United States Office of Water EPA-842-R-24-003
Environmental Protection 4304T September 2024
Agency
FINAL
FRESHWATER AQUATIC LIFE AMBIENT WATER
QUALITY CRITERIA AND ACUTE SALTWATER
BENCHMARK FOR
PERFLUOROOCTANE SULFONATE (PFOS)
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:
Amanda Jarvis, Office of Water, Office of Science and Technology, Health and Ecological
Criteria Division, Washington, DC
James R. Justice, 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
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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 also 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 viii
List of Appendices xi
Acronyms xii
Notices xv
Foreword xvi
Executive Summary xviii
1 INTRODUCTION AM) BACKGROUND 1
1.1 Previously Derived PFOS Toxicity Values and Thresholds 2
1.1.1 Previously Published Acute Water Protective Values for Direct Aqueous
Exposure 3
1.1.2 Previously Published Chronic Water Protective Values for Direct Aqueous
Exposure 3
1.1.3 Previously Published Chronic Fish Tissue Criteria 4
1.2 Overview of Per- and Polyfluorinated Substances (PFAS) 10
1.2.1 Physical and Chemical Properties of PFOS 14
2 PROBLEM FORMULATION 18
2.1 Overview of PFOS Sources 18
2.1.1 Manufacturing of PFOS 18
2.1.2 Sources of PFOS to Aquatic Environments 21
2.2 Environmental Fate and Transport of PFOS in the Aquatic Environment 23
2.2.1 Environmental Fate of PFOS in the Aquatic Environment 23
2.2.2 Environmental Transport of PFOS in the Aquatic Environment 24
2.3 Transformation and Degradation of PFOS Precursors in the Aquatic Environment.... 26
2.3.1 Degradation of perfluoroalkane sulfonamido derivatives 27
2.3.2 Perfluorooctane sulfonamide-based side-chained polymers 30
2.3.3 Fluoroalkyl surfactants used in AFFFs 30
2.4 Environmental Monitoring of PFOS in Abiotic Media 31
2.4.1 PFOS Occurrence and Detection in Ambient Surface Waters 31
2.5 Bioaccumulation and Biomagnification of PFOS in Aquatic Ecosystems 36
2.5.1 PFOS Bioaccumulation in Aquatic Life 37
2.5.2 Factors Influencing PFOS Bioaccumulation and Biomagnification in Aquatic
Ecosystems 38
2.5.3 Environmental Monitoring of PFOS in Biotic Media 40
2.6 Exposure Pathways of PFOS in Aquatic Environments 45
2.7 Effects of PFOS on Biota 46
2.7.1 Mode of Action and Toxicity of PFOS 47
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2,7,2 Potential for Interactions with Other PFAS 49
2.8 Conceptual Model of PFOS in the Aquatic Environment and Effects 51
2.9 Assessment Endpoints 53
2.10 Measurement Endpoints 54
2.10.1 Overview of Toxicity Data Requirements 54
2.10.2 Measure of PFOS Exposure Concentrations 55
2.10.3 Measures of Effect 61
2.11 Analysis Plan 64
2.11.1 Derivation of Water Column Criteria 64
2.11.2 Consideration for the Derivation of Tissue-Based Criteria following Chronic
PFOS Exposures 65
2.11.3 Translation of Chronic Water Column Criterion to Tissue Criteria 65
3 EFFECTS ANALYSIS FOR AQUATIC LIFE 68
3.1 Toxicity to Aquatic Life 68
3.1.1 Summary ofPFOS Toxicity Studies Used to Derive the Aquatic Life Criteria 68
3.2 Derivation of the PFOS Aquatic Life Criteria 96
3.2.1 Derivation of Water Column Criteria for Direct Aqueous Exposure 96
3.2.2 Derivation of Freshwater Chronic Tissue criteria for PFOS 104
3.2.3 Translation of Chronic Water Column Criterion to Tissue Criteria 104
3.3 Summary of the PFOS Freshwater Aquatic Life Criteria and Acute Estuarine/Marine
Benchmark Ill
4 EFFECTS CHARACTERIZATION FOR AQUATIC LIFE 114
4.1 Additional Analyses Supporting the Derivation of the Chronic Water Column
Criterion for Freshwater 114
4.2 Influence of Using Non-North American Resident Species on PFOS Criteria 122
4.2.1 Freshwater Acute Water Column Criterion with Native and Established
Organisms (Species Not Resident to North America removed from dataset) 122
4.2.2 Freshwater Chronic Water Criterion with Native and Established Organisms
(Species Not Resident to North America removed from dataset) 125
4.3 Qualitatively Acceptable Water Column-Based Toxicity Data 127
4.3.1 Consideration of Qualitatively-Acceptable Acute Data 128
4.3.2 Consideration of Qualitatively-Acceptable Chronic Data 130
4.4 Acute-to-Chronic Ratios 134
4.5 Comparison of Empirical Tissue Concentrations to Translated Tissue Criteria 135
4.5.1 Comparison of Quantitative Studies and Tissue-Based Criteria 138
4.5.2 Comparison of Qualitative Studies and Tissue-Based Criteria 142
4.6 Effects on Aquatic Plants 144
4.7 Protection of Threatened and Endangered Species 145
4.7.1 Quantitatively Acceptable Acute Toxicity Data for Listed Species 145
4.7.2 Quantitatively Acceptable Chronic Toxicity Data for Listed Species 146
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4,7,3 Qualitatively Acceptable Toxicity Data for Listed Species 146
4.8 Summary of the PFOS Aquatic Life Criterion and the Supporting Information 147
5 REFERENCES 148
List of Tables
Table Ex-1. Recommended Perfluorooctane Sulfonate (PFOS) Ambient Water Quality
Criteria for the Protection of Aquatic Life in Freshwaters xx
Table Ex-2. Acute Perfluorooctane Sulfonate (PFOS) Benchmark for the Protection of
Aquatic Life in Estuarine/Marine Waters xx
Table 1-1. Previously Derived PFOS Toxicity Values and Thresholds 5
Table 1-2. Two Primary Categories of PFAS1 11
Table 1-3. Classification and Chemical Structure of Perfluoroalkyl Acids (PFAAs).1 13
Table 1-4. Chemical and Physical Properties of PFOS 15
Table 2-1. Summary of Assessment Endpoints and Measures of Effect Used in the Criteria
Derivation for PFOS 63
Table 2-2. Evaluation Criteria for Screening Bioaccumulation Factors (BAFs) in the Public
Literature 67
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 PFOS 69
Table 3-2. The Four Most Sensitive Genera Used in Calculating the Acute Freshwater
Criterion (Sensitivity Rank 1-4) 71
Table 3-3. Ranked Freshwater Genus Mean Acute Values 75
Table 3-4. The Four Most Sensitive Acute Estuarine/Marine Genera 78
Table 3-5. Ranked Estuarine/Marine Water Genus Mean Acute Values 81
Table 3-6. The Four Most Sensitive Genera Used in Calculating the Chronic Freshwater
Criterion 83
Table 3-7. Ranked Freshwater Genus Mean Chronic Values 90
Table 3-8. The Four Ranked Estuarine/Marine Genus Mean Chronic Values 93
Table 3-9. Freshwater Final Acute Value and Criterion Maximum Concentration 97
Table 3-10. Freshwater Final Chronic Value and Criterion Continuous Concentration 100
Table 3-11. Summary Statistics for PFOS BAFs in Fish and Invertebrates1 105
Table 3-12. Recommended Perfluorooctane Sulfonate (PFOS) Ambient Water Quality
Criteria for the Protection of Aquatic Life in Freshwaters 112
Table 4-1. Additional Analyses Supporting the Derivation of the Freshwater Chronic Water
Column Criterion 116
Table 4-2. GMCVs Used in Derivation of Chronic Criterion and Additional Analyses
Supporting the Chronic Criterion for Freshwater 120
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Table 4-3. Ranked Freshwater Genus Mean Acute Values with Native and Established
Organisms, excluding Species Not Resident to North America 123
Table 4-4. Calculation of Freshwater Acute Water Column Concentration with Native and
Established Organisms (Species Not Resident to North America Removed from
Dataset) 124
Table 4-5. Ranked Freshwater Genus Mean Chronic Values with Native and Established
Organisms 126
Table 4-6. Calculation of Freshwater Chronic Water Column Concentration with Native and
Established Organisms 127
Table 4-7. Comparison of Empirical Tissue Concentrations to Chronic Tissue Criteria and
Additional Tissue Values 136
Table L-l. Surrogate Species Measured Values for PFOS and Corresponding Number of
ICE Models for Each Surrogate L-8
Table L-2. Comparison of ICE-predicted and measured values of PFOS for species using
both scaled values (entered as mg/L) and values potentially beyond the model
domain (entered as (J,g/L) L-l 1
Table L-3. All ICE Models Available in web-ICE v3.3 for Saltwater Predicted Species
Based on Surrogates with Measured PFOS L-l5
Table L-4. ICE-Estimated Species Sensitivity to PFOS L-17
Table L-5. Ranked Estuarine/Marine Genus Mean Acute Values L-20
Table L-6. Estuarine/Marine Final Acute Value and Protective Aquatic Acute Benchmark. ...L-21
Table N-l. Global Sediment Concentration of PFOS N-18
Table P-l. Characteristics of adult fish sampled for the calculation of PFOS reproductive
tissue BAFs P-2
Table P-2. Summary Statistics for PFOS BAFs in Additional Fish Tissues1 P-3
Table P-3. PFOS Concentrations for Additional Fish Tissue.1'2 P-4
List of Figures
Figure 1-1. Chemical Structure of Linear Perfluorooctane Sulfonate (PFOS) 14
Figure 2-1. Synthesis of PFOS by electrochemical fluorination (ECF) 19
Figure 2-2. Aerobic Biodegradation of EtFOSE in Activated Sludge 29
Figure 2-3. Map Indicating Sampling Locations for Perfluorooctane Sulfonate (PFOS)
Measured in Surface Waters across the United States (U.S.) 32
Figure 2-4. Distribution of the Minimum and Maximum Concentrations (ng/L) of
Perfluorooctane Sulfonate Measured in Surface Waters for Each State or
Waterbody (excluding the Great Lakes) with Reported Data in the Publicly
Available Literature 34
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Figure 2-5. Conceptual Model Diagram of Sources, Compartmental Partitioning, and Trophic
Transfer Pathways of Perfluorooctane Sulfonate (PFOS) in the Aquatic Environment
and its Bioaccumulation and Effects in Aquatic Life 52
Figure 3-1. Freshwater Acute PFOS GMAVs Fulfilling the Acute MDRs 77
Figure 3-2. Acceptable Estuarine/Marine GMAVs 82
Figure 3-3. Ranked Freshwater Chronic PFOS Used Quantitatively to Derive the Criterion 92
Figure 3-4. Acceptable Estuarine/Marine GMCVs 96
Figure 3-5. Ranked Freshwater Acute PFOS GMAVs Used Quantitatively to Derive the
Criterion 98
Figure 3-6. Ranked Freshwater Chronic PFOS GMCVs Used Quantitatively to Derive the
Criterion 100
Figure L-l. Example ICE Model for Rainbow Trout (surrogate) and Atlantic Salmon
(predicted) L-4
Figure L-2. Ranked Estuarine/Marine Acute PFOS GMAVs used for the Aquatic Life
Acute Benchmark Calculation L-21
Figure L-3. Americamysis bahia (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-25
Figure L-4. Americamysis bahia (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-25
Figure L-5. Americamysis bahia (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values L-26
Figure L-6. Danio rerio -embryo (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-26
Figure L-7. Danio rerio - embryo (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-27
Figure L-8. Danio rerio - embryo (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values L-27
Figure L-9. Daphnia magna (X-axis) and Americamysis bahia (Y-axis) regression model
used for ICE predicted values L-28
Figure L-10. Daphnia magna (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values L-28
Figure L-l 1. Daphnia magna (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-29
Figure L-12. Daphnia magna (X-axis) and Oncorhynchus mykiss (Y-axis) regression model
used for ICE predicted values L-29
Figure L-13. Daphnia magna (X-axis) and Pimephales promelas (Y-axis) regression model
used for ICE predicted values L-30
Figure L-14. Lampsilis siliquoidea (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-30
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Figure L-15. Lampsilis siliquoidea (X-axis) and Ligumia recta (Y-axis) regression model
used for ICE predicted values L-31
Figure L-16. Lampsilis siliquoidea (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-31
Figure L-17. Lampsilis siliquoidea (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values L-32
Figure L-18. Ligumia recta (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values L-32
Figure L-19. Lithobates catesbeianus (X-axis) and Daphnia magna (Y-axis) regression
model used for ICE predicted values L-33
Figure L-20. Lithobates catesbeianus (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-33
Figure L-21. Lithobates catesbeianus (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values L-34
Figure L-22. Oncorhynchus mykiss (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values L-34
Figure L-23. Oncorhynchus mykiss (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-3 5
Figure L-24. Oncorhynchus mykiss (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values L-3 5
Figure L-25. Oncorhynchus mykiss (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-36
Figure L-26. Oncorhynchus mykiss (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values L-36
Figure L-27. Pimephales promelas (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values L-37
Figure L-28. Pimephales promelas (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values L-37
Figure L-29. Pimephales promelas (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values L-3 8
Figure L-30. Pimephales promelas (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values L-3 8
Figure L-31. Pimephales promelas (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values L-39
Figure L-32. Pimephales promelas (X-axis) and Xenopus laevis (Y-axis) regression model
used for ICE predicted values L-39
Figure L-33. Xenopus laevis (X-axis) and Pimephales promelas (Y-axis) regression model
used for ICE predicted values L-40
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List of Appendices
Appendix A Acceptable Freshwater Acute PFOS Toxicity Studies A-l
Appendix B Acceptable Estuarine/Marine Acute PFOS Toxicity Studies B-l
Appendix C Acceptable Freshwater Chronic PFOS Toxicity Studies C-l
Appendix D Acceptable Estuarine/Marine Chronic PFOS Toxicity Studies D-l
Appendix E Acceptable Freshwater Plant PFOS Toxicity Studies E-l
Appendix F Acceptable Estuarine/Marine Plant PFOS Toxicity Studies F-l
Appendix G Other Freshwater PFOS Toxicity Studies G-l
Appendix H Other Estuarine/Marine PFOS Toxicity Studies H-l
Appendix I Acute to Chronic Ratios I-1
Appendix J Unused PFOS 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 PFOS Benchmarks for Estuarine/Marine
Waters through a New Approach Method (NAM): WeblCE L-l
Appendix M Environmental Fate of PFOS in the Aquatic Environment M-l
Appendix N Occurrence of PFOS in Abiotic Media N-l
Appendix O Bioaccumulation Factors (BAFs) Used to Calculate PFOS Tissue Values 0-1
Appendix P Translation of Chronic Water Column Criterion into Other Fish Tissue Types
(liver, blood, reproductive tissues) P-l
Appendix Q Example Data Evaluation Records (DERs) Q-l
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Acronyms
6:2 Cl-PFESA 6:2 chlorinated polyfluorinated ether sulfonate
ACR Acute-to-Chronic Ratio
AFFF Aqueous film-forming foams
AIC Akaike information criteria
AMV Acute Maximum Value
ASW artificial sea water
AWQC National Recommended Ambient Water Quality Criteria
BAF Bioaccumulation factor
C8-PFPA Perfluorooctyl phosphonic acid
C8/C8-PFPiA Bis(perfluorooctyl) phosphinic acid
CAS/CASRN Chemical Abstracts Service Registry Numbers
CC Chronic Criterion
CCC Criterion Continuous Concentration
C-F carbon-fluorine
CMC Criterion Maximum Concentration
C-R concentration-response
C-S carbon-sulfur
CWA Clean Water Act
DER Data Evaluation Record
DMSO dimethyl sulfoxide
dpf days post fertilization
drc dose-response curve
dw dry weight
ECF Electrochemical fluorination
ECOTOX ECOTOXicology database
ELS Early life-stage
EPA U.S. Environmental Protection Agency
EtFASAAs A'-ethyl perfluoroalkane sulfonamidoacetic acids
EtFASAs A'-ethyl perfluoroalkane sulfonamides
EtFOSAA TV-ethyl perfluorooctane sulfonamidoacetic acid
EtFOSE TV-ethyl perfluorooctane sulfonamidoethanol
FACR Final Acute-to-Chronic Ratio
FASAAs Perfluoroalkyl sulfonamidoacetic acids
FASAs Perfluoroalkane sulfonamids
FASEs perfluoroalkyl sulfonamidoethanols
FAV Final Acute Value
FCV Final Chronic Value
FFTG Canadian Federal Fish Tissue Guideline
FIFRA Federal Insecticide, Fungicide, and Rodenticide Act
FOSA Perfluorooctane sulfonamide
FWQG Federal Water Quality Guideline
GLI U.S. EPA Great Lakes Initiative
GMAV Genus Mean Acute Value
GMCV Genus Mean Chronic Value
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HC1
1% Hazardous Concentration
GSD
genus sensitivity distribution
hpf
hours post fertilization
ICE
Interspecies Correlation Estimation
Kow
n-octanol-water partition co-efficient
LOD
limit of detection
LOEC
Lowest Observed Effect Concentration
LOQ
limit of quantification
MATC
Maximum Acceptable Toxicant Concentration
MC
Maximum Criterion
MDL
Method Detection Limit or Minimum Detection Limit
MDRs
minimum data requirements
NAMs
New Approach Methods
NCCA
National Coastal Condition Assessment
NOEC
No Observed Effect Concentration
NPDES
National Pollutant Discharge Elimination System
NRSA
National Rivers and Streams Assessment
OCSPP
Office of Chemical Safety and Pollution Prevention
OECD
Organization for Economic Co-operation and Development
ORD
Office of Research and Development
OSF
Octane sulfonyl fluoride
OW
Office of Water
PFAAs
Perfluoroalkyl acids
PFAS
Per- and polyfluorinated substances
PFCA
Perfluoroalkyl carboxylic acids or Perfluoroalkyl carboxylates
PFDA
Perfluorodecanoate or Perfluorodecanoic acid
PFdiCAs
Perfluoroalkyl dicarboxylic acids
PFdiSAs
Perfluoroalkane disulfonic acids
PFECAs
Perfluoroalkylether carboxylic acids
PFESAs
Perfluoroalkylether sulfonic acids
PFDoA
Perfluorododecanoate or Perfluorododecanoic acid
PFOA
Perfluorooctanoic acid or Perfluorooctanoate
PFOS
Perfluorooctane sulfonate or Perfluorooctane sulfonate acid
PFOSI
Perfluorooctane sulfinic acid
PFOS-K
PFOS potassium salt
PFOS-Li
PFOS lithium salt
PFPAs
Perfluoroalkyl phosphonic acids
PFPiAs
Perfluoroalkyl phosphinic acids
PFSAs
Perfluoroalkane sulfonic acids or Perfluoroalkyl sulfonates
PFSiAs
FASA iV-glucuronides or Perfluoroalkyl sulfinic acids
pKa
Acid dissociation constant
POSF
Perfluorooctanesulfonyl fluoride
PPAR-a
Nuclear peroxisome proliferator activated receptor-alpha
ppt
parts per thousand
SMACR
Species Mean Acute-to-Chronic Ratio
SMAV
Species Mean Acute Value
Xlll
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SMCV
Species Mean Chronic Value
SNUR
Significant New Use Rules
SOP
Standard Operating Procedure
SSD
Species Sensitivity Distribution
TMDLs
Total Maximum Daily Loads
TSCA
Toxic Substances Control Act
U.S.
United States
UCMR
Unregulated Contaminant Monitoring Rule
web-ICE
Web-based Interspecies Correlation Estimation
WQS
Water Quality Standards
WW
wet weight
WWTPs
Wastewater treatment plants
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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 Perfluorooctane sulfonate (PFOS). 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 PFOS that protect aquatic life, the PFOS 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.eov/wqc/aqiiatic4ife-criteria-perfliiorooctane-siilfonate-pfos.
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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-reducing-use-vertebrate-animals-chemicaO.
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
xvii
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Executive Summary
The U.S. Environmental Protection Agency (EPA) developed the recommended
perfluorooctane sulfonate (PFOS) 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 PFOS ambient water
quality criteria recommendations to protect aquatic life. The EPA has derived the recommended
PFOS 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" (i.e., EPA's 1985 Guidelines; U.S. EPA 1985) and EPA's
OCSPP's Ecological Effects Test Guidelines (U.S. EPA 2016b).
PFOS is an organic, human-made perfluorinated compound, consisting of an eight-carbon
backbone and a sulfonate functional group. PFOS (and other related chemicals that are
perfluoroalkane sulfonic acids) is used in a variety of industrial and commercial products,
including surface treatments of soil, surface treatments of textiles, paper, and metals, and in
specialized applications such as in firefighting foams. This document provides a critical review
of all aquatic toxicity data identified in the EPA's literature search for PFOS, including the
anionic form (CAS No. 45298-90-6), the acid form (CAS No. 1763-23-1), potassium salt (CAS
No. 2795-39-3), an ammonium salt (CAS No. 56773-42-3), sodium salt (CAS No. 4021-47-0),
and a lithium salt (CAS No. 29457-72-5). It also quantifies the toxicity of PFOS to aquatic life
and provides criteria to protect aquatic life in freshwater from the acute and chronic toxic effects
of PFOS.
The Aquatic Life Ambient Water Quality Criteria for PFOS document includes water
column-based acute and a water column-based chronic criteria, as well as chronic tissue-based
-------
criteria for freshwaters. Quantitatively-acceptable estuarine/marine toxicity data only fulfilled
five of the eight minimum data requirements (MDRs) for deriving acute estuarine/marine criteria
and four of the eight MDRs for deriving chronic estuarine/marine criteria per the 1985
Guidelines. The EPA did, however, include an acute aquatic life benchmark for 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/) (Raimondo et al. 2010). 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 PFOS acute toxicity data (Appendix L).
The freshwater acute water column-based criterion magnitude is 0.071 mg/L, and the
chronic water column-based criterion magnitude is 0.00025 mg/L (250 ng/L). The final chronic
freshwater criterion also contains tissue-based criteria with magnitudes of 0.201 mg/kg wet
weight (ww) for fish whole-body, 0.087 mg/kg ww for fish muscle tissue, and 0.028 mg/kg ww
for invertebrate whole-body tissue. All criteria are intended to be equally protective against
adverse PFOS 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 (BAFs) 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 are expected to protect the freshwater aquatic
community.
xix
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Table Ex-1. Recommended Perfluorooctane Sulfonate (PFOS) Ambient Water Quality
Criteria for the Protection of Aquatic Life in Freshwaters. i
Chronic
Acute Waid-
Chronic Waid-
Clironic
l-ish
Chronic
Co! ii m n
Co! u mil
Inverlebrale
Whole-
l-ish
Type/Media
(CMC)14
(CCCy5
Whole-Bodv'~
Body1-2
Muscle12
Magnitude
0.071 mg/L
0.00025 mg/L
0.028
mg/kg ww
0.201
mg/kg ww
0.087
mg/kg ww
Duration
One-hour
average
Four-day
average
Instantaneous3
Not to be
Not to be
exceeded more
exceeded more
Frequency
than once in
than once in
Not to be exceeded6
three years on
three 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 are derived from the chronic water-column criterion magnitude (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 PFOS 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 PFOS 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 Perfluorooctane Sulfonate (PFOS) Benchmark for the Protection of
Aquatic Life in Estuarine/Marine Waters.
Type/Media
Aculc Wnlcr Column Benchmark
Magnitude
0.55 mg/L
Duration
One hour on average
Frequency
Not to be exceeded more than once in three years on average
xx
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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 before adopting these into standards. After adoption,
states/authorized Tribes submit new and revised WQS to EPA for review and approval or
disapproval. When approved by EPA, the state's/Tribes WQS become the application 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 Eliminations System (NPDES) programs. 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 the EPA 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 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 assessment provides a critical review of all aquatic toxicity data identified in the
EPA's literature search for PFOS, including the anionic form (CAS No. 45298-90-6), the acid
1
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form (CAS No. 1763-23-1), a potassium salt (CAS No. 2795-39-3), an ammonium salt (CAS No.
56773-42-3), a sodium salt (CAS No. 4021-47-0), and a lithium salt (CAS No. 29457-72-5). It
quantifies the toxicity of PFOS to aquatic life and provides criteria to protect aquatic life in
freshwater from the acute and chronic toxic effects of PFOS.
The EPA derived the recommended criteria using the best available data to reflect the
latest scientific knowledge on the toxicological effects of PFOS to 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 PFOS freshwater criteria, if adopted and implemented, are
expected to be protective of most aquatic organisms, including species listed as threated and
endangered, in the community and are derived to be protective of aquatic life designated uses
established by states and Tribes for freshwaters. The estuarine/marine benchmarks are also
intended to be protective of aquatic life designated uses, but as they are based on fewer empirical
PFOS data have greater inherent uncertainty. The criteria recommendations presented herein are
the EPA's best estimate of the concentrations of PFOS, with associated frequency and duration
specifications, that would protect sensitive aquatic life from unacceptable acute and chronic
effects.
1.1 Previously Derived PFOS Toxicity Values and Thresholds
Within the U.S., no states or Tribes have CWA Section 303(c) approved water quality
standards for the protection of aquatic life from the exposure to PFOS. However, several states
have published draft/interim acute and chronic ecological screening level values/benchmarks for
the protection of aquatic life. As such, previously published PFOS acute and chronic criteria,
benchmarks, and thresholds developed by states and international regulatory authorities were
2
-------
identified, that included values for both freshwater and marine systems, and are summarized
below.
1.1.1 Previously Published Acute Water Protective Values for Direct Aqueous Exposure
Previously published freshwater acute values were available for four states (Florida,
Michigan, Minnesota, and Texas) and one geographic region (Europe). These publicly available
values for other jurisdictions ranged from 0.021 mg/L in Texas (Giesy et al. 2010; TCEQ 2021)
to 0.78 mg/L in Michigan (EGLE 2010). The EPA's freshwater acute PFOS criterion (0.071
mg/L) falls into the middle of the range of state- and European-derived values.
There were two previously derived estuarine/marine acute values, with a
benchmark/criterion of 0.0072 mg/L in Europe (RIVM 2010) and 0.21 mg/L in Florida (Stuchal
and Roberts 2019) (Table 1-1). These values were derived using safety factors. Consequently,
these state values were both lower than the EPA's PFOS acute estuarine/marine benchmark (0.55
mg/L), which used measured and estimated acute toxicity data (Appendix L).
1.1.2 Previously Published Chronic Water Protective Values for Direct Aqueous Exposure
Previously published freshwater chronic values were available for five states (California,
Florida, Michigan, Minnesota, and Texas) and three countries or geographic regions
(Australia/New Zealand, Canada, and Europe). The publicly available state-derived values
ranged from 0.00056 mg/L in California (RWQCB 2020; SERDP 2019; 99% species protection)
to 0.14 mg/L in Michigan (EGLE 2010). Overall, the EPA's chronic water column-based PFOS
criterion (0.00025 mg/L) is lower than chronic state-derived values because the EPA's chronic
PFOS criterion was based on recently published and sensitive insect data.
Internationally, chronic PFOS protective values were 0.000023 mg/L in Europe (RIVM
2010), 0.00013 mg/L in Australia/New Zealand (CRCCare 2017; EPAV 2017) 95% species
protection level), and 0.00680 mg/L in Canada (ECCC 2018) (Table 1-1). The EPA's chronic
3
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water-column based PFOS criterion (0.00025 mg/L) is in the middle of the range of chronic
PFOS values used in different international jurisdictions.
Previously published estuarine/marine chronic values were available for three states
(California, Florida, and Texas) and two geographic regions (Australia/New Zealand and
Europe). These publicly available values for other jurisdictions ranged from 0.000294 mg/L for
Texas (CRCCare 2017; TCEQ 2021) to 0.013 mg/L in Florida (Stuchal and Roberts 2019) and
were 0.0000046 mg/L in Europe (RIVM 2010) and 0.00013 mg/L in Australia/New Zealand
(95% species protection; CRCCare 2017; EPAV 2017). The EPA did not derive a chronic PFOS
criterion or benchmark for estuarine/marine water because of data limitations.
1.1.3 Previously Published Chronic Fish Tissue Criteria
There was a single previously derived fish tissue value for other jurisdictions. This value
was a Canadian Federal Fish Tissue Guideline (FFTG) of 9.4 mg/kg whole-body wet weight
(ww) (ECCC 2018). This value was derived by multiplying Canada's Federal Water Quality
Guideline of 6.8 [j,g/L by a BAF of 1,378 L/kg. Canada's fish whole-body based Federal Water
Quality Guideline (9.4 mg/kg ww) is significantly larger than the EPA's PFOS fish whole-body
tissue criterion (0.201 mg/kg ww) because the EPA's value considered more recently published
and relatively sensitive toxicity data that were not available at the time Canada's fish whole-body
Water Quality Guideline was derived.
4
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Table 1-1. Previously Derived PFOS Toxicity Values and Thresholds.
State /
Country of
Applicability
Aquatic l.il'e Protective Value
unless otherwise iiulicitled)
Criteria or Benchmark and Calculation Approach
Source
Freshwater Acute
Some
European
Countries
0.036
Maximum Acceptable Concentration calculated using the lowest acute
(LC50) value of 3.6 mg/L for mysid (Americamysis bahia) by
assessment factor of 100. Dataset includes freshwater and marine aquatic
species, combined.
RIVM
(2010)
Texas
0.021
Acute surface water benchmark calculated using U.S. EPA Great Lakes
Initiative (GLI; (U.S. EPA 1995)) Tier I Methodology as reported in
(Giesy et al. 2010). This is an acute surface water benchmark and does
not represent a CWA Section 303(c) approved water quality standard for
PFOS.
Giesy et al.
(2010);
TCEQ
(2021)
Minnesota
0.085
Maximum Criterion (MC) calculated as the acute curve-fitted and
extrapolated 10-d EC50 for midge (Chironomus tentans) of 170 |ig/L,
which serves as the Final Acute Value or FAV followed by ^ 2. This
draft value does not represent a CWA Section 303(c) approved water
quality standard for PFOS.
STS/MPCA
(2007)
Florida
0.53
Secondary Acute Value (SAV) calculated using U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995) Tier II Methodology. FAV calculated as
the lowest GMAV (unspecified) divided by a safety factor of 6.1. 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)
5
-------
State /
Country (ป!'
Applicability
Aquatic Life Protective Value
unless otherwise inilicntecl)
Criteria or Benchmark and Calculation Approach
Source
Michigan
0.78
FAV of 1,557 |ig/L calculated as the lowest Genus Mean Acute Value
(GMAV) of 9,500 |ig/L for fathead minnow (Pimephales promelas) by
a safety factor of 6.1 (following U.S. EPA Great Lakes Initiative [(GLI;
U.S. EPA 1995)]. The Acute Maximum Value (AMV) of 0.78 mg/L was
then calculated as the FAV ^ 2. 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 PFOS.
EGLE (2010)
Marine Acute
Some
European
Countries
0.0072
Maximum Acceptable Concentration calculated using the lowest acute
value (LCso) of 3.6 mg/L for a mysid (Americamysis bahia) by an
assessment factor of 500. Dataset includes freshwater and marine aquatic
species, combined.
RIVM
(2010)
Florida
0.21
Secondary Acute Value (SAV) calculated using U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995) Tier II Methodology. FAV calculated as
the lowest GMAV (unspecified) divided by a safety factor of 21.9. 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)
Freshwater Chronic
Some
European
Countries
0.000023
Maximum Permissible Concentration calculated using the lowest value
(LOEC) of 0.0023 mg/L for Chironomus tentans (MacDonald et al.
2004) by an assessment factor (100). Dataset includes freshwater and
marine aquatic species, combined.
RIVM
(2010)
6
-------
Stsite /
Country (ป!'
Applicability
Aqusitic l.il'e Protective Vsiluc
(inซ/1. unless otherwise iiuliciited)
('rilcrisi or licnchmsirk ;iihI Csilciihition Approach
Source
Canada
0.00680
Federal Water Quality Guideline (FWQG) calculated as the fifth
percentile value from a Species Sensitivity Distribution (SSD) consisting
of 20 species-specific values representing fish (5), amphibians (2),
invertebrates (5), and plants and algae (8).
ECCC
(2018)
Australia,
New Zealand
0.00000023
(99% species protection - high
conservation value systems)
0.00013
(95% species protection - slightly to
moderately disturbed systems)
0.002
(90% species protection - highly
disturbed systems)
0.031
(80% species protection - highly
disturbed systems)
Guidelines calculated from Species Sensitivity Distribution (SSD)
consisting of 18 species-specific values for fish, amphibians, insects,
crustaceans, and algae following the guidance of Warne et al. (2018) and
Batley et al. (2014)
CRCCare
(2017);
EPAV
(2017);
HEPA
(2020)
California
0.00056
(99% species protection)
HCi calculated from an acute and chronic NOEC-based SSD as reported
in SERDP Project ER18-1614 (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 PFOS.
RWQCB
(2020);
SERDP
(2019)
Texas
0.0051
Acute surface water benchmark calculated using U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995) Tier I Methodology as reported in Giesy
et al. (2010). This is a chronic surface water benchmark and does not
represent a CWA Section 303(c) approved water quality standard for
PFOS.
Giesy et al.
(2010);
TCEQ
(2021)
7
-------
State /
Country (ป!'
Applicability
Aquatic Life Protective Value
unless otherwise iiuliciiled)
Criteria or Benchmark and Calculation Approach
Source
Minnesota
0.019
Chronic Criterion (CC) calculated as the FAV (170 |ig/L) ^ FACR (9.12)
per Minnesota Rules Chapter 7050. Two species-specific ACRs and a
default ACR were used to calculate the FACR. This draft value does not
represent a CWA Section 303(c) approved water quality standard for
PFOS.
STS/MPCA
(2007)
Florida
0.037
Secondary Chronic Value (SCV) calculated using U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995) Tier II Methodology with acute-to-
chronic (ACR) of 14.5. SCV = SAV (530 |ig/L) - ACR (14.5) = 37
|ig/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)
Michigan
0.14
Final Chronic Value (FCV) calculated as the FAV (1,557 |ig/L) FACR
(11.35) per U.S. EPA Great Lakes Initiative (GLI; U.S. EPA 1995). Two
species-specific ACRs and a default ACR were used to calculate the
FACR. 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 PFOS.
EGLE (2010)
Marine Chronic
Australia,
New Zealand
0.00000023
(99% species protection - high
conservation value systems)
Guidelines calculated from SSD following the guidance of Warne et al.
(2018) and Batley et al. (2014) and consisting of nine species-specific
values representing fish (2), echinoderms (2), crustaceans (2), mollusc
(1), and algae (2).
Note: Per HEPA (2020) 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
CRCCare
(2017);
EPAV
(2017);
HEPA
(2020)
0.00013
(95% species protection - slightly to
moderately disturbed systems)
0.002
(90% species protection - highly
disturbed systems)
0.031
(80% species protection - highly
disturbed systems)
8
-------
Stsite /
Country (ป!'
Applicability
Aqusitic l.ilc Protective Value
(inซ/1. unless otherwise iiuliciiled)
('rilcrisi or licnchmsirk ;iihI Calculation Approach
Source
Some
European
Countries
0.0000046
Maximum Permissible Concentration calculated using the lowest value
(LOEC) of 0.0023 mg/L for Chironomus tentans divided by an
assessment factor (500). Dataset includes freshwater and marine aquatic
species, combined.
RIVM
(2010)
Texas
0.000294
Default guidelines calculated from SSD following the guidance of Warne
et al. (2018) and Batley et al. (2014) and consisting of nine of 16 species-
specific values as reported in CRCCare (2017). This is a chronic surface
water benchmark and does not represent a CWA Section 303(c)
approved water quality standard for PFOS.
CRCCare
(2017);
TCEQ
(2021)
California
0.0026
(99% species protection)
HCi calculated from an acute and chronic NOEC-based SSD as reported
in SERDP Project ER18-1614 (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 PFOS.
RWQCB
(2020);
SERDP
(2019)
Florida
0.013
Secondary Chronic Value (SCV) calculated using U.S. EPA Great Lakes
Initiative (GLI; U.S. EPA 1995) Tier II Methodology with acute-to-
chronic (ACR) of 15.6. SCV = SAV (210 |ig/L) - ACR (15.6) = 13
|ig/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)
Fish Tissue
Canada
9.4 mg/kg whole body ww fish
tissue
Federal Fish Tissue Guideline (FFTG) where FFTG of 9.4 mg/kg ww =
(FWQG of 6.8 (J,g/L) * (BAFgeomean of 1378 L/kg)
ECCC
(2018)
9
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1.2 Overview of Per- and Polvfluorinated Substances (PFAS)
PFOS, 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 2021a).
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 practical guidance regarding the terminology of PFAS (U.S.
EPA 2021a). 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:
10
-------
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).
Table 1-2. Two Primary Categories of PFAS1.
PI-"AS Non-polymers
Structural Klemenls
Kxample PI-AS lamilies
Compounds in which all carbon-
, hydrogen bonds, except those on
Pertluoroalkyl acids c ฐ . , , ,
J 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
Compounds in which all carbon-
, , . , hydrogen bonds on at least one
Polyfluoroalkyl acids , n A A , ,
J carbon (but not all) are replaced
with carbon-fluorine bonds
polyfluoroalkyl carboxylic acids,
polyfluoroalkylether carboxylic
and sulfonic acids
PI AS Polvmers
Structural Klemenls
Kxample PI'AS l-'amilies
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).
PFOS belongs to the perfluoroalkyl acids (PFAAs) of the non-polymer perfluoroalkyl
substances category of PFAS. PFAAs are among the most researched PFAS (Wang et al. 2017).
11
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The PFAA family 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). The
protonated and anionic forms may have different physiochemical properties.
12
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Table 1-3. Classification and Chemical Structure of Perfluoroalkyl Acids (PFAAs).1
Classification
I'linclional Croup
r.xamplcs
Perfluoroalkyl carboxylic
acids (PFCAs)
Or
Perfluoroalkyl carboxylates
(PFCAs)
-COOH
Perfluorooctanoic acid (PFOA)
-COO"
Perfluorooctanoate (PFOA)
Perfluoroalkane sulfonic acids
(PFSAs)
Or
Perfluorokane sulfonates
(PFSAs)
-so3h
Perfluorooctane sulfonic acid (PFOS)
-so3-
Perfluorooctane sulfonate (PFOS)2
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)
( F<(O( F2)n( OOH
Periluoro (3,5,7-trioxaoctanoic) acid
Perfluoroalkylether sulfonic
acids (PFESAs)
CF3(0CF2)S03H
6:2 chlorinated polyfluorinated ether
sulfonate (6:2 Cl-PFESA)
Perfluoroalkyl dicarboxylic
acids (PFdiCAs)
hooc-cf2-cooh
9:3 Fluorotelomer betaine
Perfluoroalkane di sulfonic
acids (PFdiSAs)
ho3s-cf2-so3h
Perfluoro-l,4-disulfonic acid
1 Modified from Buck et al. (2011); OECD (2021).
2 The anionic form is most prevalent in the aquatic environment.
Perfluoroalkane (or -alkyl) sulfonic acids (PFSAs), including PFOS, consist of a general
chemical structure (of CnF2n+iS03H for PFOS; see Figure 1-1). This chemical structure makes
PFOS extremely strong and stable, and resistant to hydrolysis, photolysis, microbial degradation,
and metabolism (see Section 2.3) (Ahrens 2011; Beach et al. 2006; Buck et al. 2011).
Furthermore, PFOS has been classified as persistent, bioaccumulative, and toxic (Ahrens 2011;
Buck et al. 2011; Lindstrom et al. 2011; OECD 2002).
13
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o
Figure 1-1. Chemical Structure of Linear Perfluorooctane Sulfonate (PFOS).
Source: United States EPA Chemistry Dashboard; https://comptox.epa.gov/dashboard
1.2.1 Physical and Chemical Properties of PFOS
Physical and chemical properties along with other reference information for PFOS are
provided in Table 1-4. These physical and chemical properties help to define the environmental
fate and transport of PFOS in the aquatic environment.
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Table 1-4. Chemical and Physical Properties of PFOS.
Property
PI'"OS. acidic form1
Source
Chemical Abstracts Service
Registry Number (CAS No.)
1763-23-1
Chemical Abstracts Index
Name
1,1,2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-
heptadecafluoro-1 -
octanesulfonic acid
Synonyms
Perfluorooctane sulfonic acid;
heptadecafluoro-1 -octane
sulfonic acid; PFOS acid;
perfluorooctane sulfonate
Chemical Formula
C8HF17O3S
Molecular Weight (grams
per mole [g/moll)
500.13
Lewis (ed. 2004); HSDB
(2012); SRC (2016)
Color/Physical State
White powder (potassium salt)
OECD (2002)
Boiling Point
258-260 ฐC
SRC (2016)
Melting Point
No data
Vapor Pressure
2.0 x 10"3 millimeters Mercury
(mm Hg) at 25ฐC (estimate)
HSDB (2012)
Henry's Law Constant
Not measurable; not expected
to volatilize from aqueous
solution (< 2.0 x 10"6)
AT SDR (2015)
Kow
Not measurable
EFSA (2008); AT SDR
(2015)
Organic carbon water
partitioning coefficient (Koc)
2.57
Higgins and Luthy (2006)
Estimated pKa
3.27 (no empirical
measurements available)
Brooke et al. (2004)
Solubility in Water
680 mg/L
OECD (2002)
Half-Life in Water
Stable
UNEP (2006)
Half-Life in Air
Stable
UNEP (2006)
1 PFOS is commonly produced as a potassium salt (CAS No. 2795-39-3). Properties specific to the salt are not
included.
PFOS is moderately water soluble, nonvolatile, and stable (Beach et al. 2006; Young and
Mabury 2010). PFOS is solid at room temperature with a low vapor pressure. No direct
measurement of the acid dissociation constant (pKa) is available. However, PFOS is considered
to have a low pKa, which is based on a calculated pKa of 3.27 provided from Finland in a
comment to Brooke et al. (2004). Therefore, PFOS is deemed to be a strong acid (Brooke et al.
2004). PFOS introduced as a salt will dissociate into ionic components when in natural water at a
15
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neutral pH and is commonly present as a PFOS anion in solution (Beach et al. 2006; Giesy et al.
2010; Young and Mabury 2010). The PFOS anion forms strong ion pairs with many cations,
resulting in less solubility in waters that contain great amounts of dissolved solids. Thus, PFOS
solubility in saltwater is approximately 12 mg PFOS/L compared to 589 mg PFOS/L in pure
water (Beach et al. 2006). PFOS is reported to have a mean solubility of 56 mg PFOS/L in pure
octanol (OECD 2002). These solubility data suggest that any form of PFOS discharged into a
water source tends to remain dissolved, unless the PFOS was sorbed to particulate matter or
assimilated by organisms (which are both discussed further in Sections 2.2 and 2.5, respectively)
(OECD 2002).
Due to the surfactant properties of PFOS, it forms three layers when added to octanol and
water in a standard test system used to measure an n-octanol-water partition co-efficient (Kow),
thus preventing direct measurement (Giesy et al. 2010; OECD 2002). Although a Kow cannot be
directly measured, a Kowfor PFOS has been estimated from its individual water and octanol
solubilities (Giesy et al. 2010); however, the veracity of such estimates is uncertain (OECD
2002). Lacking a reliable Kow for PFOS 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
PFOS would bring into question the use of Kow as a predictor of environmental behavior. For
example, bioaccumulation of PFOS is thought to be mediated via binding to proteins rather than
partitioning into lipids (Giesy et al. 2010; OECD 2002), the latter being the theoretical basis for
Kow-based prediction of bioaccumulation.
PFOS is not expected to volatilize from aqueous solution based on its vapor pressure and
predicted Henry's law constant < 2.0 x 10"6 (Beach et al. 2006). In 2002, OECD classified PFOS
16
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as a type 2, non-volatile chemical that has a very low or possibly negligible volatility (Beach et
al. 2006; Giesy et al. 2010; OECD 2002).
17
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2 PROBLEM FORMULATION
A problem formulation provides a strategic framework for water quality criterion
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 is consistent with the EPA's Guidelines for Ecological Risk Assessment
(U.S. EPA 1998).
2.1 Overview of PFOS Sources
2.1.1 Manufacturing of PFOS
PFOS is used in a variety of products including surface treatments for soil and stain
resistance, coating of paper as part of a sizing agent formulation, and in specialized applications
such as firefighting foams. PFOS is produced through electrochemical fluorination (ECF) in
which an organic raw material, such as octane sulfonyl fluoride (OSF; C8H17SO2F) in the case of
PFOS, undergoes electrolysis in anhydrous hydrogen fluoride solution. This electrolysis leads to
the replacement of all the hydrogen atoms by fluorine atoms and results in
perfluorooctanesulfonyl fluoride (POSF; C8F17SO2F), which is the major raw material used to
manufacture PFOS (Figure 2-1; Buck et al. 2011). The base-catalyzed hydrolysis of POSF
results in PFOS and its salts (Lehmler 2005). ECF results in a mixture of linear and branched
chain perfluorinated isomers and homologues, with ratios of linear to branched perfluorinated
carbon chains of roughly 70 to 80% linear and 20 to 30% branched for PFOS synthesis
depending on how the process is controlled (De Voogt 2010). All compounds produced from
POSF and other neutral PFAS with sufficient chain length and a sulfur group have the potential
to degrade or transform into PFOS, and therefore have been considered to be "PFOS
equivalents" and potential sources of PFOS to the aquatic environment (see Section 2.4)(Ahrens
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2011; Lindstrom et al. 2011). PFOS is used in a variety of products including surface treatments
for soil and stain resistance, coating of paper as part of a sizing agent formulation, and in
specialized applications such as firefighting foams.
Figure 2-1. Synthesis of PFOS by electrochemical fluorination (ECF).
Modified from Buck et al. (2011).
The manufacture of PFOS started in 1949 with Minnesota Mining and Manufacturing
(name changed later to the 3M Company) (3M Company 1999). Prior to 2000, the 3M Company
was the major producer of POSF, the raw material used to make PFOS (Figure 2-1), with smaller
producers in Europe and Asia (Paul et al. 2009; U.S. EPA 2000a). In 2000, the 3M company
manufactured approximately 78% of the estimated global POSF production (approximately
19
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3,665 tons of the 4,650 tons produced globally; (approximately 3,665 tons of the 4,650 tons
produced globally; OECD 2002). The estimated total cumulative production of POSF is between
44,000 and 96,000 tons (Paul et al. 2009; Prevedouros et al. 2006; Smithwick et al. 2006).
Information on previous and current production of POSF from Asia and other production sources
is limited (Paul et al. 2009; Prevedouros et al. 2006; Smithwick et al. 2006).
In May 2000, following negotiations between the EPA and 3M, the 3M Company agreed
to voluntary phase out and find substitutes for PFOS chemistry used to produce all but a few
small applications (i.e., aqueous film-forming foams (AFFF), and hard chrome plating mist
suppression) across their range of products by 2002 (Lindstrom et al. 2011; U.S. EPA 2000a).
Starting around the same time, a series of Significant New Use Rules (SNUR) were also put into
place by the EPA to restrict the production and use of materials that contain PFOS and its
precursors in the U.S. (Lindstrom et al. 2011). In 2009, PFOS and related compounds were listed
under Annex B of the Stockholm Convention on Persistent Organic Pollutants; restricting global
manufacturing and use of PFOS (Ahrens 2011; OECD 2002). Homologues, neutral precursor
compounds, and new classes of PFAS continue to be produced and therefore, are potential
sources of PFOS (Ahrens 2011). Assuming there was no step-up production of PFOS and its
precursors to offset the phase-out by the 3M Company, the production is estimated to be
approximately 1,000 tons from 2002 and onward (Paul et al. 2009). However, while
industrialized countries, like the U.S., phased-out the use of PFOS and its precursors, producers
in other countries, such as China and Brazil, have scaled up their production to fill remaining
demand (Wang et al. 2013). Despite the wide use in an array of industrial and consumer products
globally, information on the sources, volumes, and emission of PFOS and its precursors are
limited (Paul et al. 2009; Zhang et al. 2016).
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2,1,2 Sources of PFOS to Aquatic Environments
Aquatic environments and soil are thought to serve as a reservoir of PFOS, with 42,000
tons emitted to aquatic environments compared to 235 tons released into to air between 1980 and
2002 (Paul et al. 2009). Unlike other contaminants commonly found in aquatic ecosystems, such
as metals for example, PFAS are synthetic compounds with no natural source. Thus, the
occurrence of any PFAS in the environment is an indication of anthropogenic sources (Ahrens
2011). The occurrence of PFOS in aquatic environments can be attributed to both point and non-
point sources, entering aquatic environments from industrial and consumer products during
manufacturing, along supply chains, and during product use and/or disposal (Ahrens 2011;
Ahrens and Bundschuh 2014; Kannan 2011; Paul et al. 2009). However, quantitative
assessments of PFOS production, point and non-point source discharges, and environmental
measurements are limited compared to other persistent, bioaccumulative pollutants (Ahrens and
Bundschuh 2014; Zhang et al. 2016).
Potential point sources of PFOS to the aquatic environment include both industrial
facilities and municipal wastewater treatment plants (WWTPs). Additional point sources may
include surface water runoff from industrial use sites such as metal plating facilities, areas that
have received AFFF applications, landfills, and contaminated soils. Of these, industrial facilities,
specifically those for fluorochemical manufacturing and other use facilities, are a primary source
of PFOS to aquatic systems (Ahrens et al. 201 la; Houtz et al. 2016; Sedlak et al. 2017).
Estimated total global releases to water arising from discharge of PFOS during manufacturing
from 1970 to 2002 ranged between 230 and 1,450 tons (Paul et al. 2009).
Potential non-point PFOS sources to aquatic environments include: dry and wet
atmospheric deposition, runoff from contaminated soils, runoff from metal plating facilities, the
runoff or discharge of contaminated groundwater, particularly from the use of fire-fighting
21
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foams, and land application of contaminated biosolids (Ahrens 2011; Kannan 2011; OECD
2002; Paul et al. 2009). Identification of non-point PFOS sources and understanding their
relative contribution to aquatic ecosystems is difficult due to the lack of sufficient measured
environmental data (Ahrens 2011; Paul et al. 2009). Overall, the presence of non-point PFOS
sources and their relative contributions are reported to be dependent on the aquatic system, air,
groundwater, and soil levels, and nearby land uses. For example, concentrations of PFAS,
including PFOS, have been influenced by urban land use (Ahrens 2011; Zhang et al. 2016).
Overall, PFOS occurrence in aquatic environments is driven by legacy PFOS sources since
PFOS use in the United States was voluntarily phased out by 2002 and SNUR were put into
place by the EPA to restrict the production and use of PFOS and its precursors (Lindstrom et al.
2011). However, PFAS concentrations in the environment in general continue to be positively
correlated with human population density. For example, PFOS was detected in aquatic systems
at elevated concentrations (ranging between 97 and 1,371 ng/L) in densely populated areas of the
U.S. and Europe (Zhang et al. (2016) and Loos et al. (2009); respectively), and Paul et al. (2009)
estimated the total global PFOS emissions to air and water from 1970 to 2009 resulting from
consumer use and disposal to be between 420 and 2,100 tons.
Importantly, PFAS are still produced that can transform or degrade into compounds
belonging to the PFSAs family, including PFOS (Ahrens 2011). PFAS precursors such as
perfluoroalkyl sulfonamidoacetic acids (FASAAs) and perfluoroalkyl sulfonamidoethanols
(FASEs) are known to metabolically transform and degrade to PFOS, respectively (Ahrens and
Bundschuh 2014; Benskin et al. 2009; Boulanger et al. 2005b; Buck et al. 2011; Lange 2000; Liu
and Mejia Avendano 2013; Plumlee et al. 2008; Rhoads et al. 2008; Wang et al. 2017). However,
the understanding of these transformation processes is limited, and additional work is needed to
22
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fully understand these processes and their role as sources of PFOS to aquatic environments
(Buck et al. 2011; Lau et al. 2007; Liu and Mejia Avendano 2013; Wang et al. 2017).
Degradation of precursors represents a potentially significant source of PFOS to the aquatic
environment, particularly since PFOS production within the U.S. has not occurred since 2002
(Buck et al. 2011; Liu and Mejia Avendano 2013). Nevertheless, PFOS-treated articles, such as
fabrics, paper, and other treated materials, are still being imported into the U.S. and are
ultimately, at least in part, released into the environment (Allred et al. 2015; Lang et al. 2016;
Liu et al. 2014d). The importation of PFOS treated articles is considered as production under the
Toxic Substances Control Act (TSCA) (U.S. EPA 2020).
2.2 Environmental Fate and Transport of PFOS in the Aquatic Environment
2.2,1 Environmental Fate of PFOS in the Aquatic Environment
PFOS has low volatility in ionized form but can adsorb to particles in air where it can be
transported globally, including remote locations (Benskin et al. 2012; Butt et al. 2010). PFOS is
water soluble and has been found in surface water, ground water, and drinking water. Because of
the relatively low Koc of PFOS, it does not easily adsorb to sediments and tends to stay in the
water column (Ahrens 2011; Beach et al. 2006; Giesy et al. 2010; Higgins and Luthy 2006).
PFOS can be re-emitted to aquatic environments from PFOS contaminated soil,
groundwater, ice, and sediment (see Section 2.3). Sediment may be an important sink of PFOS in
the aquatic environment (Ahrens 2011). The movement of PFOS between groundwater, surface
water, and sediment depends on the chemical properties of PFOS and site-specific
physiochemical characteristics (including pH, temperature, organic carbon content, and salinity)
of the aquatic environment. In general, PFOS may sorb to sediments (with a Kd greater than 1
mL/g; (with a Kd greater than 1 mL/g; Giesy et al. 2010). However, this sorption to sediments is
limited and PFOS has a Koc of 2.57 indicating that PFOS is relatively mobile in water and the
23
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physicochemical characteristics of the sediment ultimately influence the sorption of PFOS
(Ahrens 2011; Higgins and Luthy 2006). While the release of PFOS from the transformation of
other PFAS and historical products still in use (e.g., consumer goods manufactured, imported
and/or obtained before the PFOS discontinuation and regulations) are expected to continue into
the future, the re-emissions of PFOS from existing sinks are assumed to be decreasing since the
restrictions and regulations of PFOS have gone into place (Ahrens 2011; Ahrens and Bundschuh
2014; Paul et al. 2009; Washington and Jenkins 2015; Washington et al. 2015).
In the water column, and other environmental compartments, PFOS is stable and resistant
to hydrolysis, photolysis, volatilization, and biodegradation (see Appendix M)(Beach et al. 2006;
OECD 2002). The persistence of PFOS 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 PFOS are known. The physiochemical properties discussed in Table 1-4 result in
PFOS being highly persistent in the aquatic environment (Ahrens 2011). In aquatic
environments, the only dissipation mechanisms for PFOS are physical mechanisms, such as
environmental dilution, offsite transport, plant uptake, and sorption.
2.2.2 Environmental Transport of PFOS in the Aquatic Environment
The environmental fate of PFOS, outlined in the previous section (Section 2.2.1) plays a
role in the environmental transport of PFOS (Ahrens 2011). PFOS is either distributed in biota
(via bioaccumulation discussed in Section 2.5) or abiotic matrices (such as water and sediment).
Sediment in particular can act as a sink for PFOS. However, the role of sediment as a sink or
source by resuspension is not well understood (Ahrens 2011).
The distribution of PFOS is widespread, including to remote regions despite the limited
number of manufacturing facilities and/or small population sizes typically found in these areas
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(Benskin et al. 2012; Butt et al. 2010). PFOS has been detected in water, sediment, and biota
samples from aquatic environments in remote areas (Butt et al. 2010; Giesy and Kannan 2001;
Houde et al. 2006b; Yamashita et al. 2008). To date, the dominant transport pathway for PFOS to
remote regions has not been conclusively characterized and much of the focus has been on
marine systems, with few studies in freshwater environments (Ahrens 2011; Butt et al. 2010;
Giesy and Kannan 2002). Additionally, the relative importance of each potential transport
pathway is difficult to accurately determine (Butt et al. 2010; Young and Mabury 2010). Many
researchers suggest that the dominant mechanism of PFOS transport occurs through water as the
anionic form of PFOS, which is the most commonly found form in the aquatic environment, is
less volatile (see Section 2.2.1 above) and has a high water solubility. These characteristics make
partitioning to and transport through the air less likely (Butt et al. 2010; Giesy and Kannan
2002). However, PFOS transport through water is likely the dominant mechanism on more local
scales (e.g., within a waterbody or watershed), and is likely not the prevailing transport pathway
of PFOS to remote regions given the considerations of the long distances. Instead, atmospheric
transport is likely the main mechanism of PFOS transport to remote regions. Another potential
source to remote regions is the indirect formation of PFOS through transformation of other
PFAS, particularly volatile precursors (see Section 2.3)(Butt et al. 2010; Wang et al. 2015;
Young and Mabury 2010).
Volatile PFOS precursors, which may reach remote locations via atmospheric deposition
themselves, may subsequently be metabolized to PFOS in aquatic organisms (Giesy and Kannan
2002). In all likelihood, the continued presence of PFOS in remote areas may be due to multiple
exposure pathways, including those caused by direct production and use of PFOS itself as well as
degradation and transformation of precursor compounds (Armitage et al. 2009). To better
25
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comprehend both environmental transport and exposure to PFOS, the following needs to be
better understood: 1) the potential transformation, metabolism, and bioaccumulation of PFOS
and its precursors (particularly partitioning behavior, such as tissue distribution and
lipophilicity); 2) explicit biotransformation pathways and pharmacokinetics; and 3) atmospheric
fate and transport of PFOS and its precursors (Armitage et al. 2009).
2.3 Transformation and Degradation of PFOS Precursors in the Aquatic
Environment
Transformation and degradation processes of various PFAS are potential sources of
PFOS to the aquatic environment (see Section 2.1.2 above). PFAS are still produced that can
transform or degrade into compounds belonging to the PFSAs family of PFAS, including PFOS
(Ahrens 2011). Thus, transformation and degradation of PFAS should be considered as an
ongoing potential source of PFOS to the aquatic environment. Currently, the understanding of
these transformation and degradation processes is limited, particularly for PFOS. There is little
understanding of which PFAS and how much of each has been or will be released into the
aquatic environment (Liu and Mejia Avendano 2013; Wang et al. 2017). Additional work is
needed to fully understand the details of these processes and the occurrence of the compounds to
better comprehend their role as a source of PFOS to aquatic environments (Lau et al. 2007).
These transformation and degradation pathways are dependent on environmental
conditions, degradation kinetics, and the chemical structures and properties of the individual
PFAS precursors and volatile PFAS (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). The most well studied PFAS precursors are
fluorotelomer-based compounds, which are produced through telomerization technology and are
26
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associated with PFOA as the final product (Buck et al. 2011; Liu and Mejia Avendano 2013). In
contrast, perfluoroalkane sulfonamido derivatives and other PFAS, such as side-chain-
fluorinated polymers, are not as well studied.
It is essential to understand the biodegradation of volatile PFAS, such as perfluroalkane
sulfonamido derivatives as their degradation is directly linked with PFOS generation in the
environment (Liu and Mejia Avendano 2013). Most published studies on the degradation of
perfluoroalkane sulfonamido derivatives focus on those with eight fluorinated carbons since
PFOS is a final product (Buck et al. 2011). TV-ethyl perfluorooctane sulfonamidoethanol
(EtFOSE) in particular is the most commonly studied.
2.3,1 Degradation of perfluoroalkane sulfonamido derivatives
Perfluoroalkane sulfonamido derivatives, including perfluoroalkane sulfonamides,
sulfonamidoethanols, sulfonamidoethyl acrylates, and sulfonamidoethyl methacrylates, are final
products on their own and are important building blocks for further synthesis (Buck et al. 2011).
The various derivatives have been found to degrade into PFSAs, such as PFOS when sufficient
chain length is present, and are intermediates along the transformation pathway. These
derivatives include members of the TV-ethyl perfluoroalkane sulfonamidoacetic acids
(EtFASAAs), TV-ethyl perfluoroalkane sulfonamides (EtFASAs), perfluoroalkane
sulfonamidoacetic acids (FASAAs), perfluoroalkane sulfonamids (FASAs), FASA N-
glucuronides, and perfluoroalkane sulfinic acids (PFSiAs; Buck et al. 2011). Additionally, in the
environment TV-alkyl perfluoroalkane sulfonamidoethyl acrylates and methacrylates (and
polymers based on them) may undergo hydrolysis of the ester linkages to produce TV-alkyl
perfluoroalkane sulfonamidoethanols (FASEs; Buck et al. 2011).
In particular, several studies have demonstrated that EtFOSE, a member of the A-alkyl
perfluoroalkane sulfonamidoethanols of the perfluoroalkane sulfonamido substances, degrades
27
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into PFOS (Benskin et al. 2009; Boulanger et al. 2005b; Hatfield 2001; Lange 2000; Plumlee et
al. 2009; Rhoads et al. 2008). EtFOSE was a product of ECF and was a precursor compound for
the synthesis of other products such as phosphate esters that were used to manufacture paper
protectors (3M Company 1999). Several studies have investigated the degradation of EtFOSE
and all found that it is prone to degradation (Benskin et al. 2009; Boulanger et al. 2005b;
Hatfield 2001; Lange 2000; Plumlee et al. 2009; Rhoads et al. 2008).
The overall pathway of EtFOSE degradation was determined to be the major difference
between these studies (Figure 2-2). Rhoads et al. (2008) determined that EtFOSA could undergo
direct dealkylation to form perfluorooctane sulfonamide (FOSA; as shown by the red arrow in
Figure 2-2). Lange (2000) suggested that PFOA could be formed as a minor end product through
an abiotic one-electron transfer mechanism from perfluorooctane sulfinic acid (PFOSI;
demonstrated by the blue arrow in (Figure 2-2). In contrast, the other studies did not find PFOA
to be a degradation product (Benskin et al. 2013; Boulanger et al. 2005b; Rhoads et al. 2008).
Further, in the aerobic biodegradation studies, the rate limiting step was determined to be the
degradation of TV-ethyl perfluorooctane sulfonamidoacetic acid (EtFOSAA) and consequently
EtFOSAA was the major degradation product rather than PFOS (Liu and Mejia Avendano 2013).
Nevertheless, the degradation of EtFOSE resulted in the formation of PFOS as one of the final
degradation products. In contrast, in the abiotic degradation studies, PFOS and PFOSI were
either present at trace concentrations or were not observed (Hatfield 2001; Plumlee et al. 2009).
Instead FOSA was considered to be the stable end product (Plumlee et al. 2009). The differences
in the degradation pathways observed in the literature can likely be attributed to environmental
conditions (Buck et al. 2011; Liu and Mejia Avendano 2013). Nevertheless, these pathways
demonstrated that degradation of EtFOSE resulted in the formation of PFOS and should be
28
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considered a potential source of PFOS to the aquatic environment. However, currently the
relative contribution of this potential source to the aquatic environment cannot be quantified
(Buck et al. 2011; Liu and Mejia Avendano 2013).
OH
pFRFF.FF.FO ,
FFFFFFFF
EtFOSE
~ ~ ~
-------
2.3.2 Perfluorooctane sulfonamide-based side-chained polymers
In contrast to some other PF AS described in Section 1.2, fluorinated side-chain polymers
do not have the per- or polyfluorinated backbone. Instead, fluorinated side-chain polymers
consist of a variable composition with per- and polyfluoroalkyl side chains (Buck et al. 2011).
The side chains of each of these polymer types may serve to transform into PFAS. Currently,
little is known about these transformation processes (Liu and Mejia Avendano 2013). Given the
high production volume of perfluorooctane-sulfonamide-based side-chain polymers prior to
2002, these fluorinated side-chain polymers may contribute to the levels of PFAS in the
environment. It remains unknown how much these polymers contribute to the PFSAs in the
environment (Liu and Mejia Avendano 2013). However, this transformation process is expected
to occur over a long period of time (e.g., > 1,000 years) and may be a relatively small contributor
of PFAS, including PFOS, in the environment (Buck et al. 2011).
2.3.3 Fluoroalkyl surfactants used in AFFFs
The release of AFFF during firefighting activities has been determined to be a substantial
source of PFOS to the aquatic environment (see Section 2.1.2). Since 2002, fluorinated
alternatives to PFAAs have been used to manufacture AFFF (Buck et al. 2011; Wang et al.
2013). The ten classes of AFFF chemicals have been identified and show that the new
formulations of AFFF include the eight carbon perfluoroalkyl moiety (Place and Field 2012).
Some of these fluorinated alternatives may undergo transformation and degradation processes
and therefore may contribute to the levels of PFOS occurring in the aquatic environment (Liu
and Mejia Avendano 2013). However, additional details about the transformation and
degradation processes, including specific transformation pathways, the time to undergo
transformation to produce a final product, and the influence of the environmental condition, are
lacking at this time (Liu and Mejia Avendano 2013; Wang et al. 2013).
30
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2.4 Environmental Monitoring of PFOS in Abiotic Media
PFOS 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 medium in which PFOS is found (Lau et al. 2007). Occurrence and
detection of PFOS in surface waters is described below and occurrence in other abiotic media is
described in Appendix N.
2.4.1 PFOS Occurrence and Detection in Ambient Surface Waters
2.4.1.1 Summary of PFOS occurrence and concentrations across the U.S.
PFOS is one of the dominant PFAS detected in aquatic ecosystems, along with PFOA
(Ahrens 2011; Benskin et al. 2012; Dinglasan-Panlilio et al. 2014; Nakayama et al. 2007;
Remucal 2019; Zareitalabad et al. 2013). Despite its wide use and persistence in the aquatic
environment, current information on the distribution of PFOS in surface waters of the U.S. is
relatively limited (Jarvis et al. 2021). Available data are largely collected from freshwater
systems in eastern states, with most of the current, published PFOS occurrence data focused on a
handful of study areas with known manufacturing or industrial uses of PFAS and among areas of
known AFFF use, such as fire-training areas on military bases (Figure 2-3 and Appendix N).
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/
Figure 2-3. Map Indicating Sampling Locations for Perfluorooctane Sulfonate (PFOS)
Measured in Surface Waters across the United States (U.S.).
Based on data reported in the current, publicly available literature. Sampling locations for the Colorado data were
not available. Therefore, dash marks were used to indicate that measured PFOS surface water concentrations were
available for Colorado; however, the exact sampling locations within the state were not publicly available. Detailed
information on sampling locations, including references, coordinates (with the exception of Colorado), and sampling
site identification numbers and names, are provided in Appendix N.
Modified from: Jarvis et al. (2021).
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Concentrations of PFOS in surface waters across the U.S. appear to vary widely, with
observed concentrations ranging over eight orders of magnitude. PFOS is generally detected in
the picogram and nanogram per liter range with reported concentrations in microgram per liter
(or part-per-trillion) ranges (Ahrens 2011; Zareitalabad et al. 2013). For the purposes of this
overview, all concentrations reported here are in nanogram per liter (ng/L). Measured surface
water concentrations of PFOS in peer-reviewed journal articles and publicly available industry
and government reports range between 0.074 and 8,970,000 ng/L with an arithmetic mean
concentration of 786.77 ng/L and a median concentration of 3.6 ng/L (Figure 2-4) (Jarvis et al.
2021). However, it should be noted that the mean and median concentrations reported in Jarvis et
al. (2021) were calculated from the reported concentrations for individual samples and therefore,
are not fully representative of all the measured PFOS concentrations in U.S. surface waters.
Additionally, as demonstrated by the median concentration of 3.6 ng/L, a majority (roughly
91%) of measured PFOS concentrations were found to fall below 300 ng/L (Jarvis et al. 2021).
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C/3
o
u.
P-,
re
S
-------
PFOS in water is a result of the presence of an anthropogenic source, a transport pathway (air,
surface water, or ground water), and the persistence and mobility of the PFOS in the
environment. Therefore, PFOS concentrations in surface water tend to be dependent on the
presence of a nearby source and generally increase with levels of urbanization.
Further, there are insufficient data to quantitatively evaluate temporal trends of PFOS in
surface waters across the U.S. (Remucal 2019). However, recent studies have suggested that
PFOS concentrations in surface waters with limited sampling sites in northeastern states appear
to have decreased since the voluntary phase out of PFOS in 2002 (Pan et al. 2018; Zhang et al.
2016). While these studies observed lower measured PFOS concentrations in surface waters
compared to those reported in earlier reports (Hansen et al. 2002; Nakayama et al. 2007), few
studies have measured PFOS concentrations from the same sampling sites over time (Jarvis et al.
2021). Eight studies (six focused on the Great Lakes and two in New York on the Hudson River)
measured PFOS in the same waterbody over time (Appendix N). Thus, the observed lower
concentrations reported in recent literature could be due to trends of PFOS concentrations
decreasing since the 2002 PFOS phaseout, differences in sampling site locations, and/or
advances in analytical methods for detecting PFOS that reduced detection limits (Jarvis et al.
2021).
Despite the wide use and persistence of PFOS in aquatic ecosystems and unlike the
extensive sampling of PFOS in drinking water sources1, groundwater, and fish tissue
monitoring2, current information on the environmental distribution of PFOS in ambient surface
1 EPA's database for the Unregulated Contaminant Monitoring Rule (UCMR) that includes data for treated surface
waters (https://www.epa.gov/dwucmr').
2 EPA's National Rivers and Streams Assessment (NRSA; https://www.epa.gov/national-aquatic-resource-
surveys/ncca) and the Great Lakes Human Health Fish Tissue Study component of the EPA National Coastal
Condition Assessment (NCCA/GL).
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waters across the U.S. remains very limited. More recent sampling efforts indicate that PFOS
occurrence may be more widespread. PFOS was detected in almost all collected surface water
samples, which can likely be attributed to improvements in analytical methods that lowered the
PFOS detection limit compared to older analytical methods (Gewurtz et al. 2013).
Thus, from the currently available data, which were largely collected from freshwater
systems in eastern states and in the Upper Midwest with known manufacturing or industrial uses
of PFAS or use of AFFF, PFOS concentrations measured in U.S. surface waters appear to vary
widely, across eight orders of magnitude (Jarvis et al. 2021). PFOS concentrations in remote
areas (i.e., areas with little to no PFAS manufacturing and/or industrial uses) range between
0.074 to 23.23 ng/L (Jarvis et al. 2021). This contrasts with PFOS concentrations measured in
areas with known PFAS manufacturing, industrial use, and/or application of AFFF, which vary
widely and reach up to the maximum observed concentration of 8,970,000 ng/L at a site
impacted by AFFF (Appendix N). While current PFAS occurrence data illustrate the prevalence
and quantify concentrations of PFOS in surface waters across the U.S., additional data,
particularly in central, southwestern, and western freshwaters as well as saltwater systems, are
needed to better understand PFOS occurrence in aquatic ecosystems across the U.S. (Jarvis et al.
2021). See Appendix N for further discussion of PFOS occurrence in surface waters and other
abiotic media such as aquatic sediments, groundwater, air, and ice.
2.5 Bioaccumulation and Biomagnification of PFOS in Aquatic Ecosystems
PFAS, including PFOS, are 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 (see Section 2.2), facilitating their long-range transport to aquatic ecosystems in
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the remote arctic and mid-oceanic islands (see Section 2.2.2) (Haukas et al. 2007; Houde et al.
2006b). Several physical-chemical properties of PFOS contribute to bioaccumulation within
aquatic species once they have entered an ecosystem.
2.5.1 PFOS Bioaccumulation in Aquatic Life
In contrast to many persistent organic pollutants, which tend to partition to fats, PFOS
preferentially binds to proteins. Within an organism PFOS 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; Martin et al. 2003b). PFOS also binds to
ovalbumin, and the transfer of PFOS to such albumin in eggs can be an important mechanism for
depuration in female oviparous species, as well as a mechanism for maternal transfer of PFOS to
offspring (Jones et al. 2003; Kannan et al. 2005).
The stability of PFOS contributes to its bioaccumulation potential, as it has not been
found to undergo biotransformation within the organism (Martin et al. 2003a; Martin et al.
2003b). Within an organism, PFOS undergoes enterohepatic recirculation, in which PFOS is
excreted from the liver in bile to the small intestine, then reabsorbed and transported back to the
liver (Goecke-Flora and Reo 1996). This process becomes increasingly more efficient the longer
the perfluorinated chain length is, resulting in longer biological half-lives for chemicals like
PFOS with a relatively long chain length, as they are less readily excreted. PFAS with sulfonate
head groups, such as PFOS, are more efficiently resorbed by the small intestine than carboxylate
PFAS such as PFOA, resulting in higher bioaccumulation levels (Hassell et al. 2020; Jeon et al.
2010; Martin et al. 2003a).
Sex differences in the elimination rates of PFOS in addition to the transfer of PFOS to
albumin in eggs (e.g., Jones et al. 2003; Kannan et al. 2005) have not been well studied. Some
research suggests lower PFOS elimination rates in female rats than in male rats (Butenhoff et al.
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2012; Chang et al. 2012; Pizzurro et al. 2019), suggesting potentially longer retention of PFOS in
females. However, this difference was not observed in mice, rabbits, monkeys, or humans
(Pizzurro et al. 2019). In contrast to PFOS, PFOA elimination rates are higher in females than in
males for both female fathead minnows (Lee and Schultz 2010) and rats (Pizzurro et al. 2019),
suggesting potential longer retention of PFOA in males. These data indicate further research
across species and genders for PFAS elimination rates may be useful.
The structure of PFOS also affects its bioaccumulation potential, with linear forms being
more bioaccumulative than branched forms (Fang et al. 2014; Hassell et al. 2020). The
preferential accumulation of linear PFOS occurs because the elimination rate of branched
isomers of PFOS is higher, particularly across gill surfaces (Hassell et al. 2020). This pattern has
also been observed in the field, as the proportion of branched isomers was higher in 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 PFOS Bioaccumulation and Biomagnification in Aquatic Ecosystems
Because of their affinity for binding to proteins, PFAS can enter the base of the food web
through sorption to organic matter in sediments or biofilms (Higgins and Luthy 2006; Jeon et al.
2010; Penland et al. 2020), or can bind to blood proteins at gill surfaces of aquatic organisms
through respiration (De Silva et al. 2009; Hassell et al. 2020; Martin et al. 2003a; Martin et al.
2003b).
PFAS 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 sulfonate or carboxylate head groups (Higgins
and Luthy 2006)(see Section 2.2 for further details on the sorption of PFOS). Overall, these
results suggest that sorption to sediments should be an important mechanism for PFOS entry into
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an aquatic ecosystem, but that subsequent dietary uptake from benthic feeding organisms will be
more important for PFOS than PFOA.
The importance of the sediment pathway for PFOS bioaccumulation in aquatic
ecosystems has been demonstrated in laboratory studies with Chironomus riparius (Bertin et al.
2014), C. plumosus (Wen et al. 2016), Gammarus fossarum and G. pulex (Bertin et al. 2016),
and Lumbriculus variegatus (Lasier et al. 2011), where PFOS concentrations were positively
correlated between sediments and whole-body tissue samples of benthic feeding organisms. The
sediment pathway has also been demonstrated in several field studies, where PFOS was
measured in sediments and biofilms, and was higher in benthic-feeding invertebrates relative to
pelagic-feeding invertebrates (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 fish (Thompson et al.
2011) than they were to their distribution in pelagic organisms.
PFAS can also enter aquatic organisms directly from the water column through
respiration. Because of its binding affinity to proteins, PFOS can enter the body of gill-breathing
organisms by binding to proteins in the blood at gill surfaces (Jones et al. 2003; Martin et al.
2003a; Martin et al. 2003b). The relative distribution of PFOS in tissues is related to the primary
route of exposure (dietary or respiratory). In rainbow trout, the rank order of PFOS
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 PFOS in the intestines of green
eel goby, soft tissues, shell, and legs of shore crabs; and gills and intestines of oysters, suggesting
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bioaccumulation through both dietary and aqueous uptake in invertebrates, and primarily dietary
uptake in fish.
In addition to being bioaccumulative, PFOS has been shown to biomagnify with
increasing trophic level in a variety of freshwater ecosystems (Kannan et al. 2005; Martin et al.
