exposure
Kadlec et al. 2024
G-6
-------
Species (lil'es(;iปc)
Method'
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Midge (Instar, 3 days),
Chironomus dilutus
R, M
7 days
PFOA
99.5%
6.96
(6.86-
7.11)
23.3
(23.0-
23.8)
EC20
(growth - biomass)
-
94.0
Duration too
short for a
chronic exposure
Kadlec et al. 2024
Midge (Instar, 3 days),
Chironomus dilutus
R, M
7 days
PFOA
99.5%
7.39
(7.26-
7.60)
22.9
(20.3-
24.0)
EC20
(growth - biomass)
-
94.2
Duration too
short for a
chronic exposure
Kadlec et al. 2024
Midge
(multi-generational),
Chironomus riparius
S,M
-20-38 days /
generation
PFOA
Pure
(unspecified)
7.8-
8.2
20
NOEC
(emergence,
reproduction, sex ratio)
-
0.0089
Only one
exposure
concentration,
static chronic
exposure
Stefani et al. 2014
Midge
(multi-generational),
Chironomus riparius
S,M
-20-38 days /
generation
PFOA
Pure
(unspecified)
7.8-
8.2
20
NOEC
(increased mutation
rate)
-
0.0089
Only one
exposure
concentration,
static chronic
exposure
Stefani et al. 2014
Midge (larva, 1st instar),
Chironomus riparius
S,M
-1 yearb
PFOA
Unreported
7.5-
8.2
20.1
LOEC
(F10 developmental
time, adult weight,
exuvia length)
<0.0098-
0.0098
0.0098
Only one
exposure
concentration,
static chronic
exposure, low
control survival
in 4 of the 10
generations
Marziali et al.
2019
Midge (0.05 g, 1.2 cm),
Chironomus plumosus
S,M
96 hours
PFOA
99%
7
22
LC50
-
402.24
Atypical source
of organisms
Yang et al. 2014
Midge
(larva, instar),
Chironomus plumosus
S,M
10.33 days
PFOA
96%
-
25
NOEC
(mortality)
-
9.8 ng/g
(sediment)
10-day
subchronic
sediment
exposure and no
effects at the
highest
concentration
tested resulting in
a relatively low
greater than value
that does not
inform species
sensitivity
Zhai et al. 2016
Rainbow trout
(juvenile, 40-50 mm),
Oncorhynchus mykiss
S,U
96 hours
APFO
99.7%
6.0-
8.5
13-17
LC50
-
10T
High ammonia
concentration in
treatment
solutions
Colombo et al.
2008
G-7
-------
Species (lil'cs(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Rainbow trout (fry),
Oncorhynchus mykiss
Diet, U
70 days
PFOA
Unreported
14
MATC
(liver somatic index)
200-1,800
(mg/kg)
600
(mg/kg
diet)
Non-apical
endpoint
Tilton et al. 2008
Rainbow trout (fry),
Oncorhynchus mykiss
Diet, U
6 months
PFOA
Unreported
-
14
MATC
(palmitoyl CoA P-
oxidation - liver
enzyme)
200-1,800
(mg/kg)
600
(mg/kg
diet)
Non-apical
endpoint
Tilton et al. 2008
Rainbow trout
(juvenile),
Oncorhynchus mykiss
Diet, U
15 days
PFOA
Unreported
-
12
MATC
(increase plasma
vitellogenin)
5-50
(mg/kg)
15.81
(mg/kg
diet)
Test design and
lack of exposure
details
Benninghoff et al.
2011
Rainbow trout (fry, 10-
15 weeks old),
Oncorhynchus mykiss
Diet, U
6 months
PFOA
Unreported
-
12
LOEC
(increase tumor
multiplicity and size)
<2,000-
2,000
2,000
(mg/kg
diet)
Test design and
lack of exposure
details
Benninghoff et al.
2012
Rainbow trout
(oocyte, ova),
Oncorhynchus mykiss
S,M
3 hours
PFOA
>97%
-
6
NOEC
(accumulation residue)
-
10.58
Duration too
short for an acute
test
Raine et al. 2021
Rainbow trout
(oocyte, ova),
Oncorhynchus mykiss
S,M
3 hours
PFOA
>97%
8.5
6
NOEC
(accumulation residue)
-
10.11
Duration too
short for an acute
test
Raine et al. 2021
Rainbow trout
(juvenile, 7 months),
Oncorhynchus mykiss
R, M
10 days
PFOA
Unreported
-
12
LOEC
(7-Benzyloxy-4-
trifluoromethylcoumarin
O-debenzylase levels)
-
0.0005
Test was too long
for an acute
exposure and too
short for a
chronic exposure;
non-apical
endpoint
Solan et al. 2022
Atlantic salmon
(embryo-larval),
Salmo salar
F, U
52 days
PFOA
95%
-
5-7
NOEC-LOEC
(growth - weight and
length)
0.1->0.1
>0.1
No effects at the
highest
concentration
tested resulting in
a relatively low
greater than value
that does not
inform species
sensitivity
Spachmo and
Arukwe 2012
Atlantic salmon
(embryo),
Salmo salar
R, U
49 days
Perfluoro-n-
octanoic
acid
Unreported
-
5-7
NOEC
(growth - weight and
length)
0.1->0.1
>0.1
Lack of treatment
concentrations
and lack of
effects observed
at a relatively
low NOEC
Arukwe et al.
2013
G-8
-------
Species (lil'cs(;iiic)
Method'
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
( hrunic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Goldfish (6.0 g, 7.0
cm),
Carassius auratus
S,M
96 hours
PFOA
99%
7
22
LC50
-
606.61
Atypical source
of organisms
Yang et al. 2014
Goldfish
(juvenile, 27.85 g),
Carassius auratus
S,M
96 hours
PFOA
>98%
7.25
23
Antioxidant
enzyme activity
-
>4.931ฐ
Only two
exposure
concentrations,
non-apical
endpoint
Feng et al. 2015
Goldfish
(juvenile, 12 months),
Carassius auratus
F, M
7 days
PFOA
>96%
9
-
LC50
-
>471.5
Test was too long
for an acute
exposure and too
short for a
chronic exposure
Dong et al. 2023
Goldfish
(juvenile, 12 months),
Carassius auratus
F, M
7 days
PFOA
>96%
9
-
LC50
-
>456.3
Test was too long
for an acute
exposure and too
short for a
chronic exposure
Dong et al. 2023
Goldfish
(juvenile, 12 months),
Carassius auratus
F, M
7 days
PFOA
>96%
9
-
LC50
-
>411.7
Test was too long
for an acute
exposure and too
short for a
chronic exposure
Dong et al. 2023
Goldfish
(juvenile, 12 months),
Carassius auratus
F, M
7 days
PFOA
>96%
9
-
LC50
-
>433.2
Test was too long
for an acute
exposure and too
short for a
chronic exposure
Dong et al. 2023
Goldfish
(juvenile, 12 months),
Carassius auratus
F, M
7 days
PFOA
>96%
9
-
LC50
-
>451.1
Test was too long
for an acute
exposure and too
short for a
chronic exposure
Dong et al. 2023
Goldfish
(juvenile, 12 months),
Carassius auratus
F, M
7 days
PFOA
>96%
9
-
LC50
-
>440.1
Test was too long
for an acute
exposure and too
short for a
chronic exposure
Dong et al. 2023
Common carp
(juvenile, ~12 cm, ~20
g)>
Cyprinus carpio
F, M
96 hours
PFOA
99.8%
6.9
23
LOEC
(vitellogenin (VTG)
activity)
-
6.582
Broad range of
test treatments,
non-apical
endpoint
Kim et al. 2010
G-9
-------
Species (lilVsliiuo)
Method"
Tcsl Dui'iilion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
r.lTeel
(one.
(iiiii/l.)
Deficiencies
Reference
Common carp
(adult - 2 years old),
Cyprinus carpio
F, U
(tissue)
56 days
PFOA
96%
6.7-
8.0
10-15
LOEC
(PCNA-positive
hepatocyte abundance)
0.0002-2
2
Poor test design,
only two
exposure
concentrations,
non-apical
endpoint
Giari et al. 2016
Common carp
(adult - 2 years old),
Cyprinus carpio
F, U
(tissue)
56 days
PFOA
96%
6.7-
8.0
10-15
LOEC
(liver biomarkers)
0.0002-2
2
Only two
exposure
concentrations,
non-apical
endpoint
Manera et al. 2017
Common carp
(adult - 2 years),
Cyprinus carpio
F, U
56 days
PFOA
Unreported
6.7-
8.0
10-15
LOEC
(number of rodlet cells)
-
0.0002
Non-apical
endpoint
Manera et al.
2022b
Common carp
(adult - 2 years),
Cyprinus carpio
F, U
56 days
PFOA
Unreported
-
-
LOEC
(histological effects)
-
0.0002
Non-apical
endpoint
Manera et al.
2022a
Common carp,
Cyprinus carpio
R, M
14 weeks
PFOA
Unreported
7.71
21.8
NOEC
(mortality, length and
weight)
-
>0.0921
Greater than low
value
Petre et al. 2023
Zebrafish (embryo),
Danio rerio
R,U
96 hours
APFO
98%
-
26
LC50
-
386.3ฐ
Inability to verify
LC50
Ding et al. 2012c,
2013
Zebrafish (embryo),
Danio rerio
S,U
72 hours
PFOA
95%
8.3
28.5
LC50
-
262
Duration too
short for acute
test
Zheng et al. 2012
Zebrafish (embryo),
Danio rerio
S,U
96 hours
PFOA
95%
8.3
28.5
EC50
(malformation)
-
198
Non-apical
endpoint
Zheng et al. 2012
Zebrafish
(embryo, 4 hpf),
Danio rerio
R, U
120 hours
PFOA
Unreported
-
28
NOEC-LOEC
(increase relative
mRNA expression of
hhex and pax 8)
0.1-0.2
-
Duration too
short for a
chronic test and
too long for an
acute test, non-
apical endpoint,
only three
exposure
concentrations
Du et al. 2013
Zebrafish (adult),
Danio rerio
R, U
(tissue)
28 days
PFOA
96%
-
26
NOEC
(reproduction:
fecundity, fertility and
hatching)
1->1
1
Not a true ELS
test
Hagenaars et al.
2013
Zebrafish (adult),
Danio rerio
R, U
(tissue)
28 days
PFOA
96%
-
26
LOEC
(alterations of gene
transcripts)
<0.1-0.1
0.1
Not a true ELS
test, non-apical
endpoint
Hagenaars et al.
2013
G-10
-------
Species (lil'es(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITecl
(one.
(iiiii/l.)
Deficiencies
Reference
Zebrafish
(embryo, 4 cell stage),
Danio rerio
s,u
Fertilization to
144 hours post-
fertilization
(6 days)
PFOA
Unreported
7.2-
7.6
26
ECso
(lethal and sublethal
endpoint)
-
350
Static chronic
exposure
Ulhaq et al. 2013
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hours
APFO
Unreported
-
-
LOEC
(mortality)
<0.02759-
0.02759
0.02759ฐ
Duration too long
for acute test
Truong et al. 2014
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hours
PFOA
Unreported
-
-
NOEC
(mortality)
26.50-
>26.50
26.50ฐ
Duration too long
for acute test
Truong et al. 2014
Zebrafish (adult),
Danio rerio
R,U
21 days
PFOA
96%
-
27
MATC
(decrease in
inflammatory cytokines
(IL-lfi and IL-21) in
spleen)
0.05-0.1
0.0707
Duration too
short for a
chronic test, non-
apical endpoint
Zhang et al. 2014a
Zebrafish
(embryo, 2 days pf),
Danio rerio
s,u
72 hours
PFOA
96%
-
28.5
LC50
-
157.3ฐ
Duration too
short for an acute
test
Kalasekar et al.
2015
Zebrafish (embryo),
Danio rerio
s,u
72 hours
PFOA
Unreported
-
26
NOEC
(embryo toxicity)
-
132.5ฐ
Only one
exposure
concentration and
duration too short
for an acute test
Bouwmeester et
al. 2016
Zebrafish
(gastrula stage, 4.5
hpf),
Danio rerio
R,U
Embryo
development to
28 days post-
hatch
PFOA
Unreported
-
28
LOEC
(swim bladder
development)
-
4.7
Unconventional
test design, diet
and water
concentrations
were not
measured
Godfrey et al.
2017b
Zebrafish (3 hpf),
Danio rerio
s,u
5-day +
9-day
observation
PFOA
Unreported
7.2-
7.7
26-28
MATC
(growth - total body
length, interocular
distance, yolk sac area)
0.2-2
0.6325
Duration too long
for an acute test,
Jantzen et al.
2017a
Zebrafish (3 hpf),
Danio rerio
s,u
5-day +
9-day
observation
PFOA
Unreported
7.2-
7.7
26-28
MATC
(swimming activity -
distance traveled)
0.02-0.2
0.06325
Duration too long
for an acute test,
non-apical
endpoint
Jantzen et al.
2017a
Zebrafish
(embryo, 72 hpf),
Danio rerio
S,M
48 hours
PFOA
Unreported
-
27
LC50
-
>500
Duration too
short for an acute
test
Rainieri et al.
2017
Zebrafish
(embryo, <2 hpf),
Danio rerio
S,U
96 hours
PFOA
96%
-
28
LC50
-
<50
Results only
reported
graphically,
control mortality
not reported
Weiss-Errico et al.
2017
G-ll
-------
Species (lil'es(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITecl
(one.
(iiiii/l.)
Deficiencies
Reference
Zebrafish
(embryo, 3 hpf),
Danio rerio
s,u
117 hours + 9
days observation
PFOA
Unreported
7.2-
7.7
27
MATC
(morphology)
0.008281-
0.08281
0.02619
Duration too long
for an acute test
Annunziato 2018
Zebrafish (embryo, 6
hpf),
Danio rerio
R, M
7 days
PFOA
95%
7.8
25
LOEC
(aromatase, CYP19A1
mRNA, estrogen
receptor alpha mRNA,
17-beta estradiol,
androgen binding
protein)
<0.4969-
0.4969
0.4969ฐ
Duration too long
for an acute test
and too short for
a chronic test,
non-apical
endpoint
Chen et al. 2018
Zebrafish (embryo),
Danio rerio
S,U
168 hours
PFOA
Unreported
-
-
LC50
-
362.5
Duration too long
for an acute test
Stinckens et al.
2018
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hours
PFOA
Unreported
-
-
Benchmark Dose at
10% extra effect
(abnormal development)
-
27.75ฐ
Duration too long
for an acute test
and too short for
a chronic test,
non-apical
endpoint
Thomas et al.
2019
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
18 hours
APFO
Unreported
-
-
Benchmark Dose at
10% extra effect
(immobility)
-
27.59ฐ
Duration too long
for an acute test,
non-apical
endpoint
Thomas et al.
2019
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hours
APFO
Unreported
-
-
Benchmark Dose at
10% extra effect
(mortality)
-
24.50ฐ
Duration too long
for an acute test
and too short for
a chronic test,
non-apical
endpoint
Thomas et al.
2019
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
118 hours
PFOA
>96%
-
28
EC50
-
210.8ฐ
Duration too long
for an acute test,
no true
replication
Vogs et al. 2019
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
118 hours
PFOA
>96%
-
28
EC20
(deformities)
-
147.2ฐ
Duration too long
for an acute test,
no true
replication
Vogs et al. 2019
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
66 hours
PFOA
96%
-
28
NOEC
(survival and
development)
20.70-
>20.70
20.70ฐ
Duration too
short for an acute
test, only one
exposure
concentration
Dasgupta et al.
2020
Zebrafish
(embryo, 3 dpf),
Danio rerio
S,U
24 hours
PFOA
Unknown
-
28
LC50
-
96.06ฐ
Duration too
short for an acute
test, missing
exposure details
Gebreab et al.
2020
G-12
-------
Species (lil'cs(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Zebrafish
(embryo, 3 dpf),
Danio rerio
s,u
7 days
PFOA
Unknown
28
LC50
-
>41.41 to
<82.81ฐ
Duration too long
for an acute test
and too short for
a chronic test,
missing exposure
details
Gebreab et al.
2020
Zebrafish
(embryo, 4-64 stage),
Danio rerio
S,M
144 hours
PFOA
>98%
7.2-
7.6
26
NOEC
(survival and
development)
61 ->61
61
Duration too long
for an acute test
and too short for
a chronic test
Menger et al.
2020
Zebrafish
(embryo, 1 hpf),
Danio rerio
s,u
48 hours
PFOA
Unreported
-
28
LC50
-
300
Duration too
short for an acute
test
Pecquet et al.
2020
Zebrafish
(embryo, 1 hpf),
Danio rerio
S,M
24 hours
PFOA
Unreported
-
28
LOEC
(increase neutrophil
migration)
<0.685-
0.685
0.685
Duration too
short for an acute
test, atypical
endpoint
Pecquet et al.
2020
Zebrafish
(embryo, 5-6 hpf),
Danio rerio
R,U
90-91 hours
PFOA
>99%
-
28
LC50
-
57.6
Duration too
short for an acute
test
Wasel et al. 2020
Zebrafish
(embryo, 5-6 hpf),
Danio rerio
R,U
90-91 hours
PFOA
>99%
7
28
LC50
-
487.4
Duration too
short for an acute
test
Wasel et al. 2020
Zebrafish (embryo),
Danio rerio
R, M
5 days
APFO
>95%
-
25
NOEC
(mortality)
-
>1.582ฐ
Duration too long
for an acute test
and too short for
a chronic test
Hanetal. 2021
Zebrafish (embryo),
Danio rerio
R, M
5 days
PFOA
>95%
-
25
NOEC
(mortality)
-
>1.631ฐ
Duration too long
for an acute test
and too short for
a chronic test
Hanetal. 2021
Zebrafish
(embryo, 4 hpf),
Danio rerio
R, U
5 days
PFOA
96%
7.6-
7.8
25.3-
25.7
NOEC
(abnormal development,
growth and survival)
30->30
30
Duration too long
for an acute test
and too short for
a chronic test
Kim et al. 2021
Zebrafish
(embryo, 6-8 hpf),
Danio rerio
S,U
112-114 hours
PFOA
>98%
-
28
NOEC
(mortality)
-
0.2484ฐ
Duration too long
for an acute test
and too short for
a chronic test
Rericha et al.
2021
Zebrafish
(embryo, 6-8 hpf),
Danio rerio
S,U
112-114 hours
PFOA
>98%
-
28
MATC
(combined effects)
2.538-
6.451ฐ
4.047ฐ
Duration too long
for an acute test
and too short for
a chronic test
Rericha et al.
2021
G-13
-------
Species (lil'es(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITecl
(one.
(iiiii/l.)
Deficiencies
Reference
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
166 hours
PFOA
Unreported
28.5
LOEC
(behavior, gene
expression)
<0.00882-
0.00882
0.00882
Duration too long
for an acute test
and too short for
a chronic test,
non-apical
endpoints
Yuetal. 2021
Zebrafish,
Danio rerio
R,U
21 days
PFOA
Unreported
-
-
LOEC
(mRNA gene expression
in kidneys)
0.05-0.1
0.1
Non-apical
endpoint
Zhang et al. 2021
Zebrafish (adult, 4-5
months old),
Danio rerio
R, U
7-day exposure
+
1 day of
observation
PFOA
Unreported
7.2
27
NOEC
(immobility)
1.0->1.0
1.0
Duration too
short for a
chronic exposure,
atypical endpoint
Adedara et al.
2022
Zebrafish
(embryo, <4 hpf),
Danio rerio
R, U
5 days
PFOA
95%
-
27
NOEC
(mortality,
development)
-
>0.0007
Duration too long
for an acute test
and too short for
a chronic test
Haimbaugh et al.
2022
Zebrafish
(embryo, <4 hpf),
Danio rerio
R, U
5 days + 4-6
week
observation
PFOA
95%
-
27
NOEC
(reproduction, growth)
-
>0.0007
Duration too long
for an acute test
and too short for
a chronic test
Haimbaugh et al.
2022
Zebrafish
(embryo, 3 hpf),
Danio rerio
S,U
117 hours
PFOA
Unreported
7.2
28
LC50
-
>20.10ฐ
Duration too long
for an acute test
and too short for
a chronic test
Lindqvist and
Wincent 2022
Zebrafish
(embryo, 5 hpf),
Danio rerio
S,U
19 hr
PFOA
Unreported
7.5
28
LC50
-
33.89ฐ
Duration too
short
Satbhai et al. 2022
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOA
Unreported
-
28
NOEC
(mortality, development
and morphology)
-
>41.41ฐ
Duration too long
for an acute test
and too short for
a chronic test
Truong et al. 2022
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
APFOA
Unreported
-
28
NOEC
(mortality, development
and morphology)
-
>43.11ฐ
Duration too long
for an acute test
and too short for
a chronic test
Truong et al. 2022
Zebrafish
(embryo 2 hpf),
Danio rerio
S,M
70 hours
PFOA
Unreported
7.2-
7.4
-
NOEC
(growth-length,
mortality, and hatch)
0.1->0.1
0.1
Duration too
short for an acute
test
Yu et al. 2022
Zebrafish
(embryo, 6 hpf),
Danio rerio
R, U
90 hours
PFOA
Unreported
-
28
NOEC
(development and
growth)
-
>33.13ฐ
Duration too
short
Phelps et al. 2023
Zebrafish (larva, 72
hpf),
Danio rerio
R, U
72 hours
PFOA
>98%
-
28.5
LC50
-
132.9ฐ
Duration too
short
Sun et al. 2023a
G-14
-------
Species (lifcshiiic)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
( hrollic
l.illlil^
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
r.lTeel
(one.
(iiiii/l.)
Deficiencies
Reference
Zebrafish (larva, 72
hpf),
Danio rerio
R, U
72 hours
PFOA
>98%
28.5
LOEC
(growth - length)
-
41.41ฐ
Duration too
short
Sun et al. 2023a
Zebrafish
(embryo, <4 hpf),
Danio rerio
R, M
116 hours
PFOA
Unreported
-
-
LOEC
(immunological
responses)
0.0001
0.0001
Duration too long
for an acute test
and too short for
a chronic test,
atypical endpoint
Tang et al. 2023
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
94 hours
PFOA
>96%
-
27
NOEC
(mortality and
development)
-
0.4915
Duration too
short
Wang et al. 2023
Zebrafish
(embryo, 1 hpf),
Danio rerio
S,U
119 hours
PFOA
95%
7.2
28
LC50
-
561.01
Duration too long
for an acute test
and too short for
a chronic test
Wasel et al. 2022
Zebrafish
(embryo, 5 hpf),
Danio rerio
R, U
122 hours
PFOA
Unreported
6.8-
7.0
28
LC50
-
>41.41ฐ
Duration too long
for an acute test
and too short for
a chronic test
Hawkey et al.
2023
Zebrafish
(embryo, 5 hpf),
Danio rerio
R, M
91 hours
PFOA
96%
7.0-
8.0
28.5
LC50
-
~2
Duration too
short for an acute
test
Liu et al. 2023b
Zebrafish (2 months),
Danio rerio
R, U
14 days
APFOA
Unreported
7
28
LOEC
(biochemical and
enzymatic effects)
-
0.03
Non-apical
endpoint,
duration too long
for an acute test
and too short for
a chronic test
Liu et al. 2023a
Rare minnow
(male, 9 months old,
1.4 g, 47.7 cm),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
25
MATC
(increase relative
mRNA expression of
AhR in gills)
3.0-10
5.477
Non-apical
endpoint
Liu et al. 2008b
Rare minnow
(female, 9 months old,
1.4 g, 47.7 cm),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
25
MATC
(decrease relative
mRNA expression of
CYPla and increase
relative mRNA
expression of PXR in
gills)
10.0-30
17.32
Non-apical
endpoint
Liu et al. 2008b
Rare minnow
(adult, 9 months old),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
25
LOEC
(polymerase chain
reaction (PCR)
alterations of genes in
liver)
<3-3
3
Non-apical
endpoint
Wei et al. 2008b
G-15
-------
Chronic
Limits
Reported
(NOI'.C-
HITecl
( hem ic;il /
Temp.
i.or.C)
(one.
Species (lil'es(;iปc)
Method"
lost Dui'iiiion
Pn ril
pll
(ฐC)
r.nvci
(111 Si/I.)
(iiiii/l.)
Deficiencies
Reference
Rare minnow
(adult, 9 months old),
F, U
28 days
PFOA
98%
_
MATC
(change in m-RNA M-
3.0-10
5.477
Non-apical
endpoint
Wei et al. 2008a
Gobiocypris rarus
H-FABP)
Rare minnow
(adult, 9 months old),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
-
LOEC
(protein spots identified
by MALD1-TOF/TOF)
<3-3
3
Non-apical
endpoint
Wei et al. 2008a
Rare minnow
(9 months old female,
1.4 g, 47.7 mm),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
25
MATC
(increase relative
mRNA expression of
PPARy in gills)
3.0-10
5.477
Non-apical
endpoint
Liu et al. 2009
Rare minnow
(9 months old male, 1.4
g, 47.7 mm),
Gobiocypris rarus
F, U
28 days
PFOA
98%
-
25
MATC
(increase relative
mRNA expression of
PPARy and PPARa in
gills and CYP4T11 in
liver)
10.0-30
17.32
Non-apical
endpoint
Liu et al. 2009
Rare minnow (9 months
old male, 1.3 g),
Gobiocypris rarus
F, U
14 days
PFOA
98%
-
25
LOEC
(apolipoprotein gene
expression)
-
3
Duration too
short for a
chronic test, non-
apical endpoint
Fang et al. 2010
Test substance
was not
characterized but
considered a
mixture of APFO
and other PFAS
Fathead minnow
(juvenile),
Pimephales promelas
s,u
96 hours
APFO
Unknown
7.2-
7.9
21.8-
22.5
LC50
2,470
impurities; test
substance purity
unknown and not
used
quantitatively
because other
tests from 3M
Company
(2000a) indicated
test purity as low
as 78%.
3M Company
2000a
Fathead minnow,
Pimephales promelas
s,u
96 hours
PFOA
96.5-100%
-
-
LC50
-
440
Lack of exposure
details, possible
mixture effects of
other perfluoro
homologue
compounds; only
one replicate;
3M Company
2000a
G-16
-------
Species (lil'cs(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Fathead minnow,
Pimephales promelas
s,u
96 hours
PFOA
96.5-100%
-
LC50
-
140
Possible mixture
effects of other
perfluoro
homologue
compounds and
the amount of
isopropanol, low
initial pH (3.0-
4.3) in highest
test concentration
3M Company
2000a
Fathead minnow,
Pimephales promelas
s,u
96 hours
APFO
96.5-100%
-
-
LC50
-
70
Highest test
concentration
produced 20%
effect, LC50
based on
extrapolation
3M Company
2000a
Fathead minnow,
Pimephales promelas
s,u
96 hours
APFO
96.5-100%
7.9-
8.0
19
LC50
-
776
Lack of
replication
3M Company
2000a/Elnabarawy
1980
Fathead minnow,
Pimephales promelas
s,u
96 hours
APFO
96.5-100%
7.9-
8.0
19
LC50
-
754
Lack of
replication
3M Company
2000a/Elnabarawy
1980
Fathead minnow,
Pimephales promelas
s,u
96 hours
APFO
78-93%
7.7-
8.0
20
LC50
-
301
Possible mixture
effects of other
perfluoro
analogue
compounds; low
test substance
purity
3M Company
2000a
Fathead minnow
(juvenile),
Pimephales promelas
s,u
96 hours
APFO
Unreported
7.4-
8.4
21-22
LC50
-
>1,000
Highest treatment
did not produce
>50% effect; test
substance purity
unknown and not
used
quantitatively
because other
tests from 3M
Company
(2000a) indicated
test purity as low
as 78%.
3M Company
2000a
Fathead minnow
(embryo, 48 hpf),
Pimephales promelas
F, U
30 days post
hatch
APFO
96.5-100%
7.0-
7.3
25
NOEC
(hatch, survival and
growth)
100->100
100
Lack of
replication for
early life stage
test
3M Company
2000a
G-17
-------
Species (lil'es(;iปc)
Method'
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITccl
(one.
(iiiii/l.)
Deficiencies
Reference
Fathead minnow
(64 days old),
Pimephales promelas
S,M
13 days
APFO
96.5-100%
-
BCF
-
1.8
(L/Kg)
Steady state not
documented,
static uptake
study; exposure
duration is too
short for a
chronic test
3M Company
2000a
Fathead minnow
(adult, 6-8 months),
Pimephales promelas
F, M
96 hours
PFOA
>95%
8.2
25
LC50
-
>16.62
LC50 not
reported,
inferred, other
definitive data
available for the
species
Villeneuve et al.
2023
Fathead minnow
(adults),
Pimephales promelas
S,M
39 days
PFOA
19.4%f
8.5
10.6-
26.6
LOEC
(mean total egg
production)
74.1-
>74.1
>74.1
Atypical
exposure, started
with adults, not a
true ELS test
Oakes et al. 2004
Topmouth gudgeon
(4.0 g, 4.0 cm),
Pseudorasbora parva
S,M
96 hours
PFOA
99%
7
22
LC50
-
365.02
Atypical source
of test organisms
Yang et al. 2014
Topmouth gudgeon
(4.0 g, 4.0 cm),
Pseudorasbora parva
R, M
30 days
PFOA
99%
7
22
EC10
(survival)
-
11.78
Not a true ELS
test (started with
older life stage),
atypical source of
organisms
Yang et al. 2014
Nile tilapia,
Oreochromis niloticus
Dietary
6 days
PFOA
Unreported
-
22
MATC
(survival and growth)
1.0-5.0
(mg/g)
2.236
(mg/g)
Duration too
short for a
chronic test,
dietary exposure,
water
concentration not
reported
Hanetal. 2011
Bluegill,
Lepomis macrochirus
S,U
96 hours
APFO
96.5-100%
7.8-
8.0
18-19
LC50
-
569
Only one
replicate per
treatment
3M Co. 2000a
Mosquitofish,
Gambusia affinis
S,U
96 hours
PFOA
>98%
7.2
25
LOEC
(genetics: AKT
serine/threonine kinase
3 mRNA)
-
0.50
Non-apical
endpoint
Liu et al. 2022
G-18
-------
Species (lil'es(;iปc)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
HITecl
(one.
(iiiii/l.)
Deficiencies
Reference
Murray River
rainbowfish
(male, adult, 1 year
old),
Melanotaenia fluviatilis
R, M
14 days
PFOA
>96%
7.1-
7.4
23
NOEC
(growth and mortality)
9.0->9.0
9.0
Duration too long
for acute test and
too short for a
chronic test, not
NA species
Miranda et al.