2004; Penland et al. 2020; Xu et al. 2014) and saltwater ecosystems (de Vos et al. 2008; Houde
et al. 2006b; Loi et al. 2011; Powley et al. 2008; Tomy et al. 2004) in North America, Europe,
and Asia. PFOS is often the most abundant PFAS in aquatic organisms, and this high relative
abundance is at least partially explained by the biotransformation of PFOS precursor chemicals
into PFOS (see Section 2.3)(Haukas et al. 2007; Kannan et al. 2005; Kelly et al. 2009; Martin et
al. 2004; Tomy et al. 2004). Higher trophic level organisms have a greater capacity to metabolize
PFOS precursor chemicals, which have been found in lower concentrations in increasing trophic
level (Fang et al. 2014; Kannan et al. 2005; Martin et al. 2004). This suggests that in addition to
biomagnification, some of the trophic-level increase in PFOS can be explained by the
biotransformation of precursor chemicals.
2.5,3 Environmental Monitoring of PFOS in Biotic Media
PFOS is one of the dominant PFAS detected in aquatic ecosystems, along with PFOA
(Ahrens 2011; Benskin et al. 2012; Dinglasan-Panlilio et al. 2014; 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 then, the global distribution of PFAS in tissues of
aquatic 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).
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In lentic surface waters of the U.S., one of the most comprehensive studies of PFOS
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, PFOS was measured in fish collected at every site,
with a median concentration of 15.2 ng/g ww (Stahl et al. 2014). Lake trout (31% of sampled
species), smallmouth bass (14%), and walleye (13%) were the most commonly-sampled species
from the Great Lakes samples, and the average PFOS concentrations in lake trout muscle were
more than twice as high as PFOS concentrations in muscle of smallmouth bass and walleye
(Stahl et al. 2014).
Martin et al. (2004) measured PFOS in whole body samples of invertebrates and fish in
Lake Ontario, near the town of Niagara-on-the-Lake. PFOS concentrations were much higher in
the benthic amphipod Diporeia hoya (280 ng/g ww) than in the more pelagic My sis relicta (13
ng/g ww), suggesting sediments are an important source of PFOS in this area (Martin et al.
2004). Among the four fish species sampled, whole body PFOS concentrations were highest in
the slimy sculpin (450 ng/g ww), whose preferred food source is D. hoya (Martin et al. 2004).
Although adult lake trout occupy the highest trophic level at this site, based on nitrogen stable
isotope analysis, their PFOS concentrations were less than half (170 ng/g ww) of those measured
in sculpin, as their food web is largely pelagic, and not affected by the high sediment PFOS
concentrations. Based on stomach content analysis, 90% of the adult lake trout diet consists of
alewife, which feed primarily on the more pelagic M. relicta, and have the lowest average PFOS
concentration (46 ng/g ww) among all fish species (Martin et al. 2004).
Guo et al. (2012) measured PFOS in lake trout muscle tissues in Canadian waters of Lake
Superior, Huron, Erie, and Ontario. Average PFOS concentrations correlated with watershed
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urbanization, and were 0.85, 8.3, 27, and 46 ng/g ww, respectively (Guo et al. 2012). Delinsky et
al. (2010) measured PFOS in bluegill, black crappie, and pumpkinseed muscle tissue in 59 lakes
in Minnesota, including four lakes in the Minneapolis-St. Paul metropolitan area. PFOS was
detected in muscle tissues of fish collected in 13 of the 59 lakes, and concentrations ranged from
1.08 to 52.4 ng/g ww in lakes where it was detected. In the four lakes in the Minneapolis-St. Paul
metropolitan area, PFOS concentrations in fish muscle tissues ranged from 4.39 to 47.3 ng/g ww
(Delinsky et al. 2010).
In flowing surface waters of the U.S., one of the most comprehensive fish PFOS
monitoring studies included the collection of fish muscle tissue data from 164 urban river sites
(5th order or higher) across the conterminous U.S. selected following a probabilistic design. The
study was part of the 2008 - 2009 National Rivers and Streams Assessment and the National
Coastal Condition Assessment (Stahl et al. 2014). PFOS was detected in 73% of the urban river
sites, with a median concentration of 10.7 ng/g ww (Stahl et al. 2014). Largemouth bass (34% of
sampled species), smallmouth bass (25%), and channel catfish (11%) were the most commonly
sampled species from the urban stream sites, and PFOS concentrations in the muscles of
largemouth bass were approximately twice as high as concentrations in the muscles of
smallmouth bass (Stahl et al. 2014).
Ye et al. (2008) reported average PFOS concentrations of 83.1, 84.6, and 147 ng/g ww
from whole body composite samples of multiple fish species from the Mississippi River,
Missouri River, and Ohio River, respectively. Delinsky et al. (2010) sampled PFOS in bluegill,
black crappie, and pumpkinseed muscle tissue at several locations along the upper Mississippi
River in 2007, and found concentrations ranging from 3.06 ng/g ww at unimpacted sites to 2,000
ng/g ww at Pool 2, a heavily impacted site in the Minneapolis-St. Paul metropolitan area
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(Delinsky et al. 2010). Malinsky et al. (2011), as reported in Stahl et al. (2014), measured PFOS
concentrations ranging from 41.7 to 180 ng/g ww in fish muscle samples collected along the
Mississippi River, with the lowest concentration reported for sauger and the highest reported for
bluegill.
Kannan et al. (2005) measured PFOS in invertebrates and vertebrates from two rivers in
Southern Michigan (Raisin River, St. Claire River), and one in Northern Indiana (Calumet
River). PFOS concentrations were similar across sites for the different taxa and suggested trophic
biomagnification for PFOS. Among invertebrate taxa, zebra mussel PFOS soft tissue whole body
concentrations ranged from below detection to 3.1 ng/g ww, amphipod whole body
concentrations ranged from below detection to 2.9 ng/g ww, and crayfish whole body
concentrations ranged from 2.4 to 4.3 ng/g ww. Among fish, PFOS concentrations in round goby
whole body samples ranged from 6.6 to 21.5 ng/g ww, and smallmouth bass muscle samples
ranged from below detection to 41.3 ng/g ww (Kannan et al. 2005).
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. PFOS was measured in whole body tissues of snails (6.47 ng/g ww) but was not detected
in whole body tissues of Asian clam, unionid mussels, or crayfish. The highest concentrations in
invertebrates were measured in aquatic insect whole body samples (132.8 ng/g ww) and was
hypothesized to result from dietary uptake of aquatic biofilms. PFOS was measured in muscle
tissue of all 11 sampled fish species and ranged from 11.42 ng/g ww in channel catfish to 37.36
ng/g in whitefin shiner. The highest concentration that Penland et al. (2020) measured was 482.9
ng/g ww, from the eggs of a single robust redhorse sample, underscoring the preferential binding
of PFOS to ovalbumin.
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Houde et al. (2006a) measured whole body PFOS in six fish species in Charleston
Harbor, South Carolina, and whole body PFOS in zooplankton and five fish species in Sarasota
Bay, Florida. Charleston Harbor was the more developed of the two sites and had higher overall
PFOS concentrations. Average PFOS concentrations in Charleston Harbor ranged from 19 ng/g
ww in pinfish to 92 ng/g in spot. In Sarasota Bay, PFOS concentrations averaged 0.2 ng/g ww in
zooplankton and ranged from 3.1 ng/g ww in pigfish to 8.8 ng/g ww in spotted seatrout,
suggesting evidence of trophic biomagnification.
Lescord et al. (2015) measured PFOS in chironomids, zooplankton, and juvenile and
adult arctic char in six high arctic lakes in Canada. Two of these lakes had been contaminated by
PFAS from a nearby airport while the other lakes were free from point source contamination.
PFOS in chironomid whole body samples was high at the two contaminated lakes, ranging from
28 to 445 ng/g ww, compared to 5.3 to 14 ng/g ww at the reference lakes (Lescord et al. 2015).
Whole body concentrations in pelagic zooplankton were relatively lower, ranging from 49 to 60
ng/g ww, compared to 0.12 to 2.0 ng/g ww at the reference lakes. The higher concentrations of
PFOS in the benthic chironomids indicate the importance of sediments as a route of exposure
into the base of the food web. PFOS in whole body samples of juvenile char (181 to 224 ng/g
ww) and muscle tissue of adult char (24 to 117 ng/g ww) at the two contaminated lakes were
lower than whole body PFOS in chironomids, indicating a lack of trophic biomagnification.
Additionally, PFOS in whole body samples of juvenile char (0.001 to 15 ng/g ww) and muscle
tissues of adult char (below detection to 2 ng/g ww) at the four reference lakes was also lower
than whole body PFOS in chironomids at the four reference lakes.
Tomy et al. (2004) measured PFOS in whole body samples of zooplankton (Calamus
hyperboreus), shrimp (Pandalus sp.), clams (Nya truncata and Serripes groenlandica), and arctic
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cod (Boreogadus saida); and liver samples of deepwater redfish (Sebastes mentella) collected
from unimpacted marine locations in the Canadian Arctic. PFOS concentrations were low for all
taxa, with the lowest concentrations measured in shrimp (0.3 ng/g ww) and clams (0.04 ng/g
ww). PFOS concentrations were similar in zooplankton (1.8 ng/g ww), arctic cod (1.3 ng/g ww),
and redfish (1.4 ng/g ww), indicating little, if any biomagnification from invertebrates to fish
(Tomy et al. 2004). Haukas et al. (2007) found the average liver PFOS concentration (2.02 ng/g
ww) in arctic cod B. saida collected in the Barents Sea off the coast of Svalbard in 2004 to be
similar to whole body concentrations for this species reported by Tomy et al. (2004). The
average whole body PFOS concentration (3.85 ng/g ww) in ice amphipod (Gammarus wilkitzkii)
samples was higher than the average liver PFOS concentration in in arctic cod, indicating no
biomagnification from invertebrates to fish in this ecosystem (Haukas et al. 2007).
Current data indicate that PFOS concentrations measured in aquatic biota vary widely,
approximately across four orders of magnitude for both fish (ranging between 0.85 and 2,000
ng/g ww) and aquatic invertebrates (ranging between 0.04 and 445 ng/g ww). Like ambient
surface water concentrations, PFOS concentrations in aquatic biota inhabiting remote areas (i.e.,
areas with little to no PFAS manufacturing and/or industrial uses) appear to be lower than those
in areas with known PFAS manufacturing, industrial use, and/or application of AFFF. While
current PFAS monitoring data illustrate the prevalence and quantify concentrations of PFOS in
aquatic biota across the U.S., additional data are needed to better understand PFOS occurrence
and potential bioaccumulation in aquatic ecosystems across the U.S.
2.6 Exposure Pathways of PFOS in Aquatic Environments
There are multiple exposure pathways of PFOS in the aquatic environment, including: 1)
direct (dermal and respiratory) aqueous exposure; 2) direct exposure to contaminated sediment
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(for benthic organisms); 3) dietary and biomagnification; and 4) maternal-transfer (Ankley et al.
2020). Exposure of PFOS 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). Upon entering an organism, PFAS such as PFOS tend
to bind to proteins, and concentrate preferentially within the blood and protein rich tissues, such
as liver (Haukas et al. 2007; Xia et al. 2013b). The affinity of PFOS to bind to proteins
contributes to the bioaccumulation and biomagnification of PFOS (see Section 2.5 above),
resulting in increasing concentrations of PFOS in higher trophic level organisms, such as
predatory fish and birds (Custer et al. 2019; Haukas et al. 2007; Xu et al. 2014). However, as
noted previously in Section 1.2.1, the lack of a meaningful Kow for PFOS due to its binding
primarily to protein, not lipids, precludes application of Kow-based models that are commonly
used to estimate bioconcentration factors and predict bioaccumulation for many other important,
environmental contaminants (e.g., PCBs). Lastly, elevated PFOS concentrations in eggs and
young of aquatic life suggests that PFOS may be maternally transferred to offspring. This
exposure pathway may be particularly important among egg-laying species because of the
preferential binding of PFOS to egg albumin (Kannan et al. 2005). In summary, PFOS exposure
has been found to occur through multiple exposure routes, including via water, sediment, diet,
and maternal transfer (Jones et al. 2003; Kannan et al. 2005; Sharpe et al. 2010; Wang et al.
2011).
2.7 Effects of PFOS on Biota
The number of PFOS ecotoxicity studies and data are increasing and study designs are
evolving to expand the understanding of the effects of PFOS. Currently, PFOS ecotoxicity
studies are primarily focused on fish, aquatic invertebrates, plants, and algae. Fewer studies are
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being conducted on aquatic-dependent birds, reptiles, and mammals. Sections 3 and 4 provide
study summaries of individual publicly available high quality aquatic life toxicity studies, and
Appendix A through Appendix H summarize current PFOS aquatic life ecotoxicity data, both
studies used here and unused studies due to quality issues.
2.7,1 Mode of Action and Toxicity of PFOS
The mechanism(s) underpinning the toxicity of PFOS is not well-understood and is an
active area of research. Toxicity literature indicate that PFOS causes a wide range of adverse
effects in aquatic organisms, including reproductive effects, developmental toxicity, and
estrogen, androgen and thyroid hormone disruption (see Sections 3 and 4 and Appendices A.l
through H. 1). However, a great deal of research is still needed from a mechanistic perspective to
better understand how the different modes of action elicit specific biological responses. Some
potential PFOS modes of action in aquatic life appear to include: 1) oxidative stress (Li et al.
2017; Sant et al. 2018; Shi and Zhou 2010); 2) autophagic cell death or apoptosis (Sant et al.
2018; Shi et al. 2008); 3) endocrine modulation of estrogen and thyroid receptors (Benninghoff
et al. 2011; Chen et al. 2018; Du et al. 2013; Kim et al. 2011; Shi et al. 2008); 4) interference at
the mitochondrial level through the uncoupling of oxidative phosphorylation (ECCC 2018); 5)
interference with the homeostasis of DNA metabolism (Hoff et al. 2003); and 6) activation of the
nuclear peroxisome proliferator activated receptor-alpha (PPAR-a) pathways (Arukwe and
Mortensen 2011; Cheng et al. 2016; Fang et al. 2013; Fang et al. 2012; Yang et al. 2014).
Following exposure to PFOS, molecular level events can perturb estrogen-, androgen-
and thyroid-related endocrine systems, as well as neuronal-, lipid-, and carbohydrate-metabolic
systems and lead to cellular- and organ-level disturbances and ultimately result in effects on
reproduction, growth, and development at the individual organism-level(see Ankley et al. 2020
and Lee et al. 2020 for the latest reviews on the subject) for the latest reviews on the subject).
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The mechanisms of PFOS toxicity to fish in particular appear to be related to oxidative stress,
apoptosis, thyroid disruption, and alterations of gene expression during development (Lee et al.
2020). Additionally, published research suggested that many of these molecular pathways
interact with each other and could be linked. For example, oxidative stress following exposure to
PFOS was 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 PFOS induced oxidative stress operates still requires additional study, but
increased B-oxidation of fatty acids and mitochondrial toxicity are proposed triggers (Ankley et
al. 2020; Lee et al. 2020). Thus, the alteration of multiple biological pathways is a plausible
explanation for the diversity of observed effects of PFOS reported in the literature (Lee et al.
2020). However, the available data did not allow for a defined adverse outcome pathway-based
understanding of the ultimate reductions to survival, growth, and reproduction in the various
aquatic taxa in which these effects have been observed or may be expected to occur. Thus,
further mechanistic research is warranted.
Notably, PFOS appeared to be related to the disruption of the sex hormone-related
endocrine system at the molecular, tissue, and organ levels, resulting in observed adverse
reproductive outcomes in freshwater and saltwater fish and invertebrates alike. Further, these
effects have been reported after exposure via multiple exposure routes (i.e., waterborne, dietary,
maternal; (i.e., waterborne, dietary, maternal; Lee et al. 2020). And these reproductive effects
also appeared to be trans-generational, as observed in a multi-generational zebrafish (Danio
rerio) study by Wang et al. (2011) (see study summary in Appendix Section C.2.5).
PFOS is one of the most studied PFAS in the ecotoxicity literature, with reported adverse
effects on survival, growth, and reproduction. However, a great deal of additional research is
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needed to better understand the modes of action of PFOS. Specifically, additional research from
a mechanistic perspective is needed to better understand how the different modes of action elicit
specific biological responses in fish, aquatic invertebrates, and amphibians. Potential effects of
PFOS involving multiple biological pathways are a research challenge for PFOS and PFAS in
general.
2.7.2 Potential for Interactions with Other PFAS
PFAS 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 (see also Section
2.4.1)(Ahrens 2011; Ahrens and Bundschuh 2014; Giesy and Kannan 2002; Houde et al. 2006b;
Keiter et al. 2012; Wang et al. 2017). Although the EPA's PFOS recommended aquatic life water
quality criteria are based solely on single chemical exposures to aquatic life, it is recognized that
PFAS are often introduced into the aquatic environment as end-use formulations comprised of
mixtures of PFAS and/or PFAS-precursors. However, the ecological effects of these potential
PFAS mixtures are poorly understood (Ankley et al. 2020). It was useful, therefore, to briefly
summarize the types of interactions that might be expected based on the few PFAS mixture
studies involving PFOS and one or more PFAS to date. It should be noted 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 below are as reported by the study authors without
any additional interpretation or analysis of uncertainty. At both the organismal and cellular
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levels, studies on zebrafish (D. rerio\ Ding et al. 2013), a water flea (D. magna; Yang et al.
2019), mosquito (Aedes aegypti; Olson 2017), a bioluminescent cyanobacteria (Anabaena sp.;
Rodea-Palomares et al. 2012), or with cultured hepatocytes of the cyprinid (Gobiocypris rarus;
Wei et al. 2009), demonstrated that the effects observed from in vivo and in vitro tests on PFAS
mixtures vary and can have unpredictable exposure and species-specific effects. 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) 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, D. magna, to single and binary mixtures of PFOS and PFOA. The
authors reported synergism in acute and chronic toxic effects. Conversely, Rodea-Palomares et
al. (2012) showed a binary PFOS and PFOA mixture as having an antagonistic interaction across
the whole range of effect levels tested using the bioluminescent cyanobacterium, Anabaena.
Olson (2017) exposed larvae of the mosquito, A. aegypti, to PFOS and perfluorohexane sulfonate
(PFHxS) separately and as a mixture and reported increased toxicity in a manner greater than
would be predicted by additivity.
In tests with cultured hepatocytes of the cyprinid, G. rarus, co-exposure of PFOS with a
mixture of five other PFAS [PFOA, Perfluorononanoate or Perfluorononanoic acid (PFNA),
Perfluorodecanoate or Perfluorodecanoic acid (PFDA), Perfluorododecanoate or
Perfluorododecanoic acid (PFDoA), and 8:2 FTOH] resulted in highly complex interactions (Wei
et al. 2009). A number of genes differentially expressed in the mixture were not differentially
expressed in the exposure to the individual chemicals, potentially indicating different modes of
action for the mixture compared to the individual chemicals. In this case, the authors reported no
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additive responses for the mixture. Consistent with the possible mechanisms of toxicity of PFOS
(see Section 2.7.1), the genes identified in the study are involved in multiple biological functions
and processes, including fatty acid metabolism and transport, xenobiotic metabolism, immune
response, and oxidative stress (Wei et al. 2009). Finally, U.S. EPA (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 PFOS 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 2-5 summarizes sources, potential pathways of PFOS exposure for aquatic
life and aquatic-dependent wildlife, and possible toxicological effects.
PFOS initially enters the aquatic environment through point sources, including municipal
and industrial dischargers and landfill leachate and non-point sources, including land application
of contaminated biosolids (see Section 2.1.2). PFOS enters the aquatic environment in dissolved
and particle-bound forms and may sorb to surfaces, such as sediment and particulate matter in
the water column (see Section 2.2.2), which is depicted in the conceptual model (Figure 2-5).
The conceptual model depicted in Figure 2-5 shows exposure pathways for the biological
receptors of concern (i.e., aquatic life) and potential effects (e.g., on survival, growth, and
reproduction) in those receptors. Both direct (i.e., exposure from the water column which is
represented by **) and indirect (i.e., dietary exposure via the food web *) pathways are
represented in the conceptual model.
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aii
a
u
I*-*
U
PFOS Source
Point Sources
(from landfill leachate, municipal &
industrial dischargers, and
applications such as surfactants,
textile stain, & soil
repellents )
PFOS id Water
Dissolved & Particle-Bound
Degradation &
Metabolism of
Other PFASs
PFOS iu Sediment
PFOS Source
Nonpoint Sources
(from Aerial deposition AFFF,
land application of paper residuals
contaminated biosolids)
Aquatic Life
Producers
l:t Trophic Transfer
(from phytoplankton, periphyton, macrophytes; e.g., algae, cyanobacteria, waterweecL'c ommon eelgrass)
ir
o
cL
ai
0
it
01
Consumers
2nd Trophic Transfer
(to zooplankton. macroinvertebrates:
e.g., cladocerans'copepods &
mayflies/ribbed mussels)
Consumers
3rd Trophic Transfer
(to predatory fish:
e.g., longnose. dace/
American shad)
*ฆ
Consumers
4th Trophic Transfer
(to predatory fish;
e.g., largemouth bass'
striped bass)
~
Figure 2-5. Conceptual Model Diagram of Sources, Compartmental Partitioning, and Trophic
Transfer Pathways of Perfluorooctane Sulfonate (PFOS) in the Aquatic Environment and its
Bioaccumillation and Effects in Aquatic Life.
PFOS sources represented in ovals, compartments within the aquatic ecosystem represented by rectangles, and effects in
pentagons. Examples of organisms in each trophic transfer provided as freshwater/marine. Movement of PFOS from water to
receptors indicated by two separate pathways: bioconcentration by producers (*) and direct exposure (**) to all trophic levels
within box. Relative proportion of PFOS 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 "explicit expressions of the actual environmental
value that is to be protected" and are defined by an ecological entity (species, community, or
other entity) and its attribute or characteristics (U.S. EPA 1998). Assessment endpoints may be
identified at any level of organization (e.g., individual, population, community). In the context of
the CWA, aquatic life criteria for toxic pollutants are typically determined based on the results of
toxicity tests with aquatic organisms in which unacceptable effects on growth, reproduction, or
survival occurred. 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.
The use of laboratory toxicity tests to protect aquatic species was based on the concept
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 diversified, the
1985 Guidelines recommended acceptable data be available for at least eight genera with a
specified taxonomic diversity (the standard eight-family minimum data requirements, or MDRs).
The intent of the eight-family MDR was to serve as a 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 protect 95 percent of a
group of diverse taxa, and any commercially and recreationally important species; (the intent is
to protect 95 percent of a group of diverse taxa, and any commercially and recreationally
important species; U.S. EPA 1985).
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For more details on aquatic life assessment endpoints for PFOS see Section 3.1 below.
This criteria derivation for aquatic life was developed using a genus sensitivity distribution
(GSD), which represents the potential for impact to the survival, growth, or reproductive effects
on taxa across aquatic communities.
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
compiles acute toxicity test data from a minimum of eight diverse taxonomic groups.
Acute freshwater criterion require data from the following taxonomic groups:
a. fish in the family Salmonidae in the class Osteichthyes
b. a second family of fish 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 criterion 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 animal components of the aquatic
ecosystem from long term exposures, chronic toxicity test data are recommended from the same
eight diverse taxonomic groups that are recommended for acute criteria. If the eight diverse
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taxonomic groups are not available to support the chronic criterion derivation using a genus
distribution approach, the chronic criterion may be derived using an acute-to-chronic ratio
(ACR) approach. To apply an ACR approach to derive a chronic criterion, a minimum of three
taxa are recommended, with at least one chronic test being from an acutely sensitive species. To
calculate ACRs, chronic aquatic life criteria require data from the following taxonomic groups:
a. At least one fish
b. At least one invertebrate
c. At least one acutely sensitive freshwater species, for freshwater chronic criterion
(the other two may be saltwater species)
d. At least one 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 were examined to
determine whether aquatic plants are likely to be adversely affected by the concentration
expected to be protective for other aquatic organisms (see Appendix E for freshwater plant
toxicity studies).
2.10.2 Measure of PFOS Exposure Concentrations
This PFOS aquatic life AWQC document provides a critical review of all data identified
in the EPA's literature search for PFOS, including all forms of PFOS used in toxicity literature
(such as the anionic form and salts) and identified in the ECOTOX database:
the anionic form (CAS No. 45298-90-6)
the acid form (CAS No. 1763-23-1)
potassium salt (CAS No. 2795-39-3)
an ammonium salt (CAS No. 56773-42-3)
sodium salt (CAS No. 4021-47-0)
and a lithium salt (CAS No. 29457-72-5)
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Typically, studies do not complete an analysis or provide enough information regarding
isomer delineation to determine if the PFOS tested was purely linear or branched. However,
several PFOS toxicity studies stated that the linear PFOS isomer was used for dosing with fewer
studies indicating that the branched isomer was used. Studies reported by researchers that
conducted PFOS-only exposures were considered for possible inclusion. For most the EPA
aquatic life criteria documents with non-bioaccumulative substances, organisms are exposed to
contaminated water but fed a diet grown in uncontaminated media (not spiked with the toxicant
prior to introduction into the exposure chambers). Such tests were reviewed, and tests of
sufficient quality are included in these PFOS criteria. Toxicity tests conducted with PFOS-spiked
diet were also reviewed and considered suitable for deriving a criterion for this bioaccumulative
pollutant; however, these toxicity tests were limited in the current PFOS toxicity literature.
Consequently, toxicity tests with direct aqueous, dietary, and maternal transfer were included in
the EPA's derivation of aquatic life criterion for PFOS (see Section 3). Studies not included in
the numeric criteria derivation, including some studies with other PFOS exposures (i.e., in vitro
studies), were considered qualitatively as supporting information if they were deemed to be of
sufficient quality, and are described in the Effects Characterization section below (Section 4.3).
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; Elonen 2020). The search terms are often
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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:
a. The toxic effects are related to single chemical exposure (unless the study is being
considered as part of a mixture effects assessment)
b. 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
c. Chemical test concentrations are reported
d. There is an explicit duration of exposure
e. Toxicology information that is relevant to EPA Office of Water (OW) is reported for
the chemical of concern
f. The paper is published in the English language
g. The paper is available as a full article (not an abstract)
h. The paper is publicly available
i. The paper is the primary source of the data
j. A calculated endpoint is reported or can be calculated using reported or available
information
k. Treatment(s) are compared to an acceptable control
1. The location of the study (e.g., laboratory vs. field) is reported
m. The tested species is reported (with recognized nomenclature)
Following inclusion in the ECOTOX database, toxicity studies were subsequently
evaluated by EPA OW. All studies were evaluated for data quality as described by U.S. EPA
(1985) and in 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). Office of Water
completed a Data Evaluation Record (DER) for each species by chemical combination from the
PFOS 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
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Appendix Q with the intent to convey the meticulous level of evaluation, review, and
documentation each PFOS study identified by ECOTOX was subject to.
The 1985 Guidelines document indicates that tests used in criteria should be for North
American resident species. Due to the EPA's interest in using all available quality data,
particularly for data-sparse PFOS (relative to cadmium or ammonia, for example), PFOS toxicity
studies were considered for possible inclusion regardless of the test species residential status in
North America, as with other published aquatic life criteria. This approach was also based on the
relative similarity in sensitivities between resident and non-resident species (see Sections 3 and
4). Moreover, non-North American resident species serve as taxonomically-related surrogate test
organisms for the thousands of untested resident 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 4.2). These supporting analyses provided an additional line-of-
evidence that further suggested it is appropriate to consider non-resident species in PFOS criteria
derivation because of their minimal influence of the criteria magnitudes.
Additionally, a substantial number of PFOS toxicity tests reported only nominal, or
unmeasured, PFOS concentrations. For PFOS, 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 PFOS 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 PFOS, with the exception of
saltwater tests and freshwater studies that contained substrate. Nominal and measured
concentrations in the analysis generally displayed a high degree of linear correlation (>0.95
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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 Conversely, 65.50%) of
measured PFOS saltwater concentrations differed from corresponding nominal concentrations by
>20%o (EPA's OCSPP's Ecological Effects Test Guidelines (U.S. EPA 2016b) consider tests
acceptable when measured concentrations are within 20% of nominal, and Rewerts et al. (2021)
suggested that PF AS-specific toxicity tests may even be acceptable if measured and nominal
concentrations do not differ by up to 30%>). Potential dosing errors, differences in experimental
design, and/or the presence of substrate were hypothesized to be the primary contributor to these
discrepancies (Jarvis et al. 2023). Therefore, when available, measured PFOS concentrations
were used; however, for several studies measured PFOS concentrations were not reported, and
nominal concentrations were utilized, especially if a concentration-response relationship was
observed in another medium where PFOS was measured from the same study (e.g., diet, blood,
or eggs).
Typically, per the 1985 Guidelines, acute toxicity data from all measured flow-through
tests would be used to calculate species mean acute values (SMAV), 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 (an exception being for renewal tests with
cladocerans where test concentrations were measured). In the case of PFOS, static, renewal, and
flow-through experiments were considered for possible inclusion for both species mean acute
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and chronic values regardless of whether PFOS concentrations were measured because PFOS is
a highly stable compound (see Section 1.2.1), resistant to hydrolysis, photolysis, volatilization,
and biodegradation (see Section 2.3)(Giesy et al. 2010).
Additionally, chronic values were based on endpoints and durations of exposure that
were appropriate to the species. Thus, both life- and partial life-cycle tests were utilized for the
derivation of the chronic criteria. However, it should be noted that typically, per the 1985
Guidelines, life-cycle chronic tests would be preferred for invertebrates. The chronic studies used
in the derivation of the PFOS criteria followed taxa-specific exposure duration requirements
from various test guidelines (i.e., EPA's 1985 Guidelines and EPA's OCSPP's Ecological
Effects Test Guidelines, (i.e., EPA's 1985 Guidelines and EPA's OCSPP's Ecological Effects
Test Guidelines, U.S. EPA 2016b) when available. For example, the EPA's 1985 Guidelines
states that daphnid tests should begin with young < 24 hours old and last for not less than 21
days; and this chronic test duration was applied to the consideration of all chronic daphnid tests.
When taxa-specific exposure duration requirements were not available for a particular test
organism in the PFOS toxicity literature, both life- and partial life-cycle tests were considered in
the derivation of the chronic criteria.
PFOS toxicity in aquatic life can be manifested as effects on survival, growth, and/or
reproduction. Measurements of fish tissue, such as whole-body, muscle, and eggs, were most
closely linked to the chronic adverse effects of PFOS, since PFOS is highly persistent and
bioaccumulative. The following subsection of this problem formulation describes the approaches
used to establish PFOS effect concentrations in aquatic life in relation to the various criteria
derived, including for water and tissue.
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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 toxicity test data are 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 (Table 2-1). 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 (e.g., immobilization) in 50 percent of the test
organisms. The IC50 is the concentration of a chemical that is estimated to inhibit some
biological process (e.g., enzyme activity associated with an apical endpoint such as mortality) in
50 percent of the test organisms.
2.10.3.2 Chronic Measures of Effect
The measure of effect for chronic exposures of PFOS was the effect concentration
estimated to produce a chronic effect on survival, growth, or reproduction in 10 percent of the
test organisms (EC10; Table 2-1). The EPA selected an EC10 to estimate a low level of effect that
would be both different from controls and not expected to be severe enough to cause severe
effects at the population level for a bioaccumulative contaminant, such as PFOS. The use of the
EC10, instead of anEC2o, 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
2016c), and is consistent with the harmonized guidelines from OECD and the generally preferred
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effect level for other countries such as Canada, Australia and New Zealand (CCME 2007; OECD
2001; Warneetal. 2018).
Regression analysis was used preferentially to characterize a concentration-response (C-
R) relationship and to estimate concentrations at which chronic effects are expected to occur.
Author-reported No Observed Effect Concentrations (NOECs) and Lowest Observed Effect
Concentrations (LOECs) were only used for the derivation of 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 are used, a Maximum Acceptable Toxicant Concentration (MATC,
geometric mean of the NOEC and LOEC) is calculated. For the calculation of the chronic
criteria, 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 provided
additional information that is difficult to determine with an MATC, such as a measure of effect
level across a range of tested concentrations.
In conformity with the 2013 Ammonia Freshwater Aquatic Life Criteria (U.S. EPA
2013), a decision rule was also applied to the PFOS toxicity data when an author-reported NOEC
or LOEC was used. The decision rule was not to use "greater than" values for concentrations of
low magnitude or "less than" values for concentrations of high magnitude because they added
little significant information to the analyses. Conversely, if data from studies with only low
concentrations indicated a significant effect (suggesting the test material was highly toxic) or
studies with high concentrations only found an incomplete response for a chronic endpoint
(indicating low toxicity of the test material), those data did significantly enhance the
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understanding of PFOS toxicity. Thus, the decision rule was applied as follows: "greater than"
(>) high toxicity values and "less than" (<) low toxicity values were included (U.S. EPA 2013).
Data that met the quality objectives and test requirements were utilized quantitatively in deriving
these 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 PFOS.
Assessment Kndpoinls lor (lie
Aquatic ( omniunhv
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. LC50, 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 EC 10 could not be calculated for a
genus.
Note: only chronic exposures were consideredfor
derivation of the tissue-based criteria since PFOS is a
bioaccumulative chemical. These chronic tissue-based
criteria are expected to be protective of acute effects,
because acute effects were observed at much greater
concentrations than chronic effects.
LC50 = 50% Lethal Concentration
EC50 = 50% Effect Concentration
IC50 = 50% Inhibitory Concentration
NOEC = No-observed-effect-concentration
LOEC = Lowest-observed-effect-concentration
EC10 = 10% Effect Concentration
2.10,3,3 Summary of Independent Calculation of Toxicity Values
Where data were available, toxicity values, including LC50 and EC 10 values, were
independently calculated using data from the toxicity studies meeting the inclusion criteria
described above, via independent statistical analysis conducted by the EPA. Occasionally,
individual replicate-level data or treatment-level data were acquired from the study authors to
independently calculate toxicity values. All data were analyzed using the statistical software
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program R (version 3.6.2) and the associated dose-response curve (drc) package. The R drc
package has several models available for modeling a C-R 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 PFOS aquatic life criteria are
included in each quantitative study summary below and were utilized to derive these criteria for
aquatic life, where available (for the acute criterion see genus mean values in Table 3-9 and for
the chronic criterion, see genus mean values in Table 3-10).
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 are reviewed
to identify: 1) data from acceptable tests that meet data quality standards; and 2) whether the
acceptable data meet the minimum data requirements (MDRs) as outlined in the EPA's 1985
Guidelines (U.S. EPA 1985). The MDRs described in Section 2.10.1 were met for acute and
chronic freshwater criteria derivation. Acute and chronic MDRs for PFOS estuarine/marine
criteria derivation were not met. Consequently, the EPA used the available toxicity data and the
EPA's New Approach Methods (NAMs) to generate protective estuarine/marine benchmarks. A
minimal number of tests from acceptable studies of aquatic algae and vascular plants were also
available. The relative sensitivity of freshwater plants to PFOS exposures indicates plants are
less sensitive than aquatic vertebrates and invertebrates so plant criteria were not developed.
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2.11.2 Consideration for the Derivation of Tissue-Based Criteria following Chronic PFOS
Exposures
Chronic toxicity studies (both laboratory and field studies) were further screened to
ensure that they contained the relevant chronic PFOS exposure routes for aquatic organisms (i.e.,
dietary, maternal, or dietary and waterborne PFOS exposure), measurement of chronic effects,
and measurement of PFOS in tissue(s). The EPA considered deriving tissue-based criteria using
empirical toxicity tests with studies that exposed test organisms to PFOS via water, diet, and/or
maternal transfer and reported exposure concentrations based on measured tissue concentrations.
This approach generally corresponded with the 2016 Selenium Aquatic Life Freshwater
Criterion, which is the only other EPA 304(a) recommended aquatic life criterion with tissue-
based criteria (U.S. EPA 2016c). However, currently, the freshwater chronic PFOS toxicity
dataset with measured tissue concentrations is somewhat limited. There were 14 total chronic
aquatic life studies considered, six quantitative (three fish, one invertebrate, and two amphibian
studies) and eight qualitative studies (see Section 4.5). The quantitative studies provided data for
three of the eight MDRs. The qualitative studies provided supporting information for only one
additional MDR. Therefore, it was concluded that there are currently insufficient data to derive a
chronic tissue criterion using a GSD approach from empirical tissue data from toxicity studies.
Thus, the EPA used a Bioaccumulation Factor (BAF) approach for chronic tissue criteria
development.
2.11.3 Translation of Chronic Water Column Criterion to Tissue Criteria
To enable use of fish tissue measurements of PFOS in protecting designated uses, chronic
tissue criteria for PFOS were derived by translating the chronic freshwater water column
criterion (summarized in Section 2.11.1 above) into tissue criteria using bioaccumulation factors
(BAFs) and the following equation:
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Tissue Criteria = Chronic Water Column Criterion x BAF (Eq. 1)
The resulting tissue criteria correspond to the tissue type serving as the basis of the BAF
used in the equation.
2,11,3,1 Aquatic Life Bioaccumulation Factors (BAFs)
A BAF is determined from field measurements and is calculated using the equation:
BAF = ฃbiota_ ^Eq 2)
Cwater
Where:
CUota = PFOS concentration in organismal tissue(s)
Cwater = PFOS concentration in water where the organism was collected
Given that a BAF is determined from field measurements [as opposed to controlled
experiments designed to measure bioconcentration of PFOS using specific test guidelines;
(OECD 2001; U.S. EPA 2016c)], a BAF is an expression of all exposure routes, i.e., dietary,
water, maternal transfer, and contact with water and sediments via skin and ingestion. Depending
upon the tissue residue measurement, BAFs can be based upon residues in the whole organisms,
muscle, liver, or any other tissue.
The literature search for reporting on PFOS bioaccumulation was implemented by
developing a series of chemical-based search terms. These terms included chemical names and
Chemical Abstracts Service registry numbers (CASRN or CAS3), 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 U.S. EPA's ECOTOX database). The literature search yielded
numerous citations and the citation list was further refined by excluding citations on analytical
3 Chemical Abstracts Service registry number (CASRN or CAS) for PFOS is 1763-23-1.
66
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methods, human health, terrestrial organisms, bacteria, and where PFOS 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. Data from papers with appropriate
information were extracted into a PFOS dataset. The studies meeting these inclusion criteria
were also 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. For further details on BAFs
compilation and ranking, see .
Table 2-2 below outlines the screening criteria for study evaluation and ranking. Only
BAFs of high and medium quality were 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.
Table modified from Burkhard (2021).
Screening .Factor
High Quality
Medium Quality
Low Quality
Number of Water Samples
>3
2-3
1
Number of Organism
Samples1
>3
2-3
1
Temporal Coordination
Concurrent
collection
Within one year
Collection period > 1
year
Spatial Coordination
Co-located
collection
Within 1-2 km
Significantly
different locations
(> 2 km)
General Experimental Design
Mixed species tissues
samples
1 Organismal samples from the same species and tissue type.
67
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3 EFFECTS ANALYSIS FOR AQUATIC LIFE
3.1 Toxicity to Aquatic Life
All available, reliable studies relating to the acute and chronic toxicological effects of
PFOS on aquatic life were considered in the derivation of the national recommended PFOS
criteria. Data for possible inclusion in the PFOS criteria were obtained from published literature
reporting acute and chronic exposures of PFOS that were associated with mortality, survival,
growth, and reproduction. This set of published literature was identified by the EPA's public
ECOTOX 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. Studies were then further reviewed by the
EPA OW to determine test acceptability for use in the criteria derivation. Additional literature
searches were also conducted to ensure all available toxicity data were captured. The latest
search was conducted through March 2024.
3.1.1 Summary of PFOS Toxicity Studies Used to Derive the Aquatic Life Criteria
Quantitative data for acute PFOS toxicity were available for 29 freshwater species,
representing 20 genera and 17 families in five phyla, and six estuarine/marine species,
representing six genera and five families in four phyla. Chronic PFOS toxicity data were
available for 19 freshwater species, representing 17 genera and 15 families in four phyla, and
five estuarine/marine species, representing five genera and five families in three phyla (Table
3-1).
68
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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 Datasets for PFOS.
rrcshwalcr
MIJR
(imay
SMAY
(IMC Y
SMCY
Family Salmonidae in tile class
i
1
1
1
Osteichthyes
Second family in the class Osteichthyes,
preferably a commercially or
recreationally important warmwater
2
2
2
2
species
Third family in the phylum Chordata (may
be in the class Osteichthyes or may be an
amphibian, etc.)
5
10
3
4
Planktonic Crustacean
2
5
3
4
Benthic Crustacean
3
3
2
2
Insect
1
1
3
3
Family in a phylum other than Arthropoda
or Chordata (e.g., Rotifera, Annelida, or
5
6
2
2
Mollusca)
Family in any order of insect or any
phylum not already represented
1
1
1
1
Total
20
29
17
19
Saltwater'
MIJR
(imay
SMAY
(IMC A
SMCY
Family in the phylum Chordata
l
1
1
1
Family in the phylum Chordata
0
0
0
0
Either the Mysidae or Penaeidae family
2
2
1
1
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
1
1
1
1
Family in a phylum other than Chordata
0
0
1
1
Any other family
0
0
0
0
Total
6
6
5
5
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 Web-based Inter-species Correlation
Estimation (web-ICE) tool. These benchmarks are provided in Appendix L.
Below are the summarized studies that provided key acute and chronic freshwater
toxicity data with effect values that were used quantitatively in deriving the acute and chronic
freshwater criteria to protect aquatic life from harmful exposure to PFOS. Study summaries are
69
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also provided for the estuarine/marine toxicity data that could be used quantitatively to derive
acute and chronic estuarine/marine criteria if the MDRs were met.
Study summaries for the most sensitive taxa are grouped by acute or chronic exposure
and sorted by sensitivity to PFOS. Study data were summarized in tabular form in Appendix A
(freshwater acute studies), Appendix B (estuarine/marine acute studies), Appendix C (freshwater
chronic studies), and Appendix D (estuarine/marine chronic studies). Key acute and chronic
toxicity studies used qualitatively as supporting information are described in the Effects
Characterization (Section 4) below and corresponding data are listed in Appendix E, Appendix
F, Appendix G and Appendix H while the remaining, unused studies are listed in Appendix J.
Acute and chronic values were presented as reported by the study authors for each
individual study. The EPA independently calculated toxicity values if sufficient raw data were
available to conduct statistical analyses. All toxicity values, such as LCs, ECs, NOECs, LOECs,
and species- and genus-mean values, were given to four significant figures to prevent round-off
error in subsequent calculations, not to reflect the precision of the value. The author-reported
toxicity values and the EPA's independently-calculated values (where available) were included
for each study throughout the document (in the quantitative data study summaries and
appendices as applicable), and the specific value utilized to derive the criteria were identified
along with a justification. The EPA's independently-calculated toxicity values were used
preferentially, where available.
3.1.1.1 Summary of Acute PFOS Toxicity Studies Used to Derive the Freshwater Aquatic Life
Criterion
Acute toxicity data were available for all of the freshwater MDRs. Acceptable data on the
acute effects of PFOS in freshwater were available for a total of 29 species representing 20
genera and 17 families in five phyla (Appendix A: Acceptable Freshwater Acute PFOS Toxicity
70
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Studies). More specifically, quantitative data for acute PFOS toxicity were available for three
freshwater fish species (two of the eight MDRs), 16 freshwater invertebrate species (five of the
eight MDRs), and 10 freshwater amphibian species (one of the eight MDRs). Ranked genus
mean acute values (GMAVs) for PFOS in freshwater based on acute toxicity were identified in
Table 3-2 (4 most sensitive genera) and Table 3-3 (all genera) and plotted in Figure 3-1.
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
Genus
Species
CMAV
(mg/l.)
1
Neocloeon
Mayfly,
N. triangulifer
0.07617
2
Moina
Cladoceran,
M. microcopa and
M. micrura
3.075
3
Pimephales
Fathead minnow,
P. promelas
6.950
4
Oncorhynchus
Rainbow trout,
O. mykiss
7.515
3.1.1.1.1 Most Sensitive Freshwater Genus for Acute Toxicity: Neocloeon (mayfly)
Soucek et al. (2023) conducted a 96-hour acute toxicity test to determine the effects of
PFOS-K (PFOS potassium salt, CAS # 2795-39-3, 98% purity) in water on the parthenogenetic
mayfly, Neocloeon triangulifer. The test was performed under static, nonrenewal conditions
beginning with < 24 hour old nymphs. Mayflies were fed live diatom biofilm scraping beginning
on day 0. Feeding only occurred on day 0. 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 was able to independently calculate a 96-hour LCso
of 0.07617 mg/L (0.06546 - 0.08688 mg/L; 95% CI) for this study. The EPA's independently-
71
-------
calculated LC50 is in line with the author-reported LC50 of 0.08 mg/L. Therefore, the
independently-calculated LC50 of 0.07617 mg/L was used quantitatively to derive the freshwater
acute water column criterion for PFOS.
3.1.1.1.2 Second Most Sensitive Freshwater Genus for Acute Toxicity: Moina (cladoceran)
Ji et al. (2008) performed a 48-hour static, unmeasured acute test of PFOS (purity
unreported) with Moina macrocopa. The test followed the EPA's Methods for measuring the
acute toxicity of effluents and receiving waters to freshwater and marine organisms [U.S.
EPA/600/4-90/027F; (U.S. EPA 2002)]. The test involved four replicates of five neonates each in
five nominal test concentrations plus a negative control. Nominal concentrations were 0
(negative control), 6.25, 12.5, 25, 50 and 100 mg/L. Survival of organisms in the negative
control was not reported in the paper. However, raw data were obtained by the EPA from the
study authors and control survival was 100% in the acute test. The study authors reported a 48-
hour EC50 value of 17.95 mg/L forM macrocopa. The 48-hour EC50 value was independently-
calculated by the EPA as 17.20 mg/L. The independently-calculated acute toxicity value was
quantitatively used in the derivation of the freshwater acute water column criterion.
Razak et al. (2023) tested the acute toxicity of perfluorooctanesulfonate (PFOS, >98%
purity) on Moina micrura for 48 hours in a measured, static 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. Test water was filtered lake water. 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. The lethal effect concentrations (LC) were calculated using Probit analysis, and the 48-
72
-------
hour LC50 value of 549.6 |ig/L, or 0.5496 mg/L was determined to be acceptable for quantitative
use.
The geometric mean of the two SMAVs for Moina macrocopa (17.20 mg/L) and Moina
micrura (0.5496 mg/L) were used to calculate the GMAV of 3.075 mg/L for the genus Moina. If
the EPA excluded theM micrura SMAV on the basis of it being an overly sensitive outlier
(relative toM macrocopa and the overall acute data except for N. triangulifer) that would result
in the final PFOS acute criterion potentially being underprotective of untested sensitive
invertebrate or other taxa, considering that the available data serve as surrogate information for
the thousands of untested freshwater species. Conversely, excluding the M. macrocopa SMAV
on the basis of it being a tolerant outlier (relative to M micrura) would result in the Moina
GMAV being highly influenced by a single test/species with an LC50 that was relatively sensitive
(i.e., M. micrura SMAV = 0.5496 mg/L) compared to the overall acute data with the exception
of N. triangulifer. Averaging theM micrura andM moina SMAVs resulted in a GMAV (3.075
mg/L) that was the second most sensitive GMAV, still in the general range of the data overall.
3.1.1.1.3 Third Most Sensitive Freshwater Genus for Acute Toxicity: Pimephales (fathead
minnow)
Drottar and Krueger (2000c) evaluated the acute effects of PFOS-K (PFOS potassium
salt, CAS# 2795-39-3, Lot # 217 (T-6295) obtained from the 3M Company, 90.49% purity) on
juvenile fathead minnows (Pimephalespromelas) during a 96-hour measured, static study.
Researchers followed protocols from U.S. EPA Series 850, OPPTS 850.1075 and OECD
Guideline 203. All fish used in the test were from the same source and year class, and the total
length of the longest fish was no more than twice the length of the shortest. The authors reported
an LC50 of 9.5 mg/L PFOS. The EPA's independently-calculated 96-hour LC50 was 9.020 mg/L
and was used quantitatively to derive the freshwater acute water column criterion for PFOS. 3M
73
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Company (2000) provides the results of a 96-hour static, unmeasured acute toxicity test with the
fathead minnow and PFOS-Li (PFOS lithium salt, CAS # 29457-72-5). Fish were 79 days old at
test initiation with an average length of 2.1 cm and weight of 0.069 g. No mortality occurred in
the control treatment and 100% mortality was observed in the highest treatment (56 mg/L). The
study authors reported that the test sample containing 24.5% PFOS-Li exhibited a 96-hour LCso
of 19 mg/L, which equates to 4.655 mg/L as PFOS. The independently-calculated 96-hr LCso
value was 21.86 mg/L, which equates to 5.356 mg/L as PFOS, and was used quantitatively to
derive the freshwater acute water column criterion for PFOS.
The geometric mean of the two acute toxicity values described above for P. promelas
(9.020 and 5.356 mg/L) were used to calculate an SMAV and GMAV of 6.950 mg/L, which
represents the second most sensitive GMAV in the EPA's freshwater acute dataset for PFOS.
3.1.1.1.4 Fourth Most Sensitive Freshwater Genus for Acute Toxicity: Oncorhynchus (trout)
Sharpe et al. (2010) evaluated the acute effects of PFOS-K (potassium salt, CAS # 2795-
39-3, 98%) purity) to Oncorhynchus mykiss, rainbow trout, via a 96-hour renewal exposure with
measured concentrations (renewal was not stated in paper, but assumed based on other
information provided, including the test Guideline protocol that the authors cited as the protocol
that was used). There were limited details in the publication about the test protocol; however, it
was noted that the Organization for Economic Co-operation and Development (OECD)
Guideline 203 was followed, and the study authors did not identify any deviations from these test
guidelines. The EPA obtained clarification from the study authors on the experimental design
regarding the biomass loading rate, which was 1 to 1.5 g/L (based on four fish weighing a total
of 2 to 3 g per 2 L tank; personal communication with Greg Goss and Rainie Sharpe, March
2021). This biomass loading rate was slightly higher than that stated in OECD Guidelines of 0.8
74
-------
g/L (OECD 1992). The author-reported 96-hour LC50 for the study was 2.5 mg/L. The authors do
not specify if this concentration was nominal or measured. Given the clarifications regarding the
biomass loading, the LC50 from this study was used quantitatively to derive the freshwater acute
water column criterion for PFOS.
Palmer et al. (2002a) evaluated the acute effects of PFOS-K (potassium salt, identified
as FC-95 obtained from 3M Company) to rainbow trout via a 96-hour static exposure with
measured concentrations. The study author-reported 96-hour LC50 for the study was 22 mg/L.
The independently-calculated 96-hour LC50 value was 22.59 mg/L. The independently-calculated
LC50 was used quantitatively to derive the freshwater acute water column criterion for PFOS.
The geometric mean of the two toxicity values described above (2.5 and 22.59 mg/L),
was used to calculate the SMAV and GMAV of 7.515 mg/L for rainbow trout, 0. mykiss. The
GMAV of 7.515 mg/L is consistent with the acute rainbow trout studies cited in OECD's 2002
PFOS Hazard Assessment, from which the LC50 values for rainbow trout range from 7.8 to 22
mg/L (OECD 2002).
Table 3-3. Ranked Freshwater Genus Mean Acute Values.
Rank11
(/MAY
(ing/L PI-OS)
MI)U
(roup1'
Genus
Species
SMAV1'
(mg/L PIOS)
1
0.07617
F
Neocloeon
Mayfly,
Neocloeon triangulifer
0.07617
2
3.075
D
Moina
Cladoceran,
Moina macrocopa
17.20
Cladoceran,
Moina micrura
0.5496
3
6.950
B
Pimephales
Fathead minnow,
Pimephales promelas
6.950
4
7.515
A
Oncorhynchus
Rainbow trout,
Oncorhynchus mykiss
7.515
5
13.50
G
Ligumia
Black sandshell,
Ligumia recta
13.50
75
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Ksiuk"
Ci.MAV
(ing/L PI-OS)
MI)U
(roup1'
(0I1IIS
Species
S.MAV1'
(mg/l. PI-OS)
6
15.61
E
Neocaridina
Japanese swamp shrimp,
Neocaridina denticulata
15.61
7
15.99
C
Xenopus
African clawed frog,
Xenopus laevis
15.99
8
16.50
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
16.50
9
19.88
C
Hyla
Gray treefrog,
Hyla versicolor
19.88
10
22.48
G
Dugesia
Planaria,
Dugesia japonica
22.48
11
27.86
B
Danio
Zebrafish,
Danio rerio
27.86
12
43.15
D
Daphnia
Cladoceran,
Daphnia carinata
11.56
Cladoceran,
Daphnia magna
51.86
Cladoceran,
Daphnia pulicaria
134.0
13
47.40
C
Ambystoma
Jefferson salamander,
Ambystoma jeffersonianum
51.71
Small-mouthed salamander,
Ambystoma texanum
30.00
Eastern tiger salamander,
Ambystoma tigrinum
68.63
14
48.81
E
Pontastacus
Crayfish,
Pontastacus leptodactylus
48.81
15
56.49
C
Anaxyrus
American toad,
Anaxyrus americanus
56.49
16
59.87
E
Procambarus
Crayfish,
Procambarus fallax f virginalis
59.87
17
61.80
H
Brachionus
Rotifer,
Brachionus calyciflorus
61.80
18
64.35
G
Elliptio
Eastern elliptio,
Elliptio complanata
64.35
19
109.2
C
Lithobates
American bullfrog,
Lithobates catesbeiana
133.3
Green frog,
Lithobates clamitans
113.0
Northern leopard frog,
Lithobates pipiens
72.72
Wood frog,
Lithobates sylvatica
130.0
76
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Rank3
GMAV
(mg/L PFOS)
MDR
Group0
Genus
Species
SMAVb
(mg/L PFOS)
20
172.1
G
Physella
Bladder snail,
Physella acuta
183.0
Snail,
Physella heterostropha pomilia
161.8
a Ranked from the most sensitive to the most tolerant based on Genus Mean Acute Value.
b From Appendix A: Acceptable Freshwater Acute PFOS Toxicity Studies.
0 MDR Groups identified by list provided in Section 2.10.1 above.
1.0 T
0.9
0.8 --
0.7 -
0.6 -
a
_o
O
53
a
>
"3
E
3 ฐ.5
X
ง 0.4
C*
a
at
a
-------
available empirical data supplemented with toxicity values generated through the use of NAMs,
specifically through the use of the EPA Office of Research and Development's peer-reviewed
publicly-available Web-based Interspecies Correlation Estimation (WeblCE) tool (Raimondo et
al. 2010). These benchmarks are provided in Appendix L. Table 3-4 below shows the four most
sensitive acute estuarine/marine genera that could be used quantitatively to derive acute criteria
if the MDRs were met. Ranked GMAVs for saltwater organisms based on acceptable acute
toxicity values were identified in Table 3-5 and plotted in Figure 3-2.
Table 3-4. The Four Most Sensitive Acute Estuarine/Marine Genera.
Ranked Below from Most to Least Sensitive.
Uank
(ienus
Species
(/MAN
(ing/L PI OS)
1
Mytilus
Mediterranean mussel,
M. galloprovincialis1
1.1
2
Strongylocentrotus
Purple sea urchin,
S. purpuratus
1.7
3
Paracentrotus
Sea urchin,
P. lividus2
1.795
4
Americamysis
Mysid,
A. bahia
4.914
1 Not a resident species in North America, but other species in this genus are resident and commercially or
ecologically important species.
2 Not a resident species in North America, but other species in this family (Echinidae) are common ecotoxicity test
species that serves as a surrogate for untested urchin species residing in North America.
3.1.1.2.1 Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Mytilus (mussel)
The acute toxicity of perfluorooctane sulfonate (PFOS, purity not provided) on the
Mediterranean mussel, Mytilus galloprovincialis was evaluated by Fabbri et al. (2014). This
species is not resident to North America, but is a surrogate for North American mussel species,
including the widespread, commercially and ecologically important blue mussel, Mytilus edulis.
The test endpoint was the percentage of normal D-larvae in each well, including malformed
larvae and pre-D stages, at test termination (48 hours). The acceptability of test results was based
78
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on controls for a percentage of normal D-shell stage larvae, > 75% (ASTM 2004b). 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 a 50% reduction in malformations in the highest test concentration
(1 mg/L), 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
acceptable for quantitative use.
Hayman et al. (2021) report the results of a 48-hour static, measured test on the effects
of PFOS-K (PFOS potassium salt, CAS # 2795-39-3, 98% purity) onM galloprovincialis.
Authors noted that tests followed U.S. EPA (1995) and ASTM (2004a) protocols. Larvae were
enumerated for total number of larvae that were alive at the end of the test as well as number of
normally-developed D-shaped larvae. There were no significant differences between solvent
control and filtered seawater, suggesting no adverse effects of methanol. The author-reported 48-
hr ECso, based on normal development, was 1.1 mg/L. The EPA was not able to independently
calculate a 48-hour EC50 value as the curve fitted model did not result in a good fit. Therefore,
the author-reported EC50 of 1.1 mg/L was considered for quantitative use.
The two EC50 values from the two studies both indicated sensitivity of the Mediterranean
mussel to acute exposure of PFOS is above 1 mg/L. However, the EC50 forM galloprovincialis
from Fabbri et al. (2014) was unbounded while the EC50 from Hayman et al. (2021) was
definitive, and therefore the latter EC50 (1.1 mg/L) serves as the basis for the SMAV and GMAV
used to derive the acute estuarine/marine benchmark for PFOS.
79
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3.1.1.2.2 Second Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Strongylocentrotus
(sea urchin)
The Hayman et al. (2021) study also included the results of a 96-hour static, measured
test on the effects of PFOS-K (PFOS potassium salt, CAS # 2795-39-3, 98% purity) on the
purple sea urchin, Strongylocentrotuspurpuratus. Authors noted that tests followed U.S. EPA
(1995) and ASTM (2004a) protocols. At test termination (96 hours), the first 100 larvae were
enumerated and observed for normal development (4-arm pluteus stage). As with the other tests
in the study with different species, 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 1.7 mg/L. The EPA was not able to independently
calculate a 96-hour EC50 value as the curve fitted model did not result in a good fit. Therefore,
the author-reported EC50 of 1.7 mg/L mg/L was thus applied for quantitative use and was utilized
as the SMAV and GMAV to derive the acute estuarine/marine benchmark for PFOS.
3.1.1.2.3 Third Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Paracentrotus (sea
urchin)
A 72-hour static, unmeasured PFOS (purity not provided) toxicity test with the sea
urchin, Paracentrotus lividus (a non-North American species) was conducted by Gunduz et al.
(2013) The 72-hour EC50 based on normal development to the pluteus stage was 1.795 mg/L
PFOS and was acceptable for quantitative use; however, additional consideration needs to be
given to the use of this value in benchmark derivation due to the short test duration.
3.1.1.2.4 Fourth Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Americamysis
(mysid)
Along with the Mediterranean mussel and purple sea urchin, Hayman et al. (2021)
conducted a 96-hour static, measured test to determine the effects of PFOS-K on the mysid,
Americamysis bahia. Authors noted that tests followed U.S. EPA (1995) and ASTM (2004a)
80
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protocols. Only two of the sixty organisms (3.3%) were found dead in the controls at test
termination. The author-reported 96-hour LC50 is 5.1 mg/L PFOS-K. The independently-
calculated 96-hr LC50 value was 4.914 mg/L and is acceptable for quantitative use in the
derivation of the acute estuarine/marine benchmark for PFOS.
Table 3-5. Ranked Esi
tuarine/Marine Water Genus
Mean Acute Values.
Rank1
(/MAY
(ing/l. PI-OS)
MI)U
(roup-'
(ienus
Species
S.MAY2
(ing/l. PIOS)
1
1.1
D
Mytilus
Mussel,
Mytilus galloprovincialis
1.1
2
1.7
F
Strongylocentrotus
Purple sea urchin,
Strongylocentrotus
purpuratus
1.7
3
1.795
E
Paracentrotus
Sea urchin,
Paracentrotus lividus
1.795
4
4.914
C
Americamysis
Mysid,
Americamysis bahia
4.914
5
6.9
C
Siriella
Mysid,
Siriella armata
6.9
6
>15
A
Cyprinodon
Sheepshead minnow,
Cyprinodon variegatus
>15
1 Ranked from the most sensitive to the most tolerant based on Genus Mean Acute Value.
2 From Appendix B: Acceptable Estuarine/Marine Acute PFOS Toxicity Studies.
3 MDR Groups identified by list provided in Section 2.10.1 above.
81
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1.0
0.9
0.8
0.7
0.6
3
s
g 0.5
u,
s
53
X
a
u
0- 0.2
0.1
0.0
0.1
Cyprinodon (non-definitive, greater than value) ~
Siriella
Aniericaniysis
Paracentrotus
Strongylocentrotus
Mytilus
ฆ Invertebrate (Other)
Invertebrate (Mollusk)
~ Fish
1 10
Genus Mean Acute Value (mg/L PFOS)
100
Figure 3-2. Acceptable Estuarine/Marine GMAVs.
3.1.1.3 Summary of Chronic PFOS Toxicity Studies Used to Derive the Freshwater Aquatic
Life Criterion
Chronic toxicity data were available for all of the freshwater MDRs. Chronic PFOS
toxicity data were available for 19 freshwater species, representing 17 genera and 15 families in
four phyla. More specifically, quantitative data for acute PFOS toxicity were available for four
freshwater fish species, representing four genera and three families (two of the eight MDRs), 11
freshwater invertebrate species, representing 11 genera and ten families (five of the eight
MDRs), and three amphibian species, representing two genera in two families (one of the eight
MDRs). Ranked GMCVs for PFOS in freshwater based on chronic toxicity are listed in Table
3-6 (four most sensitive genera) and Table 3-7 (all genera) and plotted in Figure 3-3.
82
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Table 3-6. The Four Most Sensitive Genera Used in Calculating the Chronic Freshwater
Criterion.
Ranked Below from Most to Least Sensitive.
Uank
(ienus
Species
(;m( v
(ing/l. PI-OS)1
1
Neocloeon
Mayfly,
Neocloeon triangulifer
0.000226
2
Chironomus
Midge,
Chironomus dilutus
0.005198
3
Lampsilis
Fatmucket,
Lampsilis siliquoidea
0.01768
4
Enallagma
Blue damselfly,
Enallagma cyathigerum
0.03162
1 Other values were used in additional analyses supporting the criterion calculation to examine the effects of less
certain toxicity studies and non-resident species on the chronic freshwater criterion. See Section 4.1 below for more
details.
3.1.1.3.1 Most Sensitive Freshwater Genus for Chronic Toxicity: Neocloeon (mayfly)
Soucek et al. (2023) conducted a chronic life-cycle test to determine the effects of PFOS-
K (PFOS potassium salt, CAS # 2795-39-3, 98% purity) on the parthenogenetic mayfly,
Neocloeon triangulifer. The test was performed under renewal conditions over 27 days
beginning with < 24 hour old nymphs. There were sixteen (with one mayfly per replicate)
replicates per test concentration and control. Replicates one through eight were destructively
sampled on day 14 and replicates nine through sixteen continued until the end of the test (when
all mayflies either molted into imago stage or died). The endpoints that were evaluated included
survival for all replicates, 14-d length and calculated dry weight (using a previously published
body dry weight equation; Besser et al. 2021) for replicates 1 through 8, and percent survival to
pre-emergent nymph (PEN) stage, number of days until PEN stage, percent emergence (to imago
stage), and pre-egg laying live weight of imago for replicates 9 through 16. Percent survival in
the control after 14 days was 100%. Percent survival of mayflies after 14 days in the remaining
seven test concentrations ranged from 79 to 100%. The most sensitive endpoint was 14-day dry
weight. The study authors reported three different 14-day dry weight ECio values that were
83
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calculated using various point-estimation approaches. The author-reported 14-day dry weight
ECio values produced by the various approaches were relatively similar to one another, ranging
from 0.000226 (using TRAP [2 parameter, threshold sigmoidal curve]) to 0.000272 mg/L (using
log-linear regression, controls excluded). The EPA was not able to fit a reliable model with
significant model parameters to the 14-day dry weight C-R dataset and, therefore, relied on the
author-reported ECio of 0.000226 mg/L (based on TRAP) as the primary effect concentration.
The EPA selected the TRAP-based EC io preferentially over the ECio values based on the two
other point estimation approaches (i.e., log-linear regression with and without controls) because
the TRAP-based model (1) considered control responses; (2) was more fundamentally consistent
with the maximum likelihood regression approaches used by the EPA to assess C-R datasets
throughout this PFOS aquatic life AWQC document, and; (3) relied on replicate-level data,
which the EPA used preferentially over treatment-mean data in assessing C-R datasets
throughout the PFOS aquatic life criteria derivation process. The author-reported ECio of
0.000226 mg/L (TRAP-based) was used quantitatively to derive the freshwater chronic water
column criterion for PFOS.
3.1.1.3.2 Second Most Sensitive Freshwater Genus for Chronic Toxicity: Chironomus (midge)
MacDonald et al. (2004) conducted sub-chronic, partial-life cycle tests on larva and
chronic life-cycle tests to determine the effects of PFOS-K (PFOS potassium salt, 95% purity) on
the midge, Chironomus dilutus (formally known as Chironomus tentans). The test was
performed under renewal conditions over 10 days for the larval test and greater than 50 days for
the life-cycle test. The tests followed the general guidance given by EPA-600-R99-064 (U.S.
EPA 2000b) and ASTM E 1706-00 (ASTM 2002).
84
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The author-reported 10-day growth and survival ECios for the study were 0.0492 and
0.1079 mg/L, respectively. The study authors also reported NOECs of 0.0491 mg/L, LOECs of
0.0962 mg/L, and MATCs of 0.0687 mg/L for both endpoints. The author-reported 20-day ECios
for growth, survival, and total emergence were 0.0882, 0.0864, and 0.0893 mg/L, respectively.
The study authors also reported NOECs of 0.0217 mg/L for growth and survival and < 0.0023
mg/L for emergence, LOECs of 0.0949 mg/L for growth and survival and 0.0271 mg/L for
emergence, and MATCs of 0.0454 mg/L for growth and survival and 0.0071 mg/L for
emergence. It is noted here that the paper reported contrasting NOECs for 20-day survival. The
text in the paper stated that the NOEC was 0.0271 mg/L and Table 2 of the paper provided a
value of 0.0949 mg/L. The EPA assumed the NOEC in Table 2 of the paper was not correct and
that 0.0217 mg/L was the correct NOEC based on the data presented in Figure 3 A of the paper,
since the EPA was unable to gain confirmation from the study authors. This assumption was
applied to the summary of the study results presented in this PFOS aquatic life AWQC
document. The EPA was able to independently calculate an ECio for 10-day growth of 0.05896
mg/L for the study. The independently-calculated 10-day ECio value for growth of the midge
was used quantitatively to derive the freshwater chronic water column criterion for PFOS.
McCarthy et al. (2021) conducted two chronic toxicity tests with PFOS (98% purity) on
the midge, C. dilutus, a 10-day and a 20-day exposure, following standard protocols from U.S.
EPA (2000b) and ASTM (2002) with slight modifications. The 10-day exposure was considered
a range finding test, with concentrations spaced by ~100x and only mortality measured, whereas
the 20-day exposure measured both survival and growth and was termed an "abbreviated full life
cycle test" by the study authors. The 20-day exposure is less than the recommended 50 - 65 day
full-life cycle method outlined in U.S. EPA (2000b) and used in MacDonald et al. (2004), and
85
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since exposures of midges started on day two or four, the actual exposure duration is only 16 or
19 days long. The most sensitive endpoint was survival from the "abbreviated full life cycle test"
with an author-reported 16-day ECio of 0.00136 mg/L PFOS. Additionally, the study authors
reported ECios of 0.00162 and 0.00323 mg/L PFOS for growth as mean biomass and mean
weight, respectively. The EPA was unable to independently calculate ECios for survival and
mean weight. However, the EPA was able to independently calculate an ECio value for mean
biomass of 0.001588 mg/L PFOS. The independently-calculated 16-day ECio for mean biomass
was used quantitatively to derive the freshwater chronic water column criterion for PFOS.
Krupa et al. (2022) conducted a partial-life cycle chronic toxicity test with the midge, C.
dilutus, and PFOS-K (perfluorooctanesulfonate potassium salt, > 98% purity, CAS No. 2795-39-
3. The larvae were exposed to PFOS for 16 days. The measured exposure concentrations were <
the limit of detection (LOD), 0.001, 0.0025, 0.004, 0.0075, 0.016 and 0.03 mg/L. Attest
termination, larval survival was assessed, and ash-free dry weight (AFDW) was determined
following ASTM (2019). The AFDW of five groups of 12 larvae was measured at test initiation
to establish a baseline for growth. The author-reported 16-day growth ECio was 0.0015 mg/L
PFOS-K. The EPA was unable to fit a reliable model for any of the chronic endpoints from this
test. Therefore, the author-reported ECio value of 0.0015 mg/L for growth presented in the paper
was used quantitatively to derive the freshwater chronic water column criterion for PFOS.
The most sensitive endpoints from the two toxicity studies with C. dilutus that could be
independently-calculated (see details in Appendix C.2.2) were for 10-day growth with an ECio of
0.05896 mg/L (MacDonald et al. 2004) and 16-day mean biomass with an ECio of 0.001588
mg/L (McCarthy et al. 2021). The EPA could not independently calculate the 16-day growth
ECio of 0.0015 mg/L (Krupa et al. 2022). Although over an order of magnitude difference exists
86
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between MacDonald et al. (2004) and the other two studies, all three ECios were used
quantitatively to derive the chronic aquatic life criterion with a SMCV and GMCV equal to the
geometric mean of the three values or 0.005198 mg/L.
As mentioned in the Bots et al. (2010) summary (Section 3.1.1.3.4), the observed effects
of PFOS on aquatic insects appeared to be consistent across the available data for chironomids
and odonates. However, Bots et al. (2010) did not measure the effects of PFOS on nymph growth
and therefore, the observed effects in that study cannot be compared with the results of
MacDonald et al. (2004), McCarthy et al. (2021), and Krupa et al. (2022). The remainder of the
toxicity values available for aquatic insects were used as supporting information to corroborate
the toxicity value used to derive the freshwater chronic criterion and to better understand the
effects of PFOS on aquatic insects in general. No other quantitative toxicity values were
available for this species or genus.
3.1.1.3.3 Third Most Sensitive Freshwater Genus for Chronic Toxicity: Lampsilis (mussel)
Hazelton (2013); Hazelton et al. (2012) conducted a test of the long-term effects of
PFOS (acid form, > 98% purity) on glochidia and juvenile life stages from the mussel Lampsilis
siliquoidea using a unique experimental design for which standard methods have not been
established. The test exposed brooding glochidia (in marsupia) for 36 days followed by a 24-
hour exposure of free glochidia in a factorial design. As such, the free glochidia from the control
group of the marsupia exposure were divided between a control and the two PFOS treatments
and the PFOS treatments were split into control and the same PFOS treatment group as the
marsupia exposure. This factorial design allowed for the comparison of PFOS effects in two
different life-stages. See Appendix C.2.3 for additional details on the experimental design and
considerations for the utilization of this study in the criterion derivation.
87
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The data presented in the paper for metamorphosis success were considered for
quantitative use in the derivation of the chronic criterion for PFOS (see Appendix C.2.3). The
author-reported NOEC was 0.0045 mg/L and LOEC was 0.0695 mg/L. The reduction in
metamorphosis success at the LOEC was estimated to be 35.4%. However, this was not a
definitive test in that both the study design (which only included two treatment groups) and
level of data presented (which are only presented graphically in Figure 2 of the paper) in the
publication lack the details needed to fully understand the effects of chronic PFOS exposures to
the glochidia and juvenile life stages of L. siliquoidea. Additionally, as there were only two
PFOS treatment groups and the gap in these exposure concentrations is large (about 15-fold),
the EPA was not able to fit a curve to estimate an ECio in a manner similar to the other toxicity
studies used to derive this criterion. Instead, both the use of an MATC and an estimated ECio
were considered for the chronic value. An ECio was estimated by assuming the 0.0695 mg/L
treatment represents an EC35.4 and estimating the ECio using the exposure response slope from
another PFOS toxicity study focused on another mussel species (Perna viridis). Specifically, the
chronic exposure of Perna viridis reported by Liu et al. (2013), which is summarized in Section
3.1.1.4.1, was used to derive a EC10/EC35.4 ratio from that study, which was 0.0033/0.0186, or
0.1770. Applying this ratio to Hazelton et al. (2012) yields an estimated ECio of 0.0123 mg/L.
Given the similarity between this ECio and the author-reported MATC for Hazelton et al.
(2012), the MATC of 0.01768 mg/L was used to derive the chronic criterion for PFOS. The
EPA hopes to further refine the estimated ECio by obtaining the treatment level data from the
study authors and exploring additional exposure response slopes from the PFOS dataset. No
other quantitative chronic toxicity values were available for this species or genus; therefore, the
88
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MATC of 0.01768 mg/L was used quantitatively to derive the freshwater chronic water column
criterion for PFOS.
3.1.1.3.4 Fourth Most Sensitive Freshwater Genus for Chronic Toxicity: Enallagma (damselfly)
Bots et al. (2010) conducted a 320-day partial life-cycle study under renewal test
conditions to examine the effects of PFOS (tetraethylammonium salt, 98% purity) on the
damselfly Enallagma cyathigerum. Approximately 40% of the nymphs in the control treatment
died during the first 60 days and similar mortality levels were observed in the other treatments.
However, it appeared that control survival plateaued between 60 and 200 days, with 82.57% of
the remaining nymphs in the control treatment surviving during this time, indicating that survival
settled out during this phase of the experiment. The initial drop in nymph survival could likely be
attributed to the handling of the test organisms between the various phases of the experiment.
This would explain the observed plateau between 60 and 200 days, as the nymphs were not
handled during this time. The observed control survival in this test was consistent with other
odonate tests and excessive mortality of nymphs is typically expected within the first 200 days
given the difficulty in maintaining odonates in a lab setting (Abbott and Svensson 2007; Rice
2008). Therefore, the observed control survival for this study was considered within the
acceptable range for this species up to the 200-day exposure duration. Further, the control
survival observed in this study was largely consistent with the toxicity testing guidelines for
chironomids (requiring 70% control survival: ASTM 2002; U.S. EPA 2000b), which are
currently the only test guidelines for an emergent aquatic insect as there currently is no test
guideline for odonates. Therefore, considerations regarding the use of these data for chronic
criterion derivation were based on best scientific judgement and were restricted to the first 200
days of the experiment.
89
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The observed effects of PFOS on E. cyathigerum reported in the paper by the study
authors include decreased survival over the exposure duration and decreased metamorphosis
success. The MATC based on metamorphic success was less sensitive than for survival. As such,
the MATC author-reported value of 0.03162 mg/L for nymph survival was considered
quantitatively in the derivation of the aquatic life criteria. The remainder of the toxicity values
were used as supporting information to corroborate the toxicity value used to derive the
freshwater chronic criterion and to better understand the effects of PFOS on aquatic insects. As
no other quantitative toxicity values were available for this species or genus, the author-reported
MATC of 0.03162 mg/L served directly as the SMCV/GMCV. Additionally, the EPA ran
additional analyses with some of the other toxicity values for E. cyathigerum to understand the
influence of this study on the overall chronic criterion (see Section 4.1 below).
Table 3-7. Ranked Freshwater Genus Mean C
ironic Values.
Uank11
(i.MCV
(ing/l. PI-OS)
MI)U
(roup1'
Genus
Species
S.MCV1'
(ing/l. PI-OS)
1
0.000226
F
Neocloeon
Mayfly,
Neocloeon triangulifer
0.000226
2
0.005198
F
Chironomus
Midge,
Chironomus dilutus
0.005198
3
0.01768
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
0.01768
4
0.03162
F
Enallagma
Blue damselfly,
Enallagma cyathigerum
0.03162
5
0.03217
B
Danio
Zebrafish,
Danio rerio
0.03217
6
0.06519
D
Daphnia
Cladoceran,
Daphnia carinata
0.003162
Cladoceran,
Daphnia magna
1.344
7
>0.1
A
Salmo
Atlantic salmon,
Salmo salar
>0.1
8
0.1098
B
Pimephales
Fathead minnow,
Pimephales promelas
0.1098
90
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(i.MCY
MI)U
S.MCV1'
Ksiuk"
(mg/l. PI OS)
(roup1'
(CI1IIS
Species
(uiii/l, PI OS)
9
0.167
E
Procambarus
Crayfish,
Procambarus fallax f virginalis
0.167
10
0.1789
D
Moina
Cladoceran,
Moina macrocopa
0.1789
11
0.25
H
Brachionus
Rotifer,
Brachionus calyciflorus
0.25
12
0.5997
C
Xiphophorus
Swordtail fish,
Xiphophorus helleri
0.5997
13
0.7507
C
Xenopus
African clawed frog,
Xenopus laevis
>0.7160
Clawed frog,
Xenopus tropicalis
0.7871
14
1.316
C
Lithobates
Northern leopard frog,
Lithobates pipiens
1.316
15
2.899
E
Hyalella
Amphipod,
Hyalella azteca
2.899
16
8.527
G
Physella
Snail,
Physella heterostropha pomilia
8.527
17
8.640
D
Ceriodaphnia
Cladoceran,
Ceriodaphnia dubia
8.640
a Ranked from the most sensitive to the most tolerant based on Genus Mean Chronic Value.
b From Appendix C: Acceptable Freshwater Chronic PFOS Toxicity Studies
0 MDR Groups identified by list provided in Section 2.10.1 above.
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1.0 T
0.9 --
a
ฃ
O
C3
-------
Table 3-8. The Four Ranked Estuarine/Marine Genus Mean Chronic Values.
Ranked Below from Most to Least Sensitive.
Uank
Genus
Species
G\l( V
(mg/l.
PI OS)
Coniinenls
1
Perna
Asian green
mussel,
Perna viridis
0.0033
Not a resident species in North America
2
Austrochiltonia
Amphipod,
Austrochiltonia
subtenuis
0.01118
Not a resident species in North America,
but other species in this Order
(Amphipoda) are common ecotoxicity test
species that serves as a surrogate for
untested amphipod species residing in
North America.
3
Americamysis
Mysid,
Americamysis
bahia
0.3708
North American resident species
4
Tigriopus
Copepod,
Tigriopus
japonicus
0.7071
Not a resident species in North America,
but other species in this genus (Tigriopus)
are common ecotoxicity test species that
serves as a surrogate for untested copepod
species residing in North America.
3.1.1.4.1 Most Sensitive Estuarine/Marine Genus: Perna (mussel)
Liu et al. (2013) evaluated the chronic effects of PFOS-K (PFOS potassium salt, CAS#
2795-39-3, 98% purity) on green mussels, Perna viridis, via a 7-day measured, static-renewal
study. Mussels were exposed at a salinity of 25 ppt (artificial seawater) and a temperature of
25ฐC. PFOS concentrations were verified through water and muscle tissue samples via liquid
chromatography-tandem mass spectrometry. Weights and lengths were determined on days zero
and seven. An author-reported NOEC of 0.0096 mg/L and a LOEC of 0.106 mg/L was
determined for the growth condition index. The EPA's independently-calculated ECio for growth
condition index is 0.0033 mg/L. This ECio is used quantitatively to represent the chronic
sensitivity of this species to PFOS exposure in the marine/estuarine aquatic life dataset.
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3.1.1.4.2 Second Most Sensitive Estuarine/Marine Genus: Austrochiltonia (amphipod)
Sinclair et al. (2022) tested perfluorooctane sulfonic acid (PFOS, purity not reported) on
amphipods (Austrochiltonia subtenuis) in a 7-day unmeasured, static experiment. The 7-day
experiment consisted of five controls, one solvent control (methanol 0.25 mL/L), and five
nominal PFOS concentrations (0.04, 0.2, 1.0, 5.0, 25 |ig/L). Test vessels were 600 mL beakers
with 400 mL of test material and a 2x2 cm gauze substrate. Each test vessel included 20
amphipods, and all test material was dissolved in modified standard artificial media. The NOEC
and LOEC for 7-day survival were 0.005 mg/L and 0.025 mg/L, respectively, and the resulting
MATC of 0.01118 mg/L was determined to be acceptable for quantitative use.
3.1.1.4.3 Third Most Sensitive Estuarine/Marine Genus: Americamysis (mysid)
Drottar and Krueger (2000h) reported the results of a 35-day flow-through, measured
life-cycle test of PFOS-K (potassium salt, 90.49% purity) with Americamysis bahia (formerly
Mysidopsis bahia). The 35-day NOEC (reproduction and growth) was 0.25 mg/L, and the
corresponding 35-day LOEC was 0.55 mg/L. An independently-calculated ECio could not be
defined at this time given the level of data that was presented in the paper (Appendix D). The
calculated MATC for the test was 0.3708 mg/L. This chronic value was considered acceptable
for quantitative use despite the control survival of 78% because it was only slightly below the
80%) survival threshold, and because there were no other deficiencies in the study design.
3.1.1.4.4 Fourth Most Sensitive Estuarine/Marine Genus: Tigriopus (copepod)
A 20-day renewal, unmeasured full life-cycle test with PFOS (analytical grade) was
conducted on the copepod, Tigriopus japonicus (non-North American species) by Han et al.
(2015). The development of the copepod's growth from nauplii to copepodite and from nauplii to
adults was determined daily based on morphological characteristics. Results were presented as
94
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the number of days needed to reach the normal development stages. The highest test
concentration (1 mg/L PFOS) significantly increased the amount of time it took the copepods to
reach the development stage. Additionally, the authors assessed the reproduction of the copepods
by counting the nauplii produced by eight ovigerous females for 10 days in each well exposed to
PFOS. However, it was unclear if this was a subsampling of the organisms used in the 20-day
developmental test or if an independent assay with adult females was run. Results are presented
graphically as daily nauplii production/individual. There was a statistically significant decrease
in production (daily nauplii production/individual) in the 0.25, 0.5 and 1.0 mg/L PFOS
concentrations compared to the control. Production was decreased by approximately 50% in the
highest concentration (1 mg/L). The 20-day MATC based on time to reach development stage
was 0.7071 mg/L and was acceptable for quantitative use in the marine/estuarine chronic aquatic
life dataset.
95
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1.0 T
0.9 -
I ฐ'8
*-*
U
ฃ 0.7
4J
>
]g 0.6 4
s
ฃ
s 0.5
0.4 -
a
ca
Cฃ
0.2
0.1 -
0.0
0.001
Oryzias (non-definitive, greater than value) ~
Tigriopus ~
Austrochiltonia ~
Pema
Ainericamysis ~
ฆ Invertebrate (Mollusk)
~ Invertebrate (Other)
~ Fish
0.01 0.1 1
Genus Mean Chronic Value (mg/L PFOS)
10
Figure 3-4. Acceptable Estuarine/Marine GMCVs.
3.2 Derivation of the PFOS Aquatic Life Criteria
3.2.1 Derivation of Water Column Criteria for Direct Aqueous Exposure
3.2.1.1 Derivation of Acute Water Column Criterion for Freshwater
The PFOS acute dataset for freshwater based on direct aqueous exposures contained 20
genera representing all eight MDRs. GMAVs for the 20 freshwater genera are provided in Table
3-3, and the four most sensitive genera were within a factor of-100 of each other. The lowest
acute value for the mayfly, Neocloeon tricmgirtifer, is 40 times lower than the second most
sensitive genus (Figure 3-5). The freshwater FAV, the 5th percentile of the genus sensitivity
distribution, for PFOS was 0.1413 mg/L, and was calculated using the procedures described in
the 1985 Guidelines (U.S. EPA 1985). The FAV was lower than all of the GMAVs of the tested
species, except the mayfly, Neocloeon tricmgirtifer. The FAV was then divided by two to obtain a
96
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concentration yielding minimal effects (see Section 2.9). The FAV/2, which is the acute
freshwater criterion (or criterion maximum concentration, CMC), was 0.071 mg/L PFOS
(rounded to two significant figures), and is expected to be protective of approximately 95% of
freshwater genera potentially exposed to PFOS via direct aqueous exposure, under short-term
duration conditions of one-hour, when the criterion magnitude is not exceeded more than once in
three years on average (Table 3-9).
Table 3-9. Freshwater Final Acute Value and Criterion Maximum Concentration.
Calculated l-'rcshwater l\W
based on 4 lowest values: Total Number of (iMAVs in Dalasel 2d
(/MAY
Uank
(ienus
(mg/l.)
ln((;.\l.\Y)
ln((;MA\ )2
P=U/(N+I)
sqrl(P)
1
Xcol/ocoii
0.07017
-2.57
0.03
0.048
0.218
2
Moina
3.075
1.12
1.26
0.095
0.309
3
Pimephales
6.950
1.94
3.76
0.143
0.378
4
Oncorhynchus
7.515
2.02
4.07
0.190
0.436
ฃ (Sum):
2.50
15.72
0.48
1.34
S2 =
534.58
S = slope
L =
-7.127
L = X-axis intercept
A =
-1.957
A = InFAV
FAV =
0.1413
P = cumulative probability
CMC =
0.071 mg/L PFOS
(rounded to two significant figures)
97
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1.0
0.9 -
a
_o
+ฆ>
U
S 0.7
3 0.6 4
"3
S
3 0.5
a
a
cs
Pi
a<
0.4 -
0.3 -
a
0)
t
u
0.2
0.1 4
0.0
0.01
ฆ Invertebrate (Mollusk)
~ Invertebrate (Other)
ฆ Physella
~ Lithobates
Insect
ฆ Elliptic
~ Fish
~ Brachionus
~ Amphibian
~ Procambarus
CMC
~ Anaxyrus
~ Pontastacus
~ Ambystoma
~ Daphnia
~ Danio
~ Dugesia
~ Hyla
ฆ Lampsilis
~ Xenopus
~ Neocaridina
ฆ Ligumia
~ Oncorhynchus
~ Pimephales
~ Moina
ft Neocloeon
0.1 1 10
Genus Mean Acute Value (mg/L PFOS)
100
1000
Figure 3-5. Ranked Freshwater Acute PFOS GMAVs Used Quantitatively to Derive the
Criterion.
3.2.1.2 Derivation of Acute Water Column Criterion for Estuarine/Marine Water
The estuarine/marine acute dataset for PFOS contained six genera (Table 3-5 and
Appendix B) representing only five of the eight taxonomic MDR groups. The missing MDR
groups included one family in the phylum Chordata, a family in a phylum other than Chordata,
and another family not already represented. The GMAVs of the four most sensitive definitive
estuarine/marine genera were within a factor of 4.5 of each other (Table 3-5).
Because data were available for only five of eight MDRs, the EPA developed an
estuarine/marine acute benchmark using the available empirical data supplemented with toxicity
values generated through the use of NAMs, specifically through the use of the EPA Office of
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Research and Development's peer-reviewed publicly-available web-ICE tool (Raimondo et al.
2010). This benchmark is provided in Appendix L.
3.2.1.3 Derivation of Chronic Water Column Criterion for Freshwater
The PFOS chronic dataset based on direct aqueous exposures contained data for all eight
MDRs, thus the Final Chronic Value (FCV) can be calculated directly without the use of an
ACR. There were GMCVs for 17 freshwater genera (Table 3-7). The four most sensitive genera
were within a factor of 140 of each other. The lowest chronic value for the mayfly, Neocloeon
triangulifer, is over an order of magnitude lower than the second most sensitive genus (Figure
3-6). The freshwater FCV for PFOS of 0.0002491 mg/L was calculated using the procedures
described in the 1985 Guidelines (U.S. EPA 1985). The FCV is the 5th percentile of the genus
sensitivity distribution and is intended to be protective of 95 percent of the genera. The FCV was
lower than all of the GMCVs of the tested species, except the mayfly, Neocloeon triangulifer.
Unlike the FAV, the FCV was not divided by two, as it already represents a low effect level, and
was equal to the water column chronic criterion (or criterion continuous concentration, CCC;
Table 3-10). The freshwater CCC had a magnitude 0.00025 mg/L PFOS (rounded to two
significant figures), or 0.25 |ig/L, and is expected to be protective of 95% of freshwater genera
potentially exposed to PFOS through direct aqueous exposure under long term conditions of four
days, if not exceeded more than once every three years on average (Table 3-10).
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Table 3-10. Freshwater Final Chronic Value and Criterion Continuous Concentration.
Calculated Freshwater FCV based on 4 lowest values: Total Number of GMCVs in Dataset =17
GMCV
Rank
Genus
(mg/L)
ln(GMCV)
ln(GMCV)2
P=R/(N+1)
sqrt(P)
1
Neocloeon
0.000226
-8.39
70.48
0.056
0.236
2
Chironomus
0.005198
-5.26
27.66
0.111
0.333
3
Lampsilis
0.01768
-4.04
16.28
0.167
0.408
4
Enallagma
0.03162
-3.45
11.93
0.222
0.471
L (Sum):
-21.14
126.35
0.56
1.45
S2 =
472.36
S = slope
L =
-13.157
L = X-axis intercept
A =
-8.297
A = InFCV
FCV =
0.0002491
P = cumulative probability
ccc =
0.00025 mg/L PFOS
(rounded to two significant figures)
a
o
+ฆ>
0
1
QJ
>
9
s
3
o
a
S3
<&
a
a
-------
3,2,1,4 Deriving A Protective Duration Component of the Chronic Water Column-Based
Criterion
Effects to sensitive life stages was a primary reason why the 1985 Guidelines (U.S. EPA
1985) recommended a 4-day duration for most water column-based criteria. U.S. EPA (1985)
states, "An averaging period offour days seems appropriate for use with the CCC for two
reasons. With one of the two reasons specify being, "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."
The SMCV for C. dilutus, representing the Chironomus GMCV (second most sensitive
genus) is based on the EPA's independently-calculated ECio of 0.05896 mg/L for a 10-day larval
growth endpoint by MacDonald et al. (2004), an ECio of 0.001588 mg/L for a 16-day larval
mean biomass endpoint by McCarthy et al. (2021), and an ECio of 0.0015 mg/L for a 14-day
larval growth endpoint by Krupa et al. (2022). The ECio for a 10-day larval growth by
MacDonald et al. (2004) is slightly higher than the author-reported ECio for this effect in the
study. The author-reported ECios for the 20-day test by MacDonald et al. (2004) were higher
than those for the 10-day test, which is an atypical outcome, and were not used for criteria
derivation. Consequently, there was no clear influence of exposure time on the effects of PFOS
on this species.
The SMCV for L. siliquoidea, representing the Lampsilis GMCV (third most sensitive
genus) is based on a 36-day study by Hazelton (2013); Hazelton et al. (2012) using glochidia and
juvenile life stages. The test exposed brooding glochidia (in marsupia) for 36 days followed by a
24-hour exposure of free glochidia. The 24-hour free glochidia exposure consisted of a factorial
design, such that free glochidia from the control group of the marsupia exposure were divided
between a control and the two PFOS treatments and the PFOS treatments were split into control
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and the same PFOS treatment group as the marsupia exposure. This factorial design allowed for
the comparison of PFOS effects in two different life-stages.
Given the limitations of time points that could be discerned by the test, it appeared that
for reduced viability and or metamorphosis success of free glochidia to occur at concentrations
near the chronic value for the test (0.01768 mg/L), the test's 36-day exposure period would also
be needed. For example, the study authors determined that the in-marsupia (36-day) exposure
held the greatest weight of evidence and explained 78% of the variability in the glochidia
viability (AIC = 22843, wi = 0.78) and 83% of the metamorphosis success (AIC = 21955, wi =
0.83). As a result, this species appears to be protected by the chronic 4-day duration component
of the water column criterion. It should also be noted, brief PFOS exposures at elevated
concentrations consistent with the magnitude and four-day duration of the chronic criterion are
not expected to cause effects to free swimming glochidia based on the 24-hour acute toxicity
data for glochidia.
The SMC V for Enallagma cyathigerum, representing the Enallagma GMCV (fourth most
sensitive genus) are based on a 320-day partial life-cycle test by Bots et al. (2010). Only a single
treatment, 0.1 mg/L, showed partial effects. The treatment 10X higher (i.e., 1 mg/L) yielded
100%) mortality within 20 days. The treatment 10X lower (0.01 mg/L) showed no effects over
the entire test. The authors provided the time course of mortality throughout the entire test. At
0.1 mg/L a marked reduction in survival began at 130 days, and reached zero survival at 250
days, suggesting a relatively long time-to-effect. Because 0.1 mg/L is more than 3-fold higher
than the estimated chronic value for the test, 0.03162 mg/L, it is postulated that the time course
of mortality observed at 0.1 mg/L would be substantially faster than what would be expected to
occur at 0.03162 mg/L. Given the relatively slow manifestation of chronic effects observed in
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this study, this species appears to be protected by the chronic 4-day duration component of the
water column criterion.
No chronic PFOS toxicity tests specifically 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 PFOS exposures, or provided insight into the potential for latent toxicity. However,
chronic tests, including life-cycle tests with relatively sensitive species, suggested chronic effects
may occur at durations shorter than those of standard chronic toxicity tests (e.g., 28-day ELS
tests) and a chronic 4-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
PFOS chronic water column criterion at four days to reflect the chronic criterion duration
recommended in the 1985 Guidelines. This 4-day duration component of the chronic water
column is also consistent with U.S. EPA (1991), which considered the default 4-day chronic
averaging period as "the shortest duration in which chronic effects are sometimes observedfor
certain species and toxicants'', and concludes that 4-day averaging "should be fully protective
even for the fastest acting toxicants."
3,2,1.5 Derivation of Chronic Water Column Criterion for Estuarine/Marine Water
The estuarine/marine chronic dataset for PFOS contained GMCVs for five genera.
GMCVs for five estuarine/marine genera are summarized in Section 3.1.1.4 and shown in Figure
3-4. The eight-family taxonomic (MDR) requirement was not met by the chronic dataset, as
acceptable chronic studies for species representing three MDR groups are not available (one
family in the phylum Chordata, a family in a phylum other than Arthropoda or Chordata, and
another family not already represented). The 1985 Guidelines allow the use of a Final Acute-
Chronic Ratio (FACR) to convert a FAV to an FCV (i.e., FAV/FACR = FCV), which is
equivalent to a CCC. However, since an FAV could not be calculated with the available data, an
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FCV also could not be calculated. Consequently, the EPA could not derive estuarine/marine
chronic criteria for PFOS (see Appendix L for derivation of acute estuarine/marine benchmarks).
3.2.2 Derivation of Freshwater Chronic Tissue criteria for PFOS
Currently, the freshwater chronic PFOS toxicity data with measured tissue concentrations
were somewhat limited. There are 14 total freshwater aquatic life studies considered for either
quantitative (six studies - three fish, one invertebrate, and two amphibian studies) or qualitative
(eight studies) use in this aquatic life criterion. The quantitative studies only comprised data for
three of the eight MDRs. The qualitative studies provided supporting information for only one
additional MDR. Therefore, it was concluded that there is currently insufficient data to derive a
chronic tissue criterion using a GSD approach from empirical tissue data from toxicity studies.
However, these studies provided context to the translation of tissue criteria as described in
Section 3.2.3 below. This comparison is provided in the Effects Characterization (Section 4.5).
3.2.3 Translation of Chronic Water Column Criterion to Tissue Criteria
As described in Section 3.2.2 above, there are currently insufficient freshwater chronic
toxicity data with measured tissue concentrations to derive a chronic PFOS tissue criterion using
a GSD approach. Therefore, the chronic tissue criteria for PFOS were derived by translating the
chronic freshwater water column criterion (see Section 3.2.1.3) into tissue criteria using
bioaccumulation factors (summarized in Section 3.2.3.1 below) and the following equation:
Tissue Criteria = Chronic Water Column Criterion x BAF (Eq. 1)
The resulting tissue criteria corresponded to the tissue type associated with the BAF used in the
equation.
3.2,3.1 PFOS Bioaccumulation Factors (BAFs)
Section 2.11.3.1 above summarizes the literature search, calculation, and evaluation of
the PFOS BAFs for aquatic life. These BAFs were compiled by and can be found in Burkhard
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(2021). BAFs used in the derivation of the PFOS tissue 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 and to limit the effects of site-specific
differences in BAFs, 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 PFOS BAFs used in the
criteria derivation are presented in Table 3-11 and individual BAFs are provided in Appendix O.
Table 3-11. Summary Statistics for PFOS BAFs in Fish and Invertebrates1.
20,h
(ieomelric
Median
Cenlile
Mean liAl
liAl
liAl
Minimum
.Maximum
Category
n
(l./kซ-\v\v)
(l./kซ-\v\v)
(l./kซ-\v\v)
(l./kซ-\v\v)
(l./kซ-\v\v)
Invertebrates
28
771.6
924
111.5
2.69
100,000
Fish (Whole-Body)
28
3,739
5,905
803.9
4.79
46,098
Fish (Muscle)
21
1,069
1,048
346.4
8.72
50,234
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).
The fish tissue criteria were developed for muscle and whole-body to accommodate the
most commonly sampled tissue types in monitoring programs. Additional tissue values for
various other tissue types (e.g., liver and blood) were also calculated and can be found in
Appendix P.
3,2,3.2 Deriving Protective Tissue Concentrations from the Chronic Water-Column Criterion
Invertebrate whole-body and fish muscle and whole-body tissue criteria were derived
separately by multiplying the freshwater chronic water-column criterion (see Section 3.2.1.3) by
the respective 20th centile of the distribution of BAFs using Equation (Eq. 1) from Section 3.2.3.
The 20th centile BAF was used to derive tissue-based criteria as a relatively conservative BAF
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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 PFOS and resultant tissue-based exposures is relatively low as
well as those conditions with the bioaccumulation potential of PFOS is relatively high.
The invertebrate whole-body tissue criterion was calculated by multiplying the 20th
centile BAF of 111.5 L/kg ww by the PFOS freshwater chronic water criterion of 0.00025 mg/L,
resulting in an invertebrate whole-body tissue criterion of 0.028 mg/kg ww. The fish whole-body
tissue criterion was calculated by multiplying the 20th centile BAF of 803.9 L/kg ww by the
PFOS freshwater chronic water criterion of 0.00025 mg/L, resulting in a fish whole-body tissue
criterion of 0.201 mg/kg ww. The fish muscle tissue criterion was calculated by multiplying the
20th centile BAF of 346.4 L/kg ww by the PFOS freshwater chronic water criterion of 0.00025
mg/L, resulting in a fish muscle tissue criterion of 0.087 mg/kg ww. The chronic tissue-based
criteria are expected to be protective of 95% of freshwater genera potentially exposed to PFOS
under long-term exposures if the tissue-based criteria are not exceeded. The duration component
of the tissue-based criteria is expressed as an instantaneous duration because the tissue-based
criteria are protective of long-term conditions and represent an integrated measure of
bioaccumulated PFOS concentrations over time.
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 PFOS to the water column) into tissue concentrations through field-measured
bioaccumulation data of paired water and tissue concentrations in waterbodies. Nevertheless, the
chronic water column criterion is based on chronic toxicity tests where test organisms were fed.
In these tests, PFOS can directly affect species based on direct water column exposure and/or
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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- and dietary-based PFOS exposure and subsequent accumulation in
tissues. The chronic tissue criteria will allow states, Tribes, and stakeholders monitoring PFOS in
freshwater lentic and lotic systems to evaluate potential effects to aquatic organisms based on the
aquatic tissue monitoring data collected. Quantitatively acceptable data on the effects of dietary
exposures to aquatic species were relatively limited, thus the EPA chose to develop protective
values for freshwater aquatic organisms based on the observed relationship between water
column concentrations and tissue concentrations and observed PFOS toxicity in chronic tests
where PFOS was only added directly to the water column.
3.2.3,3 Deriving A Protective Duration and Exceedance Frequency for the Tissue-based
Chronic Criteria
3.2.3.3.1 Duration: Chronic Tissue-Based Criteria
PFOS concentrations in tissues are generally expected to change only gradually over time
in response to environmental fluctuations. The chronic tissue-based criteria averaging period, or
duration, was therefore specified as instantaneous, because tissue data provide point, or
instantaneous, measurements that reflect integrative accumulation of PFOS over time and space
in population(s) at a given site.
3.2.3.3.2 Frequency: Chronic Tissue-Based Criteria
The PFOS tissue-based criteria frequencies are set as "not to be exceeded" to ensure
protection of freshwater aquatic organisms. The "not to exceed" condition for frequency is meant
to account for the many variables influencing ecological recovery and uncertainty due to the
complete lack of PFAS-specific ecological recovery case studies available to inform recovery
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rates following elevated PFOS concentrations in aquatic biota. Ecological recovery times
following chemical disturbances in general are situational-specific, 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 the chemical
and potential for residual effects.
PFOS-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 PFOS 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 PFOS in adult rainbow trout exposed to PFOS for 28 days
via the diet followed by 28 days depuration was estimated to be 8.4 days in muscle tissue (Falk
et al. 2015), while the terminal half-life in rainbow trout receiving a one-time intra-arterial
injection of PFOS was 86.8 days (Consoer 2017). Additionally, the depuration half-life in
northern leopard frog tadpoles via a 40-day aqueous exposure to 0.01 mg/L PFOS was estimated
to be 2.2 days (Hoover et al. 2017). It's unclear whether PFOS half-life in aquatic organism
tissues is the mechanistic result of rapid depuration or an artifact of these measurements taken
during relatively short testing times (e.g., 28 days) where steady state between PFOS and water
and tissues has not occurred. Long-term uptake and subsequent excretion rates of PFOS has been
extensively studied in humans relative to aquatic life. Li et al. (2018) reported a median PFOS
half-life of 3.4 years in human serum following exposure to PFOS in drinking water, which the
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authors stated was in the range of previously published estimates. Due to chemical retention in
tissues, ecosystems impacted by discharges of bioaccumulative pollutants (such as selenium)
generally recover from chemical disturbances at relatively slow rates. For example, Lemly
(1997) concluded that although water quality in Belews Lake in 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 selenium discharges 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) 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 versus
persistence may also provide a mechanism influencing ecological recovery rates. The persistence
of PFOS in the environment has been attributed to the strong C-F bond, with no known
biodegradation or abiotic degradation processes for PFOS (refer to Section 2.3). Somewhat
similarly to PFOS, 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 = 1 year; 75th centile ~ 3
years; n = 20). Unlike metals, however, PFOS 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 PF AS into PFOS may represent an additional source of
PFOS in aquatic systems that further delays recovery.
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The persistence and bioaccumulative/human-made nature of PFOS in aquatic systems, in
combination with the documented recovery times of pollutants with similar chemical attributes
(Gergs et al. 2016; Lemly 1997; Mebane 2022), suggests aquatic systems may recover from
PFOS tissue criteria exceedances on the order of 5 to 10 years, if sources were eliminated.
However, recovery times could be longer, if the sources of PFOS and other PFAS that degrade
into PFOS have not been removed. Specifying a time interval associated with ecological
recovery from exceedances of PFOS tissue criteria, then, is highly uncertain given the lack of
PFOS-specific examples of ecological recovery and the many situational-specific factors
influencing recovery (Mebane 2022). For example, the lack of PFOS degradation in the
environment, and the fact that other PFAS in the environment can degrade into PFOS could act
as ongoing PFOS sources that further delay recovery. Given these uncertainties, the PFOS 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 PFOS has likely built up
through the food web and PFOS source reservoirs are likely to exist, representing a broad level
of PFOS 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 an exceedance of the PFOS tissue-based criteria
under the right conditions. Ecological recovery from such an exceedance could begin once PFOS
source reservoirs existing within the ecosystem are eliminated or decreased, including other
PFAS that degrade into PFOS; became permanently isolated from possible uptake by the
ecosystem (e.g., long-term burial with no benthic disturbances under certain environmental
conditions); and, unaffected organisms were able to repopulate the system through immigration
and/or reproductive events that yield generations that are no longer exposed to PFOS.
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Evaluation of PFOS concentrations in tissues would likely include evaluating the central
tendency 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 the PFOS Freshwater Aquatic Life Criteria and Acute
Estuarine/Marine Benchmark
The PFOS aquatic life criteria were developed to protect freshwater aquatic life against
adverse effects, such as mortality, altered growth, and reproductive impairments, associated with
acute and chronic exposure to PFOS. The nationally recommended criteria include water column
based acute and chronic criteria for fresh waters. The freshwater acute water column-based
criterion magnitude is 0.071 mg/L, and the chronic water column-based criterion magnitude is
0.00025 mg/L (Table 3-12). The chronic freshwater tissue-based criteria magnitudes are 0.201
mg/kg wet weight (ww) for fish whole-body, 0.087 mg/kg ww for fish muscle tissue and 0.028
mg/kg ww for invertebrate whole-body tissue. These PFOS aquatic life criteria are expected to
be protective of aquatic life on a national basis (Table 3-12). All of these water column and tissue
criteria are intended to be independently applicable and no one criterion takes primacy. All of the
recommended criteria (acute and chronic water column and tissue criteria) are intended to be
protective of aquatic life. Acute and chronic water column criteria for estuarine/marine waters
could not be derived at this time due to data limitations; however, an estuarine/marine acute
benchmark protective of aquatic life is provided in Appendix L.
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
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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 Perfluorooctane Sulfonate (PFOS) Ambient Water Quality
Criteria for the Protection of Aquatic Life in Freshwaters.
Clironic
Chronic
Acute Waid-
Chronic \Yaid-
Inverlebrale
I'ish
Chronic
Co! ii in n
Co! u mil
Whole-
Whole-
I'ish
Type/Media
(CMC)14
(CCC)15
Body12
Body12
Muscle1-2
Magnitude
0.071 mg/L
0.00025 mg/L
0.028
mg/kg ww
0.201
mg/kg ww
0.087
mg/kg ww
Duration
One-hour
average
Four-day
average
Instantaneous3
Not to be
Not to be
Not to be exceeded6
exceeded more
exceeded more
Frequency
than once in
three years on
average
than once in
three years on
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 are derived from the chronic water-column criterion magnitude (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 PFOS 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 PFOS 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
PFOS 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 0.55 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
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programs. 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.eov/chemical-research/epa-new-approach-methods-work-plan-rediicine-iise-
vertebrate-animals-chemical). 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 Perfluorooctane Sulfonate (PFOS) Benchmark for the Protection of
Aquatic Life in Estuarine/Marine Waters.
Type/Media
Acute Wnlcr Column Benchmark
Magnitude
0.55 mg/L
Duration
One hour on average
Frequency
Not to be exceeded more than once in three years on average
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4 EFFECTS CHARACTERIZATION FOR AQUATIC LIFE
The purpose of this section was to describe the supporting information for the derivation
of the PFOS aquatic life AWQC that contributed to the weight-of-evidence for the derivation.
This section includes: (1) additional analyses supporting the criteria that were used as part of the
lines-of-evidence discussion to better understand the influence of using less certain toxicity data
(Section 4.1); (2) an assessment of the influence of including non-North American resident
species in water column criteria derivation (i.e., species not resident to North America removed
from dataset; Section 4.2); (3) summaries of the toxicity studies with apical endpoints (e.g.,
effects on survival, growth, or reproduction) that were not used directly to derive the water
column criteria, but were used qualitatively to support them (Section 4.3); (4) a discussion of
acute to chronic ratios (Section 4.4); (5) a comparison of empirical tissue concentrations to
translated tissue criteria (Section 4.5); (6) a discussion of the effects of PFOS on aquatic plants
(Section 4.6); and (7) a discussion of the effects of PFOS on threatened and endangered species
(Section 4.7). The EPA derived the final national recommended PFOS aquatic life AWQC
described in the Effects Analysis Section (see Section 3 above). The additional analyses
presented here are solely intended to support the PFOS criteria through a weight-of-evidence
approach that evaluated the influence of data variation and uncertainties on the PFOS criteria.
4.1 Additional Analyses Supporting the Derivation of the Chronic Water
Column Criterion for Freshwater
In addition to the EPA's recommended freshwater chronic water column criterion of
0.00025 mg/L PFOS described above in Section 3.2.1.3, eight additional analyses supporting the
derivation of the chronic criterion were examined as part of a line-of-evidence evaluation to
consider the effect of including less certain toxicity data (i.e., the chronic toxicity values for
damselfly, fatmucket, and mayfly) on the magnitude of the freshwater chronic water column
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criterion. The data considered to be less certain generally centered around two specific areas: (1)
the difficultly in reliably estimating a chronic toxicity value given the wide spacing (up to 15-
fold difference) of the treatment concentrations (e.g., for fatmucket in Hazelton et al. (2012) and
damselfly in Bots et al. (2010) (see Section 3.1.1.3.3 and 3.1.1.3.4, respectively) and (2) the
uncertainty in the chronic toxicity values given the level of data presented in the papers
associated with the mayfly (N. triangulifer), fatmucket (L. siliquoidea), and damselfly (E.
cyathigerum) - see study summaries in Appendices C.2.1, C.2.3, and C.2.4, respectively).
The eight additional analyses presented below involved either changing or excluding
toxicity values from the three toxicity studies (Table 4-1). The additional analyses presented here
are solely intended to support the PFOS chronic criterion through a weight-of-evidence approach
that evaluated the influence of data variation on the criterion derivation process. Based on these
additional analyses, the EPA decided to retain the mayfly, fatmucket, and damselfly values as
presented in Section 3.1.1.3, to ensure protection of these sensitive taxa as well as the many
untested species for which these species may serve as representative taxonomic surrogate
species. The availability of additional toxicity data for these particular taxa would reduce the
uncertainty in the analysis. The criteria presented in Section 3.3 are the EPA's best estimate of
the maximum concentrations of PFOS that will support protection of sensitive aquatic life from
unacceptable chronic exposures.
115
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Table 4-1. Additional Analyses Supporting the Derivation of the Freshwater Chronic
Water Column Criterion.
Presented in the order that is summarized in the text below. *
Order of
Addilioiiiil
\ii;il\scs
Purpose of
Addilioiiiil An;il\sis
Doliiils ol' Addilioiiiil An;il\sis
Chronic Wiiior
(dill 111 11
( oiiceiit rii I ion lor
Addilioiiiil
Aiiiil> sis
(inii/l.)
Siud\
1
To explore the impact
of using the various
author reported
toxicity values for
damselfly
Used 10-day MATC of 0.3162
mg/L for damselfly instead of 150-
day MATC of 0.03162 mg/L
0.00025
Bots et al.
(2010)
2
Used 60-day NOEC of 0.1 mg/L
for damselfly instead of 150-day
MATC of 0.03162 mg/L
0.00025
3
Used 320-day NOEC of 0.01
mg/L for damselfly instead of 150-
day MATC of 0.03162 mg/L
0.00027
4
To explore the impact
of using the MATC
for fatmucket
Removed MATC of 0.01768 mg/L
for fatmucket
0.00021
Hazelton et al.
(2012)
5
To explore the impact
of using both the
ECio for fatmucket
and the 150-day
MATC for damselfly
Removed both MATC of 0.01768
mg/L for fatmucket and 150-day
MATC of 0.03162 mg/L for
damselfly
0.00016
Hazelton et al.
(2012) and Bots
et al. (2010),
respectively
6
To explore the impact
of using the ECiofor
fatmucket
Use estimated ECio of 0.0123
mg/L for fatmucket instead of
MATC of 0.01768 mg/L
0.00025
Hazelton et al.
(2012)
7
To explore the impact
of using the various
author-reported
toxicity values for
mayfly
Use ECio of 0.000272 mg/L for
mayfly calculated with linear
regression without control
0.00029
Soucek et al.
(2023)
8
Use ECio of 0.000232 mg/L for
mayfly calculated with linear
regression without control
0.00026
*Final derived freshwater chronic water column criterion was 0.00025 mg/L PFOS.
In the first additional analysis, instead of using the 150-day MATC of 0.03162 mg/L for
Enallagma cyathigerum as described in the calculation of the final freshwater chronic water
column criterion described above in Section 3.2.1.3, the 10-day MATC of 0.3162 mg/L was used
(Table 4-2) (Bots et al. 2010), yielding a freshwater FCV for PFOS of 0.0002486 mg/L. This
chronic water column concentration of 0.00025 mg/L (rounded to two significant figures) is the
same as the final chronic value of 0.00025 mg/L derived above. This first additional analysis
indicated that there is little difference in the calculated chronic criterion based either on the 150-
116
-------
day or 10-day MATC for is. cyathigerum. However, as the 150-day MATC was more
comparable to the other aquatic insect data and more representative of life-cycle effects than the
10-day MATC, the EPA has concluded that the 150-day MATC should be used quantitatively to
derive the freshwater chronic criterion.
In the second analysis, instead of using the 150-day MATC of 0.03162 mg/L for E.
cyathigerum, the 60-day NOEC of 0.1 mg/L from the same test was used (Table 4-2) (Bots et al.
2010), also yielding an FCV of 0.0002486 mg/L. Similar to the first analysis, there is no
difference in the calculated chronic criterion based either on the 150-day or 60-day NOEC for E.
cyathigerum. However, since the 150-day MATC was more comparable to the other aquatic
insect data and representative of life-cycle effects than the 10-day MATC, the EPA has
concluded that the 150-day MATC should be used quantitatively to derive the freshwater chronic
criterion.
In the third analysis, instead of using the 150-day MATC of 0.03162 mg/L for E.
cyathigerum, the 320-day NOEC of 0.01 mg/L from the same test was used (Table 4-2) (Bots et
al. 2010), yielding an FCV of 0.0002661 mg/L, or 0.00027 mg/L (rounded to two significant
figures). This analysis indicated that there is a slightly higher FCV (less stringent) in the
calculated chronic criterion if the 320-day NOEC for E. cyathigerum is used. However, as there
were concerns with the control survival of test organisms (reported as roughly 60% in the first 60
days), the EPA has determined that the 150-day MATC should be used quantitatively to derive
the freshwater chronic water column criterion since this toxicity value still represents a life-cycle
effect and control survival of test organisms was determined to be acceptable at this time point in
the test.
117
-------
In the fourth analysis, the MATC for fatmucket (L. siliquoidea) of 0.01768 mg/L was
removed from the chronic dataset to understand the influence of this toxicity value on the
criterion magnitude (Table 4-2). This additional analysis placed the GMCV of 0.03217 mg/L for
Danio among the four most sensitive genera, and yielded an FCV of 0.0002097 mg/L or 0.00021
mg/L (rounded to two significant figures) (Section 3.2.1.3; U.S. EPA 1985). The removal of the
chronic toxicity value for L. siliquoidea has only a modest influence on the calculated chronic
criterion magnitude (criteria became more stringent) but would eliminate mollusks from the
chronic PFOS dataset. The EPA decided to retain the fatmucket value to ensure representation
and protection of this sensitive taxon.
In the fifth analysis, the 150-day MATC of 0.03162 mg/L for damselfly (E. cyathigerum)
and MATC for fatmucket (L. siliquoidea) of 0.01768 mg/L were removed since these values are
less certain compared to other quantitative studies in the chronic criterion dataset (Table 4-2). As
noted above, these toxicity values were considered to be less certain due to (1) the difficultly in
reliably estimating a chronic toxicity value given the wide spacing (15-fold difference in
Hazelton et al. (2012) for L. siliquoidea and 10-fold difference in Bots et al. (2010) for is.
cyathigerum) of the treatment concentrations, and (2) the uncertainty in the chronic toxicity
values given the level of data presented in the papers. This fifth analysis yielded a freshwater
FCV for PFOS of 0.0001621 mg/L, or 0.00016 mg/L (rounded to two significant figures). The
calculated chronic criterion magnitude was reduced 1.6-fold. The EPA decided to retain the
damselfly and fatmucket values as presented in Section 3.1.1.3 in the criterion derivation to
ensure representation and protection of these sensitive taxa.
In the sixth analysis, the estimated ECio for fatmucket of 0.0123 mg/L was used in the
chronic dataset to understand the influence of this estimated toxicity value on the criterion
118
-------
derivation (Table 4-2), particularly since the EPA was not able to fit a curve to estimate an ECio
given that there were only two PFOS treatment groups and the gap in these exposure
concentrations is large (about 15-fold). This additional analysis yielded an FCV of 0.0002476
mg/L, or 0.00025 mg/L. This additional analysis indicated that the estimated toxicity value from
L. siliquoidea has no influence on the calculated chronic criterion. Since the estimated toxicity
value had no influence on the recommended CCC value, the author-reported MATC was used
instead.
In the seventh analysis, another author-reported ECio for mayfly of 0.000272 mg/L
(based on log-linear regression without controls) was used in the chronic dataset to understand
the influence of this alternate toxicity value on the criterion derivation (Table 4-2). This
additional analysis yielded a FCV was 0.0002938 mg/L, or 0.00029 mg/L, indicating that this
other toxicity value for N. triangulifer has little influence on the final calculated chronic
criterion. Since this other toxicity value had limited influence on the recommended CCC value,
the author-reported ECio (0.000226 mg/L, two parameter, threshold sigmoidal curve) described
in Section 3.1.1.3.1 in the criterion derivation was used instead.
Lastly, in the eighth analysis, a second author-reported ECio for mayfly of 0.000232
mg/L (based on log-linear regression with controls) was used in the chronic dataset to understand
the influence of this third toxicity value on the chronic criterion derivation (Table 4-2). This
additional analysis yielded a FCV was 0.0002550 mg/L, or 0.00026 mg/L, indicating that this
third possible toxicity value for N. triangulifer has limited influence on the calculated chronic
criterion. Since this alternate toxicity value had limited influence on the recommended CCC
value, the author-reported ECio (0.000226 mg/L, two parameter, threshold sigmoidal curve)
described in Section 3.1.1.3.1 in the criterion derivation was used instead.
119
-------
Table 4-2. GMCVs Used in Derivation of Chronic Criterion and Additional Analyses Supporting the Chronic Criterion for
Freshwater.
Ml)|{
< .|MU|>'
( .CIIUv
SptTir*
( hiiinii
( lili iinii
\ilililinn:il Vimlwio
<;\l< \
(III!! 1
I'l osr
1 ii-r
Vi'imil5
1 hi ril!
liiiirlli1
1 Mill
SiMh"
V\ rill If
i.iiihiir
F
Neocloeon
Mayfly,
Neocloeon triangulifer
0.000226
0.000226
0.000226
0.000226
0.000226
0.000226
0.000226
0.000272
0.000232
F
Chironomus
Midge,
Chironomus dilutus
0.005198
0.005198
0.005198
0.005198
0.005198
0.005198
0.005198
0.005198
0.005198
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
0.01768
0.01768
0.01768
0.01768
-
-
0.0123
0.01768
0.01768
F
Enallagma
Blue damselfly,
Enallagma cyathigerum
0.03162
0.3162
0.1
0.01
0.03162
-
0.03162
0.03162
0.03162
B
Danio
Zebrafish
Danio rerio
0.03217
0.03217
0.03217
0.03217
0.03217
0.03217
0.03217
0.03217
0.03217
D
Daphnia
Cladoceran,
Daphnia carinata
0.06519
0.06519
0.06519
0.06519
0.06519
0.06519
0.06519
0.06519
0.06519
Cladoceran,
Daphnia magna
A
Salmo
Atlantic salmon,
Salmo salar
>0.1
>0.1
>0.1
>0.1
>0.1
>0.1
>0.1
>0.1
>0.1
B
Pimephales
Fathead minnow,
Pimephales promelas
0.1098
0.1098
0.1098
0.1098
0.1098
0.1098
0.1098
0.1098
0.1098
E
Procambarus
Crayfish,
Procambarus fallax f. virginalis
0.167
0.167
0.167
0.167
0.167
0.167
0.167
0.167
0.167
D
Moina
Cladoceran,
Moina macrocopa
0.1789
0.1789
0.1789
0.1789
0.1789
0.1789
0.1789
0.1789
0.1789
H
Brachionus
Rotifer,
Brachionus calyciflorus
0.25
0.25
0.25
0.25
0.25
0.25
0.25
0.25
0.25
C
Xiphophorus
Swordtail fish,
Xiphophorus helleri
0.5997
0.5997
0.5997
0.5997
0.5997
0.5997
0.5997
0.5997
0.5997
C
Xenopus
African clawed frog,
Xenopus laevis
0.7507
0.7507
0.7507
0.7507
0.7507
0.7507
0.7507
0.7507
0.7507
Clawed frog,
Xenopus tropicalis
C
Lithobates
Northern leopard frog,
Lithobates pipiens
1.316
1.316
1.316
1.316
1.316
1.316
1.316
1.316
1.316
E
Hyalella
Amphipod,
Hyalella azteca
2.899
2.899
2.899
2.899
2.899
2.899
2.899
2.899
2.899
120
-------
Ml)|{
< .|MU|>'
( .CIIUv
SptTir*
( hiiinii
( lili iinii
\ilililinn:il Vimlwio
<;\l< \
(III!! 1
I'l osr
1 ir-.|:
Vi'iillil5
1 hi ril!
liiiirlli1
1 Mill
Mull"
V\ rill If
i.iiihiir
Ci
I'hy.sclLi
Snail.
I'liysclla liclcmsirnplhipnmiliu
8.527
8.527
8.527
8.527
8.527
8.527
8.527
8.527
8.527
D
Ceriodaphnia
Cladoceran,
Ceriodaphnia dubia
10.69
10.69
10.69
10.69
10.69
10.69
10.69
10.69
10.69
(,'hi'onir WntiT Column C,'oiicenti'iition
0.00025
0.00025
0.00025
0.00027
0.00021
0.00016
0.00025
0.00029
0.00026
1MDR Groups identified according to the list provided in Section 2.10.1 above.
2GMCVs as presented in Table 3-7 in Section 3.1.1.3. Genera presented in order of rank according to the chronic criterion derivation. The rank order of GMCVs
was not changed for the additional analyses.
3 Additional analysis with changes to toxicity value for E. cyathigerum.
4 Additional analysis with the exclusion of L. siliquoidea.
5 Additional analysis with the exclusion of L. siliquoidea and E. cyathigerum.
6 Additional analysis with the changes to toxicity value for L. siliquoidea.
1 Additional analysis with the changes to toxicity value for N. triangulifer.
121
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4.2 Influence of Using Non-North American Resident Species on PFOS
Criteria
The EPA conducted two additional analyses of the freshwater criteria by analyzing the
effect of reducing the limited toxicity datasets to only organisms that are resident to, or have
been introduced and have established populations in the conterminous U.S. These analyses were
conducted for illustrative purposes, to indicate the effects on the criteria magnitude of the
inclusion of data for taxa that are not resident species to North America but serve as surrogates
for other sensitive organisms. This analysis was conducted for both the acute and chronic
freshwater datasets only, since the estuarine/marine datasets are limited even when all species are
included.
4,2,1 Freshwater Acute Water Column Criterion with Native and Established Organisms
(Species Not Resident to North America removed from dataset)
For the purpose of illustrating the effect of including non-resident species in the acute
criterion calculation, additional analyses were made. For this illustrative analysis, four species
were removed from the freshwater acute water column criterion calculation to ensure that only
native, reproducing, or established organism in the conterminous U.S. were included: Japanese
swamp shrimp (Neocaridina denticulata), planarian (Dugesia japonica), crayfish (Pontastacus
leptodactylus) and cladoceran {Daphnia carinata). Removal of these species truncated the
freshwater acute dataset to 25 species (Table 4-3). None of the non-resident species were among
the four most sensitive, with the cladoceran (Daphnia carinata), being the most sensitive SMAV
of the four (Table 3-3). The acute water column concentration was 0.048 mg/L PFOS (Table 4-4)
when using the reduced dataset which was slightly lower than the recommended CMC of 0.071
mg/L. This value is lower than all of the GMAVs in Table 3-3. The EPA decided to retain the
full acute dataset and associated acute criterion for PFOS of 0.071 mg/L in order to have the
122
-------
largest, high-quality dataset to serve as surrogate species for the broad range of the thousands of
untested species present in the freshwater environment in the U.S.
Table 4-3. Ranked Freshwater Genus Mean Acute Values with Native and Established
Organisms, excluding Species Not Resident to
Vorth America.
Rank11
GMAV
(ing/l. PI-OS)
MI)U
(roup1'
Genus
Species
SMAY1'
(mg/L PI-OS)
1
0.07617
F
Neocloeon
Mayfly,
Neocloeon triangulifer
0.07617
2
3.075
D
Moina
Cladoceran,
Moina macrocopa
17.20
Cladoceran,
Moina micrura
0.5496
3
6.950
B
Pimephales
Fathead minnow,
Pimephales promelas
6.950
4
7.515
A
Oncorhynchus
Rainbow trout,
Oncorhynchus mykiss
7.515
5
13.5
G
Ligumia
Black sandshell,
Ligumia recta
13.5
6
15.99
C
Xenopus
African clawed frog,
Xenopus laevis
15.99
7
16.5
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
16.5
8
19.88
C
Hyla
Gray treefrog,
Hyla versicolor
19.88
9
27.86
B
Danio
Zebrafish,
Danio rerio
27.86
10
47.40
C
Ambystoma
Jefferson salamander,
Ambystoma jeffersonianum
51.71
Small-mouthed salamander,
Ambystoma texanum
30.00
Eastern tiger salamander,
Ambystoma tigrinum
68.63
11
56.49
C
Anaxyrus
American toad,
Anaxyrus americanus
56.49
12
59.87
E
Procambarus
Crayfish,
Procambarus fallax f virginalis
59.87
13
61.8
H
Brachionus
Rotifer,
Brachionus calyciflorus
61.8
14
64.35
G
Elliptio
Eastern elliptio,
Elliptio complanata
64.35
123
-------
Rank11
Ci.MAV
(mg/L PI-OS)
MI)U
(roup1'
Genus
Species
S.MAV1'
(mg/L PI-OS)
15
83.36
D
Daphnia
Cladoceran,
Daphnia magna
51.86
Cladoceran,
Daphnia pulicaria
134
16
109.2
C
Lithobates
American bullfrog,
Lithobates catesbeiana
133.3
Green frog,
Lithobates clamitans
113
Northern leopard frog,
Lithobates pipiens
72.72
Wood frog,
Lithobates sylvatica
130
17
172.1
G
Physella
Bladder snail,
Physella acuta
183.0
Snail,
Physella heterostropha pomilia
161.8
a Ranked from the most sensitive to the most tolerant based on Genus Mean Acute Value.
b From Appendix A: Acceptable Freshwater Acute PFOS Toxicity Studies.
0 MDR Groups identified by list provided in Section 2.10.1 above.
Table 4-4. Calculation of Freshwater Acute Water Column Concentration with Native and
Established Organisms (Species Not Resident to North America Removed from Dataset).
Calculated Freshwater FAV based on 4 lowest values: Total Num
jer of GMAVs in Dataset = 1
7
GMAV
Rank
Genus
(mg/L)
ln(GMAV)
In(GMAV)2
P=R/(N+1)
sqrt(P)
1
Neocloeon
0.07617
-2.57
6.63
0.056
0.236
2
Moina
3.075
1.12
1.26
0.111
0.333
3
Pimephales
6.950
1.94
3.76
0.167
0.408
4
Oncorhynchus
7.515
2.02
4.07
0.222
0.471
ฃ (Sum):
2.50
15.72
0.56
1.45
S2 =
458.21
S = slope
L =
-7.127
L = X-axis intercept
A =
-2.340
A = InFAV
FAV =
0.0963
P = cumulative probability
Acute Water
Column
Concentration =
0.048 mg/L PFOS
(rounded to two significant figures)
124
-------
4,2,2 Freshwater Chronic Water Criterion with Native and Established Organisms (Species Not
Resident to North America removed from dataset)
For the purpose of illustrating the effect of including non-resident species in the chronic
criterion calculation, additional analyses were made. For this illustrative analysis, two species
were removed from the chronic freshwater dataset that are not native or established organism in
the conterminous U.S.: the cladoceran (Daphnia carinata) and the clawed frog (Xenopus
tropicalis). Removal of these species truncated the freshwater chronic dataset to 17 species
representing 17 genera (Table 4-5). The revised freshwater chronic dataset consisted of all eight
MDRs. The cladoceran and clawed frog GMCVs were not among the four most chronically
sensitive species. Removal of the species that are not resident to North America had no effect on
the chronic water column concentration (Table 4-6) because other species were available from
the same genera and neither of the non-resident species were from the four most chronically
sensitive genera. The chronic water column concentration was 0.0002491 mg/L PFOS when
using the reduced dataset, which was the same as the recommended chronic criterion of 0.00025
mg/L. Therefore, the EPA decided to retain the full chronic dataset and associated chronic water
column criterion for PFOS of 0.00025 mg/L in order to have the largest, high quality dataset to
serve as surrogate species for the broad range of the thousands of untested species present in the
freshwater environment in the U.S.
125
-------
Table 4-5. Ranked Freshwater Genus Mean Chronic Values with Native and Established
Organisms.
Uank11
GM( V
(ing/l. pros)
MI)U
Group1
Genus
Species
S.MCV1'
(ing/l. PI OS)
1
0.000226
F
Neocloeon
Mayfly,
Neocloeon triangulifer
0.000226
2
0.005198
F
Chironomus
Midge,
Chironomus dilutus
0.005198
3
0.01768
G
Lampsilis
Fatmucket,
Lampsilis siliquoidea
0.01768
4
0.03162
F
Enallagma
Blue damselfly,
Enallagma cyathigerum
0.03162
5
0.03217
B
Danio
Zebrafish,
Danio rerio
0.03217
6
>0.1
A
Salmo
Atlantic salmon,
Salmo salar
>0.1
7
0.1098
B
Pimephales
Fathead minnow,
Pimephales promelas
0.1098
8
0.167
E
Procambarus
Crayfish,
Procambarus fallax f virginalis
0.167
9
0.1789
D
Moina
Cladoceran,
Moina macrocopa
0.1789
10
0.25
H
Brachionus
Rotifer,
Brachionus calyciflorus
0.25
11
0.5997
C
Xiphophorus
Swordtail fish,
Xiphophorus helleri
0.5997
12
>0.7610
C
Xenopus
African clawed frog,
Xenopus laevis
>0.7610
13
1.316
C
Lithobates
Northern leopard frog,
Lithobates pipiens
1.316
14
1.344
D
Daphnia
Cladoceran,
Daphnia magna
1.344
15
2.899
E
Hyalella
Amphipod,
Hyalella azteca
2.899
16
8.527
G
Physella
Snail,
Physella heterostropha pomilia
8.527
17
8.640
D
Ceriodaphnia
Cladoceran,
Ceriodaphnia dubia
8.640
a Ranked from the most sensitive to the most tolerant based on Genus Mean Chronic Value.
b From Appendix C: Acceptable Freshwater Chronic PFOS Toxicity Studies
0 MDR Groups identified by list provided in Section 2.10.1 above.
126
-------
Table 4-6. Calculation of Freshwater Chronic Water Column Concentration with Native
and Established Organisms.
Calculated Freshwater FCV based on 4
owest values: Total Number of GMCVs in Dataset = 17
GMCV
Rank
Genus
(mg/L)
ln(GMCV)
In(GMCV)2
P=R/(N+1)
sqrt(P)
1
Neocloeon
0.000226
-8.39
70.48
0.056
0.236
2
Chironomus
0.005198
-5.26
27.66
0.111
0.333
3
Lampsilis
0.01768
-4.04
16.28
0.167
0.408
4
Enallagma
0.03162
-3.45
11.93
0.222
0.471
ฃ (Sum):
-21.14
126.35
0.56
1.45
S2 =
472.36
S = slope
L =
-13.157
L = X-axis intercept
A =
-8.297
A = InFCV
FCV =
0.0002491
P = cumulative probability
Chronic Water
Column
Concentration =
0.00025 mg/L PFOS
(rounded to two significant figures)
4.3 Qualitatively Acceptable Water Column-Based Toxicity Data
Several studies were identified as either not meeting the EPA's data quality guidelines for
inclusion in the criteria derivation or did not have data available to support the independent
calculation of a toxicity value (e.g., LC50 and/orECio). However, these studies were used
qualitatively as supporting information to the PFOS criterion derived to protect aquatic life and
provide additional evidence of the observed toxicity and effects of PFOS, including the relative
sensitivities of surrogate, untested species. The key studies with apical endpoints (e.g., effects on
survival, growth, or reproduction) that were used qualitatively in the derivation of the PFOS
water column criteria are summarized below, grouped as either acute or chronic exposures and
sorted by relative sensitivity of genera following the previous study summaries included in the
Effects Analysis (Section 3). Qualitative study summaries within a factor of two of the final
acute and chronic values were also included and arranged according to taxonomic relatedness.
NOEC and LOEC values were provided in several of the study summaries for comparison to the
toxicity values summarized in the Effects Analysis section. The toxicity values summarized
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below as part of this Effects Characterization were not used quantitatively to derive the acute or
chronic PFOS freshwater criteria. Results of each individual study (as well as the rationale why a
study was not quantitatively acceptable) were considered relative to the corresponding
freshwater acute or chronic criterion magnitude to ensure the water column-based PFOS criteria
were not underproductive and to provide additional supporting evidence of the potential toxicity
of PFOS to aquatic organisms. Tabulated data for the studies summarized below, as well as
additional qualitative studies of less sensitive taxa, were listed in Appendix G.
4.3.1 Consideration of Qualitatively-Acceptable Acute Data
4.3.1.1 Qualitatively Acceptable Acute Data for Species Among the Four Most Sensitive
Genera Used to Derive the Acute Water Column Criterion
4.3.1.1.1 Most acutely sensitive genus, Neocloeon
There were no qualitatively-acceptable acute tests with the genus, Neocloeon.
4.3.1.1.2 Second most acutely sensitive genus, Moina
There were no qualitatively-acceptable acute tests with the genus, Moina.
4.3.1.1.3 Third most acutely sensitive genus, Pimephales
There were no qualitatively-acceptable acute tests with the genus, Pimephales.
4.3.1.1.4 Fourth most acutely sensitive genus, Oncorhynchus
Raine et al. (2021) exposed unfertilized rainbow trout (Oncorhynchus mykiss) oocytes
for three hours to PFOS (perfluorooctanesulfonic acid, > 97% pure, CAS No. 1763-23-1,
obtained from SynQuest Laboratories). The authors reported a residue accumulation NOEC of
0.87 mg/L and LOEC of 7.47 mg/L PFOS. This test is considered for qualitative use since the
exposure duration was too short for both acute and chronic test exposure according to the test
guidelines. Also, no apical endpoints were reported.
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4,3,1,2 Consideration of Relatively Sensitive Tests with Freshwater Species based on
Qualitatively-Acceptable Acute Data
4.3.1.2.1 Genus: Danio (zebrafish)
Cormier et al. (2019) evaluated the acute effects of (1,1,2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-
heptadecafluorooctane-1-sulfonic acid (PFOS, purity > 98%, CAS No. 2785-37-3, purchased
from Sigma-Aldrich in St. Louis, MO) on zebrafish (Danio rerio) via a 96-hour measured, static-
renewal study. Zebrafish embryos were collected and tested according to OECD TG 236. The
authors reported a 96 hour NOEC value of 700 ng/L PFOS (or 0.0007 mg/L) for hatching
success, embryo mortality, and developmental deformations. Since the value represents a greater
than low value (see description of decision rule in Section 2.10.3.2) (U.S. EPA 2013), the study
is only used qualitatively in the acute criterion.
Haimbaugh et al. (2022) evaluated the acute toxic effects of low-level (<2,400 ng/L)
PFOS on zebrafish from zero to five days post fertilization. Unmeasured test concentrations (24,
240, or 2,400 ng/L PFOA) were renewed daily. At test termination, the highest test concentration
(2,400 ng/L or 0.0024 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. This study also
represents a greater than low value.
The noted toxicity values provided above (>0.0007 mg/L and >0.0024 mg/L), indicated
that this genus might be more sensitive to acute exposures of PFOS than the quantitative data for
the genus (with a GMAV of 27.86 mg/L). However, this non-definitive qualitative value does
not provide any clarity on true sensitivity. All eight of the quantitatively-acceptable acute tests
for this species reported LCso values (range = 3.502 - 71.12 mg/L; geometric mean = 27.86
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mg/L; n = 8) that were more than an order of magnitude greater than the FAV, leading the EPA
to conclude that/), rerio is not a sensitive species to acute PFOS exposures.
4,3,2 Consideration of Qualitatively-Acceptable Chronic Data
4.3.2.1 Qualitatively Acceptable Chronic Data for Species Among the Four Most Sensitive
Genera Used to Derive the Chronic Water Column Criterion
4.3.2.1.1 Most chronically sensitive genus, Neocloeon
There were no qualitatively-acceptable chronic tests with the genus, Neocloeon.
4.3.2.1.2 Second most chronically sensitive genus, Chironomus
Zhai et al. (2016) exposed Chironomusplumosus larvae to PFOS (perfluorooctane
sulfonate, obtained from Tokyo Chemical Industries, Tokyo, Japan, 98% pure) spiked in
sediment for 10.3 days. The sediment was collected from the upstream region of the Yongding
River in Beijing, China. Spiking involved adding 1 mL of PFOS methanol solution (20 mg/L) to
the sediment to obtain a concentration of 100 ng/g PFOS and thoroughly mixing in a fume hood.
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. At the end of the experiment, the
surviving larvae were counted. The 10-day mortality NOEC was 0.00985 mg/L PFOS, the only
concentration tested. This study was considered for qualitative use for the following reasons: (1)
the exposure duration was relatively short when comparing to the test guidelines for aquatic
invertebrates and considered a sub-chronic exposure, (2) the sediment and test organisms used
appear to have been previously exposed to low levels of PFAS (albeit low exposures) based on
the measured concentrations reported in the paper, and (3) the apical endpoint for mortality
results in a > NOEC that is a low value, which provides little information to the relative
sensitivity of midge and was not used to derive the chronic PFOS criterion based on the data use
rules established for these criteria (see Section 2.10.3.2).
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Stefani et al. (2014) conducted a chronic (10 generation) test of PFOS (form and purity
not reported) with the midge, Chironomus riparius. The NOEC and LOEC were 0.0035 and >
0.0035 mg/L (as time-weighted average) as there were no effects on emergence, reproduction, or
sex ratio at this concentration. The results from this study were not acceptable for quantitative
use because only a single test concentration was used, the chronic value is a greater than low
value and not informative for criterion development, and there was a lack of details pertaining to
the characteristics of the sediment used in the exposure, including details regarding any
differences in measured concentrations over the duration of the exposure. Since this study was
focused on the chronic effects of PFOS to a relatively sensitive species, however, consideration
of the greater than chronic value from this study (> 0.0035 mg/L) in the context of other values
for the midge was prudent. The ECios for Chironomus dilutus of 0.05896 mg/L, 0.001588 mg/L,
and 0.0015 mg/L from MacDonald et al. (2004), McCarthy et al. (2021), and Krupa et al. (2022),
respectively, that were used quantitatively in the chronic criterion derivation are more robust
values than the toxicity value reported in Stefani et al. (2014), and likely a better estimation of
the sensitivity of C. riparius. The chronic value reported by Stefani et al. (2014), although
slightly lower than the Chironomus GMCV, is higher than the final recommended chronic
criterion, and was expressed as a NOEC, as no effects were observed at 0.0035 mg/L (as time-
weighted average).
In a companion paper to Stefani et al. (2014), Marziali et al. (2019) similarly conducted a
chronic (10 generation) test of PFOS (form and purity not reported) with C. riparius. The LOEC
based on F1 developmental time and F1 adult weight was < 0.004 mg/L (time-weighted
average). The were no effects on F1 exuvia length at this concentration. The results from this
study were not considered for quantitative use because only a single test concentration was used,
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there was a lack of consistent observed effects in both the control and the treatment groups
across the generations, and details pertaining to the characteristics of the sediment used in the
exposure were lacking, including details regarding any differences in measured concentrations
over the duration of the exposure. Again, it is prudent to consider the less than chronic value
from this study (< 0.004 mg/L) in the context of the more robust and definitive values for midge.
Similar to the determination above, the chronic values established for the related midge, C.
dilutus, from MacDonald et al. (2004) and McCarthy et al. (2021) are deemed more reliable and
definitive values representing the sensitivity of the genus in the freshwater chronic water column
criterion dataset.
4.3.2.1.3 Third most chronically sensitive genus, Lampsilis
There were no qualitatively-acceptable chronic tests with the genus, Lampsilis.
4.3.2.1.4 Fourth most chronically sensitive genus, Enallagma
Van Gossum et al. (2009) conducted a chronic, approximately 4-month renewal test of
PFOS (tetraethylammonium salt, 98% purity) with damselfly, Enallagma cyathigerum. The test
organisms were larvae that had reached the F2 instar stage. Dilution water was dechlorinated tap
water. Photoperiod was 16 hours light and 8 hours dark. Light intensity was not reported. A
primary stock solution was prepared and proportionally diluted with dilution water to prepare the
test concentrations. Exposure vessels were plastic containers (15 cm x 10 cm x 11 cm) with a 2
cm depth of test solution. The test employed 19-20 larvae each in two test concentrations plus a
negative control. Nominal concentrations were 0 (negative control), 0.01, 0.1, 1, and 10 mg/L.
All larvae were housed (and presumably tested) in temperature-controlled rooms at 21ฑ1.3ฐC.
No other water quality parameters were reported as having been measured in test solutions.
Negative control mortality was said to be much lower than the 100% mortality that occurred at 1
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and 10 mg/L but was not reported. The 4-month NOEC (behavioral - including general activity,
swimming performance, foraging success) was 0.010 mg/L. The 4-month LOEC was 0.100
mg/L. The calculated MATC was 0.03163 mg/L. The chronic value was acceptable for
qualitative use because only non-apical endpoints were reported.
4,3,2.2 Consideration of Relatively Sensitive Tests with Freshwater Species based on
Qualitatively-Acceptable Chronic Data
4.3.2.2.1 Genus: Lithobates (leopardfrog)
Flynn et al. (2021) evaluated the chronic effects of perfluorooctane sulfonic acid (PFOS,
CAS# 1763-23-1, > 96% purity, purchased from Sigma-Aldrich) on Northern Leopard frogs,
Lithobatespipiens (formerly, Ranapipiens), via a 30-day sediment-spiked measured, static
mesocosm study. The study authors reported a 30-day NOEC of 0.016 mg/L for weight, snout-
vent length and mortality and a 30-day LOEC of 0.00006 mg/L for developmental stage
(measured as Gosner stage). Independently-calculated ECios could not be calculated as the EPA
was unable to fit a model with significant parameters. Therefore, given this was an outdoor
mesocosm with spiked sediment that included the addition of algal and zooplankton
communities, and the EPA was unable to independently calculate toxicity values based on the
replicate level data provided by the study authors, this study was used qualitatively to derive the
final recommended chronic water column criterion.
Hoskins et al. (2022) evaluated the chronic effects of PFOS alone and in mixture with
PFHxS on northern leopard frogs (Lithobates pipiens) from Gosner Stage 25 tadpoles through
metamorphosis (Gosner Stage 46) in a static-renewal measured exposure. No effects were
observed on survival, mass index and snout-vent length at 0.000934 mg/L PFOS (the highest test
concentration at test termination (day 120). The test results in a greater than NOECs (>0.000934
mg/L) and provides little information to the relative sensitivity of the species. Therefore, this
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study was not used to derive the chronic PFOS criterion based on prior data use rules (U.S. EPA
2013).
4.4 Acute-to-Chronic Ratios
The 1985 Guidelines allow the use of a final acute-to-chronic ratio (FACR) to convert the
FAV to the FCV as an alternative approach to derive the chronic criterion instead of the direct
calculation to determine the FCV (as described in Section 2.10.1) when the eight MDRs are not
met (U.S. EPA 1985). While this alternative approach was not needed for the derivation of the
chronic PFOS criterion, which was derived from empirical chronic data with all of the eight
MDRs met, the possibility of calculating a scientifically-defensible FACR is as follows, for
illustrative purposes only. Seventeen ACRs for eight invertebrate species and two fish species
can be calculated from the quantitative acute and chronic toxicity data (Appendix A and
Appendix C). Appendix I includes the ACRs for freshwater aquatic species with quantitative
chronic values for which comparable quantitative acute values were reported from the same
study or same investigator and laboratory combination. For each species where more than a
single ACR was calculated, species mean acute-to-chronic ratios (SMACRs) were also
calculated as the geometric mean value of individual ACRs for a species. In the case of a single
ACR within a species, that ACR was the SMACR.
The ACRs ranged from 4.110 to 12,877 across all tests (a factor of 3,133), which occurs
within the Daphnia magna SMACR. There was little explanation for the extreme range in ACRs
among paired tests with D. magna. However, the ACR of 12,877 from paired tests conducted by
Lu et al. (2015) appears to be an outlier. Excluding the 12,877 outlier ACR from the paired tests
with D. magna reported by Lu et al. (2015) and from the paired test with Daphnia carinata
(Logeshwaran et al. 2021) produced an SMACR range of 16.32 to 1,030. This range was greater
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than a factor of 10 with no relationship between SMACR and SMAV apparent. The 1985
Guidelines do not provide for calculation of a FACR under these circumstances.
4.5 Comparison of Empirical Tissue Concentrations to Translated Tissue
Criteria
Measured PFOS tissue data were reported in 14 publications focused on freshwater
species, six of which were quantitatively acceptable and eight of which were qualitatively
acceptable (Table 4-7). The six quantitatively acceptable studies included data for one
invertebrate, two fish, and one amphibian species, and the eight qualitatively acceptable studies
included data for two invertebrate and four fish species. Results of these studies are summarized
in Section 4.5.1 and Section 4.5.2, below.
Tissue concentration data from these toxicity studies were compared to the translated
tissue criteria (fish whole body, fish muscle, invertebrate whole body) and supplemental fish
tissue values (blood, liver, reproductive tissue) to better understand the protectiveness of the
chronic aquatic life tissue criteria. Although tissue concentrations from the toxicity literature
were limited, translated tissue criteria and supplemental fish tissue values were lower than tissue-
based PFOS concentrations from chronic toxicity studies where toxic effects were observed,
suggesting that the tissue criteria (and supplemental fish tissue values) are protective. Finally,
while no amphibian tissue criteria are available, tissue concentrations from two amphibian
toxicity tests suggest that the fish tissue criteria are protective of amphibians.
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Table 4-7. Comparison of Empirical Tissue
Concentrations to Chronic Tissue Criteria ant
Additional Tissue Values.
Species
Kndpoinl
Percent I-liecl
Obsened
Measured
Tissue
Coiieeiiliitlioii
(inป/kป \\\\)'
C;ilcul:ilcd
Chronic
Tissue
\ ill lies2
(inป/kป ww)
Tissue Type
Reference
Quantitative Studies
Fatmucket
(Lampsilis siliquoidea)
Probability of
successful
metamorphosis of
glochidia
33.5%
LOEC: 0.248
0.028
Invertebrate Whole-
body (adult)
Hazelton et
al. (2012)
Zebrafish
(Danio rerio)
F1 survival
21%
LOEC: 4.8
0.201
Whole-body (adult
female)
Wang et al.
(2011)
LOEC: 6.0
0.201
Whole-body (adult
male)
Fathead minnows
(Pimephales
promelas)
Fecundity
49% (LOEC)
NOEC: 7.1 -
LOEC 19.4
0.616
Liver concentrations
(adult male)3
Ankley et al.
(2005)
NOEC: 31.8-
LOEC 82.9
0.616
Liver concentrations
(adult female)
NOEC: 8.8 -
LOEC 19.9
1.293
Gonad concentrations
(adult male)3
NOEC: 33.1 -
LOEC 81.6
1.29
Gonad concentrations
(adult female)
Growth (weight in Fl)
18%
LOEC: 37.9
1.293
Gonad concentrations
(adult F0 male)3
Suski et al.
(2021)
LOEC: 37.4
1.29
Gonad concentrations
(adult F0 female)
LOEC: 84.5
0.616
Liver (adult F0 male)
LOEC: 68.2
0.616
Liver (adult F0
female)
Northern leopard frog
(Lithobates pipiens)
Length at
metamorphosis (GS
424)
17%
LOEC: 66.6
None
Available
Whole-body (before
metamorphosis, day
54)
Ankley et al.
(2004)
Gosner stage after 40
days
5%
LOEC: 14.36
Whole-body
Hoover et al.
(2017)
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C;ilcul:itcd
Measured
Chronic
Tissue
Tissue
Percent KITecl
C'oncenlriilion
\ ill lies2
Species
Kndpoinl
Ohsened
\\\\)'
ww)
Tissue Type
Reference
Qualitative Studies
Red worms
(Limnodrilus
hoffmeisteri)
Reduction in
superoxide dismutase5
13%
LOEC: 1,757
(dw)
0.028
Invertebrate Whole-
body
Liu et al.
(2016)
Great pond snails
(Lymnaea stagnalis)
Survival
31%
LOEC: 2,877
0.028
Invertebrate Whole-
body
Olson (2017)
European eels
(Anguilla anguillci)
Survival
0%
NOEC: >
5.037
0.616
Liver
Roland et al.
(2014)
Goldfish
(Carassius auratus)
Survival
0%
NOEC: >
39.91 (dw)
0.087
Muscle
Feng et al.
(2015)
Common carp
(Cyprinus carpio)
Condition factor
3%
LOEC: 168.4
0.616
Liver
Hagenaars et
al. (2008)
Zebrafish
(Danio rerio)
Swimming distance5
18%
LOEC: 21.6
0.201
Whole-body
Spulber et al.
(2014)
1 Measured tissue concentrations are author-reported values. The EPA did not independently calculate toxicity values for tissue
concentrations.
2 Chronic tissue value concentrations represent chronic tissue criteria (invertebrates, fish muscle, fish whole body) or additional
tissue values (fish blood, fish liver, fish reproductive tissue) calculated from BAFs for a given tissue type. See Section 3.2.3
and Appendix P for details.
3 Fish reproductive tissue value based on female reproductive tissue.
4 Gosner stage (GS) associated with this endpoint is not specifically reported by the study authors. However, the authors define
complete metamorphosis as emergence of the forelimbs, which is GS 42 according to Taylor and Kollros (1946).
5 Non-apical endpoint.
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4.5.1 Comparison of Quantitative Studies and Tissue-Based Criteria
Tissue concentration data from these toxicity studies were compared to the translated
tissue values for invertebrates and fish to better understand the protectiveness of the aquatic life
tissue criteria. Hazelton et al. (2012) exposed adult fatmucket (Lampsilis siliquoideci) to aqueous
PFOS for 36 days. Measured PFOS water concentrations in the control and exposure treatments
averaged 0.0021, 0.0045, and 0.0695 mg/L, respectively. Corresponding tissue concentrations
were 0.009, 0.015 and 0.248 mg/kg wet weight. A statistically significant decrease in the
probability of successful metamorphosis of glochidia to the juvenile stage was observed in the
highest PFOS exposure concentration.
Wang et al. (2011) exposed larval (8 hpf) zebrafish (Danio rerio) to aqueous PFOS for
five months. Fish were exposed to three nominal PFOS concentrations (0.005, 0.05, and 0.25
mg/L, respectively). Whole-body PFOS tissue concentrations measured after five months in the
two highest exposure concentrations averaged 6.0 and 11.2 mg/kg wet weight, respectively, in
males, and 4.8 and 7.8 mg/kg wet weight, respectively, in females. PFOS was also measured in
embryos produced from exposed parents and averaged 5.69 and 11.35 ng/embryo wet weight in
the two highest exposure concentrations. Weights of embryos were not reported by the study
authors, so concentrations could not be calculated to compare embryo tissue concentrations to
the translated tissue criteria. However, given the study design included tissue measurements in
the parental (F0) generation and the exposure to the offspring generation (Fl) was via maternal
transfer, the tissue concentration in the F0 generation associated with the Fl survival LOEC of
0.05 mg/L was a whole-body tissue concentration of 6.2 and 4.0 mg/kg wet weight (ww) in male
and females, respectively.
Ankley et al. (2005) exposed sexually mature adult fathead minnows (Pimephales
promelas) to aqueous PFOS for 21 days during which time they were allowed to reproduce, and
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then the resulting offspring were held for an additional 24 days in the same exposure
concentrations. Aqueous measured PFOS concentrations in the control and exposure treatments
averaged <0.001, 0.0276, 0.101, 0.281, and 0.818 mg/L, respectively. PFOS was measured in the
plasma, livers, and gonads of adult males and females after 21 days, in embryos, and in whole-
body larval samples after 12 and 24 days. Tissue measurements were not made in organisms
from the highest exposure concentration, where exposed adults were either dead or listless after
14 days. Plasma PFOS concentrations in adult organisms exposed to 0.0276, 0.101, and 0.281
mg/L PFOS averaged 26.9, 135, and 354 mg/L in males, and 47.1, 177, and 471 mg/L in
females. Liver PFOS concentrations in adults averaged 7.1, 19.4, and 109 mg/kg wet weight in
males, and 31.8, 82.9, and 261 mg/kg wet weight in females, respectively. Similarly, gonad
PFOS concentrations averaged 8.8, 19.9, and 108 mg/kg wet weight in males, and 33.1, 81.6, and
263 mg/kg wet weight in females. PFOS concentrations in embryos from parents exposed to
0.0276, 0.101, and 0.281 mg/L PFOS were 9.3, 11.5, and 28.6 mg/kg, respectively. Larval PFOS
concentrations measured after 12 and 24 days of exposure were similar, with whole-body
concentrations corresponding to the 0.0276, 0.101, and 0.281 mg/L exposures were 19.8, 48.0,
and 57.5 mg/kg wet weight after 12 days, and 17.8, 49.0, and 83.5 mg/kg wet weight after 24
days. The most sensitive apical endpoint was fecundity, with an aqueous ECio of 0.051 mg/L. No
corresponding tissue-based ECio was calculated, but the corresponding liver concentrations
would be expected to fall between 7.1 and 19.4 mg/kg in males and 31.8 and 82.9 mg/kg in
females, and the corresponding gonad concentrations would be expected to fall between 8.8 and
19.9 mg/kg in males and 33.1 and 81.6 mg/kg in females. No muscle or whole-body
measurements in adults are available to perform a direct comparison to the tissue criteria.