2020
Medaka (<6 hpf),
Oryzias latipes
R, U
Embryo
development to
48 hours post-
hatch
PFOA
Unreported
-
25
LOEC
(swim bladder
development)
-
4.7
Only one
exposure
concentration
Godfrey 2017
Medaka (adult, male),
Oryzias latipes
R, U
14 days
PFOA
Unreported
-
25
NOEC
(adult survival, GSI%,
HSI%, K%)
10->10
10
Duration too long
for acute test and
too short for a
chronic test
Ji et al. 2008
Medaka (adult, female),
Oryzias latipes
R, U
14 days
PFOA
Unreported
-
25
NOEC
(adult survival, GSI%,
HSI%, K%)
10->10
10
Duration too long
for acute test and
too short for a
chronic test
Ji et al. 2008
Medaka (F1 generation,
<12 hours old,
embryo),
Oryzias latipes
R, U
7-14 days
(assumed)
PFOA
Unreported
-
25
NOEC
(% hatchability)
10->10
10
Duration too long
for acute test and
too short for a
chronic test
Ji et al. 2008
Medaka (F1 generation,
<12 hours old,
embryo),
Oryzias latipes
R, U
7-14 days
(assumed)
PFOA
Unreported
-
25
MATC
(time to hatch)
1.0-10
3.162
Duration too long
for acute test and
too short for a
chronic test
Ji et al. 2008
Medaka (F1 generation,
<12 hours old,
embryo),
Oryzias latipes
R, U
28 days post-
hatch
(assumed)
PFOA
Unreported
-
25
NOEC
(swim up success)
10->10
10
Pseudoreplication
that occurred at
hatching stage
Ji et al. 2008
Medaka (F1 generation,
<12 hours old,
embryo),
Oryzias latipes
R, U
100 days post-
hatch
PFOA
Unreported
-
25
NOEC
(growth - length and
weight)
10->10
10
Pseudoreplication
that occurred at
hatching stage
Ji et al. 2008
Medaka (F1 generation,
<12 hours old,
embryo),
Oryzias latipes
R, U
100 days post-
hatch
PFOA
Unreported
-
25
NOEC
(male/female GSI% and
HSI%)
0.1->0.1
0.1
Pseudoreplication
that occurred at
hatching stage
Ji et al. 2008
G-19
-------
Species (lilVsliiuo)
Method"
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
l.imils
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
r.lTeel
(one.
(iiiii/l.)
Deficiencies
Reference
Medaka (F1 generation,
<12 hours old,
embryo),
Oryzias latipes
R,U
28 days post-
hatch
PFOA
Unreported
25
LOEC
(larval survival)
<0.1-0.1
0.1
Pseudoreplication
that occurred at
hatching stage
Ji et al. 2008
Medaka,
Oryzias latipes
s,u
96 hours
PFOA
Unreported
7.0-
7.5
25
NOEC
(mortality)
-
473
Missing exposure
details, only one
exposure
concentration
Godfrey et al.
2019
Medaka
(embryo, <6 hpf),
Oryzias latipes
R,U
10 days
PFOA
Unreported
7.0-
7.5
25
NOEC
(growth)
-
4.7
Duration too
short for a
chronic test, only
one exposure
concentration
Godfrey et al.
2019
Medaka
(adult, 16 weeks old,
0.38 g),
Oryzias latipes
R, U
21 days
PFOA
96%
-
25
LOEC
(fecundity)
<10-10
10
Only one
exposure
concentration,
tolerant LOEC
value
Kang et al. 2019
Medaka (embryo),
Oryzias latipes
S,U
3 dph
PFOA
>98.0%
7.3
27
LOEC
(genetic and hormonal
changes)
-
25
Non-apical
endpoint
Myosho et al.
2022
Medaka (fry),
Oryzias latipes
S,U
3 dph
PFOA
>98.0%
7.3
27
NOEC
(genetic changes)
-
2.5
Non-apical
endpoint
Myosho et al.
2022
Northern leopard frog
(larva, Gosner 26),
Lithobates pipiens
(formerly, Rana
pipiens)
R, M
40 days
PFOA
96%
-
20
NOEC
(snout-vent length
and Gosner stage at 40
d)
1->1
1
No effects at the
highest
concentration
tested resulting in
a relatively low
greater than value
that does not
inform species
sensitivity
Hoover etal. 2017
Northern leopard frog
(larva, Gosner 25),
Lithobates pipiens
S,M
30 days
PFOA
>96%
7.8
26.2
NOEC
(survival and growth)
0.066-
>0.066
0.066
Mesocosm
exposure
Flynnetal. 2021
Northern leopard frog
(larva, Gosner stage
25),
Lithobates pipiens
R, M
30 days
PFOA
>96%
-
20
MATC
(growth - weight)
0.1251-
1.376
0.4149
Highly variable
dose-response
Flynn et al. 2022
G-20
-------
Species (lil'es(;iปc)
Method'
lost Dui'iiiion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
r.nvci
Chronic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
I.ITcct
(one.
(inti/l.)
Deficiencies
Reference
Black-spotted pond
frog (adult),
Pelophylax
nigromaculatus
(formerly, Rana
nigromaculatus)
S,U
21 days
PFOA
>98%
-
LOEC
(gene expression)
-
0.01
Non-apical
endpoint
Lin et al. 2022a
Black-spotted pond
frog,
Pelophylax
nigromaculatus
R,U
21 days
PFOA
>98%
6.5
20
LOEC
(cholesterol and
triglyceride levels)
-
0.0025
Non-apical
endpoint
Lin et al. 2022b
Black-spotted pond
frog (adult),
Pelophylax
nigromaculatus
S,U
21 days
PFOA
>98%
6.5
20
LOEC
(reduced glutathione)
-
0.91
Non-apical
endpoint
Lin et al. 2022c
Black-spotted pond
frog,
Pelophylax
nigromaculatus
S,U
21 days
PFOA
>98%
6.5
20
MATC
(biochemical, genetic
and immunological
effects)
0.00111-
0.01094
0.003485
Non-apical
endpoint
Liu et al. 2023c
Gray treefrog (larva,
Gosner 40),
Hyla versicolor
S,U
96 hours
PFOA
Unreported
-
21
LC50
-
191
Poor control
survival, low
number of
individuals per
treatment.
Tornabene et al.
2021
Tiger salamander
(larva, Harrison stage
46),
Ambystoma tigrinum
R, M
30 days
PFOA
>96%
-
20
NOEC
(snout-vent length,
weight, and mortality)
0.8677-
>0.8667
0.8677
Highly variable
dose-response
Flynn et al. 2022
American toad (larva,
Gosner stage 25),
Anaxryrus americanus
R, M
26-45 days
PFOA
>96%
-
20
MATC
(days to metamorphosis)
0.09403-
0.8728
0.2865
Highly variable
dose-response
Flynn et al. 2022
American toad (larva,
Gosner stage 25),
Anaxryrus americanus
R, M
26-45 days
PFOA
>96%
-
20
NOEC
(growth, mortality)
0.8728-
>0.8728
0.8728
Highly variable
dose-response
Flynn et al. 2022
Asiatic toad (tadpole,
1.8 cm, 0.048 g),
Bufo gargarizans
R, M
96 hours
PFOA
99%
7
22
LC50
-
114.74
Atypical source
of organisms
Yang et al. 2014
G-21
-------
Species (lil'cs(;iปc)
Method'
lost Diimlion
( hem ic;il /
Pn ril
pll
Temp.
(ฐC)
I'.ITecl
( limnic
Limits
(NOI'.C-
i.or.C)
(111 Si/I.)
Reported
r.lTecl
(one.
(inti/l.)
Deficiencies
Reference
Asiatic toad (tadpole,
1.8 cm, 0.048 g),
Bufo gargarizans
R, M
30 days
PFOA
99%
7
22
ECio
(survival)
-
5.89
Not a true ELS
test, atypical
source of
organisms
Yang et al. 2014
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b CI year corresponds to the total duration of the 10-generations study. Most generations did not show statistically significant effects.
0 Reported in moles converted to milligram based on a molecular weight of 414.07 g/mol PFOA or 431.1 g/mol APFO.
d Chemical concentrations made in a side-test representative of exposure and verified stability of concentrations of PFOA in the range of concentrations tested under similar
conditions. Daily renewal of test solutions.
6 Concentration of APFO determined as the anion (PFO-).
G-22
-------
Appendix H Other Estuarine/Marine PFOA Toxicity Studies
H.l Summary Table of Qualitative Estuarine/Marine PFOA Toxicity Studies
Species (lil'cs(;iปc)
Method'1
Tcsl
Diinilion
( lii-m ic;il /
Piiriu
pll
1 om p.
(ฐC)
Siilinilt
(Dl)ll
r.fiwi
Chronic
Limits
(NOI'.C-
i.or.ci
(lllป/l.)
Reported
r.lTecl
Cone.
(lliu/l - >
Deficiencies
Reference
Bacterium,
Vibrio fischeri
S,M
15 minutes
PFOA
96%
18
ECso
(bioluminescence
inhibition)
-
524
Duration too short for
a plant test, missing
some exposure
details, non-apical
endpoint
Rosal et al.
2010
Cyanobacterium,
Anabaena sp.
S,M
24 hours
PFOA
96%
-
28
-
ECso
(bioluminescence
inhibition)
-
72.3
Duration too short for
a plant test, missing
some exposure
details, non-apical
endpoint
Rosal et al.
2010
Cyanobacterium,
Geitlerinema
amphibium
S,U
72 hours
PFOA
Unreported
7.6-
7.8
20
8
ECso
(growth)
-
248.4b
Duration too short for
a plant test, missing
some exposure details
Latala et al.
2009
Dinoflagellate,
Pyrocystis lunula
S,M
24 hours
PFOA
95%
-
19
-
EC50
(bioluminescence
inhibition)
-
18
Duration too short for
a plant test, atypical
endpoint
Hayman et al.
2021
Golden brown alga,
Isochrysis galbana
S,U
72 hours
PFOA
96%
-
20
-
ECso
(growth inhibition)
-
163.6
Duration too short for
a plant test, missing
some exposure details
Mhadhbi et al.
2012
Green alga,
Chlorella vulgaris
S,U
72 hours
PFOA
Unreported
7.6-
7.8
20
8
ECso
(growth)
-
977.2b
Duration too short for
a plant test, missing
some exposure details
Latala et al.
2009
Diatom,
Skeletonema marinoi
S,U
72 hours
PFOA
Unreported
7.6-
7.8
20
8
ECso
(growth)
-
368.5b
Duration too short for
a plant test, missing
some exposure details
Latala et al.
2009
Purple sea urchin
(fertilized eggs),
Paracentrotus lividus
S,U
48 hours
PFOA
96%
-
20
-
ECso
(growth inhibition)
-
110.0
Duration too short for
an acute test
Mhadhbi et al.
2012
Blue mussel,
Mytilus edulis
R, U
21 days
PFOA
Unreported
-
16-19
-
LOEC
(catalase activity)
<0.2-0.2
0.2
Atypical endpoint
Li et al. 2021a
H-l
-------
Species (lil'cs(;iปc)
Method'1
Tcsl
Diinilion
( lii-m ic;il /
Piiriu
nil
Temp.
(ฐC)
Siilinilt
(Dl)ll
r.riwi
Chronic
Limits
(NOI'.C-
i.or.ci
(lliu/l - >
Reported
F.ITecl
Cone.
(mป/l.)
Deficiencies
Reference
Mediterranean
mussel (adult),
Mytilus
galloprovincialis
R, U
16 days
PFOA
>98%
7.98
15.2
37.9
NOEC
(biochemistry,
enzymatic changes)
0.100-
>0.100
0.100
Duration too short for
a chronic exposure,
non-apical endpoints
Geng et al.
2021
Green mussel
(60-65 mm),
Perna viridis
R,M
7 days
PFOA
96%
-
25
25
MATC
(relative condition
factor)
0.0114-
0.099
0.03359
Exposure duration too
short for chronic test
and too long for acute
test, non-apical
endpoint
Liu et al.
2013;2014c
Green mussel (adult),
Perna viridis
R, M
7 days + 7
days
observation
PFOA
96%
-
25
30
ECso
(integrative
genotoxicity)
0.093-
0.950
0.5940
Exposure duration too
short for chronic test
and too long for acute
test, non-apical
endpoint
Liu et al.
2014a
Green mussel (adult),
Perna viridis
R, M
7 days
PFOA
96%
-
25
25
MATC
(CAT and SOD
activity)
0.099-
1.12
0.3330
Exposure duration too
short for chronic test
and too long for acute
test, non-apical
endpoint
Liu et al.
2014b
Green mussel,
Perna viridis
R, M
7 days + 7
days
observation
PFOA
96%
8
25
30
MATC
(hemocyte cell
viability)
0.0114-
0.099
0.03359
Exposure duration too
short for chronic test
and too long for acute
test, non-apical
endpoint
Liu and Gin
2018
Manila clam
(3.64 cm),
Ruditapes
philippinarum
R, M
21 days
PFOA
Unreported
-
12
35
NOEC
(mortality)
0.00093-
>0.00093
0.00093
Only one exposure
concentration, apical
endpoints are not the
focus of study
Bernardini et
al. 2021
Manila clam,
Ruditapes
philippinarum
R, M
21 days
APFO
>98%
-
12
35
NOEC
(biochemistry, cell
diameter, cell
volume,
hematocyte count)
0.00093-
>0.00093
0.00093
Only one exposure
concentration, non-
apical endpoints
Fabrello et al.
2021
Dolphinfish (embryo,
<8 hr post spawn),
Coryphaena hippurus
S,M
48 hours
PFOA
96%
-
27
-
LC50
-
4
Duration too short for
an acute test, high
control mortality
Gebreab et al.
2022
Japanese medaka
(adult),
Oryzias latipes
R, U
7 days
PFOA
ammonium salt
98%
-
25
-
NOEC
(survival, condition
factor)
100->100
100
Exposure duration too
short for chronic test
and too long for acute
test
Yang 2010
H-2
-------
Species (lil'es(;iปc)
Method'1
Tesl
Diinilion
( lii-m ic;il /
Piiriu
nil
Temp.
(ฐC)
Siilinilt
(Dl)ll
r.fiwi
Chronic
limits
(NOI'.C-
i.or.ci
(lliu/l - >
Reported
l-'.ITed
Cone.
(lliu/l - >
Deficiencies
Reference
Dusky rockcod
(adult),
Trematomus newnesi
S,U
10 days
PFOA
Unreported
-
-1
-
LOEC
(gene expression)
-
1.5
Duration too long for
an acute test and too
short for a chronic
test, atypical endpoint
Pacchini et al.
2023
Marbled flounder
(adult),
Pseudopleuronectes
yokohamae
R,M
14 days
PFOA
Unreported
-
15
37
LOEC
(vitellogenin levels)
<0.197-
0.197
0.197
Atypical duration,
non-apical endpoint
Li et al. 2018b
Turbot (embryo),
Scophthalmus
maximus
(formerly, Psetta
maxima)
R, U
6 days
PFOA
96%
-
18
-
LC50
-
11.9
Exposure duration too
short for chronic test
and too long for acute
test
Mhadhbi et al.
2012
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b Reported in moles converted to milligram based on a molecular weight of 414.07 mg/mmol.
H-3
-------
Appendix I Acute-to-Chronic Ratios
1.1 Acute-to-Chronic Ratios from Quantitatively Acceptable Tests.
Species
C'heiniciil /
Purity
Acute
lest
l)ii riit ion
Chronic
lest
Duriition
Acute
I! fleet
Chronic K I'fect
Acute
I! fleet
(011c.
(niป/l.)
Chronic
I! fled
Cone.
(niป/l.)
AC K
SMACK
Reference
Rotifer,
Brachionus
calyciflorus
PFOA
96%
24 hours
Up to 200
hours
LC50
EC10
(intrinsic rate of
natural increase)
150
0.5015
299.1
299.1
Zhang et al.
2013a
Water flea,
Daphnia carinata
PFOA
95%
48 hours
21 days
EC50
MATC
(# of average
number of
offspring per brood
and total # of living
offspring)
66.8
0.03162
2,113
2,113
Logeshwaran
et al. 2021
Cladoceran,
Daphnia magna
APFO
99.7%
48 hours
21 days
EC50
EC10
(average # of live
young)
480
20.26
23.69
-
Colombo et
al. 2008
Cladoceran,
Daphnia magna
PFOA
Unreported
48 hours
21 days
EC50
(immobility)
EC10
(# young/starting
female)
542.5
7.853
69.08
-
Ji et al. 2008
Cladoceran,
Daphnia magna
PFOA
>98%
48 hours
21 days
LC50
EC10
(# young/starting
female)
193.3a
12.89
15.00
-
Li 2009,
2010
Cladoceran,
Daphnia magna
PFOA
99%
48 hours
21 days
LC50
EC10
(survival)
222.0
5.458
40.67
-
Yang et al.
2014
Cladoceran,
Daphnia magna
PFOA
98%
48 hours
21 days
EC50
MATC
(growth and
reproduction)
114.6
0.07155
l,602b
-
Lu et al. 2016
Cladoceran,
Daphnia magna
PFOA
Unreported
48 hours
21 days
LC50
EC10
(# of offspring)
117.2
8.084
14.50
27.05
Yang et al.
2019
i-l
-------
Species
Chemical /
I'liritv
Acute
Test
Dm nit ion
Chronic
Test
Dui'iition
Acule
Kffccl
Chronic Kfleet
Acute
I! fleet
Cone.
Chronic
Klfect
Cone.
(111 ซ/l,)
ACR
SMACK
Reference
( kidoccmn.
Moina macrocopa
PI O A
I nivpoi k'il
4S hours
7 da\ s
IX'SO
(immobilily)
i:c
Miiean \n1111g aduli;
1 hh 3
:
75 So
75 So
Ji cl al :00s
Mayfly,
Neocloeon
triangulifer
PFOA
95%
96 hours
23 days
LC50
NOEC
(survival, weight,
emergence)
13.05
>3.085
<4.229
<4.229
Soucek et al.
2023
American bullfrog,
Lithobates
catesbeiana
PFOA
Unreported
96 hours
72 days
LC50
LOEC
(snout vent length)
1,006
0.288
3,493
3,493
Flynn et al.
2019
a Geometric mean of three LC50 values.
b Value not used in the SMACR calculation, because the value is an order of magnitude greater than other ACRs for the species.
1-2
-------
Appendix J Unused PFOA Toxicity Studies
J,1 Summary of Unused PFOA Toxicity Studies
Author
( iliilion
Koiison I misod
Arukwe, A. and A.S. Mortensen
2t>I I Lipid pei'iiNidalKin and o\idali\ c slicss responses of salmon led a diel
containing perfluorooctane sulfonic- or perfluorooctane carboxylic acids. Comp.
Biochem. Physiol. Part C 154: 288-295.
Fui'uc-lcd (oral gavagej, only one expusiut
concentration
Consoer, D.M.
2017. A mechanistic investigation of perfluoroalkyl acid kinetics in rainbow trout
{Oncorhynchus mykiss). A dissertation submitted to the faculty of the University
of Minnesota.
Injected toxicant; only one exposure
concentration
Consoer, D.M., A.D. Hoffman, P.N.
Fitzsimmons, P. A. Kosian and J.W. Nichols
2014. Toxicokinetics of perfluorooctanoate (PFOA) in rainbow trout
{Oncorhynchusmykiss). Aquat. Toxicol. 156: 65-73.
All fish were surgically altered (dorsal
aortic cannula, plus a urinary catheter); no
controls; non-apical endpoints only
Cui, Y., W. Liu, W. Xie, W. Yu, C. Wang
and H. Chen
2015. Investigation of the effects of perfluorooctanoic acid (PFOA) and
perfluorooctane sulfonate (PFOS) on apoptosis and cell cycle in a zebrafish
(Danio rerio) liver cell line. Int. J. Environ. Res. Public Health 12(12): 15673-
15682.
Excised cells (liver cell line)
De Silva, A.O., P.J. Tseng and S.A. Mabury
2009. Toxicokinetics of perfluorocarboxylate isomers in rainbow trout. Environ.
Toxicol. Chem. 28(2): 330-337.
Study involved a mixture of ECF PFOA,
linear PFNA, and isopropyl PFNA added to
diet
Dragojevic, J., P. Marie, J. Loncar, M.
Popovic, I. Mihaljevic, and T. Smital
2020. Environmental Contaminants Modulate Transport Activity of Zebrafish
Organic Anion Transporters Oatl and Oat3. Comp. Biochem. Physiol. C Toxicol.
Pharmacol.231:8 p.
The test duration was extremely brief (~10
mins), the exposure was in vitro, and no
apical effects were evaluated.
Elnabarawy, M.T.
1980. Aquatic Toxicity Testing: FC-143 (Lot-37) L.R. 5626 S. Report No. 037;
OPPT Administrative Record AR226-0504:3780-3786.
Results reported in another publication
Fernandez-Sanjuan, M., M. Faria, S.
Lacorte and C. Barata
2013. Bioaccumulation and effects of perfluorinated compounds (PFCs) in zebra
mussels (Dreissena polymorpha). Environ. Sci. Pollut. Res. 20:2661-2669.
Mixture
Garoche, C., A. Boulahtouf, M. Grimaldi,
B. Chiavarina, L. Toporova, M.J. Den
Broeder, J. Legler, W. Bourguet and P. Ba.
2021. Interspecies differences in activation of peroxisome proliferator-activated
receptor gamma by pharmaceutical and environmental chemicals. Environ. Sci.
Technol. 55(24): 16489-16501.
In vitro
Gonzalez-Naranjo, V. and K. Boltes
2014. Toxicity of ibuprofen and perfluorooctanoic acid for risk assessment of
mixtures in aquatic and terrestrial environments. Int. J. Environ. Sci. Technol. 11:
1743-1750.
Severe lack of exposure details (cannot
judge against data quality objectives)
Gorrochategui, E., S. Lacorte, R. Tucker
and F.L. Martin
2016. Perfluoroalkylated substance effects inXenopus laevis A6 kidney epithelial
cells determined by ATR-FTIR spectroscopy and chemometric analysis. Chem.
Res. Toxicol. 29: 924-932.
The tests were performed on cell cultures
obtained from an outside source; whole
organisms were not investigated
Holth, T.F., M. Yazdani, A. Lenderink and
K. Hyllan
2012. Effects of fluoranthene and perfluorooctanoic acid (PFOA) on immune
functions in Atlantic cod (Gadus morhua). Abstracts Comp. Biochem. Physiol.
Part A. 163: S39-S42.
Abstract only; cannot judge against data
quality objectives
J-l
-------
Author
( iliilion
Ko;isoii I iiusod
Hoover, G.M.
2018. Effects of Per/Polyfluoroalkyl Substance Exposure on Larval Amphibians.
Ph.D.Thesis, Purdue University, West Lafayette, IN
Test is unused because the source material
is a PhD Thesis and has not undergone peer
review. Only peer-reviewed data are
considered in deriving aquatic life criteria
Hoover, G., S. Kar, S. Guffey, J.
Leszczynski, and M.S. Sepulveda
2019. In Vitro and In Silico Modeling of Perfluoroalkyl Substances Mixture
Toxicity in an Amphibian Fibroblast Cell Line. Chemosphere233:25-33.
In vitro exposure
Jantzen, C.E., K.M. Annunziato and K.R.
Cooper
2016. Behavioral, morphometric, and gene expression effects in adult zebrafish
(Danio rerio) embryonically exposed to PFOA, PFOS, and PFNA. Aquatic
Toxicology. 180:123-130.
Single concentration test where exposure to
PFOA was of an acute (117-hours) duration
but endpoints were measured at 6 months of
age
Jantzen, C.E., F. Toor, K.M. Annunziato
and K.R. Cooper
2017b. Effects of chronic perfluorooctanoic acid (PFOA) at low concentration on
morphometries, gene expression, and fecundity in zebrafish (Danio rerio).
Reproduct. Toxicol. 69: 34-42.
Unable to determine dietary exposure
concentration
Khan, E.A., X. Zhang, E.M. Hanna, F.
Yadetie, I. Jonassen, A. Goksoyr and A.
Arukwe
2021. Application of Quantitative Transcriptomics in Evaluating the Ex Vivo
Effects of Per- and Polyfluoroalkyl Substances on Atlantic Cod (Gadus morhua)
Ovarian Physiology. Sci. Total Environ.755(l): 11 pp.
In vitro exposure
Lee, W. and Y. Kagami
2010. Effects of perfluorooctanoic acid and perfluorooctane sulfonate on gene
expression profiles in medaka (Oryzias latipes). Abstracts. Toxicol. Letters 196S:
S37-S351.
Abstract only, cannot judge against data
quality objectives
Li, M.H.
2011. Changes of cholinesterase and carboxylesterase activities in male guppies,
Poecilia reticulata, after exposure to ammonium perfluorooctanoate, but not to
perfluorooctane sulfonate. Fresenius Environ. Bull. 20(8a): 2065-2070.
Each treatment group for PFOA was run
two times at separate times (not
simultaneously) and the sample size for
each treatment group was unclear.
Liang, X. and J. Zha
2016. Toxicogenomic applications of Chinese rare minnow (Gobiocypris rarus) in
aquatic toxicology. Comp. Biochem. Physiol. PartD 19: 174-180.
Review paper
Liu, C., Y. Du and B. Zhou
2007a. Evaluation of estrogenic activities and mechanism of action of
perfluorinated chemicals determined by vitellogenin induction in primary cultured
tilapia hepatocytes. Aquat. Toxicol. 85: 267-277.
In vitro, cultured hepatocytes
Liu, C., K. Yu, X. Shi, J. Wang, P.K.S.
Lam, R.S.S. Wu and B. Zhou
2007b. Induction of oxidative stress and apoptosis by PFOS and PFOA in primary
cultured hepatocytes of freshwater tilapia (Oreochromis niloticus). Aquat.
Toxicol. 82: 135-143.
In vitro, cultured hepatocytes
Mahapatra, C.T., N.P. Damayanti, S.C.
Guffey, J.S. Serafin, J. Irudayaraj and M.S.
Sepulveda
2017. Comparative in vitro toxicity assessment of perfluorinated carboxylic acids.
J. Appl. Toxicol. 37: 699-708.
In vitro exposure, zebrafish liver cell
cultures
Martin, J.W., S.A. Mabury, K.R. Solomon
andD.C.G. Muir
2003a. Bioconcentration and tissue distribution of perfluorinated acids in rainbow
trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 22: 196-204.
Bioaccumulation (steady state not
documented); only 12 days
Martin, J.W., S.A. Mabury, K.R. Solomon
andD.C.G. Muir
2013. Progress toward understanding the bioaccumulation of perfluorinated alkyl
acids. Environ. Toxicol. Chem. 32(11): 2421-2423
Review paper
Mortensen, A.S., R.J. Letcher, M.V.
Cangialosi, S. Chu and A. Arukwe
2011. Tissue bioaccumulation patterns, xenobiotic biotransformation and steroid
hormone levels in Atlantic salmon (Salmo salar) fed a diet containing
perfluoroactane sulfonic or perfluorooctane carboxylic acids. Chemosphere. 83:
1035-1044.
One dietary dosage level provided over a 6-
day period; not intended as a toxicity test
J-2
-------
Author
( iliilion
Ko;isoii I iiusod
Otero-Sabio, C., M. Giacomello, C.
Centelleghe, F. Caicci, M. Bonato, A.
Venerando, J.M. Graic, S. Mazzariol, L.
Finos
2022. Cell Cycle Alterations Due to Perfluoroalkyl Substances PFOS, PFOA,
PFBS, PFBA and the New PFAS C604 on Bottlenose Dolphin (Tursiops
truncatus) Skin Cell. Ecotoxicol. Environ. Saf.244:10 p.
In vitro
Padilla, S., D. Coram, B. Padros, D.L.
Hunter, A. Beam, K.A. Houck, N. Sipes, N.
Kleinstreuer, T. Knudsen, D.J. Nix and
D.M. Reif
2012. Zebrafish developmental screening of the ToxCastTM Phase I chemical
library. Reprod. Toxicol. 33: 174-187.
Severe lack of exposure details, only one
exposure concentration
Popovic, M, R. Zaja, K. Fent and T. Smital
2014. Interaction of environmental contaminants with zebrafish organic anion
transporting polypeptide, Oatpldl (Slcoldl). Toxicol. Appl. Pharmacol. 280(1):
149-158.
Excised cells
Prosser, R.S., K. Mahon, P.K. Sibley, D.
Poirier and T. Watson-Leung
2016. Bioaccumulation of perfluorinated carboxylates and sulfonates and
polychlorinated biphenyls in laboratory-cultured Hexagenia spp., Lumbriculus
variegatus and Pimephales promelas from field-collected sediments. Sci. Total
Environ. 543: 715-726.
Mixture (filed collected sediment, contained
PFAS mixtures and PCBs)
Rotondo, J.C, L. Giari, C. Guerranti, M.
Tognon, G. Castaldelli, E.A. Fano and F.
Martini
2018. Environmental doses of perfluorooctanoic acid change the expression of
genes in target tissues of common carp. Environ. Toxicol. Chem. 37(3): 942-948.
Two exposure concentrations 10,000-fold
apart; atypical endpoint
Sanderson, H., T.M. Boudreau, S.A.
Mabury and K.R. Solomon
2003. Impact of perfluorooctanoic acid on the structure of the zooplankton
community in indoor microcosms. Aquat. Toxicol. 62: 227-234.
Poor experimental design/performance
Stevenson, C.N., L.A. MacManus-Spencer,
T. Luckenbach, R.G. Luthy and D. Epel
2006. New perspectives on pefluorochemical ecotoxicology: inhibition and
induction of an efflux transporter in marine mussel, Mytilus californianus.
Environ. Sci. Technol. 40: 5580-5585.
Excised cells (gills)
Sun, X., Y. Xie, X. Zhang, J. Song, and Y.
Wu
2023b. Estimation of Per- and Polyfluorinated Alkyl Substance Induction
Equivalency Factors for Humpback Dolphins by Transactivation Potencies of
Peroxisome Proliferator-Activated Receptors. Environ. Sci. Technol.57(9): 3713-
3721.
In vitro; mammalian study
Tang, J., X. Jia, N. Gao, Y. Wu, Z. Liu, X.
Lu, Q. Du, J. He, N. Li, B. Chen, J. Jiang,
W. Liu, Y. Ding, W. Zhu and H. Zhang
2018. Role of the Nrf2-ARE pathway in perfluorooctanoic acid (PFOA)-induced
hepatotoxicity inRana nigromaculata. Environ. Pollut. 238: 1035-1043.
No apical endpoints were measured; control
survival was not reported; test duration of
14 days relatively short for a chronic
amphibian study; not NA species
Thienpont, B., A. Tingaud-Sequeira, E.
Prats, C. Barata, P.J. Babin and D. Raldua
2011. Zebrafish eleutheroembryos provide a suitable vertebrate model for
screening chemicals that impair thyroid hormone synthesis. Environ. Sci. Technol.
45(17): 7525-7532.
Only one exposure concentration; no apical
endpoints
Ulhaq, M., S. Orn, G. Carlsson, J. Tallkvist
and L. Norrgren
2012. Perfluorooctanoic acid toxicity in zebrafish (Danio rerio). Abstracts.
Toxicol. Letters 21 IS: S43-S216.
Abstract only, cannot judge against data
quality objectives
Williams, T.D., A. Diab, F. Ortega, V.S.
Sabine, R.E. Godfrey, F. Falciani, J.K.
Chipman and S.G. George
2008. Transcriptomic Responses of European Flounder (Platichthys flesus) to
Model Toxicants. Aquat. Toxicol. 90(2): 83-91.
Injected toxicant; only one exposure
concentration
Xia, X., X. Chen, X. Zhao, H. Chen and M.
Shen
2012. Effects of carbon nanotubes, chars, and ash on bioaccumulation of
perfluorochemicals by Chironomus plumosus larvae in sediment. Environ. Sci.
Technol. 46: 12467-12475.
PFCs mixed in sediment
J-3
-------
Author
( iliilion
Ko;isoii I iiusod
Xia. X., A.H. Rabeai'isoa, X. Jiang and Z.
Dai
2n| i 1 ikiaiAMimulaikiii iifpcillikiioalks 1 substances h\ 1 'ii,lmi.i ni.igu.i in water
with different types and concentrations of protein. Environ. Sci. Technol. 47:
10955-10963.
Bioaccuiiiulalioii (steady stale not
documented); unmeasured test; only 3 days
Xia, X., Z. Dai, A.H. Rabearisoa, P. Zhao
and X. Jiang
2015a. Comparing humic substance and protein compound effects on the
bioaccumulation of perfluoroalkyl substances by Daphnia magna in water.
Chemosphere. 119: 978-986.
Bioaccumulation (steady state not
documented); unmeasured test; only 3 days
Xia, X., A.H. Rabaerisoa, Z. Dai, X. Jiang,
P. Zhao and H. Wang
2015b. Inhibition effect of Na+ and Ca2+ on the bioaccumulation of
perfluoroalkyl substances by Daphnia magna in the presence of protein. Environ.
Toxicol. Chem. 34(2): 429-436.
Bioaccumulation (steady state not
documented); unmeasured test; only 3 days
Yang, Z., L. Fu, M. Cao, F. Li, J. Li, Z.
Chen, A. Guo, H. Zhong, W. Li, Y. Liang,
and Q. Luo.
2023. PFAS-Induced Lipidomic Dysregulations and Their Associations with
Developmental Toxicity in Zebrafish Embryos. Sci. Total Environ.861:9 p.
Injected toxicant
Yuan, Z. J. Zhang, C. Tu, Z. Wang and W.
Xin
2016a. The protective effect of blueberry anthocyanins against perfluorooctanoic
acid-induced disturbance inplanarian (Dugesia japonica). Ecotoxicol. Environ.
Saf. 127: 170-174.
Only one exposure concentration; not NA
species; non-apical endpoints; static,
unmeasured chronic exposure
Zhang, J., B. Wang, B. Zhao, Y. Li, X.
Zhao and Z. Yuan
2019a. Blueberry anthocyanin alleviates perfluorooctanoic acid-induced toxicity
in planarian (Dugesia japonica) by regulating oxidative stress biomarkers, ATP
contents, DNA methylation and mRNA expression. Environ. Pollut. 245: 957-964.
Only one exposure concentration; not NA
species; non-apical endpoints
Zhang, H., J. He, N. Li, N. Gao, Q. Du, B.
Chen, F. Chen, X. Shan, Y. Ding, W. Zhu,
Y. Wu, J. Tang and X. Jia
2019b. Lipid accumulation responses in the liver of Rana nigromaculata induced
by perflurooctanoic acid (PFOA). Ecotoxicol. Environ. Saf. 167: 29-35.
No apical endpoints were measured; control
survival was not reported; test duration of
14 days relatively short for a chronic
amphibian study; not NA species
J-4
-------
Appendix K EPA Methodology for Fitting Concentration-
Response Data and Calculating Effect Concentrations
Toxicity values, including LC50 and EC 10 values, were independently calculated from the
data presented in the toxicity studies meeting the inclusion criteria described above (see Section
2.10.3.3) and when adequate concentrations-response data were published in the study or could
be obtained from authors. When concentration-response data were not presented in toxicity
studies, concentration-response data were requested from study authors to independently
calculate toxicity values. In cases where study authors did not respond to the EPA's request for
data or were unable to locate concentration-response data, the toxicity values were not
independently calculated by the EPA, and the reported toxicity values were retained for criteria
deviation. The EPA also retained author-reported effect concentrations when data availability did
not support effect concentration calculation by the EPA. This retention was done to be consistent
with use of author-reported toxicity values in previous criteria documents and retain informative
toxicity values (that would have otherwise not been used only on the basis of lacking the
underlying C-R data). Where concentration-response data were available, they were analyzed
using the statistical software program R (version 3.6.2) and the associated dose-response curve
(drc) package.
In some cases, the author-reported toxicity values were different than the corresponding
effect concentrations calculated by the EPA. Overall, the magnitude of such discrepancies was
limited and largely occurred for several potential reasons such as: (1) instances where authors
were presumed to calculate effect concentrations using replicate level data, but the EPA only had
access to treatment mean data; (2) the model selected to fit a particular set of C-R data, and; (3)
the software used to fit a model to C-R data and calculate an effect concentration.
K-l
-------
K.l Fitting Concentration Response Data in R
Concentration-response data were obtained from quantitatively-acceptable toxicity
studies when reported data were available. In many scenarios, toxicity studies report treatment-
level mean concentrations and mean organismal responses; however, individual-replicate data
may also be reported. When fitting C-R curves, replicate-level data were preferred over
treatment-level data, if both types of data were available. Within R, the drc package can fit a
variety of mathematical models to each set of C-R data.
K.l.l Fitting Acute Mortality Data
K. 1.1.1 Dichotomous Data
Dichotomous data are binary in nature (e.g., live/dead or 0/1) and are typical of survival
experiments. They are usually represented as a proportion survived.
K.1.2 Fitting Chronic Growth. Reproduction, and Survival Data
K. 1.2.1 Continuous Data
Continuous data take on any value along the real number line (e.g., biomass).
K.l. 2.2 Count Data
Count data take on only integer values (e.g., number of eggs hatched).
K. 1.2.3 Dichotomous Data
Dichotomous data are binary in nature (e.g., live/dead or 0/1) and are typical of survival
experiments. They are usually represented as a proportion survived.
K.2 Determining Most Robust Model Fit for Each C-R curve
The R drc package was used to fit a variety of models to each individual C-R dataset. A
single model was then selected from these candidate models to serve as the representative C-R
model. The selected model represented the most statistically-robust model available. To
K-2
-------
determine the most-statistically-robust model for a C-R dataset, all individual model fits were
assessed on a suite of statistical metrics.
K.2.1 Selecting Candidate Models
Initially, models were ranked according to the Akaike information criteria (AIC). The AIC
provides a measure of the amount of information lost for a given model by balancing goodness
of fit with model parsimony. The models with the lowest AIC, relative to other models based on
the same data, tend to be optimal. In some instances, however, the model with the lowest AIC
possessed a questionable characteristic that suggested said model was not the most appropriate.
Rather than selecting a model based solely on the lowest AIC, the initial ranking step was only
used to identify a subset of candidate models that were more closely examined before selecting a
model fit for each C-R dataset.
K.2.2 Assessment of Candidate Models to Determine the Most Appropriate Model
Candidate models (i.e., models with low AIC scores relative to other models produced for
a particular C-R dataset) were further evaluated based on additional statistical metrics to
determine a single, statistically robust curve for each quantitatively-acceptable toxicity test.
These additional statistical metrics were evaluated relative to the other candidate curve fits
produced for each C-R dataset. Of these statistical metrics, residual standard errors, confidence
intervals relative to effects concentration estimates, and confidence bands carried the most
weight in determining the most appropriate model to be representative of an individual C-R
dataset. These additional statistical metrics included:
K. 2.2.1 Comparison of residual standard errors
As with AIC, smaller values were desirable. Residual standard errors were judged
relative to other models.
K-3
-------
K. 2.2.2 Width of confidence intervals for EC estimates
Confidence intervals were assessed on standard error relative to estimate and confirming
that the intervals were non-negative. Judged in absolute and relative to other models.
K. 2.2.3 Width of confidence bands around the fitted model
A general visual inspection of the confidence bands for the fitted model. Wide bands in
the area of interest were undesirable. Judged in absolute and relative to other models.
K. 2.2.4 P-values of parameters estimates and goodness of fit tests
Hypothesis tests of parameter values to determine whether an estimate is significantly
different from zero. Goodness of fit tests were used to judge the overall performance of the
model fit. Typically, the level of significance was set at 0.05. There may have been occasional
instances where the 0.05 criterion may not be met, but there was little recourse for choosing
another model. Judged in absolute terms.
K.2.2.5 Residual plots
Residuals were examined for homoscedasticity and biasedness. Judged in absolute and
relative to other models.
K. 2.2.6 Overly influential observations
Observations were judged based on Cook's distance and leverage. When an observation
was deemed overly influential, it was not reasonable to refit the model and exclude any overly
influential observations given the limited data available with typical C-R curves. Judged in
absolute terms.
K.3 Determining Curve Acceptability for use in Criteria Derivation
The final curve fits selected for each of the quantitatively-acceptable toxicity tests were
further evaluated and classified to determine whether the curves were: 1) quantitatively-
K-4
-------
acceptable for use, 2) qualitatively acceptable for use, or 3) unacceptable. To determine curve
acceptability for use in deriving an effect concentration, each individual curve was considered
based on the statistical metrics described above and assessed visually to compare how the
calculated effect concentration aligned with the underlying raw C-R data. Instead of evaluating
curves fits relative to other curve fits for the same data (as was previously described to select the
most-robust curve for each test), curve fit metrics were used to assign each curve a score:
Quantitatively Acceptable Model. Model performed well on most/all statistical
metrics and resultant effect concentrations were typically used in a quantitative
manner.
Qualitatively Acceptable Model. Model generally performed well on statistical
metrics; however, the model presented some characteristic(s) that called estimates
into question. Such models were considered with caution. These problems may have
consisted of any number of issues such as a parameter with a high p-value, poor
goodness of fit p-value, wide confidence bands for fit or estimate interval, or
residuals that indicate model assumptions are not met. Broadly, effect concentrations
from models that were deemed qualitatively acceptable were not used numerically in
criteria derivation if quantitatively acceptable models for different endpoints or tests
from the same publication were available. If quantitatively acceptable models for
different endpoints or tests from the same publication were not available, effect
concentrations from the qualitatively acceptable model were used numerically in
criteria derivation on a case-by-case basis.
Unacceptable Model. Model poorly fit the data. These models were not used for
criteria derivation.
K-5
-------
No single statistical metric can determine a given model's validity or appropriateness. Metrics
should be considered as a whole. As such, there is a slightly subjective component to these
evaluations. That said, this assessment methodology was developed to aid in evaluating models
as to their quantitative or qualitative attributes in a transparent and relatively repeatable manner.
K-6
-------
Appendix L Derivation of Acute Protective PFOA Benchmarks
for Estuarine/Marine Waters through a New Approach
Method (NAM)
L.l Background
The 1985 Guidelines for Deriving Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and Their Uses (U.S. EPA 1985) recommend that data for a
minimum of eight families be available to fulfill taxonomic minimum data requirements (MDRs)
to calculate criteria values, including to calculate estuarine/marine aquatic life criteria. Acute
estuarine/marine test data are currently available for only three families, a mysid (Siriella and
Americamysis), a sea urchin (Strongylocentrotus), and a mussel (Mytilus), addressing only three
of the eight MDRs; thus, the EPA was not able to derive an acute estuarine/marine criterion
element for PFOA based on the 1985 Guidelines MDR specifications (Section 2.10.1). However,
the EPA was able to develop an acute PFOA protective benchmark using a New Approach
Methods (NAMs) process, via the application of Interspecies Correlation Estimation (ICE)
models (Raimondo et al. 2010). Although not a criterion based on 1985 Guidelines
specifications, because of gaps in available data for several of the taxonomic MDRs listed in the
1985 Guidelines for the derivation of aquatic life criteria, this benchmark represents an aquatic
life value derived to be protective of aquatic communities. The ICE model predictions
supplement the available test dataset to help fill missing MDRs and allow the derivation of acute
an estuarine/marine benchmark for aquatic life using procedures consistent with those in the
1985 Guidelines. This is important as it provides an approach by which values that are protective
of aquatic life communities can be developed, even when MDRs are not fulfilled by direct PFOA
test data. This approach is consistent with both the 1985 Guidelines "good science" clause, the
EPA's interest in providing useful information to states and Tribes regarding protective values
L-l
-------
for aquatic life, and the EPA's intention to reduce the use of animal testing via application of
NAMs (https://www.epa.gov/chemical-research/epa-new-approach-methods-work-plan-
reducing-use-animals-chemical-testing).
L.l.l Introduction to Web-ICE
ICE models, developed by the EPA's Office of Research and Development, are log-linear
regressions of the acute toxicity (EC50/LC50) of two species across a range of chemicals, thus
representing the relationship of inherent sensitivity between those species (Raimondo et al.
2010). Each model is derived from an extensive, standardized database of acute toxicity values
by pairing each species with every other species for which acceptable toxicity data are available.
Once developed, ICE models can be used predict the sensitivity of an untested taxon (predicted
taxa are represented by the y-axis) from the known, measured sensitivity of a surrogate species
(represented by the x-axis; Figure L-l).
ICE models have been developed for a broad range of different chemicals (e.g., metals
and other inorganics, pesticides, solvents, and reactive chemicals) and across a wide range of
toxicity values. There are approximately 3,400 significant ICE models for aquatic animal and
plant species in the most recent version of web-ICE (v3.3, https://www3.epa.gov/webiceA last
updated June 2016; Raimondo et al. 2015).
Models were validated using leave-one-out cross validation, which formed the basis for
the analyses of uncertainty and prediction robustness. For this process, each datapoint within the
model (representing the relative sensitivity of two species for a particular chemical) is
systematically removed, one at a time. The model is then redeveloped with the remaining data
(following each removal) and the removed value of the surrogate species is entered into the
model. The estimated value for the predicted species is then compared to the measured value for
that species (Raimondo et al 2010; Willming et al. 2016).
L-2
-------
ICE models have high prediction accuracy when values are derived from models with
robust parameters (e.g., mean square error, R2), that fall within a defined range of acceptability,
and with close prediction confidence intervals that facilitate evaluating the fit of the underlying
data (Brill et al. 2016; Raimondo et al. 2010; Willming et al. 2016). Results of these analyses
provide the basis of the user guidance for selecting ICE predicted toxicity with high confidence
(Box 1).
ICE models have undergone extensive peer review and their use has been supported for
multiple applications, including direct toxicity estimation for endangered species (NRC 2013;
Willming et al. 2016) and development of Species Sensitivity Distributions (SSDs) (Awkerman
et al. 2014; Bejarano et al. 2017; Dyer et al. 2006, 2008; Raimondo et al. 2010, 2020). The
application of ICE-predicted values to develop protective aquatic life values by multiple
independent, international groups confirms that values developed from ICE-generated SSDs
provides a level of protection that is consistent with using measured laboratory data (Dyer et al.
2008; Feng et al. 2013; Fojut et al. 2012a, 2012b; Palumbo et al. 2012; Wu et al. 2015, 2016;
Wang et al. 2020; Zhang et al. 2017). A recent external review of ICE models additionally
supports their use in regulatory applications based on the reliability of underlying data, model
transparency, statistical robustness, predictive reliability, proof of principle, applicability to
probabilistic approaches, and reproducibility of model accuracy by numerous independent
research teams (Bejarano and Wheeler 2020).
L-3
-------
Figure L-l. Example ICE Model for Rainbow Trout (surrogate) and Atlantic Salmon
(predicted).
Each model datapoint is a common chemical that was tested in both species to develop a log-linear regression.
Box 1. ICE Model User Guidance Recommended for Listed Species (Willming et al 2016):
Close taxonomic distance (within class)
Low Mean Squared Error (<~ 0.95)
High R2(>~ 0.6)
High slope (>~ 0.6)
Prediction confidence intervals should be used to evaluate the prediction using
professional judgement for the application (Raimondo et al. 2024).
For models between vertebrates and invertebrates, using those with lower MSE or
MOA-speciftc models (not available for PFAS) has been recommended for listed
species predictions (Willming et al. 2016).
L.1.2 Application of Web-ICE with PFOA
As previously discussed, ICE models have been developed using a diversity of
compounds (e.g., metals and other inorganics, pesticides, solvents, and reactive chemicals)
L-4
-------
across a wide range of toxicity values; however, PFAS are not included in Web-ICE v3.3 due to
the lack of available toxicity data at that time. PFAS acute values (typically reported as mg/L)
can be greater than those used to develop an ICE model (ICE database toxicity range IE"4 to IE8
(j,g/L) such that the input PFAS value of the surrogate would be outside the model domain. In
these cases, a user can either enter the value as [j,g/L and allow the model to extrapolate beyond
its range or enter the toxicity as a "scaled" value (i.e., estimate the value as mg/L). The principal
assumptions of ICE models are: 1) they represent the relationship of inherent sensitivity between
two species, which is conserved across chemicals, mechanisms of action, and ranges of toxicity
and; 2) the nature of a contaminant that was tested on the surrogate reflects the nature of the
contaminant in the predicted species (e.g., effect concentration [EC50] or lethal concentration
[LC50]), percentage of active ingredient, technical grade (Raimondo et al. 2010). While neither of
these assumptions are violated by either extrapolating beyond the range of the model or using
scaled toxicity data, the uncertainty of using ICE models in either manner had not been
thoroughly evaluated. Additionally, since PFAS were not included in the database used to
develop Web-ICE v3.3, the validation of ICE models to accurately and specifically predict to
these compounds has not been previously explored. We address both these topics in the sections
below.
L.2 Prediction Accuracy of Web-ICE for Scaled Toxicity and Values
Beyond the Model Domain
The accuracy of using scaled toxicity data as input into ICE models was evaluated using
an analysis with the existing ICE models and is described in detail in Raimondo et al. (2024).
Briefly, ICE models containing a minimum of 10 datapoints and spanning at least five orders of
magnitude were separated into two subsets: 1) a lower subset that contained all paired chemical
data corresponding to values below the 75th percentile of surrogate species values and; 2) an
L-5
-------
upper subset containing paired chemical data above the 75th percentile of surrogate values. The
lower subset was used to develop "truncated" ICE models. The surrogate species values in the
upper subset were converted to mg/L and entered into the truncated ICE models. The predicted
mg/L value was compared to the respective value of the measured predicted species. Prediction
accuracy was determined as the fold difference (maximum of the predicted/measured and
measured/predicted) between the predicted and the measured value, consistent with previously
published evaluations of ICE models (Raimondo et al. 2010; Willming et al. 2016). Accuracy of
using scaled toxicity as input into ICE models was compared to overall ICE prediction accuracy
as previously reported and prediction accuracy of the respective upper subset data points that
were entered into the models as |ig/L (i.e., values beyond the model domain). A total of 3,104
datapoints from 398 models were evaluated. A match-paired comparison showed that the
average fold differences of toxicity values predicted using scaled toxicity was not significantly
different than the respective average fold differences of all cross-validated data points reported in
Willming et al. (2016) (Wilcoxon paired rank sum test, V = 42741, p-value 0.11). Additionally,
Raimondo et al. (2010) and Willming et al (2016) showed a consistent and reproducible
relationship between the taxonomic distance of the predicted and surrogate species, which was
also reproduced using scaled values; the percentage of datapoints predicted using scaled toxicity
was within five-fold of the measured value for over 94% of all validated datapoints for species
pairs within the same order, with a reduction in accuracy coinciding with decreasing taxonomic
relatedness (Raimondo et al. 2024). Comparison of scaled values with those predicted from [j,g/L
values beyond the model domain showed that predicted values varied by a factor of 10 for
models with slopes ranging from 0.66 - 1.33. Toxicity values predicted from models with slopes
within this range had a median fold difference of 2.4 using mg/L values and 2.8 using |ig/L
L-6
-------
values (Wilcoxon paired rank sum test, V = 1334749, p-value 0.77). These results and a detailed
review of ICE model assumptions are provided in Raimondo et al. (2024)2.
L.3 Direct Comparison of Web-ICE and Measured Toxicity Values
Since limited PFOA toxicity test data are available for estuarine/marine species, the
ability of ICE models to predict PFOA toxicity was evaluated using direct comparisons of
freshwater species sensitivity as reported in the criteria document and predicted by Web-ICE. In
this comparison, the measured SMAVs for PFOA reported in Appendix A.l and Appendix B.l
were used as values for surrogate species to predict to all possible species that also had a
measured PFOA SMAV reported. The available SMAVs for PFOA that could be used as ICE
surrogate values, along with the number of ICE models corresponding to each surrogate, are
shown in Table L-l.
2 Use of scaled toxicity values and the use of surrogate toxicity values beyond the bounds of the ICE model that are
input as |ig/L are two approaches that both make extrapolations beyond the bounds of the underlying data. Actual
predictions resulting from the two approaches from the same ICE model begin to deviate from one another the
further the slope of the ICE model deviates from 1.0 (which is a primary reason why scaled toxicity data were only
employed on ICE models with slopes ranging from 0.66 - 1.33). Overall, use of the scaled approach compared to
direct extrapolation results a negligible change in the final estuarine/marine benchmark, primarily because the four
most sensitive estuarine/marine GMAVs were based on direct toxicity test results, and secondarily, because only a
subset of ICE models required use of scaled toxicity data to account for predicting beyond the bounds of the
underlying ICE model. For example, the final acute PFOA estuarine/marine benchmark was 7.0 mg/L (see section
L.5). Had the values in Table L-4 been predicted using unsealed data that were input as |ig/L only (and the model
slope requirement of 0.66 -1.33 been retained), the final acute estuarine/marine benchmark would remain unchanged
at 7.0 mg/L. Had the values in Table L-4 been predicted using unsealed data input as |ig/L only (and the model slope
requirement of 0.66 -1.33 was removed), the final acute estuarine/marine benchmark would increase slightly to 7.5
mg/L. While both approaches contain uncertainty, use of the scaled approach resulted in a more protective acute
PFOA estuarine/marine benchmark (i.e., CMC = 7.0 mg/L) than an exploratory benchmark that used acute toxicity
data estimated through direct extrapolation, with the model slope requirement of 0.66 -1.33 removed (i.e.,
exploratory CMC = 7.5 mg/L).
L-7
-------
Table L-l. Surrogate Species Measured Values for PFOA and Corresponding Number of
ICE Models for Each Surrogate.
For example there are 53 species for which P.ij'hni.i m
-------
difference" between the measured and the predicted species, such that fold difference is the
maximum of the ratio of the predicted LCso/measured LC50 or measured LCso/predicted LC50.
Analyses of ICE prediction accuracy have shown that ICE models over- and under-estimate
toxicity values randomly, i.e., there is no systematic bias associated with the models (Table L-2,
Raimondo et al. 2010; Raimondo et al. 2024). For accuracy assessments, the fold difference
provides a simplified metric to easily see how close predictions are to measured values at a
glance. A five-fold difference has been demonstrated to be the average interlaboratory variability
of acute aquatic toxicity tests and represents a conservative amount of variance under
standardized test conditions for a given life stage (Fairbrother 2008; Raimondo et al. 2010). This
inter-test variation can increase significantly where experimental variables differ between tests;
however, all ICE models are based on standardized life stages to minimize extraneous variability
(Raimondo et al. 2010).
L-9
-------
Table L-2. Comparison of ICE-predicted and measured values of PFOA for species using both scaled values (entered as mg/L) and values
potentially beyond the model domain (entered as jig/L).
Measured SMAVs are for the predicted species as listed in Appendix A.l, Appendix B.l and Table L-l. Footnotes indicate where predictions or models do not meet one or more
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351962.71
47028.33 -
2634108.91
2 gab
1085.27
358.55-3284.91
1.06
Bullfrog
Daphnid
(Daphnia magna)
1020000
250407.71
42580.36 -
1472604.12
4.07
1020
265.51
41.71 - 1689.91
3.84
(Lithobates catesbeianus)
Fathead minnow
(Pimephales promelas)
530164.94
178236.61 -
1576976.01
1.92
841.35
368.83- 1919.18
1.21
Rainbow trout
(Oncorhynchus mykiss)
3874927.81
1530344.80-
9811557.16
3.8
8676.1
4670.29- 16117.78
8.51
African clawed frog
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377000
317645.95
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19236405.56
1.19ab
377
854.79
96.46-7574.84
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Bluegill
(Lepomis macrochirus)
101324.58
39595.31 -
259290.08
4.22b
88.82
61.09- 129.12
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Mysid
Daphnid
(Daphnia magna)
24000
30882.91
15730.51 -
60630.86
1.29
24
95.62
67.73- 134.98
3.98
(Americamysis bahia)
Fathead minnow
(Pimephales promelas)
25492.89
4248.75 -
152959.48
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34.09
14.21-81.79
1.42c
Rainbow trout
(Oncorhynchus mykiss)
381425.97
119491.45 -
1217541.25
15.89bc
655.82
412.11 - 1043.64
27.33c
Bluegill
(Lepomis macrochirus)
159433.9
88808.47 -
286224.59
1.38c
256.05
191.88-341.67
1.16ฐ
Bullfrog
(Lithobates catesbeianus)
482481.09
76477.24 -
3043886.00
2.19
1174.46
256.08-5386.36
5.34
Fathead minnow
(Pimephales promelas)
94631.7
46980.95 -
190612.55
2.32c
183.63
115.53-291.86
1.2C
Daphnide
(Daphnia magna)e
Fatmucket
(Lampsilis siliquoidea)
220000e
331623.04
78520.34 -
1400577.72
1.51
220e
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31.65-389.26
1.98
Mysid
(Americamysis bahia)
22344.14
12674.03 -
39392.40
9.85
78.04
54.00- 112.78
2.82
Rainbow trout
(Oncorhynchus mykiss)
516125.21
246588.64 -
1080281.84
2.35c
1642.27
1213.56-2222.43
7.46c
Zebrafish embryo
(Danio rerio)
87831.04
13398.62-
575752.58
2.5C
55.49
7.49-410.87
3.96ac
Bullfrog
(Lithobates catesbeianus)
1501135.53
114395.51 -
19698393.38
2.26ab
601.13
166.16 - 2174.76
1.1
Bluegill
Daphnid
(Daphnia magna)
664000
81671.31
52906.69 -
126074.85
8.13
664
801.26
627.74-1022.74
1.21
(Lepomis macrochirus)
Fathead minnow
(Pimephales promelas)
218193.44
111739.02-
426067.60
3.04
357.51
244.06-523.68
1.86
Fatmucket
(Lampsilis siliquoidea)
408413.42
32355.12-
5155335.38
1,63ac
295.92
21.91 -3996.50
2.24ac
L-10
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23663.56
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40553.53
28.06
266.08
195.07-362.94
2.5d
Rainbow trout
(Oncorhynchus mykiss)
3129709.92
2269044.40 -
4316832.32
4.71
4818.01
4250.43-5461.39
7.26
Zebrafish embryo
(Danio rerio)
271173.18
34573.71 -
2126901.59
2.45a
214.38
35.96- 1278.11
3.1
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(Lepomis macrochirus)
353682.57
274243.78-
456131.98
11.31
524.72
466.57-590.13
7.63
Bullfrog
(Lithobates catesbeianus)
781571.06
313951.07-
1945695.88
5.12
376.93
177.31 -801.30
10.61
Daphnid
(Daphnia magna)
56931.33
35836.37-
90443.76
70.28c
630.92
485.50-819.91
6.34cd
Rainbow trout
(Oncorhynchus mykiss)
Fathead minnow
(Pimephales promelas)
4001000
206200.6
127654.61 -
333075.99
19.4
4,001
255.44
192.66-338.66
15.66
Fatmucket
(Lampsilis siliquoidea)
736514.01
14043.59-
38626348.23
5.43abc
201.6
39.91-1018.32
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(Americamysis bahia)
17330.67
9740.08 -
30836.72
230.86
192.8
137.35-270.65
20.75d
Zebrafish embryo
(Danio rerio)
309598.05
62068.02 -
1544288.83
12.92
91.52
30.44-275.13
43.72
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(Xenopus laevis)
524824.76
4876.97-
56477848.18
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257.81
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Bluegill
(Lepomis macrochirus)
387148.61
201381.26-
744279.98
1.53
1427.45
1024.71 - 1988.48
2.4
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(Lithobates catesbeianus)
1012520.93
299871.50-
3418793.08
1.71
758.77
320.24- 1797.81
1.28
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(Daphnia magna)
593600
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189590.43
5.31c
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1209.11 -2417.01
2.88cd
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675590
54952.40 -
8305767.53
1.14ac
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576.51 -20651.62
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34114.93
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635.95
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5.55
8360.63
6502.51 - 10749.71
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400424.37
203694.32 -
787158.28
1.48
724.18
302.07- 1736.11
1.22
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153582.08
28669.47 -
822737.61
1.07
109.67
9.40- 1278.86
1.5a
Fatmucket
Bluegill
(Lepomis macrochirus)
164400
97837.12
12334.24 -
776059.00
1,68ac
164.4
733.53
111.00-4847.23
4.46c
(Lampsilis siliquoidea)
Daphnid
(Daphnia magna)
69944.45
23967.21 -
204121.63
2.35
406.14
160.97- 1024.70
2.47
Fathead minnow
(Pimephales promelas)
23355.28
2752.96 -
198138.64
7.04ac
104.49
11.80-925.23
1.57ac
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Rainbow trout
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41091.9
3178.59 -
531223.75
^abc
1143.35
381.34-3428.01
6.95cd
Black sandshell
(Ligumia recta)
Fatmucket
(Lampsilis siliquoidea)
161000
171536.98
34311.02-
857594.23
1.07
161
237.72
25.46-2219.07
1.48a
a Confidence interval >1.5 order magnitude.
b Input data outside model range.
0 Guidance for model mean square error, R2, and/or slope not met.
d Does not meet slope criteria for using scaled toxicity (0.66-1.33).
e D.magna species mean acute value used in Appendix L is 220.0 mg/L, which differs slightly from the D. magna species mean acute value reported in Appendix A.l (i.e., 213.9
mg/L). This discrepancy was the result of including late-breaking data ahead for the release of the 2024 Final PFOA Aquatic Life Criteria document. The discrepancy in the D.
magna SMAV used here vs. Appendix A.l has no influence on the 2024 PFOA acute estuarine/marine benchmark (7.0 mg/L) because the four most sensitive genera used to derive
the PFOA acute estuarine/marine are based entirely on empirical, laboratory-based toxicity data.