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Suski et al. (2021) reported the chronic toxicity of PFOS-K (PFOS potassium salt, CAS#
2795-39-3, > 98%,) on the fathead minnow, Pimephalespromelas. Measured PFOS
concentrations in water were 0.00014 (control), 0.044, 0.088, 0.14, and 0.231 mg/L. The most
sensitive endpoint from the study was a significant decrease in the mean mass of individuals in
the larval F1 generation with the author-reported NOEC and LOEC, based on growth in the F1
generation, being 0.044 (6% reduction in growth compared to controls) and 0.088 mg/L PFOS
(associated with an 18% reduction in growth), respectively. The calculated MATC based on
mean mass of individuals in the larval F1 generation is 0.06222 mg/L. The F1 larval LOEC was
associated with measured gonad and liver concentrations in F0 male and females of 37.9, 37.4,
84.5, and 68.2 mg/kg ww, respectively. No corresponding tissue-based ECio was calculated, but
the corresponding gonad and liver ECio concentrations would be expected to be greater than the
translated reproductive tissue concentration of 1.29 mg/kg ww and the translated liver tissue
concentration of 0.616 mg/kg ww. No muscle or whole-body measurements in adults are
available to perform a direct comparison to the fish tissue criteria.
Ankley et al. (2004) exposed Northern leopard frogs (Lithobatespipiens) to PFOS from
Gosner stage 8/9 embryos through metamorphosis. The time to metamorphosis ranged from 60-
112 days. All frogs in the highest exposure concentration died before metamorphosis. The most
sensitive apical endpoint was length at metamorphosis, which was significantly lower (p < 0.05)
in the second highest exposure relative to the control. The measured aqueous PFOS
concentrations in the NOEC and LOEC exposure concentrations averaged 0.957 and 3.42 mg/L
over the full exposure duration. PFOS in whole body tissue was analyzed as dry weight but
reported by the authors as wet weight normalized. Corresponding whole-body tissue NOEC and
LOEC concentrations measured in tadpoles exposed for 54 days (before metamorphosis) were
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16.9 and 66.6 mg/kg wet weight normalized. Whole body concentrations were also measured
after 35 days and were higher than the 54-day measurements at the 0.957 mg/L exposure (22.1
mg/kg wet weight normalized) and 3.42 mg/L exposures (117.0 mg/kg wet weight normalized).
Tadpole moisture content was not reported.
In a separate study with L. pipiens. Hoover et al. (2017), exposed juvenile (Gosner stage
26) northern leopard frogs to three PFOS concentrations (0.008, 0.078, and 0.884 mg/L
measured PFOS, respectively) for 40 days. Survival, growth (snout-vent length), and
developmental time (Gosner stage after 40 days) were measured, and the most sensitive apical
endpoint was developmental time (Gosner stage after 40 days), with a NOEC of 0.008 mg/L and
a LOEC of 0.078 mg/L. Whole body PFOS concentrations in frogs exposed to 0.008 mg/L PFOS
in solution averaged 10.45 mg/kg dry weight after 40 days, and concentrations in frogs exposed
to 0.078 mg/L averaged 51.46 mg/kg dry weight after 40 days. Tadpole moisture content was not
reported in this study. 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
criteria, 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 Fish
Database (usgs.gov)). Corresponding 40-day NOEC and LOEC wet weight PFOS tissue
concentrations were 2.92 and 14.36 mg/kg wet weight, respectively.
In all of the studies described above, the translated tissue criteria and supplemental fish
tissue concentrations were lower than the measured tissue concentrations where toxicity was
observed, suggesting that the tissue criteria are protective. As noted above, tissue concentrations
associated with the LOEC in both Ankley et al. (2004) and Hoover et al. (2017) are higher than
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the fish whole-body tissue criterion, suggesting that the fish tissue criteria is protective of
amphibians.
4.5.2 Comparison of Qualitative Studies and Tissue-Based Criteria
Like the comparison with the quantitative studies, tissue concentration data from these
qualitative toxicity studies were compared to the translated tissue values for invertebrates and
fish to better understand the protectiveness of the aquatic life tissue criteria. Liu et al. (2016)
exposed 4-5 cm body length red worms (Limnodrilus hoffmeisteri) to two aqueous
concentrations of PFOS for 10 days at pH 8.0 and measured oxidative stress biomarker activity.
Measured exposure concentrations were 0.567 and 5.494 mg/L PFOS, and corresponding whole-
body tissue PFOS concentrations were 89.5 and 1,757 mg/kg dry weight. Moisture content was
not reported. A significant (P < 0.05) reduction in superoxide dismutase was observed in the
highest treatment concentration after 10 days. Apical endpoints were not reported for this
exposure. In a separate study with L. hoffmeisteri, Qu et al. (2016) calculated 48-hour ECsos in
response to PFOS at three pH values (6.2, 7.0, 8.0). PFOS was not measured in water. However,
whole body tissue concentrations were measured after 48-hours in the control, 0.2 mg/L, and 2.0
mg/L nominal PFOS exposures. Whole-body tissue concentrations in the 2.0 mg/L exposure
were 23.41 mg/kg dry weight at pH 6.2 and 12.61 mg/kg dry weight at pH 8.0. After 48 hours, a
significant (P < 0.05) increase in superoxide dismutase was observed in both the 0.2 mg/L and
2.0 mg/L PFOS treatments; however, no significant differences were observed at pH 8.0 for
either treatment level.
Olson (2017) exposed adult great pond snails (Lymnaea stagnalis) to PFOS for 21 days.
The most sensitive apical endpoint was survival, with aNOEC of 3 mg/L PFOS nominal, and a
LOEC of 6 mg/L PFOS nominal. Whole-body PFOS tissue concentrations at the NOEC and
LOEC after 21 days were 8,969 mg/kg dry weight and 9,820 mg/kg dry weight, respectively.
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Percent moisture was not reported by the study authors, so dry weights were converted to wet
weights using the average whole soft body % moisture content of 70.7% for the snail species
Achatina achatina (Achaglinkame et al. 2020) in order to more directly compare L. stagnalis
tissue concentrations from this study to the invertebrate chronic whole body tissue criterion.
Resulting wet weight PFOS concentrations at the NOEC and LOEC were 2,628 and 2,877
mg/kg, respectively.
Roland et al. (2014) exposed juvenile European eels {Anguilla anguilla) to PFOS for 28
days. Measured PFOS water concentrations were 0.00001 mg/L in the control and 0.00081 and
0.011 mg/L in the two aqueous exposure concentrations. Corresponding liver tissue PFOS
concentrations after 28 days were 0.0338 and 5.037 mg/kg wet weight, respectively. The study
authors noted that during the study, there was no mortality, and no significant differences in
growth across either PFOS treatment. Significant (p < 0.05) changes in protein expression were
reported for both exposure concentrations.
Feng et al. (2015) conducted a 96-hour study with juvenile goldfish (Carassius auratus)
and measured the effects of PFOS on mortality or antioxidant enzyme activity. Measured PFOS
in the two exposure concentrations were 1.04 |imol/L (0.520 mg/L) and 10.18 |imol/L (5.09
mg/L). Liver, gill, and muscle PFOS concentrations were 32.81, 42.13, and 33.08 mg/L dry
weight, respectively, at the lower exposure level, and 58.37, 69.02, and 39.91 mg/L dry weight,
respectively, at the higher exposure level. No mortality occurred during the test. Among the
antioxidant enzyme activity endpoints, glutathione peroxidase activity was significantly (p <
0.05) lower in the highest exposure concentration than the control.
Hagenaars et al. (2008) exposed juvenile common carp (Cyprinus carpio) to three
exposure concentrations of PFOS plus a control for 14 days and measured relative condition
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factor and several non-apical endpoints related to liver function. Nominal PFOS exposure
concentrations were control, 0.1, 0.5, and 1 mg/L. Corresponding liver PFOS concentrations
after the 14-day exposure were 0.97, 35.97, 168.4, and 283.0 mg/kg wet weight. The most
sensitive endpoint was condition factor, which was significantly (p < 0.0001) lower than controls
at the 0.5 mg/L nominal aqueous exposure concentration. Hepatosomatic index was also
significantly (P < 0.05) lower at the 0.5 mg/L concentration compared to the control. The
corresponding liver tissue PFOS concentration at this effect concentration (LOEC) was 168.4
mg/kg wet weight.
Spulber et al. (2014) exposed Danio rerio embryos (2 hpf) to 0.1 mg/L and 1.0 mg/L
nominal PFOS concentrations for seven days. Corresponding whole-body PFOS concentrations
in 7-day-old larvae were 21.6 and 213.5 mg/kg wet weight, respectively. Spulber et al. (2014)
reported no effects of PFOS on viability, time to hatch, or deformities. The most sensitive
endpoint was swimming distance, where fish exposed to the 0.1 and 1.0 mg/L PFOS treatments
exhibited lower levels of activity (p < 0.05) in response to a pulse of darkness.
The translated tissue criteria and supplemental fish tissue concentrations were lower than
the measured tissue concentrations where toxicity was observed for all of the qualitative studies.
Although tissue concentrations from the toxicity literature were limited, available data suggest
that the tissue criteria are protective.
4.6 Effects on Aquatic Plants
Available data for aquatic plants and algae were reviewed to determine if aquatic plants
were likely to be more sensitive than aquatic animals to aqueous PFOS exposure (see Appendix
E). Toxicity values for freshwater plants were well above the freshwater chronic water column
criterion. Effect concentrations for freshwater plants and algae ranged from 0.19 to 252 mg/L
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compared to the range in animal chronic values of 0.000226 to 16.35 mg/L (Appendix C:
Acceptable Freshwater Chronic PFOS Toxicity Studies). Therefore, it was not necessary to
develop a criterion based on the toxicity of PFOS to aquatic plants. The PFOS freshwater acute
and chronic criteria are expected to be protective of freshwater plants.
4.7 Protection of Threatened and Endangered Species
The PFOS acute and chronic datasets are 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 PFOS toxicity data for listed species indicating that the 2024 PFOS criteria are
protective of these listed species, based on available scientific data.
4,7,1 Quantitatively Acceptable Acute Toxicity Data for Listed Species
Quantitatively acceptable acute toxicity test data evaluating the effects of PFOS on
threatened and endangered freshwater species were available for rainbow trout (Oncorhynchus
mykiss) with a SMAV of 7.515 mg/L PFOS (Palmer et al. 2002a; Sharpe et al. 2010). The
SMAV is over 100 times higher than the recommended acute criterion (CMC) of 0.071 mg/L,
indicating the acute criterion is protective of rainbow trout and is expected to be protective of
other listed salmonid species.
Quantitatively acceptable acute data were also available for the Eastern tiger salamander
(A. tigrinum). While the species is not considered to be a federally listed species, it is considered
endangered in Delaware, Maryland, New Jersey, New York, and Virginia (Smith 2003),
threatened in North Carolina (Smith 2003), and critically imperiled in Louisiana (2024). The
Eastern tiger salamander is also closely related to the endangered California tiger salamander
(U.S. FWS 2016a; U.S. FWS 2016b; U.S. FWS 2017). The A tigrinum SMAV of 68.63
(Tornabene et al. 2021) is almost an order of magnitude above the recommended acute criterion
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(CMC) of 0.071 mg/L, indicating the acute criterion is protective of the Eastern tiger salamander
and the federally-listed California tiger salamander. There were no acceptable acute toxicity data
for endangered or threatened estuarine/marine aquatic species.
4.7.2 Quantitatively Acceptable Chronic Toxicity Data for Listed Species
Quantitatively acceptable chronic toxicity test data evaluating the effects of PFOS on
threatened and endangered freshwater species were available for the listed Atlantic salmon
{Salmo salar) with a SMCV of >0.1 mg/L PFOS (Spachmo and Arukwe 2012). The SMCV is
400 times higher than the recommended chronic criterion (CCC) of 0.00025 mg/L, indicating the
acute criterion is protective of Atlantic salmon and other listed salmonid species. There were no
acceptable chronic toxicity data for endangered or threatened estuarine/marine aquatic species.
4.7.3 Qualitatively Acceptable Toxicity Data for Listed Species
Focusing on qualitatively acceptable tests with apical endpoints and water column
exposures, there were toxicity data available for two fish species, rainbow trout and Atlantic
salmon. For rainbow trout, the NOEC for mortality was 1 mg/L in a 12-day microcosm exposure
and 3 mg/L in a 14-day early life stage static laboratory test (Oakes et al. 2005). For Atlantic
salmon {Salmo salar), no adverse effects for growth were observed at the highest treatment
concentration (0.1 mg/L PFOS) following a 49-day exposure (Arukwe et al. 2013). For both
species, the qualitative NOECs were orders of magnitude greater than the recommended acute
criterion (CCC) of 0.00025 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.
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4.8 Summary of the PFOS Aquatic Life Criterion and the Supporting
Information
The PFOS aquatic life AWQC were developed to protect aquatic life against adverse
effects, such as mortality, altered growth, and reproductive impairments, associated with acute
and chronic exposure to PFOS. The national recommended criteria include water column-based
acute and chronic criteria for fresh waters. The freshwater acute water column-based criterion
magnitude is 0.071 mg/L, and the chronic water column-based criterion magnitude is 0.00025
mg/L (0.25 |ig/L), The chronic freshwater criterion also contains tissue-based criteria expressed
as 0.201 mg/kg wet weight (ww) for fish whole-body, 0.087 mg/kg ww for fish muscle tissue,
and 0.028 mg/kg ww for invertebrate whole-body tissue. These PFOS aquatic life criteria are
expected to be protective of all freshwater aquatic life on a national basis. Although empirical
PFOS 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
using available estuarine/marine species toxicity data and a NAMs application of the EPA
ORD's peer-reviewed web-ICE tool (see Appendix L). The estuarine/marine acute water
column-based benchmark magnitude is 0.55 mg/L and is expected to protect estuarine/marine
aquatic life from acute aqueous PFOS exposures. The EPA conducted additional analyses
supporting the derivation of the water column criteria for PFOS (as summarized above in
Sections 4.1 and 4.2) and confirmed that the criteria and benchmark calculations presented in this
document accurately reflect the latest and best available scientific knowledge.
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Appendix A Acceptable Freshwater Acute PFOS Toxicity Studies
A.1 Summary Table of Acceptable Quantitative Freshwater Acute PFOS Toxicity Studies
Species (lilVsliiuo)
MciiHxr
Tesl
Dnnilion
( hcmic;il /
Piiriu
Pll
Temp.
(ฐC)
r.riw-i
Anllior
Reported
IITecl
(ttne.
(inii/l.)
r.PA
( iileuliiled
r.lTecl
(one.
(in Si/1 - >
liiiiil
r.iTeei
(one.
(niti/l.)'
Speeies
Mesin
Aeule
Ysilue
Reference
Planaria (0.9 cm),
Dugesia japonica
s,u
96 hr
PFOS-K
>98%
25
LC50
17
-
17
-
Li (2008)
Planaria (0.9 ฑ0.1 cm),
Dugesia japonica
s,u
96 hr
PFOS-K
>98%
-
25
LC50
23
22.68
22.68
-
Li (2009)
Planaria (10-12 mm),
Dugesia japonica
R, U
96 hr
PFOS-K
>99%
-
20
LC50
29.46
-
29.46
22.48
Yuanetal. (2014)
Eastern elliptio
(76.5 g, 48.7 mm).
Elliptio complanata
(formerly, Unio
complamatus)
R,M
96 hr
PFOS-K
90.49%
7.9-8.5
21.8-
23.7
LC50
59
64.35
64.35
64.35
Drottar and
Krueger (2000f)
Fatmucket
(glochidia, <24 hr),
Lampsilis siliquoidea
S,M
24 hr
PFOS
>98%
8.46
20
EC50
(viability)
16.5
-
16.5
-
Hazelton (2013);
Hazelton et al.
(2012)
Fatmucket
(juvenile, 4-6 wks),
Lampsilis siliquoidea
R, M
96 hr
PFOS
>98%
8.46
20
LC50
158.1
-
158.ld
16.5
Hazelton (2013);
Hazelton et al.
(2012)
Black sandshell
(glochidia, <24 hr),
Ligumia recta
S,M
24 hr
PFOS
>98%
8.46
20
EC50
(viability)
13.5
-
13.5
-
Hazelton (2013);
Hazelton et al.
(2012)
Black sandshell
(juvenile, 4-6 wk),
Ligumia recta
R, M
96 hr
PFOS
>98%
8.46
20
LC50
141.7
-
141.7d
13.5
Hazelton (2013);
Hazelton et al.
(2012)
Bladder snail
(mixed age),
Physella acuta
(formerly, Physa acuta)
S,U
96 hr
PFOS-K
>98%
-
25
LC50
178
183.0
183.0
183.0
Li (2009)
A-l
-------
Species (li IVsl ;iLto t
Method'
losl
Dui'iilion
( hcinic;il /
PuriU
pll
IVmp.
(ฐC)
r.iTcci
Author
Reported
r.riw-i
Cone,
(inii/l.)
i:p\
( iilciiliilod
l.llecl
(one.
(inii/l.)
liiiiil
r.riw-i
(one.
liii"/!.)'
Species
Mciin
Acute
Yiilue
(iii'^/l.)
Reference
Snail (adult, 4 mo.),
Physella heterostropha
pomilia
(formerly, Physa
pomilia)
S,M
96 hr
PFOS-K
>98%
25
LC50
161.77
-
161.8
161.8
Funkhouser
(2014)
Rotifer
(<2 hr old neonates),
Brachionus calyciflorus
s,ub
24 hr
PFOS-K
>98%
-
20
LC50
61.8
-
61.8
61.8
Zhang et al.
(2013)
Cladoceran (6-12 hr),
Daphnia carinata
s,u
48 hr
PFOS-K
>98%
-
21
LC50
8.8
11.56
11.56
11.56
Logeshwaran et
al. (2021)
Cladoceran (<24 hr),
Daphnia magna
S,M
48 hr
PFOS-K
90.49%
8.2-8.6
19.3-
20.2
EC50
61
58.51
58.51
-
Drottar and
Krueger (2000g)
Cladoceran (<24 hr),
Daphnia magna
S,U
48 hr
PFOS-K
95%
-
21
EC50
(immobility)
67.2
-
67.2
-
Boudreau (2002);
Boudreau et al.
(2003a)
Cladoceran (<24 hr),
Daphnia magna
S,U
48 hr
PFOS
Unreported
-
21
EC50
(immobility)
37.36
35.46
35.46
-
Ji et al. (2008)
Cladoceran (<24 hr),
Daphnia magna
S,U
48 hr
PFOS-K
>98%
7.82-
7.91
25
EC50
63
55.40
55.40
-
Li (2009)
Cladoceran (<24 hr),
Daphnia magna
s,u
48 hr
PFOS-K
>98%
7.82-
7.91
25
EC50
63
72.70
72.70
-
Li (2009)
Cladoceran (<24 hr),
Daphnia magna
s,u
48 hr
PFOS-K
>98%
7.82-
7.91
25
EC50
63
64.60
64.60
-
Li (2009)
Cladoceran (<24 hr),
Daphnia magna
S,M
48 hr
PFOS-K
99%
7
22
LC50
78.09
-
78.09
-
Yang et al. (2014)
Cladoceran (<24 hr),
Daphnia magna
S,U
48 hr
PFOS
98%
7.2
20
EC50
(death/immobility)
23.41
-
23.41
-
Lu et al. (2015)
Cladoceran (<24 hr),
Daphnia magna
s,u
48 hr
PFOS-K
>98%
7
20
EC50
(death/immobility)
79.35
94.58
94.58
-
Liang et al. (2017)
Cladoceran (12-24 hr),
Daphnia magna
s,u
48 hr
PFOS-K
98%
-
-
LC50
22.77
22.43
22.43
51.86
Yang et al. (2019)
Cladoceran (<24 hr),
Daphnia pulicaria
s,u
48 hr
PFOS-K
95%
-
21
EC50
(immobility)
134
-
134
134
Boudreau (2002);
Boudreau et al.
(2003a)
A-2
-------
Species (li IVsl ;iLto t
Method'
losl
Dui'iilion
( hcinic;d /
PuriU
pll
IVmp.
(ฐC)
r.iTcci
Author
Reported
r.riw-i
Cone,
(inii/l.)
i:p\
( iilciiliilod
l.llecl
(one.
liiiiil
r.riw-i
(one.
liii"/!.)'
Species
Mciin
Acnle
\ illllO
Reference
Cladoceran (<24 hr),
Moina macrocopa
S,U
48 hr
PFOS
Unreported
-
25
EC50
(immobility)
17.95
17.20
17.20
17.20
Ji et al. (2008)
Cladoceran (<48 hr),
Moina micrura
S,M
48 hr
PFOS
>98%
-
27
LC50
0.5496
-
0.5496
0.5496
Razak et al.
(2023)
Crayfish (intermolt),
Pontastacus
leptodactylus
(formerly, Astacus
leptodactylus)
R,M
96 hr
PFOS-K
>98%
6.79
21
LC50
48.81
-
48.81
48.81
Belek et al. (2022)
Crayfish (juvenile, 2
wks, 0.041 g),
Procambarus fallax f.
virginalis
S,M
96 hr
PFOS-K
>98%
-
25
LC50
59.87
59.87
59.87
59.87
Funkhouser
(2014)
Japanese swamp
shrimp,
Neocaridina
denticulata
S,U
96 hr
PFOS-K
>98%
-
25
LC50
108
12.91
12.91
-
Li (2009)
Japanese swamp
shrimp,
Neocaridina
denticulata
S,U
96 hr
PFOS-K
>98%
-
25
LC50
108
28.55
28.55
-
Li (2009)
Japanese swamp
shrimp,
Neocaridina
denticulata
S,U
96 hr
PFOS-K
>98%
-
25
LC50
108
10.32
10.32
15.61
Li (2009)
Mayfly (<24 hr larva),
Neocloeon triangulifer
S,M
96 hr
PFOS-K
98%
-
23
LC50
0.08
0.07617
0.07617
0.07617
Soucek et al.
(2023)
Rainbow trout
(juvenile),
Oncorhynchus mykiss
S,M
96 hr
PFOS-K
86.9%
-
11.3-
12.9
LC50
22
22.59
22.59
-
Palmer et al.
(2002a)
A-3
-------
Species (li IVsl ;iLto t
MciiHxr
losl
Dui'iilion
( hcinic;il /
PuriU
pll
IVmp.
(ฐC)
r.iTcci
Author
Reported
r.riw-i
Cone,
(inii/l.)
i:p\
( iilciiliilod
l.llecl
(one.
(niii/l.)
liiiiil
r.riw-i
(one.
liii"/!.)'
Species
Mciin
Acute
Value
(niii/l.)
Reference
Rainbow trout (parr),
Oncorhynchus mykiss
R,M
96 hr
PFOS-K
98%
10
LC50
2.5
-
2.5
7.515
Sharpe et al.
(2010)
Zebrafish (embryo),
Danio rerio
R, U
96 hr
PFOS
Unreported
7-8.5
26
LC50
71
-
71
-
Ye et al. (2007)
Zebrafish (embryo),
Danio rerio
S,U
96 hr
PFOS-K
>97%
7.2-7.5
26
LC50
58.47
-
58.47
-
Hagenaars et al.
(2011a)
Zebrafish (adult),
Danio rerio
R, M
96 hr
PFOS-K
98%
-
26
LC50
22.2
-
22.2
-
Sharpe et al.
(2010)
Zebrafish
(3 mo., 2.2 cm),
Danio rerio
R, U
96 hr
PFOS-K
Unreported
-
23
LC50
17.0
-
17.0
-
Wang et al. (2013)
Zebrafish (embryo),
Danio rerio
S,U
96 hr
PFOS-K
98%
8.3
28.5
LC50
68
71.12
71.12
-
Li et al. (2015)
Zebrafish (embryo),
Danio rerio
S,U
96 hr
PFOS-K
98%
-
28
LC50
3.502
-
3.502
Du et al. (2016a);
Du et al. (2017)
Zebrafish
(embryo, 1 hpf),
Danio rerio
R, U
96 hr
PFOS
Unreported
-
26
LC50
34.2
38.82
38.82
-
Stengel et al.
(2017b)
Zebrafish (embryo),
Danio rerio
R, M
96 hr
PFOS-K
>98%
-
26
LC50
23.99ฐ
-
23.99
27.86
Nilen et al. (2022)
Fathead minnow
(juvenile),
Pimephales promelas
S,M
96 hr
PFOS-K
90.49%
8.2-8.5
22
LC50
9.5
9.020
9.020
-
Drottar and
Krueger (2000c)
Fathead minnow (79 d),
Pimephales promelas
S,U
96 hr
PFOS-Li
24.5%
8.0-8.4
19.2-
19.5
LC50
4.655f
5.356f
5.356
6.950
3M Company
(2000)
American toad (larva,
Gosner stage 26),
Anaxyrus americanus
S,U
96 hr
PFOS
Unreported
-
21
LC50
62g
63.41
63.41d
-
Tornabene et al.
(2021)
American toad (larva,
Gosner stage 41),
Anaxyrus americanus
S,U
96 hr
PFOS
Unreported
-
21
LC50
62g
56.49
56.49
56.49
Tornabene et al.
(2021)
Gray treefrog (larva,
Gosner stage 26),
Hyla versicolor
S,U
96 hr
PFOS
Unreported
-
21
LC50
79
78.33
78.33d
-
Tornabene et al.
(2021)
A-4
-------
Author
i:p\
Speeies
Reported
( iileuliiled
liiiiil
Mean
r.ricct
I'.ITeel
I'.ITeel
Aeule
Tesl
( hcmic;il /
Temp.
Cone.
(one.
(one.
\ nine
Speck's (lircshiiic)
Method'
Dui'iilion
Piiriu
pll
(ฐC)
r.iTcci
(inii/l.)
(niii/l.)
(ill"/!.)'
Reference
Gray treefrog (larva,
Gosner stage 40),
Hyla versicolor
S,U
96 hr
PFOS
Unreported
21
LC50
24
19.88
19.88
19.88
Tornabene et al.
(2021)
American bullfrog
(tadpole, Gosner stage
PFOS
Unreported
25),
Lithobates catesbeiana
s,u
96 hr
-
21
LC50
144
154.8
154.8d
-
Flynnetal. (2019)
(formerly, Rana
catesbeiana)
American bullfrog
(larva, Gosner stage
26),
s,u
96 hr
PFOS
Unreported
-
21
LC50
163
133.3
133.3
133.3
Tornabene et al.
(2021)
Lithobates catesbeiana
Green frog (larva,
Gosner stage 26),
Lithobates clamitans
(formerly, Rana
s,u
96 hr
PFOS
Unreported
-
21
LC50
113
-
113
113
Tornabene et al.
(2021)
clamitans)
Northern leopard frog
(larva, Gosner stage
26),
s,u
96 hr
PFOS
21
LC50
73
72.72
72.72
72.72
Tornabene et al.
Lithobates pipiens
Unreported
(2021)
(formerly, Rana
pipiens)
Wood frog (larva,
Gosner stage 26),
Lithobates sylvatica
(formerly, Rana
s,u
96 hr
PFOS
Unreported
-
21
LC50
130
-
130
130
Tornabene et al.
(2021)
sylvatica)
African clawed frog
(embryos),
Xenopus laevis
R,M
96 hr
PFOS-K
86.9%
7.3
24
LC50
13.8
15.53
15.53
-
Palmer and
Krueger (2001)
A-5
-------
Author
i:p\
Speeies
Reported
( iileuliiled
liiiiil
Mean
r.ricct
I'.ITeel
I'.ITeel
Aeule
Tesl
( hcmic;il /
Temp.
Cone.
(one.
(one.
Value
Speck's (lircshiiic)
Method'
Dui'iilion
Piiriu
pll
(ฐC)
r.iTcci
(inii/l.)
(niii/l.)
(ill"/!.)'
(niii/l.)
Reference
African clawed frog
(embryos),
Xenopus laevis
R,M
96 hr
PFOS-K
86.9%
7.27
24
LC50
17.6
18.04
18.04
-
Palmer and
Krueger (2001)
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.26
24
LC50
15.3
14.6
14.60
15.99
Palmer and
Krueger (2001)
Jefferson salamander
(larva, Harrison stage
40),
Ambystoma
S,U
96 hr
PFOS
Unreported
-
21
LC50
64
51.71
51.71
51.71
Tornabene et al.
(2021)
jeffersonianum
Small-mouthed
salamander
(larva, Harrison stage
40),
s,u
96 hr
PFOS
Unreported
-
21
LC50
41g
46.71
46.71d
-
Tornabene et al.
(2021)
Ambystoma texanum
Small-mouthed
salamander
PFOS
Unreported
Tornabene et al.
(2021)
(larva, Harrison stage
46),
s,u
96 hr
-
21
LC50
41g
30.00
30.00
30.00
Ambystoma texanum
Eastern tiger
salamander
PFOS
Unreported
Tornabene et al.
(2021)
(larva, Harrison stage
40),
s,u
96 hr
-
21
LC50
73
68.63
68.63
68.63
Ambystoma tigrinum
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 PFOS in the range of concentrations tested under similar
conditions. Daily renewal of test solutions.
0 Reported in moles converted to milligram based on a molecular weight of 500.13 mg/mmol or 538.22 mg/mmol PFOS-K.
dNot used in SMAV calculation; only the most sensitive life-stage used.
e Values in bold used the in the SMAV calculation.
f Author-reported LC50 of 19 mg/L x 24.5% PFOS = 4.655 mg/L PFOS; EPA-calculated LC50 of 21.86 mg/L x 24.5% PFOS = 5.356 mg/L PFOS.
g Author pooled tests or life-stages.
A-6
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A.2 Detailed PFOS Acute Freshwater Toxicity Study Summaries and
Corresponding Concentration-Response Curves (when calculated for
the most sensitive genera)
The purpose of this section was to present detailed study summaries for tests that were
considered quantitatively acceptable for criteria derivation, with summaries grouped and ordered
by genus sensitivity. Concentration-response (C-R) models developed by the EPA that were used
to determine acute toxicity values used for water column criterion derivation are also presented
for the most sensitive genera when available. C-R models included here with study summaries
were those for the five most sensitive genera (consistent with Section 3.1.1.1). When required,
the EPA also included models for non-resident species that were more sensitive than the fourth
most sensitive North American resident genus. In many cases, authors did not report C-R data in
the publication/supplemental materials and/or did not provide C-R data upon the EPA request. In
such cases, the EPA did not independently calculate a toxicity value and the author-reported
effect concentrations were used in the derivation of the criterion.
A.2.1 Most Sensitive Freshwater Genus for Acute Toxicity: Neocloeon (mayfly)
Soucek et al. (2023) conducted a 96-hour acute toxicity test to determine the effects of
PFOS-K (PFOS potassium salt, CAS # 2795-39-3, 98% purity) on the parthenogenetic mayfly,
Neocloeon triangulifer. The test was performed under static, nonrenewal conditions beginning
with < 24-hour old nymphs. Exposures consisted of five mayfly nymphs per 30 mL
polypropylene cup filled with 20 mL test solution. The control and each of six PFOS test
concentrations were replicated five times for a total of 25 test organisms per treatment. Nominal
test concentrations were 0 (control), 0.0156, 0.0313, 0.0625, 0.125, 0.250, and 0.500 mg/L
PFOS. Mean measured PFOS concentrations (EPA Analytical Method 1633; LC-MC/MS) were
0.0002 (control), 0.017, 0.046, 0.052, 0.103, 0.253, and 0.358 mg/L PFOS, respectively. Animals
were exposed at 23 ฑ 1ฐC under a 16:8 hour light (~110 - 300 lux):dark cycle and fed live diatom
A-7
-------
biofilm scraping beginning on Day 0. Percent survival in the control treatment after 96 hours was
100%. A uniform C-R pattern for percent survival was observed decreasing from 100 and 96% in
the lowest treatments to 4 and 0% in the highest test treatments. The EPA was able to
independently calculate a 96-hour LCso of 0.07617 mg/L (0.06546 - 0.08688 mg/L; 95% CI) for
this study. The EPA's independently-calculated LCso is in line with the author-reported LCso of
0.08 mg/L. Therefore, the independently-calculated LCso of 0.07617 mg/L was acceptable for
quantitative use in the derivation of the freshwater acute water column criterion for PFOS.
A-8
-------
A.2.1.1 Soucek el al. (2023) Concentration Response Curve - Neocloeon (mayfly)
Publication: Soucek et al. (2023)
Species: Mayfly, Neocloeon triangulifer
Genus: Neocloeon
EPA-Calculated LCso: 0.07617 (0.06546 - 0.08688) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
3.4747
0.5921
58679
4.414 e9
e
0.07617
0.0063135
12.0644
< 2.2 e"16
Concentration-Response Model Fit:
USGS unpublished
Neoclcon triangulifer
Log-Logistic, 2 para
0.001 o.oio o.ioo
PFOS ( mg/L )
A.2.2 Second Most Sensitive Freshwater Genus for Acute Toxicity: Moina (cladoceran)
Ji et al. (2008) performed a 48-hour static, unmeasured acute test of PFOS (acid form,
CAS # 1763-23-1, purity unreported) with Moina macrocopa. The test followed the EPA's
Methods for measuring the acute toxicity of effluents and receiving waters to freshwater and
marine organisms [U.S. EPA/600/4-90/027F; (U.S. EPA 2002)]. M. macrocopa used for testing
were obtained from brood stock cultured at the Environmental Toxicology Laboratory at Seoul
A-9
-------
National University (South Korea). Test organisms were less than 24 hours old at test initiation.
Dilution water was moderately hard reconstituted water (hardness typically 80-100 mg/L as
CaCCb). Experiments were conducted in glass jars of unspecified size and fill volume. The
photoperiod for the test was assumed by the reviewers to have been 16:8-hours light:dark, as was
used for daphnid culture in tests by the same authors. Preparation of test solutions was not
described. The test involved four replicates of five neonates each in five nominal test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 6.25,
12.5, 25, 50 and 100 mg/L. Test temperature was maintained at 25 ฑ 1ฐC. Authors note water
quality parameters (pH, temperature, conductivity, and D.O.) were measured 48 hours after
exposure, but the information was not reported. Survival of organisms in the negative control
was not reported in the paper. However, raw data were obtained by the EPA from the study
authors and control survival was 100% in the acute test, meeting the EPA/600/4-90/027F
requirement of at least 90% survival for test acceptability. The study authors reported a 48-hour
ECso value of 17.95 mg/L (C.I. 14.72 - 21.18) forM macrocopa. The 48-hour EC50 value was
independently-calculated by the EPA as 17.20 (13.73 - 20.66) mg/L. The independently-
calculated acute toxicity value was quantitatively used in the derivation of the freshwater acute
water column criterion.
Razak et al. (2023) tested the acute toxicity of perfluorooctanesulfonate (PFOS) on
Moina micrura for 48 hours in a measured, static experiment. PFOS (>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
A-10
-------
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
PFOS 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. 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 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,
LCso, LC75, LC90) were calculated using Probit analysis, and the 48-hour LC50 value of 549.6
|ig/L, or 0.5496 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.2.2.1 Ji et al. (2008) Concentration Response Curve - Moina (cladoceran)
Publication: Ji et al. (2008)
Species: Cladoceran, Moina macrocropa
Genus: Moina
EPA-Calculated LCso: 17.20 (13.73 - 20.66) mg/L
Concentration-Response Model Estimates:
Parameter
Kslimale
Sul. Krror
1-slal
p-value
b
2.18
0.47
4.60
4.16 e06
e
20.34
2.06
9.87
<2.2 e16
A-ll
-------
Concentration-Response Model Fit:
Ji et al. 2008
Moina macrocopa
Weibull type 1, 2 para
1 10 100
PFOS-K ( mg/L)
A.2.3 Third Most Sensitive Freshwater Genus for Acute Toxicity: Pimeyhales (fathead
minnow)
3M Company (2000) provides the results of a 96-hour static, unmeasured acute toxicity
test with the fathead minnow, Pimephalespromelas, and PFOS-Li (perfluorooctanesulfonate
lithium salt, CAS # 29457-72-5). A stock solution was made with carbon-filtered well water at a
test sample concentration of 400 mg/L and where the test sample was reported as a mixture of
PFOS-Li (24.5%) in water (75.5%). Fish were obtained from a commercial supplier (Aquatic
Biosystems, Fort Collins, CO) and were 79 days old at test initiation with an average length of
2.1 cm and weight of 0.069 g. Exposure vessels were 2 L glass beakers containing 1 L of
solution and 10 fish per beaker (0.69 g fish/L). Each test treatment was replicated twice with
nominal test concentrations (control, 3.2, 5.6, 10.0, 18.0, 32.0 and 56.0 mg/L test sample).
Throughout the experiment the dissolved oxygen (D.O.) ranged from 4.8 - 7.9 mg/L, pH 8.0 - 8.4
and a test temperature of 19.2 - 19.5ฐC. The low D.O. of 4.8 mg/L was only observed in one
A-12
-------
replicate of the highest test concentration at 96 hours; D.O. was >6.0 mg/L for all other
treatments and replicates. No mortality occurred in the control treatment and 100% was observed
in the highest treatment (56 mg/L). The study authors reported that the test sample containing
24.5% PFOS-Li exhibited a 96-hour LCso of 19 mg/L, which equates to 4.655 mg/L as PFOS.
The independently-calculated 96-hour LCso value was 21.86 (17.63 - 26.08) mg/L, which
equates to 5.356 mg/L as PFOS and is acceptable for quantitative use in the derivation of the
acute freshwater criterion for PFOS.
Drottar and Krueger (2000c) evaluated the acute effects of PFOS-K (CAS# 2795-39-3,
Lot #217 (T-6295) obtained from the 3M Company, 90.49% purity, stored at ambient room
temperature) on juvenile fathead minnows (Pimephalespromelas) during a 96-hour measured,
static study. Researchers stated they followed protocols by U.S. EPA Series 850 (OPPTS
850.1075), OECD Guideline 203, and ASTM E729-88a. A primary stock solution was prepared
at 27 mg/L and mixed with an electric mixer for 22 hours prior to use in testing to ensure
solubilization of the test substance. After mixing, the primary stock solution was proportionally
diluted with dilution water to prepare the four additional test concentrations. Test fish were
obtained from cultures at Wildlife International Ltd. in Easton, Maryland. The minnows were
held for approximately 126 days prior to testing and were acclimated to test conditions for 48
hours prior to test initiation. Fish were fed a commercially-prepared diet prior to the 48-hour
acclimation period. All fish used in the test were from the same source and year class, and the
total length of the longest fish was no more than twice the length of the shortest. Fathead
minnows were randomly distributed among mean measured test concentrations of 0 (control),
3.3, 5.6, 9.5, 17 and 28 mg/L, with 10 fish per 25-L polyethylene aquarium provided in
duplicate. Aquaria were filled with 15 L of test solution with an observed D.O. of 7.7 - 8.4 mg/L,
A-13
-------
temperature of 22 ฑ 2ฐC, pH of 8.2 - 8.5, and a total hardness of 131 mg/L as CaCCb. Fathead
minnows were subjected to a 16:8-hour light:dark photoperiod at 391 lux. Sand and 0.45 um
filtered well water from a 40 m deep well on site served as both the culture water and the testing
media. The authors reported an LCso of 9.5 mg/L PFOS. The EPA's independently-calculated
96-hour LCso was 9.020 (7.146 - 10.89) mg/L, rounded to four significant figures, and was used
quantitatively to derive the freshwater acute water column criterion.
A.2.3.1 3M Company (2000) Concentration Response Carve - Pimephales (fatheadminnow)
Publication: 3M Company (2000)
Species: Fathead minnow, Pimephalespromelas
Genus: Pimephales
EPA-Calculated LCso: 21.86 (95% C.I. 17.63 - 26.08) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.8756
0.3016
6.2191
4.999 e10
e
26.5720
2.5673
10.3501
< 2.2 e"16
Concentration-Response Model Fit:
Pimephales promelas
Weibull type 1,2 para
1 00
PFOS (mg/L )
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A.2.3.2 Drottar andKrueger (2000c) Concentration Response Curve - Pimephales (fathead
minnow)
Publication: Drottar and Krueger (2000c)
Species: Fathead minnow, Pimephalespromelas
Genus: Pimephales
EPA-Calculated LCso: 9.020 (95% C.I. 7.146 - 10.89) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
2.9683
0.4433
6.6965
2.134 e11
d
1.0503
0.0174
60.4403
< 2.2 e"16
e
9.0196
0.8923
10.1084
< 2.2 e"16
Concentration-Response Model Fit:
Drottar and Krueger 2000
Pimephales promelas
Log Logistic type 1, 3 para
PFOS ( mg/L )
A.2.4 Fourth Most Sensitive Freshwater Genus for Acute Toxicity: Oncorhynchus (trout)
Sharpe et al. (2010) evaluated the acute effects of perfluorooctane sulfonate (PFOS,
potassium salt, CAS #2795-39-3, 98% purity) to Oncorhynchus mykiss, rainbow trout, via a 96-
hour renewal measured exposure (renewal not stated in paper, but assumed based on other
information provided, including the test Guideline protocol). Limited details about the test
protocol were provided in the publication, but the authors noted they followed OECD Guideline
203, and did not identify any deviations from these test guidelines. Trout eggs were obtained
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from Raven Trout Hatchery, transported immediately postfertilization to the University of
Alberta aquatics facility, and kept in dechlorinated City of Edmonton water. The eggs were
reared until hatching in Heath trays in a recirculating, temperature-controlled system at 10ฐC
with a 12:12-hour light:dark photoperiod (the same conditions are assumed for the toxicity test).
The rainbow trout used in the study were parr (2-3 g, the fourth stage of the salmon life cycle) at
test initiation. The dilution water was dechlorinated City of Edmonton water, and dissolved
oxygen content and temperature were monitored daily, but physico-chemical results were not
reported. PFOS was dissolved in MeOH, and all vehicle controls received a volume of MeOH
equal to that present in the highest PFOS dose of that experiment (final MeOH content 0.2%
v/v). The concentration of PFOS in any experiment was always well below its reported solubility
in water (-500 mg/L). Trout toxicity tests were performed using food-grade 2 L plastic tanks
with four fish per tank, and two tanks per dose. The EPA obtained clarification from the study
authors regarding the experimental set-up pertaining to the biomass loading rate, which was 1 to
1.5 g/L (based on four fish weighing a total of 2 to 3 g per 2 L tank (personal communication
with Greg Goss and Rainie Sharpe, March 2021). This biomass loading rate was nearly two-fold
higher than that stated in OECD Guidelines of 0.8 g/L (OECD 1992). The trout were randomly
assigned to doses defined as control (0 mg/L PFOS); vehicle control (0 mg/L PFOS, 0.2%
MeOH v/v); and 0.78, 1.56, 3.12, 6.25, and 12.5 mg/L PFOS. The authors indicated that
measured PFOS concentrations averaged 88% of nominal but did not indicate whether LCsos
were based on measured or nominal concentrations. Given the clarifications regarding the
biomass loading, this study was considered for quantitative use in the derivation of the acute
PFOS freshwater criterion. The author-reported 96-hour LCso for the study of 2.5 mg/L (authors
did not specify if this concentration was nominal or measured) was acceptable for quantitative
A-16
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use and was among other toxicity values used for this species to calculate the SMAV/GMAV
(see further details at the end of the study summaries in this section) that was utilized to derive
the freshwater acute water column criterion.
Palmer et al. (2002a) evaluated the acute effects of PFOS (PFOS, potassium salt,
identified as FC-95 obtained from 3M Company) to Oncorhynchus mykiss, rainbow trout, via a
96-hour static exposure with measured concentrations. The test organisms were obtained from
Thomas Fish Company in Anderson, California and were reported as juveniles with a mean
weight of 0.34 g and mean total length of 3.6 cm. All test organisms were from the same source
and year class, and the length of the longest fish was no more than twice the length of the
shortest. The fish were held for approximately five weeks prior to the initiation of the test. This
acclimation was done in water from the same source and at the same temperature as the test.
During the acclimatation period, no mortalities or signs of disease were observed. Test
organisms were only fed a commercially-prepared diet (reported from Zeigler Brothers Inc.)
during a 14-day holding period after which point fish were no longer fed through the acclimation
period (at least 48 hours prior to the test) or during the test. The test water was obtained from a
well located near the testing facility and was characterized as moderately-hard water. The target
test temperature was 12 ฑ 1ฐC and a 16:8-hour light:dark photoperiod was maintained through
the holding, acclimatation, and testing periods. Dissolved oxygen and pH measurements were
made on water samples collected at test initiation followed by 24-hour intervals for each
replicate test chamber of each treatment and control. Test chambers were 25-L polyethylene
aquaria containing 15 L of test solution. At the initiation of the test, rainbow trout were
indiscriminately moved from the acclimation tank and distributed two at a time to the test
chambers until each contained ten fish. The resulting biomass loading rate was 0.23 g fish/L of
A-17
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test water. A 40-L stock solution was prepared in dilution water at a concentration of 150 mg
PFOS/L. Nominal concentrations were 3.1, 6.3, 13, 25, and 50 mg/L. Two replicates of each test
solution were prepared at nominal concentrations by adding the appropriate volume of stock
solution to dilution water in the test aquaria to achieve the final volume of 15 L. Measured test
concentrations at the end of the test ranged from 97 to 100% of nominal with concentrations of
3.0, 6.3, 13, 25, and 50 mg/L. Results from this study were based on measured concentrations.
Mortality and other signs of toxicity were observed daily. Trout in the control group appeared
normal and healthy throughout the test period. Additionally, test organisms in the lowest
treatment groups (3.0 and 6.3 mg/L) appeared healthy with no mortalities or other signs of
toxicity. After 96-hours of exposure, mortality in the 13, 25, and 50 mg/L treatment groups was
20, 50, and 100%, respectively. The author-reported 96-hour LCso for the study was 22 mg/L.
This study was considered acceptable for quantitative use in the derivation of the acute PFOS
freshwater water column criterion. The independently-calculated 96-hour LCso value was 22.59
mg/L.
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A.2.4.1 Sharpe el al. (2010) Concentration Response Curve - Oncorhynchus (trout)
Publication: Sharpe et al. (2010)
Species: Rainbow trout, Oncorhynchus mykiss
Genus: Oncorhynchus
EPA-Calculated LCso: Not calculable, concentration-response data not available
A.2.4.2 Palmer et al. (2002a) Concentration Response Curve - Oncorhynchus (trout)
Publication: Palmer et al. (2002a)
Species: Rainbow trout, Oncorhynchus mykiss
Genus: Oncorhynchus
EPA-Calculated LCso: 22.59 (95% C.I. 14.53 - 30.65) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
2.3775
0.5634
4.2189
2.455 e5
d
1.0339
0.0140
73.6910
< 2.2 e"16
e
26.3557
5.7449
4.5877
4.482 e6
Concentration-Response Model Fit:
Palmer et al. 2002
Oncorhynchus mykiss
Weibull type 1.3 para
PFOS (mg'L)
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A.2.5 Fifth Most Sensitive Freshwater Genus for Acute Toxicity: Ligumia (mussel)
Hazelton (2013); Hazelton et al. (2012) evaluated the acute effects of PFOS (acid form,
> 98% purity) on two freshwater mussels: Ligumia recta and Lampsilis siliquoidea. The tests
yielded the 4th and 7th most sensitive genera values (respectively) in the PFOS freshwater acute
criterion dataset (the L. siliquoidea results are reported below). Acute toxicity was observed
under either static or renewal conditions over a 24-hour period (< 24-hour old glochidia; static
exposure) or a 96-hour period (4-6-week-old juveniles; renewal exposure). The tests followed
the (ASTM 2006) test method. Dilution water was hard reconstituted water (total hardness
typically 160-180 mg/L as CaCCb). 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 control treatment). Nominal concentrations were 0 (negative
control), 0.005, 0.05, 0.5, 5, 50, and 500 mg/L; respective measured concentrations were < limit
of quantitation (LOQ; specifics not provided), 0.0054, 0.0514, 0.456, 4.68, 47.2, and 490 mg/L.
Recovery of PFOS standards ranged from 85.3-123% over all experiments. For all acute tests,
alkalinity ranged from 97 to 110 mg CaC03/L with a mean of 104.4 mg CaC03/L; total hardness
ranged from 132 to 162 mg CaC03/L with a mean of 149.6 mg CaC03/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 D.O. 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 in environmental
chambers set at a temperature of 20ฐC (glochidia tests), or in dilution water maintained at 20ฐC
(juvenile tests). Survival of mussels in the negative control was > 90% in all exposures. The 24-
hour ECso reported by the study authors for glochidia of L. rectawas 13.5 mg/L (C.I. 5.7-31.8).
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The 96-hour LC50 reported by the study authors for juvenile L. recta was 141.7 mg/L (C.I. 80.4-
249.6). The 24-hour EC50 for L. recta glochidia represented an acute value acceptable for
quantitative use. The juvenile life stage was less sensitive, such that its LC50 is not used
quantitatively in the SMAV for the species. An independently-calculated toxicity value could not
be calculated at this time given the lack of data presented in the paper. The study author reported
values are currently used quantitatively to derive the freshwater acute water column criterion.
A.2.5.1 Hazelton et al. (2012) Concentration Response Curve - Ligumia (mussel)
Publication: Hazelton et al. (2012)
Species: Black sandshell, Ligumia recta
Genus: Ligumia
EPA-Calculated LC50: Not calculable, concentration-response data not available
A.2.6 Sixth Most Sensitive Freshwater Genus for Acute Toxicity: Neocaridina (shrimp)
Li (2009) conducted three independent repeats of a 96-hour static test on PFOS
(potassium salt, >98% purity) with the freshwater shrimp species, Neocaridina denticulata (a
non-North American species). 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. 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 shrimp each
in at least five test concentrations (the first repeated experiment had one additional PFOS
treatment group at 10 mg/L compared to the other two experimental repeats) plus a negative
control. Each treatment was tested three independent times. Nominal test concentrations were in
the range of 5-200 mg/L PFOS. The test temperature was maintained at 25ฑ2ฐC. Water quality
parameters including pH, conductivity, and D.O. were reported as having been measured at the
A-21
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beginning and end of each test, but the information was not reported. Survival of negative
control animals was 90%. The study reported 96-hour LC50 was 10 mg/L (C.I. 9-12). The
toxicity test was acceptable for quantitative use. The independently-calculated LC50 values for
the three independent experimental repeats were 12.91 (10.29 - 15.53), 28.55 (15.05 - 42.05),
10.32 (7.788 - 12.85) mg/L, respectively. These independently-calculated LC50 values were used
to calculate the SMAV/GMAV value (as the geometric mean of the three LC50 values previously
mentioned) of 15.61 mg/L and was used to derive the freshwater acute water column criterion.
A.2.6.1 Li (2009) Concentration Response Carve - Neocaridina (shrimp)
Publication: Li (2009)
Species: Japanese Swamp Shrimp, Neocaridina denticulata
Genus: Neocaridina
EPA-Calculated LCsos: 12.91 (95% C.I. 10.29 - 15.53), 28.55 (95% C.I. 15.05 - 42.05),
10.32 (95% C.I. 7.788 - 12.85) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
-1.5141
0.1920
-7.8879
3.091 e15
e
10.1360
1.0252
9.8865
< 2.2 e"16
Concentration-Response Model Fit: In order of LCsos listed immediately above
Li 2009
Neocaridina denticulate
Weibufl type 2,2 para
PFOS (mg'L )
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Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.0404
0.2369
4.3919
1.124 e5
d
0.8880
0.0571
15.5493
< 2.2 e16
e
40.6105
7.5714
5.3637
8.156 e8
10 100
PFOS(mg'L)
: 0.50 -
OJ25
Li 2009
Neocaridina denticulate
Weibufl type 1, 3 para
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Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
0.7749
0.1332
5.8195
5.903 e9
e
16.5563
3.3654
4.9196
8.672 e7
Li 2009
Neocaridina denticulate
WeibuH type 1. 2 para
PFOS ( mg'L )
A.2.7 Seventh Most Sensitive Freshwater Genus for Acute Toxicity: Xenopus (frog)
Palmer and Krueger (2001), associated with Wildlife International, conducted three
good laboratory practice (GLP) renewal definitive assays with the potassium salt of PFOS
(86.9% purity) using the frog embryo teratogenesis assay - Xenopus (FETAX) with Xenopus
laevis. A primary PFOS stock solution was prepared in FETAX solution at a concentration of 48
mg/L, and subsequently diluted with FETAX solution to prepare the six nominal test
concentrations (1.82, 3.07, 5.19, 8.64, 14.4 and 24.0 mg PFOS/L). Eggs were obtained from
breeding colonies of X laevis at the University of Maryland Wye Research and Education
Center. Adults were held in flow-through polyethylene aquaria with 10 cm of dechlorinated tap
water (23.5ฑ0.5ฐC) and a maximum of 10 adults/chamber and photoperiod of 16:8-hours
light:dark. They were bred in the dark following injection of human chorionic gonadotropin to
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dorsal lymph sac of males and females. During each assay, X. laevis embryos (between NF
stages 8-11) were exposed to PFOS for 96 hours. Two replicate test chambers were maintained
in each treatment group and four replicates were maintained in each control group from the three
separate assays. Each test chamber contained 25 embryos for a total of 50 embryos per treatment
group and 100 embryos per control group. Tests were conducted at 24ฐC, pH of 7.26-7.30,
estimated total hardness of 75 mg/L as CaCCb, D.O. of 7.8-8.1 mg/L and a 12:12-hour light:dark
photoperiod (60-85 foot candles). PFOS concentrations were measured at the initiation and
termination of all three assays. The authors reported 96-hour LCso values for mortality of 13.8,
17.6 and 15.3 mg/L PFOS, teratogenesis ECsos of 12.1, 17.6 and 16.8 mg/L PFOS, and minimum
concentration to inhibit growth values (effectively a LOEC) of >14.7, 7.97 and 8.26 mg/L for the
same three tests, respectively. Independently-calculated 96-hour LCso values for mortality were
15.53 (13.86 - 17.21), 18.04 (15.33 - 20.74), and 14.60 (12.65 - 16.55) mg/L for the three
assays, respectively. No additional quantitative, acute toxicity data were available for this
species. Therefore, these independently-calculated LCso values were used to calculate the
SMAV/GMAV value (as the geometric mean of the three LCso values) of 15.99 mg/L that was
used to derive the freshwater acute water column criterion.
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A.2.7.1 Palmer andKrueger (2001) Concentration Response Curve - Xenopus (frog)
Publication: Palmer and Krueger (2001)
Species: Frog, Xenopus laevis
Genus: Xenopus
EPA-Calculated LCso: 15.53 (95% C.I. 13.86 - 17.21), 18.04 (95% C.I. 15.33 - 20.74),
14.60 (95% C.I. 12.65 - 16.55) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
4.1306
1.0464
3.9475
7.897 e5
d
0.9633
0.0149
64.7242
< 2.2 e"16
e
16.9770
0.7991
21.2452
< 2.2 e"16
Concentration-Response Model Fit: In order ofLCsos listed immediately above
Palmer and Krueger 2001
Xenopus laevis
WeibuD type 1. 3 para
PFOS ( mg'L )
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Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
1.8800
0.3458
5.4366
5.431 e8
d
0.9868
0.0127
77.7694
< 2.2 e"16
e
21.9190
1.9259
11.3812
< 2.2 e"16
Palmer and Krueger 2001
Xenopus laevis
Weibufl type 1, 3 para
PFOS ( mg'L)
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Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
-1.9934
0.2667
-7.4757
7.681 e14
e
12.1461
0.7197
16.8763
< 2.2 e"16
Palmer and Krueger 2001
Xenopus laevis
WeibuD type 2, 2 para
PFOS ( mgl)
A.2.8 Eighth Most Sensitive Freshwater Genus for Acute Toxicity: Lampsilis (mussel)
Hazelton (2013); Hazelton et al. (2012) evaluated the acute effects of PFOS (acid form,
> 98% purity) on two freshwater mussels, as noted above: Lampsilis siliqaoidea and Ligumia
recta. The L. siliquoidea studies yielded the 7th most sensitive genus value in the PFOS
freshwater acute criterion dataset. Hazelton et al.'s experimental design and study conditions for
L. siliquoidea were reported above under the description of the fourth most sensitive taxa,
Ligumia. The 24-hour EC so reported by the study authors for glochidia of L. siliquoidea was 16.5
mg/L (C.I. 8.0-33.9). The 96-hour LCso reported by the study authors for juvenile L. siliquoidea
was 158.1 mg/L (C.I. not calculable). The 24-hour ECso for L. siliquoidea glochidia represented
an acute value acceptable for quantitative use for the mussel species. Because the juvenile life
stage was less sensitive, only the glochidia LCso was used to calculate the SMAV. An
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independently-calculated toxicity value could not be calculated at this time given the data
presented in the paper. The study author reported value was used quantitatively to derive the
freshwater acute water column criterion.
A.2.9 Ninth most Sensitive Freshwater Genus for Acute Toxicity: Hyla (frog)
Tornabene et al. (2021) conducted acute toxicity tests with the gray treefrog, Hyla
versicolor, and PFOS (purchased from Sigma Aldrich, Catalog # 77282-10G; purity not
provided). The acute tests followed standard 96-hour acute toxicity test guidance (ASTM 2017;
U.S. EPA 2002). The frog was 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 (GS) 40
was run to determine if toxicity varied between life stages. Before test initiation larvae were
acclimated to test conditions (21ฐC and 12:12-hour light:dark photoperiod) for 24 hours. A stock
solution of PFOS (500 mg/L) was made in UV-filtered well water and diluted with well water to
reach test concentrations (ranged from 0 - 500 mg/L PFOS). 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 solutions and were not fed during the exposure period. The number of replicates
varied by life stage and treatment; 10 replicates for each treatment for GS 26 larva, and five to
six replicates for each treatment for GS 40 frogs. No mortality occurred in the controls of the GS
26 test and two of the six frogs died in the controls of the GS 40 test. The author reported 96-
hour LCsos were 79 and 24 mg/L PFOS for GS 26 and 40, respectively. The independently-
calculated 96-hr LC50 values were 78.33 and 19.88 mg/L and are acceptable for quantitative use.
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Given that GS 40 frogs appear to be a more sensitive life-stage the LC50 of 19.88 (13.80 - 25.95)
mg/L was utilized in the derivation of the freshwater acute water column criterion.
A.2.10 Tenth Most Sensitive Freshwater Genus for Acute Toxicity: Dugesia (planarian)
Li (2008) conducted three independent repeats of a 96-hour static, unmeasured acute
toxicity test on the potassium salt of PFOS (CAS # 2795-39-3, > 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. Dilution water was dechlorinated tap water. The photoperiod consisted of 12 hours of
illumination at an unreported intensity. A primary stock solution was prepared in dilution water.
Exposure vessels were polypropylene beakers of unreported dimensions and 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 10-200 mg/L PFOS. The test
temperature was maintained at 25 ฑ1ฐC. No other water quality parameters were reported for test
solutions. Survival of negative control animals was not reported. The author-reported 96-hour
LC50 was 17 mg/L (C.I. 16-18). The toxicity value could not be independently calculated at this
time given the level of data that was presented in the paper. The author-reported value was used
quantitatively to derive the freshwater acute water column criterion.
Li (2009) conducted three independent repeats of a second 96-hour static, unmeasured
acute test of PFOS (potassium salt, > 98% purity) with Dugesia japonica. The tested individuals
were also 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 0.9ฑ0.1 cm at test initiation. Dilution water was dechlorinated tap water. The
photoperiod consisted of 12 hours of illumination at an unreported intensity. A primary stock
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solution was prepared in dilution water. Exposure vessels were polypropylene beakers of
unreported dimensions and 50 mL fill volume. Each of the three independent repeats employed
three replicates of 10 planarians each in at least five test concentrations plus a negative control.
Nominal test concentrations were in the range of 5-200 mg/L PFOS. The test temperature was
maintained at 25ฑ1ฐC. 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. Survival of negative control animals was also not reported. The author-reported 96-
hour LCso was 23 mg/L (C.I. 20-25). An independently-calculated LC50 could not be estimated
for the first and second independent tests (as EPA was unable to fit a model with significant
parameters), but was estimated for the third independent test as 22.68 (18.27 - 27.10) mg/L. This
acute value was acceptable for quantitative use and was used to derive the freshwater acute water
column criterion.
Yuan et al. (2014) also conducted a 96-hour unmeasured acute test on PFOS (potassium
salt, > 99% purity) with Dugesia japonica, under daily renewal conditions. D. japonica used for
testing were originally collected from a fountain in Quan HetouBoshan, China, and cultivated in
the laboratory for an unspecified time period before use. The planarians had a body length of 10-
12 mm at test initiation. Dilution water was aerated tap water. No details were provided
regarding photoperiod or light intensity. A primary stock solution was prepared by dissolving the
salt in DMSO. The control and exposed planarians received 0.005% DMSO (v/v). Exposure
vessels were beakers of unreported material type and dimensions with 50 mL fill volume. The
test employed three replicates of 10 planarians each in six test concentrations plus a solvent
control. Nominal test concentrations were 0 (solvent control), 10, 30, 35, 37.5, 40 and 45 mg/L
PFOS. The test temperature was reported as 20ฐC. No other water quality parameters were
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reported. Survival of solvent control animals was not reported. The author-reported 96-hour LC50
was 29.46 mg/L (C.I. 25.80 - 33.12). An independently-calculated toxicity value could not be
calculated at this time given the level of data that was presented in the paper. The author-
reported value was used quantitatively to derive the freshwater acute water column criterion.
The noted toxicity values provided in each study summary above (17.00, 22.68, and
29.46 mg/L), comprising of both author-reported and independently-calculated LC50 values,
were used to calculate the SMAV/GMAV value (as the geometric mean of the three LC50 values
previously mentioned) of 22.48 mg/L, which was used to derive the freshwater acute aquatic life
criterion.
A.2.11 Eleventh Most Sensitive Freshwater Genus for Acute Toxicity: Danio (zebrafish)
Ye et al. (2007) evaluated the acute effects of perfluorooctanesulfonic acid (PFOS, purity
not reported) to Danio rerio via a 96-hour static-renewal unmeasured exposure. The PFOS stock
solution of 480 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. Each test solution, at the selected concentration, was prepared by diluting the stock solution
with reconstituted water. No added solvents were used. The fish (AB strain) used in this
experiment 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
14:10-hour light:dark period in aquaria containing aerated natural water. The pH ranged from 7
to 8.5 and water temperature was maintained at 26 ฑ 1ฐC. Embryos were obtained from
spawning adults, usually 5 male and 3 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 test solution which were completely renewed daily by
transferring embryos into newly cleaned wells. Nominal exposure concentrations were 0
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(control), 10, 85, 100, 110, 160, 200 and 240 mg/L PFOS. 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 71 mg/L PFOS. The EPA was unable to independently calculate a 96-hour
LC50 value based on the level of data provided by the study authors in the paper. Therefore, the
author-reported LC50 value was used quantitatively to derive the freshwater acute water column
criterion.
Hagenaars et al. (2011b) exposed/), rerio embryos to the potassium salt of PFOS (CAS
# 2795-39-3, purity >97%) under static unmeasured conditions for 96 hours. The PFOS was
dissolved in medium-hard reconstituted laboratory water, which was aerated and kept at 26ฐC
until use (no solvent). Adult wild-type 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. Fertilized eggs were collected in egg traps within 30 minutes of
spawning. Eggs were transferred to the test solutions (nominal PFOS concentrations of 0.1, 0.5,
1, 5, 10 mg/L in the definitive ELS test and 1, 5, 10, 25, 50 and 100 mg/L in the 96-hour range-
finding test) 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 egg 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
treatment at the beginning of the experiment. The 24-well plates were covered with a self-
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adhesive foil, placed in an incubation chamber at 26 ฑ 0.3ฐC and subjected to a 14:10-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
58.47 mg/L PFOS, based on the results of the range-finding test. The author-reported value was
used quantitatively to derive the freshwater acute water column criterion.
Sharpe et al. (2010) examined the toxicity of PFOS-K (potassium salt, CAS # 2795-39-
3, 98% purity) and bioaccumulation of PFOS and its isomers on Danio rerio through three
different tests, a 96-hour renewal toxicity test on adults, a 48-hour renewal toxicity test on
embryos, and a chronic exposure test that evaluated maternal transfer and fecundity of PFOS
isomers. The 96-hour test is described in this appendix, as these results were used quantitatively
to derive the freshwater acute water column criterion. The 48-hour and chronic toxicity tests
were used qualitatively (Appendix G). The authors provided little detail about the test protocol,
other than following OECD Guideline 203. Adult zebrafish were obtained from a local pet store.
They were acclimated and held in 70 L glass aquaria in an environmental chamber set to 26ฐC
under a 14:10-hour light:dark photoperiod for six to 10 months prior to use in experiments.
Conditioned zebrafish water (ZF water) was obtained from the Biological Sciences Zebrafish
Facility at the University of Alberta, where an automated reverse osmosis system regulated both
the salinity and hardness (160 mg/L total hardness and 20 mg/L calcium carbonate hardness) of
the water. A stock solution of 25 mg/ml PFOS in MeOH was used for dosing in all experiments.
All vehicle controls received a volume of MeOH equal to that present in the highest PFOS
treatment of that experiment (final MeOH content 0.65% v/v). The concentration of PFOS in any
experiment was always well below its reported solubility in water (approximately 500 mg/L).
Zebrafish toxicity tests were performed using food grade 2 L plastic tanks with four fish per
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tank, and two tanks per concentration. Fish were randomly assigned to nominal concentrations
defined as control (0 mg/L PFOS); vehicle control (0 mg/L PFOS, 0.4% MeOH v/v); and 6.25,
12.5, 2, 50 and 100 mg/L PFOS. Authors indicated that measured PFOS concentrations averaged
88% of nominal but did not indicate whether the LCso was measured or nominal. The adult 96-
hour acute test followed OECD 203 protocol and was acceptable for quantitative use. The
author-reported LCso was 22.2 ฑ 4.6 mg/L for PFOS. The author-reported value was used
quantitatively to derive the freshwater acute water column criterion.
Evaluated the acute effects of perfluorooctane sulfonate, potassium salt (PFOS-K, CAS#
2795-39-3 purchased from Wellington Laboratories Inc., Ontario, Canada) on zebrafish (Danio
rerio) during a 96-hour unmeasured, static-renewal study. Zebrafish were purchased from a local
market at approximately three months in age and 2.2 cm in length. Fish were allowed to
acclimate for seven days and were fed three times per week until 24 hours before the test was
started. Water used for the testing was aerated for 48 hours before testing began, and testing
followed OECD Guideline 203. Observed exposure water characteristics were total hardness of
180-220 mg/L as CaC03, temperature of 23 ฑ 1ฐC, D.O. of 7.0 - 8.6 mg/L and a photoperiod of
12:12-hours light:dark. Each 2-L beaker was filled with 1,500 mL of test solution at nominal
concentrations of 0 (control), 2.87, 5, 8.7, 15.14, 26.34, 45.83 and 79.74 mg/L PFOS. There were
three replicates per concentration, and seven fish per beaker. Test solutions were renewed at 48
hours. The author-reported 96-hour LCso was 17.0 mg/L PFOS based on a sigmoidal three-
parameter regression. The EPA was unable to independently calculate a 96-hour LCso value
based on the level data provided in the paper. Therefore, the author-reported LCso value of 17.0
mg/L PFOS was used quantitatively to derive the freshwater acute water column criterion.
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Li et al. (2015) evaluated the acute effects of PFOS (CAS # 2795-39-3, 98% purity) to
Danio rerio via a 96-hour static unmeasured exposure. Solutions for waterborne exposure were
formulated with medium used to rear embryos (reconstituted laboratory water). Adult, wild-type
zebrafish were obtained from the Institute of Hydrobiology, at the Chinese Academy of Sciences
(Wuhan, China), and kept in treated tap water at 26 - 29ฐC. Fish were reared with a female/male
ratio of 1:2 under 14:10-hour light:dark photoperiod, with 1/3 of the water exchanged daily.
Spawning and fertilization took place within 30 minutes after the lights were turned on in the
morning, with fertilized embryos collected and cleaned with embryo rearing water. Immediately
after fertilization, embryos were examined, and damaged or unfertilized embryos were discarded.
Zebrafish embryos were exposed in 24-well cell culture plates (assume plastic) to a series of
PFOS concentrations (6.25, 12.5, 25.0, 50.0, 100.0 and 200.0 mg/L). Twenty, normally shaped,
fertilized embryos were assigned to each treatment (three replicates) and 2 mL of test solution
per well; the four remaining wells were assigned with control embryos. Embryos were exposed
in an incubator at 28.5ฐC, pH of 8.3 and a 14:10-hour light:dark photoperiod. Toxicological
endpoints included whether embryos were clear or opaque, exhibited edema at 4, 8, 24, 48, 72, or
96 hpf, or developed structural malformations at 72 or 96 hpf until hatching. Coagulated
embryos before hatching are opaque, milky white, and appear dark under the microscope.
Coagulation of embryos was recorded and used for the calculation of an LCso value. The author
reported 96-hour LCso was 68.0 mg/L PFOS. The independently-calculated LCso value was
71.12 (56.82 - 85.42) mg/L PFOS. This toxicity value is acceptable for quantitative use and was
used to derive the freshwater acute water column criterion.
Du et al. (2017); Du et al. (2016b) exposed Danio rerio to
heptadecafluorooctanesulfonic acid (PFOS, potassium salt, CAS# 2795-39-3, 98% purity) using
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static unmeasured procedures. Although the study focused on the protective effects of zinc
nanoparticles (ZnO-NPs) on PFOS toxicity (development and damage to DNA), data were also
reported for PFOS-only exposures. Adult AB strain zebrafish were purchased from State Key
Laboratory of Freshwater Ecology and Biotechnology, Chinese Academy of Sciences (Wuhan,
China). Fish were maintained and tested at 28ฐC under a 14:10-hour light:dark cycle. Male and
female fish were paired in spawning boxes overnight in rearing water and spawning was
completed at the beginning of the light cycle. Eggs were collected from the spawn traps and
transferred to clean rearing water prior to testing. The embryos were exposed to PFOS (1, 2, 4, 8
and 16 mg/L in a preliminary test to determine the LCso, and at 0.4, 0.8 and 1.6 mg/L in a later
test with ZnO-NPs solutions to evaluate mortality (at 96 hours), body length (at 96 hours), hatch
rate (at 72 hours), heart rate (at 48 hours), and malformation rate (at 96 hours). Embryos were
kept in 24-well multi-plates at two embryos/well, in which 20 wells contained 2 mL
reconstituted water test solution and four wells contained 2 mL of culture medium as the control.
Each plate contained 40 embryos for exposure testing and eight embryos as internal water
controls. For each concentration and water control, three 24-well plates (replicates) were
included. The study authors reported a 96-hour LCso of 3.502 mg/L for PFOS. The EPA was
unable to independently calculate a 96-hour LCso value based on the level data provided in the
paper. Therefore, the author-reported LCso value of 3.502 mg/L PFOS was used quantitatively to
derive the freshwater acute water column criterion.
Stengel et al. (2017b) exposed 1 hpf Danio rerio embryos to PFOS for 96 hours using
renewal unmeasured procedures as specified in (OECD 2013) guidelines. PFOS 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
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Aquatic Ecology and Toxicology breeding facilities at the University of Heidelberg. Details of
zebrafish maintenance, egg production and embryo rearing were as described previously
(Kimmel et al. 1995; Kimmel et al. 1988; Nagel 2002; Spence et al. 2006; Wixon 2000) but
included updates for the purpose of the zebrafish embryo toxicity test by Lammer et al. (2009).
Embryos no older than 1 hpf were exposed in glass vessels, which had been preincubated
(saturated) for at least 24 hours, to a series of PFOS dilutions (0, 3.125, 6.25, 12.5, 25, 50 mg/L).
After confirmation of fertilization success, embryos were individually transferred to the wells of
24-well plates, which had been pre-incubated with 2 mL of the test solutions per well for 24
hours prior to the test start and kept in an incubator at 26.0 ฑ1,0ฐC under a 14:10-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. Both lethal
and sublethal effects were used for the determination of EC values. The author-reported 96-hour
LCso and EC50 (combination of lethal and sublethal effects) values were 34.2 and 21.4 mg/L
PFOS, respectively. The independently-calculated LC50 was 38.82 (36.67 - 40.98) mg/L PFOS.
The independently-calculated LC50 was considered quantitative and used to derive the freshwater
acute water column criterion.
Nilen et al. (2022) evaluated the acute effects of perfluorooctanesulfonate, potassium salt
(PFOS-K, CAS No. 2795-39-3, > 98% purity, purchased from Sigma-Aldrich, Stockholm,
Sweden) on zebrafish (Danio rerio) during a 96-hour measured, static-renewal test. PFOS stock
solutions (464 mM) were prepared in DMSO. Adult zebrafish (AB strain) were purchased from
Karolinska Institute (Stockholm, Sweden). Fish were kept at a 14:10-hour light:dark cycle in a
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recirculating system. Fish embryo toxicity tests were performed following the OECD Technical
Guideline No. 236 (OECD 2013). Prior to testing, 96-well plates were pre-incubated for 24 hours
with the chemical solutions to limit sorption to the well plates. On the day of exposure, five
different exposure concentrations were prepared in glass beakers through a serial dilution (1:2)
using ISO water. The DMSO content was adjusted to 0.09% (v/v) in all solutions and all tests
included three control groups: negative control (ISO-water), solvent control (0.09 % DMSO in
ISO-water) and positive control (4 mg/L DCA). Test concentrations were selected based on
previous studies about lethality and sub-lethality. One embryo (64-cell stage) with 250 |iL of test
solution was added to each well on a 96-well plate. For each concentration, 24 embryos were
used, and each test contained a total of 192 embryos. The plates were covered with self-adhesive
film and incubated at 26 ฑ 1ฐC in a 14:10-hour light:dark cycle. An experiment was considered
valid by the study authors if the mortality in the negative control and the solvent control was less
than 10%. In addition, the mortality of the positive control had to be > 30%. Each experiment
was repeated three to five times representing independent replicates and the embryos were
exposed for 96 hpf. The exposure solutions were sampled before the start and at the end of the
exposure period. Nominal exposure concentrations were 0 (control), 0 (solvent control), 7, 14,
28, 56 and 111.5 |iM PFOS-K. Equivalent PFOS concentrations were 3.768, 7.535, 15.07, 30.14
and 60.01 mg/L PFOS-K, respectively, based on a molecular weight of 538.22 g/mol PFOS-K.
Measured concentrations were between 1.3 and 23% of nominal at 96 hours. The author-reported
96-hour LCso was 44.57 |iM PFOS-K, or 23.99 mg/L PFOS-K (based on a molecular weight of
538.22 g/mol; 44.57 x 538.22 ^ 1,000). The EPA was unable to independently calculate a 96-
hour LC50 value based on the level of data provided in the paper by the study authors. Therefore,
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the author-reported LC50 value of 23.99 mg/L PFOS was used quantitatively to derive the
freshwater acute water column criterion.
A.2.12 Twelfth Most Sensitive Freshwater Genus for Acute Toxicity: Dayhnia (cladoceran)
Logeshwaran et al. (2021) conducted acute and chronic toxicity tests with the
cladoceran, Daphnia cannula, and PFOS-K (perfluorooctanesulfonate potassium salt, > 98%
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 16:8-hour
light:dark photoperiod. The acute test protocol followed OECD (2000) with slight modifications.