L-12
-------
These comparisons are consistent with web-ICE user guidance, previously published
reports on ICE model accuracy (Raimondo et al. 2010; Willming et al. 2016), and the above
described uncertainty analysis of using scaled toxicity as model input. ICE models predict with
acceptable accuracy for PFOA when invertebrates were used to predict to invertebrate species
and vertebrates were used to predict to vertebrate species in these comparisons. Models validated
across a wide range of species, chemicals, and toxicity values show an acceptable level of
prediction accuracy (>90% values predicted within five-fold of measured value) when adhering
to the model guidance listed in Box 1 (Raimondo et al. 2010; Willming et al. 2016). The results
summarized in Sections L. 1 and L.2 demonstrate that the relationship of inherent sensitivity is
preserved across taxa, chemicals, and range of toxicity values when using robust ICE models.
While the current analysis uses freshwater species to predict to estuarine/marine species,
previous model validation and uncertainty analyses did not indicate the habitat of the species to
be an influential source of ICE model uncertainty (Raimondo et al. 2010; Willming et al. 2016).
L.4 Prediction of Estuarine/Marine Species Sensitivity to PFOA
A value of PFOA sensitivity was predicted with web-ICE v3.3 for all possible species
using all available surrogate species (Table L-l). Predicted values were obtained by entering all
available surrogate species into the web-ICE SSD generator, which predicts to all possible
species from all available surrogates simultaneously and exports results into an excel
spreadsheet. Web-ICE results were generated using both mg/L and [j,g/L values to evaluate the
full set of possible predictions using both units of measure against the model domain, confidence
intervals, and model parameters. First, all available models were evaluated based on the
parameter (MSE, R2, slope) guidance in Box 1, which are the same for an ICE species pair
regardless of input value (Table L-3). Models that did not meet the parameter criteria in Box 1
L-13
-------
were rejected in this first pass. In the next step, values that were predicted using |ig/L were
evaluated against the model domain and selected for the next tier of evaluation when the
surrogate value was within the range of data used to develop the model. If the surrogate value
reported as [j,g/L was beyond the model domain, the mg/L value was evaluated if it was within
the model domain and if the model slope was between 0.66-1.33 (Raimondo et al. 2024). Cases
in which both units were outside the model domain were not included quantitatively, but the
value with the narrowest confidence intervals was included for qualitative considerations. Values
(using either [j,g/L or mg/L input value) were excluded quantitatively from the SMAVs but
retained for qualitative consideration if an evaluation of confidence intervals, model parameters,
and the model domain indicated the relationship between surrogate and predicted species was not
informed by robust underlying data. At this stage, specific predictions should be based on
holistic evaluation of all available information provided by the model, confidence interval, and
data used to develop the model. Decisions to exclude a prediction from the SMAV are clarified
in footnotes. Because the sensitivity of a single species can be predicted by multiple surrogates,
we calculated the SMAV where multiple robust models were available for a predicted species.
Each predicted species was then assigned to the appropriate saltwater MDRs as defined in the
1985 Guidelines:
a) Family in the phylum Chordata
b) Family in the phylum Chordata
c) Either the Mysidae or Penaeidae family
d) Family in a phylum other than Arthropoda or Chordata
e) Family in a phylum other than Chordata
f) Family in a phylum other than Chordata
g) Family in a phylum other than Chordata
h) Any other family
L-14
-------
The acute sensitivity of estuarine/marine species to PFOA is presented in Table L-4. A
total of 48 models representing 21 estuarine/marine species were available in Web-ICE to predict
the toxicity of PFOA to saltwater species (Table L-4). Of these, 14 models were initially rejected
based on model parameters not meeting the guidance in Box 1, reducing the number of predicted
species to 19 represented by 34 models. Further evaluation of ICE predictions resulted in 13
SMAVs representing crustaceans, molluscs, and fish and demonstrated the potential to meet the
saltwater MDRs. The range of sensitivity for the predicted taxa is consistent with the range of
sensitivity of freshwater species for this compound.
L-15
-------
Table L-3. All ICE Models Available in Web-ICE v3.3 for Saltwater Predicted Species Based on Surrogates with Measured PFOA.
Model parameters are used to evaluate prediction robustness. Cross-validation success is the percentage of all model data that were predicted within 5-fold of the measured value
through leave-one-out cross-validation (Willming et al. 2016). Taxonomic distance describes the relationship between surrogate and predicted species (e.g., 1 = shared genus, 2 =
I'mlkl.-.l Siici-ir-
Slll l iiliiili- S|iri ir-
Slope
llllrl\r|il
1 h-iilVl1*
Ml
1- iciildin
<\-2>
|i-Milllc
Menu
S(|u;iic
I.ITMI
(MSI.)
suiTimiiir
Mndcl
Minimum
\ nine
(II!! 1 )
Sim rii-iiilr
Mndcl
M :i\i iiiiiiii
\ nine (llii 1 )
( l'((sv
\;ili(l;iliiin
SlKTCNN
("")
1 ;i\Miioiiii(
DivliUHv
1 m- in ( l ilci hi
Acartia tonsa
Daphnia magna
0.59
1.31
2
0.91
0.0443
0.17
2.24
38514.7
50
5
Rejected
Allorchestes compressa
Daphnia magna
0.83
1.59
3
0.8
0.039
0.12
5
184.54
100
5
Accepted qualitatively
Allorchestes compressa
Pimephales promelas
0.84
0.15
3
0.96
0.0028
0.02
163.05
26895.72
100
6
Accepted
Americamysis bahia
Daphnia magna
0.83
0.02
160
0.68
0
0.93
0.07
840000
64
5
Accepted
Americamysis bahia
Lepomis macrochirus
1.01
-0.92
138
0.66
0
0.95
0.13
290000
59
6
Accepted
Americamysis bahia
Oncorhynchus mykiss
0.92
-0.5
150
0.6
0
1.08
0.06
1100000
57
6
Rejected
Americamysis bahia
Pimephales promelas
0.95
-1.12
46
0.55
0
1.75
2.27
70200000
35
6
Rejected
Chelon labrosus
Lampsilis siliquoidea
1.27
1.5
1
0.99
0.0403
0
19.01
281
na
6
Accepted
Chelon macrolepis
Pimephales promelas
1.51
-1.04
2
0.97
0.0114
0.05
26
2533.38
100
4
Accepted qualitatively
Crassostrea virginica
Americamysis bahia
0.44
1.76
114
0.34
0
0.88
0.003
117648.2
55
6
Rejected
Crassostrea virginica
Daphnia magna
0.44
1.54
116
0.28
0
1.08
0.08
137171.43
58
6
Rejected
Crassostrea virginica
Lampsilis siliquoidea
0.82
-0.28
3
0.95
0.0041
0.06
30
22000
100
4
Accepted
Crassostrea virginica
Lepomis macrochirus
0.66
0.71
112
0.51
0
0.64
0.36
290000
69
6
Rejected
Crassostrea virginica
Oncorhynchus mykiss
0.59
0.97
120
0.5
0
0.68
0.02
570000
68
6
Rejected
Crassostrea virginica
Pimephales promelas
0.75
0.44
24
0.61
0
0.68
1.24
206300.75
69
6
Accepted
Cyprinodon bovinus
Lepomis macrochirus
0.66
0.7
1
0.99
0.0326
0
7.43
7326.2
na
4
Accepted
Cyprinodon bovinus
Oncorhynchus mykiss
0.72
0.8
2
0.91
0.0427
0.08
4.93
1637.92
100
4
Accepted qualitatively
Cyprinodon bovinus
Pimephales promelas
0.67
0.65
2
0.99
0.0043
0
10.49
7847.42
100
4
Accepted
Cyprinodon variegatus
Americamysis bahia
0.57
1.88
88
0.56
0
0.67
0.0025
182000
64
6
Rejected
Cyprinodon variegatus
Daphnia magna
0.53
1.79
84
0.49
0
0.72
0.08
304000
64
6
Rejected
Cyprinodon variegatus
Lampsilis siliquoidea
0.72
0.76
1
0.99
0.0392
0
30
22000
na
6
Accepted
Cyprinodon variegatus
Lepomis macrochirus
0.74
0.87
82
0.65
0
0.47
0.77
157000
82
4
Accepted
Cyprinodon variegatus
Oncorhynchus mykiss
0.75
0.9
87
0.65
0
0.56
0.82
12700000
75
4
Accepted
Cyprinodon variegatus
Pimephales promelas
0.69
0.98
24
0.74
0
0.43
2.27
16500000
77
4
Accepted
Farfantepenaeus
duorarum
Americamysis bahia
1.03
0.06
6
0.81
0.0022
0.55
0.01
720
50
4
Accepted
Farfantepenaeus
duorarum
Daphnia magna
1.08
0.14
16
0.76
0
1.32
0.04
65686.02
44
5
Rejected
Farfantepenaeus
duorarum
Lepomis macrochirus
1.16
-1.21
15
0.67
0
1.88
2.32
130000
35
6
Rejected
L-16
-------
I'mlklcd Spri-ir-
SuiTii!i;lll- S|li-cir-
Sliipr
IllUTl'Clll
1 h-iilVl1*
i>l
1' IVClllllll
<\-2>
l<-
|l-MllllC
Menu
Si|ii:iiv
I.ITMI
(MSI.)
suiTimiiir
Mi ii Ul
Minimum
\ nine
(II!! 1 )
Sim rii-iiilr
MmiIcI
Mil\iiiiiiiii
\ lllllc (II!* 1 )
( rosv
Miliiliilinii
Sllnc^
("ฆ.)
1 :i\MHMiiiii
DivliUHv
1 m- in ( rilrrin
Farfantepenaeus
duorarum
Oncorhynchus mykiss
1.2
-1.36
15
0.72
0
1.54
0.57
221000
47
6
Rejected
Fenneropenaeus
merguiensis
Daphnia magna
0.82
1.43
4
0.66
0.0473
0.4
5
1251.41
67
5
Accepted
Gasterosteus aculeatus
Lepomis macrochirus
1.15
0
3
0.95
0.0039
0.09
0.36
340
80
4
Accepted qualitatively
Gasterosteus aculeatus
Oncorhynchus mykiss
1.05
0.29
4
0.9
0.0038
0.18
0.61
890
83
4
Accepted qualitatively
Hydroides elegans
Daphnia magna
0.49
1.59
2
0.96
0.0182
0.01
5
1251.41
100
6
Rejected
Hydroides elegans
Oncorhynchus mykiss
0.2
2.3
1
0.99
0.0179
0
1.84
13390.93
na
6
Rejected
Lagodon rhomboides
Lepomis macrochirus
1.61
-2.02
1
0.99
0.0301
0
110
760
na
3
Accepted qualitatively
Litopenaeus stylirostris
Americamysis bahia
1.04
0.01
5
0.6
0.0401
0.29
0.58
24.09
57
4
Accepted qualitatively
Menidia beryllina
Lepomis macrochirus
0.79
0.9
5
0.89
0.0012
0.19
12.3
93800
86
4
Accepted
Menidia menidia
Lepomis macrochirus
1.05
-0.35
4
0.96
0.0005
0.14
2.85
97000
83
4
Accepted
Menidia menidia
Oncorhynchus mykiss
1.28
-1.4
3
0.94
0.005
0.23
11.24
91000
60
4
Accepted
Menidia peninsulae
Americamysis bahia
0.63
0.91
3
0.88
0.0162
0.32
0.01
1160
80
6
Accepted qualitatively
Menidia peninsulae
Lepomis macrochirus
0.9
-0.1
3
0.97
0.0012
0.06
0.77
2480
100
4
Accepted
Menidia peninsulae
Oncorhynchus mykiss
1.01
-0.36
2
0.91
0.0421
0.35
0.82
1600
50
4
Accepted qualitatively
Metamysidopsis
insularis
Daphnia magna
0.86
0.93
3
0.94
0.0057
0.18
6.97
317472.74
80
5
Accepted
Metamysidopsis
insularis
Lampsilis siliquoidea
1.03
0.62
2
0.99
0.0027
0.02
19.01
87705.88
75
6
Accepted
Mugil cephalus
Lepomis macrochirus
1.06
-0.15
3
0.92
0.0093
0.09
0.77
118.76
100
4
Accepted qualitatively
Mugil cephalus
Oncorhynchus mykiss
1.44
-0.37
3
0.89
0.0144
0.12
0.82
29.18
100
4
Accepted qualitatively
Tigriopus japonicus
Lepomis macrochirus
0.6
1.73
3
0.92
0.0095
0.1
1.2
11202.42
80
6
Accepted qualitatively
Tigriopus japonicus
Pimephales promelas
0.81
1.12
5
0.76
0.0103
0.11
195.14
27000
86
6
Accepted
Tisbe battagliai
Daphnia magna
0.86
1.25
2
0.94
0.0289
0.08
0.61
184.54
100
5
Accepted qualitatively
NA = Not Available
L-17
-------
Table L-4. ICE-estimated Species Sensitivity to PFOA.
Values in bold and underlined are used for SMAV.
( (imiiKiii Niimo
Prcdiclcd Species
Siimiป;i(c Species
Inpiil
I nil
I'lsliniiiled 1 o\ici(\
(inti/l.)
95V'ii ( onridence
lnler\iils imu/1.)
SMAV
Calanoid copepod
Acartia tonsa
Daphnia magna
Hg/L
30.97abc
(0.84- 1138.99)
NA
Amphipod
Allorchestes compressa
Daphnia magna
mg/L
3608.27b
(604.53 -21536.48)
306.14
Pimephales promelas
mg/L
306.14
(167.3 -560.19)
Mysid
Americamysis bahia
Daphnia magna
Hg/L
30.88
(15.73 -60.63)
52.37
Lepomis macrochirus
mg/L
88.82
(61.09- 129.12)
Oncorhynchus mykiss
Hg/L
381.43bc
(119.49- 1217.54)
Pimephales promelas
Hg/L
25.49ฐ
(4.25 - 152.96)
Thicklip mullet
Chelon labrosus
Lampsilis siliquoidea
mg/L
21448.8
(5726.99-80330.23)
21448.8
Bigscale mullet
Chelon macrolepis
Pimephales promelas
mg/L
1475.56d
(400.78 - 5432.58)
NA
Eastern oyster
Crassostrea virginica
Americamysis bahia
Hg/L
5.08ฐ
(2.55-10.1)
111.23
Daphnia magna
Hg/L
8.16bc
(3.27-20.37)
Lampsilis siliquoidea
mg/L
34.96
(13.16-92.9)
Lepomis macrochirus
Hg/L
37.78bc
(15.7-90.87)
Oncorhynchus mykiss
Hg/L
86.67bc
(31.82-236.06)
Pimephales promelas
mg/L
353.9
(161.95-773.33)
Leon springs pupfish
Cyprinodon bovinus
Lepomis macrochirus
mg/L
385.84
(100.03 - 1488.19)
363.56
Oncorhynchus mykiss
mg/L
2637. lab
(135.14-51456.4)
Pimephales promelas
mg/L
342.56
(204.07-575.01)
Sheepshead minnow
Cyprinodon variegatus
Americamysis bahia
Hg/L
24.42ฐ
(12.11 -49.24)
377.3
Daphnia magna
Hg/L
43.97ฐ
(18.29- 105.67)
Lampsilis siliquoidea
mg/L
236.12
(42.61 - 1308.27)
Lepomis macrochirus
mg/L
975.2
(695.11 - 1368.14)
Oncorhynchus mykiss
Hg/L
834.95
(288.9-2413.1)
Pimephales promelas
Hg/L
105.41
(29.42 - 377.63)
Pink shrimp
Farfantepenaeus duorarum
Americamysis bahia
mg/L
31.16
(5.31 - 182.77)
31.16
Daphnia magna
Hg/L
825.2abฐ
(47.76 - 14258.5)
Lepomis macrochirus
Hg/L
350.34abฐ
(11.8- 10398.14)
Oncorhynchus mykiss
Hg/L
4I87 i7abฐ
(97.76- 179343.3)
Banana prawn
Fenneropenaeus merguiensis
Daphnia magna
mg/L
2395.84
(373.14- 15382.93)
2395.84
Threespine stickleback
Gasterosteus aculeatus
Lepomis macrochirus
mg/L
1867.23ab
(239.79- 14539.92)
NA
Oncorhynchus mykiss
mg/L
12299.55ab
(605.28-249929.11)
L-18
-------
( (imiiKiii Niimo
Prcdiclcd Species
Siimiป;i(c Species
Inpiil
I nil
I'lsliniiiled 1 o\ici(\
(inii/l.)
95V'ii ( onridence
lnler\iils img/l.)
SMAY
Polychaete
Hydroides elegans
Daphnia magna
Hg/L
17.26abc
(1.79- 166.56)
NA
Oncorhynchus mykiss
Hg/L
4.67bc
(2.22 - 9.84)
Pinfish
Lagodon rhomboides
Lepomis macrochirus
mg/L
342.29d
(121.48-964.44)
NA
Blue shrimp
Litopenaeus stylirostris
Americamysis bahia
mg/L
28.4a
(3.76-214.22)
NA
Inland silverside
Menidia beryllina
Lepomis macrochirus
mg/L
1373.42
(509.99-3698.68)
1373.42
Atlantic silverside
Menidia menidia
Lepomis macrochirus
mg/L
419.03
(151.91 - 1155.88)
859.43
Oncorhynchus mykiss
mg/L
1762.7
(281.3 - 11045.23)
Tidewater silverside
Menidia peninsulae
Americamysis bahia
mg/L
61.5 3ad
(8.51 -444.5)
279.81
Lepomis macrochirus
mg/L
279.81
(97.91 -799.61)
Oncorhynchus mykiss
mg/L
1974.29ab
(15.61-249609.1)
Mysid
Metamysidopsis insularis
Daphnia magna
mg/L
894.87
(209.96-3814.04)
853.17
Lampsilis siliquoidea
mg/L
813.42
(361.18- 1831.9)
Striped mullet
Mugil cephalus
Lepomis macrochirus
mg/L
730.34ab
(36.69- 14537.49)
NA
Oncorhynchus mykiss
mg/L
65159.7abd
(132.99- 31925133.65)
Harpacticoid copepod
Tigriopus japonicus
Lepomis macrochirus
mg/L
2812.24d
(976.85 - 8096.09)
2432.2
Pimephales promelas
mg/L
2432.2
(825.63 -7164.93)
Harpacticoid copepod
Tisbe battagliai
Daphnia magna
mg/L
1923.22ab
(204.62 - 18075.62)
NA
NA = Not Available
a Both confidence intervals >1.5 order magnitude
b Input data outside model range
0 Guidance for model mean square error, R2, and/or slope not met
d Does not meet slope criteria for using scaled toxicity (0.66-1.33)
L-19
-------
L.5 Derivation of Acute Water Benchmark for Estuarine/Marine Water
The Web-ICE predicted acute data set for PFOA contains 10 genera representing the
eight MDR groups that would be necessary for developing an estuarine/marine criterion.
However, the EPA supplemented this dataset with acceptable quantitative study data (discussed
in Section 3.1.1.2). In scenarios where both empirical LC50 values and estimated LC50 values
were available for the same species, only the empirical data were used to derive the species mean
acute value. The ranked GMAVs for these combined data along with the MDR met by each
GMAV is summarized in Table L-5. From this dataset, an acute benchmark was calculated using
procedures consistent with the 1985 Guidelines and with those used for the derivation of
freshwater criterion values for PFOA. GMAVs for the four most sensitive genera were within a
factor of 1.5 of each other (Table L-5). The estuarine/marine FAV (the 5th percentile of the genus
sensitivity distribution) for PFOA is 14.07 mg/L (Table L-6). The FAV was lower than all of the
GMAVs for both the tested species and for values derived using Web-ICE. The FAV is then
divided by two to obtain a concentration yielding a minimal effects acute effect value. Based on
the above, the FAV/2, which is the estuarine/marine acute water column benchmark magnitude,
is 7.0 mg/L PFOA (rounded to two significant figures) and is expected to be protective of 95% of
saltwater genera potentially exposed to PFOA under short-term conditions of one-hour of
duration, if the one-hour average magnitude is not exceeded more than once in three years
(Figure L-2).
L-20
-------
Table L-5. Ranked Estuarine/Marine Genus Mean Acute Values.
Values in bold are derived from empirical PFOA toxicity tests with the species.
MDR
(.roup
( (imiiKiii Name
Species
SM.W
(niii/l.)
(iM.W
(inii/l.)
Rank
Percenlile
C
Mysid
Siriella armata
15.5
15.5
1
0.07
D
Mediterranean mussel
Mytilus galloprovincialis
17.6
17.6
2
0.14
F
Purple sea urchin
Strongylocentrotus purpuratus
20.63
20.63
3
0.21
C
Mysid
Americamysis bahia
24
24
4
0.29
F
Pink shrimp
Farfantepenaeus duorarum
31.16
31.16
5
0.36
D
Eastern oyster
Crassostrea virginica
111.2
111.2
6
0.43
E
Amphipod
Allorchestes compressa
306.1
306.1
7
0.50
A
Leon springs pupfish
Cyprinodon bovinus
363.6
370.4
8
0.57
Sheepshead minnow
Cyprinodon variegatus
377.3
B
Inland silverside
Menidia beryllina
1,373
691.2
9
0.64
Atlantic silverside
Menidia menidia
859.4
Tidewater silverside
Menidia peninsulae
279.8
C
Mysid
Metamysidopsis insularis
853.2
853.2
10
0.71
G
Harpacticoid copepod
Tigriopus japonicus
2,432
2,432
11
0.79
F
Banana prawn
Fenneropenaeus merguiensis
2,396
2,396
12
0.86
H
Thicklip mullet
Chelon labrosus
21,449
21,449
13
0.93
1: Estuarine/Marine MDR Groups
a) Family in the phylum Chordata
b) Family in the phylum Chordata
c) Either the Mysidae or Penaeidae family
d) Family in a phylum other than Arthropoda or Chordata
e) Family in a phylum other than Chordata
f) Family in a phylum other than Chordata
g) Family in a phylum other than Chordata
h) Any other family
L-21
-------
Table L-6. Estuarine/Marine Final Acute Value and Protective Aquatic Acute Benchmark.
Calculated Estuarine/Marine FAV based on 4 lowest values; n=13
Rank
Genus
GMAV
(mg/L)
ln(GMAV)
ln(GMAV)2
P=R/(N+1)
sqrt(P)
1
Siriella
15.5
2.74
7.51
0.071
0.267
2
Mytilus
17.6
2.87
8.22
0.143
0.378
3
Strongylocentrotus
20.63
3.03
9.16
0.214
0.463
4
Americcimysis
24
3.18
10.10
0.286
0.535
E (Sum):
11.81
35.00
0.71
1.64
S2 =
L =
A =
FAV =
PVAL=
2.73 S = slope
2.275 L = X-axis intercept
2.644 A = InFAV
14.07 P = cumulative probability
7.0 mg/L PFOA (rounded to two significant figures)
1 T
0.9 -ฆ
| 0.8 +
U
a
&
ฃ 0.7
| 0.6
P6
3 0.4 +
a
u
I 0.3 -F
Cm
0.2 -
0.1 -ฆ
.
O Chelon
.
~ Fenneropenaeus
~ Tigriopus
--
~ Metamysidopsis
:
O Menidia
ฆ
O Cyprinodon
~ Allorchestes
-
A Crassostrea
"
~ Farfantepenaeus
O Fish (WeblCE)
"
ฆ Americamysis
ฆ Invertebrate (Empirical)
-
"
~ Invertebrate (WeblCE)
ฆ Strongylocentrotus
~ Mollusk (Empirical)
!
~ Mytilus
A Mollusk (WeblCE)
"
ฆ Siriella
Acute Benchmark
ฆi iii
i i i i i i i I i . . i i . . i l
10 100 1,000
Genus Mean Acute Value (mg/L PFOA)
10,000
100,000
Figure L-2. Ranked Estuarine/Marine Acute PFOA GMAVs Used for the Aquatic Life
Acute Benchmark Calculation.
L-22
-------
L.5.1 Estuarine Marine/Benchmark Uncertainty
The epistemic uncertainty of individual ICE estimates used for SMAV calculation was
quantified through the calculation of corresponding 95% confidence intervals for each ICE
estimate. Of the individual models and resultant ICE-estimated LC50 values from the available
and quantitatively acceptable models (see bolded and underlined values in Table L-4; n =21), the
range of individual 95% CIs (i.e., 95% CI range = upper 95% CI - lower 95% CI) as a percent of
the corresponding LC50 estimate (i.e., = [95% CI range/LC50 estimate]*100) ranged from
69.01%) to 626.49%). The ICE model with the lowest 95% CI range relative to the LC50 estimate
(i.e., 69.01%) employed Lepomis macrochirus as the predictor species and Cyprinodon
variegatus as the predicted species. The ICE model with the largest 95% CI range relative to the
LC50 estimate (i.e., 626.49%) employed Daphnia magna as the predictor species and
Fenneropenaeus merguiensis as the predicted species. Nineteen of the 21 ICE-predicted values
in Table L-4 that were used for SMAV calculation had 95% CI ranges that were greater than the
corresponding LC50 estimate (i.e., 95% CI range was >100% of the LC50 estimate). The
relatively wide ranging 95% CIs demonstrate the underlying uncertainty in the PFOA
estuarine/marine benchmark.
Four of the 13 GMAVs used to derive the acute PFOA estuarine/marine benchmark were
based on empirical toxicity tests. The four GMAVs based on empirical data were also the four
most sensitive GMAVs in the GSD (Figure L-2), meaning final estuarine/benchmark magnitude
was primarily based on relatively certain empirical toxicity tests and the inherent uncertainty in
the ICE models had little influence on the final acute estuarine/marine benchmark magnitude.
The estuarine/marine benchmark appears adequately protective based on the available
high quality empirical data (Appendix B.l). The acute PFOA estuarine/marine benchmark (i.e.,
7.0) is more than two times lower than the lowest GMAV (i.e., 15.5 mg/L), which was based on
L-23
-------
empirical data for Siriella. The EPA further evaluated the appropriateness of the
estuarine/marine benchmark by comparing it to empirical, but qualitatively acceptable, data for
estuarine/marine species. The EPA specifically focused on qualitatively-acceptable
estuarine/marine tests reported in Table H.l that (1) tested an animal species, (2) exposed test
organisms to PFOA for a duration that was reasonably similar to standard acute exposures (e.g.,
48 hours to seven days), (3) reported acute apical effects, and (4) reported effect concentrations
that were lower than the acute estuarine/marine benchmark final acute value (i.e., 14 mg/L). The
EPA identified three individual tests in Table H. 1 as meeting the previous criteria:
1. Liu et al. (2013, 2014c) evaluated the chronic effects of PFOA (96% purity,
purchased from Sigma-Aldrich) on green mussels, Perna viridis, via a seven-day
measured, static-renewal study. A NOEC of 0.0114 mg/L and a LOEC of 0.099
mg/L was determined for a decrease in the relative condition factor (RCF). The
study was acceptable for qualitative use only because of the atypical test duration,
which is too long for an acute test and too short for a chronic test. Additionally,
the PFOA test displayed a questionable concentration-response patten where there
was no difference between the RCF at the LOEC (i.e., 0.099 mg/L) and the
highest test concentration, which contained a PFOA concentration that was more
than 10X greater (i.e., 1.120 mg/L). The large magnitude between these two
concentrations in combination with the lack of effects to the RCF observed
between the LOEC and the highest treatment concentration suggests a true
concentration-response relationship was not observed for PFOA in this test.
2. Bernardini et al. (2021) reported the results of a 21-day chronic study with the
Manila clam, Ruditapesphilippinarum. Subsamples of clams (n=20) were also
L-24
-------
collected on test day seven. No significant effects of mortality were observed in
the single treatment group throughout the exposure, including at test day seven.
The seven-day NOEC, based on mortality, was 0.00093 mg/L PFOA. Although
the seven-day NOEC is less than the acute estuarine/marine benchmark, the
authors did not report any significant effects to mortality and this study was not
useful in understanding the relative protectiveness of the acute PFOA
estuarine/marine benchmark.
3. Mhadhbi et al. (2012) conducted a six-day acute test on the turbot, Scophthalmus
maximus (a non-North American species). Endpoints included dead embryos,
malformation, hatch success at 48-hours and larvae survival (missing heartbeat
and a non-detached tail) at six days. The reported six-day LCso of 11.9 mg/L
PFOA was not used quantitatively because the test duration was longer than the
standard 96 hour acute exposure. Nevertheless, this six-day test suggests early life
stages of S. maximus may be sensitive to acute PFOA exposures. The EPA
concluded the acute PFOA estuarine/marine benchmark to be protective on the
six-day LCso reported by Mhadhbi et al. (2012) because (1) it was reasonably
similar to the most sensitive GMAV used to derive the acute estuarine/marine
benchmark (i.e., Siriella GMAV = 15.5 mg/L) and (2) the 96-hour LCso that
corresponds to the six-day LCso reported by Mhadhbi et al. (2012) was
hypothesized to be greater than or equal to the six-day LCso under the premise
that acute effect concentrations typically decrease with exposure time (until an
incipient lethal concentration is reached).
L-25
-------
4. Gebreab et al. (2022) exposed the marine mahi-mahi fish, Coryphaena hippurus
to PFOA in a 48-hour static acute test with measured PFOA treatments. Authors
stated that the test followed a new embryotoxicity assay for mahi mahi that was
adapted from assays previously developed and validated for zebrafish
embryotoxicity (Berry et al. 2007; Gebreab et al. 2020; Weiss-Errico et al. 2017).
The author-reported 48-hour LCso (i.e., 4 mg/L; 95% C.I. = 2-6 mg/L) was less
than the acute estuarine/marine FAV of 14 mg/L. Results of this test were not
used quantitatively primarily because there was 20-30% mortality in the negative
control at the end of the exposure period. The high control mortality suggests the
exposed organisms were stressed in the laboratory setting, which compounded
any effects of PFOA and likely produced an artificially low LCso value.
Therefore, this qualitatively acceptable test did not provide sufficient evidence to
conclude the estuarine/marine benchmark was underprotective of C. hippurus.
Overall, results of quantitatively- and qualitatively- acceptable empirical toxicity studies with
estuarine/marine organisms do not provide any evidence that the aquatic estuarine/marine
community will experience unacceptable chronic effects at the acute estuarine/marine PFOA
benchmark.
L-26
-------
L.6 ICE Regressions Supporting the Acute Estuarine/Marine Benchmark
-2 0 2 4 6
Americamysis bahia
(Log LC50)
Figure L-3. Americamysis bahia (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
-2 0 2 4
Americamysis bahia
(Log LC50)
Figure L-4. Americamysis bahia (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values.
L-27
-------
-2 0 2 4
Americamysis bahia
(Log LC50)
Figure h-5. Americamysis bahia (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
-2 0 2 4 6
Americamysis bahia
(Log LC50)
Figure L-6. Americamysis bahia (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values.
L-28
-------
2 3 4 5 6 7
Danio rerio- embryo
[Log LC50)
Figure L-7. Danio rerio - embryo (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
Danio rerio- embryo
(Log LC50)
Figure L-8. Danio rerio - embryo (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values.
L-29
-------
Danio rerio- embryo
[Log LC50)
Figure L-9. Danio rerio - embryo (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
2 3 4 5 6 7
Danio rerio- embryo
(Log LC50)
Figure L-10. Danio rerio - embryo (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
L-30
-------
Daphnia magna
(Log LC50)
Figure L-ll. Daphnia magna embryo (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values.
Daphnia magna
(Log LC50)
Figure L-12. Daphnia magna (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values.
L-31
-------
o o ^r
" Y\
| "
ฆg O
o
Cl
-------
-2 0 2 4 6
Daphnia magna
(Log LC50)
Figure L-15. Daphnia magna (X-axis) and Oncorhynchus mykiss (Y-axis) regression model
used for ICE predicted values.
ra
CD
E
O o
-s
ill
Daphnia magna
(Log LC50)
Figure L-16. Daphnia magna embryo (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
L-33
-------
2 3 4 5 6
Lampsilis siliquoidea
(Log LC50)
Figure L-l 7. Lampsilis siliquoidea (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
2 3 4 5 6
Lampsilis siliquoidea
(Log LC50)
Figure L-l 8. Lampsilis siliquoidea (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values.
L-34
-------
23 4
5
6
Lampsilis siliquoidea
(Log LC50)
Figure L-19. Lampsilis siliquoidea (X-axis) and Ligumia recta (Y-axis) regression model
used for ICE predicted values.
2 3 4
Lampsilis siliquoidea
(Log LC50)
Figure L-20. Lampsilis siliquoidea (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
L-35
-------
2 3 4 5 6
Lampsilis siliquoidea
(Log LC50)
Figure L-21. Lampsilis siliquoidea (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values.
-1 0 1 2 3 4 5
Lepomis macrochirus
(Log LC50)
Figure L-22. Lepomis macrochirus (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values.
L-36
-------
0 2 4 6
Lepomis macrochirus
(Log LC50)
Figure L-23. Lepomis macrochirus (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
CO -
03
CD
TD kL~ -
O _
=> O
ct m
= ฐ
CO I ฆ^1"
cd
= o
_l
!_L '
ฃ ^ -
03
CN -
2 4 6
Lepomis macrochirus
(Log LC50)
Figure L-24. Lepomis macrochirus (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values.
L-37
-------
CD
ซ S
CD ^
CD ^ CO
O 1
to CD
Q) O
I
CD ^
J~1
O
CM
Lepomis macrochirus
(Log LC50)
Figure L-25. Lepomis macrochirus (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values.
O
CO Li~
2 Q
ฃ=
>-
Lepomis macrochirus
(Log LC50)
Figure L-26. Lepomis macrochirus (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
L-38
-------
0
2
4
6
Lepomis macrochirus
(Log LC50)
Figure L-27. Lepomis macrochirus embryo (X-axis) and Pimephales promelas (Y-axis)
regression model used for ICE predicted values.
"E
o _
=3 O
cr ljd
= ฐ
co |
.52 o
o
E
ro en
Ligumia recta
(Log LC50)
Figure L-28. Ligumia recta (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values.
L-39
-------
2 4 6
Lithobates catesbeianus
(Log LC50)
Figure L-29. Lithobates catesbeianus (X-axis) and Daphnia magna (Y-axis) regression
model used for ICE predicted values.
1 2 3 4 5
Lithobates catesbeianus
(Log LC50)
Figure L-30. Lithobates catesbeianus (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values.
L-40
-------
2
4
6
Lithobates catesbeianus
(Log LC50)
Figure L-31. Lithobates catesbeianus (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
2 4 6
Lithobates catesbeianus
(Log LC50)
Figure L-32. Lithobates catesbeianus (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
L-41
-------
0 2 4 6
Oncorhynchus mykiss
(Log LC50)
Figure L-33. Oncorhynchus mykiss (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values.
Oncorhynchus mykiss
(Log LC50)
Figure L-34. Oncorhynchus mykiss (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
L-42
-------
1 2 3 4 5
Oncorhynchus mykiss
(Log LC50)
Figure L-35. Oncorhynchus mykiss (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values.
O O
o ซ
CT3 O
E
CO ^
E ฐ N
o
Q.
OJ
Oncorhynchus mykiss
(Log LC50)
Figure L-36. Oncorhynchus mykiss (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values.
L-43
-------
0
2
4
6
Oncorhynchus mykiss
(Log LC50)
Figure L-37. Oncorhynchus mykiss (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values.
0 2 4 6
Oncorhynchus mykiss
(Log LC50)
Figure L-38. Oncorhynchus mykiss (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
L-44
-------
X *
*
[ฆ
4
Pimephales prornelas
(Log LC50)
Figure L-39. Pimephales prornelas (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values.
CD
O
CD
O
Li"
c O
CD _l
ฃZ CD
jz o
Cl _l
CD " i
Q
Pimephales prornelas
(Log LC50)
Figure L-40. Pimephales prornelas (X-axis) and Daphrtia magna (Y-axis) regression model
used for ICE predicted values.
L-45
-------
3 4 5 6 7
Pimephales promelas
(Log LC50)
Figure L-41. Pimephales promelas (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values.
0 2 4 6
Pimephales promelas
(Log LC50)
Figure L-42. Pimephales promelas (X-axis) and Lepomis macrochirus (Y-axis) regression
model used for ICE predicted values.
L-46
-------
0 2 4 6
Pimephales prornelas
(Log LC50)
Figure L-43. Pimephales prornelas (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values.
0 2 4 6
Pimephales prornelas
(Log LC50)
Figure L-44. Pimephales prornelas (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
L-47
-------
2
3
4
5
Pimephales prornelas
(Log LC50)
Figure L-45. Pimephalespromelus (X-axis) and Xenopus laevis (Y-axis) regression model
used for ICE predicted values.
Xenopus laevis
(Log LC50)
Figure L-46. Xenopus laevis (X-axis) and Pimephales promelus (Y-axis) regression model
used for ICE predicted values.
L-48
-------
Appendix M Occurrence of PFOA in Abiotic Media
M.l PFOA Occurrence in Surface Waters
Table M-l. Measured Perfluorooctanoic acid (PFOA) Concentrations in Surface Waters
Across the United States.
Sliilo
W;i(erl>o(l\1
ArillinuMic
Moiin PI-OA
( oiiconli-iilioii
diji/l.)
Merihin PI-OA
( nniTiilmlion
(Wi/I.)
Riinjie ol PI-OA
Conccnli'iilions
(nป/l.)
Reference
Lake Erie
18.33
15
13-27
Sinclair and Kannan
2006
35.75
33.5
30-46
Boulanger et al. 2004
5.460
5.852
3.367-7.16
De Silva et al. 2011
1.9
1.9
1.6-2.2
Furdui et al. 2008
Lake Huron
3.222
3.475
0.656-4.72
De Silva et al. 2011
0.592
0.433
0.1-1.1
Furdui et al. 2008
Lake Michigan
1.840
1.840
0.28-3.4
Simcik and
Dorweiler 2005
4.100
3.788
3.579-5.243
De Silva et al. 2011
Lake Ontario
not provided
21
18-34
Sinclair and Kannan
2006
44.75
48.5
27-55
Boulanger et al. 2004
4.465
4.150
3.226-6.710
De Silva et al. 2011
3.773
2.900
1.8-6.7
Furdui et al. 2008
Lake Superior
0.255
0.236
0.095-0.395
De Silva et al. 2011
0.233
0.3
0.1-0.3
Furdui et al. 2008
0.246
0.124
0.074 - 0.996
Scott et al. 2010
Alabama
Waterbody in Decatur
7.5
-------
Ai-illum-lic
Moiin PI-OA
Modiiin PI-OA
R;inปe ol PI- OA
( oiiconli-iilioii
( uiiitm 1 ml ion
('oiicoiili'iilioiis
Sliilo
W;i(crho(l\1
diji/l.)
(Wi/I.)
(nji/l.)
Reference'
Big Thompson River
2.90
2.90
2.90
Blue River
1.40
1.40
1.40
Boulder Feeder Canal
<0.71
<0.71
<0.71
Boyd Lake
1.50
1.50
1.50
Cache la Poudre River
6.68
6.68
<0.72-13.0
Clear Creek
3.05
3.05
3.00-3.10
Colorado River
0.77
0.93
<0.77-1.00
Coon Creek
<0.76
<0.76
<0.76
Eagle River
1.40
1.40
1.40
East Plum Creek
<0.67
<0.67
<0.67
Erie Lake
0.81
0.81
0.81
Fairmount Reservoir
<0.81
<0.81
<0.81
Fountain Creek
11.3
13.0
4.30-15.0
Fraser River
1.10
1.10
1.10
Gore Creek
2.00
2.00
2.00
Gunnison River
<0.78
<0.78
<0.78
Horsetooth Reservoir
<0.71
<0.71
<0.71
Jackson Creek
<0.70
<0.70
<0.70
Jerry Creek
<0.77
<0.77
<0.76-<0.78
Kannah Creek
Flowline
<0.78
<0.78
<0.78
Lakewood Reservoir
<0.71
<0.71
<0.71
Little Fountain Creek
<0.73
<0.73
<0.73
Maple Grove
Reservoir
8.50
8.50
8.50
Marstron Reservoir
<0.75
<0.75
<0.75
McBroom Ditch
3.10
3.10
3.10
Mclellen Reservoir
2.20
2.20
2.20
Mesa Creek
<0.77
<0.77
<0.77
Michigan River
<0.72
<0.72
<0.72
Molina Power Plant
Tail
<0.78
<0.78
<0.78
North Fork Gunnison
River
<0.75
<0.75
<0.75
Purdy Mesa Flowline
<0.77
<0.77
<0.77
Purgatoire River
<0.73
<0.73
<0.73
Ralston Reservoir
<0.72
<0.72
<0.72
Rio Grande
<0.75
<0.75
<0.75
Roaring Fork River
0.88
0.88
0.88
San Juan River
<0.69
<0.69
<0.69
M-2
-------
Sliilo
W;i(erl>o(l\1
ArillinuMic
Moiin PI-OA
( oiiconli-iilioii
diji/l.)
Merihin PI-OA
( nniTiilmlion
(Wi/I.)
Riinjie ol PI- OA
Conccnli'iilions
(nji/l.)
Reference
Sand Creek
14.25
14.25
5.50-23.0
Severy Creek
<0.74
<0.74
<0.74
Somerville Flowline
<0.75
<0.75
<0.75
South Boulder Creek
0.74
0.74
0.74
South Platte River
9.68
11.0
4.60-14.0
St. Vrain River
5.40
5.40
5.40
Strontia Springs
<0.80
<0.80
<0.80
Taylor River
<0.70
<0.70
<0.70
Uncompahgre River
(delta)
0.73
0.73
0.73
Welton Reservoir
1.20
1.20
1.20
White River
<0.73
<0.73
<0.73
Yampa River
<0.74
<0.74
<0.74
DE, NJ, PA
Delaware River
5.95
5.24
2.12-14.9
Pan et al. 2018
Florida
Waterbody in
Pensacola
<7.5
<7.5
<7.5
3M Company 2001
Pond in Pensacola
<7.5
<7.5
<7.5
Waterbody in Port St.
Lucie
7.5
-------
Sliilo
W;i(erl>o(l\1
ArillinuMic
Moiin PI-OA
( oiiconli-iilioii
diji/l.)
Merihin PI-OA
( nniTiilmlion
(Wi/I.)
Riinjie ol PI- OA
Conccnli'iilions
(nji/l.)
Reference
Lake Calhoun
19.44
19.44
19.44
Lake Harriet
3.38
3.38
3.38
Minnesota River
1.2
1.2
1.2
Lake Tettegouche
0.47
0.47
0.47
Simcik and
Dorweiler 2005
Lake Nipisiquit
0.14
0.14
0.14
Lake Loiten
0.7
0.7
0.7
Little Trout Lake
0.31
0.31
0.31
Echo Lake Reservoir
4.9
4.9
4.9
Passaic River
13.55
13.55
13-14.1
Raritan River
8.7
8.7
8.7
Metedeconk River
31.1
31.1
28.3-33.9
Pine Lake
13.6
13.6
13.6
Horicon Lake
1.9
1.9
1.9
NJDEP 2019
New Jersey
Little Pine Lake
25.9
25.9
25.9
Mirror Lake
13.2
13.2
13.2
Woodbury Creek
7.2
7.2
7.2
Fenwick Creek
10.5
10.5
10.5
Cohansey River
4.6
4.6
4.3-4.9
Harbortown Road
3.738
3.738
3.738
Zhang et al. 2016
Passaic River
18.65
13.24
0.871-47.25
Alamogordo Domestic
Water System
<1
<1
<1
Animas River
<0.97
<0.95
<0.89-
-------
Sliilo
W;i(erl>o(l\1
ArillinuMic
Moiin PI-OA
( oiiconli-iilioii
diji/l.)
Merihin PI-OA
( nniTiilmlion
(Wi/I.)
Riinjie ol PI- OA
Conccnli'iilions
(nji/l.)
Reference
Mountain Orchard
MDWCA
<0.89
<0.89
<0.89
Pecos River
0.628
<0.96
<0.94-0.936
Rio Chama
<0.98
<0.98
<0.96-<1.0
Rio Grande
0.791
0.474
<0.86-1.95
Rio Puerco
<1.3
<1.3
<1.3
San Juan River
<1.06
<0.96
<0.89-<1.9
Tularosa Water
System
<0.89
<0.89
<0.89
New York
Washington Park
I .like
10.1
10.4
4.83-15.8
Kim and Kannan
2007
Rensselaer Lake
6.79
7.2
3.27-10.6
Iroquois Lake
7.365
not provided
not provided
Unnamed lake 1
outside Albany, NY
2.246
not provided
not provided
Unnamed lake 2
outside Albany, NY
4.341
not provided
not provided
Niagara River
19.67
19
18-22
Sinclair and Kannan
2006
Finger Lakes
not provided
14
11-20
Lake Onondaga
50.67
49.00
39-64
Lake Oneida
19
19
19
Erie Canal
38.0
30.0
25-59
Hudson River
not provided
35
22-173
Lake Champlain
not provided
24
10-46
Lower NY Harbor
2.02
2.02
2.02
Zhang et al. 2016
Staten Island
4.049
4.049
4.049
Hudson River
7.333
7.333
2.805-11.86
North Carolina
Cape Fear River
43.4
12.6
<0.2-287
Nakayama et al. 2007
Rhode Island
Narragansett Bay
1.2
1.2
1.2
Benskin et al. 2012
Allen Cove Inflow
3.784
3.784
3.784
Zhang et al. 2016
Bristol Harbor
1.168
1.170
1.014-1.320
Brook at Mill Cove
36.81
36.81
36.81
Buckeye Brook
8.455
8.455
8.455
Chickasheen Brook
1.006
1.006
1.006
EG Town Dock
1.972
1.972
1.972
Fall River
0.64
0.64
0.64
Green Falls River
0.6470
0.6470
0.5860-0.7080
Hunt River
6.978
6.978
6.978
Mill Brook
9.237
9.237
9.237
Narrow River
0.9850
0.8985
0.6630-1.480
Pawcatuck River
16.98
16.98
14.99-18.97
M-5
-------
Slsile
W;i(erl>o(l\1
ArillinuMic
Moiin PI-OA
( oiiconli-iilioii
diji/l.)
Merihin PI-OA
( nniTiilmlion
(Wi/I.)
Riinjie ol PI- OA
Conccnli'iilions
(nji/l.)
Reference
Pawtuxet River
7.546
7.546
7.546
Queens River
0.898
0.898
0.898
Sand Hill Brook
6.905
6.905
6.905
Secret Lake - Oak
Hill Brook
0.849
0.849
0.849
Slack's Tributary
2.363
2.363
2.363
South Ferry Road Pier
0.267
0.267
0.267
Southern Creek
10.08
10.08
10.08
Woonasquatucket
River
7.034
7.034
5.236-8.832
South Carolina
Charleston Harbor
9.5
not provided
not provided
Houde et al. 2006b
Tennessee
Waterbody near
Cleveland
7.5
-------
in Lake Huron (Furdui et al. 2008; De Silva et al. 2011; Simcik et al. 2005). The most western
and upstream lake within the great lakes system, Lake Superior, consistently had lower PFOA
concentrations than the other Great Lakes, with mean concentrations reported in the literature
ranging from 0.23 to 0.25 ng/L (Furdui et al. 2008; Scott et al. 2010; De Silva et al. 2011).
The higher PFOA concentrations in Lakes Erie and Ontario are likely due to higher levels
of urbanization around these lakes (Boulanger et al. 2004; Remucal 2019). A mass balance
constructed for Lake Ontario by Boulanger et al. (2005) indicated that surface water and
wastewater effluent are the major sources of PFOA to the lake. In contrast, inputs from Canadian
tributaries and atmospheric deposition of PFAS were the major contributing sources of PFOA to
Lake Superior. Inputs from Canadian tributaries and atmospheric deposition were estimated to
contribute 59% and 35% of PFOA inputs into Lake Superior, respectively (Scott et al. 2010).
Within the Great Lakes, Remucal (2019) noted there were limited PFOA data to evaluate
temporal trends in surface waters. If the dataset was limited to Lake Ontario, which is one of the
most well-studied waterbodies for PFOA occurrence in the U.S. (with data from 2002 to 2010)
there appears to be a mild decrease in PFOA concentrations over time. This decrease was likely
due to the reduction in PFOA manufacturing; however, the downward PFOA trend in Lake
Ontario was statistically insignificant, with authors noting additional data over longer time scales
were needed to fully inform conclusions (reviewed in Remucal 2019).
M-7
-------
100
10
U 0.1
0.01
Lake Lake Lake Lake Lake
Erie Huron Michigan Ontario Superior
Figure M-l. Distribution of the minimum and maximum concentrations (ng/L) of
Perfluorooctanoic acid (PFOA) measured in surface water samples collected from the
Great Lakes as reported in the publicly available literature.
This distribution is arranged alphabetically by waterbody.
M.1.2 PFOA Occurrence and Concentrations in the Midwestern U.S.
Similar PFOA concentrations are reported in the publicly available literature for
waterbodies in urban areas across the Midwest and northeastern U.S. along with lower PFOA
concentrations associated with remote areas in the same states (Newsted et al. 2017; NJ DEP
2019; Simcik and Dorweiler 2005; Sinclair et al. 2006). In Minnesota, Simcik and Dorweiler
(2005) observed PFOA concentrations ranging between 0.46 and 19 ng/L in urban areas near
Minneapolis and concentrations ranging between 0.14 to 0.7 ng/L in remote areas in northern
Minnesota.
M.1.3 PFOA Occurrence and Concentrations in the Northeastern U.S.
In New York, Sinclair et al. (2006) quantified (limit of quantification = 2 ng/L) PFOA in
all waters sampled across the state. Unlike PFOS, PFOA was detected at relatively elevated
M-8
-------
concentrations across all sites with comparatively little variability (median PFOA concentrations
across nine sites ranged from 14 to 49 ng/L) (Sinclair et al. 2006). Additionally, the New Jersey
Department of Environmental Protection (NJ DEP) measured PFOA in surface water samples
collected from 14 different sites across the state. PFOA concentrations ranged from 1.9 ng/L to a
high of 33.9 ng/L, which was quantified in the Metedeconk River downstream of a wastewater
treatment plant. NJ DEP (2019) also indicated the Metedeconk River is impacted by PFOA-
contaminated groundwater originating from an industrial source. Zhang et al. (2016) reported the
surface water median PFOA concentration was 4.05 ng/L (n = 9; Zhang et al. 2016) in 2014 in
the New York City Metropolitan area, including sites in New Jersey.
M.1.4 PFOA Occurrence and Concentrations in the Southeastern U.S.
Measured PFOA concentrations in surface waters among southeastern states of the U.S.
are highly variable with some of the highest observed concentrations occurring in specific
waterbodies near areas with PFOA manufacturing. In 2001 the 3M Company conducted a multi-
city study measuring PFOA concentrations across waterbodies with known manufacturing and/or
industrial uses of PFOA. In the 3M Company's 2001 report, PFOA concentrations from sites
with known PFOA manufacturing uses were compared to PFOA concentrations in waterbodies
with no known sources of any PFAS (3M Company 2001). In this comparison study, cities with
PFOA manufacturing uses included Mobile and Decatur, Alabama, Columbus, Georgia, and
Pensacola, Florida. Measured PFOA concentrations in surface waters, including lentic systems,
ranged from not detected (with a detection limit of 7.5 ng/L stated in the report; 3M Company
2001) to 83 ng/L in the cities with known PFOA use in manufacturing processes. These PFOA
concentrations were compared to those measured in control cities with no known PFOA
applications in manufacturing. These control cities were Cleveland, Tennessee and Port St.
Lucie, Florida. PFOA concentrations ranged from not detected to not quantified (limit of
M-9
-------
quantification = 25 ng/L; 3M Company 2001) in flowing surface waters. PFOA concentrations,
however, ranged from 97 ng/L to 748.5 ng/L in lentic systems (i.e., ponds, lakes, and reservoirs;
3M Company 2001) in St. Lucie, Florida. Lentic water samples were not collected at the other
city described as a "control," Cleveland, Tennessee. At the time of the report from the 3M
Company, the source of PFOA in lentic waters near Port St. Lucie, Florida was unknown;
however, the report noted the presence of visible litter, a greenish film on the water, and
contributions from a culvert creating a grayish plume as it entered the waterbody (3M Company
2001). Aside from the samples collected in Port St. Lucie, Florida, this report demonstrated that
measured PFOA concentrations in surface waters tend to be higher in areas with PFOA
manufacturing and/or industrial use (3M Company 2001).
In separate studies, PFOS and PFOA concentrations were measured in surface waters by
Hansen et al. (2002) near Decatur, Alabama and Konwick et al. (2008) in Georgia. Hansen et al.
(2002) studied a stretch of the Tennessee river near Decatur, Alabama and Konwick et al. (2008)
focused on the Conasauga River in Georgia as areas with known PFOA exposure and use.
Hansen et al. (2002) reported discharge from a fluorochemical manufacturing facility entered the
Tennessee River towards the middle of the sampling area of the study, allowing for a comparison
of PFOA concentrations in relation to the fluorochemical manufacturing facility. In contrast,
Konwick et al. (2008) compared the PFOA concentrations measured in the Conasauga River
with those from reference sites (i.e., not impacted) along the Altamaha River. In both studies,
mean PFOA concentrations were higher in the study areas near manufacturing sources of organic
fluorochemicals. Specifically, Hansen et al. (2002) did not detect PFOA above the limit of
quantification (i.e., 25 ng/L) upstream of the fluorochemical manufacturing facility. Downstream
of the facility, PFOA concentrations ranged from below the limit of quantification at two sites
M-10
-------
immediately downstream of the facility to 598 ng/L with a mean concentration of 335.2 ng/L.
Similarly, Konwick et al. (2008) observed higher measured PFOA concentrations in the
Conasauga River, which ranged from 32.4 to 1,150 ng/L, compared to those in the Altamaha
River, ranging between 3.0 and 3.1 ng/L. Consistent with the report from the 3M Company
summarized above, the effluent from manufacturing facilities were determined to be the source
of increased PFOA concentrations in both the Tennessee and Conasauga rivers (Hansen et al.
2002; Konwick et al. 2008). PFOA concentrations in contaminated areas of the Conasauga River
and Altamaha River were relatively consistent with those measured in Alabama and Georgia
(3M Company 2001).
Similarly, Nakayama et al. (2007) and Cochran (2015) measured PFAS, including PFOA,
in the Cape Fear Drainage Basin in North Carolina and waterbodies on Barksdale Air Force Base
in Bossier City, Louisiana; respectively. PFOA and PFOS were found to be the dominant PFAS
detected in both studies. Nakayama et al. (2007) reported PFOA exceeded the level of
quantification (i.e., 1 ng/L) in 82.3% of samples. PFOA concentrations in the Cape Fear
Drainage Basin ranged between < 1 (the lower limit of quantification) and 287 ng/L with a mean
concentration of 43.4 ng/L. Cochran (2015) detected PFOA in 64% of all water samples
collected with an average concentration of 62.67 ng/L. As in other studies summarized above,
lower PFAS concentrations were found in the smallest upland tributaries and highest in the
middle reaches of the Cape Fear River. WWTP effluents were identified as the source of PFAS
in the study area (Nakayama et al. 2007).
M.1.5 PFOA Occurrence and Concentrations in the Western U.S.
PFOA concentrations in urbanized areas of western U.S. states were consistent with
concentrations measured in northeastern states (Sinclair et al. 2006; Zhang et al. 2016) but
remained lower than contaminated areas of southeastern states (3M Company 2001). Plumlee et
M-ll
-------
al. (2008) measured PFOA in surface water samples collected from Coyote Creek and a tributary
of Upper Silver Creek in San Jose, California. PFOA concentrations in Coyote Creek ranged
from below the detection limit (4 ng/L) to 13 ng/L and from 10 ng/L to 36 ng/L in Upper Silver
Creek. Plumlee et al. (2008) postulated a combination of atmospheric deposition of volatile
precursors and surface runoff are likely sources of PFOA to both Coyote and Upper Silver
Creeks.
Dinglasan-Panlilio et al. (2014) measured PFOA concentrations along the Puget Sound in
Washington, as well as Clayoquot and Barkley Sounds in British Columbia, Canada. Broadly,
sampling locations spanned these inland marine systems and included freshwaters and
estuarine/marine waters. Overall, PFOA was detected at all sampling locations (PFOA
concentration range = 0.16 ng/L - 8.2 ng/L), but concentrations were lower than those observed
from sites with known manufacturing and/or industrial PFOA uses. These concentrations were
consistent with those reported in the publicly available literature for remote areas, such as the
northern Great Lakes and rural Minnesota (Simcik and Dorweiler 2005). Dinglasan-Panlilio et al.
(2014) speculate these specific regional sources as well as atmospheric deposition are likely
contributors of PFOA to these remote areas (Dinglasan-Panlilio et al. 2014).
M.1.6 Summary of PFOA Occurrence and Concentrations across the U.S.
Despite the wide use and persistence of PFOA in aquatic ecosystems, current information
on the environmental distribution of PFOA in surface waters across the U.S. remains limited.
Present data are largely focused from freshwater systems collected along the east coast,
southeast, and upper Midwest, focusing primarily on study areas with known manufacturing or
industrial uses of PFAS. Current data indicate that PFOA concentrations measured in U.S.
surface waters vary widely across five orders of magnitude (Figure M-2). Additional data,
M-12
-------
particularly in saltwater systems, are needed to better understand PFOA occurrence in aquatic
ecosystems.
M-13
-------
10000
1000
100
10
-------
M.1.7 Comparison to Global Occurrence of PFOA in Surface Waters
Similar to PFAS occurrence in surface waters in the U.S., generally PFOS and PFOA
were the dominant PFAS detected in surface waters around the world (Ahrens 2011). On a global
scale, concentrations of PFOA measured in surface waters generally range between pg/L and
ng/L with some concentrations in the |ig/L range. PFOA concentrations measured in the U.S.
appear to be similar to those detected in studies with sampling sites in other countries. Global
surface water PFOA concentrations reported in the public literature range between not detected
and 11,300 ng/L near a PFAS spill site (as summarized below), and Zareitalabad et al. (2013)
reported a median PFOA concentration in surface water of 24 ng/L across Canada, Europe, and
Asia.
In Canada elevated concentrations of PFOA in surface waters were generally distributed
among urbanized areas, suggesting that urban and industrial areas with high population densities
contributed to the elevated concentrations of PFOA in surface waters (Gewurtz et al. 2013; Scott
et al. 2009). In a systematic, cross-Canada study of PFAS in surface waters, Scott et al. (2009)
observed that PFOS and PFOA were the predominant PFAS detected and that generally PFOS
concentrations were higher overall, ranging between < 0.02 and 34.6 ng/L, than PFOA
concentrations, which ranged between 0.044 and 9.9 ng/L. From the systems sampled in Canada,
Scott et al. (2009) indicated that PFOA concentrations measured in Canadian surface waters
were lower than those measured in the U.S., Europe, and Asia. However, elevated PFOA
concentrations were observed in Etobicoke Creek, a tributary to Lake Ontario, after an accidental
spill of a fire-retardant foam containing perfluorinated surfactants at L.B. Pearson International
Airport in Toronto, Ontario in June 2000. The measured concentrations of PFOA ranged
between not quantified (with a quantification limit of 9 ng/L) to 11,300 ng/L (Moody et al.
2002).
M-15
-------
PFOA concentrations measured in surface waters across Europe are similar to those
observed in the U.S. Specifically, in a European Union (EU)-wide study of polar organic
persistent pollutants, Loos et al. (2009) detected PFOA in 97% of samples with a median
concentration of 3 ng/L in surface waters sampled across a wide range of sampling sites
(including contaminated and pristine rivers and streams of various sizes). However, relatively
high PFOA concentrations of nearly 200 ng/L were detected in the Po River, Italy. Mean PFOA
concentrations observed by Pan et al. (2018) were similar to those reported in Loos et al. (2009)
and across the U.S. with mean surface water concentrations from waterbodies across western
Europe, specifically the Thames River, Malaren Lake, and Rhine River, ranging between 2.31
ng/L to 8.51 ng/L, with a maximum concentration of 11.7 ng/L detected in the Thames River.
Kwadijk et al. (2010) detected PFOA in all surface water samples collected from 20 locations
across the Netherlands, with concentrations ranging from 6.5 to 43 ng/L. Huset et al. (2008)
measured similar PFOA concentrations in three rivers in the Glatt Valley Watershed,
Switzerland, and reported averages from three rivers ranging from 7.0 to 7.6 ng/L. Like in the
U.S. and Canada, elevated concentrations of PFOA in surface waters across Europe are higher in
urbanized areas and sources have been attributed to waste water treatment plant effluent, AFFF
spills, and fluorochemical manufacturing facilities (Ahrens 2011; Huset et al. 2008; Kwadijk et
al. 2010; Loos et al. 2007 and 2009).
PFOA concentrations observed in surface waters across eastern Asia were broadly similar
to the U.S., Canada, and Europe. In Japan, Saito et al. (2003) observed PFOA concentrations
ranging between 0.1 and 456 ng/L in surface waters samples collected from various locations.