A PFOS stock solution (20 mg/mL) was prepared in dimethylformamide and diluted with
deionized water to achieve a concentration of 200 mg/L PFOS. Cladoceran culture medium was
used to prepare the PFOS stock and test solutions. Ten daphnids (6-12 hours old) were
transferred to polypropylene containers containing one of 14 nominal test concentrations (0, 0.5,
1, 2.5, 5, 10, 15, 20, 25, 30, 35, 40, 45 and 50 mg/L PFOS). 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 8.8 mg/L PFOS. The independently-calculated
48-hour EC50 value was 11.56 (10.06 - 13.07) mg/L and is acceptable for quantitative use in the
derivation of the freshwater acute water column criterion.
Drottar and Krueger (2000g) reported the results of a 48-hour static, measured acute
toxicity test on PFOS (potassium salt, CAS # 2795-39-3, 90.49% purity) with Daphnia magna.
The GLP test was conducted at Wildlife International, Ltd. In Easton, MD in February, 1999.
The test followed OECD (2004); (U.S. EPA 1996d). The test organisms were less than 24 hours
old at test initiation. Dilution water was 0.45 |im filtered well water [hardness: 132 (128-136)
mg/L as CaC03; alkalinity: 178 (176-178) mg/L as CaC03; pH: 8.3 (8.2-8.3); TOC: < 1.0 mg/L;
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conductivity: 313 (310-315) |imhos/cm]. Photoperiod was 16:8-hours light:dark with a 30
minute transition period. Light was provided at an intensity of approximately 359 lux. A primary
stock solution was prepared in dilution water at 91 mg/L. It was mixed for -19.5 hours prior to
use. After mixing, the primary stock was proportionally diluted with dilution water to prepare the
four additional test concentrations. Exposure vessels were 250 mL plastic beakers containing 240
mL of test solution. The test employed two replicates of 10 daphnids each in five measured test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 12, 20,
33, 55, 91 mg/L. Mean measured concentrations were
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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. 2003a). The test followed ASTM
(1999a). Daphnids used for testing were less than 24 hours old at test initiation. D. 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.
Hardness was softened by addition of distilled deionized water to achieve a range of 200-225
mg/L of CaC03. Photoperiod was 16:8-hours light:dark 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 filled with 150 mL of test solution. All
toxicity testing involved four replicates of 10 daphnids each in five nominal test concentrations
plus a negative control. Nominal concentrations were 0 (negative control), 31, 63, 125, 250 and
450 mg/L. Experiments were conducted in environmental chambers at a test temperature of 21
ฑ1ฐC. Authors note temperature and pH were measured at beginning and end of study, but the
information was not reported. Survival of daphnids in the negative control was also not reported,
although ASTM E729-96 requires at least 90% survival for test acceptability. The author-
reported 48-hour ECso for D. magna was 61.2 mg/L (C.I. 31.3-88.5). The author-reported 48-
hour EC50 for D. pulicaria was 134 mg/L (C.I. 103-175). Independently-calculated toxicity
values could not be calculated given the level of data that was presented in the paper. The study
author reported values were used quantitatively to derive the freshwater acute water column
criterion.
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Ji et al. (2008) performed a 48-hour static, unmeasured acute test of PFOS (acid form,
CAS # 1763-23-1, purity unreported) with Daphnia magna. The test followed the EPA's
Methods for measuring the acute toxicity of effluents and receiving waters to freshwater and
marine organisms (U.S. EPA 2002). D. magna used for testing were obtained from brood stock
cultured at the Environmental Toxicology Laboratory at Seoul National University (South
Korea). Test organisms were less than 24 hours old at test initiation. Dilution water was
moderately hard reconstituted water (hardness typically 80-100 mg/L as CaCCb). Experiments
were conducted in glass jars of unspecified size and fill volume. Photoperiod was assumed by the
reviewers to have been 16:8-hours light:dark. Preparation of test solutions was not described.
The test involved four replicates of five daphnids each in five nominal test concentrations plus a
negative control. Nominal concentrations were 0 (negative control), 6.25, 12.5, 25, 50 and 100
mg/L. Test temperature was maintained at 21 ฑ 1ฐC. Authors note water quality parameters (pH,
temperature, conductivity, and D.O.) were measured 48 hours after exposure, but the information
is not reported. Survival of daphnids in the negative control was not reported in the paper.
However, raw data were obtained by the EPA from the study authors and control survival was
100% and therefore meets the EPA/600/4-90/027F requirement of at least 90% survival for test
acceptability. The author-reported 48-hour EC so value for the study was 37.36 mg/L (C.I. 30.72-
43.99) for D. magna. The 48-hour ECso value was independently-calculated as 35.46 (28.26 -
42.66) mg/L. The independently-calculated acute toxicity value was included in the derivation of
the freshwater acute water column criterion.
Li (2009) conducted three independent repeats of a 48-hour static acute test on PFOS
(potassium salt, > 98% purity) with Daphnia magna. The test followed OECD (1984). D. magna
used for the test were less than 24 hours old at test initiation. Dilution water was dechlorinated
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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 at least five test concentrations plus a negative control.
Each treatment was tested three independent times. Based on water solubility of test chemicals
and preliminary toxicity results, nominal test concentrations were in the range of 10 - 400 mg/L
for PFOS. Water quality parameters including water pH, conductivity and D.O. were measured
at the beginning and at the end of each test. Initial values of pH were 7.82ฑ0.12 and 7.91ฑ0.03
after 48 hours. At the start of the bioassays, D.O. and specific conductivity were 67.7ฑ6.8% and
101.8ฑ6.8 |iS/cm, After the 48-hour testing period, D.O. and specific conductivity were
55.6ฑ1.26% and 109.1ฑ3.5 |iS/cm. The D.O. after 4 8-hours of testing was lower than the test
guideline recommendation of >60% (ASTM 2002; U.S. EPA 2016a; U.S. EPA 2016b); however,
it was not low enough to change the use of the study. The test was conducted in a temperature
incubator at 25 ฑ2ฐC. None of the control animals became immobile at the end of the test. The
author-reported 48-hour EC so was 63 mg/L (C.I. 58-69) which was an average LCso of the three
tests. The independently-calculated LCso values for the three independent experimental repeats
were 55.40 (45.97), 72.70 (61.63 - 83.77) and 64.60 (49.53 - 79.66) mg/L, respectively. The
three independently-calculated LCso values were used to calculate the SMAV for D. magna and
derive the freshwater acute water column criterion.
Yang et al. (2014) conducted a 48-hour static acute test of PFOS (potassium salt, CAS #
2795-39-3, 99% purity) with Daphnia magna, following ASTM (1993). I). magna 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,
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7.0ฑ0.5; D.O., 7.0ฑ0.5 mg/L; total organic carbon, 0.02 mg/L; and hardness, 190.0ฑ0.1 mg/L as
CaC03). Photoperiod was 12:12-hours light:dark at an unreported light intensity. A primary
stock solution was prepared by dissolving PFOS in deionized water and cosolvent DMSO. The
primary stock was proportionally diluted (0.56x) with dilution water to prepare the test
concentrations. Exposure vessels were 200 mL beakers of unreported material type containing
200 mL of test solution. The test employed three replicates of 10 daphnids each in six test
concentrations (measured in low and high treatments) plus a negative and solvent control.
Nominal concentrations were 0 (negative and solvent controls), 20.00, 36.00, 64.80, 116.64,
209.95 and 377.91 mg/L. Mean measured concentrations before and after renewal were
respectively 18.43 and 19.80 and mg/L (lowest concentration) and 341.74 and 372.35 mg/L
(highest concentration). Analyses of test solutions were performed using HPLC/MS and negative
electrospray ionization. The concentration of PFOS 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 105%. The
temperature, D.O., and pH were reported has having been measured every day during the acute
test, but results are not reported. Negative control survival was > 96%. Solvent control survival
was 100%). The author reported 48-hour LCso was 78.09 mg/L (C.I. 54.38-112.13) and was used
quantitatively to derive the freshwater acute water column criterion.
Lu et al. (2015) conducted a 48-hour static test on PFOS (purity 98%>) with Daphnia
magna, following OECD (2004). D. magna used for the test were originally obtained from the
Chinese Center for Disease Control and Prevention (Beijing, China) and cultured in the
laboratory according to the International Organization for Standardization (ISO 1996). Daphnids
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were less than 24 hours old at test initiation. Dilution water was the same used for daphnid
culture and was reconstituted according to OECD (2004) with a hardness of 250 mg/L as CaCCb,
as calculated based on the recipe provided, and pH ranging from 7.7 to 8.4. Photoperiod was
16:8-hours light:dark at an unreported light intensity. The test solution was prepared immediately
prior to use by diluting the stock solution with a daphnia culture medium. Exposure vessels were
100 mL glass beakers containing 45 mL of test solution. The test employed three replicates of 10
daphnids each in six nominal test concentrations plus a negative control. Nominal concentrations
were 0 (negative control), 1,3, 10, 30, 100 and 300 mg/L. Exposure water quality was checked
daily and maintained at a temperature of 20 ฑ 1ฐC, pH of 7.2 ฑ 0.3, and D.O. of 6.5 ฑ 0.5 mg/L.
One hundred percent survival was observed at 48 hours in the negative control. The author-
reported 48-hour ECso was 23.41 mg/L (LC50 = 49.27) and used quantitatively to derive the
freshwater acute water column criterion.
Liang et al. (2017) conducted a 48-hour static test on PFOS (potassium salt, CAS #
2795-39-3, > 98% purity) with Daphnia magna. The test followed OECD (2004). I). magna used
for the test were originally obtained from State Key Laboratory of Environmental Aquatic
Chemistry (Eco-Environmental Sciences of Chinese Academy of Sciences, Beijing) and cultured
in the laboratory according to Revel et al. (2015). Daphnids were less than 24 hours old at test
initiation. Dilution water was artificial medium (M4) at 20ฐC and pH 7 (Revel et al. 2015).
Photoperiod was 16:8-hours light:dark at an unreported light intensity. The test solution was
prepared immediately prior to use by diluting the stock solution with M4 medium. Exposure
vessels were 80 mL beakers of unreported material type containing an unspecified volume of test
solution. The test employed five replicates of five daphnids each in six nominal test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 30, 44,
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66, 100 and 150 mg/L. No mention was made of water quality being checked during the
exposure. 100% survival was observed at 48 hours in the negative control. The author reported
48-hour ECso was 79.35 mg/L. The independently-calculated toxicity value was 94.58 (94.20 -
94.96) mg/L and was used quantitatively to derive the freshwater acute water column criterion.
Yang et al. (2019) evaluated the acute effects of perfluorooctane sulfonate, potassium
salt (PFOS-K, CAS# 2795-39-3, 98% purity, purchased from Sigma-Aldrich in St. Louis, MO)
on Daphnia magna via a 48-hour unmeasured, static mortality test. D. magna cultures were
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 laid out in
OECD Guideline 202, and all testing followed the guideline. Cultures were fed green algae daily
and were acclimated for two to three weeks before testing. Acute test concentrations included 0
(control), 0.0000156, 0.0000234, 0.0000349, 0.0000788 and 0.000118 mol/L (or 0 (control),
8.396, 12.59, 18.78, 28.31, 42.41, and 63.51 mg/L given the molecular weight of the form of
PFOS used in the study, CAS # 2795-39-3, of 538.22 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 LC50 of 22.77 mg/L. The EPA's independently-calculated 48-hour LC50 was 22.43
(15.74 - 29.12) mg/L PFOS and was used quantitatively to derive the freshwater acute water
column criterion.
A.2.13 Thirteenth Most Sensitive Freshwater Genus for Acute Toxicity: Ambystoma
(salamander)
Tornabene et al. (2021) conducted acute toxicity tests with three species of salamanders
in the genus Ambystoma and PFOS (purchased from Sigma Aldrich, Catalog # 77282-10G;
purity not provided). Acute tests followed standard 96-hour acute toxicity test guidance (ASTM
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2017; U.S. EPA 2002). The three test species (Jefferson salamander, Ambystomajeffersonianum\
small-mouthed salamander, A. texanum\ eastern tiger salamander, A. tigrinum) 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 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 12:12-hour
light:dark photoperiod) for 24 hours. An additional acute test with Harrison stage 46 small-
mouthed salamanders was run to determine if toxicity varied between life stages. A stock
solution of PFOS (500 mg/L) was made in UV-filtered well water and diluted with well water to
reach test concentrations (ranged from 0-500 mg/L PFOS). 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 solutions 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 46 small-mouthed salamander, 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. No mortality occurred in any of the control groups.
Author-reported 96-hour LCsos were 64, 41 and 73 mg/L PFOS for the Jefferson salamander,
small-mouthed salamander and eastern tiger salamander, respectively. The authors did not find a
significant difference between the life stages of small-mouthed salamander so results of the two
tests were pooled. The independently-calculated 96-hour LCso values were 51.71 (40.84 - 62.58)
for Jefferson salamander; 46.71 (34.33 - 59.09) and 30.00 (27.14 - 32.86) for small-mouthed
salamander, Harrison stage 40 and 46; and 68.63 (61.90 - 75.37) mg/L for eastern tiger
A-48
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salamander, respectively. In general, the independently-calculated toxicity values were
acceptable for quantitative use and used to derive the freshwater acute water column criterion for
PFOS. However, only the LC50 value of 30.00 mg/L for Harrison stage 46 small-mouthed
salamander was used for this species as this life stage was determined to be the most sensitive.
A.2.14 Fourteenth Most Sensitive Freshwater Genus for Acute Toxicity: Pontastacus (crayfish)
Belek et al. (2022) tested heptadecafluorooctanesulfonic acid potassium salt (PFOS-K)
on narrow-clawed crayfish (Pontastacus leptodactylus; formerly, Astacus leptodactylus) for 96
hours in an unmeasured, static experiment. Heptadecafluorooctanesulfonic acid potassium salt
(>98% purity) was obtained from Sigma-Aldrich (USA). Crayfish were obtained from a local
breeder in Lake Egirdir, Turkey during the inter-molt stage, and averaged 29.1ฑ0.39 g weight
and 10.27ฑ0.05 cm length. Crayfish were acclimated in the laboratory at the Biology Department
of Gazi University for two weeks in glass aquariums filled with aerated, dechlorinated tap water
for two weeks at a temperature of 21ฐC, and were fed a daily ration of raw trout. Testing
parameters followed animal care guidance from NRC (1996). A 96-hour acute test was
conducted to determine the LC50, and a subsequent 21-day chronic test was conducted with a
water only control, a DMSO solvent control, and two treatments set to 0.5 and 5 mg/L, which
approximated 1% and 10% of the acute LC50. The chronic test measured biochemical and
enzymatic responses to PFOS. Mean water quality parameters for test water were as follows:
21ฑ1ฐC temperature, 6.79ฑ0.31 pH, 117.38ฑ17.20 |iS/cm specific conductivity, 0.01ฑ0.001
mg/L total ammonia nitrogen, and 6.72ฑ0.05 mg/L dissolved oxygen. Organisms were kept
under a 16:8 light:dark cycle in 15 L of aerated, dechlorinated tap water during testing. The LC50
was calculated using Probit analysis using the U.S. EPA LC50 Software Program version 1.00.
The author reported 96-hour LC50 value of 48.81 mg/L was determined to be acceptable for
quantitative use.
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A.2.15 Fifteenth Most Sensitive Freshwater Genus for Acute Toxicity: Anaxyrus (toad)
Tornabene et al. (2021) conducted acute toxicity tests with the American toad, Anaxyrus
americanus, and PFOS (purchased from Sigma Aldrich, Catalog # 77282-10G; purity not
provided). The acute tests followed standard 96-hour acute toxicity test guidance (ASTM 2017;
U.S. EPA 2002). The frog was 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 toads reached Gosner stage (GS) 26, defined as
when larvae are free swimming and feeding. An additional acute test with GS 41 was run to
determine if toxicity varied between life stages. Before test initiation larvae were acclimated to
test conditions (21ฐC and 12:12-hour light:dark photoperiod) for 24 hours. A stock solution of
PFOS (500 mg/L) was made in UV-filtered well water and diluted with well water to achieve test
concentrations ranging from 0 - 500 mg/L PFOS. 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
solutions and were not fed during the exposure period. The number of replicates varied by life
stage, and treatment; 10 replicates for each treatment for GS 26 toads, and six to 10 replicates for
each treatment for GS 41 toads. No mortality occurred in any of the control groups. The author-
reported 96-hour LCso was 62 mg/L PFOS. The authors did not find a significant difference
between the life stages of the American toad, so results of the two tests were pooled. The
independently-calculated 96-hour LC50 values were 63.41 (62.32 - 64.51) mg/L for the GS 26
toads and 56.49 (49.10 - 63.90) mg/L for GS 41 toads. Given that the GS 41 toads appear to be a
more sensitive life-stage the LC50 of 56.49 mg/L was considered acceptable for quantitative use
and used in the derivation of the freshwater acute water column criterion.
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A.2.16 Sixteenth Most Sensitive Freshwater Genus for Acute Toxicity: Procambarus (crayfish)
Funkhouser (2014) conducted a 7-day static acute test on PFOS (potassium salt, >98%
purity) with the crayfish species, Procambarus fallax (f. virginalis), as part of a Master's thesis
at the Texas Tech University, Lubbock, TX. Juvenile P. fallax (2-week old, 0.041 g) used for the
test were originally purchased from a private collector. The crayfish reproduced for several
generations before being used for experiments. Based on an average reproductive age of 141-255
days, an interclutch period of 50-85 days, and a brooding time of 22-42 days, the author
estimated the experimental animals to be stage F4-F6 (Seitz et al. 2005). Dilution water was
moderately hard reconstituted laboratory water (3.0 g CaSC>4, 3.0 g MgSC>4, 0.2 g KC1, and 4.9 g
NaFtCCb added to 50 L deionized water). Photoperiod was 14:10-hours light:dark at an
unreported light intensity. PFOS was dissolved in dilution water to prepare the test
concentrations. Exposure vessels were 1 L polypropylene containers containing 500 mL of test
solution. The test employed two replicates of three snails each in five test concentrations plus a
negative control. Nominal concentrations were 0 (negative control), 40, 80, 120, 160, and 200
mg/L. Exposure concentrations were reportedly measured, but concentrations were not reported.
Analyses of test solutions were performed using LC-MS/MS. Standards were used as part of the
analytical method, but details were not reported. The reporting limit was 0.010 mg/L.
Experiments were conducted in an incubator at 25 ฑ1ฐC and covered with plastic opaque
sheeting to limit evaporation. No other water quality parameters were reported as having been
measured in test solutions. Negative control survival was 100% after seven days. The author
reported 96-hour LCso was reported as 59.87 mg/L. For comparison, the 7-day LCso was 39.71
mg/L. The independently-calculated 96-hour LCso value was also 59.87 mg/L (C.I. 54.29-65.45).
This independently-calculated LCso value of 59.7 mg/L was used quantitatively to derive the
freshwater acute water column criterion.
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A.2.17 Seventeenth Most Sensitive Freshwater Genus for Acute Toxicity: Brachionus (rotifer)
Zhang et al. (2013) performed a 24-hour static test of PFOS (potassium salt, CAS #
2795-39-3, >98% purity) with Brachionus calyciflorus. Organisms 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 (16:8-hours 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 PFOS (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, 6-well cell culture plates (assumed plastic) each containing at total of
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), 40, 50, 60, 70, 80, 90, and
100 mg/L. PFOS 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 PFOS
measured every eight hours over a 24-hour period in rotifer medium with green algae incurs
minimal change in the concentration range 0.25 to 2.0 mg/L. The acute test was conducted
without green algae added to the exposure medium. One hundred percent survival was observed
at 24 hours in the negative control. The author-reported 24-hour LCso was 61.8 mg/L. The
author-reported value was used quantitatively to derive the freshwater acute water column
criterion.
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A.2.18 Eighteenth Most Sensitive Freshwater Genus for Acute Toxicity: Elliptio (mussel)
Drottar and Krueger (2000f) reported the results of a 96-hour renewal, measured test on
the effects of PFOS (potassium salt, CAS # 2795-39-3, 90.49% purity) on Elliptio complanata
(formerly known as Unio complanatus). The good laboratory practice (GLP) test was conducted
at Wildlife International, Ltd. In Easton, MD in August, 1999, using a protocol based on
procedures outlined in U.S. EPA (1996b). E. complanata (76.5 g and 48.7 mm) used for the test
were purchased from Carolina Biological Supply Company in Burlington, NC, after being caught
in the wild. They were of an unspecified age at test initiation. Dilution water was 0.45 |im
filtered well water [total hardness: 126 (120-132) mg/L as CaCCb; alkalinity: 174 (170-178)
mg/L as CaCCb; pH: 8.3 (8.1-8.5); TOC: <1.0 mg/L; conductivity: 21 (310-330) |imhos/cm].
Photoperiod was 16:8-hours light:dark with a 30-minute transition period. Light was provided at
an intensity of approximately 369 lux. A primary stock solution was prepared in dilution water at
91 mg/L. It was mixed for -24 hours prior to use. After mixing, the primary stock was
proportionally diluted with dilution water to prepare the four additional test concentrations.
Exposure vessels were 25 L polyethylene aquaria containing 20 L of test solution. The test
employed two replicates of 10 mussels each in five measured test concentrations plus a negative
control. Nominal concentrations were 0 (negative control), 5.7, 11, 23, 46, and 91 mg/L. Mean
measured concentrations were less <0.115 mg/L, 5.3, 12, 20, 41, and 79 mg/L, respectively.
Analyses of test solutions were performed at Wildlife International, Ltd. using high performance
liquid chromatography with mass spectrometric detection (HPLC/MS). The mean percent
recovery of matrix fortifications analyzed concurrently during sample analysis was 94.7%.
Concentrations measured at test initiation averaged 86% of nominal. Concentrations measured
prior to renewal at 48 hours averaged 89% of nominal. Concentrations measured at 96 hours
averaged 100% of nominal. Dissolved oxygen in control and the high-test concentration (79
A-53
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mg/L) respectively ranged from 5.8-8.5 mg/L and 5.0-8.6 mg/L; pH ranged from 8.0-8.4 and 7.9-
8.5. Test temperature ranged from 21.4-21.8ฐC and 21.8-23.7ฐC. Mussels in the negative control,
the 5.3, 12, and 20 mg/L treatments appeared healthy and normal throughout the test with no
mortality, immobility or overt clinical signs of toxicity. The author-reported 96-hour LCso was
59 mg/L (C.I. 51-68). The independently-calculated LCso value was 64.35 (56.22 - 72.48) mg/L
and used to derive the freshwater acute water column criterion.
A.2.19 Nineteenth Most Sensitive Freshwater Genus for Acute Toxicity: Lithobates (frog)
Flynn et al. (2019) evaluated the acute effects of perfluorooctanesulfonic acid (PFOS,
CAS# 1763-23-1, purchased from Sigma-Aldrich) on the American bullfrog (Lithobates
catesbeiana, formerly, Rana catesbeiana) during a 96-hour unmeasured, static 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 TetraMinฎ ad libitum and were acclimated to
laboratory conditions for 24 hours before testing. A 500 mg/L PFOS stock solution was prepared
with RO water to create nominal test concentrations of 0 (control), 10, 25, 50, 75, 100, 150, 300
and 500 mg/L. Each treatment contained 10 replicates with one Gosner Stage 25 tadpole in each
250 mL plastic tub maintained at 21ฐC and a 12:12-hour light:dark photoperiod. Mortality was
monitored twice daily. The author-reported LCso value was 144 mg/L PFOS. The EPA's
independently-calculated 96-hour LCso was 154.8 (108.7 - 200.9) mg/L PFOS and used
quantitatively to derive the freshwater acute water column criterion. However, this value was not
used in the SMAV calculation because a more sensitive life stage was available for the species.
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Tornabene et al. (2021) conducted acute toxicity tests with four species of frogs in the
genus Lithobates (formerly, Rana) and PFOS (purchased from Sigma Aldrich, Catalog # 77282-
10G; purity not provided). Acute tests followed standard 96-hour guidance (ASTM 2017; U.S.
EPA 2002). The four test species (American bullfrog, Lithobates catesbaeiana; 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
test initiation larvae were acclimated to test conditions (21ฐC and 12:12-hour light:dark
photoperiod) for 24 hours. A stock solution of PFOS (500 mg/L) was made in UV-filtered well
water and diluted with well water to reach test concentrations ranging from 0 - 500 mg/L PFOS.
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 solutions and were not fed during the exposure
period. The number of replicates varied by species, and treatment; 20 replicates in the control
and five to 10 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. Author-reported 96-hour LCsos were 163, 113, 73 and 130 mg/L PFOS for the American
bullfrog, green frog, northern leopard frog, and wood frog, respectively. The independently-
calculated 96-hr LCso values for American bullfrog and northern leopard frog were 133.23
(95.75 - 170.8), and 72.72 (63.88 - 81.55) mg/L, respectively. The EPA was unable to
independently calculate LC50 values for green frog and wood frog as a curve could not be fit with
significant parameters. Therefore, the independently-calculated LC50 values for American
A-55
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bullfrog (133.3 mg/L) and northern leopard frog (72.72 mg/L) were used quantitatively to derive
the freshwater acute water column criterion for PFOS, while the author-reported LC50 values for
green frog (113 mg/L) and wood frog (130 mg/L) were used. The author-reported toxicity values
were consistent with the independently-calculated LC50 values for other species included in the
study.
A.2.20 Twentieth Most Sensitive Freshwater Genus for Acute Toxicity: Physella (snail)
Li (2009) conducted three independent repeats of a 96-hour static acute test on PFOS
(potassium salt, > 98% purity) with the bladder snail species, Physella acuta (Note: formerly
known as 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. P. acuta of mixed ages were used 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. Exposure vessels were polypropylene
beakers of unreported dimensions and 1 L fill volume. The test employed 5-6 replicates of six
snails each in at least five test concentrations plus a negative control. Each treatment was tested
three independent times. Nominal test concentrations were in the range of 25-300 mg/L PFOS.
The test temperature was maintained at 25ฑ2ฐC. 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 reported. Survival of negative control animals was also not
reported. The author-reported 96-hour LC50 was 178 mg/L (C.I. 167-189) and represented an
average of the LC50S for each test. Only one of three independent experiments could be fitted.
The independently-calculated LC50 value was 183.0 (161.4 - 204.7) mg/L and was used
quantitatively to derive the freshwater acute water column criterion.
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Funkhouser (2014) conducted a 96-hour static test on PFOS (potassium salt, >98%
purity) with the physid snail, Physella heterostrophapomilia (Note: formerly known as Physa
pomilia), as part of a Master's thesis at the Texas Tech University, Lubbock, TX. Adult P.
pomilia (4 month old) used for the test were field collected from two different collections from
the North Fork of the Double Mountain Fork of the Brazos River near Lubbock, TX. Offspring
from both collections were reared in 12, 10-gallon aquaria with lab water for several generations
prior to use in the test. Dilution water was moderately hard reconstituted laboratory water (3.0 g
CaSC>4, 3.0 g MgSC>4, 0.2 g KC1, and 4.9 g NaFtCCb added to 50 L deionized water).
Photoperiod was 12:12-hours light:dark at an unreported light intensity. PFOS was dissolved in
dilution water to prepare the test concentrations. Exposure vessels were 400 mL polypropylene
containers containing 200 mL of test solution. The test employed two replicates of four snails
each in six test concentrations plus a negative control. Nominal concentrations were 0 (negative
control), 100, 150, 200, 250, 300, and 375 mg/L. Exposure concentrations were reportedly
measured, but concentrations were not reported. Analyses of test solutions were performed using
liquid chromatography/ tandem mass spectrometry (LC-MS/MS). Standards were used as part of
the analytical method, but details were not reported. The reporting limit was 0.010 mg/L.
Experiments were conducted in incubators set to 25ฐC, which did not vary more than 1ฐC during
the course of the test. No other water quality parameters were reported as having been measured
in test solutions. Negative control survival was not reported specifically for the test but was
reported to be 85-100% across all experiments. The author-reported 96-hour LCso was reported
as 161.77 mg/L. An independently-calculated toxicity value could not be calculated given the
level of data that was presented in the paper. The study author reported value was used
quantitatively to derive the freshwater acute water column criterion.
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Appendix B Acceptable Estuarine/Marine Acute PFOS Toxicity Studies
B.l Summary Table of Acceptable Quantitative Estuarine/Marine Acute PFOS Toxicity Studies
Species (lilVsliiuo)
Method'
Tesl
Dunilion
( hcmic;il /
Piiriu
pll
Temp
(ฐC)
Siilinilt
(DDli
r.nw-i
Author
Reported
I'.ITeel
(one.
(niii/l.)
l-'.PA
( iileuliiled
HITccl
(one.
(iii^/l.)
I'iiiiil
Klleel
(one.
Speeies
Me.in
Aeule
\ nine
(inii/l.)
Reference
Sea urchin (larvae),
Paracentrotus lividus
S,U
72 hr
PFOS
Unreported
18
35
EC50
(malformation)
1.795
-
1.795
1.795
Gunduz et
al. (2013)
Purple sea urchin
(embryo),
Strongylocentrotus
purpuratus
S, M
96 hr
PFOS-K
98%
-
15
30
EC50
(normal
development)
1.7
-
1.7
1.7
Hayman et
al. (2021)
Mediterranean mussel
(larva),
Mytilus galloprovincialis
S,U
48 hr
PFOS
Unreported
7.9-
8.1
16
36
EC50
(malformation)
>1
-
>lc
-
Fabbri et al.
(2014)
Mediterranean mussel
(embryo),
Mytilus galloprovincialis
S, M
48 hr
PFOS-K
98%
-
15
30
EC50
(normal
development)
1.1
-
1.1
1.1
Hayman et
al. (2021)
Mysid (3 d),
Americamysis bahia
S, M
96 hr
PFOS-K
98%
-
20
30
LC50
5.1
4.914
4.914
4.914
Hayman et
al. (2021)
Mysid (neonate, <24 hr),
Siriella armata
S,U
96 hr
PFOS
98%
-
20
-
LC50
6.9
-
6.9
6.9
Mhadhbi et
al. (2012)
Sheepshead minnow
(3.0 cm, 0.44 g),
Cyprinodon variegatus
R, M
96 hr
PFOS-K
86.9%
-
22
20
LC50
>15
-
>15
>15
Palmer et
al. (2002b)
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 the in the SMAV calculation
0 Not used in SMAV calculations, because a definitive value is available
B-l
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B.2 Detailed PFOS Acute Saltwater Toxicity Study Summaries and
Corresponding Concentration-Response Curves (when calculated for
the most sensitive genera)
The purpose of this section was to present detailed study summaries for acute saltwater
tests that were considered quantitatively acceptable for criterion derivation, with summaries
grouped and ordered by genus sensitivity. The data available for saltwater invertebrates fulfilled
three of the eight MDRs. The EPA could not, therefore, develop acute estuarine/marine criteria
following the 1985 Guidelines methods. In the interest of providing informations to states/Tribes
on protective values, the EPA developed an estuarine/marine acute benchmark using the
available empirical data supplemented with toxicity values generated through the use of New
Approach Methods (NAMs), specifically through the use of the EPA Office of Research and
Development's peer-reviewed publicly-available web-ICE tool (Raimondo et al. 2010). These
benchmarks are provided in Appendix L.
B.2.1 Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Mytilus (mussel)
The acute toxicity of perfluorooctane sulfonate (PFOS, purity not provided) on the
Mediterranean mussel, Mytilus galloprovincialis was evaluated by Fabbri et al. (2014). This
species is not resident to North America, but is a surrogate for North American mussel species,
including the widespread, commercially and ecologically important blue mussel, Mytilus edulis.
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 ฑ1ฐC. It was
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 16:8-hour light:dark photoperiod for 48 hours and exposed to one of six
B-2
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nominal PFOS concentrations (0.00001, 0.0001, 0.001, 0.01, 0.1, and 1 mg/L) or control (ASW)
water. The PFOS stock was made with ethanol, and ASW control samples run in parallel
included ethanol at the maximal final ethanol 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 the control group exhibiting >75% of normal D-shell stage larvae (ASTM
2004b). Authors noted that controls had >80% normal D-larvae across all tests. PFOS was only
measured once in one treatment which was similar to the nominal concentration; that is,
0.000085 mg/L versus the nominal concentration of 0.0001 mg/L. PFOS was below the limit of
detection (LOD) in the control ASW (0.06 ng/L or 0.00000006 mg/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 a 50% reduction in 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 acceptable for quantitative use.
Hayman et al. (2021) report the results of a 48-hour static, measured test on the effects
of PFOS-K (potassium salt, CAS # 2795-39-3, 98% purity, purchased from Sigma-Aldrich, St.
Louis, MO) on the Mediterranean mussel, Mytilus galloprovincialis. Authors note tests followed
U.S. EPA (1995) and ASTM (2004b) protocols. Mussels were collected in the field (San 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 16:8-hour light:dark photoperiod. Six test solutions were
B-3
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made in 0.45 |im filtered seawater (North San Diego Bay, CA) with PFOS-K dissolved in
methanol. The highest concentration of methanol was 0.1% (v/v). Measured test concentrations
ranged from 0.52 - 2.50 mg/L. Control mussel embryos were exposed to 100% filtered 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 and negative (100%
filtered seawater) control groups, suggesting no adverse effects of methanol. The author-reported
48-hr ECso, based on normal development, was 1.1 mg/L PFOS. The EPA was not able to
independently calculate a 48-hour EC50 value as the curve fitted model did not result in a good
fit. Therefore, the author-reported EC50 of 1.1 mg/L mg/L was considered for quantitative use.
B.2.2 Second Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Strongylocentrotus
(sea urchin)
Hayman et al. (2021) report the results of a 96-hour static, measured test on the effects
of PFOS-K (potassium salt, CAS # 2795-39-3, 98% purity, purchased from Sigma-Aldrich, St.
Louis, MO) on the purple sea urchin, Strongylocentrotuspurpuratus. Authors note that tests
followed U.S. EPA (1995) and ASTM (2004b) protocols. Sea urchins were collected in the field
(San Diego Bay, CA) and conditioned in 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 16:8-hour light:dark photoperiod.
Seven test solutions were made in 0.45 |im filtered seawater (North San Diego Bay, CA) with
PFOS dissolved in methanol. The highest concentration of methanol was 0.1% (v/v). Measured
test concentrations ranged from 0.52 - 10.0 mg/L. Control urchin embryos were exposed to
B-4
-------
100% filtered 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 (4-arm
pluteus stage). There were no significant differences between solvent and negative (100%
filtered seawater) control groups, suggesting no adverse effects of methanol. The author-reported
96-hour ECso, based on normal development, was 1.7 mg/L PFOS. The EPA was not able to
independently calculate a 96-hour EC50 value as the curve fitted model did not result in a good
fit. Therefore, the author-reported EC50 of 1.7 mg/L mg/L was considered for quantitative use.
B.2.3 Third Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Paracentrotus (sea
urchin)
A 72-hour static, unmeasured PFOS (purity not provided) toxicity test with the sea
urchin, Paracentrotus lividus (a non-North American species) was conducted by Gunduz et al.
(2013). Adult sea urchins were collected from the Aegean coast of Turkey, in an area the authors
noted as clean and lacking domestic and industrial wastewater inputs. Filtered natural seawater
from the same area was used as the dilution water. Adult sea urchins were cultivated in the same
filtered natural sea water with a salinity of 35 ppt and 18ฐC. Zygote suspensions (1 mL) were
added to the controls or 9 mL of the various PFOS treatments. This ensured that there were about
30 fertilized embryos/mL or approximately 300 embryos per treatment. The experiments were
conducted in six-well TPP culture plates with six replicates per treatment. PFOS stock solutions
were made with dimethyl sulfoxide (DMSO) and diluted with seawater to obtain five nominal
treatments (0.5, 1.0, 3.0, 5.0 and 10 mg/L PFOS). In addition to a natural seawater control,
experiments also included a DMSO solvent control equal to the amount in the highest test
concentration. The embryos were incubated in a growth chamber at 18 ฑ2ฐC from 10 minutes
after fertilization to up to the 72-hour pluteus larval stage. At test termination, 100 individuals
were selected randomly from each treatment and evaluated for normal plutei, retarded plutei,
B-5
-------
pathologic malformed plutei, pathologic embryos unable to differentiate up to the pluteus larval
stages, and dead embryos/larvae. There was 97.75% and 91% frequency of normal larvae in the
control and solvent control, respectively with no deaths reported in the controls or any PFOS
treatments. The 72-hour EC50 based on normal development to the pluteus stage was 1.795 mg/L
PFOS and is acceptable for quantitative use; however, additional consideration needs to be given
to the short test duration.
B.2.4 Fourth Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Americamysis
(mysid)
Hayman et al. (2021) report the results of a 96-hour static, measured test on the effects
of PFOS (potassium salt, CAS # 2795-39-3, 98% purity, purchased from Sigma-Aldrich, St.
Louis, MO) on the mysid, Americamysis bahia. Authors note that tests followed U.S. EPA
(1995); U.S. EPA (2002); and ASTM (2004b) protocols. Mysids were purchased from a
commercial supplier (Aquatic Research Organisms, Hampton, NH) and acclimated to test
conditions (30 ppt, 20ฐC and a 16:8-hour light:dark photoperiod). Five test solutions were made
in 0.45 |im filtered seawater (North San Diego Bay, CA) with PFOS-K dissolved in methanol.
The highest concentration of methanol was 0.1% (v/v). Measured test concentrations ranged
from 0.95 - 16 mg/L. Control mysids were exposed to 100% filtered seawater and the acute test
also included a solvent control. Five mysids (3 days old, which is older than the typical age of <
24 hours at test initiation) were added to 120 mL polypropylene cups and 100 mL of test
solutions with six replicates per treatment. Living mysids were counted and dead organisms were
removed daily. There were no significant differences between solvent and negative (100%
filtered seawater) control groups, suggesting no adverse effects of methanol. Only two organisms
were found dead in the controls at test termination. The author-reported 96-hour LC50 was 5.1
B-6
-------
mg/L PFOS. The independently-calculated 96-hr LC50 value was 4.914 (3.578 - 6.250) mg/L
and is acceptable for quantitative use.
B.2.5 Fifth Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Siriella (mysid)
Mhadhbi et al. (2012) performed a 96-hour static, unmeasured acute test with PFOS
(98% purity) and the mysid, Siriella armata. Mysids were collected from the same source as the
dilution water [filtered sea water from the Ria of Vigo (Iberian Peninsula)] 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 D.O.
6 mg/L). A stock solution of PFOS was made either with filtered sea water from the Ria of Vigo
for low exposure concentrations, or with DMSO for high PFOS concentrations (a final maximum
DMSO concentration of 0.01% (v/v) in the test medium). However, the authors did not indicate
what was considered a high-test concentration. If DMSO was used, a solvent control was also
included. Twenty neonates (<24 hours old) were used per each treatment. Mysids were exposed
to one of five nominal PFOS treatments (1.25, 2.5, 5, 10 and 20 mg/L). 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 mortality was recorded after 96 hours. The 96-hour LC50 was 6.9 mg/L PFOS and was
acceptable for quantitative use.
B.2.6 Sixth Most Sensitive Estuarine/Marine Genus for Acute Toxicity: Cyprinodon
(sheepshead minnow)
Palmer et al. (2002b) conducted a 96-hour static-renewal measured acute test with
PFOS-K (perfluorooctanesulfonate potassium salt, 86.9% purity from the 3M Company) on the
sheepshead minnow, Cyprinodon variegatus. The test followed standard guidance for acute
toxicity tests outlined in U.S. EPA (1985) and ASTM (1994). Sheepshead minnows were
purchased from a commercial supplier (Aquatic Biosystems, Fort Collins, CO) and held for
B-7
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several weeks prior to testing. Fifty-one hours before testing fish were acclimated to test
conditions (16:8-hour light:dark photoperiod, salinity of 20 ppt and 22ฐC). Natural seawater
(Indian River Inlet, Delaware) was filtered and diluted with well water to 20 ppt and was used
for culturing and testing. A nominal PFOS stock solution (40 mg/L) was made by dissolving
PFOS in methanol and diluting it with seawater to achieve the nominal test concentration (20
mg/L). A solvent control (0.5 mL/L methanol) and a sea water control were also included. Ten
minnows (3.0 cm, 0.44 g) were added to 25 L polyethylene aquaria with 15 L of test solution
(loading was 0.30 g fish/L of test water). Test treatments were replicated three times. PFOS
concentrations were measured daily at test solution renewal with averaged measured
concentrations in the control and solvent control less than the LOQ (5 mg/L) and PFOS-spiked
seawater, 15 mg/L. At test termination (96 hours) none of the minnows died in the controls or
single PFOS test treatment, therefore the author-reported LCso was >15 mg/L. The EPA was
unable to independently calculate the LCso value as this test only consisted of one treatment
group. The author-reported LCso >15 mg/L is being used quantitatively for the acute saltwater
benchmark because it is a greater than high value which adds value to the assessment of potential
sensitivity of this species to acute PFOS exposure (see Section 2.10.3.2).
B-8
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Appendix C Acceptable Freshwater Chronic PFOS Toxicity Studies
C.l Summary Table of Acceptable Quantitative Freshwater Chronic PFOS Toxicity Studies
Species ( lilosliiuo)
Method'
liM
Duriilion
( Ih-iii ic;il /
Pu lit \
pll
1 cm p.
(ฐC)
Chronic Yiiluc
l.nripoinl
Author
Kcporlc
(1
Chronic
Yiiluc
(inป/l.)
i:p\
( iilculiilcd
Chronic
Yiiluc
(inป/l.)
I'iiiiil
Chronic
Yiiluc
(inป/l.)'
Species
Mean
Chronic
Yiiluc
(inป/l.)
Kcl'crcncc
Fatmucket (adult),
Lampsilis siliquoidea
R, M
36 d
PFOS
>98%
7.6-8.5
14.6-
16.1
MATC
(metamorphosis
success)
0.01768
0.0123
0.01768
0.01768
Hazelton et al.
(2012);
Hazelton
(2013)
Snail (egg),
Physella heterostropha
pomilia
(formerly, Physa pomilia)
R, M
44 d
PFOS-K
>98%
-
25
EC10
(clutch size)
14.14
8.527
8.527
8.527
Funkhouser
(2014)
Rotifer (<2 hr old neonates),
Brachionus calyciflorus
R,Ub
Up to
158 hr
PFOS
>98%
-
20
LOEC
(reduced net
reproductive rate)
<0.25
-
0.25
0.2500
Zhang et al.
(2013)
Cladoceran (<24 hr),
Ceriodaphnia dubia
R, M
6 d
PFOS-K
98%
6.91-
8.02
24.0-
25.9
EC10
(neonates / female)
6.9
10.69
10.69
-
Krupa et al.
(2022)
Cladoceran (<8 hr),
Ceriodaphnia dubia
R, M
7 d
PFOS-K
97.5%
7.73
(7.64-
7.86)
24.5
(23.8-
25.2)
EC10
(neonates / female)
10.0
8.371
8.371
-
Kadlec et al.
(2024)
Cladoceran (<8 hr),
Ceriodaphnia dubia
R, M
7 d
PFOS-K
97.5%
7.76
(7.71-
7.81)
24.4
(24.1-
24.8)
EC10
(neonates / female)
14.5
9.205
9.205
-
Kadlec et al.
(2024)
Cladoceran (<8 hr),
Ceriodaphnia dubia
R, M
7 d
PFOS-K
97.5%
7.58
(7.49-
7.67)
24.9
(24.3-
25.7)
EC10
(neonates / female)
9.8
6.766
6.766
8.640
Kadlec et al.
(2024)
Cladoceran (6-12 hr),
Daphnia carinata
R, U
21 d
PFOS-K
>98%
-
21
MATC
(days to first brood)
0.00316
2
-
0.003162
0.003162
Logeshwaran
et al. (2021)
Cladoceran (<24 hr),
Daphnia magna
R, M
21 d
PFOS-K
90.49%
8.1-8.5
19.4-
20.1
EC10
(cumulative young)
16.97
11.19
11.19
-
Drottar and
Krueger
(2000e)
Cladoceran (<24 hr),
Daphnia magna
R, U
21 d
PFOS-K
95%
-
21
EC10
(survival)
35.36
16.35
16.35
-
Boudreau
(2002);
C-l
-------
Species (lilesljiiicM
Method'
Test
Dui'iilioii
( hemiciil /
Pu lit \
pll
Temp.
(ฐC)
Chronic Y;iluc
l.mlpoinl
Author
Rcporlc
(1
Chronic
Value
(inii/l.)
r.PA
( iilcuhiled
Chronic
Value
(inii/l.)
l-'inal
Chronic
Value
(ni"/!.)'
Species
Mesin
Chronic
Yiilnc
img/l.)
Reference
Boudreau et al.
(2003a)
Cladoceran (<24 hr),
Daphnia magna
R,U
21 d
PFOS
Unreported
-
21
EC10
(number of young /
brood)
1.768
1.051
1.051
-
Ji et al. (2008)
Cladoceran (<24 hr),
Daphnia magna
R,U
21 d
PFOS-K
>98%
-
20
EC10
(total neonates/female)
2.236
3.030
3.030
-
Li (2010)
Cladoceran (<24 hr),
Daphnia magna
R, M
21 d
PFOS-K
99%
7
22
EC10
(survival)
4.17
2.610
2.610
-
Yang et al.
(2014)
Cladoceran (<24 hr),
Daphnia magna
R, U
21 d
PFOS
98%
7.2
20
EC10
(number of offspring /
brood / female)
0.0179
0.001818
0.001818
-
Lu et al.
(2015)
Cladoceran (<24 hr),
Daphnia magna
R, U
21 d
PFOS-K
>98%
7
20
EC10
(survival)
5.657
3.596
3.596
-
Liang et al.
(2017)
Cladoceran (12-24 hr),
Daphnia magna
R, U
21 d
PFOS-K
98%
-
20
EC10
(growth-length)
0.8218
0.9093
0.9093
-
Yang et al.
(2019)
Cladoceran (<24 hr),
Daphnia magna
R, U
21 d
PFOS
>99%
7.5
23
MATC
(number of young)
1.581
-
1.5815
1.344
Seyoum et al.
(2020)
Cladoceran (<24 hr),
Moina macrocopa
R, U
7 d
PFOS
Unreported
-
25
EC10
(number of young /
starting adult)
<0.3125
0.1789
0.1789
0.1789
Ji et al. (2008)
Amphipod (7-8 d, juvenile),
Hyalella azteca
R, M
42 d
PFOS-K
98%
7.77-
8.10
22.1-
22.8
EC10
(survival)
<4.8
2.899
2.899
2.899
Krupa et al.
(2022)
Crayfish
(4 wk juvenile, 0.056 g),
Procambarus fallax f.
virginalis
R, M
28 d
PFOS-K
>98%
-
25
LC20
0.1670
-
0.1670
0.1670
Funkhouser
(2014)
Blue damselfly (nymph),
Enallagma cyathigerum
R, U
320 d
Perfluorooctanes
ulfonic acid
tetraethylammoni
um
>98%
>7.5
21
MATC
(survival at 150 days)
0.03162
-
0.03162
0.03162
Bots et al.
(2010)
Midge
(newly hatched larva),
Chironomus dilutus
R, M
10 d
PFOS-K
95%
-
21-23
EC10
(growth at 10 days)
0.04920
0.05896
0.05896
-
MacDonald et
al. (2004)
C-2
-------
Species (lilesljiiicM
Method"
Test
Dui'iilioii
( hemiciil /
Pu lit \
pll
Temp.
(ฐC)
Chronic Y;iluc
r.nripoini
Author
Rcporlc
(1
Chronic
Value
(inii/l.)
r.PA
( iilcnhiled
Chronic
Value
(inii/l.)
l-'inal
Chronic
Value
(ni"/!.)'
Species
Mesin
Chronic
Yiilnc
(mii/l.)
Reference
Midge (4-day old larvae),
Chironomus dilutus
R, M
16 d
PFOS
98%
6.8-8.7
20.0-
23.2
EC10
(mean biomass)
0.00162
0
0.001588
0.001588
-
McCarthy et
al. (2021)
Midge (4-day old larvae),
Chironomus dilutus
R, M
16 d
PFOS-K
98%
7.28-
7.75
22.0-
22.9
EC10
(growth)
0.0015
-
0.0015
0.005198
Krupa et al.
(2022)
Mayfly (<24 hr larva),
Neocloeon triangulifer
R, M
14 d
PFOS-K
98%
-
23
EC10
(dry weight at day 14)
0.00022
6
-
0.000226
0.000226
Soucek et al.
(2023)
Atlantic salmon
(embryo-larval),
Salmo salar
F, U
49 d
PFOS
98%
-
5.0-7.0
LOEC
(growth - weight and
length)
>0.1
-
>0.1
>0.1
Spachmo and
Arukwe
(2012)
Zebrafish (8 hpf),
Danio rerio
R, U
Life-cycle
PFOS
>96%
7.0-7.5
28
EC10
(F1 offspring: %
survival)
0.01581d
0.01650
0.01650
-
Wang et al.
(2011)
Zebrafish (male, 3.5 mo),
Danio reio
R, U
21 d
PFOS
Unknown
7.0-7.4
28
EC10
(mean body length)
0.05657
0.06274
0.06274
0.03217
Guo et al.
(2019)
Fathead minnow
(embryo, <24 hpf),
Pimephales promelas
F, M
47 d
PFOS-K
90.49%
8.2
24.5
EC10
(survival)
0.4243
0.4732
0.4732
-
Drottar and
Krueger
(2000d)
Fathead minnow (adult),
Pimephales promelas
F, M
21 d
PFOS
>98%
7.3
25
EC10
(fecundity)
0.4794
0.05101
0.05101
-
Ankley et al.
(2005)
Fathead minnow
(adult, 5 mo.),
Pimephales promelas
R, M
42 d
PFOS-K
>98%
7.9
24.96
EC10
(F1 larval growth -
weight)
0.06223
0.0549
0.0549
0.1098
Suski et al.
(2021)
Swordtail fish
(adult female 6-7 mo),
Xiphophorus helleri
R, U
42 d
PFOS-K
>98%
-
27
EC10
(female survival)
1.118
0.5997
0.5997
0.5997
Han and Fang
(2010)
Northern leopard frog (stage
8/9 embryo),
Lithobates pipiens
F, M
35 d
PFOS-K
98%
-
20
LC50
6.210
-
6.21
Ankley et al.
(2004)
Northern leopard frog (stage
8/9 embryo),
Lithobates pipiens
F, M
112 d
PFOS-K
98%
-
20
MATC
(growth - length)
1.732
-
1.732
-
Ankley et al.
(2004)
C-3
-------
Species < lil'eslii^e)
Method"
Test
Dui'iilioii
( hemiciil /
Pu lit \
pll
Temp.
(ฐC)
Chronic Y;iluc
r.nripoini
Author
Kcporlc
(1
Chronic
Value
(inii/l.)
r.PA
( iilcnhiled
Chronic
Value
(inii/l.)
l-'inal
Chronic
Value
(ni"/!.)'
Species
Mesin
Chronic
Yiilnc
(mii/l.)
Rcl'crcncc
Northern leopard frog (larva,
Gosner stage 26),
Lithobates pipiens
R, M
40 d
PFOS
>98%
20
MATC
(Gosner stage at 40 d)
0.0316
-
0.03162
-
Hoover et al.
(2017)
Northern leopard frog (larva,
Gosner stage 26),
Lithobates pipiens
R, M
40 d
PFOS
>98%
-
20
LOEC
(growth - snout-vent
length)
>1
-
>1
1.3161
Hoover et al.
(2017)
African clawed frog
(larvae, NF 46/47 - 5 dpf),
Xenopus laevis
R, M
4 mo.
PFOS
98%
6.5-7.0
22
LOEC
(survival, weight, sex
ratio/intersex)
>0.7160
-
> 0.7160
>0.7160
Lou et al.
(2013)
Clawed frog
(embryo, NF 10),
Xenopus tropicalis
(formerly, Silurana
tropicalis)
F, M
150 dpost
metamorp
hosis
PFOS
>98%
7.5
26
MATC
(weight at
metamorphosis)
0.7871
-
0.7871
0.7871
Fort 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 PFOS in the range of concentrations tested under similar
conditions. Daily renewal of test solutions.
0 Values in bold used in SMCV calculation. SMCVs are calculated as the geometric mean of all bold-faced values for a species. See section 2.10.3.2 (Chronic Measures of Effect)
for decision rules regarding use of greater (>) and less than (<) values in SMCV calculations.
d Author-reported value based on a different test endpoint than the EPA-calculated value.
C-4
-------
C.2 Detailed PFOS Chronic Freshwater Toxicity Study Summaries and
Corresponding Concentration-Response Curves (when calculated for
the most sensitive genera)
The purpose of this section was to present detailed study summaries for tests that were
considered quantitatively acceptable for freshwater chronic water column criterion 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 for the most sensitive genera when available. The C-R models included with the study
summaries were those for the four most sensitive genera (consistent with Section 3.1.1.3). When
required, the EPA also included models for non-resident species that were more sensitive than
the fourth most sensitive North American resident genus. In many cases, authors did not report
C-R data in the publication/supplemental materials and/or did not provide C-R data upon the
EPA request. In such cases, the EPA did not independently calculate a toxicity value and the
author-reported effect concentrations were used in the derivation of the criterion.
C.2.1 Most Sensitive Freshwater Genus for Chronic Toxicity: Neocloeon (mayfly)
Soucek et al. (2023) conducted a chronic life-cycle test to determine the effects of PFOS-
K (PFOS potassium salt, CAS # 2795-39-3, 98% purity) on the parthenogenetic mayfly,
Neocloeon triangulifer. The test was performed under renewal conditions over 27 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 then water was 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 water) and to 300 mL tall form glass beakers for emergence. There were sixteen
replicates (with one mayfly per replicate) per test concentration and control. Replicates one
C-5
-------
through eight were destructively sampled on day 14 and replicates nine through sixteen
continued until the end of the test (when all mayflies either molted into imago stage or died). The
endpoints that were evaluated included survival for all replicates, 14-d length and calculated dry
weight (using a previously published body length dry weight equation; Besser et al. 2021) for
replicates 1 through 8, and percent survival to pre-emergent nymph (PEN) stage, number of days
until PEN stage, percent emergence (to imago stage), and pre-egg laying live weight of imago
for replicates 9 through 16. Nominal test concentrations were 0 (control), 0.00016, 0.00031,
0.00063, 0.00125, 0.0025, 0.005, and 0.010 mg/L PFOS. Mean measured PFOS concentrations
(EPA Analytical Method 1633; LC-MC/MS) were 0.000056 (control), 0.000205, 0.000418,
0.000764, 0.001143, 0.002057, 0.003892, and 0.006789 mg/L PFOS, 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 100%. Percent survival
of mayflies after 14 days in the remaining seven test concentrations ranged from 79 to 100%.
The most sensitive endpoint was 14-day dry weight. The study authors reported three different
14-day dry weight ECio values that were calculated using various point-estimation approaches.
The author-reported 14-day dry weight ECio values produced by the various approaches were
relatively similar to one another, ranging from 0.000226 (using TRAP [2 parameter, threshold
sigmoidal curve]) to 0.000272 mg/L (using log-linear regression, controls excluded). The EPA
was not able to fit a reliable model with significant model parameters to the 14-day dry weight
C-R dataset and, therefore, relied on the author-reported ECio of 0.000226 mg/L (based on
TRAP) as the primary effect concentration. The EPA selected the TRAP-based ECio
preferentially over the ECio values based on the two other point estimation approaches (i.e., log-
C-6
-------
linear regression with and without controls) because the TRAP-based model (1) considered
control responses; (2) was more fundamentally consistent with the maximum likelihood
regression approaches used by the EPA to assess the C-R datasets evaluated in this document,
and; (3) relied on replicate-level data, which the EPA used preferentially over treatment-mean
data in assessing the C-R datasets. The author-reported ECio of 0.000226 mg/L (TRAP-based)
was used in the derivation of the freshwater PFOS chronic water column criterion.
Current toxicity literature indicates that aquatic insects, specifically mayfly, midge, and
odonates, are sensitive to PFOS exposures. Further, given that recent research has led to the
development of successful culturing methods for mayflies that are now being used in laboratory-
based toxicity studies (Soucek and Dickinson 2015; Soucek et al. 2023), new toxicity studies
indicate that mayflies (Neocloeon triangulifer) are the most sensitive taxa to acute and chronic
PFOS exposures. This finding is consistent with other chemical exposures, including 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 (Cormier et al. 2018; U.S. EPA
2011). Many of these laboratory-based toxicity tests used mayflies capable of adapting to
laboratory settings, therefore it is hypothesized that mayfly species unable to survive in
laboratory settings may also be even more sensitive to contaminant exposures than the relatively
hardy mayfly species (e.g., N. triangulifer) now commonly being used to toxicity testing. Thus,
inclusion of the mayfly toxicity data (Soucek et al. 2023) is important to ensuring the
protectiveness of the PFOS aquatic life AWQC.
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Publication: Soucek et al (2023)
Species: Mayfly, Neocloeon triangulifer
Genus: Neocloeon
EPA-Calculated ECio: unable to fit a reliable model with significant model parameters;
author-reported TRAP-based ECio used
C.2.2 Second Most Sensitive Freshwater Genus for Chronic Toxicity: Chironomus (midge)
MacDonald et al. (2004) conducted larval sub-chronic partial-life cycle and chronic life-
cycle tests to determine the effects of PFOS (potassium salt, 95% purity) on the midge,
Chironomus dilutus (formally known as Chironomus tentans). The test was performed under
renewal conditions over 10 days for the larval sub-chronic partial-life cycle test and 60 days for
the chronic life-cycle test, with four of the twelve replicates terminated following 20 days of
exposure to evaluate survival and growth. The tests followed the general guidance given by
EPA-600-R99-064 (U.S. EPA 2000b) and ASTM E 1706-00 (ASTM 2002). These methods are
for measuring the toxicity and bioaccumulation of sediment-associated contaminants with
freshwater invertebrates and have different exposure durations than those typically considered
for invertebrate aqueous exposures, as well as different control survival requirements and
recommendations. C. dilutus used for the tests were 10-day old larvae for the 10-day exposure
(larval sub -chronic partial-life cycle test) and newly-hatched larvae at test initiation for the
chronic life-cycle test (both 60-d and 20-day exposures). Dilution water was reconstituted hard
water consistent with ASTM (2002) with unspecified total hardness, but typically 160-180 mg/L
as CaCC>3 (based on ASTM 2002), with alkalinity 110-120 mg/L as CaCC>3, and pH 7.6-8.0. The
photoperiod was 16:8-hours light:dark. Light intensity was not reported. A primary stock
solution was proportionally diluted with dilution water to prepare the test concentrations.
Exposure vessels were 250 mL polypropylene beakers containing 240 mL of test solution and a
sediment substrate. The 10-day exposure test employed at least two replicates with 10
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individuals all of which were obtained from four large C-shaped egg cases that were distributed
among seven test solutions plus a negative control. The life-cycle test (60-day exposure)
employed 12 replicates of 12 midges each in five measured test solutions plus a negative control.
From these 12 replicates, four were randomly terminated following a 20-day exposure to
measured survival and growth endpoints (thereby referred to as the 20-day exposure henceforth)
The remaining eight replicates were monitored over the test duration for emergence and
reproduction. Nominal test concentrations for the 10-day test were 0 (negative control), 0.001,
0.005, 0.010, 0.020, 0.040, 0.080, and 0.150 mg/L. The nominal test concentrations for the 20-
day exposure were 0 (negative control), 0.001, 0.005, 0.010, 0.050, and 0.100 mg/L. Mean
measured concentrations for the 10-day test were 0 (LOQ not reported), 0.0008, 0.00460, 0.0115,
0.241, 0.0491, 0.0962, and 0.1501 mg/L, respectively. Mean measured concentrations for the 20-
day exposure were 0 (LOQ not reported), 0.0023, 0.0144, 0.0217, 0.0949, and 0.149 mg/L,
respectively. Analyses of test solutions were performed using LC-MS. The mean percent
recovery and detection limits were not reported. Measured values of test concentrations in the
20-day exposure were 2 to 2.5-fold higher than nominal concentrations. Temperature and D.O.
concentrations were measured in at least two replicates per treatment on a daily basis for the 10-
day test and up to day 20 in the life-cycle test. Afterwards they were measured every other day
(on alternate days from test solution renewal) from days 21 to 60 for the life-cycle test. The
frequency of monitoring was reduced during this period, because both parameters consistently
remained within acceptable ranges (21.0-23.0ฐC; D.O. >5.0 mg/L). Survival of negative control
animals was >75%, which was considered acceptable for a full life-cycle exposure per ASTM
(2002). The study authors reported ECios and NOECs; however, specific details pertaining to the
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curve fitting process (including statistical output from the models and the curves) were not
provided in the paper and therefore, limit independent interpretation of the toxicity values.
The observed effects of PFOS on C. dilutus reported in the paper by the study authors
include survival and growth as weight (measured as mg of ash-free dry mass per individual) for
both the 10-day and 20-day exposure durations and emergence and reproduction over the 60-day
exposure duration. Significant reductions in larval weight were observed after 10 days of
exposure to PFOS in the 0.0962 and 0.1501 mg/L treatment groups (roughly 0.38 and 0.19 mg,
respectively) compared to the control group (roughly 0.88 mg). These differences resulted in
roughly a 56.8 and 78.4% decline in midge weight in these treatment groups compared to those
observed in the control. In contrast, there were no significant differences reported for survival
between any of the PFOS treatments (with percent survival ranging between roughly 69.7% in
the highest treatment group and 100% in the lowest) and the control (with roughly 100%
survival). However, the authors noted that there was a 30% decline of midge survival in the
highest PFOS treatment group with a measured concentration 0.1501 mg/L. The author-reported
10-day growth and survival ECios for the study were 0.0492 and 0.1079 mg/L, respectively. The
study authors also reported NOECs of 0.0491 mg/L, LOECs of 0.0962 mg/L, and MATCs of
0.0687 mg/L for both endpoints.
Similar to the 10-day exposure results summarized above, there was a general decline in
growth (as ash-free dry mass per individual) across the PFOS treatment groups (ranging roughly
between 29.2 and 47.2% reduction compared to controls) in the 20-day exposure (chronic life-
cycle test). However, only the decline in the 0.0949 mg/L treatment group was significantly
different (roughly 0.29 mg) compared to the control (roughly 0.89 mg) and there was not a C-R
relationship across the PFOS treatment groups. Additionally, midge survival was reduced after
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20 days of exposure to PFOS in the 0.0949 and 0.149 mg/L treatment groups (29.2 and 0%
survival, respectively) compared to the control (75% survival). Survival was determined to be
not significantly different across the rest of the PFOS treatment groups (ranging roughly between
56.5 and 75% survival) compared to the control. However, it should be noted that there was a
25% decline in survival in the 0.0217 mg/L PFOS treatment group compared to the control that
was determined not to be significantly different. The author-reported 20-day ECios for growth,
survival, and total emergence were 0.0882, 0.0864, and 0.0893 mg/L, respectively, and the study
authors also reported NOECs of 0.0217 mg/L for growth and survival and < 0.0023 mg/L for
emergence, LOECs of 0.0949 mg/L for growth and survival and 0.0217 mg/L for emergence,
and MATCs of 0.0454 mg/L for growth and survival and 0.0071 mg/L for emergence. Also, it
should be noted, the paper reports contrasting NOECs for 20-day survival. The text in the paper
stated that the NOEC was 0.0217 mg/L for growth and survival and Table 2 of the paper stated
0.0949 mg/L. The EPA assumed the NOEC in Table 2 of the paper was not correct and that
0.0217 mg/L was the correct NOEC based on the data presented in Figure 3 A of the paper. This
assumption was applied to the summary of the study results presented in this document.
Independent statistical analyses were conducted for both the 10-day (larval sub-chronic
partial-life cycle test) and 60- and 20-day (chronic life-cycle test) exposure durations using data
that were estimated (using Web plot digitizer) from the figures presented in the paper. The EPA
could not fit a curve to independently verify the 10-day survival (due to a lack of a specific
sample size for this endpoint as the number of replicates was not stated in the paper; however,
the number of replicates was between two and four and the EPA sought to obtain clarification
and treatment level data from the study authors) or the 20-day growth toxicity values (due to a
lack of an observed C-R for this endpoint). However, the EPA-calculated 10-day ECio for
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growth was 0.05896 mg/L, which was slightly higher than the growth-based ECio of 0.0492
mg/L reported in the paper. The chronic life-cycle ECios for larval survival (following 20 days of
exposure) and emergence (with 60 days of exposure) were 0.0171 and 0.0102 mg/L,
respectively. These chronic life-cycle ECios were much lower than those reported in the paper of
0.0864 and 0.0893 mg/L, respectively. The ECios for survival and emergence endpoints from the
chronic life-cycle test were not considered to be reliable endpoints at this time given the
disparities in the calculated ECios and the level of data that was presented in the paper, which
made independent verification of the toxicity values less accurate. Specifically, for the 20-day
survival endpoint, there appeared to be overdispersion (i.e., observed data display a larger
variability than would be expected given an assumed statistical distribution about the mean
response) in the data as it was presented in the paper (in Figure 3 A of the paper), which adds
uncertainty around the independently-calculated ECio of 0.0171 mg/L and may explain the
disparity between the reported ECio and the EPA's independently-calculated value. As for the
60-day emergence endpoint, there was a lack of a C-R relationship and there were very similar
levels of observed effects (which ranged between 42.6 and 50.1%) despite the more than nine-
fold increase in the mid-range treatment concentrations (0.0023, 0.0144, 0.0217 mg/L,
respectively). Lastly, the toxicity values from the observed effects from the chronic life-cycle
exposure were considered to be less certain given the relatively large difference between the
nominal and measured concentrations for this test. The dosing of the chronic life-cycle test (20-
and 60-day exposure) was more of a concern than the larval sub-chronic partial-life cycle test
(10-day exposure), which had measured concentrations that were much more in line with the
expected nominal concentrations. Thus, the survival and emergence endpoints from the chronic
life-cycle test were not considered for quantitative use in the derivation of the freshwater chronic
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water column criterion. Instead, these endpoints were considered as supporting information until
detailed replicate level data can be obtained from the study authors.
The most sensitive endpoint from the remaining toxicity values that could be
independently calculated was for 10-day growth with an ECio of 0.05896 mg/L. As mentioned in
the Bots et al. (2010) summary in Section C.2.4 below, the observed effects of PFOS on aquatic
insects appears to be consistent across the available data for chironomids and odonates.
However, Bots et al. (2010) did not measure the effects of PFOS on nymph growth and
therefore, the observed effects in MacDonald et al. (2004) on larval weight cannot be compared
across the two studies. The ECio of 0.05896 (0.05769 - 0.06023) mg/L for 10-day growth was
used quantitatively to derive the freshwater chronic water column criterion. The remainder of the
toxicity values were used as supporting information to corroborate the toxicity value used and to
better understand the effects of PFOS on aquatic insects.
McCarthy et al. (2021) conducted a 10-day range-finding toxicity test and a separate 20-
day (note, based on age of starting organisms, this test was actually 16 or 19 days of exposure)
"abbreviated partial-life cycle" toxicity test with PFOS (98% purity, purchased from Sigma-
Aldrich) on the midge, Chironomus dilutus. PFOS stock solution was dissolved in reconstituted
moderately hard water without the use of a solvent and stored in polyethylene at room
temperature until use. Two chronic exposures with PFOS were run, a 10-day and a 20-day
exposure, following standard protocols (U.S. EPA 2000b) with slight modifications. The 10-day
exposure was considered a range-finding test, with concentrations spaced by ~ 1 OOx and only
mortality measured, whereas the 20-day exposure measured both survival and growth. The 20-
day exposure is less than the recommended 65-day full-life cycle method outlined in U.S. EPA
(2000b) and since exposures of midges started on day two or four the actual exposure duration is
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only 16 or 19 days long; therefore, the study authors referred to this test as an "abbreviated
partial-life cycle test". Exposure vessels for both experiments were 1 L high-density
polyethylene beakers containing natural-field collected sediment. The 10-day exposure had 60
mL of sediment and 105 mL of test solution and the 20-day exposure had 100 mL of sediment
and 175 mL of test solution. PFOS in test solutions was added via pipette to the beakers with the
tip just above the sediment substrate. Nominal test concentrations for the 10-day and 20-day
exposure were 0, 0.0004086, 0.33, 33, 100 and 350 mg/L PFOS and 0, 0.001, 0.005, 0.01, 0.05
and 0.1 mg/L PFOS, respectively. Egg cases were obtained from outside suppliers (Aquatic
Biosystems or USGS Columbia Environmental Research Center) and held for 10 days in the 10-
day test or held for four days before testing in the 20-day exposure (in test vessels). In the 20-day
exposure the test organism age (four-day old larvae) was greater than the protocol
recommendation (< 24 hour) because earlier experiments had control survival issues (< 70%). In
both tests each beaker held 12 organisms with five replicates per exposure treatment. Solutions
were renewed every 48 - 72 hours in the 10-day exposure and daily for the 20-day exposure.
Water samples of test concentrations were measured on day one and day 10 in the 10-day
exposure and day 10, 15 and 20 in the 20-day exposure. In the 10-day exposure measured test
concentrations ranged from 7 - 62% of nominal. In the 10-day exposure, the author-reported
LOEC, based on mortality, of 0.4086 |ig/L (0.0004086 mg/L PFOS) is reported as a nominal
concentration. Mean PFOS concentrations in the 20-day exposure were 0 (control), 0.000447,
0.00209, 0.0042, 0.0231 and 0.0463 mg/L PFOS. Results of the C. tetans 16- to 19-day
"abbreviated full life cycle test" were used preferentially over the results of the 10-day range-
finding test to inform the chronic sensitivity of C. tetans because: (1) the 10-day range-finding
test only measured survival, (2) the 10-day range-finding test had exposure concentrations that
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differed by up to a factor of 100, which make C-R modeling more difficult than the dilution
series in thel6- to 19-day test, and (3) the 16- to 19-day test was a longer exposure duration that
was more akin to a full life-cycle test. Consequently, the results of the 10-day range finding test
were not used quantitatively to derive the freshwater chronic water column criterion, but they
were retained for qualitative use.
From the 16- to 19-day "abbreviated full life cycle test" percent survival in the control
and lowest test concentration were 77% with no survivors reported in the highest two test
concentrations. The most sensitive endpoint appeared to be survival with an author-reported 16-
day reported ECio of 1.36 |ig/L (0.00136 mg/L PFOS). Additionally, the authors reported ECios
of 1.62 |ig/L (0.00162 mg/L PFOS) and 3.23 |ig/L (0.00323 mg/L PFOS) for growth as mean
biomass and mean weight, respectively. The EPA was unable to independently calculate ECios
for survival and mean weight. However, the 16- to 19-day independently-calculated ECio value
for mean biomass was 0.001588 (0.00118 - 0.00200) mg/L PFOS. This independently-calculated
ECio was acceptable for quantitative use and was used in the derivation of the freshwater chronic
water column criterion.
Krupa et al. (2022) conducted a partial-life cycle chronic toxicity test with the midge,
Chironomus dilutus, and PFOS-K (perfluorooctanesulfonate potassium salt, > 98% purity, CAS
No. 2795-39-3, purchased from Sigma-Aldrich). C. dilutus egg masses obtained from Aquatic
Biosystems (Fort Collins, Colorado, USA) were placed in 12-inch glass culture bowls (2-3 egg
mases per dish) containing carbon-filtered municipal tap water and examined daily for viability
and hatch. Hatching began after 2 days and larvae typically left the egg case 24-hours after the
first hatch. The larvae were fed daily finely ground TetraMinฎ fish food flakes (-150 mg/dish as
a slurry) with partial water changes as necessary until the larvae were the appropriate age for
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initiating the test (4-days old). The larvae were exposed to PFOS for 16 days. A 100 mg/L
PFOS-K stock solution was prepared by dissolving PFOS-K salt into ultrapure water using a stir
bar and stir plate for >20-hours. Controls and PFOS solutions were then prepared using carbon
filtered dechlorinated tap water. Each test concentration was individually spiked rather than
serially diluted to reduce PFOS waste generated. All test solutions for water renewals were
prepared the day before test initiation. The measured exposure concentrations were < LOD,
0.001, 0.0025, 0.004, 0.0075, 0.016 and 0.03 mg/L, with 5 replicates of 12 animals per replicate
for each concentration. Tests were conducted in 300 mL polycarbonate beakers containing
approximately 200 mL of exposure solution under a 16:8-hour light:dark cycle in an
environmental chamber maintained at 22.5 ฑ 1ฐC. Each beaker received 50 mL of clean silica
sand (250 - 500 jam) as substrate. Aeration was provided and larvae were fed finely ground
TetraMinฎ (6 mg/day, added as a slurry in dechlorinated water). Partial water renewals (150 mL
of solution was exchanged) were conducted three times per week, after 48 or 72-hour periods. At
test termination, larval survival was assessed, and ash-free dry weight (AFDW) was determined
following ASTM (2019). The AFDW of five groups of 12 larvae was measured at test initiation
to establish a baseline for growth. The temperature, D.O., pH, and total hardness test values
ranged from 22.0 - 22.9ฐC, 7.29 - 8.05 mg/L, 7.28 - 7.75 SU and 52 - 62 mg/L as CaC03,
respectively. Water samples for verification of PFOS concentrations were collected at test
initiation (day 0) and termination (day 16), in addition to before (out-water) and after (in-water)
every renewal of test solution. Water samples from days 0, 6, 10, and 16 were analyzed to verify
PFOS concentrations. The author-reported 16-day growth ECio was 0.0015 mg/L PFOS-K. The
EPA was unable to fit a reliable model for any of the chronic endpoints from this test. Therefore,
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the author-reported ECio value of 0.0015 mg/L for growth was used to derive the freshwater
chronic water column criterion.
C.2.2.1 MacDonald et al. (2004) Concentration Response Curve - Chironomus (midge)
Publication: MacDonald et al. (2004)
Species: Midge (Chironomus dilutus)
Genus: Chironomus
EPA-Calculated ECio: 0.05896 (95% C.I. 0.05769 - 0.06023) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
-2.6770
0.6384
-4.1933
0.0057
e
0.0805
0.0090
8.9243
0.0001
Concentration-Response Model Fit:
MacDonald et al. 2004
Chironomus dilutus
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C.2.2.2 McCarthy el al. (2021) Concentration Response Curve - Chironomus (midge)
Publication: McCarthy et al. (2021)
Species: Midge Chironomus dilutus
Genus: Chironomus
EPA-Calculated ECio: 0.001588 (95% C.I. 0.00118 - 0.00200) mg/L
Concentration-Response Model Estimates:
Parameter
Estimate
Std. Error
t-stat
p-value
b
5.2881
1.0432
5.0693
0.0148
d
1.0372
0.0238
43.4942
2.675 e5
e
0.0024
0.0001
21.9936
0.0002
Concentration-Response Model Fit:
Chironomus dilutus
Log Logistic type 1, 3 para
PFOS {mg/L)
C.2.2.3 Krupa et al. (2022) Concentration Response Curve - Chironomus (midge)
Publication: Krupa et al. (2022)
Species: Midge, Chironomus dilutus
Genus: Chironomu
EPA-Calculated ECio: unable to fit a reliable model with significant model parameters;
author-reported ECio used
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C.2.3 Third Most Sensitive Freshwater Genus for Chronic Toxicity: Lampsilis (mussel)
Hazelton (2013); Hazelton et al. (2012) conducted a test of the long-term effects of
PFOS (acid form, > 98% purity) on glochidia and juvenile life stages from the mussel Lampsilis
siliquoidea. To initiate the PFOS partial life-cycle test, brooding females were collected from
Perche Creek, Missouri and shipped over night to the test laboratory. The length of time between
collection from Perche Creek and shipment was not reported, and authors were unable to recall
such details (R. Bringolf, personal comm.); however, the EPA did not believe storage, shipping,
and handling compromised test results since study authors only relied on those mussels with
>70% glochidia viability. Dilution water was dechlorinated tap water. Mean total hardness (47.5
ฑ 9.2 mg CaCCb/L) and alkalinity (34.8 ฑ4.1 mg CaCCb/L) were measured by titration twice
weekly (n = 8) prior to water changes. Replicates used for water quality measurements were
changed daily to allow measurements from all four replicates every four days. For all treatments,
water temperature ranged from 14.6 to 16.1ฐC, D.O. ranged from 6.1 to 7.3 mg/L, and pH ranged
from 7.6 to 8.5, but did not differ across treatments. Photoperiod and light intensity were not
reported. No details were provided regarding primary stock solution and test solution
preparation. The test exposed brooding glochidia (in marsupia) for 36 days followed by a 24-
hour exposure of free glochidia. Experiments were conducted in 3.8 L glass jars of unspecified
fill volume. The 36-day in marsupia exposure test employed four replicates individually
containing single brooding females for each of the two PFOS treatment groups plus the control.