Similarly, Nguyen et al. (2011) reported PFOA concentrations ranging between 5 and 31 ng/L
collected from an urbanized section of the Marina catchment in Singapore. Pan et al. (2018)
M-16
-------
reported PFOA concentrations from 112 samples across eastern Asia ranging from 0.15 ng/L to
52.8 ng/L. These 112 samples were collected from eight different water bodies, including; the
Yangtze River (median PFOA = 12.2 ng/L; n = 35), Yellow River (median PFOA = 2.45 ng/L; n
= 15), Pearl River (median PFOA = 1.82 ng/L; n = 13), Liao River (median PFOA = 9.39 ng/L;
n = 6), Huai River, (median PFOA = 6.01 ng/L; n = 9), Han River (median PFOA = 3.69 ng/L; n
= 6), Chao Lake (median PFOA = 8.17 ng/L; n = 13) and Tai Lake (median = 17.95 ng/L; n =
15).
Overall, these studies show the widespread distribution and variability of PFOA
concentration in surface waters around the world and that surrounding land use and urbanization
with high population densities have a large influence on the overall occurrence of PFOA in
surface waters (Ahrens 2011; Gewurtz et al. 2013; Loos et al. 2007; Loos et al. 2009; Scott et al.
2009).
M.2 PFOA Occurrence and Detection in Aquatic Sediments
Although aquatic sediments are not anticipated to be a major PFOA sink (Ahrens 2011;
Ahrens et al. 2009), PFOA has been detected in aquatic sediments in relatively remote regions.
For example, maximum PFOA concentrations of 1.7 |ig/kg dry weight (dw) were detected in
lake sediments in the Canadian Artie (Butt et al. 2010). Typically, in the U.S., soil and sediment
measurements of PFOA near contaminated sites and manufacturing plants occur in the |ig/kg dry
weight range. For example, PFOA ranged from below the limit of detection (0.017 |ig/kg) to 700
|ig/kg in sediments near a fluorochemical manufacturer in West Virginia (Lau et al. 2007).
Across ten U.S. Air Force bases where there is a known historic use of AFFF, Anderson
et al. (2016) measured sediment (0-1 foot below top of the sediment) samples between March-
September of 2014 at the ten locations with PFOA concentrations detected in 67% of samples.
M-17
-------
The median concentration of PFOA across all sites was 2.45 |ig/kg, with a maximum
concentration of 950 |ig/kg (Anderson et al. 2016). Lasier et al. (2011) measured PFOA in
sediment from the Coosa River, Georgia watershed, upstream and downstream of a land-
application site of municipal/industrial wastewater, at concentrations ranging from 0.06-1.97
|ig/kg dry weight. Values reported in various locations across San Francisco Bay ranged from
below detection to 0.292 |ig/kg dry weight (San Francisco Bay RWQCB 2020, Sedlak et al.
2017). Internationally, values ranged from below detection in areas with relatively low
population density to |ig/kg wet weight in areas of higher human population density, including
-------
high as ~ 23 |ig/L in ground water near a PFAS disposal site, with concentrations decreasing to <
0.1 |ig/L 1.4 km from the PFAS disposal site (Xaio et al. 2015). Despite not having been an
active-fire training area, PFOA was still present on various U.S. Air Force Installations where
there is a known history of use of AFFF to extinguish hydrocarbon-based fires. Anderson et al.
(2016) measured groundwater samples between March and September of 2014 at the ten
locations with PFOA concentrations detected in 90% of samples. The median concentration of
PFOA across all sites was 0.41 |ig/L, with a maximum concentration of 250 |ig/L (Anderson et
al. 2016). These concentrations are consistent with groundwater samples from Holloman Air
Force Base in New Mexico measured in 2017 with PFOA groundwater concentrations in
evaporation ponds and fire training areas ranging from 0.746 -254 |ig/L (NMED 2021).
PFOA was detected in groundwater samples across Minnesota in 2006/2007,
approximately five years after the 3M Corporation phased out PFOS production in Minnesota in
2002 (MPCA 2008). Analyses of samples collected from vulnerable, shallow aquifers in both
urban and agricultural areas across Minnesota, with a variety of potential contamination sources
(i.e., industrial and municipal stormwater, pesticides, land application of contaminated biosolids
and atmospheric deposition), indicated that perfluorinated chemicals were present in
concentrations of potential concern in areas beyond the disposal sites and aquifers associated
with them (MPCA 2008). Groundwater samples ranged from <0.001 to 0.0324 |ig/L with a
reporting limit of 0.025 |ig/L.
M.4 PFOA Occurrence and Detection in Air and Rain
Air concentrations of PFOA in the atmosphere is widely distributed globally. In an urban
area in Albany, NY perfluorinated acids were measured in air samples in both the gas and
particulate phase in May and July 2006 (Kim and Kannan 2007). PFOA in the gas phase had a
M-19
-------
mean concentration of 3.16 pg/m3 (range: 1.89-6.53) and in the particulate phase had a mean
concentration of 2.03 pg/m3 (range: 0.76-4.19) (Kim and Kannan 2007). Kim and Kanaan (2007)
also reported mean PFOA concentrations of 2.53 ng/L and 4.89 ng/L in rain and snow,
respectively. In an urban area in Minneapolis, MN, PFOA was measured in both the particulate
and gas phase. PFOA in the particulate phase ranged from 1.6-5.1 pg/m3 and from 1.7-16.1
pg/m3 in the gas phase across the five samples (MPCA 2008). The mean concentration value
reported from a location in Resolute Bay, Nunavut, Canada was 1.4 pg/m3 (Stock et al. 2007).
These concentrations are greater than PFOA concentrations measured in the particle phase of air
samples measured in Zeppelinstasjonen, Svalbard (Butt et al. 2010). PFOA was measured in
September and December of 2006 and August and December of 2007, with mean concentrations
of 0.44 pg/m3 (Norwegian Institute for Air Research, 2007a,b).
M.5 PFOA Occurrence and Detection in Ice
Very little information is provided about PFOA concentrations in ice. The PFOA
concentration from a Russian Arctic ice core sampled in 2007 was 131 pg/L (Saez et al. 2008;
Martin et al. 2010). During the spring of 2005 and 2006 surface snow was collected and the
following values were reported for the Canadian Arctic and Greenland, respectively: 13.1-53.7
pg/L and 50.9-520 pg/L (Young et al. 2007).
M-20
-------
Appendix N Translation of The Chronic Water Column
Criterion into Other Fish Tissue Types
The PFOA freshwater aquatic life criteria (summarized in Section 3.3) includes chronic
tissue criteria for fish whole body, fish muscle, and invertebrate whole-body tissues. Additional
values for fish liver, fish blood, and fish reproductive tissues were also calculated by
transforming the freshwater chronic water column criterion (i.e., 0.10 mg/L) into representative
tissue concentrations using tissue-specific bioaccumulation factors (BAFs). Fish liver, fish blood,
and fish reproductive BAFs were identified following the same approaches used to identify fish
wholebody, mush muscle, and inverterbrate whole body BAFs, which are described in detial in
section 2.11.3.1. Briefly, BAFs were determined from field measurements and calculated using
the equation:
BAF = (Equation N-l)
Cwater
Where:
CUota = PFOA concentration in organismal tissue(s)
Cwater = PFOA concentration in water where the organism was collected
To identify the representative BAFs, a literature search for reporting on PFOA
bioaccumulation was implemented by developing a series of chemical-based search terms to
identify studies that were reviewed for reported BAFs and/or reported concentrations in which
BAFs could be calculated for both freshwater and estuarine/marine species. BAFs from both
freshwater and estuarine/marine species were considered because; (1) inclusion of
estuarine/marine BAFs expanded the relatively limited PFOA BAF dataset and (2) Burkhard
(2021) did not specifically observe notable differences in PFAS BAFs between freshwater and
estuarine/marine systems, instead stating additional research is needed to formulate conclusions.
N-l
-------
Sources with relevant BAF information were further screened to determine if the reported BAF
information from each source was of low, medium, or high quality. Only BAFs of high and
medium quality were used to derive the tissue-specific BAFs and corresponding tissue-based
values described below.
BAFs based on reproductive tissues identified by Burkhard (2021) were further screened
to ensure only BAFs based on adult females were considered, because female reproductive
tissues are most relevant to potential maternal transfer to offspring. This subset of reproductive-
based BAFs and corresponding species and sampling locations are described in Table N-l.
Table N-l. Characteristics of adult fish sampled for the calculation of PFOA reproductive
tissue BAFs.
All sampled fish were adults, and all reproductive tissues identified as gonad. Weights, lengths, and BAFs are
Author
Species
Collection
Dale
il
Sex
Age
(\r.)
Weight
(g-\v\v)
Length
(cm)
liAl
(1 -/kg)
Ahrens et al.
2015
European perch
(Perca fluviatilis)
10/12/2012
3
F
7-9
N.R.
N.R.
3.1
Shi et al.
2015,2018
Crucian carp
(Carassius carassius)
July 2014
30
24 F
6 M
N.R.
79.4 (F)
60.5 (M)
15.0 (F)
13.7 (M)
6.59
Shi et al.
2015,2018
Crucian carp
(Carassius carassius)
July 2014
13
9 F
4 M
N.R.
352.3 (F)
320.7 (M)
24.6 (F)
24.8 (M)
4.64
Wang et al.
2016
Crucian carp
(Carassius carassius)
April 2014
8
N.R.
N.R.
(16.8-
65.1)1
(10.0-
14.7)1
85.4
N.R. = Not Reported
1 Range.
Additional details on BAF compilation and ranking can be found in Section 2.11.3.1 and
Burkhard (2021). The distributions of fish liver, fish blood, and fish reproductive BAFs
identified in the literature used to calculate tissue-specific BAFs were determined in the same
manner as invertebrate, fish muscle, and fish whole body BAFs (Section 3.2.2). Briefly,
distributions of BAFs used to derive additional tissue values were based on the lowest species-
level BAF reported within a waterbody. When more than one BAF was available for the same
N-2
-------
species at the same site, the species-level BAF was calculated as the geometric mean of all BAFs
for that species at that site. Summary statistics for the PFOA BAFs used in the derivation of the
additional tissue-based values are presented below (Table N-2) and individual BAFs are
provided in Appendix O.
Table N-2. Summary Statistics for PFOA Freshwater BAFs in Additional Fish Tissues1.
20,h
Ceomelric
Median
Cenlile
Mean li \ 1
liAl
IJAI
Mini in ii in
Maxim ii in
(l./kg-we!
(1 ./kg-w el
(1 ,/kg-wel
(1 ./kg-we!
(1 ,/kg-w el
Category
n
weight)
weight)
weigh!)
weigh!)
weigh!)
Li\ or
13
15.59
1U
2.349
0.732
1,109
Blood
5
80.71
34.1
14.90
14
636
Reproductive
Tissue
4
9.488
5.62
3.1
3.1
85.41
1 Based on the lowest species-level BAF measured at a site (i.e., when two or more BAFs were available for the
same species at the same site, the species-level geometric mean BAF was calculated, and the lowest species-level
BAF was used).
Equation N-2 was used to transform the chronic freshwater column criterion (see Section
3.2.1.3) into tissue values using the 20th centile BAFs from the distributions of BAFs described
above and summarized in Table N-2:
Tissue Value = Chronic Water Column Criterion x 20th Centile BAF (Equation N-2)
The resulting tissue values that correspond to the 20th centile tissue-specific BAFs used in
Equation N-2 are reported in Table N-3. The values reported in Table N-3 represent tissue-based
concentrations that offer a level of protection that is equal to the magnitude components of the
chronic water column criterion as well as the fish whole body, fish muscle, and invertebrate
whole body tissue-based criteria; however, the tissue-based values reported in Table N-3 are only
presented for comparative purposes and are not recommended criteria.
N-3
-------
Table N-3. PFOA Concentrations for Additional Fish Tissue Values.1'2
Category
PI-OA Concentration (nig/kซ ww)
Liver
0.2349
Blood
1.490
Reproductive Tissue
0.3100
1 These PFOA concentrations are provided as supplemental information and are not recommended criteria.
2 Tissue concentrations are expressed as wet weight (ww) concentrations.
N-4
-------
Appendix O Bioaccumulation Factors (BAFs) Used to Calculate PFOA Tissue Values
O.l Summary Table of PFOA BAFs used to calculate tissue criteria and supplemental fish tissue values
( (iiniiKiii Niiino
Scioulillc Name
Tissue'
I5AI-
(l./kii-
ซซ )
Silo
Species
IJAI-
(l./kii-
ซ ซ
KiiukiiiU
l.ociiliou
Reference
gokllish
Carassius auratus
Blood
oil u
011.0
lugh
1" silos in si\ niaior n\ crs, Korea
Lam cl al. J"0l4
mandarin
Siniperca scherzeri
Blood
739.0
739.0
high
17 sites in six major rivers, Korea
Lam et al. 2014
Crucian carp
Carassius carassius
Blood
635.6
635.6
high
Beijing Airport, China
Wang et al. 2016
European perch
Perca fluviatilis
Blood
14.00
14.00
medium
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. 2015
Cruician carp
Carassius carassius
Blood
18.50
18.50
high
Tangxun Lake
Shi et al. 2018
Cruician carp
Carassius carassius
Blood
34.06
34.06
high
Xiaoqing River, China
Shi et al. 2018
common carp
Cyprinus carpio
Blood
85.11
85.11
high
Xiaoqing River, China
Pan et al. 2017
Crucian carp
Carassius carassius
Gonad
85.41
85.41
high
Beijing Airport, China
Wang et al. 2016
European perch
Perca fluviatilis
Gonad
3.100
3.100
medium
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. 2015
Cruician carp
Carassius carassius
Gonad
4.641
4.641
high
Tangxun Lake
Shi et al. 2018
Cruician carp
Carassius carassius
Gonad
6.594
6.594
high
Xiaoqing River
Shi et al. 2018
common carp
Carassius auratus
Liver
134.0
134.0
high
17 sites in six major rivers, Korea
Lam et al. 2014
mandarin
Siniperca scherzeri
Liver
601.0
601.0
high
17 sites in six major rivers, Korea
Lam et al. 2014
Mozambique
tilapia
Oreochromis mossambicus
Liver
7.017
7.017
medium
Matikulu at N2 Bridge
Fauconier et al.
2020
Cape stumpnose
Rhabdosargus holubi
Liver
2.714
2.714
medium
Matikulu at N2 Bridge
Fauconier et al.
2020
Crucian carp
Carassius carassius
Liver
1109
1109
high
Beijing Airport, China
Wang et al. 2016
tilapia
tilapia
Liver
67.00
58.02
medium
Key River, Taiwan
Linetal. 2014
tilapia
tilapia
Liver
53.00
medium
Key River, Taiwan
Linetal. 2014
tilapia
tilapia
Liver
55.00
medium
Key River, Taiwan
Linetal. 2014
European perch
Perca fluviatilis
Liver
7.100
7.100
medium
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. 2015
European chub
Leuciscus cephalus
Liver
10.00
10.00
high
Orge River, near Paris, France
Labadie and
Chevreuil 2011
goldfish
Carassius auratus
Liver
1000
1000
high
Pearl River Delta, China
Pan et al. 2014
O-l
-------
( (iiniiKiii Niiino
Scioulillc Name
Tissue1
I5AI-
-------
Common Niimo
Scioulillc Niimo
Tissue1
I5AI-
ซซ )
Silo
Species
IJAI-
(l./kป-
ซซ)"'
KiiukiiiU
l.ociilion
Reference
European perch
Perca fluviatilis
Muscle
37.00
37.00
high
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. 2015
Brown Bullhead
Ameiurus nebulosus
Muscle
10.00
10.00
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Common Carp
Cyprinus carpio
Muscle
12.59
12.59
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Northern Pike
Esox lucius
Muscle
7.943
7.943c
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Channel Catfish
Ictalurus punctatus
Muscle
10.00
10.00
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Pumpkinseed
Lepomis gibbosus
Muscle
10.00
10.00
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Smallmouth Bass
Micropterus dolomieu
Muscle
7.943
7.943c
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Largemouth Bass
Micropterus salmoides
Muscle
7.943
7.943c
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Yellow Perch
Percaflavescens
Muscle
7.943
7.943c
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
White Crappie
Pomoxis annularis
Muscle
12.59
12.59
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Black Crappie
Pomoxis nigromaculatus
Muscle
25.12
25.12
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
2016
Adult char
Salvelinus alpinus
Muscle
5.882
5.882
high
Meretta Lake, Canadian High Arctic
Lescord et al.
2015
eel
Anguilla anguilla
Muscle
13.18
13.18
high
Netherlands
Kwadijk et al.
2010
Adult char
Salvelinus alpinus
Muscle
37.23
37.23
high
Resolute Lake, Canadian High Arctic
Lescord et al.
2015
goby
Gobio gobio
Muscle
655.6
655.6
medium
Roter Main, Upper Franconia,
Germany
Becker etal. 2010
Silver perch
Bidyanus bidyanus
Muscle
9.000
9.000
high
Shoalhaven region, Australia
Terechovs et al.
2019
crucian
Carassius cuvieri
Muscle
97.24
97.24
high
Taihu Lake, China
Fang et al. 2014
Lake Saury
Coilia mystus
Muscle
313.0
313.0
high
Taihu Lake, China
Fang et al. 2014
0-3
-------
Common Niimo
Scioulillc Niimo
Tissue1
I5AI-
-------
Common Niimo
Scioulillc Niimo
Tissue1
I5AI-
WW )
Silo
Species
IJAI-
(l./kป-
ซซ)"'
KiiukiiiU
l.ociilion
Reference
Chameleon goby
Tridentiger
trigonocephaly
WB
16273
16273
medium
Gulf Park, Xiamen Sea, China
Dai and Zeng
2019
Chinese icefish
Neosalanx tangkahkeii
taihuensis
WB
61.90
61.90
medium
Lake Chaohu, China
Pan et al. 2019
lake trout
Salvelinus namaycush
WB
31.00
156.9
high
Lake Erie
De Silva et al.
2011
Lake Trout
Salvelinus namaycush
WB
794.3
medium
Lake Erie
Furdui et al. 2007
walleye
Sander vitreus
WB
266.0
266.0
high
Lake Erie
De Silva et al.
2011
European perch
Perca fluviatilis
WB
1.000
1.000
medium
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. 2015
Lake Trout
Salvelinus namaycush
WB
3981
3981
medium
Lake Huron
Furdui et al. 2007
Lake Trout
Salvelinus namaycush
WB
2512
2512
medium
Lake Michigan
Furdui et al. 2007
lake trout
Salvelinus namaycush
WB
402.0
400.0
high
Lake Ontario
De Silva et al.
2011
Lake Trout
Salvelinus namaycush
WB
398.1
medium
Lake Ontario
Furdui et al. 2007
lake trout
Salvelinus namaycush
WB
55.00
331.3
high
Lake Superior
De Silva et al.
2011
Lake Trout
Salvelinus namaycush
WB
1995
medium
Lake Superior
Furdui et al. 2007
Juvenile char
Salvelinus alpinus
WB
77.06
77.06
high
Meretta Lake, Canadian High Arctic
Lescord et al.
2015
grass goby
Zosterisessor
ophiocephalus
WB
97.07
97.07
medium
NC Site, Orbetell lagoon, Italy
Renzi et al. 2013
Yellowfin goby
Acanthogobius flavimanus
WB
2441
2441
medium
Omuta River mouth and estuary,
Japan
Kobayashi et al
2018
Sea Bass
Lateolabrax
WB
294.1
294.1
medium
Omuta River mouth and estuary,
Japan
Kobayashi et al
2018
Grey mullet
Mugil cephalus
WB
382.4
382.4
medium
Omuta River mouth and estuary,
Japan
Kobayashi et al
2018
Juvenile char
Salvelinus alpinus
WB
5387
5387
high
Resolute Lake, Canadian High Arctic
Lescord et al.
2015
medaka
Oryzias latipes
WB
330.0
330.0
high
Seven locations across Japan
Iwabuchi et al.
2015
0-5
-------
Common Niimo
Scioulillc Niimo
Tissue1
I5AI-
ซซ )
Silo
Species
IJAI-
(l./kป-
ซซ)"'
KiiukiiiU
l.ociilion
Reference
common shiner
Notropis cornutus
WB
7.586
7.586
high
Spring/Etobicoke Creek, Toronto,
Canada
Awad et al. 2011
Cruician carp
Carassius carassius
WB
2.775
2.775
high
Tangxun Lake
Shi et al. 2018
Bleak
Alburnus alburnus
WB
199.5
199.5
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
Common carp
Cyprinus carpio
WB
158.5
158.5
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
Mullet
Liza sp.
WB
100.0
100.0
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
Roach
Rutilus rutilus
WB
125.9
125.9
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
Rudd
Scardinius erythrophtalmus
WB
79.43
79.43
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
European catfish
Silurus glanis
WB
125.9
125.9
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
Ebro chub
Squalius laietanus
WB
100.0
100.0
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
2017
Cruician carp
Carassius carassius
WB
3.713
3.713
high
Xiaoqing River, China
Shi et al. 2018
Snail
Gastropoda
Invert
10.58
10.58
medium
Matikulu, N2 Bridge
Fauconier et al.
2020
zooplankton
zooplankton
Invert
199.5
199.5
medium
Baltic Sea
Gebbink et al.
2016
Ghost Crab
Ocypode stimpsoni
Invert
7440
7440
medium
Fenglin, Xiamen Sea, China
Dai and Zheng
2019
Copepods
Copepoda
Invert
358.0
29.92
medium
Gironde estuary, SW France
Munoz et al. 2017
Copepods
Copepoda
Invert
2.500
high
Gironde estuary, SW France
Munoz et al. 2019
Brown shrimp
Crangon crangon
Invert
458.0
33.94
medium
Gironde estuary, SW France
Munoz et al. 2017
brown shrimp
Crangon crangon
Invert
2.500
high
Gironde estuary, SW France
Munoz et al. 2019
Oyster
Crassostrea gigas
Invert
20.00
20.00
high
Gironde estuary, SW France
Munoz et al. 2017
Gammarids
Gammarus
Invert
1250
1250
medium
Gironde estuary, SW France
Munoz et al. 2017
Mysids
Mysidacea
Invert
108.0
16.43
medium
Gironde estuary, SW France
Munoz et al. 2017
mysids
Mysidacea
Invert
2.500
high
Gironde estuary, SW France
Munoz et al. 2019
White shrimp
Palaemon longirostris
Invert
377.0
29.45
medium
Gironde estuary, SW France
Munoz et al. 2017
0-6
-------
Common Niimo
Scioulillc Niimo
Tissue1
I5AI-
-------
(0111111011 Niimo
Scioulillc Niimo
Tissue1
I5AI-
-------
0.2 Summary of PFOA BAFs used to calculate tissue criteria and
supplemental fish tissue values
Field measured BAFs used to calculate fish and invertebrate PFOA tissue criteria (fish
muscle, fish whole body, invertebrate whole body) and supplemental fish tissue values (blood,
reproductive tissue, liver) are shown in Appendix 0.1. Summary statistics for the BAFs from this
table used to derive tissue criteria and additional tissue values (i.e., lowest species-level BAF
from each site) are reported in Table 3-11 and Table N-2, respectively. Rankings for individual
BAFs were determined by Burkhard (2021), who devised a ranking system based on five
characteristics: 1) number of water samples; 2) number of tissue samples; 3) spatial coordination
of water and tissue samples; 4) temporal coordination of water and tissue samples; and 5) general
experimental design. For the first four characteristics, a score of one to three was assigned, based
on number of samples or how closely the water and tissue measurements were paired. For the
experimental design characteristic, a default value of zero was assigned; unless the measured
tissues were composites of mixed species, in which case it was assigned a three (Burkhard 2021).
These sub-scores were then summed and assigned a rank based on the final score. Studies with
high quality rankings had scores of four or five, studies with medium quality rankings had scores
of five or six, and studies with low quality rankings had scores of seven or higher (Burkhard
2021). Parameters for the scores assigned to the five characteristics are listed in Table 2-2, and
additional details can be found in Burkhard (2021). Only BAFs from studies with high or
medium quality rankings were included for the final BAF geometric mean calculations used to
derive tissue criteria (Table 3-12) and supplemental tissue values (Table N-3).
0-9
-------
0.3 PFOA BAFs References
Ahrens, L., K. Norstrom, T. Viktor, A.P. Cousins and S. Josefsson. 2015. Stockholm Arlanda
Airport as a source of per- and polyfluoroalkyl substances to water, sediment and fish.
Chemosphere. 129: 33-38.
Awad, E., X. Zhang, S.P. Bhavsar, S. Petro, P.W. Crozier, E.J. Reiner, R. Fletcher, S.A.
Tittlemier and E. Braekevelt. 2011. Long-term environmental fate of perfluorinated compounds
after accidental release at Toronto Airport. Environ. Sci. Tech. 45: 8081-8089.
Becker, A.M., S. Gerstmann and H. Frank. 2010. Perfluorooctanoic acid and perfluorooctane
sulfonate in two fish species collected from the Roter Main River, Bayreuth, Germany. Bull.
Environ. Contam. Toxicol. 84: 132-135.
Bhavsar, S.P., C. Fowler, S. Day, S. Petro, N. Gandhi, S.B. Gewurtz, C. Hao, X. Zhao, K.G.
Drouillard and D. Morse. 2016. High levels, partitioning and fish consumption based water
guidelines of perfluoroalkyl acids downstream of a former firefighting training facility in
Canada. Environ. Internal 94: 415-423.
Burkhard, L. P. 2021. Evaluation of published bioconcentration factor (BCF) and
bioaccumulation factor (BAF) data for per-and polyfluoroalkyl substances across aquatic
species. Environ. Toxicol. Chem. 40(6): 1530-1543.
Dai, Z. and F. Zheng. 2019. Distribution and bioaccumulation of perfluoroalkyl acids in Xiamen
coastal waters. J. Chem. 36: 1-8.
De Silva, A. O., C. Spencer, B. F. Scott, S. Backus and D. C. Muir. 2011. Detection of a cyclic
perfluorinated acid, perfluoroethylcyclohexane sulfonate, in the Great Lakes of North America.
Environ. Sci. Technol. 45(19): 8060-8066.
De Solla, S.R., A.O. De Silva and R.J. Letcher. 2012. Highly elevated levels of perfluorooctane
sulfonate and other perfluorinated acids found in biota and surface water downstream of an
international airport, Hamilton, Ontario, Canada. Environ. Internat. 39: 19-26.
Fang, S., X. Chen, S. Zhao, Y. Zhang, W. Jiang, L. Yang and L. Zhu. 2014. Trophic
magnification and isomer fractionation of perfluoroalkyl substances in the food web of Taihu
Lake, China. Environ. Sci. Tech. 48: 2173-2182.
Fauconier, G., T. Groffen, V. Wepener and L. Bervoets. 2020. Perfluorinated compounds in the
aquatic food chains of two subtropical estuaries. Sci. Total Environ. 719: 135047
Furdui, V.I., N.L. Stock, D A. Ellis, C.M. Butt, D M. Whittle, P.W. Crozier, E.J. Reiner, D.C.G.
Muir and S.A. Mabury. 2007. Spatial distribution of perfluoroalkyl contaminants in lake trout
from the Great Lakes. Environ. Sci. Technol. 41(5): 1554-1559.
O-10
-------
Gebbink, W.A., A. Bignert and U. Berger. 2016. Perfluoroalkyl acids (PFAAs) and selected
precursors in the Baltic Sea environment: Do precursors play a role in food web accumulation of
PFAAs? Environ. Sci. Technol. 50(12): 6354-6362.
Iwabuchi, K., N. Senzaki, S. Tsuda, H. Watanabe, I. Tamura, H. Takanobu andN. Tatarazako.
2015. Bioconcentration of perfluorinated compounds in wild medaka is related to octanol/water
partition coefficient. Fundam. Toxicol. Sci. 2(5): 201-208.
Kobayashi, J., Y. Maeda, Y. Imuta, F. Ishihara, N. Nakashima, T. Komorita and T. Sakurai.
2018. Bioaccumulation patterns of perfluoroalkyl acids in an estuary of the Ariake Sea, Japan.
Bull. Environ. Contam. Toxicol. 100: 536-540.
Koch, A., A. Karrman, L.W.Y. Yeung, M. Jonsson, L. Ahrens and T. Wang. 2019. Point source
characterization of per- and polyfluoroalkyl substances (PFASs) and extractable organofluorine
(EOF) in freshwater and aquatic invertebrates. Environ. Sci. Process Impacts. 21: 1887-1898.
Kwadijk, C., P. Korytar and A. Koelmans. 2010. Distribution of perfluorinated compounds in
aquatic systems in the Netherlands. Environ. Sci. Technol. 44(10): 3746-3751.
Labadie, P. and M. Chevreuil. 2011. Partitioning behaviour of perfluorinated alkyl contaminants
between water, sediment and fish in the Orge River (nearby Paris, France). Environ. Pollut. 159:
391-397.
Lam, N.-H., C.-R. Cho, J.-S. Lee, H.-Y. Soh, B.-C. Lee, J.-A. Lee, N. Tatarozako, K. Sasaki, N.
Saito, K. Iwabuchi, K. Kannan and H.-S. Cho. 2014. Perfluorinated alkyl substances in water,
sediment, plankton and fish from Korean rivers and lakes: A nationwide survey. Sci. Total
Environ. 491-492: 154-162.
Lee, Y.-M., J.-Y. Lee, M.-K. Kim, H. Yang, J.-E. Lee, Y. Son, Y. Kho, K. Choi and K.-D. Zoh.
2020. Concentration and distribution of per- and polyfluoroalkyl substances (PFAS) in the Asan
Lake area of South Korea. J. Haz. Mat. 381: 120909.
Lescord, G. L., K. A. Kidd, A. O. De Silva, M. Williamson, C. Spencer, X. W. Wang and D. C.
G. Muir. 2015. Perfluorinated and polyfluorinated compounds in lake food webs from the
Canadian High Arctic. Environ. Sci. Technol. 49: 2694-2702.
Lin, A. Y.-C., S.C. Panchangam, Y.-T. Tsai and T.-H. Yu. 2014. Occurrence of perfluorinated
compounds in the aquatic environment as found in science park effluent, river water, rainwater,
sediments, and biotissues. Environ. Monit. Assess. 186: 3265-3275.
Loi, E. I., L. W. Yeung, S. Taniyasu, P. K. Lam, K. Kannan and N. Yamashita. 2011. Trophic
magnification of poly- and perfluorinated compounds in a subtropical food web. Environ. Sci.
Technol. (45): 5506-5513.
O-ll
-------
Munoz, G., H. Budzinski, M. Babut, H. Drouineau, M. Lauzent, K.L. Menach, J. Lobry, J.
Selleslagh, C. Simonnet-Laprade and P. Labadie. 2017. Evidence for the trophic transfer of
perfluoroalkylated substances in a temperate macrotidal estuary. Environ. Sci. Technol. 51:
8450-8459.
Munoz, G., H. Budzinski, M. Babut, J. Lobry, J. Selleslagh, N. Tapie and P. Labadie. 2019.
Temporal variations of perfluoroalkyl substances partitioning between surface water, suspended
sediment, and biota in a macrotidal estuary. Chemosphere. 233: 319-326.
Pan, C.-G., J.-L. Zhao, Y.-S. Liu and Q.-Q. Zhang. 2014. Bioaccumulation and risk assessment
of per- and polyfluoroalkyl substances in wild freshwater fish from rivers in the Pearl River
Delta region, South China. Ecotox. Environ. Saf. 107: 192-199.
Pan, X., J. Ye, H. Zhang, J. Tang and D. Pan. 2019. Occurrence, removal and bioaccumulation of
perfluoroalkyl substances in Lake Chaohu, China. Int. J. Environ. Res. Public Health. 16(10):
1692.
Pan, Y., H. Zhang, Q. Cui, N. Sheng, L.W.Y. Yeung, Y. Guo, Y. Sun and J. Dai. 2017. First
report on the occurrence and bioaccumulation of hexafluoropropylene oxide trimer acid: An
emerging concern. Environ. Sci. Tech. 51: 9553-9560.
Pignotti, E., G. Casas, M. Llorca, A. Tellbuscher, D. Almeida, E. Dinello, M. Farre and D.
Barcelo. 2017. Seasonal variations in the occurrence of perfluoroalkyl substances in water,
sediment and fish samples from Ebro Delta (Catalonia, Spain). Sci. Total Environ. 607-608:
933-943.
Renzi, M., C. Guerranti, A. Giovani, G. Perra and S.E. Focardi. 2013. Perfluorinated
compounds: Levels, trophic web enrichments and human dietary intakes in transitional water
ecosystems. Mar. Pollut. Bull. 76: 146-157.
Shi, Y., R. Vestergren, Z. Zhou, X. Song, L. Xu, Y. Liang and Y. Cai. 2015. Tissue distribution
and whole body burden of the chlorinated polyfluoroalkyl ether sulfonic acid F-53B in crucian
carp (Carassius carassius): Evidence for a highly bioaccumulative contaminant of emerging
concern. Environ. Sci. and Technol. 49: 14156-14165.
Shi Y., R. Vestergren, T.H. Nost, Z. Zhou and Y. Cai. 2018. Probing the differential tissue
distribution and bioaccumulation behavior of per-and polyfluoroalkyl substances of varying
chain-lengths, isomeric structures and functional groups in crucian carp. Environ. Sci. Technol.
52: 4592-4600.
Terechovs, A. K. E., A.J. Ansari, J.A. McDonald, S.J. Khan, F.I. Hai, N.A. Knott, J. Zhou and
L.D. Nghiem. 2019. Occurrence and bioconcentration of micropollutants in Silver Perch
(Bidyanus bidyanus) in a reclaimed water reservoir. Sci. Total Environ. 650 (Part 1): 585-593.
0-12
-------
Wang, Y., R. Vestergren, Y. Shi, D. Cao, L. Xu, X. Zhao and F. Wu. 2016. Identification, tissue
distribution, and bioaccumulation potential of cyclic perfluorinated sulfonic acids isomers in an
airport impacted ecosystem. Environ. Sci. Tech. 50: 10923-10932.
Wilkinson, J.L., P.S. Hooda, J. Swinden, J. Barker and S. Barton. 2018. Spatial (bio)
accumulation of pharmaceuticals, illicit drugs, plasticisers, perfluorinated compounds and
metabolites in river sediment, aquatic plants and benthic organisms. Environ. Pollut. 234: 864-
875.
Xu, J., C. S. Guo, Y. Zhang and W. Meng. 2014. Bioaccumulation and trophic transfer of
perfluorinated compounds in a eutrophic freshwater food web. Environ. Pollut. 184: 254-261.
Zhou, Z., Y. Shi, L. Xu and Y. Cai. 2012. Perfluorinated compounds in surface water and
organisms from Baiyangdian Lake in North China: Source profiles, bioaccumulation and
potential risk. Bull. Environ. Contam. Toxicol. 89: 519-524.
0-13
-------
Appendix P Example Data Evaluation Records (DERs)
Background: This set of published literature was identified using the ECOTOXicology
database (ECOTOX; https://cfpub.epa. gov/ecotox/) as meeting data quality standards. ECOTOX
is a source of high-quality toxicity data for aquatic life, terrestrial plants, and wildlife. The
database was created and is maintained by the EPA, Office of Research and Development,
Center for Computational Toxicology and Exposure. The ECOTOX search generally begins with
a comprehensive chemical-specific literature search of the open literature conducted according to
ECOTOX Standard Operating Procedures (SOPs). The search terms are often comprised of
chemical terms, synonyms, degradates and verified Chemical Abstracts Service (CAS) numbers.
After developing the literature search strategy, ECOTOX curators conduct a series of searches,
identify potentially applicable studies based on title and abstract, acquire potentially applicable
studies, and then apply the applicability criteria for inclusion in ECOTOX. Applicability criteria
for inclusion into ECOTOX generally include:
1. The toxic effects are related to single chemical exposure (unless the study is being
considered as part of a mixture effects assessment);
2. There is a biological effect on live, whole organisms or in vitro preparation including
gene chips or omics data on adverse outcome pathways potentially of interest;
3. Chemical test concentrations are reported;
4. There is an explicit duration of exposure;
5. Toxicology information that is relevant to OW is reported for the chemical of
concern;
6. The paper is published in the English language;
7. The paper is available as a full article (not an abstract);
8. The paper is publicly available;
9. The paper is the primary source of the data;
10. A calculated endpoint is reported or can be calculated using reported or available
information;
11. Treatment(s) are compared to an acceptable control;
12. The location of the study (e.g., laboratory vs. field) is reported; and
13. The tested species is reported (with recognized nomenclature).
P-l
-------
Following inclusion in the ECOTOX database, toxicity studies are subsequently
evaluated by the Office of Water. All studies were evaluated for data quality generally as
described by U.S. EPA (1985) in the 1985 Guidelines and in the EPA's Office of Chemical
Safety and Pollution Prevention (OCSPP)'s Ecological Effects Test Guidelines (U.S. EPA
2016b), and EPA OW's internal data quality SOP, which is consistent with OCSPP's data quality
review approach (U.S. EPA 2018). These toxicity data were further screened to ensure that the
observed effects could be primarily attributed to PFOA exposure. Office of Water completed a
DER for each species by chemical combination from the PFOA studies identified by ECOTOX.
Example DERs are presented here to convey the meticulous level of evaluation, review, and
documentation each PFOA study identified by ECOTOX was subject to. Appendix P. 1 shows an
example fish DER and Appendix P.2 shows an example aquatic invertebrate DER.
P-2
-------
P.l Example Fish DER
Part A: Overview
I. Test Information
Chemical name:
CAS name: CAS Number:
Purity: Storage conditions:
Solubility in Water (units):
Controlled Experiment Field Study/Observation (Place X by One)
(imanipulated) {not manipulated)
Primary Reviewer: Date: EPA Contractor {Place X by One)
Secondary Reviewer: Date: EPA Contractor {Place X by One)
{At least one reviewer should be from EPA for sensitive taxa)
Citation: Indicate: author(s), year, study title, journal, volume, and pages.
(e.g., Slonim, A.R. 1973. Acute toxicity of beryllium sulfate to the common guppy. J. Wat. Pollut. Contr. Fed. 45(10): 2110-2122)
Companion Papers: Identify any companion papers associated with this paper using the citation format above.
{Ifyes, list file names of
Were other DERs completed for Companion Papers? Yes No DERs below)
Study Classification for Aquatic Life Criteria Development: Place X by One Based on Highest Use
Acceptable for Quantitative Use
Acceptable for Qualitative Use
Not Acceptable for Use/Unused
General Notes: Provide any necessary details regarding the study's use classification for all pertinent endpoints,
including non-apical endpoints within the study (e.g., note all study classifications for each endpoint if the use varies)
Major Deficiencies (note any stated exclusions): Check all that apply. Checking any of these items make the study "Not
Acceptable for Use"
, , ,,, ,, , . , . . No Controls (for controlled experiments
Mixture (tor controlled experiments only) only)
Excessive Control Mortality (> 10% for acute and > 20% for chronic)
, 1 1 , < Bioaccumulation: steady state not
Dilution water not adequately characterized , ,
reached
Dermal or Injection Exposure Pathway
Review paper or previously published without modification
Other: (if any, list here)
P-3
-------
POTENTIAL CHEMICAL MIXTURES : Describe any potential chemicals mixtures as characterized by study
authors (including any confirmation of chemical mixtures).
DESCRIPTION OF DILUTION WATER: Describe concerns with characterization of and/or major deficiencies
with dilution water.
General Notes:
Minor Deficiencies: List and describe any minor deficiencies or other concerns with test. These items may make the study
"Acceptablefor Qualitative Use" (exceptions may apply as noted)
For Field Studies/Observations: A field study/observation may be considered "Acceptable for Quantitative Use" if it
consisted of a range of exposure concentrations and the observed effects are justifiably contributed to a single chemical
exposure
Mixture (observed effects not justifiably contributed to single chemical exposure)
Uncharacterized Reference Sites/Conditions
POTENTIAL CHEMICAL MIXTURES PRESENT AT SITE: Describe any potential chemicals mixtures present at
the site as characterized by study authors (including any confirmation of chemicals present at study site).
EXPOSURE VARIABILITY ACROSS STUDY SITE(SV Describe any exposure variability across study site(s)
as characterized by study authors (i.e., description of study design with reference and contaminated sites).
General Notes:
Reviewer's Comments: Provide additional comments that do not appear under other sections of the DER.
P-4
-------
ABSTRACT: Copy and paste abstract from publication.
SUMMARY: Fill out and modify as needed.
Acute:
Species (lil'cMiiuc)
Method'
1 CM
Diinilioii
( hcmiciil
/ Pnriu
pll
1 cm p.
(ฐC)
lliii'dncss
(lllli/l. ilS
( ;i( Oil
or
S;ilini(\
(ppl)
DOC
(mป/l.)
r.iTcci
Reported
I'.ITccl
( oiiccnlriilioii
(mji/l.)
Verified
I'.ITccl
C oncciiI i';i 1 ion
(iiili/l.)
Chissil'iciilion
Quantitative /
Qualitative / Unused
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
Chronic:
Species (liresliiuc)
Method'
lesl
Diinilioii
( hcmicid
/ Pnriu
pll
Temp.
(ฐC)
lliirdncss
(111ii/l. ilS
CaCO.o
(ii*
S;ilinii\
(Dl)ll
!)()(
(inii/l.)
Chronic
Limits
Reported
Chronic
Value
(niii/l. or
Verified
(lironic
Value
(ni^/l. or
|AU/^)
(lironic
\ iilne
l.ndpoinl
Cliissiriciilion
Quantitative /
Qualitative /
Unused
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
P-5
-------
II. Results Provide results as reported in the publication (including supplemental materials). Include screen shots of tables and/or
figures reporting results from the article following tabulated data table in each associated results section for all studies. Complete
tabulated data tables for all studies for studies marked "Acceptable for Quantitative Use" and "Acceptable for Qualitative Use".
Water Quality Parameters: If only general summary data of water quality parameters is provided by study authors (i.e., no
specific details of water quality parameters on a treatment level is provided), summarize any information regarding water quality
parameters under General Notes below and indicate data not provided in Table A.II.l.
General Notes: For aquatic life criteria development, measured water quality parameters in the treatments nearest the toxicity
test endpoint(s), e.g., LC50, EC20, etc., are most relevant.
Table A.II.l. Measured Water Quality Parameters in Test Solutions.
Dissolved oxygen, temperature, pH and [other parameters (hardness, salinity, DOC)] in test solutions during the /A'/-day
exposure of [test organism] to [concentration of treatments)] of [test substance] under [static renewal/flow-through]
conditions.
Pa ram el er
Trealmenl
Mean
Range
Dissolved
[1]
Oxygen
[2]
(% saturation
or mg/L)
j
j
[I]
Temperature
[2]
(Q
j
j
[1]
pH
[2]
j
j
Other (e.g.,
hardness,
salinity, DOC)
[I]
[2]
j
j
P-6
-------
Chemical Concentrations: Summarize the concentration verification data from test solutions/media. Expand table to include
measured concentration data for each media type (i.e., water, diet, muscle, liver, blood, etc.).
General Notes: Provide any necessary detail regarding the measured concentrations, including any identified cause for
substantial differences between nominal and measured concentrations, if samples were collected on separate days (and if so provide
details), and any potential cross contamination.
Table A.II.2. Measured (and Nominal) Chemical Concentrations in Test Solutions/Media.
[Analytical Method] verification of test and control concentrations during an [X]-day exposure of [test organism] to [test
substance] under [static renewal/flow-through] conditions.
| Mciin |
Nil in her of
ISliindiird
Nomiiiiil
Mc.isuml
S;i in pies
l)c\ hilion or
(oiiitii trillion
( <>iK'i'iilr;ilion
Nil in hoi' til'
Non-
Ik'low \on-
Siiindiird
1 iviilmonl
(iinils)
(iinils)
S;i in pies
IK'Icci'1
Ik'kcl
I'.rrorl
Kiinuo
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
aNon-Detect: 0 = measured and detected; 1= measured and not detected; if not measured or reported enter as such
P-7
-------
Mortality: Briefly summarize mortality results (if any).
General Notes: Comment on concentrations response relationship and slope of response ifprovided. Compare mortality in
treatments with control group and/or the reference chemical.
Table A.II.3. Mean Percent [Mortality or Survival].
Mean percent mortality [or number of immobilized, survival] of [test organism] exposed to [test substance] for [test duration]
under [static/renewal/flow-through] conditions.
|Mo;in
|M;ni(l;inl lk'\ iiilion
1 iviilmonl
Mnr(;ili(\ |
or Siiindiinl l.rmr|
Control
[1]
[2]
[3]
[4]
[5]
[6]
[LCx]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
P-8
-------
Growth: Briefly summarize growth results (if any).
General Notes: Comment on concentrations response relationship and slope of response ifprovided. Compare growth endpoints
in treatments with control group and/or the reference chemical.
Table A.II.4. Mean [Growth].
Mean growth [length and/or weight] of [test organism] exposed to [test substance] for [test duration] under
[static/renewal/flow-through] conditions.
Menu (.row(h
Mcsiii IVrcenl
11 h/\\ oi vih 11
ISiiindiinl IK'\ hilinn
( hiiiiiic in | l.cn^lh/
ISiiindiird l)c\ hiiiun
1 iviilmonl
(ฆฆnils)
or Siiindiird l-'rrซir|
liioiiiiissj
or Siiindiird llrrorj
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
[ECx]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
P-9
-------
Reproductive: Briefly summarize reproduction endpoint results (if any). For multi-senerational studies, copy and paste Table
A.II.5 below for each generation with reproductive effects data.
General Notes: Comment on concentrations response relationship and slope of response ifprovided. Compare reproductive
endpoints in treatments with control group and/or the reference chemical.
Table A.II.5. Mean [Reproductive] Effect.
Mean [reproductive] effects for [generation] of [test organism] exposed to [test substance] for [test duration] under
[static/renewal/flow-through] conditions.
ISliindiinl
|Siiindiird
IMciin
ISliindiinl
ISliindiinl
l)c\ in 1 ion
l)c\ iiilion
IlillCll
l)c\ iiilion
|Moiin
Do iiilion or
IMciin
or
|Moiin
or
IV iron I
or
1 iviilmonl
Number of
Sliindiird
Nil ill hoi' til'
Sliindiird
IVrci'iH
Sliindiird
Snr\ i\ ill
Sliindiird
(units)
S|ป;i\\ lis |
I'.rrorl
1''-litis |
I'.rrorl
11 illcll |
I'.rrorl
Posll
I'.rrorl
Control
[11
[21
[31
[4]
[5]
[6]
./'
[ECxl
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
P-10
-------
Sublethal Toxicity Endpoints: Include other sublethal effect (s), including behavioral abnormalities or other signs of toxicity,
if any. Copy Table A.II.6 as needed to provide details for each sublethal effect observed.
General Notes: Briefly summarize observed sublethal effects otherwise not captured in the results table(s) below.
Table A.II.6. Mean [Sublethal] Effect.
Mean /"Sublethal effect, (e.g., behavioral abnormalities, etc.)] in [test organism] during [test duration (acute/chronic)]
exposure to [test substance] under [static/renewal/flow-through] conditions.
|Mciin Siihlolhiil
Rcsponsc'l
|S|;iikI;ii'(I l)c\ iiiiion or
1 IVillllHMII
( ii nils)
Siiindiird l.rmr|
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
[ECx]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control
Reported Statistics: Copy and paste statistical section from publication
P-ll
-------
Part B: Detailed Review
I. Materials and Methods
Protocol/Guidance Followed: Indicate ifprovided by authors.
Deviations from Protocol: If authors report any deviations from the protocol noted above indicate here.
Study Design and Methods: Copy and paste methods section from publication.
TEST ORGANISM: Provide information under Details and any relevant or related information or clarifications in Remarks.
Parameler
Details
Remarks
Species:
Common Name:
Scientific Name:
North American species?
Surrogate for North American
Taxon?
(Place X if applicable)
Strain/Source:
Wild caught from unpolluted areas [1]
o Quarantine for at least 14 days or until they are
disease free, before acclimation [1]
Must originate from same source and population [1]
Should not be used:
o If appeared stressed, such as discoloration or
unusual behavior [1]
o If more than 5% die during the 48 hours before
test initiation [1]
o If they were used in previous test treatments or
controls [2]
No treatments of diseases may be administered:
o Within 16 hour of field collection [1]
o Within 10 days or testing or during testing [11
Age at Study Initiation:
Acute:
Juvenile stages preferred [1]
Chronic:
Life-cycle test:
o Embryos or newly hatched young < 48 hours old
[2]
Partial life-cycle test:
o Immature juveniles at least 2 months prior to
active gonad development [2]
Early life-stage test:
o Shortly after fertilization [21
Was body weight or length recorded at
test initiation?
Yes No
Was body weight or length recorded at
regular intervals?
Yes No
If yes, describe regular intervals:
P-12
-------
STUDY PARAMETERS: Provide information under Details and any relevant information of deficiencies in Remarks.
Complete for both Controlled Experiments and Field Studies Observations.
Paramcler
Details
Remarks
Number of Replicates per Treatment
Control(s):
Group:
At least 2 replicates/treatment recommended for
acute tests [1]
At least 2 replicates/treatment recommended for
chronic tests 131
Treatment(s):
Number of Organisms per Replicate/
Control(s):
Treatment Group:
At least 10 organisms/treatment recommended [3]
At least 7 organisms/treatment acceptable 141
Treatment(s):
5
ฃ
ฆs-
|
ฃ
s
-s
Exposure Pathway:
(i.e., water, sediment, gavage, or diet).
Note: all other pathways (e.g., dermal, single dose via
gavage, and injection) are unacceptable.
Exposure Duration:
Acute
Should be 96 hours [2]
Chronic
Life-cycle tests:
o Ensure that all life stages and life processes are
exposed [2]
o Begin with embryos (or newly hatched young),
continue through maturation and reproduction, and
should end not less than 24 days (90 days for
salmonids) after the hatching of the next
generation [2]
Partial life-cycle tests:
o Allowed with species that require >1 year to reach
sexual maturity, so that all major life stages can be
exposed to the test material in <15 months [2]
o Begin with immature juveniles at least 2 months
prior to active gonad development, continue
through maturation and reproduction, and end not
less than 24 days (90 days for salmonids) after the
hatching of the next generation [2]
Early life-cycle tests:
o 28 to 32 day (60 day post hatch for salmonids)
exposures from shortly after fertilization through
embryonic, larval, and early juvenile development
121
Acute
Partial Life Cycle
Early Life Stage
Full Life Cycle
Other (please remark):
Test Concentrations (remember units):
Nominal:
Recommended test concentrations include at least three
Measured:
concentrations other than the control; four or more will
provide a better statistical analysis [31
Media measured in:
Observation Intervals:
Should be an appropriate number of observations
over the study to ensure water quality is being
properly maintained 141
P-13
-------
CONTROLLED EXPERIMENT STUDY PARAMETERS: Provide information under Details and any relevant
information of deficiencies in Remarks. Complete for Controlled Experiments only.
-*
Paramcler
Delails
Remarks
Acclimaliou/lloldiug:
Should be placed in a tank along with the water in
which they were transported
o Water should be changed gradually to 100%
dilution water (usually 2 or more days) [1]
o For wild-caught animals, test water temperature
should be within 5ฐC of collection water
temperature [1]
o Temperature change rate should not exceed 3ฐC
within 72 hours [1]
To avoid unnecessary stress and promote good
health:
o Organisms should not be crowded [1]
o Water temperature variation should be limited [1]
o Dissolved oxygen:
ฆ Maintain between 60 - 100% saturation [1]
ฆ Continuous gentle aeration if needed [1]
o Unionized ammonia concentration in holding and
acclimation waters should be < 35 ng/L [11
Duration:
analysis (ij any):
Feeding:
Water type:
Temperature (ฐC):
Dissolved Oxygen (mg/L):
Health (any mortality observed?):
Acclimation followed published guidance?
Describe, if any
Yes No
If yes, indicate which guidance:
Test Vessel:
Test chambers should be loosely covered [1]
Test chamber material:
o Should minimize sorption of test chemical from
water [1]
o Should not contain substances that can be leached
or dissolved in solution and are free of substances
that could react with exposure chemical [1]
o Glass, No. 316 stainless steel, nylon screen and
perfluorocarbon (e.g. Teflon) are acceptable [1]
o Rubber, copper, brass, galvanized metal, epoxy
glues, lead and flexible tubing should not come
into contact with test solution, dil. Water, or stock
[1]
Size/volume should maintain acceptable biomass
loading rates (see Biomass Loading Rate below) [11
Material:
Briefly describe the test vessel:
Size:
Fill Volume:
Test Solution Delivery System/Method:
Flow-through preferred for some highly volatile,
hydrolysable or degradable materials [2]
o Concentrations should be measured often enough
using acceptable analytical methods [2]
Chronic exposures:
o Flow-through, measured tests required [2]
Test Concentrations Measured
Yes No
Test Solution Delivery System:
Static
Renewal
Indicate Interval:
Flow-through
Indicate Type of Diluter:
Source of Dilution Water:
Freshwater hardness range should be < 5 mg/L or <
10% of the average (whichever is greater) [1]
Saltwater salinity range should be < 2 g/kg or < 20%
of the average (whichever is greater) [1]
Dilution water must be characterized (natural surface
water, well water, etc.) [3]
o Distilled/deionized water without the addition of
appropriate salts should not be used [2]
Dilution water in which total organic carbon or
particulate matter >5 mg/L should not be used [2]
o Unless data show that organic carbon or particulate
matter do not affect toxicity [21
Dilution Series (e.g., 0.5x, 0.6x, etc.):
P-14
-------
Paramcler
Details
Remarks
Dilution Water Parameters:
Measured at the beginning of the experiment or
averaged over the duration of the experiment (details of
water quality parameters measured in test solutions
should be included under the results section)
Dissolved Oxygen (mg/L):
pH:
Temperature (ฐC):
Hardness (mg/L as CaCCh):
Salinity (ppt):
Total Organic Carbon (mg/L):
Dissolved Organic Carbon (mg/L):
Aeration:
Acceptable to maintain dissolved oxygen at 60 -
100% saturation at all times [1]
Avoid aeration when testing highly oxidizable,
reducible and volatile materials [1]
Turbulence should be minimized to prevent stress on
test organisms and/or re-suspend fecal matter [1]
Aeration should be the same in all test chambers at all
times [11
Yes No
Describe Preparation of Test
Concentrations (e.g., water exposure,
diet):
Test Chemical Solubility in Water:
List units and conditions (e.g., 0.01% at 20X2)
Were concentrations in water or diet
verified by chemical analysis?
Measured test concentrations should be reported in
Table A.II.2 above.
Yes No
Indicate media:
Were test concentrations verified by
chemical analysis in tissue?
Measured test concentrations can be verified in test
organism tissue (e.g., blood, liver, muscle) alone if a
dose-response relationship is observed.
Measured test concentrations should be reported in
Table A.II.2 above.
Yes No
Indicate tissue type:
If test concentrations were verified in test organism
tissue, was a dose-response relationship observed?
Were stability and homogeneity of test
material in water/diet determined?
Yes No
Was test material regurgitated/avoided?
Yes No
Solvent/Vehicle Type (Water or Dietary):
When used, a carrier solvent should be kept to a
minimum concentration [1]
Should not affect either survival or growth of test
organisms [1]
Should be reagent grade or better [ 1 ]
Should not exceed 0.5 ml/L (static) or 0.1 ml/L (flow
through) unless it was shown that higher
concentrations do not affect toxicity 131
Negative Control:
Yes No
Reference Toxicant Testing:
Yes No
If Yes, identify substance:
Other Control: If any (e.g. solvent control)
P-15
-------
Biomass Loading Rate:
Loading should be limited so as not to affect test
results. Loading will vary depending on temperature,
type of test (static vs. flow-through), species,
food/feeding regime, chamber size, test solution
volume, etc. [1]
This maximum number would have to be determined
for the species, test duration, temperature, flow rate,
test solution volume, chamber size, food, feeding
regime, etc.
Loading should be sufficiently low to ensure:
o Dissolved oxygen is at least 60% of saturation
(40% for warm-water species) [1,5]
o Unionized ammonia does not exceed 35 ng/L [1]
o Uptake by test organisms does not lower test
material concentration by > 20% [1]
o Growth of organisms is not reduced by crowding
Generally, at the end of the test, the loading (grams of
organisms; wet weight; blotted dry) in each test
chamber should not exceed the following:
o Static tests: >0.8 g/L (lower temperatures); >0.5
g/L (higher temperatures) [1]
o Flow through tests: > 1 g/L/day or > 10 g/L at any
time (lower temperatures); >0.5 g/L/day or > 5
g/L at any time (higher temperatures) [1]
Lower temperatures are defined as the lower of 17ฐC
or the optimal test temperature for that species [11
P-16
-------
Piii'iiiiK'Icr
Deliiils
Keniiii'ks
-*
Feeding:
Unacceptable for acute tests [2]
o Exceptions:
ฆ Data indicate that the food did not affect the
toxicity of the test material [2]
ฆ Test organisms will be severely stressed if they
are unfed for 96 hours [2]
ฆ Test material is very soluble and does not sorb
or complex readily (e.g., ammonia) [21
Yes No
Lighting:
Depends on the type of test (acute or chronic) and
endpoint (e.g., reproduction) of interest.
o Embryos should be incubated under dim
incandescent lighting (< 20 fc) or total darkness
during early life-stage toxicity testing
o Embryos must not be subjected to prolonged
exposure to direct sunlight, fluorescent lighting, or
high intensity incandescent lighting
Generally, ambient laboratory levels (50-100 fc) or
natural lighting should be acceptable, as well as a
diurnal cycle consisting of 50% daylight or other
natural seasonal diurnal cycle.
Artificial light cycles should have a 15 - 30-minute
transition period to avoid stress due to rapid increases
in light intensity [11
Study Design/Methods Classification: (Place Xby One Based on Overall Study Design/Methods Classification)
Provide details of Major or Minor Deficiencies/Concerns with Study Design in Associated Sections of Part A: Overview
This classification should be taken into consideration for the overall study classification for aquatic life criteria development in Part A.
Study Design Acceptable for Quantitative Use
Study Design Acceptable for Qualitative Use
Study Design Not Acceptable for Use
Additional Notes: Provide additional considerations for the classification of study use based on the study design.
P-17
-------
OBSERVATIONS: Provide information under Details and any relevant information in Remarks. This information should be
consistent with the Results Section in Part A.
Parameler
Details
Remarks
Parameters measured including sublethal
effects/toxicity symptoms:
Common Apical Parameters Include:
Acute
EC50 based on percentage of organisms exhibiting
loss of equilibrium plus the percentage of organisms
immobilized plus percentage of organisms killed [2]
0 If not available, the 96-hr LC50 should be used [2]
Chronic
Life-cycle/Partial Life-cycle test:
0 Survival and growth of adults and young,
maturation of males and females, eggs spawned
per female, embryo viability (salmonids only), and
hatchability [2]
Early life-cycle test:
0 Survival and growth 121
List parameters:
Was control survival acceptable?
Acute
> 90% control survival at test termination [2]
Chronic
> 80% control survival at test termination [21
Yes No
Control survival (%):
Were individuals excluded from the
analysis?
Yes No
If yes, describe justification provided:
Was water quality in test chambers
acceptable?
If appropriate, describe any water quality issues
(e.g., dissolved oxygen level below 60% of
saturation)
Yes No
Availability of concentration-response
data:
Were treatment level concentration-response
data included in study publication (can be from
tables, graphs, or supplemental materials)?
Specify endpoints in remarks
Yes No
Were replicate level concentration-response
data included in study publication (can be from
tables, graphs, or supplemental materials)?
specify endpoints in remarks
Yes No
If treatment and/or replicate level
concentration-response data were included, how
was data presented? (check all that apply)
Tables
Graphs
Supplemental Files
Were concentration-response data estimated
from graphs study publication or supplemental
materials?
Yes No
If yes, indicate software used:
Yes No
Should additional concentration-response data
be requested from study authors?
If concentration-response data are available, complete
Verification of Statistical Results (Part C) for sensitive
species.
Requested by:
Request date:
Date additional data received:
P-18
-------
Part C: Statistical Verification of Results
I. Statistical Verification Information: Report the statistical methods (e.g., EPA TRAP, BMDS, R, other) used to verify the
reported study or test results for the five (5) most sensitive genera and sensitive apical endpoints (including for tests where such
estimates were not provided). If values for the LC50, LT50 and NOEC are greater than the highest test concentration, use the "> "
symbol.
Primary Reviewer: Date: EPA Contractor {Place X by One)
Secondary Reviewer: Date: EPA Contractor {Place X by One)
{At least one reviewer should be from EPA for sensitive taxa)
Endpoint(s) Verified:
Additional Calculated Endpoint(s):
Statistical Method (e.g., TRAP, BMDS, R, other):
II. Toxicity Values: Include confidence intervals if applicable
NOEC:
LOEC:
MATC:
ECs:
EC10:
EC20:
ECso or LCso
Dose-Response Curve Classification: (PlaceXby One)
This classification should be taken into consideration for the overall study classification for aquatic life criteria development in Part A
Dose-Response Curve Acceptable for Quantitative Use
Dose-Response Curve Acceptable for Qualitative Use
Dose-Response Curve Not Acceptable for Use
Summary of Statistical Verification: Provide summary of methods used in statistical verification.