The in marsupia exposure was followed by a 24-hour free glochidia exposure consisting of a
factorial design, such that free glochidia from the control group of the in marsupia exposure
were divided between a control and the two PFOS treatments and the PFOS treatments were split
into control and the same PFOS treatment group as the in marsupia exposure. This factorial
design allowed for the comparison of PFOS effects in two different life stages. However, it
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should be noted that glochidia were pooled from females within each in marsupia treatment
group, and thus the influence of parental effects could be a confounding factor that cannot be
separated from the PFOS effects. Nevertheless, the influence of the potential parental
confounding factor was likely to be minimal compared to the effects of the PFOS exposures.
Nominal concentrations throughout the exposures were 0 (negative control), 0.001 and
0.100 mg/L. Mean measured concentrations were 0.00211 (negative control), 0.00452 and
0.0695 mg/L. Analyses of test solutions were performed at the U.S. EPA National Exposure
Research Laboratory in Research Triangle Park, NC using HPLC/MS. Two standard curves were
used to quantify PFOS water concentrations during the experiment: low range (0.00005,
0.00025, 0.0005, 0.00075, 0.001, 0.0025, 0.005 mg/L) and high range (0.001, 0.005, 0.010,
0.025, 0.050, 0.100, 0.150 mg/L). Two replicate samples were measured at each standard
concentration. Accuracy (recovery) of PFOS in the low-range standard curve ranged from 89.5
to 123% (n = 7) and for the high-range standard curve accuracy was 85.3 to 123% (n = 7). Adult
mussel and glochidia survival in the negative control was 100% and > 90%, respectively. The
study authors determined that the in marsupia exposure held the greatest weight of evidence and
explained 78% of the variability in the glochidia viability (AIC = 22843, = 0.78) and 83% of
the metamorphosis success (AIC = 21955, = 0.83), and therefore it appeared that the data
presented in the study are in terms of the in marsupia exposure alone and there are no data
presented in terms of the factorial design during the 24-hour free glochidia exposure.
Additionally, the specific treatment groups of the data presented in the paper are unclear in terms
of the factorial design during the 24-hour free glochidia exposure (e.g., it is unclear if the data
presented in Figure 2 of the paper are lumped according to marsupial exposure, reducing seven
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treatments to three, or if only the data in which the in marsupia and free glochidia exposures
were the same are presented).
The test resulted in an author-reported NOEC of 0.0695 mg/L, which was associated with
a 38% reduction in the viability of free glochidia at 24 hours post removal from females, a point
when control viability of free glochidia was > 80% (author reported LOEC and MATC > 0.0695
mg/L). While a 38% reduction was observed at the NOEC (0.0695 mg/L) treatment group
compared to controls, the authors reported this reduction was not statistically different from the
control. Over time, the study authors reported significant reductions in free glochidia survival
between three- and seven-days post removal from females, indicating a potential LOEC < 0.0045
mg/L. However, it should be noted that the observed level of effect between the two PFOS
treatment groups (0.0045 and 0.0695 mg/L) were extremely similar despite the 15-fold
difference between treatment groups. Additionally, in accordance with the decision rule
described in Section 2.10.3.2 and a study by Bringolf et al. (2013), only glochidia toxicity data
within 24 hours and with survival of at least 80% in the control treatment would be considered
(U.S. EPA 2013). These specific data requirements ensured that the related effects of PFOS
exposure to the viability of glochidia were consistent with environmental exposures during this
short life stage and also take the unique life cycle of mussels into account. Therefore, the chronic
toxicity value for viability of free glochidia at 24 hours following removal from females resulted
in a NOEC of >0.0695 mg/L, which is an uncertain value and indicated that viability of free
glochidia at 24 hours was a less sensitive endpoint.
In contrast, the data presented in the paper for metamorphosis success suggest a NOEC
of 0.0045 mg/L and a LOEC of 0.0695 mg/L, or MATC = 0.01768 mg/L. The reduction in
metamorphosis success at the LOEC was estimated to be 35.4%. However, as there were only
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two PFOS treatment groups and the gap in these exposure concentrations is large (about 15-
fold), the EPA was not able to fit a curve to estimate an ECio in a manner similar to the other
toxicity studies used to derive the freshwater chronic water column criterion. Instead, both the
use of an MATC and an estimated ECio were considered for the chronic value. An ECio was
estimated by assuming the 0.0695 mg/L treatment represents an EC35.4 and estimating the ECio
using the exposure response slope from another chronic PFOS toxicity study focused on another
mussel species (Perna viridis). Specifically, the chronic exposure of Perna viridis reported by
Liu et al. (2013), which is summarized in Section 3.1.1.4.1 and D.2.1, was used to derive a ratio
of EC10/EC35.4 values from that study equal to: EC10/EC35.4 = 0.0033/0.0186 = 0.1774. Applying
this ratio to Hazelton et al. (2012) yields an estimated ECio of 0.0123 mg/L. Given the
similarity between this ECio and the author-reported MATC for metamorphosis success of
0.01768 mg/L, the latter was used to derive the freshwater chronic water column criterion.
While this MATC is currently used quantitatively to derive the chronic water column criterion,
the EPA hopes to further refine this estimated ECio by obtaining the treatment level data from
the study authors and exploring additional exposure response slopes from the study-specific
dataset.
C.2.3.1 Hazelton et al. (2012) Concentration Response Curve - Lampsilis (mussel)
Publication: Hazelton et al. (2012)
Species: Fatmucket, Lampsilis siliquoidea
Genus: Lampsilis
EPA-Calculated ECio: 0.0123 mg/L
Concentration-Response Model Fit: Concentration-response data not available
Value used Quantitatively in Criterion: Author-reported MATC of 0.01768 mg/L
C.2.4 Fourth Most Sensitive Freshwater Genus for Chronic Toxicity: Enallasma (damselflv)
Bots et al. (2010) conducted a 320-day partial life-cycle study under renewal test
conditions to assess the effects of PFOS (tetraethylammonium salt, 98% purity) on the damselfly
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Enallagma cyathigerum. Test organisms were obtained by collecting mature female E.
cyathigerum all from the same location near the edge of a fen (a groundwater fed wetland) in
northern Belgium. After collection, females were transported to the laboratory in small cages and
housed in oviposition chambers for 24 hours before eggs were collected. E. cyathigerum used for
the test were newly-hatched nymphs at test initiation. Dilution water was dechlorinated tap water
that contained only a negligible concentration of PFOS (2.64 ng/L) and no other water quality
parameters from the tap water were provided other than pH >7.5. Photoperiod was 16:8-hours
light:dark in a climate room. Light intensity was not reported. Test solutions were prepared
taking purity into account. To start the test, a total of 18,552 eggs were distributed amongst 150
exposure chambers (i.e., petri dishes of unreported size and material type). The distribution of
the total number of eggs consisted of the entire clutch from each of the 30 females being divided
into five subsamples, which were then randomly allotted to the various test treatments; thereby
ensuring that each treatment group consisted of an even distribution of test organisms from the
30 females. After hatching, a total of 7,938 nymphs continued to be exposed (10 individuals per
cup of unreported size and material type). After 10 days, seven nymphs for every female and
treatment were monitored (resulting in a total of 741 nymphs). Nominal concentrations were 0
(negative control), 0.01, 0.1, 1.0, and 10 mg/L. Actual test concentrations were not measured. All
nymphs were housed (and presumably tested) in a climate room at 21ฐC. Water quality (pH,
carbonate and total water hardness, O2, NO2, and NO3 levels) was checked weekly using
standard aquarium tests, but values are not reported. Approximately 40% of the nymphs in the
control treatment died during the first 60 days and similar mortality levels were observed in the
other treatments. Additionally, it appears that control survival plateaued between 60 and 200
days, with 82.57% of the remaining nymphs in the control treatment surviving during this time,
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indicating that survival settled out during this phase of the experiment. The initial drop in nymph
survival can likely be attributed to the handling of the test organisms between the various phases
of the experiment. This would explain the observed plateau between 60 and 200 days, as the
nymphs were not handled during this time. The observed control mortality in this test was
consistent with other odonate tests and excessive mortality of nymphs is typically expected
within the first 200 days given the difficulty in maintaining odonates in a lab setting (Abbott and
Svensson 2007; Rice 2008). Therefore, the observed control survival for this study was
considered within the acceptable range for this species up to the 200-day exposure duration.
Further, the control survival observed in this study was largely consistent with the toxicity
testing guidelines for chironomids (requiring 70% control survival)(ASTM 2002; U.S. EPA
2002), which represent the only test guidelines for an emergent aquatic insect species as similar
test guidelines for odonates are not available. Therefore, considerations regarding the use of
these data for chronic water column criterion derivation was based on best scientific judgement
and were restricted to the first 200 days of the experiment. After 200 days, nymph survival in the
control and the PFOS treatments decreased. This drop in survival likely coincided with
metamorphosis. However, control survival at the end of the exposure duration was only roughly
40% of the starting nymphs and therefore, survival after 200 days of exposure was not
considered a viable test endpoint for this particular study.
The other possible observed effects of PFOS on E. cyathigerum reported by the authors
included decreased survival over the exposure duration and decreased metamorphosis success.
Nymph survival after five days did not differ between the control, 0.01 and 0.100 mg/L
treatments and was significantly lower in the 1.0 and the 10.0 mg/L treatments. After 10 days of
exposure, 80% of the nymphs in the 1.0 mg/L treatment and all nymphs in the 10 mg/L treatment
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died. After 20 days of exposure, all nymphs in the 1.0 mg/L treatment died. However, there was
no observed statistical difference between the control and any of the other treatment groups
during this exposure time through 120 days. Between 120 and 250 days of exposure there was
not an observed difference in survival between the control and the lowest treatment group (0.01
mg/L). In contrast, nymph survival in the 0.100 mg/L treatment group started to decrease
compared to the control and the 0.01 mg/L treatment group, with 60% survival in the control
compared to 48.5% survival in the 0.100 mg/L treatment after 150 days of exposure. This
decrease was statistically significantly different from controls. All nymphs in the 0.100 mg/L
treatment group died within 250 days of exposure. While nymph survival in the control was
roughly 40% at the end of the 320-day exposure duration, there was no observed difference
between the control and the lowest treatment group of 0.01 mg/L. Lastly, the paper also reported
observed effects of PFOS on metamorphosis success stating that metamorphosis success was
lower with 15.5% success in the 0.01 mg/L treatment (the only treatment group to have nymphs
survive to this life stage) compared to the control with 92.5%. However, data for this observed
endpoint was not provided in the paper beyond the percentages observed in the control and 0.01
mg/L PFOS treatment group. The specific sample sizes for this endpoint were difficult to
ascertain from the paper as only the total number of test organisms across all test treatments was
provided.
As indicated in the summary of the results above, toxicity values through the experiment
decline with exposure duration. The EPA took all of the author-reported toxicity values between
10 (which was considered to be the start of the chronic exposure) and 200 days of exposure into
account. Independently-calculated ECio values could not be determined given the level of data
that were presented in the paper. Author-reported toxicity values after 10 days of exposure were
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a NOEC of 0.1 mg/L and a LOEC of 1.0 mg/L. The LOEC was associated with a 79% decrease
in nymph survival compared to the control at this time. This NOEC and LOEC resulted in a
MATC of 0.3162 mg/L. Author-reported toxicity values after 150 days of exposure were a
NOEC of 0.01 mg/L and a LOEC of 0.1 mg/L. The LOEC was associated with a 19% decrease
in nymph survival compared to the control at this time. This NOEC and LOEC resulted in a
MATC of 0.03162 mg/L. Lastly, the authors also reported a NOEC of 0.01 mg/L for survival and
an LOEC of < 0.01 mg/L for metamorphosis success after 320 days of exposure. Both of these
toxicity values fell outside the 200-day exposure duration and were not considered for use in the
freshwater chronic criterion calculation since control survival at this point was low (40%) and
considered unacceptable for quantitative use. Additionally, there was insufficient data provided
in the paper to evaluate the reported results for the endpoints at 320 days of exposure. Therefore,
these toxicity values were considered as supporting information and only the toxicity values
from 10 to 200 days of exposure range were considered further for chronic water column
criterion derivation.
The 150-day MATC of 0.03162 mg/L for nymph survival was similar to the author-
reported 10-day and 20-day survival and growth MATCs of 0.0687 and 0.0454 mg/L for
chironomid (MacDonald et al. 2004), and these later toxicity values were therefore more
comparable than the 10-day MATC of 0.3162 mg/L for nymph survival, which was focused on
the effects of PFOS on a much earlier instar of odonate (which has a much longer development
time and life span) in relation to the 20-day MATC of 0.0454 mg/L for chironomid. These results
indicated that PFOS effects to the two aquatic insects was likely similar; however additional data
are needed to fully understand the effects of PFOS to odonates. The MATC for nymph survival
at 150-day reported above was used quantitatively to derive the chronic water column criterion.
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Additionally, the EPA ran additional analyses with some of the other toxicity values for E.
cyathigerum to understand the influence of this study on the overall chronic criterion (see
Section 4.1).
C.2.4.1 Bots et al. (2010) Concentration Response Curve - Enallagma (damselfly)
Publication: Bots et al. (2010)
Species: Damselfly, Enallagma cyathigerum
Genus: Enallagma
EPA-Calculated ECio: Not calculable, concentration-response data not available
C.2.5 Fifth Most Sensitive Freshwater Genus for Chronic Toxicity: Danio (zebrafsh)
Wang et al. (2011) evaluated the full life-cycle effects of PFOS (> 96% purity) on Danio
rerio via a static-renewal study that reported nominal exposure concentrations. This test
evaluated the effects of PFOS on a parental (F0) generation and included breeding trials to assess
the effects of PFOS on an offspring (Fl) generation exposed via maternal transfer. PFOS stock
solutions were prepared in 100% dimethylsulfoxide (DMSO). Adult zebrafish (wild-type strain
AB) were raised and kept at standard laboratory conditions of 28ฐC with a 14:10-hour light:dark
cycle in a recirculation system according to standard zebrafish culture protocols. Water supplied
to the system was filtered by reverse osmosis (pH 7.0-7.5), and Instant Ocean salt was added to
the water to raise the conductivity to a range of 450 to 1,000 |iS/cm (system water). Zebrafish
embryos were obtained from spawning adults in tanks overnight with a sex ratio of 1:1. Embryos
were collected within one hour after spawning and rinsed in embryo medium. High-quality 8-hpf
embryos were divided into four treatment groups: DMSO vehicle control (0.01% v/v), and PFOS
concentrations of 0.005, 0.050, and 0.250 mg/L. Embryos were first exposed to PFOS in a petri
dish (100 embryos/treatment) for five days without media change, and all embryos hatched and
survived in this stage. After five days, the fish were transferred into 2 L tanks for the period of 5-
dpf to 30 dpf, and after that were raised in 9 L tanks (30 fish/tank) until the end of the
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experiment, 150 dpf. Fish were kept in a static system, and 50% water was renewed with freshly
prepared solutions every five days. Each tank was checked for morbid fish on a daily basis, and
water quality was monitored on a weekly basis. Feeding was initiated at day five. Between five
and 14 dpf, fish were fed three times daily with zebrafish larval diet (Aquatic Habitats), and after
14 dpf they were fed twice daily with freshly hatched live Artemia. The experiment was repeated
three times with embryos derived from different parental stocks. At the end of exposure period
(150 dpf or five months), all fish were checked for their sex. However, the method used for
determining sex, as either external morphology or genetic testing, was not stated in the paper.
The EPA assumed external morphology was used and concluded that the effects on sex ratio may
not be reliable since determining sex through external morphology in zebrafish is difficult. A
subsample of 10 male and 10 female fish from each batch were also measured for standard body
length and wet weight. Condition factor (K) was tabulated to determine their overall fitness, and
sperm motility in male F0 fish was also determined after chronic PFOS exposure. The most
sensitive endpoint was F0 parental male sperm density with a chronic value of <0.005 mg/L
PFOS. However, as sperm density was not typically considered an apical endpoint and the
reported effects of PFOS on sperm density did not translate to other reproductive effects (i.e.,
fertilization), this endpoint was not considered further. Instead, the most sensitive apical
endpoint for the F0 generation was considered to be male growth (length and weight) with an
author-reported MATC of 0.01581 mg/L PFOS. However, the EPA was unable to fit a C-R curve
with significant model parameters for the male growth endpoints; and therefore, was unable to
independently verify the reported toxicity value for the F0 generation.
Breeding trials were also carried out to produce F1 offspring. Six different crosses were
employed between F0 females and males to incorporate both the exposure of the same treatment
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groups throughout and crosses between the control and highest treatment group. Specifically, for
the groups exposed to the same treatment throughout the experiment, females were paired with
males in the same treatment group (DMSO control or PFOS-exposed concentrations of 0.005,
0.050, and 0.250 mg/L). For the crosses between the control and the highest treatment group,
some females from the 0.250 mg/L PFOS treatment group were paired with males from the
DMSO controls, and some females from the controls were paired with males from the 0.250
mg/L PFOS treatment group. For each of these crosses, eight randomly selected female fish were
paired with four male fish in two separate spawning tanks with four females and two males per
tank. Spawning was induced every other day for five days, and embryos were used for
monitoring their developmental progress. All eggs from each spawn were evaluated for
fertilization success. Percent fertilization was expressed as the number of fertilized eggs divided
by total number of eggs. Fifty fertilized embryos from each spawn were further monitored for
continuous development. Percent hatch was calculated at 72 hpf. Larvae were also assessed for
their morphological appearance. Percent survival was monitored until 8 dpf. Surviving larvae at
5 dpf with normal morphology were further subjected to behavior assessment (larval swimming
speeds were recorded when they responded to a 70-minute dark to light, 10-minute for each
period, transition stimulation). Following the receipt of treatment level data from the study
authors, the EPA independently calculated an ECio value of 0.0165 (0.01267 - 0.02033) mg/L
for F1 survival. While this ECio has some uncertainty given the wide spacing (lOx) of the
treatment concentrations, this toxicity value was supported by others in the PFOS toxicity
literature (see Section 4.3.2.1.1 and Appendix G).The independently-calculated ECio value of
0.01650 mg/L value for F1 survival from this study was used to derive the freshwater chronic
water column criterion.
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Guo et al. (2019) evaluated the chronic effects of perfluorooctane sulfonate (PFOS
solution of-40% in water purchased from Sigma-Aldrich) to AB strain zebrafish (Danio rerio)
males in a 21-day static-renewal, unmeasured study. Use of official test guidelines were not cited
by the authors. Approximately 3.5-month-old male adult zebrafish were purchased from Taiyuan
fish hatcheries in Shanxi Province, PR China. Prior to exposure, fish were acclimated for 15 days
in a flow-through dechlorinated tap water system (<1% mortality during the holding period) with
the following water quality characteristics and conditions: pH: 7.0-7.4, temperature: 28 ฑ 1ฐC
and a 14:10-hour light:dark photoperiod. The fish were fed a commercially available adult
zebrafish compound feed during both acclimation and exposure. Nominal concentrations of
PFOS dissolved in dechlorinated tap water were reported to be 0 (control), 0.02, 0.04 and 0.08
mg/L. Three replicates were tested at each concentration. A total of 660 fish were divided
equally among the four concentration groups. Water quality was maintained the same throughout
the experiment as during acclimation, as well as to meet the following conditions: D.O. between
5-6 mg/L and total hardness 20.0 mg/L as CaCCb. Exposure media was changed every three
days and aquaria were cleaned during testing. On days 7, 14 and 21, 50 fish from each group
were sacrificed, with 30 fish measured for length and body weight, while the other 20 dissected
on ice to evaluate PFOS concentrations in the liver. The test fish had a mean body weight of 0.19
ฑ 0.03 g and a mean length of 2.5 ฑ 0.3 cm at test initiation. On day seven fish lengths ranged
from >2 cm to <3 cm for all groups, and weights were >0.3 to <0.4 g for the control and 0.02
mg/L exposure. Mean fish weight measured for the 0.04 and 0.08 mg/L treatment groups were
significantly different from the control group after 7 days. At days 14 and 21, the length of fish
in the highest concentration (0.08 mg/L PFOS) was significantly different from the control
group, and the same effect of PFOS on mean fish weight was observed at 14 and 21 days as
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reported at seven days. Therefore, weight was the most sensitive endpoint at 21 days, with a
NOEC and LOEC of 0.02 and 0.04 mg/L PFOS, respectively. No mortality was observed in any
treatment. An independently-calculated ECio could not reliably be estimated for mean fish
weight as the data were sparse, was inconsistent with the author-reported toxicity values, and the
confidence bands were wide. Instead, the EPA's independently-calculated ECio based on mean
body length (in cm) at 21 days was 0.06274 (0.06229 - 0.06318) mg/L PFOS and used
quantitatively to derive the freshwater chronic water column criterion.
C.2.6 Sixth Most Sensitive Freshwater Genus for Chronic Toxicity: Dayhnia (cladoceran)
Logeshwaran et al. (2021) conducted acute and chronic toxicity tests with the
cladoceran, Daphnia cannula, and PFOS-K (perfluorooctancesulfonate potassium salt, > 98%
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 16:8-hour
light:dark photoperiod. The chronic test protocol followed OECD (2012). A PFOS stock solution
(20 mg/mL) was prepared in dimethylformamide and diluted with deionized water to achieve a
concentration of 200 mg/L PFOS. Cladoceran culture medium was used to prepare the PFOS
stock and test solutions. One daphnid (6-12 hours old) was transferred to 100 mL polypropylene
containers containing 50 mL of nominal test solutions (0, 0.001, 0.01, 0.1, 1.0 and 10 mg/L
PFOS). Each test treatment was replicated ten 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. At the higher test concentrations (1 and
10 mg/L) reproduction was completely inhibited. No mortality occurred in the controls and
lowest test concentration. However, reproduction was inhibited at the lowest test concentration.
The author-reported 21-day NOEC and LOEC, based on average offspring in each brood and
total live offspring, was < 0.001 and 0.001 mg/L PFOS, respectively. Additionally, the author-
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reported 21-day NOEC and LOEC based on the days to first brood was 0.001 and 0.01 mg/L,
respectively, resulting in an MATC of 0.003162 mg/L. The EPA could not independently
calculate 21-day ECio values for any of the endpoints given the level of data provided in the
paper by the study authors. And while the endpoints of mean offspring per each brood and total
living offspring appear to be more sensitive than the days to first brood, they result in less than
LOECs of 0.001 mg/L and are not consistent with other chronic toxicity values for this species.
Therefore, the author-reported MATC of 0.003162 mg/L for the days to first brood was used to
derive the freshwater chronic water column criterion.
Exclusion of the D. carinata SMCV under the basis of being an overly sensitive outlier
(relative to D. magna and the chronic data overall except for aquatic insects [A', triangulifer and
C. dilutus]) had the possibility that the Daphnia GMCV could be underproductive. Conversely,
excluding the D. 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 D. carinata chronic value was also an MATC,
calculated as the geometric mean of the NOEC (0.001 mg/L) and LOEC (0.01 mg/L) from a 10X
dilution series, meaning the MATC was influenced by a relatively low NOEC. However,
offspring-based endpoints reported by (Logeshwaran et al. 2021) suggest a LOEC of <0.001
mg/L. Using both the D. magna and D. carinata SMAVs resulted in protective & Daphnia
GMCV based on the Daphnia data as a whole. Finally, the chronic PFOS freshwater criterion
(i.e., 0.00025 mg/L) is four times lower than the offspring-based LOECs (i.e., 0.001 mg/L)
reported by (Logeshwaran et al. 2021) and should be protective of D. carinata based on all
endpoints measured by (Logeshwaran et al. 2021).
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Drottar and Krueger (2000e) reported the results of a life-cycle, 21-day renewal,
measured test of PFOS (potassium salt, CAS # 2795-39-3, 90.49% purity) with Daphnia magna.
The GLP test was conducted at Wildlife International, Ltd. in Easton, MD in February, 1999.
The test followed (U.S. EPA 1996c). D. magna used for the test were less than 24 hours old at
test initiation. Dilution water was 0.45 |im filtered and UV sterilized well water [total hardness:
124 (120-128) mg/L as CaC03; alkalinity: 169 (164-172) mg/L as CaC03; pH: 8.2 (8.0-8.3);
TOC: <1.0 mg/L; and conductivity: 329 (315-340) |imhos/cm]. Photoperiod was 16:8-hours
light:dark with a 30 minute transition period. Light was provided at an intensity of 329-383 lux.
A primary stock solution was prepared in dilution water at 46 mg/L. It was mixed until all test
substance was dissolved prior to use. After mixing, the primary stock was proportionally diluted
with dilution water to prepare the five additional test concentrations. Exposure vessels were 250
mL plastic beakers containing 200 mL of test solution. The test employed 10 replicates of one
daphnid each in six measured test concentrations plus a negative control. Nominal concentrations
were 0 (negative control), 1.4, 2.9, 5.7, 11, 23, and 46 mg/L. Mean measured concentrations
were < 0.458 mg/L (the LOQ), 1.5, 2.9, 5.6, 12, 24, and 48 mg/L, respectively. Analyses of test
solutions were performed at Wildlife International Ltd. using HPLC/MS. The mean percent
recovery of matrix fortifications analyzed concurrently during sample analysis was 104%.
Measured values of new samples ranged from 94 to 121% of nominal. Measured values from the
old solutions ranged from 90 to 108%) of nominal. PFOS was stable throughout the renewal
periods. Dissolved oxygen in new and old test concentrations ranged from 8.3-8.9 mg/L in the
negative controls and 8.3-9.0 mg/L at the NOEC of 12 mg/L. Similarly, pH ranged from 8.1-8.4
and 8.2-8.5, respectively, and test temperature from 19.4-20.1ฐC (negative control and at the
NOEC). Cumulative young in the 1.5, 2.9, 5.6, 12, and 24 mg/L treatment groups was 100, 100,
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100, 90, and 0%, respectively. After 48 hours, cumulative young of the second generation in the
negative control was 95%. The 21-day NOEC (survival, growth, and reproduction) was 12 mg/L.
The 21-day LOEC was 24 mg/L and the calculated MATC is 16.97 mg/L. No second-generation
D. magna survived the 24 mg/L treatment. The independently-calculated ECio based on
cumulative young was 11.19 (10.50 - 11.89) mg/L and used quantitatively to derive the
freshwater chronic water column criterion.
Boudreau (2002) also conducted a chronic life-cycle 21-day renewal, unmeasured test of
PFOS (potassium salt, CAS # 2795-39-3, 95% purity) with Daphnia magna 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. (2003a). The test followed ASTM (1999a). D. magna used
for testing were less than 24 hours old at test initiation. D. magna were obtained from a brood
stock (Dm99-23) at ESG International (Guelph, ON, Canada). Dilution water was clean well
water. Hardness was softened by addition of distilled deionized water to achieve a range of 200-
225 mg/L of CaC03. Photoperiod was 16:8-hours light:dark 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 120 mL of test solution. All
toxicity testing involved four replicates of three daphnids each in five nominal test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 6, 13,
25, 50, and 100 mg/L. The test was conducted in environmental chambers at 21 ฑ1ฐC. Authors
noted that temperature and pH were measured at the beginning and end of study, but the
information was not reported. Survival of daphnids in the negative control was 100%. The 21-
day NOEC (survival and reproduction) was 25 mg/L. The 21-day LOEC was 50 mg/L and the
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calculated MATC is 35.36 mg/L. The independently-calculated ECio based on survival was
16.35 (7.377 - 25.33) mg/L and was used quantitatively to derive the freshwater chronic water
column criterion.
Ji et al. (2008) conducted chronic life-cycle tests of the effects of PFOS (acid form, CAS
# 1763-23-1, purity unreported) on Daphnia magna. Tests were done under renewal conditions
over a 21-day period. The test followed OECD (1998). D. magna used for testing were obtained
from brood stock cultured at the Environmental Toxicology Laboratory at Seoul National
University (in South Korea). 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 16:8-hours light:dark as was used for daphnid culture. 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),
0.3125, 0.625, 1.25, 2.5, and 5 mg/L. Test temperature was 21 ฑ1ฐC. Authors noted water quality
parameters (pH, temperature, conductivity, and D.O.) were measured after changing the medium,
but the information was not reported. Survival of daphnids in the negative control was 100%.
The author reported D. magna 21-day NOEC for the reproductive endpoint of number of young
per brood was 1.25 mg/L. The author reported 21-day LOEC for the same endpoint was 2.5
mg/L. The calculated MATC was 1.768 mg/L. In the independent verification of the toxicity
values, the EPA recalculated the reproductive endpoint noted to be the number of young per
brood. This recalculated reproductive endpoint took the full effects of PFOS into account as it
was representative of the full life cycle. The calculated ECio for D. magna was 1.051 (0.2680 -
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1.834) mg/L. The independently-calculated ECio of 1.051 mg/L was used quantitatively to derive
the freshwater chronic water column criterion.
Li (2010) conducted a chronic life-cycle 21-day test on the effects of PFOS (potassium
salt, >98% purity) on Daphnia magna. The test followed OECD (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 ASTM medium (0.12 g/L CaS04.2H20,
0.12 g/L MgSC>4, 0.192 g/L NaHCCb, and 0.008 g/L KC1 - calculated total hardness 169 mg/L as
CaCCb). Photoperiod was 16:8-hours light:dark at an unreported light intensity. A primary stock
solution was prepared in dilution water and did not exceed 400 mg/L. 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. Nominal concentrations
were 0 (negative control), 0.5, 1, 5, 10, and 20 mg/L. Test temperature was maintained at 20
ฑ1ฐC. Water quality parameters measured in test solutions were not reported. Survival of
daphnids in the negative control was 96.7%. The D. magna 21-day NOEC (reproduction - no.
young per female) was 1 mg/L. The 21-day LOEC was 5 mg/L and the calculated MATC was
2.236 mg/L. The independently-calculated toxicity value (ECio) based on total neonates per
female was 3.030 (-1.280 - 7.340) mg/L and was used quantitatively to derive the freshwater
chronic water column criterion.
Yang et al. (2014) evaluated the chronic 21-day renewal, measured test of PFOS
(potassium salt, CAS # 2795-39-3, 99% purity) with Daphnia magna. The test followed (ASTM
1993). D. magna used for the test were donated by the Chinese Research Academy of
Environmental Sciences, and less than 24 hours old at test initiation. Dilution water was
dechlorinated tap water (pH, 7.0 ฑ 0.5; D.O., 7.0 ฑ 0.5 mg/L; total organic carbon, 0.02 mg/L;
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and total hardness, 190.0 ฑ0.1 mg/L as CaCCb). Photoperiod was 12:12-hours light:dark at an
unreported light intensity. A primary stock solution was prepared by dissolving PFOS in
deionized water and cosolvent DMSO. The primary stock was 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 10 replicates of
one daphnid each in six test concentrations (measured in low and high treatments) plus a
negative and solvent control. Nominal concentrations were 0 (negative and solvent controls),
2.00, 2.60, 3.38, 4.39, 5.71 and 7.43 mg/L. Mean measured concentrations before and after
renewal were 1.74 and 1.98 mg/L (lowest concentration) and 6.78 and 7.54 mg/L (highest
concentration). Analyses of test solutions were performed using HPLC/MS and negative
electrospray ionization. The concentration of PFOS 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 105%. Test
temperature was maintained at 22 ฑ2ฐC. The D.O. and pH were reported as having been
measured, but results are not reported. Negative and solvent control survival was 100%. The D.
magna 21-day ECio for reproduction was reported to be 2.26 mg/L from the study authors and
4.17 mg/L for survival. The independently-calculated ECio based on survival was 2.610 (1.291 -
3.929) mg/L and was used quantitatively to derive the freshwater chronic water column criterion.
Lu et al. (2015) conducted a chronic life-cycle 21-day renewal, unmeasured test of PFOS
(purity 98%) with Daphnia magna. The test followed OECD (2012). D. magna used for the test
were originally obtained from the Chinese Center for Disease Control and Prevention (Beijing,
China) and cultured in the laboratory according to the International Organization for
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Standardization (ISO 1996). Daphnids were less than 24 hours old at test initiation. Dilution
water was the same used for daphnid culture and was reconstituted according to OECD (2004)
with a total hardness of 250 mg/L as CaCCb, as calculated based on the recipe provided, and pH
ranging from 7.7 to 8.4. Photoperiod was 16:8-hours light:dark at an unreported light intensity.
The test solution was prepared immediately prior to use by diluting the stock solution with
culture medium. Exposure vessels were 100 mL glass beakers containing 45 mL of test solution.
The test employed 20 replicates of one daphnid each in six nominal test concentrations plus a
negative control. Nominal concentrations were 0 (negative control), 0.008, 0.04, 0.2, 1, and 5
mg/L. Exposure water quality was checked daily and maintained at 20 ฑ1ฐC, pH of 7.2 ฑ0.3, and
D.O. of 5.3 mg/L. Negative control survival was 100%. The author reported D. magna 21-day
NOEC (no. offspring per brood per female) was 0.008 mg/L and the 21-day LOEC was 0.04
mg/L. The calculated MATC was 0.0179 mg/L and the independently-calculated ECio was
0.001818 (-0.0000395 - 0.003675) mg/L for the same endpoint. Other endpoints, including
growth and other reproductive endpoints, could not be independently-calculated by the EPA. The
independently-calculated ECio from this study was acceptable for quantitative use to derive the
freshwater chronic water column criterion.
Liang et al. (2017) conducted a chronic life-cycle 21-day renewal, unmeasured test of
PFOS (potassium salt, CAS # 2795-39-3, >98% purity) with Daphnia magna. The test organisms
were originally obtained from State Key Laboratory of Environmental Aquatic Chemistry (Eco-
Environmental Sciences of Chinese Academy of Sciences, Beijing) and cultured in the
laboratory according to Revel et al. (2015). Daphnids were less than 24 hours old at test
initiation. Dilution water was artificial medium "M4 (Elendt)" at 20ฐC and pH 7. Photoperiod
was 16:8-hours light:dark at an unreported light intensity. The test solution was prepared
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immediately prior to use by diluting the stock solution with M4 medium. Exposure vessels were
80 mL glass beakers containing an unspecified volume of test solution. The test employed 10
replicates of one daphnid each in six nominal test concentrations plus a negative control.
Nominal concentrations were 0 (negative control), 1, 2, 4, 8, and 16 mg/L. No mention was made
of water quality being checked during the exposure. Negative control survival was 100%. The D.
magna 21-day NOEC (days to 1st brood, intrinsic rate of natural increase, r) was 4 mg/L. The
21-day LOEC was 8 mg/L and the calculated MATC was 5.657 mg/L. The independently-
calculated ECio based on survival was 3.596 (2.1207 - 5.0704) mg/L and used quantitatively to
derive the freshwater chronic water column criterion.
Yang et al. (2019) evaluated the chronic effects of perfluorooctane sulfonate, potassium
salt (PFOS-K, CAS# 2795-39-3, 98% purity, purchased from Sigma-Aldrich in St. Louis, MO)
on Daphnia magna via a 21-day unmeasured, static-renewal test that evaluated growth and
reproductive effects. D. magna cultures were 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 laid out in OECD Guideline 202 and all testing followed
OECD Guideline 211. Cultures were fed green algae daily and were acclimated for two to three
weeks before testing. The 21-day chronic study had nominal concentrations of 0 (control),
0.00000124, 0.00000188, 0.0000281 and 0.00000420 mol/L (or 0 (control), 0.6674, 1.012,
1.512, and 2.261 mg/L given the molecular weight of the form of PFOS used in the study, CAS #
2795-39-3, of 538.22 g/mol). Each neonate (12-24 hours old) was placed in a 100 mL glass
beaker, in which there were 10 replicates, each filled with 80 mL of test solution maintained at
20 ฑ1ฐC and a 16:8-hour light:dark photoperiod with a light intensity maintained at 1000 - 1500
lux. D. magna were fed green algae and test solutions were renewed every 72 hours. Test
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organisms were counted daily, with any young also removed. The author-reported NOEC and
LOEC for reproduction (measured as mean offspring proportion relative to control at 21 days)
was <0.6674 and 0.6674 mg/L PFOS, respectively. The author-reported NOEC and LOEC for
growth (measured as length) was 0.6674 and 1.012 mg/L PFOS (MATC = 0.8218 mg/L). The
independently-calculated ECio values for reproduction and growth are 0.3773 and 0.9093 mg/L,
respectively. However, the reproduction ECio of 0.3773 mg/L was determined to be less
statistically robust as the independently-calculated toxicity values were control normalized and
could not be weighted given the level of data provided by the study authors in the paper.
Therefore, the independently-calculated ECio for growth of 0.9093 (0.7423 - 1.076) mg/L was
used quantitatively to derive the freshwater chronic water column criterion freshwater.
Seyoum et al. (2020) evaluated the chronic effects of perfluorooctane sulfonic acid
(PFOS, CAS# 1763-23-1, > 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. D. magna ephippia were purchased from MicroBioTests Inc. (Belgium) and were
activated by rinsing in tap water. Ephippia were hatched by incubating at 20-22 ฐC for 72 to 90
hours in standard freshwater under a continuous light intensity (6,000 lux). Newly hatched
neonates (<24-hour old) were fed a suspension of Spirulina micro-algae two hours before testing.
Nominal concentrations of 0 (control), 1, 10 and 25 |iM (or 0 (control), 0.5001, 5.001, and 12.50
mg/L given the molecular weight of the form of PFOS used in the study, CAS # 176-23-1, of
500.13 g/mol) were prepared by mixing the respective amounts of PFOS in dimethyl sulfoxide
(DMSO). Ten <24-hour old neonates, exposed in triplicate, were placed into 250 mL
crystallization dishes with 100 mL of test solution. A mean temperature of 23ฐC, D.O. of 8 to 9
mg/L, total hardness of 250 mg/L as CaC03, pH of 7.5 ฑ0.25 and salinity of 0.02% were
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reported in the exposure water. D. magna were fed a mixture of Spirulina microalgae and yeast
(Saccharomyces cerevisiae) daily during the test, and 50% of the test solution was changed every
other day. Neonates were counted daily and removed. At day 21, neonate counts were reported to
be highest in the control with >40 to < 60 neonates, and >20 to <40 neonates were reported at the
0.5001 and 5.001 mg/L (or 1 and 10 |iM) concentrations, respectively. Neonate counts for the
12.50 mg/L (or 25 |iM) concentration were not reported. A reproductive NOEC of 0.5001 mg/L
and a LOEC of 5.001 mg/L were reported by the study authors, resulting in an MATC of 1.581
mg/L. This LOEC of 5.001 mg/L was associated with a 42.95% decrease in reproduction
(measured as the mean number of daphnids at 21 days) compared to control. An independently-
calculated ECio value could not be determined as the EPA was unable to fit a model with
significant parameters. Instead, the author-reported MATC of 1.581 mg/L PFOS was used
quantitatively to derive the chronic water column criterion for freshwater.
C.2.7 Seventh Most Sensitive Freshwater Genus for Chronic Toxicity: Salmo (salmon)
Atlantic salmon, Salmo salar, embryos were evaluated by Spachmo and Arukwe (2012)
via a 56-day unmeasured exposure to PFOS (98% purity). Eggs were obtained from Lundamo
Hatcheries, Norway (Aquagen) and transported to the Norwegian University of Science and
Technology Centre of Fisheries and Aquaculture in Trondheim, Norway. The eggs were kept in
plastic tanks (25 L) at 5-7ฐC with filtered, re-circulating and aerated water. Approximately one-
third of the water volume was changed once per week. The eggs and larvae were exposed to
PFOS (100 |ig/L) for 49 days representing the developmental period from 404 to 679-degree
days. PFOS was dissolved in methanol (carrier solvent: 0.01%) and control group was exposed
to the carrier solvent only. Hatching occurred at 20 calendar days after start of exposure, at an
effective developmental age of 504-degree days, after which riverbed environment was
simulated by tank bed gravel and continuous water flow. Fish sampling was performed at 21, 35,
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49 and 56 calendar days after exposure, or at respective developmental ages of 549, 597, 679 and
721 degree days. The exposure was terminated at 679-degree days, and 712-degree days
represents the end of a one-week exposure-free recovery period. Thus, day 49 sampling was
performed 24 hours after terminating the exposure and no exposure related differences in
hatching rate were observed. The 49-day growth NOEC and LOEC were 0.10 and >0.10 mg/L
PFOS, respectively. These data are deemed quantitative and used to derive the freshwater
chronic water column criterion.
C.2.8 Eighth Most Sensitive Freshwater Genus for Chronic Toxicity: Pimeyhales (minnow)
Drottar and Krueger (2000d), associated with Wildlife International, conducted a good
laboratory practice (GLP) 47-day flow-through measured early life-stage toxicity test with <24-
hour old P. promelas embryos. A primary stock solution was prepared by dissolving PFOS
(90.49% purity) in dilution water at a concentration of 88.4 mg a.i./L, then proportionally diluted
with dilution water to prepare five secondary stock solutions at concentrations of 44.2, 22.1,
11.0, 5.52 and 2.76 mg a.i./L. Stock solutions were prepared every three to four days during the
test. The five stocks were injected into the diluter mixing chambers (at a rate of 6.0 mL/minute)
where they were mixed with dilution water (at a rate of 116 mL/minute) to achieve the desired
test concentrations. The water used for culturing and testing was freshwater obtained from a well
approximately 40 meters deep located on the Wildlife International Ltd. site. The well water was
characterized as moderately-hard water. The well water was passed through a sand filter to
remove particles greater than approximately 25 [j,m and then pumped into a 37,800-L storage
tank where the water was aerated with spray nozzles. Prior to delivery to the diluter system, the
water again was filtered (0.45 (j,m), then passed through a UV sterilizer to remove
microorganisms and particles. Fathead minnow embryos used in this test originated from
cultures maintained by Wildlife International Ltd., Easton, MD. The embryos were removed
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from spawning substrates and examined under the dissecting microscope to select healthy
specimens at approximately the same stage of development. Embryos collected for use in the test
were from six individual spawns. Embryos were exposed to a geometric series of six test
concentrations and a negative (dilution water) control under flow-through conditions at 24.5ฐC,
pH of 8.2, total hardness of 140 mg/L as CaCCb and a photoperiod of 16:8-hours light:dark. Four
replicate test chambers (9 L glass aquaria) were maintained in each treatment and the control
group. Each test chamber contained one incubation cup with 20 embryos, resulting in a total of
80 embryos per treatment. The exposure period included a five-day embryo hatching period, and
a 42-day post-hatch juvenile growth period. Nominal test concentrations were 0.14, 0.29, 0.57,
1.1, 2.3 and 4.6 mg/L a.i. Mean measured test concentrations (0, 0.15, 0.30, 0.60, 1.2, 2.4 and 4.6
mg/L) were determined from samples of test water collected from each treatment and the control
group at the beginning of the test, on day four, at weekly intervals during the test, and at test
termination. To start the test, embryos less than 24 hours old were collected from cultures and
groups of one and two individuals were impartially distributed among incubation cups until each
cup contained 20 embryos. One cup was then placed in each treatment and control test chamber.
Twice during the next twenty-four hours and daily thereafter, all dead embryos were counted and
removed from the cups to avoid contaminating viable embryos. All eggs that remained were
considered viable. Dead embryos continued to be removed daily. After hatching, the larvae were
counted and released into the test chambers, where exposure continued until test termination.
Observations of mortality and other clinical signs were made daily during the test. Time to hatch,
hatching success, growth, and survival were monitored in each treatment and control group. The
most sensitive 47-day chronic value (MATC) of 0.4243 mg/L PFOS was based on post-hatch
survival as reported by the study authors. The independently-calculated ECio based on survival
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was 0.4732 (0.3308 - 0.6156) mg/L and used quantitatively to derive the freshwater chronic
water column criterion.
Ankley et al. (2005) also exposed Pimephalespromelas to PFOS (potassium salt, > 98%
pure) under flow-through measured conditions for 21 days. Stock solutions were prepared by
dissolving crystals in Lake Superior control water with stirring (mean measured test conditions:
25ฐC, pH of 7.3, total hardness of 46 mg/L as CaCC>3, alkalinity of 40 mg/L as CaCC>3 and D.O.
of 6.2 mg/L). Two stock solutions of approximately 9.7 and 97 mg/L were used to span the
desired range of target concentrations in test water. Final test concentrations were generated by
appropriate dilution of the PFOS stocks with Lake Superior water and were supplied to the test
tanks at a flow rate of approximately 45 mL/min. Sexually mature fathead minnows (six to seven
months old) obtained from the on-site culture facility were used for the toxicity test. Eight pairs
of fish (one male and one female) were exposed at each treatment level, 0 (control), 0.03, 0.1,
0.3, and 1.0 mg PFOS/L. Assays were conducted using glass aquaria containing 10 L of test
solution, with two pairs of fish separated by perforated nylon screening in each tank.
Reproductive viability of the fish used for the test was documented during a 27-day acclimation
phase in the same tanks in which the tests were conducted. The number of eggs spawned by each
pair was evaluated daily by inspecting the underside of a polyvinyl chloride spawning tile placed
on the bottom of the test chambers. Egg fertility was assessed using an optical microscope. The
animals were held at 25 ฑ1ฐC under 16:8-hour light:dark photoperiod and fed frozen brine
shrimp to satiation twice daily. Conditions during the 21-day reproduction phase of the PFOS
exposure were the same as during the acclimation phase. To evaluate possible early
developmental toxicity of PFOS, 50 to 75 eggs from single viable spawns were collected during
the final 7-day of the reproduction phase of the test. A subset of eggs was reserved for the
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determination of PFOS concentrations. Embryos were held in 300 mL Pyrex beakers in the same
aquaria as the parental fish. Embryos hatched within four to five days and thereafter were fed
live brine shrimp twice daily. After 12 days, fry were randomly sampled for PFOS analysis and
to reduce the number of animals per chamber to <30. Remaining fry were maintained in a larger
chamber (1 L plastic container) within the original tank. Developing fish were inspected daily to
assess survival. After 24 days, they were anesthetized and weighed. A subset of the fry was
collected for PFOS measurements, while others were preserved in Bouin's fixative for
histological analyses. The authors reported a 21-day EC so (fecundity) of 0.23 mg/L PFOS, and a
chronic value of 0.4794 mg/L PFOS for percent hatch (21-day), probability of survival, and
larval weight endpoints (21-day (F0) + 24-day (Fl)). The independently-calculated ECio was
0.05101 (0.0408 - 0.0613) mg/L based on fecundity and used quantitatively to derive the
freshwater chronic water column criterion.
Suski et al. (2021) reported the chronic toxicity of PFOS-K (perfluorooctancesulfonate
potassium salt, CAS # 2795-39-3, > 98%, purchased from Sigma-Aldrich) on the fathead
minnow, Pimephalespromelas. Adult (5-month old) fathead minnows were purchased from a
commercial supplier (Aquatic Biosystems) and were sexually mature when the test was initiated.
Fish were fed twice a day and held in dechlorinated tap water at test conditions (mean
conditions: 24.96ฐC, D.O. of 7.68 mg/L, pH 7.9 and conductivity of 347.3 |iS/cm). Stock
solutions of PFOS (150 mg/L) were made without a solvent and prepared weekly with stock
solutions shaken at 80 rpm for 24 hours to ensure mixing. Test solutions were made by diluting
the stock with dechlorinated tap water and shaking the solutions for 10 minutes prior to water
exchanges. Half of the total volume (10 L) in each exposure 5-gallon polycarbonate tank was
renewed three times per week. Measured PFOS concentrations were 0.14 (control), 44, 88, 140
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and 231 |ig/L PFOS (or 0.00014 (control), 0.044, 0.088, 0.14, and 0.231 mg/L PFOS). Each test
treatment was replicated six times for each treatment and consisted of two females and one male
per tank with exposures lasting 42 days. Tanks were expected daily for eggs and all eggs
collected were assumed to be per single female regardless of the number of females per tank. On
the last week of testing, eggs were carried through hatching in their respective test treatments,
and 20 larval fish per concentration were exposed for an additional 21 days to investigate
developmental effects. One liter polypropylene beakers were used for the F1 generation exposure
with solutions renewed daily. Survival of adult fathead minnows in the control and two lowest
test concentrations was >80% at test termination. Survival of male fish in the highest test
treatment was significantly less than male control fish, and while female survival was also less
compared to control fish, the effects were not significant. The mean number of spawning events
per female was also reduced in the two high test treatments, but the effect was only significant in
the 140 |ig/L (0.14 mg/L) treatment. Larval survival in the F1 generation was significantly
reduced in the highest test treatment. The most sensitive endpoint from the study was a
significant decrease in the mean mass of individuals in the larval F1 generation with reported
values of 3.76, 3.53, 3.09, 2.64 and 2.00 mg for the test treatments of control, 0.044, 0.088, 0.14,
and 0.231 mg/L PFOS, respectively. The author-reported NOEC and LOEC, based on growth in
the F1 generation, were 0.044 (6% reduction in growth compared to controls) and 0.088 mg/L
PFOS (associated with an 18% reduction in growth), respectively, with a MATC of 0.06223
mg/L. The independently-calculated ECio value was 0.0549 (0.0396 - 0.0701) mg/L and used in
the derivation of the freshwater chronic column criterion.
C.2.9 Ninth Most Sensitive Freshwater Genus for Chronic Toxicity: Procambarus (crayfish)
Funkhouser (2014) conducted a chronic 28-day renewal test of PFOS (potassium salt,
>98% purity) with a crayfish species, Procambarus fallax (f. virginalis). The study was
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conducted as part of a Master's thesis at Texas Tech University, Lubbock, TX. Juvenile P.fallax
(4-weeks old, 0.056 g) used for the test were originally purchased from a private collector. The
crayfish reproduced for several generations before being used for experiments. Based on an
average reproductive age of 141-255 days, an interclutch period of 50-85 days, and a brooding
time of 22-42 days, the author estimated the experimental animals to be stage F4-F6 (Seitz et al.
2005). Dilution water was moderately hard reconstituted laboratory water (3.0 g CaSC>4, 3.0 g
MgSC>4, 0.2 g KC1, and 4.9 g NaHCCb added to 50 L deionized water). Photoperiod was 14:10-
hours light:dark at an unreported light intensity. PFOS was dissolved in dilution water to prepare
the test concentrations. Exposure vessels were 1 L polypropylene containers containing 500 mL
of test solution. The test employed eight replicates of one crayfish each in five test
concentrations plus a negative control. Nominal concentrations were 0 (negative control), 0.2,
0.5, 1.3, 3.2, 8 and 20 mg/L. Exposure concentrations were reportedly measured, but
concentrations were not provided. Analyses of test solutions were performed using LC-MS/MS.
Standards were used as part of the analytical method, but details were not reported. The reporting
limit was 0.010 mg/L. Experiments were conducted in an incubator set at 25 ฑ1ฐC and covered
with plastic opaque sheeting to limit evaporation. No other water quality parameters were
reported as having been measured in test solutions. Negative control survival was 85% after 28
days. The 28-day LC20 was reported as 0.167 mg/L. An independently-calculated EC10 could not
be calculated given the level of data that were presented in the paper. The study author-reported
LC20 value was used quantitatively to derive the freshwater chronic water column criterion.
C.2.10 Tenth Most Sensitive Freshwater Genus for Chronic Toxicity: Moina (cladoceran)
Ji et al. (2008) conducted a chronic life-cycle test of the effects of PFOS (acid form,
CAS # 1763-23-1, purity unreported) on Moina macrocopa. The test was performed under
renewal conditions over a 7-day period. TheM macrocopa test followed a protocol developed
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and reported by Sutherland and Krueger (2001) that was similar to OECD (1998), but with slight
modification. M. macrocopa used for testing were obtained from brood stock cultured at the
Environmental Toxicology Laboratory at Seoul National University (in South 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 16:8-hours
light: dark as was used for daphnid culture. 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), 0.3125, 0.625, 1.25, 2.5, and
5 mg/L. Test temperature was 25 ฑ1ฐC. Authors noted that water quality parameters (pH,
temperature, conductivity, and D.O.) were measured after changing the medium, but the
information was not reported. Survival of daphnids in the negative control was 100%. The author
reported M. macrocopa 7-day LOEC for the reproductive endpoint of number of young per
surviving adult was 0.3125 mg/L. In the independent verification of the toxicity value, the EPA
recalculated the reproductive endpoint to be the number of young per starting adult (instead of
surviving adult). This recalculated reproductive endpoint took the full effects of PFOS into
account as it was representative of the full life cycle. The independently-calculated ECio forM
macrocopa was 0.1789 (0.041 - 0.399) mg/L and used to derive the freshwater chronic water
column criterion.
C.2.11 Eleventh Most Sensitive Freshwater Genus for Chronic Toxicity: Brachionus (rotifer)
Zhang et al. (2013) conducted a chronic life-cycle renewal test of PFOS (potassium salt,
CAS # 2795-39-3, >98% purity) with Brachionus calyciflorus. The test duration was five days in
a full-life cycle test (primary emphasis) and 28 days in a multi-generation population growth test
(secondary emphasis - only two exposure concentrations plus a control). Test organisms were
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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 (16:8-
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 PFOS (1,000 mg/L)
were prepared by dissolving the solid in deionized water via sonication. After mixing, the
primary stock was proportionally mixed 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 each in individual cells.
Treatments were repeated six times. Nominal concentrations were 0 (negative control), 0.25, 0.5,
1.0, and 2.0 mg/L. PFOS 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
PFOS measured every eight hours over a 24-hour period in rotifer medium with green algae
incurs minimal change in the concentration range 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 was
not provided for the life-cycle test. The B. calyciflorus 5-day LOEC (net reproductive rate and
intrinsic rate of natural increase) was 0.25 mg/L. The author-reported value (<0.25 mg/L) was
used quantitatively to derive the freshwater chronic water column criterion.
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C.2.12 Twelfth Most Sensitive Freshwater Genus for Chronic Toxicity: Xiphophorus (swordtail
fish)
The toxicity of PFOS (potassium salt, > 98% purity) to the swordtail fish, Xiphophorus
helleri, was evaluated by Han and Fang (2010). A PFOS stock solution (250 mg/L) was
prepared by dissolving crystals in dechlorinated tap water (from the same water source as that
used in fish keeping). Six- to seven-month old adult swordtails were purchased from a local fish
farm with no water pollution. The fish were separated by sex into different aquaria. Both the
males and females were acclimated for eight weeks under semi-static conditions in charcoal
filtered, aerated tap water at 27 ฑ1ฐC with a 14:10-hour light:dark photoperiod. The water in
each aquarium was completely renewed every 48 hours. The fish were fed once daily in the
morning with flake food and once daily at dusk with frozen blood worms. Adult male fish were
then randomly distributed into 30 L tanks containing 20 L dechlorinated tap water or a
corresponding PFOS solution. Swordtail fish were exposed to 0 (control), 0.1, 0.5 or 2.5 mg/L
PFOS for three weeks and then transferred into clean water for one-week recovery. Every day,
half of the water in each tank was replaced with fresh water, and the fish were exposed to the
appropriate concentrations daily. Exposure conditions were the same as those during the
acclimation period. Each aquarium housed 10 swordtails. Three aquaria were used for each
exposure concentration and for the controls, resulting in three full biological replicates for each
exposure group. Body, liver and testis weights were determined at 7, 14, 21 and 28 days after
ice-bath anaesthetization. The livers were weighed immediately, then frozen in liquid nitrogen
and stored at -80ฐC for RNA extraction. The hepatosomatic index (HSI) and gonadal somatic
index (GSI) values were also calculated. Nonpregnant adult female fish were housed under the
same exposure conditions as the males for the six-week exposure period. At the same time, to
ensure impregnation of the females, nine adult females were paired with three adult males and
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kept in each aquarium for one week, after which the males were taken out. There were also three
biological replicates for each exposure group. One pregnant female per aquarium was isolated
and housed until giving birth. Larvae were maintained in clean water for up to 14 days after birth
to calculate their survival rate. At the end of the exposure period, the survival rate, HSI and GSI
values of all groups were determined. The total number of puerperal females and females with
eggs or embryos in each group was recorded to determine their corresponding ratios. More than
100 adult swordtails (with a male:female ratio of about 1:3) were housed together to obtain at
least 240 juveniles (20-30 days old). All of the fry were then randomly separated into two
exposure groups (0 and 0.1 mg/L) and kept under the same housing conditions as the males.
Each tank contained 40 fry. There were also three biological replicates in each group. After a 90-
day exposure period, the HSI, GSI, and condition factor (CF) values and the sex ratio of each
group were calculated by sex category. Body length from the snout to the end of the caudal fin
and sword length from the distal end of the middle rays of the caudal fin to the tip of the sword
were measured for each young male. After an extended period of stable breeding, part of the
juveniles became young females and some of them were with eggs, embryos or puerperal. So,
just like adult females, the total number of puerperal females and females with eggs or embryos
in each group were recorded as a single entity to determine their corresponding ratios. The 4-
week (adult male), 6-week (adult female) and 90-day (juvenile female and male) survival chronic
values were >2.5, 1.118 and >0.1 mg/L PFOS, respectively. The study-author reported survival
chronic value for offspring of females exposed for six weeks was 0.2236 mg/L PFOS, and the
90-day growth and percent females with eggs chronic value was <0.1 mg/L PFOS. The
independently-calculated ECio for adult female survival was 0.5997 (0.2336 - 0.9658) mg/L,
which was acceptable for quantitative use.
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C.2.13 Thirteenth Most Sensitive Freshwater Genus for Chronic Toxicity: Xenopus (frog)
Lou et al. (2013) evaluated the chronic toxicity of PFOS to the African clawed frog,
Xenopus laevis. PFOS (98% purity) stock solutions (8 mg/mL) were prepared by dissolving in
DMSO every four weeks and stored at 4ฐC. Stock solutions were diluted by charcoal-filtered tap
water to prepare test water. DMSO concentrations were 0.001% (v/v) in all tanks including the
solvent control group. The same charcoal-filtered tap water (pH 6.5-7.0, D.O. >5 mg/L, and total
water hardness, as CaC03, of approximately 150 mg/L) was used to raise X laevis frogs and
tadpoles. Adult female and maleX laevis (3 years old, obtained from Nasco, USA.) were raised
separately in glass tanks at 22 ฑ2ฐC with a 12:12-hour light:dark cycle and fed with chopped
pork liver (commercial amphibian diet three times a week). A pair of X laevis was injected by
human chorionic gonadotropin to induce breeding. Fertilized eggs were incubated in the same
dechlorinated tap water at 22 ฑ2ฐC for six days (and were fed live Artemia starting on the 5th
day). On the fifth day postfertilization, tadpoles at NF stage 46/47 were exposed to PFOS
(nominal: 0.0001, 0.001, 0.100 and 1.0 mg/L; measured: 0, 0.00009, 0.001, 0.1117, 0.7160
mg/L) until two months post-metamorphosis. Each exposure group and control group consisted
of three replicated tanks. Each tank with 18 L water was assigned randomly 25 tadpoles. The
tadpoles were fed with live Arlemia three times daily. After metamorphosis, the juvenile frogs
were fed with live Artemia daily and chopped pork liver every other day. The test water (22
ฑ2ฐC) was completely replaced every other day. Fluorescent lighting provided a photoperiod of
12 hours and a light intensity ranging from 600 to 1,000 lux at the water surface. During the
exposure, the animals were observed for mortality and growth daily and dead tadpoles were
removed. At the end of exposure, the survival rate of the frogs in each tank was recorded. After
anaesthetization, the frogs were weighted and dissected. The liver tissue of each frog was
weighed and hepatosomatic index (HSI) was calculated. The sex or intersex of each frog was
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determined by examining the gross gonadal morphology with a stereo microscope. The survival,
weight and sex ratio/intersex chronic value were all > 0.7160 mg/L PFOS (or 1 mg/L PFOS as
the nominal concentration). The study-author reported value was used quantitatively to derive
the freshwater chronic water column criterion.
Fort et al. (2019) evaluated the chronic effects of perfluorooctane sulfonic acid (PFOS,
>98% purity, CAS # 1763-23-1, lot # BCBH2834V from Sigma-Aldrich in St. Louis, MO) on
clawed frogs {Xenopus tropicalis, formerly Silurana tropicalis) in a 150-day post-metamorphosis
flow-through, measured study. Stock solutions were prepared by dissolving PFOS into filtered,
dechlorinated tap water in 18 L glass carboys, which were then pumped into the master mixing
cell of the continuous flow diluter. Adult frogs were obtained from Xenopus 1 and fed salmon
starter pellets daily for 30 days during acclimation prior to breeding. Temperature during
acclimation was maintained at 26 ฑ0.5ฐC. Researchers followed the breeding guidance of Fort et
al. (2002), and added human chorionic gonadotropin the day before breeding began. Three pairs
of frogs were isolated and allowed to breed, but only a single clutch with a >70% spawn rate was
utilized for the experiment. Normal appearing dejellied embryos (Nieuwkoop and Faber Stage
10) were randomly selected, and 20 were placed in each of four aquaria, each 4-L in size, for a
total of 80 embryos per concentration. The frogs were subjected to a 12:12-hour light:dark
photoperiod with a light intensity of 600 ฑ 50 lux, and the pH was maintained naturally at 7.5
ฑ0.3. The diluter system achieved a complete volume change every 6.5 hours, and diluter
performance, flow rates, temperature, D.O. and light intensity were measured daily. Test
organisms were exposed to mean measured concentrations of <0.03 (control), 0.05, 0.13, 0.31,
0.59 and 1.05 mg/L PFOS until metamorphosis, and liquid chromatography mass-spectrometry
was used to verify differences in PFOS concentrations. At metamorphosis (NF Stage 66), weight
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and snout-vent lengths were measured. Frogs were kept an additional 150 days past
metamorphosis without PFOS to determine weights, lengths, and sex differences amongst the
organisms. Mortality data showed aNOEC value >1.05 mg/L while the pre-metamorphosis
portion of the study showed a NOEC of 0.59 mg/L and a LOEC of 1.05 mg/L for both snout-vent
length and weight (MATC = 0.7871 mg/L). The LOEC of 1.05 mg/L was associated with 5%
(snout-vent length) and 14% (weight) decrease compared to controls, respectively. A significant
increase in the median metamorphosis time was observed in the 1.05 mg/L PFOS treatment
relative to the control. The post-metamorphosis LOEC was reported as 1.05 mg/L. No LCso
value was reported in that only 5.2 percent mortality was observed in the highest exposure
concentration (1.05 mg/L) at test termination. Independently-calculated ECios could not be
calculated as the EPA was unable to fit a model with significant parameters. Instead, the author-
reported MATC of 0.7871 mg/L PFOS based on growth (measured as mean body weight at
metamorphosis) was used quantitatively to derive the freshwater chronic water column criterion.
C.2.14 Fourteenth Most Sensitive Freshwater Genus for Chronic Toxicity: Lithobates (frog)
The chronic flow-through measured toxicity of PFOS (potassium salt, 98% purity) to the
northern leopard frog, Lithobatespipiens (formerly, Ranapipiens), was investigated by Ankley
et al. (2004) Two PFOS stock solutions (708 and 21.7 mg/L) were prepared by dissolving solid
PFOS with one liter of Lake Superior water in a glass carboy for 24 hours and then brought to a
volume of 18 L for the final stock solutions. Contents were stirred at room temperature (~20ฐC)
for 24 hours prior to being used. Solutions were pumped from the carboys to the glass aquaria
through Teflonฎ tubing using fluid metering pumps equipped with stainless-steel rotary
dispensers. Target concentrations were achieved by diluting the high and low stock solutions
with an appropriate volume of the Lake Superior (control) water. The PFOS stock solutions were
renewed every seven days. Fertilized eggs were collected from Grand Lake (St. Louis County,
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MN), near a sandy shoreline with no development. Tests were initiated with stage 8/9 embryos;
animals were gently separated with a plastic spatula from the egg mass, inspected under a
microscope for viability (evidence of cell division), and randomly allocated to treatment groups.
Exposures were conducted in glass aquaria in 10 L of water, which was continually renewed at a
flow rate of about 50 mL/minute (72 L/day). Duplicate tanks at target (nominal) PFOS
concentrations of 0.03, 0.1, 0.3, 1, 3, and 10 mg/L and four replicate control aquaria were used.
Embryos (n=120) were placed in each aquarium; in addition, two of the control tanks and the
duplicate tanks at 0.1 and 1 mg PFOS/L received an extra 80 organisms (total of 200) at test
initiation to provide animals for determination of PFOS concentrations during the early part of
the assay. Although biomass varied between the tanks with 120 versus 200 tadpoles, in both
situations total loading to the system was more than two orders of magnitude lower than
guidance recommended for a test at this flow rate. Water temperature was maintained at 20 ฑ
0.5ฐC, and the photoperiod (provided by fluorescent lights) was a constant 16:8-hour light:dark
cycle. On hatching (at approximately six days), animals were fed a mixture of live brine shrimp,
ground trout chow, and Tetrafin ad libitum two times daily. Dead organisms were removed daily
and inspected for gross abnormalities. On test day 6, 10 newly-hatched (<24 hours) animals were
randomly removed from each tank, preserved in 10% neutral buffered formalin, and
subsequently examined for developmental anomalies. Groups of animals were randomly selected
from each treatment (excluding the 10 mg/L group, which had been terminated because of high
mortality) on test days 35 (10 tadpoles/tank) and 54 (three tadpoles/tank). The animals were
weighed, and developmental stage was recorded, before being processed for PFOS tissue
analysis. The first tadpoles to undergo complete metamorphosis (defined as emergence of the
forelimbs) were observed on test day 60. Metamorphs were removed from the test tanks,
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sacrificed with an overdose of MS-222, weighed, measured (total and snout-vent length), and
assessed for gross abnormalities. Metamorphosis of the tadpoles continued over the next 51 days,
until the test was terminated, when remaining tadpoles were counted, staged, and weighed. A
subset of tadpoles from the control and 3 mg PFOS/L treatments were processed for histological
analysis of the thyroid gland when they were sampled at forelimb emergence. The most sensitive
apical chronic value was the 112-day growth MATC of 1.732 mg/L PFOS, followed by the 5-
week LCso of 6.21 mg/L PFOS. These data are considered quantitative even though the control
mortality was >20% at test termination (Note: Excessive mortality of amphibian larvae should be
expected within the full duration of this experiment given the life history strategy employed by
amphibians. Therefore, the observed control survival for this study was considered within the
acceptable range for this species and the toxicity data should be limited to the first 10 weeks of
the experiment). The author-reported value (112-d growth MATC of 1.732 mg/L) was used
quantitatively to derive the chronic water column criterion.
Hoover et al. (2017) also evaluated the chronic toxicity of PFOS (>98% purity) to
Lithobatespipiens. Test solutions were renewed every four days and exposure concentrations
were measured prior to and after each water change. Stock solutions consisted of 1 g of chemical
dissolved in 2 L of Milli-Q water, then vacuum-filtered before storage in polycarbonate bottles.
Eight northern leopard frog egg masses were collected during early spring from a temporary
pond at the Purdue Wildlife Area in West Lafayette, IN, and randomly assigned to outdoor ~100
L wading pools. After hatching, larvae were checked daily for mortality and fed Purina Rabbit
Chow ad libitum. Treatments consisting of a control and PFOS at three concentrations
(nominally 0.010, 0.100, and 1.0 mg/L) were placed in two replicates on adjacent shelves within
an environmental chamber. Experimental units consisted of 15 L plastic aquaria filled with 7.5 L
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of filtered, UV-irradiated well water. Tadpoles (n=35 per aquarium) were randomly assigned to
the experimental units. Prior to addition to aquaria, a subset of animals was examined to confirm
development at Gosner stage 26, when hind limb buds start to develop. Tadpoles with visible
irregularities in morphology, coloration, or behavior were excluded. Animals were maintained at
20 ฑ2ฐC with a 12:12-hour light:dark photoperiod for 10 days to acclimate to indoor conditions
and were fed a TetraMinฎ slurry ad libitum. Water changes (100%) were conducted every four
days. Tadpoles were exposed for 40 days and were monitored daily for mortality and
abnormalities. A water sample (~5 mL) was taken immediately prior to and after each water
change to monitor concentration of test chemicals. Every 10 days, six animals were randomly
collected from each aquarium. The animals were euthanized, measured (total length at 10 days,
snout-vent length otherwise), and staged (Gosner) prior to storage at -20ฐC for chemical
analyses. After 40 days, the depuration phase was initiated by removing animals, cleaning each
aquarium with a methanol-soaked sponge, and rinsing to remove adsorbed compound. Aquaria
were refilled with clean water; animals were returned to the same aquarium and monitored as
described above. Water changes were carried out every four days with fresh water, and a water
sample was taken prior to each water change. Two tadpoles were sampled every 10 days for an
additional 30 days. The 40-day chronic value was 0.0316 mg/L PFOS based on Gosner stage
reached at test termination. This study was deemed quantitative, even though PFOS was detected
in the control organisms. While the concentrations were much lower than any of the PFOS
treatment groups (3 orders of magnitude lower), it indicated that some potential contamination
may have occurred in the controls. The author-reported value (a developmental-based MATC of
0.0316 mg/L) was used quantitatively to derive the freshwater chronic water column criterion.
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C.2.15 Fifteenth Most Sensitive Freshwater Genus for Chronic Toxicity: Hyalella (amphipod)
Krupa et al. (2022) conducted a 42-day chronic toxicity test with the amphipod,
Hyalella azleca, and PFOS-K (perfluorooctanesulfonate potassium salt, > 98% purity, CAS No.
2795-39-3, purchased from Sigma-Aldrich). H. azteca were obtained from an in-house culture
maintained according to U.S. EPA (2000b). Following published methods for conducting water-
only testing with H. azteca (Bartlett et al. 2021; Ivey et al. 2016), the methods for this chronic
42-d static-renewal test were modified from standard guidance for sediment testing (U.S. EPA
2000b) to a water-only exposure to avoid sorption of PFOS to sediment. Juvenile amphipods
approximately 7- to 8-d old were obtained from mixed-age animals passing through a 425 |im
sieve and retained on a 355 |im sieve. Animals were then acclimated to test conditions for 2-days
before the start of the exposures. Tests were conducted in 300 mL polycarbonate beakers
containing 200 mL of test water, a thin layer of clean silica sand (250 - 500 mm diameter, 5 mL
per beaker) and ten amphipods under a 16:8-hour light:dark cycle in an environmental chamber
maintained at 23 ฑ 1ฐC for 42-days. Aeration was provided at a trickle flow rate via glass
pipettes. Mean measured exposure concentrations were 0.0093 (control), 4.8, 9.3, 21, and 45
mg/L PFOS, with eight replicates per concentration. Three liters of a 300 mg/L PFOS stock
solution was made by dissolving the PFOS salt into dechlorinated tap and mixing the solution on
a stir plate for > 20-hours. A total of 10 L of each test concentration was then prepared by mixing
the stock solution into carbon filtered dechlorinated tap water containing added sodium bromide
(0.052 mg/L). Amphipods were fed 1 mL of YCT food mixture (1.8 g/L mixture) per beaker
daily and a ramped (food increased over time) ration of finely ground TetraMinฎ (0.25 mg/day
during week 1, 0.5 mg/day at week 2, 1 mg/day at week 3, and 1.5 mg/day thereafter). For
beakers with < 50 % survival, the feeding ration was halved. The YCT and TetraMinฎ were
mixed together as a slurry. The test solution was renewed every 7 days. On Day 7, a partial water
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change removing about 150 mL of solution was performed to avoid disturbing the younger//.
azteca. Complete renewal of the test solution started on Day 14 and continued weekly for the
remainder of the test. For these water changes, the entire content of a beaker was transferred to a
12-inch diameter glass culture bowl. New sand and exposure solution were then placed in the
beaker. Surviving adult amphipods were enumerated and transferred back to the beaker. On Day
21, all beakers were replaced with new ones. On days 28, 35 and 42, after the adults were
transferred back to the beakers, the former contents of the beaker (all the discarded sand and
water) were preserved in 70% ethanol with rose bengal stain for later enumeration of offspring.
At test termination, sex and combined replicate dry biomass were determined for each replicate.
To determine biomass, animals were loaded onto a pre-weighed pan made of aluminum foil and
placed in a 60ฐC oven for a minimum of 24-hours. The pans were then placed in a desiccator for
at least 1-hour before being weighed. Eight pans with ten animals each were loaded with
organisms on Day 0 to establish a baseline for growth. Water quality parameters observed during
testing ranged from 7.31 - 8.57 mg/L D.O., 7.77 - 8.10 SU pH, 22.1 - 22.8ฐC and 62 - 72 mg/L as
CaC03 total hardness. Aqueous samples for PFOS concentration verification were collected
before test initiation (day 0) and termination (day 42). Analytical samples were also collected
during test solution renewals: out-water on days 7, 14, 21, 28 and 35; and in-water on day 21.
The author- reported 42-day reproductive ECio was 0.7 mg/L PFOS. The EPA's independently-
calculated models for ECio estimation were similar; however, the survival model was the most
robust. The EPA's independently-calculated ECios for the other 2 endpoints (the growth as dry
weight and reproduction as neonates per female) were different from the author-reported values
of 0.9 and 0.7 mg/L, respectively. The study authors were not able to calculate an ECio for
survival and so was reported to be <4.8 mg/L. However, the ECsos for all endpoints were similar
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between the author-reported and the EPA's independently-calculated values. Thus, the
differences in the ECio values were considered to be a result of the difference in models.
Therefore, the EPA's calculated value for the 42-day survival endpoint (ECio = 2.899 mg/L
PFOS-K; 1.132 - 4.667, 95% CI) was used to derive the chronic freshwater chronic water
column criterion.
C.2.16 Sixteenth Most Sensitive Freshwater Genus for Chronic Toxicity: Physella (snail)
Funkhouser (2014) conducted a 44-day renewal test of PFOS (potassium salt, >98%
purity) with Physella heterostrophapomilia as part of a Master's thesis at Texas Tech
University, Lubbock, TX. Egg masses from 100 P. pomilia adults were collected from Canyon
Lake 6, Lubbock Lakes System, Lubbock, TX, in May 2013 and used for testing. Dilution water
was moderately hard reconstituted laboratory water (3.0 g CaSC>4, 3.0 g MgSC>4, 0.2 g KC1, and
4.9 g NaFtCCb added to 50 L deionized water). Photoperiod was 12:12-hours light:dark at an
unreported light intensity. PFOS was dissolved in dilution water to prepare the test
concentrations. Exposure vessels were 250 mL polypropylene containers containing 200 mL of
test solution. The test employed two replicates composed of four egg masses each with an
average of 37.25 eggs/egg mass at start, then truncated to just four snails per replicate once snails
hatched. The test consisted of seven test concentrations plus a negative control. Nominal
concentrations were 0 (negative control), 10, 20, 40, 50, 70, 80, and 90 mg/L. Exposure
concentrations were reportedly measured, but concentrations were not provided. Analyses of test
solutions were performed using LC-MS/MS. Standards were used as part of the analytical
method, but details were not reported. The reporting limit was 0.010 mg/L. Experiments were
conducted in incubators set to 25ฐC, which did not vary more than 1ฐC during the course of the
studies. No other water quality parameters were reported as having been measured in test
solutions. Negative control survival was not reported specifically for the test but was reported to
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be 85-100% across all experiments. The 44-day life-cycle MATC was 14.14 mg/L (from the
author-reported NOEC and LOEC, 10 and 20 mg/L, respectively) for mean number of eggs per
egg mass. The independently-calculated ECio for the same endpoint was 8.527 (6.170 - 10.88)
mg/L. The independent statistical analysis was conducted using data that was estimated (using
Web plot digitizer) from the figures presented in the paper. This chronic value was acceptable for
quantitative use and was used to derive the freshwater chronic water column criterion.
C.2.17 Seventeenth Most Sensitive Freshwater Genus for Chronic Toxicity: Ceriodayhnia
(cladoceran)
Krupa et al. (2022) conducted a 6-day chronic toxicity test with the cladoceran,
Ceriodaphnia dubia, and PFOS-K (perfluorooctanesulfonate potassium salt, > 98% purity,
purchased from Sigma-Aldrich). In-house cultures of daphnids were reared according to U.S.
EPA (2002) and maintained in 100 mL glass beakers filled with 80 mL of moderately hard
reconstituted water (MHRW) prior to testing so that organisms used in the test were less than 24
hours old and were all released within an 8-hour period. The chronic toxicity test was conducted
as a three-brood (6-d) static-renewal test according to standard protocol (U.S. EPA 2002). The
measured exposure concentrations were < LOD-0.0003, 1.7, 3.5, 7.1, 13, 27, and 48 mg/L PFOS.
There were 10 replicates per concentration with one organism per each replicate. The highest test
concentration was prepared by dissolving the appropriate mass of PFOS into MHRW water.
Once the compound was fully dissolved, the highest concentration was then serially diluted to
the other, lower target concentrations. The tests were conducted in 20 mL glass scintillation vials
containing 15 mL of test water under a 16:8-hour light:dark cycle in an environmental chamber
maintained at 25 ฑ 1ฐC without aeration. C. dubia were fed daily with 0.45 mL of 1:1 I\
subcapitata and YCT (yeast-cerophyll-trout chow). Survival and reproduction were recorded
each day and neonates and any dead animals were removed. Complete renewal of the test
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solution and count of survivors and neonates was performed daily. Water quality parameters
were measured at every water renewal for both the in-water and out-water and ranged from 6.1 -
12.4 mg/L D.O., 6.91 - 8.02 SU pH and 24.0 - 25.9ฐC. Water samples collected at test initiation
(day 0), before (out-water) and after (in-water) test solution renewals, and at test termination
(day 6) were analyzed to verify PFOS concentrations. The author-reported 6-day reproductive
ECio was 6.9 mg/L PFOS-K. The independently-calculated 6-day ECio value based on
reproduction was 10.69 (5.839 - 15.54) mg/L and was considered acceptable for quantitative use
to the chronic freshwater chronic water column criterion.
Kadlec et al. (2024) tested the chronic toxicity of potassium perfluorooctanesulfonate
(PFOS-K) on Ceriodaphnia dubia for 7 days in a measured, renewal experiment. 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 was supplemented with Na2SC>4, NaCl, KC1, CaCh x
2FhO, 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
PFOS concentrations, with ten replicates of each concentration, and one organism per replicate.
Test 1 mean concentrations were 0.23 (control), 2.3, 4.4, 8.8, 18, and 38 mg/L. Test 2 mean
concentrations were 0.0049 (control), 0.049, 0.090, 0.25, 0.42, 0.92, 1.9, 3.9, 8.0, 17, and 35
mg/L. Test 3 mean concentrations were 0.27 (control), 2.7, 5.9, 11, 23, and 46 mg/L. C. dubia
neonates (<8-hour) were placed in 1 oz polystyrene cups filled with 15 mL of solution. Dissolved
oxygen and pH were measured twice in each exposure per treatment and twice in stock solutions.
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Tests 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 CaCC>3 hardness, 41 mg/L as CaCC>3 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. EC20s and EC50S
for survival and young per surviving female were calculated following methods described in
Mount et al. (2016), using custom software written with Intel Visual Fortran Compiler Xe and
Winteracter 13.0. The author reported EC20s for young per surviving female for tests 1, 2, and 3
were 10.0, 14.5 and 9.8, 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 ECios for tests 1, 2, and 3 were 8.371, 9.205, and 6.766
mg/L, respectively, which were determined to be acceptable for quantitative use.
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Appendix D Acceptable Estuarine/Marine Chronic PFOS Toxicity Studies
D.l Summary Table of Acceptable Quantitative Estuarine/Marine Chronic PFOS Toxicity Studies
Species (lil'csliiue)
Met hod '
Tcsl
Dui'iilion
( hcmic;il /
PuriU
pll
Temp.
(ฐC)
S;ilinil>
IPPII
('limine
Value
l.nripoinl
Author
Kcpnrlcri
Chronic
Value
(mป/l.)
i:p\
( iilcnliilod
Chronic
\;iluo
(inu/l.)
liinil
(h ionic
\ illllC
(mป/l.)h
Species
Mciin
(lirnnic
Y;iliic
(IllSi/l.)
Reference
Asian green mussel (60-65
mm),
Perna viridis
R,M
7 d
PFOS-K
98%
25
25
EC10
(growth condition
index)
0.03190
0.0033
0.0033
0.0033
Liu et al.
(2013)
Copepod (nauplii),
Tigriopus japonicus
R, U
20 d
PFOS
Unreported
-
25
32
MATC
(developmental
stage)
0.7071
-
0.7071
0.7071
Han et al.
(2015)
Amphipod
(juvenile, 14 d),
Austrochiltonia subtenuis
S,U
7 d
PFOS
Unreported
8.12-
8.3
20.3-
21.2
-
MATC
(mortality)
0.01118
-
0.01118
0.01118
Sinclair et al.
(2022)
Mysid (< 24 hr)
Americamysis bahia
F, M
35 d
PFOS-K
90.49%
8.2-
8.4
25
19-21
MATC
(reproduction,
growth)
0.3708
-
0.3708
0.3708
Drottar and
Krueger
(2000h)
Japanese medaka,
Oryzias latipes
S,U
30 d
PFOS-K
98%
7.87
24.77
34.68
NOEC
(growth condition
index)
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 SMCV calculation.
D-l
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D.2 Detailed PFOS Acute Toxicity Study Summaries and Corresponding
Concentration-Response Curves (when calculated for the most sensitive
genera)
The purpose of this section was to present detailed study summaries for tests that were
considered quantitatively acceptable for estuarine/marine chronic criterion derivation, with
summaries grouped and ordered by genus sensitivity. Data for chronic PFOS toxicity were
available for three saltwater invertebrate species, representing three genera and three families.
Chronic PFOS toxicity data was available for one fish species. The data available fulfilled only
four of the eight MDRs, therefore the EPA could not develop a chronic estuarine/marine
criterion following the 1985 Guideline methods.
D.2.1 Most Sensitive Estuarine/Marine Genus: Perna (mussel)
Liu et al. (2013) evaluated the chronic effects of perfluorooctanesulfonate, potassium salt
(PFOS-K, CAS# 2795-39-3, 98% purity, purchased from Sigma-Aldrich) on green mussels,
Perna viridis) via a 7-day measured, static-renewal study. The mussels were obtained from a
local farm in Singapore, and subsequently acclimated to laboratory conditions for seven days
before testing. Mussels were kept at a salinity of 25 ppt (artificial seawater) and a temperature of
25ฐC. Forty mussels (60-65 mm length) per 50-L polypropylene tank in duplicate were exposed
to measured PFOS concentrations of 0 (control), 0.00012, 0.0011, 0.0096, 0.106 and 0.968
mg/L. Mussels were fed a commercial marine micro-alga purchased from Reed Mariculture on
renewal days, which occurred every two days, two hours before the solution renewal. PFOS
concentrations were verified through water and muscle tissue samples via LC/MS. Weights and
lengths were determined on days 0 and 7. A NOEC of 0.0096 mg/L and a LOEC of 0.106 mg/L
was determined for the growth condition index resulting in an MATC of 0.03190 mg/L. No LCso
value was reported. The EPA's independently-calculated ECio is 0.0033 (0.00330 - 0.00332)
mg/L and was acceptable for quantitative use.
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D.2.2 Second Most Sensitive Estuarine/Marine Genus: Austrochiltonia (amphipod)
Sinclair et al. (2022) tested perfluorooctane sulfonic acid (PFOS) on amphipods
(Austrochiltonia subtenuis) in a 7-day unmeasured, static experiment. PFOS (purity not reported)
was purchased from Sigma-Aldrich (Melbourne, VIV, Australia). Test organisms were originally
field collected from Deep Creek, Victoria, Australia, but had been maintained in a laboratory
culture for over five years at RMIT University in Melbourne, Australia. Organisms were cultured
in standard artificial media (SAM) modified from Borgmann (1996) at 21ฑ1ฐC under a 16:8
light:dark cycle, and they were fed powdered Tetramin fish food and YTC every two days. Two
days prior to testing, 14-day old amphipods were selected and held in 2 L glass beakers until test
initiation. The 7-day experiment consisted of five controls, one solvent control (methanol 0.25
mL/L), and five nominal PFOS concentrations (0.04, 0.2, 1.0, 5.0, 25 |ig/L). Test vessels were
600 mL beakers with 400 mL of test material and a 2x2 cm gauze substrate. Each test vessel
included 20 amphipods, and all test media was dissolved in SAM. Seven-day survival was
assessed using a series of two-sample Student's t-tests (assuming equal variance) comparing
survival in each treatment to its corresponding control. After seven days, PFOS in tissues was
measured to calculate bioconcentration factors, and a suite of metabolites were also measured.
Based on the results of the t-tests, there was a statistically significant (p<0.05) decrease in
survival (21% mortality) at the highest concentration compared to its control. Setting the highest
concentration 0.025 mg/L as the LOEC and the highest concentration with no adverse effect
0.005 mg/L as the NOEC, the resulting MATC was 0.01118 mg/L, which was determined to be
acceptable for quantitative use.
D.2.3 Third Most Sensitive Estuarine/Marine Genus: Americamysis (mysid)
Drottar and Krueger (2000h) reported the results of a life-cycle, 35-day flow-through,
measured test of PFOS-K (potassium salt, 90.49% purity) with Americamysis bahia (formerly
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Mysdiopsis bahia). This good laboratory practice (GLP) test was conducted at the Wildlife
International, Ltd. toxicology facility in Easton, MD in June, 1999. The test followed U.S. EPA
OPPTS 850.1350, and ASTM Standard E 1191-90 test guidelines. Mysids used for the test were
neonates less than 24 hours old at test initiation. The dilution water was filtered natural seawater
collected at Indian River Inlet, DE diluted to a salinity of approximately 20 ppt with well water
[pH: 8.3 (8.2-8.4); TOC: >5.8 mg/L; temperature: 25ฑ2ฐC], Photoperiod was 16:8-hours,
light:dark with a 30-minute transition period. Light was provided at an intensity of 623-815 lux.
A primary stock solution was prepared in dilution water at 89.5 mg/L. It was mixed until all of
the test substance was dissolved prior to use. After mixing, the primary stock was proportionally
diluted with dilution water to prepare the five additional test concentrations. Exposure vessels
were glass beakers with nylon mesh screens on each side placed in 9 L glass aquaria with 5 L of
test solution. After mysids reached sexual maturity, they were placed in pairs in glass petri dishes
to observe reproduction. The test employed four replicates of fifteen mysids each in six
measured test concentrations plus a negative control. Nominal concentrations were 0 (negative
control), 0.086, 0.17, 0.34, 0.69, 1.4, and 2.7 mg/L. Mean measured concentrations were <
0.0458 (LOQ), 0.057, 0.12, 0.25, 0.55, 1.3, and 2.6 mg/L, respectively. Analyses of test solutions
were performed at Wildlife International Ltd. using HPLC/MS. Measured values ranged from 66
to 96% of nominal. Mortality from test initiation to pairing (day 20) in the 0.057, 0.12, 0.25,
0.55, 1.3, and 2.6 mg/L treatment groups was 8, 25, 18, 17, 32 and 100%, respectively, and mean
control mortality was 22%. From pairing until test termination (day 20 to day 35) survival was
greater than 90% in the control and all but the 1.3 mg/L treatment, which had 57% survival
during that period. The 35-day NOEC (reproduction and growth) was 0.25 mg/L and
corresponding LOEC was 0.55 mg/L. The EPA-calculated MATC was 0.3708 mg/L. The ECio
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could not be calculated at this time given the level of data that was presented in the paper. The
chronic value was considered acceptable for quantitative use despite the control survival of 78%
because it was only slightly below the 80% survival threshold, and because there were no other
deficiencies in the study design.
D.2.4 Fourth Most Sensitive Estuarine/Marine Genus: Tisriopus (copepod)
A 20-day renewal, unmeasured full life-cycle test with PFOS (analytical grade) was
conducted on the copepod, Tigriopus japonicus (non-North American species) by Han et al.
(2015). Copepods were cultured and maintained in 0.2 |im filtered artificial seawater adjusted to
32 psu salinity and 25ฐC under a 12-hour photoperiod. T. japonicus were fed with green algae,
Tetraselmis suecica. PFOS (100 mg/L in MeOH) was concentrated by evaporation and re-
dissolved in DMSO to obtain a maximum stock concentration (1,000 mg/L). The PFOS stock
was diluted with artificial seawater to obtain four nominal test concentrations (0, 0.25, 0.5 and 1
mg/L PFOS). The final concentration of DMSO in seawater was 0.001% (v/v) or less for each
treatment. Ten newly-hatched nauplii (<12 hour post hatch) were allocated to each well of a 12-
well tissue culture plate with 4 mL of test solution. There were three replicates per each
treatment. Organisms were fed algae during testing and 50% of test media was replaced daily.
Over the next 20 days, the development of the copepod's growth from nauplii to copepodite and
from nauplii to adults was determined daily based on morphological characteristics. Results were
presented as the number of days needed to reach the normal development stages. The highest test
concentration (1 mg/L PFOS) significantly increased the amount of time it took the copepods to
reach the development stage. Additionally, the authors assessed the reproduction of the copepods
by counting the nauplii produced by eight ovigerous females for 10 days in each well exposed to
PFOS. However, it was unclear if this was a subsampling of the organisms used in the 20-day
developmental test or if an independent assay with adult females. Results are presented
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graphically as daily nauplii production/individual. There was a statistically significant decrease
in production (daily nauplii production/individual) in the 0.25, 0.5 and 1.0 mg/L PFOS
concentrations compared to the control. It was decreased by approximately 50% in the highest
concentration (1 mg/L). While this endpoint was more sensitive than the growth endpoint, the
publication is unclear about the method used for the reproduction test endpoint and whether it
was an independently-conducted 10-day test or a subsample of reproducing adults were observed
from the 20-day test. The EPA sought but did not receive responses to clarifying questions posed
to the authors. Additionally, the authors were asked if control survival for the test was above
80% and if the authors could provide the data. Based on the information presented in the paper
without additional information and data provided by the authors to clarify adherence to the EPA
data quality objectives and allow independent calculation and verification of point estimates, the
developmental stage is considered for quantitative use and the reproductive endpoint for
qualitative use. The 20-day MATC (based on time to reach development stage) was 0.7071 mg/L
and was acceptable for quantitative use.
D.2.5 Fifth Most Sensitive Estuarine/Marine Genus: Oryzias (fish)
Oh et al. (2013) evaluated the 30-day chronic toxicity of PFOS to the Japanese medaka,
Oryzias latipes, via an unmeasured static exposure. PFOS (98% pure, CAS No. 2795-39-3) was
dissolved in filtered seawater with the minimal concentration (< 0.001%) of dimethyl sulfoxide
(DMSO) used as a vehicle to prevent cellular damage to the fish. Chemical measurements were
not made, and nominal concentrations were used throughout the study. Prior to test initiation, the
fish were adapted to a seawater environment by first acclimation in a 50/50 freshwater/saltwater
mix, then sequentially increasing the saltwater component by replacing half of the water with the
same volume of seawater every day for 15 days. The fish were fed newly-hatched brine shrimp
(< 24 hours after hatching) and commercial flake food twice per day. No mortality was observed
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during acclimation to the seawater over one month. 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, D.O., and temperature. Fish from the third generation of O. latipes (n = 7/group) that had
adapted to seawater for over one month were used in the series of exposure experiments. Fish
were exposed for 30 days to one PFOS concentration (1 mg/L) plus a 0.22 |im filtered seawater
control and a DMSO carrier solvent control to examine biological effects (specifically, the
condition factor, K, a growth endpoint). Test conditions were maintained at an average
temperature of 24.77 ฐC, pH of 7.87 SU, D.O. of 5.90 mg/L, and salinity of 34.68 PSU with fish
fed daily.
The 30-day NOEC condition factor of 1.0 mg/L PFOS was selected as the primary
endpoint from this study. Authors also stated, "In our preliminary study, fish mortality was
altered 30 days after PFC (perfluorinated compounds) exposure, suggesting that repeated
exposure to PFCsfor 30 days at 1 ng/mL causes adverse effect on O. latipes." The statement
about results from the preliminary test is in direct conflict with the results of condition factor in
the final test. Few details are provided about the preliminary test. Results of the final test were
retained as quantitatively acceptable because they provide chronic estuarine/marine data that are
limited.
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Appendix E Acceptable Freshwater Plant PFOS Toxicity Studies
E.l Summary Table of Acceptable Quantitative Freshwater Plant PFOS Toxicity Studies
Species
Method1'
Test
Diii'iilion
( hemiciil
/ Piiriu
Pll
1 'em p.
<ฐC)
ll'lccl
Reported
ll'l'eel
( oiicciili'iilion
(mป/l.)
Reference
Cyanobacteria,
Microcystis aeruginosa
S,U
72 hr
PFOS-K
98%
7.4
23
MATC
(cell density)
0.3162
Muhammad (2023)
Cyanobacteria,
Microcystis aeruginosa
s,u
96 hr
PFOS-K
98%
-
25
EC30
(population growth rate)
77.39
Zhang et al. (2023b)
Diatom,
Navicula pelliculosa
S,M
96 hr
PFOS-K
86.9%
7.5-8.9
24
EC50
(area under growth curve)
252
Sutherland and
Krueger (2001)
Green alga,
Chlorella vulgaris
S,U
96 hr
PFOS-K
95%
-
23
IC50
(cell density)
81.6
Boudreau (2002);
Boudreau et al.
(2003a)
Green alga,
Chlorella vulgaris
S,U
96 hr
PFOS-K
98%
-
25
EC20
(population growth rate)
13.42
Zhang et al. (2023b)
Green alga,
Raphidocelis subcapitata
(formerly, Selenastrum
capricornutum and
Pseudokirchneriella subcapitata)
S,U
96 hr
PFOS-K
24-28%
-
23
EC50
(specific growth rate)
49.28b
3M Company (2000)
Green alga,
Raphidocelis subcapitata
s,u
4 d
PFOS-K
Unknown
-
23
EC50
(cell count)
77.19
3M Company (2000)
Green alga,
Raphidocelis subcapitata
s,u
7 d
PFOS-K
Unknown
-
23
EC50
(cell count)
76.68
3M Company (2000)
Green alga,
Raphidocelis subcapitata
s,u
10 d
PFOS-K
Unknown
-
23
EC50
(cell count)
83.92
3M Company (2000)
Green alga,
Raphidocelis subcapitata
s,u
14 d
PFOS-K
Unknown
-
23
EC50
(cell count)
76.78
3M Company (2000)
Green alga,
Raphidocelis subcapitata
S,M
96 hr
PFOS-K
90.49%
7.4-8.4
24
EC50
(cell density)
71
Drottar and Krueger
(2000b)
Green alga,
Raphidocelis subcapitata
s,u
96 hr
PFOS-K
95%
-
23
IC50
(cell density)
48.2
Boudreau (2002);
Boudreau et al.
(2003a)
Green alga,
Scenedesmus quadricauda
S,M
96 hr
PFOS-K
99%
7
22
EC50
(growth inhibition rate)
89.34
Yang et al. (2014)
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Species
Method'1
Test
Dui'iilioii
( hemiciil
/ I'uriu
pll
Temp.
<ฐC)
r.iTcci
Reported
i:iTcc(
( oiicciilmlion
(iiiii/l.)
Reference'
Duckweed,
Lemna gibba
S,M
7 d
PFOS-K
86.9%
7.5
25
IC10
(frond number)
18.06
Desjardins et al.
(2001b)
Duckweed,
Lemna minor
S.M
96 hr
PFOS
>95.0%
6.5
25
NOEC
(population growth rate)
9.859
Wu et al. (2023)
Water milfoil (5 cm apical shoots),
Myriophyllum sibiricum
S,M
14 d
PFOS-K
Unreported
-
-
EC10
(wet weight)
0.7
Hanson et al. (2005)
Water milfoil (5 cm apical shoots),
Myriophyllum sibiricum
S,M
28 d
PFOS-K
Unreported
-
-
EC10
(wet weight)
0.19
Hanson et al. (2005)
Water milfoil (5 cm apical shoots),
Myriophyllum sibiricum
S,M
42 d
PFOS-K
Unreported
-
-
EC10
(wet weight)
0.6
Hanson et al. (2005)
Water milfoil (5 cm apical shoots),
Myriophyllum spicatum
S,M
14 d
PFOS-K
Unreported
-
-
EC10
(plant length)
4.8
Hanson et al. (2005)
Water milfoil (5 cm apical shoots),
Myriophyllum spicatum
S,M
28 d
PFOS-K
Unreported
-
-
EC10
(dry weight)
3.3
Hanson et al. (2005)
Water milfoil (5 cm apical shoots),
Myriophyllum spicatum
S,M
42 d
PFOS-K
Unreported
-
-
EC10
(wet weight)
3.5
Hanson et al. (2005)
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, NR=not reported
b The independently-calculated EC50 value was 176.0 mg/L as the test substance, or 49.28 mg/L based on the percentage of PFOS-K (active ingredient 28%) in the test substance.
E-2
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E.2 Summary of Plant PFOS Toxicity Studies Considered in the Aquatic
Life Criterion Derivation
E.2.1 Cvanobacteria. Microcystis aeruginosa
Muhammad (2023) tested perfluorooctane sulfonic acid potassium salt (PFOS-K) on
Microcystis aeruginosa for seven days in a static, unmeasured experiment. PFOS potassium salt
(98% purity) was obtained from Sigma-Aldrich. Algae were obtained from Golder Laboratory, at
the State University of New York, USA. Algae were cultured in 500 mL Erlenmeyer flasks
containing 200mL of BG11 medium at the Phycology Laboratory of Ahmadu Bello University's
Department of Botany in Zaria. Cultures were maintained at 23ฐC, pH 7.4, under a 16:8 hour
light:dark photoperiod at a light intensity of 40 |imol/m2/s. The BG11 growth medium was
sterilized by autoclaving at 121 ฐC 24 hours before use. Cultures were changed and maintained in
a manner to insure they were experiencing exponential growth. M. aeruginosa 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 nominal concentrations of 0 (control),
0.001, 0.01, 0.1, 1 and lOmg/LPFOS, diluted from a PFOS stock solution prepared inBG-11
medium, with three replicates per treatment level. Cell density was measured on days 1, 3, and 7.
Cellular pigment content was measured on days 1 and 7, biochemical composition, antioxidant
enzyme composition, and microcystin analysis was measured on day 7. One-way ANOVA and
means comparisons tests were used to assess significant differences (p<0.05) between
treatments, and repeated measures ANOVA was used to assess significant differences between
sampling dates. All analyses were conducted using the R statistical analysis program v3.63. By
day three, cell density in the 1 mg/L treatment were significantly lower than in controls, and by
day seven, cell density in the 0.1 mg/L treatment were significantly lower than in controls.
Chlorophyll a also decreased with increasing PFOS concentrations, and biochemical and
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enzymatic endpoints also differed among PFOS treatment concentrations. The author-reported
72 hour NOEC and LOEC for cell density were 0.1 and 1 mg/L, and the resulting MATC of
0.3162 mg/L was determined to be acceptable for quantitative use.
Zhang et al. (2023b) tested perfluorooctane sulfonic acid potassium salt (PFOS-K) on
Microcystis aeruginosa for 12 days in a static, unmeasured experiment. PFOS potassium salt
(98% purity) was purchased from Sigma-Aldrich (St. Louis, USA). Algae were obtained from
the Freshwater Algae Culture Collection at the Institute of Hydrobiology in Wuhan, China.
Algae were cultured in Erlenmeyer flasks within BG-11 media in an illumination incubator at
25ฑ1ฐC on a 12:12 hour light-dark cycle at a light intensity of 4,000ฑ50 lux. M aeruginosa in
the exponential growth phase were added to 100 mL Erlenmeyer flasks containing test solution
at a density of approximately 1.0x10s cells/mL. The experimental design consisted of nominal
concentrations of 0 (solvent control), 0.01, 1, 20, 50, 100, 250 and 500 mg/L PFOS, diluted from
a PFOS stock solution prepared in BG-11 medium and containing 0.05% formaldehyde as a
solvent, with three replicates per treatment. Testing methods followed OECD guidelines with
modifications (OECD 2006). Algal cell density was measured after 4 and 12 days using a UV-vis
spectrophotometer. Algal cell densities were used to calculate growth inhibition at all
concentrations relative to controls, and effect concentrations (ECs) were calculated using
nonlinear regression modeling. Chlorophyll a and optical quantum yield were also measured
after 4 and 12 days. ANOVA followed by Tukey's multiple comparison test was used to identify
significant (p<0.05) differences between treatments and controls, using GraphPad Prism v8.0.
EC20S were calculated using a four-parameter nonlinear regression model in GraphPad Prism
v8.0. A 96-hour EC20 was not reported, so the 96-hour EC30 for growth inhibition of 77.39 mg/L
was determined to be acceptable for quantitative use.
E-4
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E.2.2 Diatom. Naviculapelliculosa
Sutherland and Krueger (2001) conducted a 96-hour static acute algal growth
inhibition test on PFOS (potassium salt, 86.9% purity) with the freshwater diatom, Navicula
pelliculosa. The good laboratory practice (GLP) test was conducted at the Wildlife International,
Ltd. in Easton, Maryland in February-March, 2000. The test followed U.S. EPA (1996a) and
ASTM (1990). The freshwater diatom was provided from in-house cultures that had been
actively growing in the culture medium for at least two weeks. The test media was prepared by
adding the stock nutrient solution to purified well water according to ASTM 1218 and adjusting
pH to 7.5. Seven measured concentrations (0, 62.3, 83.2, 111, 150, 206, 266, 335 mg/L PFOS)
were tested from one negative control and six nominal concentrations: 61.5, 81.3, 110, 147, 198,
264 and 347 mg/L based on PSOF-K purity. Solutions were stirred for approximately 24 hours
before testing. Exposures were conducted in 250 mL plastic Erlenmeyer flasks containing 100
mL solution and plugged with foam stoppers. Each flask contained lxlO4 cells/mL and each test
concentration had three replicates. Flasks were incubated in environmental chambers at 24ฑ2ฐC
under constant illumination (4,300 lux) and shaken continuously at -100 rpm. pH in the test
solutions ranged from 7.5-8.9 over the exposure period. Samples were collected daily to
determine cell density and to calculate area under the curve and growth rates. The cell density of
the control replicates increased by greater than two orders of magnitude during the test. The 96-
hour ECso, based on area under the growth curve, was 252 mg/L PFOS (NOEC<62.3 mg/L). The
plant value was acceptable for quantitative use.
E.2.3 Green alga. Chlorella vulgaris
Boudreau (2002) performed a 96-hour static acute algal growth inhibition test on PFOS
(potassium salt, 95% purity) with Chlorella vulgaris as part of a Master's thesis at the University
of Guelph, Ontario, Canada. The same information was subsequently published in the open
E-5
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literature as Boudreau et al. (2003a). The acute algal growth inhibition tests followed protocols
found in ASTM (1999b) and Geis et al. (2000). All treatment concentrations were based on the
PFOS anion (without K) and solutions were prepared in lab oratory-grade distilled water. C.
vulgaris (UTCC 266 strain) were obtained as slants from the University of Toronto Culture
Collection (UTCC; Toronto, Canada). Toxicity testing consisted of initial range-finder tests (0,
28, 56, 113, 225, and 450 mg/L) followed by at least two definitive tests (0, 12.5, 25, 50, 100,
200, and 400 mg/L). Tests were conducted in 20 mL solution in 60 x 15 mm polyethylene
disposable petri dishes. Each dish contained 1.5xl04 cells/mL and each test concentration had
four replicates. Tests were continuously illuminated with cool-white fluorescent light between
3,800 and 4,200 lux and incubated at 23ฑ1ฐC. Each dish was manually shaken twice a day
during testing. Toxicity test endpoints included cell density and chlorophyll a content. The most
sensitive endpoint, cell density, had a reported NOEC of 8.2 mg/L and an IC50 of 81.6 mg/L. The
IC50 value was considered quantitatively acceptable for use.
Zhang et al. (2023b) tested perfluorooctane sulfonic acid potassium salt (PFOS-K) on
Chlorella vulgaris for 12 days in a static, unmeasured experiment. PFOS potassium salt (98%
purity) was purchased from Sigma-Aldrich (St. Louis, USA). Algae were obtained from the
Freshwater Algae Culture Collection at the Institute of Hydrobiology in Wuhan, China. Algae
were cultured in Erlenmeyer flasks within BG-11 media in an illumination incubator at 25ฑ1ฐC
on a 12:12 hour light-dark cycle at a light intensity of 4,000ฑ50 lux. C. vulgaris in the
exponential growth phase were added to 100 mL Erlenmeyer flasks containing test solution at a
density of approximately 1.0x10s cells/mL. The experimental design consisted of nominal
concentrations of 0 (solvent control), 0.01, 1, 20, 50, 100, 250 and 500 mg/L PFOS, diluted from
a PFOS stock solution prepared in BG-11 medium and containing 0.05% formaldehyde as a
E-6
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solvent, with three replicates per treatment. Testing methods followed OECD guidelines with
modifications (OECD 2006). Algal cell density was measured after 4 and 12 days using a UV-vis
spectrophotometer. Algal cell densities were used to calculate growth inhibition at all
concentrations relative to controls, and ECsos were calculated using nonlinear regression
modeling. Chlorophyll a and optical quantum yield were also measured after 4 and 12 days.
ANOVA followed by Tukey's multiple comparison test was used to identify significant (p<0.05)
differences between treatments and controls, using GraphPad Prism v8.0. Effect concentrations
(ECs) were calculated using a four-parameter nonlinear regression model in GraphPad Prism
v8.0. The 96-hour EC20 for growth inhibition was 13.42 mg/L, and was determined to be
acceptable for quantitative use.
E.2.4 Green alga. Rayhidocelis subcayitata
3M Company (2000) provided the results of a 96-hour toxicity test completed in 1991
with the green alga, Raphidocelis subcapitata (formerly Selenastrum capricornutum), and PFOS-
K (perfluorooctancesulfonate potassium salt, CAS # 2795-39-3) in a formulated mixture with
diethylene glycol butyl ether and water (mixed product FM-3820, with 24-28% PFOS-K). Based
on this purity the author made calculations to adjust test concentrations using 28% active
ingredient, but the presence of diethylene glycol could also contribute to toxicity. The toxicity
test followed OECD test guidelines with five test concentrations and control in a static
unmeasured exposure. A stock culture of the alga was obtained from the Culture Collection of
Algae at the University of Texas at Austin. Alga were transferred to 250 mL flasks with an initial
density of lxlO4 cells/mL and 100 mL of test solution. There were three replicates for each of the
five nominal test concentrations (62.5, 125, 250, 500 and 1,000 mg/L) and control. Synthetic
nutrient medium was used as the dilution media for all test treatments. Alga were grown at 23ฐC
and continuously shaken. The author-reported EC50, based on average specific growth rate, was
E-7
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255 mg/L as the test substance, or 71 mg/L based on the percentage of PFOS-K (active
ingredient 28%) in the test substance. The independently-calculated ECso value was 176.0 mg/L
as the test substance, or 49.28 mg/L based on the percentage of PFOS-K (active ingredient 28%)
in the test substance. The plant value was acceptable for quantitative use.
3M Company (2000) provides the results of four separate toxicity tests completed in
1981 with the green alga, Raphidocelis subcapitata (formerly Selenastrum capricornutum), and
PFOS-K (perfluorooctancesulfonate potassium salt, CAS # 2795-39-3, unknown purity). The
toxicity tests followed a protocol modified from OECD (1979). There were four separate
exposure regimes: 1) four day exposure + 10 day recovery period; 2) seven day exposure + seven
day recovery period; 3) 10 day exposure + four day recovery period; and 4) 14 day continuous
exposure. A bacteria-free culture of the alga was obtained from the U.S. EPA (Corvallis, OR)
and stored in the dark until testing. Seven-day old stock cultures with an initial density of lxlO4
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 (26, 40, 61, 93, 145 and 225 mg/L) and control.
Nutrient medium was used as the dilution media for all test treatments and test solutions were not
renewed during the exposure. Alga were grown at 23 ฐC and continuously shaken at 100 rpm.
The author-reported ECso, based on cell counts, was 82, 99, 98, and 95 mg/L, for the 4-, 7-, 10-
and 14-day exposures, respectively. However, it should be noted that the authors do not specify
if the ECsos were determined after the exposure period or the post-observation period. The
independently-calculated EC50 values were 77.19, 76.68, 83.92, 76.78 mg/L and are acceptable
for quantitative use.
Drottar and Krueger (2000b) conducted a 96-hour static acute algal growth inhibition
test on PFOS (potassium salt, 90.49% purity) with the freshwater alga, Raphidocelis subcapitata
E-8
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(formerly Selenastrum capricornutum). The good laboratory practice (GLP) test was conducted
at the Wildlife International, Ltd. in Easton, Maryland in April, 1999. The test followed ASTM
(1990); OECD (1984); U.S. EPA (1996a) methodologies. The green alga was originally obtained
from the culture collection at University of Texas at Austin (or another supplier) and maintained
at Wildlife International Ltd. for a minimum of two weeks in culture medium. Algae used in tests
were in exponential growth phase. The test media was prepared by adding the stock nutrient
solution to purified well water according to ASTM 1218 and adjusting pH to 7.5. Seven
measured concentrations (< 0.115, 5.5, 11, 21, 44, 86, 179 mg/L PFOS) were tested from a
negative control and six nominal concentrations: 5.7, 11, 23, 46, 91, 183 mg/L based on PFOS-K
purity. Test concentrations were measured at test initiation, at 72 hours, and at test termination
by HPLC-MS with a mean recovery of 99.1%. Solutions were stirred for approximately 24 hours
before testing. Exposures were conducted in 250 mL polycarbonate flasks containing 100 mL
solution and plugged with foam stoppers. Each flask contained lxlO4 cells/mL and each test
concentration had three replicates. Flasks were incubated in environmental chambers at 24ฑ2ฐC
under constant illumination (4,300 lux) and shaken continuously at -100 rpm. The pH in test
solutions ranged from 7.4-8.4 over the exposure period. Samples were collected daily to
determine cell density and to calculate area under the curve and growth rates. The 96-hour ECso,
based on cell density and area under the growth curve, was 71 mg/L PFOS (NOEC=44 mg/L).
The plant value was acceptable for quantitative use.
Boudreau (2002) performed a 96-hour static acute algal growth inhibition test on PFOS
(potassium salt, 95% purity) with Raphidocelis subcapitata (formerly Pseudokirchneriella
subcapitata). The study was part of a Master's thesis at the University of Guelph, Ontario,
Canada and subsequently published in the open literature as Boudreau et al. (2003a). The acute
E-9
-------
algal growth inhibition tests followed protocols found in ASTM (1999b); Geis et al. (2000) and
Geis et al. (2000). All treatment concentrations were based on the PFOS anion (without K) and
solutions were prepared in laboratory-grade distilled water. R. subcapitata (UTCC 37 strain)
were obtained as slants from the University of Toronto Culture Collection (UTCC; Toronto,
Canada). Toxicity testing consisted of initial range-finder tests (0, 28, 56, 113, 225, and 450
mg/L) followed by at least two definitive tests (0, 12.5, 25, 50, 100, 200, and 400 mg/L). Tests
were conducted in 20 mL solutions in 60 x 15 mm polyethylene disposable petri dishes. Each
dish contained 1.5xl04 cells/mL and each test concentration had four replicates. Tests were
continuously illuminated with cool-white fluorescent light between 3,800 and 4,200 lux and
incubated at 23ฑ1ฐC. Each dish was manually shaken twice a day during testing. Toxicity test
endpoints included cell density and chlorophyll a content. The reported NOEC and IC50 based on
most sensitive endpoint, cell density, were 5.3 mg/L and 48.2 mg/L. The IC50 from the study was
acceptable for quantitative use.
E.2.5 Green alga. Scenedesmus guadricauda
Yang et al. (2014) conducted a 96-hour static, measured test on the growth effects of
PFOS (potassium salt, CAS # 2795-39-3, 99% purity) with the green alga, Scenedesmus
quadricauda. Algae were obtained from in-house cultures from the Chinese Research Academy
of Environmental Sciences. The algae used for testing were inoculated at a cell density equal to
2.0xl04 cells/mL in 50 mL beakers. PFOS was dissolved in deionized water and DMSO (amount
not provided) and then diluted with M4 medium. Alga were exposed to 0 (solvent control),
50.00, 65.00, 84.50, 109.85, 142.81 and 185.65 mg/L PFOS. Each treatment was replicated three
times. While the text implied the exposures were static, the supplemental information provided
the measured test concentrations in the highest and lowest test treatments both before and after
renewal. Measured concentrations ranged from 42.56 mg/L (before renewal) to 49.78 mg/L
E-10
-------
(after renewal) in the lowest treatment, and from 165.61 (before renewal) to 183.90 mg/L (after
renewal) in the highest treatment. The experiments were conducted at 22ฑ2ฐC with a 12:12-hour
light:dark cycle. The initial pH of the test solution was 7.0ฑ0.5, total hardness was 190ฑ0.1 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 ECso (based on growth inhibition) was 89.34
mg/L and was acceptable for quantitative use.
E.2.6 Duckweed. Lemna sp.
Desjardins et al. (2001b) performed a static, measured 7-day growth inhibition study on
the duckweed Lemna gibba with PFOS-K (perfluorooctanesulfonate potassium salt, 86.9% purity
from 3M Corporation). The test protocol from U.S. EPA, OPPTS Number 850.4400 was
followed. Duckweed was cultured and tested at Wildlife International Ltd. in 20X AAP medium
and were actively growing for at least two weeks prior to testing. The pH of the medium was
adjusted to pH 7.5 with HC1 and filtered to sterilize before use. Test chambers were covered 250
mL plastic beakers with 100 mL of culture medium or test concentration and held at 25ฐC under
continuous warm-white lighting with a target intensity of 5,000 lux. Fronds of duckweed were
exposed to six test concentrations and a control with three replicates for each treatment. PFOS
concentrations in the test medium were measured on day 0, 3, 5 and 7 with mean reported
concentrations of <4.39 (LOQ), 7.74, 15.1, 31.9, 62.5, 147 and 230 mg/L PFOS active
ingredient. Growth was defined as an increase in the total number of fronds in each replicate and
measured by direct count on day 3, 5 and 7. Frond numbers on day seven in the 147 and 230
mg/L test treatments were inhibited by 65 and 81% as compared to the control. The reported 5-
day ICio based on frond number was 30.7 mg/L PFOS. The independently-calculated 5-day ICio
value was 18.06 mg/L and was acceptable for quantitative use.
E-ll
-------
Wu et al. (2023) tested perfluorooctane sulfonate (PFOS) on duckweed, Lemna minor,
for 96 hours in a static, measured experiment. PFOS (>95.0% 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
for seven days prior to testing. Duckweed experiments were conducted in 6-well polypropylene
plates to avoid PFOS sorption to container walls. Each replicate well contained lOmL 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 PFOS. All test chambers included
DMSO solvent, and each treatment had three duplicates. PFOS concentrations measured on day
0 were 1.00ฑ0.02, 90.0ฑ0.73, and 9,859ฑ2.83 |ig/L. PFOS 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 PFOS 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 9.859 mg/L for population
growth rate was determined to be acceptable for quantitative use.
E-12
-------
E.2.7 Watennilfoil. Myriophyllum sp.
Hanson et al. (2005) conducted a 42-day toxicity study of PFOS (potassium salt, purity
not provided) with two species of submergent watermilfoils, Myriophyllum spicatum andM
sibiricum. The study was conducted in 12,000 L outdoor microcosms at the University of Guelph
Microcosm Facility located in Ontario, Canada. 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, 5 cm in length, from in-house cultures using the
same sediment were transferred to each microcosm, with three separate microcosms used for
each treatment (0, 0.3, 10 and 30 mg/L). Endpoints of toxicity that were monitored on days 1, 14,
28 and 42 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. PFOS treatments were
dissolved in the same water (well water) used to supply the microcosms. Measured
concentrations in the microcosms were reported in a companion publication (Boudreau et al.
2003b). Results from the companion paper 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. Water quality (i.e., pH, temperature, D.O., hardness,
and alkalinity) and light levels were measured regularly, but were not reported. M sibiricum was
more sensitive to PFOS thanM spicatum. The 42-day ECio (based on wet weight) was 0.6 mg/L
forM. sibiricum and 3.5 mg/L forM spicatum. The plant values were acceptable for quantitative
use.
E-13
-------
Appendix F Acceptable Estuarine/Marine Plant PFOS Toxicity Studies
F.l Summary Table of Acceptable Quantitative Estuarine/Marine Plant PFOS Toxicity Studies
Species
Method'
Tesl
Dui'iilioii
Chomiciil
/ Piiriu
pll
Temp.
(ฐC)
S;ilinii\
(|)|)l)
i.nvci
Reported
r.nwi
( oiiccnlriilion
(ni}ป/l.)
KoI'oivikt
Green alga,
Chlorella sp.
S,U
96 hr
PFOS
>98%
23
o
o
EC50
(population abundance)
77.62
Mao (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
-------
F.2 Summary of Plant PFOS Toxicity Studies Considered in the Aquatic
Life Criterion Derivation
F.2.1 Green alga. Chlorella sp.
Mao (2023) tested perfluorooctane sulfonate (PFOS) on an estuarine/marine Chlorella
sp. for seven days in a static, unmeasured experiment. PFOS (>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, and 160 mg/L PFOS. 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 10 mg/L, but growth
significantly declined at 40 mg/L and higher concentrations. Increasing PFOS concentrations
also inhibited chlorophyll a, and increased oxidative stress. The 96-hour EC so for algal growth
inhibition was 77.62 mg/L, and was determined to be acceptable for quantitative use.
F-2
-------
Appendix G Other Freshwater PFOS Toxicity Studies
G.l Summary Table of Acceptable Qualitative Freshwater PFOS Toxicity Studies
Spocics (life's!;iiicI
Method"
Tesl
Dui'iilioii
( hcmiciil /
Piiriu
pll
Temp.
(ฐC)
r.riw-i
Chronic
l.imils
iNOI'.C-
i.or.ci
Reported
I'.ITccl
(one.
Deficiencies
Reference
Unicellular protist,
Paramecium caudatum
S,U
1 hr
Heptadecafluor
ooctane
sulfonic acid
potassium salt
>98%
7.2
20-24
LC50
-
12.86e
Duration too short for
an acute test, single-
cell organism
Matsubara et al.
(2006)
Protozoa,
Tetrahymena pyriformis
s,u
2 hr
PFOS
Unreported
7.2
25
EC50
(population abundance)
-
51.51ฎ
Duration too short for
an acute test, single-
cell organism
Lim (2022)
Protozoa,
Tetrahymena pyriformis
s,u
96 hr
PFOS
Unreported
7.2
25
EC50
(population abundance)
-
13.2
Single-cell organism
Lim (2022)
Cyanobacteria,
Anabaena sp.
S,M
24 hr
PFOS
98%
-
-
EC50
(bioluminescence)
-
16.29
Duration too short for
a plant test, non-apical
endpoint
Rodea-Palomares
et al. (2012)
Cyanobacteria,
Anabaena sp.
S,U
24 hr
PFOS-K
98%
7.8
28
EC50
(bioluminescence)
-
83.51
Duration too short for
a plant test, non-apical
endpoint
Rodea-Palomares
et al. (2015)
Green alga
(7.0 x 105 cells/ml),
Chlorella vulgaris
S,M
96 hr
PFOS-K
98%
-
-
LOEC
(chlorophyll a)
-
40
Missing exposure
details
Xu et al. (2017)
Green alga,
Raphidocelis subcapitata
S,M
72 hr
PFOS-K
98%
-
21-24
EC50
(growth)
-
35
Duration too short for
a plant test, missing
some exposure details
Rosal et al. (2010)
Green alga,
Raphidocelis subcapitata
S,U
72 hr
PFOS-K
98%
-
22
EC50
(growth inhibition)
-
35
Duration too short for
a plant test
Boltes et al.
(2012)
Green alga,
Scenedesmus obliquus
S,U
72 hr
PFOS
Unreported
7.5
22
IC50
(growth rate reduction)
-
77.8e
Duration too short for
a plant test
Liu et al. (2008)
Green alga,
Scenedesmus obliquus
S,U
72 hr
PFOS
>98%
7.5
22
NOEC
(growth)
-
40
Duration too short for
a plant test
Liu et al. (2009)
Green alga,
Scenedesmus obliquus
S,U
7 d
PFOS-K
>98%
7.1
25
EC50
(biomass)
-
136.69
Missing exposure
details
Xue et al. (2022)
G-l
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcinic;il /
I'llril\
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Duckweed,
Lemna gibba
S,U
7 d
PFOS-K
95%
25
lC5u
(wet weight)
-
31.1
Culture water not
characterized, missing
some exposure details
Boudreau el al.
(2003a)
Blue green algae,
Scynechocystis sp.
S,M
2d
PFOS-K
>98%
7.5
30
NOEC
(abundance)
1->1
1
Only one exposure
concentration
Marchetto et al.
(2021)
Blue green algae,
Scynechocystis sp.
F, M
12-15 d
PFOS-K
>98%
7.5
30
NOEC
(biomass)
1->1
1
Only one exposure
concentration
Marchetto et al.
(2021)
Aquatic microcosm
(mixed invertebrate and
aquatic plant community)
S,M
35-42 d
PFOS-K
86%
8.3-
8.6
15.9-
20.5
MATC
(zooplankton community
abundance)
3.0-10
5.478
Mixed species
exposure, static
chronic exposure
Boudreau (2002);
Boudreau et al.
(2003b)
Aquatic microcosm
(mixed invertebrate and
aquatic plant community)
S,M
35 d
PFOS-K
Unreported
8.3
18
MATC
(zooplankton abundance;
Cyclops diaptomus
abundance)
1.0-10
3.162
Mixed species
exposure
Sanderson et al.
(2002)
Tubificid worm
(0.03g, 0.8cm),
Limnodrilus hoffmeisteri
S,M
96 hr
PFOS-K
99%
7
22
LC50
-
120.97
Atypical source of
organisms
Yang et al. (2014)
Tubificid worm,
Limnodrilus hoffmeisteri
S,U
24 hr
PFOS
>98%
5.0
23
LC50
-
45.26
Duration too short for
an acute test, missing
some exposure details
Liu et al. (2016)
Tubificid worm,
Limnodrilus hoffmeisteri
S,U
24 hr
PFOS
>98%
6.0
23
LC50
-
46.23
Duration too short for
an acute test, missing
some exposure details
Liu et al. (2016)
Tubificid worm,
Limnodrilus hoffmeisteri
S,U
24 hr
PFOS
>98%
7.0
23
LC50
-
60.70
Duration too short for
an acute test, missing
some exposure details
Liu et al. (2016)
Tubificid worm,
Limnodrilus hoffmeisteri
S,U
24 hr
PFOS
>98%
8.0
23
LC50
-
64.48
Duration too short for
an acute test, missing
some exposure details
Liu et al. (2016)
Tubificid worm,
Limnodrilus hoffmeisteri
S,U
24 hr
PFOS
>98%
9.0
23
LC50
-
65.74
Duration too short for
an acute test, missing
some exposure details
Liu et al. (2016)
Tubificid worm (3-4 cm),
Limnodrilus hoffmeisteri
R,U
48 hr
PFOS-K
98%
6.2
22
LC50
-
23.81
Duration too short for
an acute test, missing
some exposure details
Qu et al. (2016)
Tubificid worm (3-4 cm),
Limnodrilus hoffmeisteri
R,U
48 hr
PFOS-K
98%
7.0
22
LC50
-
35.89
Duration too short for
an acute test, missing
some exposure details
Qu et al. (2016)
Tubificid worm (3-4 cm),
Limnodrilus hoffmeisteri
R, U
48 hr
PFOS-K
98%
8.0
22
LC50
-
39.80
Duration too short for
an acute test, missing
some exposure details
Qu et al. (2016)
G-2
-------
Spocics (lilestjiiicM
Mi-llmd"
Tcsl
Dui'iilioii
( hcniic;il /
I'llril\
nil
Icm|).
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.fl'ccl
(one.
(mป/l.)
Deficiencies
Reference
Planarian (10-12 mm),
Dugesia japonica
R,U
10 d
PFOS-K
>99%
-
20
LOEC
(regeneration: decreased
appearance of auricles)
<0.5-0.5
0.5
Duration too long for
an acute test and too
short for a chronic test
Yuanetal. (2014)
Planarian,
Dugesia japonica
R,U
10 d
PFOS
>99%
-
20
LOEC
(enzymatic, gene
expression and
biochemistry changes)
-
5
Duration too long for
an acute exposure and
too short for a chronic
exposure, atypical
endpoints
Zhang et al.
(2023a)
Planarian,
Dugesia japonica
R, U
7 d
PFOS
Unreported
-
20
MATC
(gene expression)
0.5-1
0.7071
Duration too long for
an acute exposure and
too short for a chronic
exposure, atypical
endpoints
Sun et al. (2023a)
Chinese pond mussel
(1 year),
Sinanodonta woodiana
(formerly, Anodonta
woodiana)
s,u
48 hr
PFOS
Unreported
7
24
LC50
-
28.39
Duration too short for
an acute test
Xia et al. (2018)
Freshwater mussel
(6 cm),
Unio ravoisieri
R, U
96 hr
PFOS-K
>98%
8
18
LC50
-
65.9
Test species fed from
the natural freshwater
used
Amraoui et al.
(2018)
Asian clam (adult),
Corbicula fluminea
R, M
28 d
PFOS-K
98%
6.85-
7.35
23
LOEC
(biochemical, enzyme
and genetic markers)
-
0.0005082
Atypical endpoint,
only one exposure
concentration
Bi et al. (2022)
Mud snail (4.0 g, 2.0 cm)
Cipangopaludina
cathayensis
S, M
96 hr
PFOS-K
99%
7
22
LC50
-
247.14
Source of organisms
may be problematic
Yang et al. (2014)
Snail (adult),
Lymnaea stagnalis
S, M
96 hr
PFOS-K
95%
-
20
LC50
-
196
Test species fed
Olson (2017)
Snail
(0-3 week, juvenile),
Lymnaea stagnalis
S, M
96 hr
PFOS-K
95%
-
20
LC50
-
150
Test species fed
Olson (2017)
G-3
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
Ptiriu
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Snail
(0-3 week, juvenile),
Lymnaea stagnalis
R, M
21 d
PFOS-K
95%
20
NOEC
(survival, feeding rate,
mass change, length
change, carbohydrate
concentration)
50->50
>50
Duration too short for
a chronic test
Olson (2017)
Snail
(3-6 week, juvenile),
Lymnaea stagnalis
R, M
21 d
PFOS-K
95%
-
20
MATC
(mass change, length
change)
25-50
35.35
Duration too short for
a chronic test
Olson (2017)
Snail
(6-9 week, juvenile),
Lymnaea stagnalis
R, M
21 d
PFOS-K
95%
-
20
NOEC
(survival, mass change,
length change,
carbohydrate and protein
concentration)
50->50
>50
Duration too short for
a chronic test
Olson (2017)
Snail
(9-12 week, juvenile),
Lymnaea stagnalis
R, M
21 d
PFOS-K
95%
-
20
NOEC
(survival, feeding rate,
mass change, length
change, protein
concentration)
50->50
>50
Duration too short for
a chronic test
Olson (2017)
Snail (adult),
Lymnaea stagnalis
R, M
21 d
PFOS-K
95%
-
20
MATC
(survival)
3.0-6
4.243
Duration too short for
a chronic test
Olson (2017)
Snail (5 mm),
Physella heterostropha
pomilia
(formerly, Physa pomilia)
S, M
50 hr
PFOS-K
>98%
-
25
NOEC-LOEC
(avoidance)
< 30-30
-
Duration too short for
an acute test; atypical
endpoint
Funkhouser
(2014)
Snail (adult, 4 mo.),
Physella heterostropha
pomilia
(formerly, Physa pomilia)
R, M
14 d
PFOS-K
>98%
-
25
LC50
-
94.99
Duration too long for
an acute test and too
short for a chronic test
Funkhouser
(2014)
Rotifer
(< 2 hr old neonates),
Brachionus calyciflorus
R,Ub
4 d
PFOS-K
98%
-
20
MATC
(intrinsic rate of
population increase and
resting egg production)
0.125-0.25
0.1768
Atypical
concentration-response
pattern
Zhang et al.
(2014)
Cladoceran (<24 hr),
Daphnia magna
R, U
25 d
PFOS-K
99%
7.8
20
MATC
(reproduction F0
generation)
0.01-0.1
0.03162
No consistent
concentration-response
relationship
Jeong et al. (2016)
Cladoceran (<24 hr),
Daphnia magna
s,u
48 hr
PFOS-Li
24.5%
8.6
20.1-
21.0
EC50
(death/immobility)
-
51.45
Inability to verify
author-reported LC50
3M Company
(2000)
G-4
-------
Spocics (lilestjiiicM
Method'
Tcsl
Dui'iilioii
( hcinic;il /
I'llril\
pll
Temp.
(ฐC)
IIIWl
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
I'.ITccl
(one.
(mป/l.)
Deficiencies
Reference
Cladoceran (0-24 hr),
Daphnia magna
S,U
48 hr
PFOS-K
Unreported
7.6
22
EC50
27
.Ynolhor lost within the
same publication had
24.5% purity and this
test purity was
unknown, could be of
low purity.
3\1 CompaiiN
(2000)
Cladoceran (0-12 hr),
Daphnia magna
R,U
28 d
PFOS-K
Unreported
7.6
22
MATC
(reproduction)
7.0-18.0
11.22
Inability to calculate
an EC 10 and
comments by authors
3M Company
(2000)
Cladoceran
(adult, ~ 14 d),
Daphnia magna
S,M
48 hr
PFOS-K
>98%
-
21
MATC
(biochemistry changes)
13.34-
27.33
19.09
Non-apical endpoint
Labine et al.
(2022)
Amphipod (7 d),
Hyalella azteca
R, M
7 d
PFOS-K
97.5%
7.11
(6.85-
7.46)
22.8
(22.1-
23.3)
EC20
(growth - biomass)
-
7.20
Duration too short for
a chronic exposure
Kadlec et al.
(2024)
Crayfish
(3 wk juvenile, 0.048 g),
Procambarus fallax f.
virginalis
R, M
38 d
PFOS-K
>98%
-
25
MATC
(survival/growth)
0.2->0.2
>0.2
Only two organisms
per exposure
concentration; invasive
species
Funkhouser
(2014)
Crayfish
(juvenile, 2 wk, 0.041 g),
Procambarus fallax f.
virginalis
R, M
7 d
PFOS-K
>98%
-
25
LC50
-
39.71
Duration too long for
an acute test and too
short for a chronic test,
only six organisms per
exposure
concentration, test
species fed; invasive
species
Funkhouser
(2014)
Crayfish (intermolt),
Pontastacus leptodactylus
(formerly, Astacus
leptodactylus)
R, M
21 d
PFOS-K
>98%
6.79
21
MATC
(oxidative status index)
0.5-5
1.581
Non-apical endpoint
Belek et al. (2022)
Oriental river prawn
(0.30 g, 4.0 cm),
Macrobrachium
nipponense
S,M
96 hr
PFOS-K
99%
7
22
LC50
-
19.77
Source of organisms
may be problematic
Yang et al. (2014)
G-5
-------
Spocics (lil'cs(;i;ic)
Method1
Tcsl
Dunilion
( hcmiciil /
Piiriu
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Yellow l'e\ or iiuisqiiiln
(1st instar),
Aedes aegypti
S,U
48 hr
PFOS
Unreported
25
LC50
-
1.18
Duration loo short for
an acute test, missing
some exposure details
Olson (2017)
Yellow fever mosquito
(1st instar),
Aedes aegypti
R,U
-42 d
PFOS
Unreported
-
25
MATC
(average time to
emergence)
0.05-0.125
0.079
Missing some
exposure details
Olson (2017)
Blue damselfly (larva, F2
instar stage),
Enallagma cyathigerum
R,U
4 mo
Perfluorooctanesulfonic
acid
tetraethylammonium
98%
-
21
MATC
(general activity, burst
swimming, foraging
success)
0.01-0.1
0.03163
Sporadic solution
renewal, behavioral
endpoints
Van Gossum et al.
(2009)
Midge (Instar, 3 d),
Chironomus dilutus
R, M
7 d
PFOS-K
97.5%
7.01
(6.54-
7.33)
22.6
(21.7-
23.7)
EC20
(growth - biomass)
-
0.018
Duration too short for
a chronic exposure
Kadlec et al.
(2024)
Midge (Instar, 3 d),
Chironomus dilutus
R, M
7 d
PFOS-K
97.5%
7.38
(7.32-
7.47)
22.9
(20.5-
23.9)
EC20
(growth - biomass)
-
0.016
Duration too short for
a chronic exposure
Kadlec et al.
(2024)
Midge,
Chironomus dilutus
R, M
10 d
PFOS
98%
LOEC
(mortality)
-
0.0004086
Range-finding test
McCarthy et al.
(2021)
Midge (0.05 g, 1.2 cm),
Chironomus plumosus
S,M
96 hr
PFOS-K
99%
7
22
LC50
-
182.12
Source of organisms
may be problematic
Yang et al. (2014)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00985
Only one exposure
concentration,
sediment exposure
Zhai et al. (2016)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00987
Only one exposure
concentration,
sediment exposure
Zhai et al. (2016)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00987
Only one exposure
concentration,
sediment exposure
Zhai et al. (2016)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00985
Only one exposure
concentration,
sediment exposure
Zhai et al. (2016)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00985
Only one exposure
concentration,
sediment exposure
Zhai et al. (2016)
G-6
-------
Spocics (lilestjiiicM
Method"
Tesl
Dui'iilioii
( hcmic;d /
Piiriu
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
25
NOEC
(mortality)
-
0.00985
Only one exposure
concentration,
sediment exposure
Zhaietal. (2016)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00987
Only one exposure
concentration,
sediment exposure
Zhaietal. (2016)
Midge
(larva, 3rd-4th instar),
Chironomus plumosus
S,M
10.33 d
PFOS
98%
-
25
NOEC
(mortality)
-
0.00987
Only one exposure
concentration,
sediment exposure
Zhaietal. (2016)
Midge
(multi-generational),
Chironomus riparius
S,M
10
generations
(-20-28 d ea.)
PFOS
Unspecified
oo
^ 00
20
NOEC
(emergence,
reproduction, sex ratio)
-
0.0035
Only one exposure
concentration, static
chronic test
Stefani et al.
(2014)
Midge
(multi-generational),
Chironomus riparius
S,M
10
generations
(-20-28 d ea.)
PFOS
Unspecified
oo
^ 00
20
LOEC
(increased mutation rate)
-
0.0035
Only one exposure
concentration, static
chronic test
Stefani et al.
(2014)
Midge (1st instar larva),
Chironomus riparius
S,M
-36 dd
(1st of 10
generations)
PFOS
Unreported
ฆA 98%
-
20
NOEC
(survival, growth)
0.011-
>0.011
0.011
Not true ELS test (28
days beginning with
juvenile)
Roland et al.
(2014)
European eel
(juvenile, 138.3 g),
Anguilla anguilla
R, M
28 d
PFOS-K
>98%
-
20
LOEC
(proteomic growth)
<0.00081-
0.00081
0.00081
Not true ELS test (28
days beginning with
juvenile), atypical
endpoint
Roland et al.
(2014)
Rainbow trout (immature,
16.4 cm, 22.7 g),
Oncorhynchus mykiss
Microcosm
12 d
PFOS
89%
9.2
6.0-
16.5
NOEC
(mortality)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Rainbow trout (immature,
16.4 cm, 22.7 g),
Oncorhynchus mykiss
Microcosm
12 d
PFOS
89%
9.2
6.0-
16.5
LOEC
(decrease LSI and
condition index (K) in
females)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Rainbow trout (female,
mature, 34.8 cm, 511.1 g),
Oncorhynchus mykiss
s,u
14 d
PFOS
89%
-
12
NOEC
(mortality)
-
1
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
G-7
-------
Spocics (lilestjiiicM
Method'
Tcsl
Dui'iilioii
( hcinic;il /
I'llril\
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Rainbow trout (female,
mature, 34.8 cm, 511.1 g),
Oncorhynchus mykiss
S,U
14 d
PFOS
89%
12
LOEC
(decrease LSI)
-
1
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Rainbow trout (11 mo),
Oncorhynchus mykiss
Diet, U
15 d
PFOS-K
Unknown
-
12
NOEC
(growth - weight)
-
250
mg/kg bw
per day
Dietary exposure
Benninghoff et al.
(2011)
Rainbow trout (fry, 15
week),
Oncorhynchus mykiss
Diet, U
8 mo
PFOS-K
Unknown
-
12
LOEC
(survival, tumor
incidence)
-
2.5 mg/kg
bw per
day
Dietary exposure,
mixture exposure
Benninghoff et al.
(2012)
Rainbow trout
(oocyte, ova),
Oncorhynchus mykiss
S,M
3 hr
PFOA
>97%
-
6
NOEC
(accumulation residue)
-
0.87
Duration too short for
an acute test
Raine et al. (2021)
Rainbow trout
(oocyte, ova),
Oncorhynchus mykiss
S,M
3 hr
PFOA
>97%
8.5
6
LOEC
(accumulation residue)
-
7.47
Duration too short for
an acute test
Raine et al. (2021)
Rainbow trout
(juvenile, ~7 mo),
Oncorhynchus mykiss
R, M
10 d
PFOS
Unreported
-
12
LOEC
(enzymatic changes)
-
0.0008
Non-apical endpoint,
duration too short for a
chronic exposure
Solan et al. (2022)
Atlantic salmon (embryo),
Salmo salar
R, U
49 d
Sodium
perfluoro-1-
octanesulfonate
Unreported
-
5-7
NOEC
(growth - length and
weight)
0.1->0.1
0.1
Only one exposure
concentration; greater
than low value so less
informative
Arukwe et al.
(2013)
Goldfish
(6.91 g, 6.01 cm),
Carassius curatus
R, U
48 hr
PFOS-K
>99%
-
18
NOEC-LOEC
(swimming behavior:
motion distance and % of
actionless time)
2.0-8
-
Atypical endpoint and
source of organisms,
duration too short for
an acute test
Xia et al. (2013a)
Goldfish (6.0 g, 7.0 cm),
Carassius curatus
S,M
96 hr
PFOS-K
99%
7
22
LC50
-
81.18
Source of organisms
may be problematic
Yang et al. (2014)
Goldfish
(juvenile, 27.85 g),
Carassius curatus
S,M
96 hr
PFOS
>98%
7.25
23
Antioxidant enzyme
activity
-
5.oor
Atypical endpoint, no
point estimate
Feng et al. (2015)
Common carp (juvenile,
3.72g, 5.18 cm),
Cyprinus carpio
R, Uc
14 d
PFOS
>98%
-
-
NOEC
(liver protein)
1->1
1
Duration too short for
a chronic test, atypical
endpoint
Hagenaars et al.
(2008)
G-8
-------
Spocics (lilestjiiicM
Method"
Tesl
Dui'iilioii
( hcmiciil /
Piiriu
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Common carp (juvenile,
3.72g, 5.18 cm),
Cyprinus carpio
R, Uc
14 d
PFOS
>98%
-
MATC
(liver glycogen)
0.5-1
0.7071
Duration too short for
a chronic test, atypical
endpoint
Hagenaars et al.
(2008)
Common carp (juvenile,
3.72g, 5.18 cm),
Cyprinus carpio
R, Uc
14 d
Perfluorooctanesulfonic
PFOS
>98%
-
-
NOEC
(liver lipid)
1->1
1
Duration too short for
a chronic test, atypical
endpoint
Hagenaars et al.
(2008)
Common carp (juvenile,
3.72g, 5.18 cm),
Cyprinus carpio
R, Uc
14 d
PFOS
>98%
-
-
LOEC
(relative condition factor)
<0.1-0.1
0.1
Duration too short for
a chronic test
Hagenaars et al.
(2008)
Common carp (juvenile,
3.72g, 5.18 cm),
Cyprinus carpio
R, Uc
14 d
PFOS
>98%
-
-
MATC
(HSI)
0.1-0.5
0.2236
Duration too short for
a chronic test, atypical
endpoint
Hagenaars et al.
(2008)
Common carp (juvenile,
~12 cm; ~20 g),
Cyprinus carpio
F, M
96 hr
PFOS
100.3%
6.9
23
LOEC
(DNA damage)
-
5.395
Atypical endpoint
Kim et al. (2010)
Common carp (juvenile),
Cyprinus carpio
R, M
76 d
PFOS
98%
-
-
NOEC
(growth - weight)
-
0.00082
Greater than low value
Shan et al. (2022)
Zebrafish
(female fry, 14 dpf),
Danio rerio
R, U
70 d
PFOS-K
>99%
-
27
EC10
(male weight)
0.01-0.05
0.001990
Missing some
exposure details
Du et al. (2009)
Zebrafish
(embryo - blastula stage),
Danio rerio
R, U
Fert. up to
15 dpf
PFOS-K
99%
-
-
MATC
(body length and average
weight)
0.200-
0.400
0.2828
Duration too short for
a chronic test
Shi et al. (2009)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,M
114 hr
PFOS
>96%
7.0-
7.5
28
LC50
-
2.20
Duration too long for
an acute test
Huang et al.
(2010)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,M
114 hr
PFOS
>96%
7.0-
7.5
28
EC50
(malformation)
-
1.12
Duration too long for
an acute test, atypical
endpoint
Huang et al.
(2010)
Zebrafish (embryo),
Danio rerio
R, M
21 d
PFOS-K
98%
-
26
LOEC
(reduce fecundity)
<0.5-0.5
0.5
Only one exposure
concentration, control
issues
Sharpe et al.
(2010)
Zebrafish (embryo),
Danio rerio
R, M
48 hr
PFOS-K
98%
-
26
LC50
(range of 3 tests)
-
7.7-38.9
Duration too short for
an acute test, results
are not reproducible
Sharpe et al.
(2010)
Zebrafish
(embryo, 4 hpf),
Danio rerio
S,U
96 hr
PFOS-K
>99%
-
28.5
NOEC-LOEC
(increased ROS
formation)
0.2-0.4
-
Atypical endpoint,
missing exposure
details
Shi and Zhou
(2010)
G-9
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
I'llril\
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish (cilibi\o).
Danio rerio
R,U
96 hr
PFOS-K
98%
26
LC50
-
54.36e
Problems uilli
reported data to be
used for LC50 analysis
Ding el al. (.2012;,
Ding et al. (2013)
Zebrafish
(F2 embryo, 0 hpf),
Danio rerio
F, M
300-330 d
PFOS-K
>98%
8.25-
8.75
26
MATC
(F2 180 d survival)
0.1-0.3
0.1732
Poor concentration-
response, test design
complications
Keiter et al.
(2012)
Zebrafish
(embryo, 6-8 hpf),
Danio rerio
R,U
6 d
PFOS
Unreported
-
26
AC50
(toxicity score: includes
survival, hatchability, and
malformation index)
-
16.44e
Duration too long for
an acute test and too
short for a chronic test,
atypical endpoint
Padilla et al.
(2012)
Zebrafish (embryo),
Danio rerio
S,U
72 hr
PFOS
98%
8.3
28.5
LC50
-
68
Duration too short for
an acute test, missing
some exposure details
Zheng et al.
(2012)
Zebrafish (embryo),
Danio rerio
S,U
72 hr
PFOS
98%
8.3
28.5
EC50
(malformation)
-
37
Duration too short for
an acute test, atypical
endpoint, missing
exposure details
Zheng et al.
(2012)
Zebrafish
(embryo, FO generation),
Danio rerio
R, U
120 dpf
PFOS
>96%
OO
28
LOEC
Increase mortality and
malformations in the F1
generation
<0.250-
0.250
0.250e
Only one exposure
concentration
Chen et al. (2013)
Zebrafish (embryo, 4hpf),
Danio rerio
R, U
120 hr
PFOS
>98%
-
28
NOEC-LOEC
(suppression of
steroidogenic enzyme
synthesis)
0.1-0.2
-
Duration too long for
an acute test, atypical
endpoint, missing
exposure details
Du et al. (2013)
Zebrafish
(embryo, 4 hpf),
Danio rerio
R, U
116 hr
PFOS-K
Unreported
-
-
NOEC
(development, hatch,
mortality)
5->5
5
Duration too long for
an acute test and too
short for a chronic test,
only one exposure
concentration
Liu et al. (2013)
Zebrafish
(embryo - 4 cell stage),
Danio rerio
S,U
Fert. to
144 hpf
PFOS
Unreported
7.2-
7.6
26
EC50
(lethal and sublethal
effects)
-
1.5
Duration too long for
an acute test and too
short for a chronic test,
static chronic exposure
Ulhaq et al.
(2013)
Zebrafish
(embryo - 4 cell stage),
Danio rerio
S,U
Fert. to
144 hpf
PFOS
Unreported
7.2-
7.6
26
LC50
-
>10
Duration too long for
an acute test and too
short for a chronic test,
static chronic exposure
Ulhaq et al.
(2013)
Zebrafish (embryo),
Danio rerio
S,U
48 hr
PFOS
>96%
-
-
Malformation
(100%)
-
8.002e
Duration too short for
an acute test, atypical
endpoint, no point
estimate
Chen et al. (2014)
Zebrafish (embryo),
Danio rerio
R, U
6 d
PFOS-K
98%
7.5
28.5
LC50
-
6.25
Duration too long for
an acute test and too
short for a chronic test
Hagenaars et al.
(2014)
G-10
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
Ptiriu
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Kcported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish (embryo),
Danio rerio
R,U
6 d
PFOS-K
98%
7.5
28.5
EC50
(uninflated swim bladder)
-
2.29
Duration loo long ibr
an acute test and too
short for a chronic test,
atypical endpoint
Hagenaars et al.
(2014)
Zebrafish
(embryo, 2 hpf),
Danio rerio
S,U
6 d
PFOS-K
>98%
7.4
28
NOEC-LOEC
(behavior: spontaneous
swimming activity)
0.1-1.0
-
Duration too long for
an acute test and too
short for a chronic test,
atypical endpoint, only
two exposure
concentrations
Spulber et al.
(2014)
Zebrafish (embryo, 6
hpf),
Danio rerio
S,U
114 hr
PFOS-K
Unknown
-
28
LOEC
(mortality)
3.307-
33.07
33.07e
Duration too long for
an acute test and too
short for a chronic test
Truong et al.
(2014)
Zebrafish (embryo, 6
hpf),
Danio rerio
S,U
114 hr
PFOS
Unknown
-
28
LOEC
(mortality)
0.32-3.2
3.2e
Duration too long for
an acute test and too
short for a chronic test
Truong et al.
(2014)
Zebrafish
(embryo, 8 hpf),
Danio rerio
R,U
42 dpf
PFOS
>96%
7.0-
7.5
28
LOEC
(increased condition
index)
-
0.25
Only one exposure
concentration
Chen et al. (2016)
Zebrafish
(embryo, 8 hpf),
Danio rerio
R, U
150 dpf
PFOS
>96%
7.0-
7.5
28
LOEC
(increased estradiol in
male/females and
testosterone in males)
-
0.25
Only one exposure
concentration
Chen et al. (2016)
Zebrafish (larva, 120 hpf),
Danio rerio
S,M
24 hr
PFOS
>98%
7.0-
7.5
28
NOEC
(various metabolites)
-
9.700
Duration too short for
an acute test, atypical
endpoint
Huang et al.
(2016)
Zebrafish (embryo),
Danio rerio
R, U
6 d
PFOS
Unknown
-
28
LOEC
(liver size and gene
expression)
<0.0005-
0.0005
0.0005
Duration too long for
an acute test and too
short for a chronic test,
atypical endpoint
Tse et al. (2016)
Zebrafish
(embryo, 8 hpf),
Danio rerio
R, U
180 d
PFOS
>96%
7.0-
7.5
27
MATC
(altered sex ratio: female
dominance, F1 offspring
survival)
0.05-0.25
0.1118e
Non-apical endpoint
Cui et al. (2017)
Zebrafish (3 hpf),
Danio rerio
S,U
5 d + 9 d
observation
PFOS
Unreported
7.2-
7.7
26-28
MATC
(growth - total body
length)
0.02-0.2
0.06325e
Duration too long for
an acute test and too
short for a chronic test,
static chronic exposure
Jantzen et al.
(2017)
Zebrafish (3 hpf),
Danio rerio
S,U
5 d + 9 d
observation
PFOS
Unreported
7.2-
7.7
26-28
LOEC
(interoccular distance)
< 0.02-
0.02
0.02e
Duration too long for
an acute test and too
short for a chronic test,
static chronic exposure
Jantzen et al.
(2017)
Zebrafish (3 hpf),
Danio rerio
S,U
5 d + 9 d
observation
PFOS
Unreported
7.2-
7.7
26-28
MATC
(yolk sac area)
0.02-0.2
0.06325e
Duration too long for
an acute test and too
Jantzen et al.
(2017)
G-ll
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
Ptiriu
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Kcported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
short for a chronic lost,
static chronic exposure
Zebrafish (3 hpf),
Danio rerio
S,U
5 d + 9 d
observation
PFOS
Unreported
7.2-
7.7
26-28
LOEC-
(swimming activity -
crossing frequency)
< 0.02-
0.02
0.02e
Duration too long for
an acute test and too
short for a chronic test,
static chronic exposure
Jantzen et al.
(2017)
Zebrafish (embryo),
Danio rerio
S,M
48 hr
PFOS
Unknown
-
27
LC50
-
107.6
Duration too short for
an acute test
Rainieri et al.
(2017)
Zebrafish
(embryo, 3 hpf),
Danio rerio
R,U
7 d
PFOS
Unreported
-
28.5
MATC
(islet morphological
anomalies)
8.0-16.0e
11.31ฎ
Duration too long for
an acute test and too
short for a chronic test
Sant et al. (2017)
Zebrafish (sperm),
Danio rerio
S,U
20 sec
PFOS-K
>98%
8
25
NOEC-LOEC
(sperm motility)
0.09-0.9
-
Duration too short for
an acute test, atypical
endpoint
Xia and Niu
(2017)
Zebrafish (sperm/egg),
Danio rerio
S,U
2 min
PFOS-K
>98%
8
25
NOEC-LOEC
(fertilization success)
0.09-0.9
-
Duration too short for
an acute test, atypical
endpoint
Xia and Niu
(2017)
Zebrafish (embryo, 3
hpf),
Danio rerio
S,U
5 d
PFOS
Unknown
7.2-
7.7
27
LOEC
(gene expression of
Leptin A mRNA)
<0.01-0.01
0.01e
Duration too long for
an acute test and too
short for a chronic test,
atypical endpoint
Annunziato
(2018)
Zebrafish
(embryo, 1-2 hpf),
Danio rerio
R,U
96 hr
PFOS
>99%
-
25
NOEC-LOEC
(growth: body length)
<0.050-
0.050
-
Atypical endpoint,
missing exposure
details
Dang et al. (2018)
Zebrafish (embro, 2 hpf),
Danio rerio
R, U
72 hr
PFOS
Unknown
-
28.5
LOEC
(malformations)
0.5-1.0
1.0e
Duration too short for
an acute test
Ortiz-Villanueva
et al. (2018)
Zebrafish (embro, 2 hpf),
Danio rerio
R, U
72 hr
PFOS
Unknown
-
28.5
LOEC
(survival)
5.0-10
10e
Duration too short for
an acute test
Ortiz-Villanueva
et al. (2018)
Zebrafish
(embryo, 1 hpf),
Danio rerio
R, U
96 hr
PFOS
Unreported
7.6
28.5
NOEC-LOEC
(pericardial area)
8-16e
-
Atypical endpoint,
missing exposure
details
Sant et al. (2018)
Zebrafish
(embryo, 1 hpf),
Danio rerio
S,U
96 hr
PFOS
Unreported
-
26
NOEC
(enzymes, olfactory cells)
-
6.650
Atypical endpoint,
missing exposure
details
Stengel et al.
(2017a)
Zebrafish (female, 4 mo),
Danio rerio
R, U
21 d
PFOS
Unknown
7.0-
7.5
28
NOEC
(growth - length and
weight)
0.2->0.2
0.2
Inability to
independently verify
effect values, partial
life cycle test
Bao et al. (2019)
Zebrafish (embryo,
maximum of 4 hpf),
Danio rerio
R, M
96 hr
PFOS
Unknown
-
26
NOEC
(hatching success,
embryo mortality,
deformation)
0.0007-
>0.0007
0.0007
Greater than low value
Cormier et al.