Additional Notes:
Attachments:
1. Provide attachments to ensure all data used in Part C are captured, whether from study results reported in the publication
and/or from additional data requested from study authors
Data from study results of the publication should be reported in Results section of Part A
Additional data provided upon request from study authors should be reported in Table C.II.l below and original
correspondence with study authors should be included as attachments
2. Model assessment output (including all model figures, tables, and fit metrics)
3. Statistical code used for curve fitting
P-19
-------
III. Attachments: Include all attachments listed above after the table below.
Additional Data Used in Response-Curve: Provide all data used to fit dose-response curve not captured in Results section of PER above in Part A. Add rows as needed.
First row in italicized text is an example.
Table C.II.1 Additional Data Used in Dose-Response Curve.
('ur\c II)
I'lidpiiinl
liviilnunl
Ki-|>lii;ik-
|Sl;i iul;i I'd
l)i-\ iiiiiim
fir
S|;iihI;ii-(I
Kitoi'I
#of
Sun ixiirs
V
k1
11'
kl'spilllsi-
Kl'SpilllM-
l nil
ClIIH'
Chih' uiiils
Alchronicl
Ceriodaphnia dubia
#of
young/female
0
6
10
10
1
18
count
0.03
mg/L
aN = number of individuals per treatment; k = number of replicates per treatment level; n = number of individuals per replicate
P-20
-------
Part I): References to Test Guidance
1. ASTM Standard E 739, 1980. 2002. Standard guide for conducting acute toxicity tests on
test materials with fishes, macroinvertebrates, and amphibians. ASTM International,
West Conshohocken, PA.
2. Stephan, C.E., D.I. Mount, D.J. Hansen, J.H. Gentile, G.A. Chapman and W.A. Brungs.
1985. Guidelines for Deriving Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and their Uses. PB85-227049. National Technical
Information Service, Springfield, VA.
3. Stephan, C.E. 1995. Review of results of toxicity tests with aquatic organisms. Draft.
U.S. EPA, MED. Duluth, MN. 13 pp.
4. OECD 203. 1992. Test No. 203: Fish, Acute Toxicity Test. OECD Guidelines for the
Testing of Chemicals, Section 2, OECD Publishing, Paris,
https://doi. org/10.1787/9789264069961 -en.
5. American Public Health Association (APHA). 2012. Standard methods for the
examination of water and wastewater. Part 8000 - Toxicity. APHA. Washington, DC.
P-21
-------
P.2 Example Aquatic Invertebrate DER
Part A: Overview
I. Test Information
Chemical name:
CAS name:
Purity:
Solubility in Water (units):
Controlled Experiment
(imanipulated)
Primary Reviewer:
Secondary Reviewer:
(At least one reviewer should be from EPA for sensitive taxa)
CAS Number:
Storage conditions:
Field Study/Observation
(inot manipulated)
Date:
Date:
{Place X by One)
EPA
EPA
Contractor (Place X by One)
Contractor (Place X by One)
Citation: Indicate: author (s), year, study title, journal, volume, and pages.
(e.g., Keller, A.E and S.G. Zam. 1991. The acute toxicity of selected metals to the freshwater mussel, Anodonta imbecilis. Environ. Toxicol. Chem. 10(4): 539-546.)
Companion Papers: Identify any companion papers associated with this paper using the citation format above.
Were other DERs completed for Companion Papers?
Yes
(Ifyes, list file names of
No DERs below)
Study Classification for Aquatic Life Criteria Development:
Acceptable for Quantitative Use
Acceptable for Qualitative Use
Not Acceptable for Use/Unused
General N otes: Provide any necessary details regarding the study's use classification for all pertinent endpoints, including
non-apical endpoints within the study (e.g., note all study classifications for each endpoint if the use varies)
Major Deficiencies (note any stated exclusions): Check all that apply. Checking any of these items make the study "Not
Acceptable for Use"
Mixture (for controlled experiments only)
No Controls (for controlled experiments
only)
Excessive Control Mortality (> 10% for acute and > 20% for chronic)
. , . . . . Bioaccumulation: steady state not
Dilution water not adequately characterized , ,
reached
Dermal or Injection Exposure Pathway
Review paper or previously published without modification
Other: (if any, list here)
P-22
-------
POTENTIAL CHEMICAL MIXTURES : Describe any potential chemicals mixtures as characterized by study
authors (including any confirmation of chemical mixtures).
DESCRIPTION OF DILUTION WATER: Describe concerns with characterization of and/or major deficiencies
with dilution water.
General Notes:
Minor Deficiencies: List and describe any minor deficiencies or other concerns with test. These items may make the study
"Acceptablefor Qualitative Use" (exceptions may apply as noted)
For Field Studies/Observations: A field study/observation may be considered "Acceptable for Quantitative Use" if it
consisted of a range of exposure concentrations and the observed effects are justifiably contributed to a single chemical
exposure
Mixture (observed effects not justifiably contributed to single chemical exposure)
Uncharacterized Reference Sites/Conditions
POTENTIAL CHEMICAL MIXTURES PRESENT AT SITE: Describe any potential chemicals mixtures present at
the site as characterized by study authors (including any confirmation of chemicals present at study site).
EXPOSURE VARIABILITY ACROSS STUDY SITE(SV Describe any exposure variability across study site(s)
as characterized by study authors (i.e., description of study design with reference and contaminated sites).
General Notes:
Reviewer's Comments: Provide additional comments that do not appear under other sections of the template.
P-23
-------
ABSTRACT: Copy and paste abstract from publication.
SUMMARY: Fill out and modify as needed.
Acute:
Species (lil'cMiiuc)
Method'
1 CM
(liii'iilion
( hcmiciil
/ Piiriu
pll
Temp.
(ฐC)
Ihirdncss
(111 ii/1 :is
CaCO.o
or
S;ilini(\
(ppl)
DOC
(mji/l.)
r.iTcci
Reported
F.ITecl
( oiiccn 1 r;ition
(mji/l.)
Verified
l.lleel
( oiiccnlriilioii
(inii/l.)
CI;issifie;ilion
Qiiaiiuiau\ c Qualiiaii\ c
Unused
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
Chronic:
Species (lileslime)
Method'
Test
duriilion
( heniieiil
/ Pu ri( \
pll
Temp.
(ฐC)
Ihirdness
(lllfi/l. ilS
CaCO.o
or
S;ilini(\
(DDII
!)()(
(inii/l.)
( hronie
Limits
Reported
Chronic
Value
(mg/l. or
Verified
Chronic
Value
(inii/l. or
Chronic
Value
l.ndpoinl
Classificalion
Quantitative /
Qualitative / Unused
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
P-24
-------
II. Results Provide results as reported in the publication (including supplemental materials). Include screen shots of tables and/or
figures reporting results from the article following tabulated data table in each associated results section for all studies. Complete
tabulated data tables for all studies for studies marked "Acceptable for Quantitative Use" and "Acceptable for Qualitative Use".
Water Quality Parameters: If only general summary data of water quality parameters is provided by study authors (i.e., no
specific details of water quality parameters on a treatment level is provided), summarize any information regarding water quality
parameters under General Notes below and include data not provided in Table A.II.l.
General Notes: For aquatic life criteria development, measured water quality parameters in the treatments nearest the toxicity
test endpoint(s), e.g., LC50, EC20, etc., are most relevant.
Table A.II.1. Measured Water Quality Parameters in Test Solutions.
Dissolved oxygen, temperature, pH and [other parameters (hardness, salinity, DOC)] in test solutions during the /A'/-day
exposure of [test organism] to [concentration of treatments)] of [test substance] under [static renewal/flow-through]
conditions.
Parameter
Trealmenl
Mean
Range
Dissolved
[1]
oxygen
[2]
(% saturation
or mg/L)
j
j
[I]
Temperature
[2]
(Q
j
j
[1]
pH
[2]
j
j
Other (e.g.,
hardness,
salinity, DOC)
[I]
[2]
j
j
P-25
-------
Chemical Concentrations: Summarize the concentration verification data from test solutions/media. Expand table to include
each measured concentration data for each media type (i.e., muscle, liver, blood, etc.).
General Notes: Provide any necessary detail regarding the measured concentrations, including any identified cause for
substantial differences between nominal and measured concentrations, if samples were collected on separate days (and if so provide
details), and any potential cross contamination.
Table A.II.2. Measured (and Nominal) Chemical Concentrations in Test Solutions/Media.
[Analytical Method] verification of test and control concentrations during an [X]-day exposure of [test organism] to [test
substance] under [static renewal/flow-through] conditions.
| Mciin |
Nil in her of
ISliindiird
Nomiiiiil
Mc.isuml
S;i in pies
l)c\ hilion or
(oiiitii trillion
( <>iK'i'iilr;ilion
Nil in hoi' til'
Non-
Ik'low \on-
Siiindiird
1 iviilmonl
(iinils)
(iinils)
S;i in pies
IK'Icci'1
Ik'kcl
I'.rrorl
Kiinuo
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
aNon-Detect: 0 = measured and detected; l=measured and not detected; if not measured or reported enter as such
P-26
-------
Mortality: Briefly summarize mortality results (if any).
General Notes: Comment on concentrations response relations and slope of response ifprovided. Compare mortality with control
treatment and/or the reference chemical.
Table A.II.3. Mean Percent [Mortality or Survival].
Mean percent mortality [or number of immobilized] or survival of [test organism] exposed to [test substance] for [test
duration] under [static/renewal/flow-through] conditions.
|Mo;in
|M;ni(l;inl lk'\ iiilion
1 iviilmonl
Mnr(;ili(\ |
or Siiindiinl l.rmr|
Control
[1]
[2]
[3]
[4]
[5]
[6]
[LCX]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
P-27
-------
Growth: Briefly summarize growth results (if any).
General Notes: Comment on concentrations response relations and slope of response ifprovided. Compare growth endpoints with
control treatment and/or the reference chemical.
Table A.II.4. Mean [Growth].
Mean growth [length and/or weight] of [test organism] exposed to [test substance] for [test duration] under
[static/renewal/flow-through] conditions.
Menu (.row(h
Mcsiii IVrcenl
11 h/\\ oi vih 11
ISiiindiinl IK'\ hilinn
( hiiiiiic in | l.cn^lh/
ISiiindiird l)c\ hiiiun
1 iviilmonl
(ฆฆnils)
or Siiindiird l-'rrซir|
liioiiiiissj
or Siiindiird llrrorj
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
[ECX]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
P-28
-------
Reproductive: Briefly summarize reproduction endpoint results (if any). For multi-senerational studies, copy and paste Table
A.II.5 below for each generation with reproductive effects data.
General Notes: Comment on concentrations response relations and slope of response ifprovided. Compare reproduction
endpoints with control treatment and/or the reference chemical.
Table A.II.5. Mean [Reproductive] Effect.
Mean [reproductive] effects for [generation] of [test organism] exposed to [test substance] for [test duration] under
[static/renewal/flow-through] conditions.
ISliindiird
ISliindiird
| Moiin
|Sl;iil(l;inl
l)c\ in 1 ion
l)c\ iiilion
Number
lk'\ iillion oi'
IMciin
or
|Moiin
or
1 IVillllHMll
ol'
Siiindiinl
Number of
Siiindiinl
Number of
Sliindiird
(iinils)
S|ป;i\\ lis |
r.rmi'l
r.rrorl
OITsprinul
r.rrorl
Control
[11
[21
[31
[41
[51
[61
i
[ECxl
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
P-29
-------
Sublethal Toxicity Endpoints: Include other sublethal effect (s), including behavioral abnormalities or other signs of toxicity,
if any. Copy Table A.II.6 as needed to provide details for each sublethal effect observed.
General Notes: Briefly summarize observed sublethal effects otherwise not captured in the results table(s) below.
Table A.II.6. Mean [Sublethal] Effect.
Mean /"Sublethal effect, (e.g., behavioral abnormalities, etc.)] in [test organism] during [test duration (acute/chronic)]
exposure to [test substance] under [static/renewal/flow-through] conditions.
|Mciin Siihlolhiil
Rcsponsc'l
|S|;iikI;ii'(I l)c\ iiiiion or
1 IVillllHMII
( ii nils)
Siiindiird l.rmr|
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
[ECx]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control
P-30
-------
Reported Statistics: Copy and paste statistical section from publication.
P-31
-------
Part B: Detailed Review
I. Materials and Methods
PROTOCOL/GUIDANCE FOLLOWED: Indicate if provided by authors.
DEVIATIONS FROM PROTOCOL: If authors report any deviations from the protocol noted above indicate here.
Study Design and Methods: Copy and paste methods section from publication.
TEST ORGANISM: Provide information under Details and any relevant or related information or clarifications in Remarks.
Parameler
Delails
Remarks
Species:
Common Name:
Scientific Name:
North American species?
Surrogate for North American
Taxon?
(Place X if applicable)
Strain/Source:
Wild caught from unpolluted areas [1]
o Quarantine for at least 7 days or until they are
disease free, before acclimation [1]
Must originate from same source and population [1]
Should not be used:
o If appeared stressed, such as discoloration or
unusual behavior [1]
o If more than 5% die during the 48 hours before
test initiation [1]
o If they were used in previous test treatments or
controls [2]
No treatments of diseases may be administered:
o Within 16 hours of field collection [1]
o Within 10 days of testing or during testing 111
Age at Study Initiation:
Acute:
Larval stages preferred [1]
Mayflies and Stoneflies
o Early instar [1]
Daphnids/cladocerans:
o < 24-hr old [1]
Midges:
o 2ntl or 3ri instar larva [1]
Hyalella azteca (chronic exposure)
o Generally, 7-8 days old [3]
Freshwater mussels (chronic exposure)
o Generally, 2 month old juveniles [4]
Mysids (chronic exposure)
o < 24-hr old m
Was body weight or length recorded at
test initiation and/or at regular intervals?
Yes No
Was body weight or length recorded at
regular intervals?
Yes No
If yes, describe regular intervals:
P-32
-------
STUDY PARAMETERS: Provide information under Details and any relevant information of deficiencies in Remarks.
Complete for both Controlled Experiments and Field Studies/Observations.
Paramcler
Details
Remarks
Number of Replicates per Treatment
Group:
At least 2 replicates/treatment recommended for
acute tests [1]
At least 2 replicates/treatment recommended for
chronic tests 151
Control(s):
Treatment(s):
Number of Organisms per Replicate/
Treatment Group:
At least 10 organisms/treatment recommended.
Control(s):
Treatment(s):
5
Exposure Pathway:
(i.e., water, sediment, or diet). Note: all other pathways
(e.g., dermal, injection) are unacceptable.
Exposure Duration:
Acute
Cladocerans and midges should be 48 hours [2]
o Longer durations acceptable if test species not fed
and had acceptable controls [2]
Freshwater mussel glochidia should be a maximum
of 24 hours [4]
o Shorter durations (6, 12, 18 hours) acceptable so
long as 90% survival of control animals achieved
(see below) [4]
Embryo/larva (bivalve mollusks, sea urchins,
lobsters, crabs, shrimp and abalones) should be 96
hours, but at least 48 hours [2]
Other invertebrate species should be 96 hours
Acute
Chronic
Other (please remark):
ฆi
Chronic
Daphnids/cladocerans should be 21 days (3-brood
test) [2]
o Exception 7 days acceptable for Ceriodaphnia
dubia [2]
Freshwater juvenile mussels should be at least 28
days [4]
Hyalella azteca should be at least 42 days
o Beginning with 7-8 day old animals [3]
Mysids should continue until 7 days past the median
time of first brood release in the controls 141
Test Concentrations (remember units):
Nominal:
Recommended test concentrations include at least three
concentrations other than the control; four or more will
provide a better statistical analysis.
Measured:
Media measured in:
Observation Intervals:
Should be an appropriate number of observations
over the study to ensure water quality is being
properly maintained 111
P-33
-------
CONTROLLED EXPERIMENT STUDY PARAMETERS: Provide information under Details and any relevant
information of deficiencies in Remarks. Complete for Controlled Experiments only.
Paramcler
Delails
Remarks
Acclimaliou/lloldiug:
Should be placed in a tank along with the water in
Duration:
unuhsis tifony):
which they were transported [1]
o Water should be changed gradually to 100%
dilution water (usually 2 or more days) [1]
o For wild-caught animals, test water temperature
should be within 5ฐC of collection water
Feeding:
Water:
temperature [1]
o Temperature change rate should not exceed 3ฐC
Temperature (ฐC):
within 72 hours [1]
To avoid unnecessary stress and promote good
health:
o Organisms should not be crowded [1]
o Water temperature variation should be limited
o Dissolved oxygen:
ฆ Maintain between 60 - 100% saturation [1]
ฆ Continuous gentle aeration if needed [1]
o Unionized ammonia concentration in holding and
acclimation waters should be < 35 ng/L [11
Dissolved Oxygen (mg/L):
Health {any mortality observed?):
-
Acclimation followed published guidance?
Describe, if any
Yes No
If yes, indicate which guidance:
Test Vessel:
Test chambers should be loosely covered [1]
Test chamber material:
Material:
Briefly describe the test vessel here
o Should minimize sorption of test chemical from
water [1]
o Should not contain substances that can be leached
Size:
ฃ
or dissolved in solution and free of substances that
could react with exposure chemical [1]
o Glass, No. 316 stainless steel, nylon screen and
Fill Volume:
perfluorocarbon (e.g. Teflon) are acceptable [1]
o Rubber, copper, brass, galvanized metal, epoxy
glues, lead and flexible tubing should not come
into contact with test solution, dilution water or
stock [1]
Size/volume should maintain acceptable biomass
loading rates (see below) [1]
Substrate:
o Required for some species (e.g., Hyalella azteca)
[3]
o Common types: stainless steel screen, nylon
screen, quartz sand, cotton gauze and maple leaves
[3]
o More inert substances preferred over plant
material, since plants may break down during
testing and promote bacterial growth [3]
o Consideration should be given between substrate
and toxicant [3]
ฆ Hydrophobic organic compounds in particular
can bind strongly to Nitexฎ screen, reducing
exposure concentrations, especially for studies
using static or intermittent renewal exposure
methods [31
Substrate Used (if applicable)-.
P-34
-------
Paramclcr
Delails
Rem arks
Test Solution Delivery System/Method:
Flow-through preferred for some highly volatile,
hydrolyzable or degradable materials [2]
o Concentrations should be measured often enough
using acceptable analytical methods [2]
Chronic exposures:
o Flow-through, measured tests required [2]
o Exception: renewal is acceptable for daphnids [2]
Test Concentrations Measured
Yes No
Test Solution Delivery System:
Static
Renewal
Indicate Interval:
Flow-through
Indicate Type of Diluter:
Source of Dilution Water:
Freshwater hardness range should be < 5 mg/L or <
10% of the average (whichever is greater) [1]
Saltwater salinity range should be < 2 g/kg or < 20%
of the average (whichever is greater) [1]
Dilution water must be characterized (natural surface
water, well water, etc.) [2]
o Distilled/deionized water without the addition of
appropriate salts should not be used [2]
Dilution water in which total organic carbon or
particulate matter exceed 5 mg/L should not be used
o Unless data show that organic carbon or particulate
matter do not affect toxicity [2]
Dilution water for tests with Hyalella azteca
o Reconstituted waters should have at least 0.02 mg
bromide/L; natural ground or surface water
presumed to have sufficient bromide [3]
o Recommended that control/dilution waters have
chloride concentrations at or above 15 mg/L [31
Dilution Series (e.g., 0.5x, 0.6x, etc.):
Dilution Water Parameters:
Measured at the beginning of the experiment or
averaged over the duration of the experiment (details of
water quality parameters measured in test solutions
should be included under the results section)
Dissolved Oxygen (mg/L):
pH:
Temperature (ฐC):
Hardness (mg/L as CaCCh):
Salinity (ppt):
Total Organic Carbon (mg/L):
Dissolved Organic Carbon (mg/L):
Aeration:
Acceptable to maintain dissolved oxygen at 60 -
100% saturation at all times [1]
Avoid aeration when testing highly oxidizable,
reducible and volatile materials
Turbulence should be minimized to prevent stress on
test organisms and/or re-suspend fecal matter [1]
Aeration should be the same in all test chambers at all
times [11
Yes No
Describe Preparation of Test
Concentrations (e.g., water exposure,
diet):
P-35
-------
Paramcler
Details
Remarks
Test Chemical Solubility in Water:
List units and conditions (e.g., 0.01% at 20ฐC)
Were concentrations in water or diet
verified by chemical analysis?
Measured test concentrations should be reported in
Table A.II.2 above.
Yes No
Indicate media:
Were test concentrations verified by
chemical analysis in tissue?
Measured test concentrations can be verified in test
organism tissue (e.g., blood, liver, muscle) alone if a
dose-response relationship is observed.
Measured test concentrations should be reported in
Table A.II.2 above.
Yes No
If test concentrations were verified in test organism
tissue, was a dose-response relationship observed?
Indicate tissue type:
Were stability and homogeneity of test
material in water/diet determined?
Yes No
Was test material regurgitated/avoided?
Yes No
Solvent/Vehicle Type:
When used, a carrier solvent should be kept to a
minimum concentration [1]
Should not affect either survival or growth of test
organisms [1]
Should be reagent grade or better [ 1 ]
Should not exceed 0.5 ml/L (static), or 0.1 ml/L (flow
through) unless it was shown that higher
concentrations do not affect toxicity [51
Negative Control:
Yes No
Reference Toxicant Testing:
Yes No
If yes, identify substance:
Other Control: If any (e.g. solvent control)
Biomass Loading Rate:
Loading should be limited so as not to affect test
results. Loading will vary depending on temperature,
type of test (static vs. flow-through), species,
food/feeding regime, chamber size, test solution
volume, etc. [1]
This maximum number would have to be determined
for the species, test duration, temperature, flow rate,
test solution volume, chamber size, food, feeding
regime, etc.
Loading should be sufficiently low to ensure:
o Dissolved oxygen is at least 60% of saturation
(40% for warm-water species) [1,6]
o Unionized ammonia does not exceed 35 ng/L [1]
o Uptake by test organisms does not lower test
material concentration by > 20% [1]
o Growth of organisms is not reduced by crowding
Generally, at the end of the test, the loading (grams of
organisms; wet weight; blotted dry) in each test
chamber should not exceed the following:
o Static tests: >0.8 g/L (lower temperatures); >0.5
g/L (higher temperatures) [1]
o Flow through tests: > 1 g/L/day or > 10 g/L at any
time (lower temperatures); >0.5 g/L/day or > 5
g/L at any time (higher temperatures) [1]
o Lower temperatures are defined as the lower of
17ฐC or the optimal test temperature for that
species. [11
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Feeding:
Unacceptable for acute tests [2]
o Exceptions:
ฆ Data indicate that the food did not affect the
toxicity of the test material [2]
Yes No
c.
ฆ Test organisms will be severely stressed if they
are unfed for 96 hours [2]
.ง
ฆ Test material is very soluble and does not sorb
>
or complex readily (e.g., ammonia) [21
iฃr"
Lighting:
No specific requirements for lighting
Generally, ambient laboratory levels (50 - 100 fc) or
ฃ
natural lighting should be acceptable, as well as a
diurnal cycle consisting of 50% daylight or other
natural seasonal diurnal cycle
Artificial light cycles should have a 15 - 30 minute
-C.
transition period to avoid stress due to rapid increases
in light intensity [1]
Depends on the type of test (acute or chronic) and
endpoint (e.g., reproduction) of interest.
Study Design/Methods Classification: (Place Xby One Based on Overall Study Design/Methods Classification)
Provide details of Major or Minor Deficiencies/Concerns with Study Design in Associated Sections of Part A: Overview
This classification should be taken into consideration for the overall study classification for aquatic life criteria development in Part A.
Study Design Acceptable for Quantitative Use
Study Design Acceptable for Qualitative Use
Study Design Not Acceptable for Use
Additional Notes: Provide additional considerations for the classification of study use based on the study design.
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OBSERVATIONS: Provide information under Details and any relevant information in Remarks. This information should be
consistent with the Results Section in Part A.
Parameler
Details
Remarks
Parameters measured including sublethal
effects/toxicity symptoms:
Common Apical Parameters Include:
Acute
Daphnids/cladocerans:
o EC50 based on percentage of organisms
immobilized plus percentage of organisms killed
[2]
Embryo/larva (bivalve molluscs, sea urchins, lobsters,
crabs, shrimp, and abalones):
0 EC50 based on the percentage of organisms with
incompletely developed shells plus the percentage
of organisms killed [2]
ฆ If not available, the lower of the 96 hour EC50
based on the percentage of organisms with
incompletely developed shells and the 96-hr
LC50 should be used [2]
Freshwater mussel (glochidia and juveniles):
0 Glochidia: EC50 based on 100 x number closed
glochidia after adding NaCl solution - number
closed glochidia before adding NaCl solution) /
Total number open and closed glochidia after
adding NaCl solution [4]
0 Juvenile: EC50 based on percentage exhibiting foot
movement within a 5-min observation period [4]
All other species and older life stages:
0 EC50 based on the percentage of organisms
exhibiting loss of equilibrium plus the percentage
of organisms immobilized plus the percentage of
organisms killed [2]
ฆ If not available, the 96 hour LC50 should be
used [2]
Chronic
Daphnid:
0 Survival and young per female [2]
Mysids:
0 Survival, growth and young per female 121
List parameters:
Was control survival acceptable?
Acute
> 90% control survival at test termination [2]
0 Glochidia 90% after 24 hours, or, the next longest
duration less than 24 hours that had at least 90%
survival [4]
Chronic
> 80% control survival at test termination [2]
0 80% in 42 day test with Hyalella azteca, slightly
lower in tests substantially longer than 42 days [31
Yes No
Control survival (%):
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Paramcler
Delails
Remarks
Were individuals excluded from the
analysis?
Yes No
If yes, describe justification provided:
Was water quality in test chambers
acceptable?
If appropriate, describe any water quality issues
(e.g., dissolved oxygen level below 60% of
saturation)
Yes No
Availability of concentration-response
data:
Were treatment level concentration-response
data included in study publication (can be from
tables, graphs, or supplemental materials)?
specify endpoints in remarks
Yes No
Were replicate level concentration-response
data included in study publication (can be from
tables, graphs, or supplemental materials)?
specify endpoints in remarks
Yes No
If treatment and/or replicate level
concentration-response data were included, how
was data presented? (check all that apply)
Tables
Graphs
Supplemental Files
Were concentration-response data estimated
from graphs study publication or supplemental
materials?
Yes No
If yes, indicate software used:
Yes No
Should additional concentration-response data be
requested from study authors?
If concentration-response data are available, complete
Verification of Statistical Results (Part C) for sensitive
species.
Requested by:
Request date:
Date additional data received:
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Part C: Statistical Verification of Results
I. Statistical Verification Information: Report the statistical methods (e.g., EPA TRAP, BMDS, R, other) used to verify the
reported study or test results for the five (5) most sensitive genera and sensitive apical endpoints (including for tests where such
estimates were not provided). If values for the LC50, LT50 and NOEC are greater than the highest test concentration, use the "> "
symbol.
Primary Reviewer: Date: EPA Contractor {Place X by One)
Secondary Reviewer: Date: EPA Contractor {Place X by One)
{At least one reviewer should be from EPA for sensitive taxa)
Endpoint(s) Verified:
Additional Calculated Endpoint(s):
Statistical Method (e.g., TRAP, BMDS, R, other):
II. Toxicity Values: Include confidence intervals if applicable
NOEC:
LOEC:
MATC:
ECs:
EC10:
EC20:
ECso or LCso
Dose-Response Curve Classification: (PlaceXby One)
This classification should be taken into consideration for the overall study classification for aquatic life criteria development in Part A
Dose-Response Curve Acceptable for Quantitative Use
Dose-Response Curve Acceptable for Qualitative Use
Dose-Response Curve Not Acceptable for Use
Summary of Statistical Verification: Provide summary of methods used in statistical verification.
Additional Notes:
Attachments:
1. Provide attachments to ensure all data used in Part C is captured, whether from study results reported in the publication
and/or from additional data requested from study authors
Data from study results of the publication should be reported in Results section of Part A
Additional data provided upon request from study authors should be reported in Table C.II.l below and original
correspondence with study authors should be included as attachments
2. Model assessment output (including all model figures, tables, and fit metrics)
3. Statistical code used for curve fitting
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-------
III. Attachments: Include all attachments listed above after the table below.
Additional Data Used in Response-Curve: Provide all data used to fit dose-response curve not captured in Results section of PER above in Part A, rows as needed. First
row in italicized text is an example.
Table C.II.1 Additional Data Used in Dose-Response Curve.
('ur\c II)
I'lidpiiinl
liviilnunl
Ki-|>lii;ik-
|Sl;i iul;i I'd
l)i-\ iiiiiim
fir
S|;iihI;ii-(I
Kitoi'I
#of
Sun ixiirs
V
k1
11'
kl'spilllsi-
Kl'SpilllM-
l nil
ClIIH'
Chih' uiiils
Alchronicl
Ceriodaphnia dubia
#of
young/female
0
6
10
10
1
18
count
0.03
mg/L
aN = number of individuals per treatment; k = number of replicates per treatment level; n = number of individuals per replicate
P-41
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Part I): References to Test Guidance
6. ASTM Standard E 739, 1980. 2002. Standard guide for conducting acute toxicity tests on
test materials with fishes, macroinvertebrates, and amphibians. ASTM International,
West Conshohocken, PA.
7. Stephan, C.E., D.I. Mount, D.J. Hansen, J.H. Gentile, G.A. Chapman and W.A. Brungs.
1985. Guidelines for Deriving Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and their Uses. PB85-227049. National Technical
Information Service, Springfield, VA.
8. Mount, D.R. and J.R. Hockett. 2015. Issue summary regarding test conditions and
methods for water only toxicity testing with Hyalella azteca. Memorandum to Kathryn
Gallagher, U.S. EPA Office of Water. U.S. EPA Office of Research and Development.
MED. Duluth, MN. 9 pp.
9. Bringolf, R.B., M.C. Barnhart and W.G. Cope. 2013. Determining the appropriate
duration of toxicity tests with glochidia of native freshwater mussels. Submitted to
Edward Hammer. U.S. EPA. Chicago, IL, May 8, 2013. 39 pp.
10. Stephan, C.E. 1995. Review of results of toxicity tests with aquatic organisms. Draft.
U.S. EPA, MED. Duluth, MN. 13 pp.
11. American Public Health Association (APHA). 2012. Standard methods for the
examination of water and wastewater. Part 8000 - Toxicity. APHA. Washington, DC.
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