(2019)
G-12
-------
Spocics (lilestjiiicM
Method"
Tesl
Dui'iilioii
( hcmic;d /
Piiriu
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish (embryo, 2
hpf),
Danio rerio
R,U
72 hr
PFOS-K
>98%
28
LOEC
(growth - total body
length)
2.691-
5.832
5.382e
Duration too short for
an acute test
Martinez et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS-K
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
5.732e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS-K
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(mortality)
-
3.014e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
2.526e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
6.357e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
4.181e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
6.642e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
2.786e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
1.180s
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
5.211ฎ
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
G-13
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
I'llril\
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS
Unreported
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
1.370e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
4.751e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
4.641e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
4.791e
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(abnormal development)
-
5.8778
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
s,u
114 hr
PFOS
Unreported
-
-
Benchmark Dose
Value at 10% extra
effect
(mortality)
-
0.55016
Duration too long for
an acute test, atypical
test endpoint
Thomas et al.
(2019)
Zebrafish (embryo),
Danio rerio
R, M
6 d
PFOS
>98%
7.5
28.5
NOEC
(growth and survival)
-
1.339
Greater than low value
Tu et al. (2019)
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
118 hr
PFOS-K
>98%
-
28
EC50
(mortality,
malformations)
-
2.045e
Duration too long for
an acute test and too
short for a chronic test,
mixed test endpoints
Vogs et al. (2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
R, M
96 hr
Perfluoroctane
sulfonate
sodium salt
>98%
-
26
NOEC
(mortality, hatch)
-
0.4514e
Only one exposure
concentration
Yi et al. (2019)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
90-94 hr
PFOS-K
>98%
7-7.5
28
NOEC
(mortality, deformity)
-
2.06
Duration too short for
an acute test,
behavioral focus with
secondary reference to
no mortality or
deformity
Christou et al.
(2020)
G-14
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
I'llril\
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
I'.ITccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish (embryo, 6
hpf),
Danio rerio
S,U
66 tu-
PFOS
97%
28
NOEC
(survival)
25->25
25e
Duration too short for
an acute test
Dasgupta et al.
(2020)
Zebrafish
(embryo 4-64 stage),
Danio rerio
S,M
rn hr
PFOS
98%
7.2-
7.6
26
MATC
(abnormal development)
0.16-2.2
0.5933
Duration too long for
an acute test and too
short for a chronic test
Menger et al.
(2020)
Zebrafish (adult, 4.5 mo),
Danio rerio
R,U
28 d
PFOS
>95%
-
28
LOEC
(reproduction - sperm
development)
<0.2510-
0.2510
0.2501"
Atypical endpoint
Xin et al. (2020)
Zebrafish (adult, 4.5 mo),
Danio rerio
R,U
28 d
PFOS
>95%
-
28
NOEC
(reproduction - oocyte
development)
0.2510-
>0.2510
0.2501"
Atypical endpoint
Xin et al. (2020)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
90 hr
exposure +
177 d
observation
PFOS-K
>98%
7.0-
7.5
28-
28.5
NOEC
(mortality)
2.06->2.06
2.06
Atypical exposure
duration
Christou et al.
(2021)
Zebrafish
(embryo, 4 hpf),
Danio rerio
R, U
92 hr
PFOS
>98%
-
-
NOEC
(mortality)
0.5->0.5
0.5
Only one exposure
concentration, duration
too short for an acute
test
Dong et al. (2021)
Zebrafish (embryo),
Danio rerio
R, M
120 hr
PFOS-K
>95%
-
28.5
NOEC
(mortality)
-
>1.803e
Duration too long for
an acute exposure, too
short for a chronic test
Han et al. (2021)
Zebrafish (embryo),
Danio rerio
R, M
120 hr
PFOS
>95%
-
28.5
NOEC
(mortality)
-
>1.730e
Duration too long for
an acute exposure, too
short for a chronic test
Han et al. (2021)
Zebrafish
(embryo, 6 hpf),
Danio rerio
R, M
96 hr
PFOS-K
>98%
-
28.5
LOEC
(malformations,
locomotive behavior)
<20-20
20
Only one exposure
concentration; atypical
endpoint
Huang et al.
(2021)
Zebrafish
(embryo, 6 hpf),
Danio rerio
R, U
96 hr
PFOS-K
>98%
-
-
MATC
(growth, histology,
abnormal development)
10.0-20.0
14.14
Atypical endpoint for
an acute test
Huang et al.
(2021)
Zebrafish
(embryo, 4 hpf),
Danio rerio
S,U
116.83 hr
PFOS
Unreported
-
28.0
NOEC
(mortality)
10.00-
>10.00
10.00c
Atypical duration
Lee et al. (2021)
Zebrafish
(embryo, 6-8 hpf),
Danio rerio
S,U
112-114 hr
Potassium
perfluoro-1-
octanesulfonate
>98%
-
28
NOEC
(mortality)
-
0.2476e
Duration too long for
an acute test; only one
exposure concentration
Rericha et al.
(2021)
G-15
-------
Spocics (lilestjiiicM
Method"
Tesl
Dui'iilioii
( hcmic;d /
I'llril\
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish (embryo, <1
hpf),
Danio rerio
R,U
96 hr
PFOS
Unreported
28.5
LOEC
(increase lauric C12:0
and myristic C14:0 fatty
acids)
<8.002-
8.002
8.002e
Atypical endpoint
Sant et al. (2021)
Zebrafish (dechorinated
embryo, 1 dpf),
Danio rerio
R,U
30 d
PFOS
Unreported
-
28.5
NOEC
(growth - length)
16->16
16e
Growth effects not the
focus of study rather
other non-apical
endpoints
Sant et al. (2021)
Zebrafish (adult, 90 dpf),
Danio rerio
R, U
10 d
PFOS-K
>98%
7.21
28.0
LOEC
(gene expression)
<0.5-0.5
0.5
Atypical endpoint
Zhuetal. (2021)
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
120 hr
PFOS-K
Unreported
-
28
MATC
(growth - length)
0.050-
2.066
0.3214
Duration too long for
an acute exposure, too
short for a chronic test
Fey et al. (2022)
Zebrafish
(embryo, <4 hpf),
Danio rerio
R, U
5 d
PFOS
99.4%
-
27
NOEC
(mortality and
development)
-
>0.0024
Duration too long for
an acute exposure, too
short for a chronic test,
test represents a
greater than low value
(followed decision
rule; Section 2.10.3.2)
Haimbaugh et al.
(2022)
Zebrafish
(embryo, <4 hpf),
Danio rerio
R, U
4-6 wk
PFOS
99.4%
-
27
NOEC
(reproduction and
growth)
-
>0.0024
Test represents a
greater than low value
(followed decision
rule; Section 2.10.3.2)
Haimbaugh et al.
(2022)
Zebrafish
(embryo, 4 hpf),
Danio rerio
S,U
116 hr
PFOS
Unreported
-
-
NOEC
(development, mortality)
-
10.00e
Duration too long for
an acute exposure, too
short for a chronic test,
test represents a
greater than low value
(followed decision
rule; Section 2.10.3.2)
Lee et al. (2022)
Zebrafish
(embryo, 3 hpf),
Danio rerio
S,U
117 hr
PFOS-NA
>98%
-
-
LC50
-
16.47e
Duration too long for
an acute exposure, too
short for a chronic test
Lindqvist and
Wincent (2022)
Zebrafish
(embryo, 7 hpf),
Danio rerio
R, M
30 d
PFOS-K
>98%
6.86-
7.39
24.9-
25.3
EC10
(growth-weight)
0.004-
0.140
0.098
Poor control survival
(75%)
Krupa et al.
(2022)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS
Unreported
-
28
Benchmark Dose
Value at 10% extra
effect
(morphology)
-
7.753e
Duration too long for
an acute exposure, too
short for a chronic test
Truong et al.
(2022)
Zebrafish
(embryo, 6 hpf),
Danio rerio
S,U
114 hr
PFOS-K
Unreported
-
28
Benchmark Dose
Value at 10% extra
-
5.930e
Duration too long for
an acute exposure, too
short for a chronic test
Truong et al.
(2022)
G-16
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
I'llril\
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
effect
(morphology)
Zebrafish
(embryo, 2 hpf),
Danio rerio
R,U
94 hr
PFOS
98.7%
7.2
28.5
NOEC
(growth - length)
-
>0.5001e
Duration too short for
an acute exposure.
Greater than low
value.
Wang et al.
(2022)
Zebrafish
(embryo, 2 hpf),
Danio rerio
R,U
82 hr
PFOS
98.7%
7.2
28.5
NOEC
(hatching rate)
-
>o.50or
Duration too short for
an acute exposure.
Greater than low
value.
Wang et al.
(2022)
Zebrafish
(embryo, 2 hpf),
Danio rerio
R, M
118 hr
PFOS-K
>98%
-
28.0
NOEC
(mortality)
-
2.418
Duration too long for
an acute test
Wu et al. (2022)
Zebrafish
(embryo, 0.5 hpf),
Danio rerio
S,U
119.5 hr
PFOS-K
>98%
7.2
28
EC50
(abnormal development)
-
6
Duration too long for
an acute exposure, too
short for a chronic test,
atypical endpoint
Gui et al. (2023)
Zebrafish
(embryo, 5 hpf),
Danio rerio
R, U
8 mo
PFOS
Unreported
00 p
28
NOEC
(mortality)
-
>0.50018
Greater than low value
Hawkeye et al.
(2023)
Zebrafish
(embryo, 72 hpf),
Danio rerio
R, U
48.83 hr
PFOS
>98%
-
28.5
LOEC
(swimming behavior)
-
0.1
Duration too short for
an acute exposure,
atypical endpoint
Kalyn et al.
(2023)
Zebrafish (adult, 3 mo),
Danio rerio
F, U
14 d
PFOS
Unreported
7.0-
7.4
28
LOEC
(growth - weight)
<0.08-0.08
0.08
Duration too short for
a chronic exposure
Lu et al. (2024)
Zebrafish (2 mo),
Danio rerio
R, U
14 d
PFOS-K
Unreported
7
28
LOEC
(biochemistry and
enzymatic changes)
-
0.03
Non-apical endpoints,
duration too short for a
chronic exposure
Liu et al. (2023a)
Zebrafish
(embryo, 4 hpf),
Danio rerio
R, U
92 hr
PFOS-K
>98%
7.2-
7.4
27
MATC
(growth - length)
0.100-
0.500
0.2236
Non-apical endpoints,
duration too short for
an acute exposure
Mahapatra et al.
(2023)
Zebrafish
(embryo, 4 hpf),
Danio rerio
S,U
116 hr
PFOS-K
>98%
-
28.5
LC50
-
24.778
Duration too long for
an acute exposure and
too short for a chronic
exposure
Min et al. (2023)
Zebrafish
(embryo, 6 hpf),
Danio rerio
R, U
90 hr
PFOS-K
Unreported
-
28
AC50
(development)
-
1.9658
Duration too short for
an acute exposure
Phelps et al.
(2023)
Zebrafish
(embryo, 6 hpf),
Danio rerio
R, U
90 hr
PFOS-K
Unreported
-
28
AC50
(growth length)
-
1.35 Is
Duration too short for
an acute exposure
Phelps et al.
(2023)
G-17
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
Ptiriu
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Zebrafish
(embryo, 6 hpf),
Danio rerio
R,U
90 hr
PFOS-K
Unreported
28
AC50
(morphology)
-
1.1258
Duration too short for
an acute exposure
Phelps et al.
(2023)
Zebrafish
(embryo, 2 hpf),
Danio rerio
R,U
70 hr
PFOS-K
>98%
-
26
LC50
-
2.104g
Duration too short for
an acute exposure
Zhou et al. (2023)
Spottail shiner (female,
mature, 8.9 cm, 6.7 g),
Notropis hudsonius
Microcosm
14 d
PFOS
89%
-
-
NOEC
(mortality)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Spottail shiner (female,
mature, 8.9 cm, 6.7 g),
Notropis hudsonius
Microcosm
14 d
PFOS
89%
-
-
LOEC
(increase TBARS in
liver/ovary and FAO
activity in liver)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Fathead minnow
(embryo, 48 hpf),
Pimephales promelas
F, M
33 d
PFOS-K
Unreported
6.6-
7.3
22-26
EC10
(survival)
1.0-1.9
1.378
Another test within the
same publication had
24.5% purity and this
test purity was
unknown, could be of
low purity
3M Company
(2000)
Fathead minnow
(mature, 6.1 cm, 2.0 g),
Pimephales promelas
Microcosm
28 d
PFOS
89%
9.2
16.6-
22.8
LC10
-
3.5
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Fathead minnow
(embryo, <2 hpf),
Pimephales promelas
R, M
14 d
PFOS-K
>98%
8.4
27
MATC
(mortality)
1.25-2.50
1.768
Duration, atypical
exposure (5 d in
toxicant + 9 d in clean
water)
Krzykwa et al.
(2021)
Topmouth gudgeon
(juvenile female, 0.81 g,
4.03 cm),
Pseudorasbora parva
R, U
96 hr
PFOS-K
>99%
-
15
NOEC-LOEC
(spontaneous swim
behavior: swim distance)
0.5-2
-
Atypical endpoint and
source of organisms
Xia et al. (2014)
Topmouth gudgeon
(4.0 g, 4.0 cm),
Pseudorasbora parva
S,M
96 hr
PFOS-K
99%
7
22
LC50
-
67.74
Source of organisms
may be problematic
Yang et al. (2014)
Topmouth gudgeon
(4.0 g, 4.0 cm),
Pseudorasbora parva
R, M
30 d
PFOS-K
99%
7
22
EC10
(survival)
-
2.12
Not a true ELS test
(started with older life
stage), renewal chronic
exposure, source of
organisms may be
problematic
Yang et al. (2014)
G-18
-------
Spocics (lilestjiiicM
Method'
Tesl
Dui'iilioii
( hcmic;il /
Piiriu
nil
Icm|).
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.fl'ccl
(one.
(mป/l.)
Deficiencies
Reference
Creek chub
(mature, 11.8 cm, 16.3 g),
Semotilus atromaculatus
Microcosm
14 d
PFOS
89%
-
-
NOEC
(mortality)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Creek chub
(mature, 11.8 cm, 16.3 g),
Semotilus atromaculatus
Microcosm
14 d
PFOS
89%
-
-
LOEC
(increase TBARS in
liver/ovary and FAO
activity in liver)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Quinbo (juvenile, 2.77 g,
5.62 cm),
Spinibarbus sinensis
R,U
30 d
PFOS-K
>99%
OO
VO r-
18
MATC
(% mobile, % highly
mobile, swim distance,
swim speed, freq. highly
mobile, % social, resting
metabolic rate)
0.32-0.80
0.506
Test was not replicated
Xia et al. (2015b)
Quinbo (juvenile, 2.77 g,
5.62 cm),
Spinibarbus sinensis
R,U
30 d
PFOS-K
>99%
00
VO r-
18
MATC
(decrease maximum
linear acceleration)
0.32-0.80
0.506
Atypical endpoint
Xia et al. (2015c);
Xia et al. (2015a)
Quinbo (juvenile, 2.77 g,
5.62 cm),
Spinibarbus sinensis
R, U
30 d
PFOS-K
>99%
00
28
MATC
(decrease maximum
linear acceleration)
0.32-0.80
0.506
Atypical endpoint
Xia et al. (2015a);
Xia et al. (2015b)
White sucker (mature,
22.7 cm, 114.5 g),
Catostomus commersonii
Microcosm
14 d
PFOS
89%
-
-
NOEC
(mortality)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
White sucker (mature,
22.7 cm, 114.5 g),
Catostomus commersonii
Microcosm
14 d
PFOS
89%
-
-
LOEC
(decrease LSI in females)
-
3
Atypical exposure, not
a true ELS test
Oakes et al.
(2005)
Bluegill
(28.6 mm, 0.60 g),
Lepomis macrochirus
s,u
96 tu-
PFOS DEA salt
Unreported
8.2-
8.3
-
LC50
-
31
Only one replicate per
treatment
3M Company
(2000)
Nile tilapia,
Oreochromis niloticus
Diet, U
rn hr
PFOS
Unreported
-
22
MATC
(weight and survival)
1.0-5.0
(mg/g bw)
2.236
(mg/g bw)
Duration too short for
a chronic test, missing
exposure details
Han et al. (2011)
Medaka (adult, male),
Oryzias latipes
R, U
14 d
PFOS
Unreported
-
25
NOEC
(adult survival, GSI%,
HSI%, condition factor)
1->1
1
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
G-19
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
Ptiriu
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
Medaka uidull. Icmalcj.
Oryzias latipes
R,U
14 d
PFOS
Unreported
25
NOLL
(adult survival, condition
factor)
1->1
1
Duration loo long lor
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka (adult, female),
Oryzias latipes
R,U
14 d
PFOS
Unreported
-
25
LOEC
(GSI%)
<0.01-0.01
0.01
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka (adult, female),
Oryzias latipes
R, U
14 d
PFOS
Unreported
-
25
MATC
(HSI%)
0.1-1
0.3162
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka (F1 generation,
<12 hr, embryo),
Oryzias latipes
R, U
7-14 d
(assumed)
PFOS
Unreported
-
25
MATC
(% hatchability, time to
hatch)
0.1-1
0.3162
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka (F1 generation,
<12 hr, embryo),
Oryzias latipes
R, U
~28 d post-
hatch
(assumed)
PFOS
Unreported
-
25
MATC
(swim up success)
0.1-1
0.3162
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka (F1 generation,
<12 hr, embryo),
Oryzias latipes
R, U
100 d post-
hatch
PFOS
Unreported
-
25
EC10
(growth - length)
<0.01-0.01
0.0013
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka (F1 generation,
<12 hr, embryo),
Oryzias latipes
R, U
28 d post-
hatch
PFOS
Unreported
-
25
LOEC
(larval survival)
<0.01-0.01
0.01
Duration too long for
an acute test and too
short for a chronic test
Ji et al. (2008)
Medaka
(adult, 16 week, 0.38g)
Oryzias latipes
R, U
21 d
PFOS
>98%
-
25
LOEC
(fecundity)
<1.0-1.0
1
Only one exposure
concentration
Kang et al. (2019)
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.3
24
EC50
(teratogenesis)
-
12.1
Atypical acute
endpoint
Palmer and
Krueger (2001)
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.27
24
EC50
(teratogenesis)
-
17.6
Atypical acute
endpoint
Palmer and
Krueger (2001)
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.26
24
EC50
(teratogenesis)
-
16.8
Atypical acute
endpoint
Palmer and
Krueger (2001)
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.3
24
NOEC
(growth)
-
14.7
Atypical acute
endpoint
Palmer and
Krueger (2001)
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.27
24
LOEC
(growth)
-
7.97
Atypical acute
endpoint
Palmer and
Krueger (2001)
G-20
-------
Spocics (lilestjiiicM
Method"
Test
Dui'iilioii
( hcmiciil /
Ptiriu
pll
Temp.
(ฐC)
r. fleet
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.l'l'ccl
(one.
(mป/l.)
Deficiencies
Reference
African clawed frog
(embryos),
Xenopus laevis
R, M
96 hr
PFOS-K
86.9%
7.26
24
LOEC
(growth)
-
8.26
Atypical acute
endpoint
Palmer and
Krueger (2001)
African clawed frog
(tadpoles, NF stage
46/47),
Xenopus laevis
R,U
67 d
PFOS
>96%
-
22
NOEC
(survival and forelimb
emergence)
0.1->0.1
0.1
Control issues
Cheng et al.
(2011)
African clawed frog
(embryo, NF 10),
Xenopus laevis
R, M
96 hr
PFOS
>99%
-
24
LC50
-
>96
Non-definitive value
San-Segundo et
al. (2016)
American toad (larva,
Gosner stage 25),
Anaxyrus americanus
R, M
26-45 d
PFOS
>98%
-
20
NOEC
(mortality, growth,
development)
-
0.6162
Lack of dose response
Flynn et al. (2022)
Asiatic toad
(tadpole, 1.8 cm, 0.048 g),
Bufo gargarizans
R, M
96 hr
PFOS-K
99%
7
22
LC50
-
48.21
Source of organisms
may be problematic
Yang et al. (2014)
Asiatic toad
(tadpole, 1.8 cm, 0.048 g),
Bufo gargarizans
R, M
30 d
PFOS-K
99%
7
22
EC10
(survival)
-
2.00
Renewal chronic
exposure, not a true
ELS test, source of
organisms may be
problematic
Yang et al. (2014)
Bullfrog (tadpole, Gosner
stage 25),
Lithobates catesbeiana
(formerly, Rana
catesbeiana)
R, M
65 d
PFOS-K
>98%
-
23
LOEC
(growth: snout-vent
length)
-
0.002430
Potential mixture
effects, missing
exposure details (prior
exposure)
Lech et al. (2022)
Northern leopard frog
(Gosner stage 25),
Lithobates pipiens
S,M
116 d
PFOS
Unknown
7.41-
8.54
13.1-
29.8
NOEC
(survival and growth)
0.0128-
>0.0128
0.0128
Outdoor mesocosm
Foguth et al.
(2020)
Northern leopard frog
(Gosner stage 26.5, 0.109
g),
Lithobates pipiens
R, U
10 d
PFOS-K
>98%
7.9
22
NOEC
(development, growth,
survival)
0.1->0.1
0.1
Duration too long for
an acute test and too
short for a chronic test
Brown et al.
(2021)
G-21
-------
Spocics (life's!;iiicI
Method'
Tcsl
Dui'iilioii
( hcinic;il /
Piiriu
pll
Temp.
(ฐC)
r.riw-i
Chronic
limits
(NOI'.C-
i.or.ci
(lllli/l.)
Reported
r.fl'ccl
(one.
(mป/l.)
Deficiencies
Reference
Northern leopard frog
(larva, Gosner stage 25),
Lithobates pipiens
S,M
30 d
PFOS
>96%
7.8
26.2
LOEC
(developmental stage)
<0.00006-
0.00006
0.00006
Duration too long for
an acute test and too
short for a chronic test
Flynnetal. (2021)
Leopard frog (larva,
Gosner stage 25),
Lithobates pipiens
R, M
30 d
PFOS
>98%
-
20
MATC
(developmental stage,
growth-weight)
0.1219-
1.437
0.4185
Lack of dose response;
PFOS present in
controls
Flynn et al. (2022)
Leopard frog (larva,
Gosner stage 25),
Lithobates pipiens
R, M
30 d
PFOS
>98%
-
20
LOEC
(scaled mass index)
-
0.00774
Lack of dose response;
PFOS present in
controls
Flynn et al. (2022)
Leopard frog (larva,
Gosner stage 25),
Lithobates pipiens
R, M
30 d
PFOS
>98%
-
20
NOEC
(mortality)
-
1.437
Lack of dose response;
PFOS present in
controls
Flynn et al. (2022)
Leopard frog (tadpole,
Gosner stage 25),
Lithobates pipiens
R, M
120 d
PFOS
>98%
7.87
(7.40-
8.25)
20.3
(19.4-
21.0)
NOEC
(mortality, growth)
-
>0.000934
Test represents a
greater than low value
(followed data rule)
Hoskins et al.
(2022)
Black spotted frog,
Pelophylax
nigromaculatus
(formerly, Rana
nigromaculata)
R, M
21 d
PFOS
>98%
6.5
20
LOEC
(biochemistry changes
and gene expression)
-
0.01
Non-apical endpoints
Lin et al. (2022a)
Black spotted frog (adult),
Pelophylax
nigromaculatus
R, M
21 d
PFOS-K
>98%
6.5
20
LOEC
(biochemistry changes
and gene expression)
-
0.0009150
Non-apical endpoints
Lin et al. (2022b)
Black spotted frog,
Pelophylax
nigromaculatus
R, M
21 d
PFOS-K
>98%
6.5
20
LOEC
(gene expression)
-
0.00121
Non-apical endpoints
Liu et al. (2023b)
Black spotted frog (adult),
Pelophylax
nigromaculatus
R, M
21 d
PFOS
>98%
6.5
20
LOEC
(gene expression)
-
0.01114
Non-apical endpoint
Shi et al. (2023)
Tiger salamander
(larva, 46 hr),
Ambystoma tigrinum
R, M
30 d
PFOS
>98%
-
20
NOEC
(growth, survival)
0.6213
Lack of dose response
Flynn et al. (2022)
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 PFOS in the range of concentrations tested under similar
conditions. Daily renewal of test solutions.
0 Water concentrations were not measured, but PFOS concentrations were measured in the liver.
G-22
-------
d 36 days corresponds to the first of ten generations, the one with the most consistent negative response. The value at 36 days is 1/10 of the duration of this year-long 10-generation
study.
e Reported in moles converted to gram based on a molecular weight of 500.13 g/mol (PFOS); 538.22 g/mol (PFOS-K); 629.4 g/mol (PFOS-TEA); 522. Ill g/mol (PFOS-Na).
G-23
-------
Appendix H Other Estuarine/Marine PFOS Toxicity Studies
H.l Summary Table of Acceptable Qualitative Estuarine/Marine PFOS Toxicity Studies
Speck's (lircshiiic)
Mel hod'
Test
Dui'iilidii
( hcmic;il /
PnriU
pll
Temp.
<ฐC)
S;ilini(\
(ppl)
ll'lccl
('limine
Limits
(NOI-.C-
i.or.ci
(lllli/l.)
Reported
I'.ITeel
(one.
C lllli/l.)
Deficiencies
Reference
Bacterium,
Vibrio fischeri
S,M
15 min
PFOS-K
98%
18
EC50
(bioluminescence
inhibition)
-
>500
Duration too short for a
plant test, missing some
exposure details, non-
apical endpoint
Rosal et al
(2010)
Cyanobacterium,
Anabaena sp.
S,M
24 hr
PFOS-K
98%
-
28
-
EC50
(bioluminescence
inhibition)
-
143.27
Duration too short for a
plant test, missing some
exposure details, non-
apical endpoint
Rosal et al.
(2010)
Dinoflagellate,
Pyrocystis lunula
S,M
24 hr
PFOS-K
98%
-
19
-
EC50
(bioluminescence
inhibition)
-
4.9
Duration too short for a
plant test
Hayman et
al. (2021)
Dinoflagellate,
Symbiodiniaceae
R,M
7 d
L-PFOS
Unreported
-
25
-
NOEC
(population
abundance)
0.0001-
>0.0001
0.0001
Only one exposure
concentration
Bednarz et
al. (2022)
Dinoflagellate,
Symbiodiniaceae
R,M
14 d
L-PFOS
Unreported
-
25
-
LOEC
(population
abundance)
<0.0001-
0.0001
0.0001
Only one exposure
concentration
Bednarz et
al. (2022)
Dinoflagellate,
Symbiodiniaceae
R, M
7 d
L-PFOS
Unreported
-
32
-
NOEC
(population
abundance)
0.0001-
>0.0001
0.0001
Only one exposure
concentration
Bednarz et
al. (2022)
Dinoflagellate,
Symbiodiniaceae
R, M
14 d
L-PFOS
Unreported
-
32
-
LOEC
(population
abundance)
<0.0001-
0.0001
0.0001
Only one exposure
concentration
Bednarz et
al. (2022)
Dinoflagellate,
Symbiodiniaceae
R, M
28 d
L-PFOS
Unreported
-
32
-
NOEC
(population
abundance)
0.0001-
>0.0001
0.0001
Only one exposure
concentration
Bednarz et
al. (2022)
Golden brown alga,
Isochrysis galbana
S,U
72 hr
PFOS
98%
-
20
-
EC50
(growth inhibition)
-
37.5
Duration too short for a
plant test
Mhadhbi et
al. (2012)
Alga,
Ceratoneis closterium
S,U
72 hr
PFOS-K
Unknown
-
-
33
NOEC
(growth)
4.16-
>4.16
4.16
Sediment and other PFAS
present in exposure
Simpson et
al. (2021)
H-l
-------
Species (li IVsl ;iLto t
Method'
Test
Dui'iiliun
( hcmic;il /
PuriU
nil
Temp.
<ฐC)
Siilinilt
ll'lccl
( limine
l.imils
iNOI'.C-
I.OIX )
imji/l.)
Reported
1. ITcd
(one.
(niii/l.)
Deficiencies
Reference
Diatom,
Skeletonema costatum
S, M
96 hr
PFOS-K
86.9%
00 00
o
20
-30
EC50
(cell density)
-
>3.20
Only one exposure
concentration
Desjardins
et al.
(2001a)
Sandworm (adult),
Perinereis wilsoni
R,M
7 d
PFOS-K
Unreported
8.1
17.1
36
NOEC
(survival)
0.000028-
>0.000028
0.000028
Only one exposure
concentration
Sakurai et
al. (2017)
Sea urchin (adult),
Glyptocidaris
crenularis
R, U
21 d + 7 d
observation
PFOS-K
98%
8.1
13
30
NOEC
(mortality)
1.0->1.0
1.0
Not a true ELS test
(started with adults);
missing exposure details
Ding et al.
(2015)
Sea urchin (adult),
Glyptocidaris
crenularis
R, U
21 d + 7d
observation
PFOS-K
98%
8.1
13
30
LOEC
(SOD activity)
<0.01-
0.01
0.01
Not a true ELS test
(started with adults);
missing exposure details;
atypical endpoint
Ding et al.
(2015)
Purple sea urchin
(fertilized eggs),
Paracentrotus lividus
s,u
48 hr
PFOS
98%
-
20
-
EC50
(growth inhibition)
-
20
Duration too short for an
acute test
Mhadhbi et
al. (2012)
Purple sea urchin
(sperm),
Paracentrotus lividus
s,u
65 min
PFOS
Unreported
7.69
22
-
NOEC
(reproduction - egg
fertilization)
-
0.0005
Duration too short, greater
than low value
Munari et
al. (2022)
Sea urchin (embryo),
Psammechinus
miliaris
R,Ub
(tissue)
72 hr
PFOS-K
>98%
8
19
31
EC50
(morphological
abnormality)
-
>0.3999ฐ
Interpolated endpoint;
missing some exposure
details
Anselmo et
al. (2011)
Sea urchin (embryo),
Psammechinus
miliaris
R,Ub
(tissue)
16 d
PFOS-K
>98%
8
19
31
NOEC
(morphological
abnormalities, hatch
success,
development)
0.3999-
>0.3999
0.3999ฐ
Duration too short for
chronic test and too long
for acute test
Anselmo et
al. (2011)
Sea urchin (larva),
Psammechinus
miliaris
S,U
85 min.
PFOS-K
>98%
-
19
-
IC50
(cellular efflux
pump inhibition)
-
1.399ฐ
Duration too short for a
chronic test and too long
for an acute test, atypical
endpoint
Anselmo et
al. (2012)
Eastern oyster
(33.8mm),
Crassostrea virginica
S, M
96 hr
PFOS-K
90.49%
7.5-
8.1
22
20-21
EC50
(shell deposition)
-
>3.0
Lack of replication;
atypical endpoint
Drottar and
Krueger
(2000i)
H-2
-------
Species (li IVsl ;iLto t
Method'1
Test
Dui'iiliun
( Ik-iii ic;il /
PuriU
pll
Temp.
<ฐC)
Siilinilt
Il'lccl
(hi-unic
l.imils
iNOI'.C-
I.OIX )
97%
7.5
24.9
20
LOEC
(cellular lysosomal
damage)
<3-3
3
Atypical endpoint
Aquilina-
Beck et al
(2020)
Mediterranean mussel
(6.4 cm),
Mytilus
galloprovincials
R, U
30 d
PFOS
Analytical
grade
8.1
17.5
34.5
LOEC
(increase
micronuclei nuclear
aberrations in gills
cells)
<2-2
2
Atypical endpoint; missing
some exposure details
Nalbantlar
and Arslan
(2017)
Green mussel (adult),
Perna viridis
R,M
7 d + 7 d
observation
PFOS-K
98%
-
25
30
EC50
(integrative
genotoxicity)
0.00095-
0.0097
0.033
Duration too short for a
chronic test and too long
for an acute test, atypical
endpoint
Liu et al.
(2014a)
Green mussel (adult),
Perna viridis
R, M
7 d
PFOS-K
98%
-
25
25
MATC
(CAT and SOD
activity)
0.106-
0.968
0.3203
Duration too short for a
chronic test and too long
for an acute test, atypical
endpoint
Liu et al.
(2014b)
Green mussel
(60-65 mm),
Perna viridis
R, M
7 d
PFOS-K
98%
-
25
25
MATC
(relative condition
factor)
0.0096-
0.106
0.0319
Duration too short for a
chronic test and too long
for an acute test, atypical
endpoint
(Liu et al.
2014c)
Green mussel,
Perna viridis
R, M
7 d + 7 d
observation
PFOS-K
98%
8
25
30
MATC
(hemocyte cell
viability)
0.0096-
0.106
0.0319
Duration too short for a
chronic test and too long
for an acute test, atypical
endpoint
Liu and
Gin (2018)
Green mussel,
Perna viridis
R, U
7 d
PFOS
>98%
-
25
3.2
MATC
(biochemistry
changes)
0.01-0.1
0.0316
Non-apical endpoint
Xu et al.
(2022)
White sunset shell
(15.0-20.3 mm).
Soletellina alba
S,M
28 d
PFOS-K
Unreported
8
19
30
NOEC
(survival)
0.85-
>0.85
0.85
Other PFAS measured in
the sediment and water
Simpson et
al. (2021)
Bivalve
(8.1-18.9 mm),
Tellina deltoidalis
S,M
28 d
PFOS-K
Unreported
8
19
30
MATC
(growth - weight)
0.22-0.28
0.2482
Other PFAS measured in
the sediment and water
Simpson et
al. (2021)
Smooth cauliflower
coral,
Stylophora pistillata
R, U
7 d
L-PFOS
Unreported
-
32
-
NOEC
(lipid peroxidation)
0.0001-
>0.0001
0.0001
Atypical endpoint
Bednarz et
al. (2022)
H-3
-------
Species (li IVsl ;iLto t
Mclhuri'1
losl
Dui'iiliun
( Ik-iii ic;il /
PuriU
pll
Temp.
<ฐC)
Siilinilt
IITecl
Chronic
l.imils
iNOI'.C-
I.OIX )
2.0
2.0
Other PFAS measured in
the sediment and water
Simpson et
al. (2021)
Copepod (adult),
Nitocra spinipes
S, M
28 d
PFOS-K
Unreported
8.1
21
30
NOEC
(survival)
0.48-
>0.48
0.48
Other PFAS measured in
the sediment and water
Simpson et
al. (2021)
Copepod (adult,
female),
Tigriopus japonicus
R, U
10 d
PFOS
Unreported
-
25
32
MATC
(reproduction)
0.1-0.25
0.1581
Difficult to determine test
methodology
Han et al.
(2015)
Amphipod (adult),
Gammarus
insensibilis
s,u
48 hr
PFOS-K
>98%
8
19
32.5
LC50
-
9.99
Duration too short for an
acute test
Touaylia et
al. (2019)
Amphipod (adult),
Melita plumulosa
S, M
10 d
PFOS-K
Unreported
-
21
30
EC10
(reproduction)
-
0.9
Other PFAS measured in
the sediment and water
Simpson et
al. (2021)
Smooth sentinel crab
(6-15 mm carapace),
Macrophthalmus sp.
S, M
28 d
PFOS-K
Unreported
8
19
30
NOEC
(survival)
0.85-
>0.85
0.85
Other PFAS measured in
the sediment and water
Simpson et
al. (2021)
Chinese mitten crab
(11.89 g),
Eriocheir sinensis
R, U
21 d
PFOS-K
>98%
1 ,
^ 00
18-22
0.3
MATC
(total hemocyte
count)
0.01-0.1
0.03162
Duration too short for a
chronic test and too long
for an acute test
Zhang et al.
(2015)
Mud crab (3 cm),
Macrophthalmus
japonicus
R, U
96 hr
PFOS
98%
-
20
30
LC50
-
>0.03
Only three exposure
concentrations, atypical
source of organisms
Park et al.
(2015)
H-4
-------
Species (li IVsl ;iLto t
Method'1
Test
Dui'iiliun
( Ik-iii ic;il /
PuriU
pll
Temp.
<ฐC)
Siilinilt
ll'lccl
(hi-unic
l.imils
iNOI'.C-
I.OIX )
16
16
Duration too short for a
chronic test and too long
for an acute test
Fang et al.
(2012)
Marine medaka
(embryo),
Oryzias melastigma
R, M
8 d
PFOS
98%
-
28
30
LOEC
(malformation)
<1-1
1
Duration too short for a
chronic test and too long
for an acute test
Fang et al.
(2012)
Marine medaka
(embryo),
Oryzias melastigma
R, U
<21 d
PFOS
98%
-
28
30
MATC
(increase hatching
rate, decrease
hatching time)
1.0-4
2.00
Duration too short for a
chronic test, low control
hatch success, only three
exposure concentrations
Wu et al.
(2012)
Marine medaka
(embryo),
Oryzias melastigma
R, U
<21 d + 7d
observation
PFOS
98%
-
28
30
MATC
(larval survival)
1.0-4
2.00
Duration too short for a
chronic test, low control
hatch success, only three
exposure concentrations
Wu et al.
(2012)
Atlantic Cod
(juvenile),
Gadus morhua
F,Ub
(tissue)
5 d
(1 hr/day)
PFOS
Technical
grade
7.7
10
33.8
NOEC
(survival, growth)
0.2->0.2
0.20
Duration too short for a
chronic test and too long
for an acute test, only two
exposure concentrations.
Pulsed exposure.
Preus-
Olsen et al.
(2014)
European seabass
(juvenile),
Dicentrarchus labrax
D, U
21 d
PFOS-K
>98%
-
20
28
LOEC
(histology,
enzymatic and
genetic changes)
-
4.83
lig/kg
Dietary exposure, non-
apical endpoints
Espinosa-
Ruiz et al.
(2023)
Blackrock fish
(5 mo. old),
Sebastes schlegelli
R, U
6 d
PFOS
99%
8.0-
8.2
8.0-12
10
NOEC
(survival, growth)
1->1
1
Duration too short for a
chronic test and too long
Jeon et al.
(2010)
H-5
-------
Speck's (lik-shi^c)
Mel hod'
Test
Dui'iiliun
( Ik-iii ic;il /
PuriU
pll
Temp.
<ฐC)
Siilinilt
(ppn
IITecl
Chronic
l.imils
iNOI'.C-
I.OIX )
1
1
Duration too short for a
chronic test and too long
for an acute test, only two
exposure concentrations
Jeon et al.
(2010)
Blackrock fish
(5 mo. old),
Sebastes schlegelli
R, U
6 d
PFOS
99%
8.0-
8.2
8.0-12
25
NOEC
(survival, growth)
1->1
1
Duration too short for a
chronic test and too long
for an acute test, only two
exposure concentrations
Jeon et al.
(2010)
Blackrock fish
(5 mo. old),
Sebastes schlegelli
R, U
6 d
PFOS
99%
8.0-
8.2
8.0-12
34
NOEC
(survival, growth)
1->1
1
Duration too short for a
chronic test and too long
for an acute test, only two
exposure concentrations
Jeon et al.
(2010)
Turbot (embryo),
Scophthalmus
maximus
(formerly, Psetta
maxima)
R, U
6 d
PFOS
98%
-
18
-
LC50
-
0.11
Duration too short for a
chronic test and too long
for an 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 Study did not measure water concentrations, but there are measured concentrations from analysis of the tissue of organisms.
0 Reported in moles converted to gram based on a molecular weight of 500.13 g/mol (PFOS); 538.22 g/mol (PFOS-K); 629.4 g/mol (PFOS-TEA).
H-6
-------
Appendix I Acute to Chronic Ratios
1.1 Acute to Chronic Ratios from Quantitatively Acceptable Toxicity Tests.
Species
( hemiciil /
Pu lit \
Acule
Method'
Chronic
Method'
Acule
Tesl
Diii'iilion
Chronic
Tesl
Diinilion
Chronic HITccl
Acule
r.lTcd
(one.
(niii/l.)
Chronic
r.lTecl
(one.
ACR'
SMACK'
Reference
Fatmucket,
Lampsilis siliquoidea
PFOS
>98%
S, M
S, M
24 hour
36 day
MATC
(metamorphosis
success)
16.5
0.01768
933.3
933.3
Hazelton (2013);
Hazelton et al.
(2012)
Snail,
Physella heterostropha
pomilia
(formerly, Physa
pomilia)
PFOS-K
>98%
S, M
R, M
96 hour
44 day
EC10
(clutch size)
161.8
8.527
18.98
18.98
Funkhouser (2014)
Rotifer,
Brachionus calyciflorus
PFOS
>98%
S,Ub
R,Ub
24 hour
Up to
158 hours
LOEC
(reduced net
reproductive rate)
61.8
0.25
247.2
>247.2
Zhang et al. (2013)
Cladoceran,
Daphnia carinata
PFOS-K
>98%
S,U
R, U
48 hour
21 day
MATC
(days to first brood)
11.56
0.003162
3,656
3,656b
Logeshwaran et al.
(2021)
Cladoceran,
Daphnia magna
PFOS-K
90.49%
S, M
R, M
48 hour
21 day
EC10
(cumulative young)
58.51
11.19
5.229
-
Drottar and
Krueger (2000c)
Cladoceran,
Daphnia magna
PFOS-K
95%
S,U
R, U
48 hour
21 day
EC10
(survival)
67.2
16.35
4.110
-
Boudreau et al.
(2003a)
Cladoceran,
Daphnia magna
PFOS
Unreported
S,U
R, U
48 hour
21 day
EC10
(# of young/brood)
35.46
1.051
33.74
-
Ji et al. (2008)
Cladoceran,
Daphnia magna
PFOS-K
>98%
S,U
R, U
48 hour
21 day
EC10
(total neonates/female)
63.84d
3.030
21.07
-
Li (2009); Li
(2010)
Cladoceran,
Daphnia magna
PFOS-K
99%
S, M
R, M
48 hour
21 day
EC10
(survival)
78.09
2.610
29.92
-
Yang et al. (2014)
Cladoceran,
Daphnia magna
PFOS
98%
S,U
R, U
48 hour
21 day
EC10
(number of
offspring/brood/female)
23.41
0.001818
12,877b
-
Lu et al. (2015)
Cladoceran,
Daphnia magna
PFOS-K
>98%
S,U
R, U
48 hour
21 day
EC10
(survival)
94.58
3.596
26.30
-
Liang et al. (2017)
Cladoceran,
Daphnia magna
PFOS-K
98%
S,U
R, U
48 hour
21 day
EC10
(growth-length)
22.43
0.9093
24.67
16.23
Yang et al. (2019)
1-1
-------
Species
( hcmiciil /
Pu ii( \
Acule
Method'1
Chronic
Method'
Acule
Tesl
Diimlion
Chronic
Tesl
Dni'iilion
Chronic I'.ITecl
Acule
I'.ITecl
(one.
98%
S, M
R, M
96 hour
28 day
LC20
59.87
0.167
358.5
358.5
Funkhouser (2014)
Mayfly,
Neocloeon triangulifer
PFOS-K
98%
R, M
R, M
96 hour
23 day
EC10
(dry weight at day 14)
0.07617
0.000226
337.0
337.0
Soucek et al.
(2023)
Zebrafish,
Danio rerio
PFOS-K
unknown/
PFOS 96%
R, U
R, U
96 hour
Life Cycle
EC10
(F1 offspring: %
survival)
17
0.01650
1,030
1,030
Wang et al. (2011);
Wang et al. (2013)
Fathead minnow,
Pimephales promelas
PFOS-K
90.49%
S, M
F, M
96 hour
47 day
EC10
(survival)
9.020
0.4732
19.06
19.06
Drottar and
Krueger (2000c)
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
b Value appears to be an outlier and is not used in SMACR calculation.
0 Values in bold are used in the SMACR and FACR calculations.
d Geometric mean of three LC50s.
1-2
-------
1.2 Acute to Chronic Ratios from Qualitatively Acceptable Toxicity Tests.
Species
Aciilc / Chronic
( hcmic;il ;iii(l
Puriu
Anile
Method'1
Chronic
Method'1
Acule
1 esl
Dui'iilion
Chronic
1 esl
Dui'iilion
Acule
r.lTccl
Chronic r.lTccl
Acule
r.lTccl
(one.
(lliu/l - >
Chronic
r.lTccl
(one.
ACR
References
Planaria,
Dugesia japonica
PFOS-K
>99%
R, U
R, U
96 hours
10 days
LC50
LOEC
(regeneration:
decreased appearance
of auricles)
29.46
0.5
58.92
Yuanetal. (2014)
Snail,
Lymnaea stagnalis
PFOS
Unreported
S, M
R, M
96 hours
21 days
LC50
MATC
(survival)
171.5
4.243
40.41
Olson (2017)
Midge,
Chironomus sp.
PFOS-K (99%) /
PFOS Unreported
S, M
S, M
96 hours
~36 days
(1st of 10
generations)
LC50
LOEC
(F1 developmental
time, adult weight,
exuvia length)
182.12
0.004
45,530
Marziali et al.
(2019); Yang et al.
(2014)
Midge,
Chironomus sp.
PFOS-K (99%) /
PFOS-K (95%)
S, M
S, M
96 hours
Life
cycle
(>50
days)
LC50
EC10
(total emergence)
182.12
0.05896
3,089
MacDonald et al.
(2004); Yang et al.
(2014)
Yellow fever mosquito,
Aedes aegypti
PFOS
Unreported
S,U
R, U
48 hours
~42 days
LC50
MATC
(average time to
emergence)
1.18
0.079
14.94
Olson (2017)
Rainbow trout,
Oncorhynchus mykiss
PFOS-K (98%) /
PFOS (89%)
R, M
s,u
96 hours
14 days
LC50
LOEC
(decrease LSI)
2.5
1.0
2.500
Oakes et al.
(2005); Sharpe et
al. (2010)
Zebrafish,
Danio rerio
PFOS
>97%
S,U
R, U
96 hours
6 days
LC50
EC50
(uninflated swim
bladder)
58.47
2.29
25.53
Hagenaars et al.
(2014); Hagenaars
et al. (2011b)
Zebrafish,
Danio rerio
PFOS-K
98%
R, M
R, M
96 hours
21 days
LC50
LOEC
(reduced fecundity)
22.2
0.5
44.40
Sharpe et al.
(2010)
Zebrafish,
Danio rerio
PFOS
98%
S,U
R, U
96 hours
70 days
LC50
MATC
(increased
malformation &
decreased survival of
F1 fish)
3.502
0.02236
156.6
Du et al. (2016a);
Du et al. (2009)
1-3
-------
Species
Acule / Chronic
( hemiciil iind
Piiriu
Acule
Method'
Chronic
Method'
Acule
Tesl
Dui'iilion
Chronic
1 esl
Dui'iilion
Acule
I'.ITecl
Chronic HITccl
Acule
I'.ITecl
(one.
(niii/l.)
(lironic
I'.ITecl
Cone,
(inii/l.)
ACR
Uererences
African clawed frog,
Xenopus laevis
PFOS-K (86.9%) /
PFOS-K (86.9%)
R, M
R, M
96 hours
96 hours
LC50
LOEC
(growth)
15.49
8.26
1.875
Palmer and
Krueger (2001)
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
1-4
-------
Appendix J Unused PFOS Toxicity Studies
J.l Summary Table of Unused PFOS Toxicity Studies
Aullior
Ciliilion
Kciisoii I iiiisod
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.
Force-fed (oral gavage); only one exposure
concentration
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
carboxylic acids. Aquat. Toxicol. 130-131: 219-230.
Only one exposure concentration
Balbi, T., C. Ciacci, E. Grasselli, A. Smerilli, A.
Voci and L. Canesi
2017. Utilization of Mytilus digestive gland cells for the invito screening of
potential metabolic disruptors in aquatic invertebrates. Comp. Biochem.
Physiol. PartC. 191:26-35.
In vitro (excised cells)
Bilbao, E., D. Raingeard, 0. Diaz de Cerio, M.
Ortiz-Zarragoitia, P. Ruiz, U. Izagirre, A. Orbea, I.
Marigomez, M.P. Cajaraville and I. Cancio
2010. Effects of exposure to Prestige-like heavy fuel oil and to
perfluorooctane sulfonate on conventional biomarkers and target gene
transcription in the thicklip grey mullet Chelon labrosus. Aquat. Toxicol. 98:
282-296.
Only one exposure concentration; the
number of fish was not reported
Blanc, M., A. Karrman, P. Kukucka, N. Scherbak
and S. Keiter
2017. Mixture-specific gene expression in zebrafish (Danio rerio) embryos
exposed to perfluorooctane sulfonic acid (PFOS), perfluorohexanoic acid
(PFHxA) and 3,3',4,4',5-pentachlorobiphenyl (PCB126). Sci. Total Environ.
590: 249-257.
Mixture (PFOS, PFHxA and PCB126)
Blanc, M., J. Ruegg, N. Scherbak and S.H. Keiter
2019. Environmental chemicals differentially affect epigenetic-related
mechanisms in the zebrafish liver (zf-1) cell line and in zebrafish embryos.
Aquat. Toxicol. 215:105272-9999.
Control absent from test
Chen, J., L. Zheng, L. Tian, N. Wang, L. Lei, Y.
Wang, Q. Dong, C. Huang and D. Yang
2018. Chronic PFOS exposure disrupts thyroid structure and function in
zebrafish. Bull. Environ. Contam. Toxicol. 101: 75-79.
Only one treatment concentration; severe
lack of procedural details
Chen, K., N. Iwasaki, X. Qiu, H. Xu, Y. Takai, K.
Tashiro, Y. Shimasaki and Y. Oshima
2020. Adipogenesis of perfluorooctanesulfonate (PFOS) on Japanese medaka
(Oryzias latipes) embryo using ovo-nanoinjection-mRNA seq analysis. J. Fac.
Agric. Kyushu Univ. 65(2): 295-303.
Injected toxicant in ova
Cheng, J., S. Lv, S. Nie, J. Liu, S. Tong, N. Kang, Y.
Xiao, 0- Dong, C. Huang and D. Yang
2016. Chronic perfluorooctane sulfonate (PFOS) exposure induces hepatic
steatosis in zebrafish. Aquat. Toxicol. 176: 45-52.
Only one exposure concentration;
unmeasured chronic exposure
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
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)
Dale, K., F. Yadetie, T. Horvli, X. Zhang, H.G.
Froysa, O.A. Karlsen and A. Goksoyr
2022. Single PFAS and PFAS mixtures affect nuclear receptor- and oxidative
stress-related pathways in precision-cut liver slices of Atlantic cod (Gadus
morhua). Sci. Total Environ. 814: 1-12.
Invito; no apical endpoints
J-l
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Author
( iliilion
Kciison I misod
Diaz de Cerio, 0., E. Bilbao, M.P. Cajaraville and I.
Cancio
2012. Regulation of xenobiotic transporter genes in liver and brain of juvenile
thicklip grey mullets (('heIon labrosus) after exposure to Prestige-like fuel oil
and to perfluorooctane sulfonate. Gene. 498: 50-58.
Only one exposure concentration
Dorts, J., P. Kestemont, P.A. Marchand, W.
D'Hollander, M.L. Thezenas, M. Raes and F.
Silvestre
2011. Ecotoxicoproteomics in gills of the sentinel fish species, Cottus gobio,
exposed to perfluorooctane sulfonate (PFOS). Aquat. Toxicol. 103: 1-8.
Only two exposure concentrations, not
North American species
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.
In vitro; no apical endpoints
Du, J., S. Wang, H. You and Z. Liu
2016b. Effects of ZnO nanoparticles on perfluorooctane sulfonate induced
thyroid-disrupting on zebrafish larvae. J. Environ. Sci. 47: 153-164.
Only 72-75% control survival in 14-day test
Du, J., J. Tang, S. Xu, J. Ge, Y. Dong, H. Li and M.
Jin
2018. Parental transfer of perfluorooctane sulfonate and ZnO nanoparticles
chronic co-exposure and inhibition of growth in F1 offspring. Regul. Toxicol.
Pharmacol. 98: 41-49.
Excessive control mortality in the F0
generation
Fang, C., Q. Huang, T. Ye, Y. Chen, L. Liu, M.
Kang, Y. Lin, H. Shen and S. Dong
2013. Embryonic exposure to PFOS induces immunosuppression in the fish
larvae of marine medaka. Ecotox. Environ. Saf. 92: 104-111.
Excessive control mortality (-60% control
survival)
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.
Mixture
Garoche, C., A. Boulahtouf, M. Grimaldi, B.
Chiavarina, L. Toporova, M.J. DenBroeder, 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
Gorroehategui, 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.
Hagenaars A., I.J. Meyer, D. Herzke, B.G. Pardo, P.
Martinez, M. Pabon, W. De Coen and D. Knapen
2011. The search for alternative aqueous film forming foams (AFFF) with a
low environmental impact: Physiological and transcriptomic effects of two
Forafacฎ fluorosurfactants in turbot. Aquat. Toxicol. 104: 168-176.
Only one exposure concentration; missing
detail (focus is on other chemicals)
Hoff, P.T., W. VanDongen, E.L. Esmans, R. Blust
and W.M. De Coen
2003. Evaluation of the toxicological effects of perfluorooctane sulfonic acid
in the common carp (Cyprinus carpio). Aquat. Toxicol. 62 (4): 349-359.
Exposure was from a single intra-peritoneal
injection
Hoff, P.T., K. Van Campenhout, K. Van de Vijver,
A. Covaci, L. Bervoets, L. Moens, G. Huyskens, G.
Goemans, C. Belpaire, R. Blust and W. De Coen
2005. Perfluorooctane sulfonic acid and organohalogen pollutants in liver of
three freshwater fish species in Flanders (Belgium): relationships with
biochemical and organismal effects. Environ. Pollut. 137: 324-333.
Field exposure, but concentrations were not
measured so no BAFs could be calculated
Honda, M., A. Muta, T. Akasaka, Y. Inoue, Y.
Shimasaki, K. Kanna, N. Okino and Y. Oshima
2014. Identification of perfluorooctane sulfonate binding protein in the plasma
of tiger pufferfish Takifugu rubripes. Ecotox. Environ. Safety. 104: 409-413.
Injected toxicant; only one exposure
concentration
Honda, M., A. Muta, A. Shimazaki, T. Akasaka, M.
Yoshikuni, Y. Shimasaki and Y. Oshima
2018. High concentrations of perfluorooctane sulfonate in mucus of tiger
pufferfish Takifugu rubripes: a laboratory exposure study. Environ. Sci.
Pollut. Res. 25: 1551-1558.
Injected toxicant
Huang, T.S., P.A. Olsvik, A. Krovel, H.S. Tung and
B.E. Torstensen
2009. Stress-induced expression of protein disulfide isomerase associated 3
(PDIA3) in Atlantic salmon (Salmo salar L.). Comp. Biochem. Physiol. Part B
Biochem. Mol. Biol. 154(4): 435-442.
In vitro (cultured hepatocytes)
J-2
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Author
( iliilion
Kciison I misod
Huang, Q., S. Dong, C. Fang, X. Wu, T. Ye and Y.
Lin
2012. Deep sequencing-based transcriptome profiling analysis of Oryzias
melastigma exposed to PFOS. Aquat. Toxicol. 120-12: 54-58.
Only one or two exposure concentrations
Huang, Q., Y. Chen, Y. Chi, Y. Lin, H. Zhang, C.
Fang and S. Dong
2015. Immunotoxic effects of perfluorooctane sulfonate and di(2-ethylhexyl)
phthalate on the marine fish Oryzias melastiema. Fish Shell. Immunol. 44:
302-306.
Only two exposure concentrations
Huang, J., Q. Wang, S. Liu, H. Lai and W. Tu.
2022. Comparative chronic toxicities of PFOS and its novel alternatives on the
immune system associated with intestinal microbiota dysbiosis in adult
zebrafish. J. Hazard. Mater. 425: lip.
Only on exposure concentration; lack of
apical endpoints
Jacobson, T., K. Holmstrom, G. Yang, A.T. Ford, U.
BergerandB. Sundelin
2010. Perfluorooctane sulfonate accumulation and parasite infestation in a
field population of the amphipod Monoporeia affinis after microcosm
exposure. Aquat. Toxicol. 98(1): 99-106.
Dilution water not characterized, mixture
Jantzen, C.E.
2016. Toxicological Profiles of Perfluoroctanoic Acid (PFOA),
Perfluorooctane Sulfonate (PFOS) and Perfluornonanoic Acid (PFNA) in
Zebrafish (Danio rerio). Ph.D. Thesis, Rutgers, The State University of New
Jersey, New Brunswick, NJ: 177 p.
Thesis publication; separate DERs were
completed for individual components of the
study
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.
Single concentration test where exposure to
PFOS was of an acute (117 hours) duration
but endpoints were measured at 6 months of
age.
Keiter S., K. Burkhardt-Medicke, P. Wellner, B.
Kais, H. Farber, D. Skutlarek, M. Engwall, T.
Braunbeck, S.H. Keiter and T. Luckenbach
2016. Does perfluorooctane sulfonate (PFOS) act as chemosensitizer in
zebrafish embryos? Sci. Total Environ. 548-549:317-324.
Mixture
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): lip.
In-vitro study
Kim, S., K. Ji, S. Lee, J. Lee, J. Kim, S. Kim, Y. Kho
and K. Choi
2011. Perfluorooctane sulfonic acid exposure increases cadmium toxicity in
early life stage of zebrafish, Danio rerio. Environ. Toxicol. Chem. 30(4): 870-
877.
Only one exposure concentration; atypical
duration (7 days)
Kovacevic, V., A.J. Simpson and M.J. Simpson
2018. Evaluation of Daphnia magna metabolic responses to organic
contaminant exposure with and without dissolved organic matter using 1H
nuclear magnetic resonance (NMR)-based metabolomics. Ecotoxicol. Environ.
Saf. 164:189-200.
Only one exposure concentration; test not
focused on the toxicological effects of
PFOS but on the effects of dissolved
organic matter following exposure to PFOS
and other contaminants
Kovacevic, V., A.J. Simpson and M.J. Simpson
2019. The concentration of dissolved organic matter impacts the metabolic
response in Daphnia magna exposed to 17a-ethynylestradiol and
perfluorooctane sulfonate. Ecotoxicol. Environ. Saf. 170: 468-478.
Only one treatment concentration
(examined across a gradient of dissolved
organic matter concentrations); endpoints
measured were a suite of metabolic
changes; atypical design for this test
organism
Krovel, A.V., L. Softeland, B. Torstensen and P.A.
Olsvik
2008. Transcriptional effects of PFOS in isolated hepatocytes from Atlantic
salmon Salmo salar L. Comp. Biochem. Physiol., Part C. 148: 14-22.
In vitro
J-3
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Author
( iliilion
Kciison I misod
Lee, W. and Y. Kagami
2010. Effects of perfluorooctanoic acid and perfluorooctane sulfonate on gene
expression profiles inmedaka (Oryzias latipes). Abstracts. Toxicol. Let. 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 was run three times at
separate times (not simultaneously) and the
sample size for each treatment group was
unclear; control mortality not reported
Li, Y., B. Men, Y. He, H. Xu, M. Liu and D. Wang
2017. Effect of single-wall carbon nanotubes on bioconcentration and toxicity
of perfluorooctane sulfonate in zebrafish (Danio rerio). Sci. Total Environ.
607-608: 509-518.
Bioaccumulation (steady state no
documented); only 4 days; static exposure
Li, R., T. Tang, W. Qiao and J. Huang
2020. Toxic effect of perfluorooctane sulfonate on plants in vertical-flow
constructed wetlands. J. Environ. Sci. 92: 176-186.
PFOS added to a simulated wastewater
(mixture) which was not properly
characterized
Liu, C., Y. Dua and B. Zhoua
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.
Excised cells (cultured hepatocytes)
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.
Bioaccumulation (steady state no
documented); only 12 days
Martin, J.W., S.A. Mabury, K.R. Solomon and
D.C.G. Muir
2003b. Dietary accumulation of perfluorinated acids in juvenile rainbow trout
(Oncorhynchus mykiss). Environ. Toxicol. Chem. 22(1): 189-195.
Mixture
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.
Review paper
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 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
Mylroie, J.E., M.S. Wilbanks, A.N. Kimble, K.T. To,
C.S. Cox, S.J. Mcleod, KA. Gust, D.W. Moore, E.J.
Perkins and N. Garcia-Reyero
2021. Perfluorooctanesulfonic acid induced toxicity on zebrafish embryos in
the presence or absence of the chorion. Environ. Toxicol. Chem. 40(3): 780-
791.
Use of dilution medium (estradiol media) to
prepare stock solutions inconsistent with
EPA test guidelines
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.
Only one exposure concentration, no
concentration-response observed, not North
American species
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; no apical endpoints
Pablos, M.V., P. Garcia-Hortiguela and C. Fernandez
2015. Acute and chronic toxicity of emerging contaminants, alone or in
combination, in Chlorella vulgaris and Daphnia magna. Environ. Sci. Pollut.
Res. 22: 5417-5424.
Mixture
J-4
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Author
( iliilion
Kciison I misod
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)
Roland, K., P. Kestemont, L. Henuset, M.A.
Pierrard, M. Raes, M. Dieu and F. Silvestre
2013. Proteomic responses of peripheral blood mononuclear cells in the
European eel (Anguilla anguilla) after perfluorooctane sulfonate exposure.
Aquat. Toxicol. 128/129: 43-52.
In vitro (excised cells)
Shi, X., Y. Du, P.K.S. Lam, R.S.S. Wu and B. Zhou
2008. Developmental toxicity and alteration of gene expression in zebrafish
embryos exposed to PFOS. Toxicol. Appl. Pharmacol. 230(1): 23-32.
Excessive control mortality
Shi, X., L.W.Y. Yeung, P.K.S. Lam, R.S.S. Wu and
B. Zhou
2009b. Protein profiles in zebrafish (Danio rerio) embryos exposed to
perfluorooctane sulfonate. Toxicol. Sci. 110(2): 334-340.
Only one exposure concentration; atypical
duration (8 days)
Stanic, B., J. Petrovic, B. Basica, S. Kaisarevic, K.
Schirmer and N. Andric
2021. Characterization of the ERK1/2 phosphorylation profile in human and
fish liver cells upon exposure to chemicals of environmental concern. Environ.
Toxicol. Pharmacol. 88: 9 p.
In vitro
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 (mussel gill tissue)
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
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; atypical
duration (3 days)
Qiu, X., N. Iwasaki, K. Chen, Y. Shimasaki and Y.
Oshima
2019. Tributyltin and perfluorooctane sulfonate play a synergistic role in
promoting excess fat accumulation in Japanese medaka (Oryzias latipes) via in
ovo exposure. Chemosphere. 220: 687-695.
Injected toxicant into eggs, not North
American species
Wagner, N.D., A.J. Simpson and M.J. Simpson
2016. Metabolomic responses to sublethal contaminant exposure in neonate
and adult Daphnia magna. Environ. Toxicol. Chem. 36(4): 938-946.
Only one exposure concentration
Wagner, N.D., A.J. Simpson and M.J. Simpson
2018. Sublethal metabolic responses to contaminant mixture toxicity in
Daphnia magna. Environ. Toxicol. Chem. 37(9): 2448-2457.
Only one exposure concentration
Wang, S., C. Zhuang, J. Du, C. Wu and H. You
2017. The presence of MWCNTs reduces developmental toxicity of PFOS in
early life stage of zebrafish. Environ. Pollut. 222: 201-209.
The 96-hour LC50 reported in the
publication is the same as the value in Du et
al. 2016 (no details provided about this test)
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.
Mixture (PFCs mixed in sediment)
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Author
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Kciison I misod
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.
Bioaccumulation (steady state not
documented); only 3 days; test was
unmeasured
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); only 3 days; test was
unmeasured
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); only 3 days; test was
unmeasured
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
Zhang, L., Y.Y. Li, T. Chen, W. Xia, Y. Zhou, Y.J.
Wan, Z.Q. Lv, G.Q. Li and S.Q. Xu
201 la. Abnormal development of motor neurons in perfluorooctane
sulphonate exposed zebrafish embryos. Ecotoxicol. 20: 643-652.
Static, unmeasured exposure to single-
concentration (1 mg/L) from 6 hours post-
fertilization to 120 days post-fertilization
Zhang, L., Y.Y. Li, H.C. Zeng, J. Wei, Y.J. Wan, J.
Chen and S.Q- Xu
201 lb. MicroRNA expression changes during zebrafish development induced
by perfluorooctane sulfonate. J. Appl. Toxicol. 31: 210-222.
Poor control survival (>80% at 24 hour and
increasing)
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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) 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 were
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.1 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.1.2.2 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
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
-------
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. 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-3
-------
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
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
K-4
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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.
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 scheme was developed to aid in evaluating models as to
their quantitative or qualitative attributes in a transparent and relatively repeatable manner.
K-5
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Appendix L Derivation of Acute Protective PFOS Benchmarks for
Estuarine/Marine Waters through a New Approach Method
(NAM): WeblCE
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 five of the eight family MDRs (the
dataset was missing another family in the Phylum Chordata, a family in a phylum other than
Chordata, and any other family); thus, the EPA was not able to derive an acute estuarine/marine
criterion element for PFOS based on the 1985 Guidelines MDR specifications (Section 3.2.1.2).
However, the EPA was able to develop an acute PFOS protective benchmark for aquatic life
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 MDR 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 fulfill the missing MDRs and allow the
derivation of an acute 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 PFOS 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 for aquatic life, and the EPA's intention to reduce the use of animal testing via
L-l
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application of NAMs (https://www.epa.gov/chemical-research/epa-new-approach-methods-
work-plan-reducing-use-animals-chemical -testing).
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 to 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, www3.epa.gov/webice, 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
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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; Dyer et al. 2008; Raimondo et al. 2010;
Raimondo and Barron 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 provide a level of protection that is consistent with using measured
laboratory data (Dyer et al. 2008; Feng et al. 2013; Fojut et al. 2012; Palumbo et al. 2012; Wang
et al. 2020; Wu et al. 2015; Wu et al. 2016; 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
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5
c
_g
4-
2
-1 -
0 2 4
6
ฃ
Rainbow trout (log LC50)
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 MSE(<~ 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-4
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L.2 Application of Web-ICE with PFOS
ICE models are developed using a diversity of compounds (e.g., metals and other
inorganics, pesticides, solvents, and reactive chemicals) across a wide range of toxicity values;
however, PFAS are not included in web-ICE v3.3 due to the lack of available PFAS toxicity data
when web-ICE v3.3 was created. 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., enter and 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.2.1 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 (v3.3) and as described in detail in Raimondo et al.
L-5
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(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 upper subset containing paired chemical data above the 75th percentile of surrogate values.
The Raimondo et al. (2024) lower subset was used to develop "truncated" ICE models. The
surrogate values in the upper subset were converted to mg/L and entered into the truncated ICE
model. 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 [j,g/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 5-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
L-6
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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 [^g/L 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)4.
L.2.2 Direct Comparison of Web-ICE and Measured Toxicity Values
Since limited PFOS toxicity test data are available for estuarine/marine species, the
ability of ICE models to predict PFOS 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 species mean acute values (SMAVs) for PFOS reported in
Appendix A.l and Appendix B.l were used as values for surrogate species to predict all possible
species that also had a measured PFOS SMAV reported. The available SMAVs for PFOS that
could be used as ICE surrogate values along with the number of ICE models (i.e., potential
predicted species) corresponding to each surrogate are shown in Table L-l.
4 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 three
of 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 PFOS estuarine/marine benchmark was 0.55 mg/L
(see section L.2.4). 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 0.55 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 0.57 mg/L. While both approaches contain uncertainty, use of the scaled approach resulted in a
more protective acute PFOS estuarine/marine benchmark (i.e., CMC = 0.55 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 = 0.57 mg/L).
L-7
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Table L-l. Surrogate Species Measured Values for PFOS and Corresponding Number of
ICE Models for Each Surrogate.
For example, there are 53 species for uliidi P.ij'hni.i m
-------
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 5-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).
These comparisons are consistent with web-ICE user guidance (Raimondo et al. 2015),
previously published reports on ICE model accuracy (Raimondo et al. 2010; Willming et al.
2016), and the above presented uncertainty analysis of using scaled toxicity as model input. ICE
models predict with acceptable accuracy for PFOS 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 5-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.2.1 and L.2.2 demonstrate that the relationship of
inherent sensitivity represented by ICE models is preserved across taxa, chemicals, and range of
toxicity values when using robust ICE models. While the current analysis uses freshwater species
L-9
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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-10
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Table L-2. Comparison of ICE-predicted and measured values of PFOS 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
To\icil\ Yiilucs Polcnlhillv licvond Model Doniiiin
Sc;ded 1 o\icil\ \ ;dnes
Prediclcd Species
Siiitoป;Kc Species
Measured
SM.W
(iili/l.l
weh-K 1.
Prcdiclcd
(Jiti/I.)
Confidence
lnler\:ds (nji/l.)
l-old
Difference
Measured
SM.W
(nili/l.)
\\cl>-l( T.
Predicted
Confidence
lnlcr\ :d
l-old
Difference
Bullfrog
(Lithobates
catesbeianus)
Daphnid
(Daphnia magna)
133,300
59755.54
12281.24-
290746.24
2.23
133.3
63.35
7.50 -534.65
2.1a
Fathead minnow
(Pimephales promelas)
8356.68
3748.61 - 18629.28
15.95
13.26
4.13-42.57
10.05
Rainbow trout
(Oncorhynchus mykiss)
15140.53
8139.36 -28163.81
8.8
33.9
13.63 -84.26
3.93
African clawed frog
(Xenopus laevis)
Fathead minnow
(Pimephales promelas)
15,990
7034.49
800.65 -61804.35
2.27a
15.99
18.93
0.306- 1170.65
1.18ab
Daphnid
(Daphnia magna)
4,914
9221.68
5220.28 - 16290.18
1.88
4.914
28.55
19.59-41.60
5.81
Mysid
(Americamysis bahia)
Fathead minnow
(Pimephales promelas)
359.91
135.34 - 957.15
13.65ฐ
0.481
0.104-2.21
10.22ฐ
Rainbow trout
(Oncorhynchus mykiss)
1172.37
702.88 - 1955.47
4.19ฐ
2.01
1.08-3.75
2.44ฐ
Bullfrog
(Lithobates catesbeianus)
51,860
81946.04
17394.84 -
386042.67
1.58
51.86
199.47
32.95 - 1207.24
3.85
Fathead minnow
(Pimephales promelas)
1697.85
1149.29 -2508.22
30.54ฐ
3.29
1.36-7.96
15.76ฐ
Daphnid
Fatmucket
(Lampsilis siliquoidea)
23122.84
7634.81 -70030.01
2.24
7.73
1.46-40.85
6.71b
(Daphnia magna)
Mysid
(Americamysis bahia)
6096.75
3829.31 - 9706.79
8.51
21.29
13.73 - 33.02
2.44
Rainbow trout
(Oncorhynchus mykiss)
2775.45
2007.74 - 3836.72
18.69ฐ
8.83
5.26 - 14.80
5.87ฐ
Zebrafish embryo
(Danio rerio)
4515.51
1042.06 - 19566.76
11.48ฐ
2.85
0.171-47.42
18.2abฐ
Bullfrog
(Lithobates catesbeianus)
7,515
82395.25
38247.48 -
177501.32
10.96
7.515
39.73
16.29-96.91
5.29
Rainbow trout
(Oncorhynchus
mykiss)
Daphnid
(Daphnia magna)
22196.99
15080.46 - 32671.85
2.95ฐ
245.99
182.01 -332.46
32.73ฐd
Fathead minnow
(Pimephales promelas)
2771.13
2136.90 - 3593.60
2.71
3.43
2.01-5.85
2.19
Fatmucket (Lampsilis
siliquoidea)
48028.61
3264.96 -706515.68
6.39aฐ
13.14
1.03 - 167.64
1 75abฐ
L-ll
-------
To\icil\ Yiilucs Polcnlhillv licvond Model Doiii;iin
Scaled 1 o\icil\ \ iilucs
Prcdiclcd Species
Siiitoป;i1c Species
Measured
SM AY
(Mli/I.)
\\cl>-l( 1.
Prcdiclcd
(n ii/1.)
Confidence
1 nler\ills (nii/l.)
l-nld
Difference
Me;isiii-ed
SM.W
(mji/l.)
\\cl>-l< 1.
Predicled
(inu/l.)
Confidence
Inlen ;il
(inii/l.)
l-old
Difference
Mysid (Americamysis
bahia)
6169.68
3855.10 - 9873.91
1.22
68.63
45.85 - 102.73
9.13d
Zebrafish embryo
(Danio rerio)
11721
4212.88 -32610.00
1.56
3.46
0.618-19.40
2.17b
African clawed frog
(Xenopus laevis)
6,950
16080.14
1020.67 -253332.73
2.3 la
6.95
7.89
0.071 - 868.40
1.14ab
Bullfrog (Lithobates
catesbeianus)
121541.51
44334.20 -
333204.08
17.49
91.08
33.84 -245.08
13.11
Fathead minnow
(Pimephales
promelas)
Daphnid {Daphnia
magna)
46651.96
30060.61 -72400.58
6.71ฐ
712.85
474.15 -
1071.71
102.57ฐd
Fatmucket (,Lampsilis
siliquoidea)
116669.9
19477.15 -
698863.06
16.79ฐ
595.88
48.52 -7317.09
85.74abฐ
Mysid (Americamysis
bahia)
13672.93
5348.62 - 34952.76
1.97ฐ
254.88
118.25 -549.38
36.67ฐd
Rainbow trout
(Oncorhynchus mykiss)
14424.97
11028.30 - 18867.80
2.08
36.58
23.89 -56.02
5.26
Zebrafish embryo
(Danio rerio)
31446.57
17390.46 - 56863.77
4.52
56.87
17.21 - 187.92
8.18b
Black sandshell
(Ligumia recta)
16,500
11412.52
2418.09 -53863.02
1.45
16.5
8.15
0.319 -208.15
2.02ab
Fatmucket
(Lampsilis
siliquoidea)
Daphnid (Daphnia
magna)
23821.82
9341.74 -60746.62
1.44
138.32
48.08 - 397.96
8.38
Fathead minnow
(Pimephales promelas)
717.52
149.82 -3436.35
23ฐ
3.21
0.065 - 158.39
5 14abฐ
Rainbow trout
{Oncorhynchus mykiss)
1585.37
485.38 -5178.24
10.41ฐ
44.11
9.18-211.95
2.67ฐd
Black sandshell
(Ligumia recta)
Fatmucket (Lampsilis
siliquoidea)
13,500
19191.22
4438.79 - 82973.68
1.42
13.5
26.59
1.49-472.22
1 97^
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).
L-12
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L.2.3 Prediction of Estuarine/Marine Species Sensitivity to PFOS
A value of PFOS 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
were rejected in this first pass. In the next step, values that were predicted using [j,g/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
L-13
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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.
Saltwater MDRs:
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
The acute sensitivity of estuarine/marine species to PFOS is presented in Table L-4. A
total of 36 models representing 19 estuarine/marine species were available in web-ICE to predict
the toxicity of PFOS to saltwater species (Table L-3). Of these, 12 models were initially rejected
based on model parameters not meeting the guidance in Box 1, reducing the number of predicted
species to 17 represented by 24 models. Further evaluation of ICE predictions resulted in 12
SMAVs. The range of sensitivity for the predicted taxa is consistent with the range of sensitivity
of freshwater species for this compound.
L-14
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Table L-3. All ICE Models Available in web-ICE v3.3 for Saltwater Predicted Species Based on Surrogates with Measured PFOS.
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 =
SuiTiiisilr
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liTur
\ nine
Miixiiiiiiin
Si kll'NN
1 jMiiiciiiiii
I'mlklril Spi'iir-.
SuiTiiisilr S|iriir-
Mn|ir
lllUTir|il
<\-2>
K-
|l-\ ฆline*
(MSI.)
(II!! 1 )
\ nine (II!! 1 )
("..)
l)hl;MHi-
1 m- in ( rilrrhi
Acartia tonsa
Daphnia magna
0.59
1.31
2
0.91
0.0443
0.17
2.24
38514.70
50
5
Rejected
Allorchestes compressa
Daphnia magna
0.83
1.59
3
0.8
0.039
0.12
5.00
184.54
100
5
Accepted
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
<0.001
0.93
0.07
840000.00
64
5
Accepted
Americamysis bahia
Oncorhynchus mykiss
0.92
-0.5
150
0.6
<0.001
1.08
0.06
1100000.00
57
6
Rejected
Americamysis bahia
Pimephales promelas
0.95
-1.12
46
0.55
<0.001
1.75
2.27
70200000.00
35
6
Rejected
Chelon labrosus
Lampsilis siliquoidea
1.27
1.5
1
0.99
0.0403
0
19.01
281.00
NA
6
Accepted qualitatively
Chelon macrolepis
Pimephales promelas
1.51
-1.04
2
0.97
0.0114
0.05
26.00
2533.38
100
4
Accepted qualitatively
Crassostrea virginica
Americamysis bahia
0.44
1.76
114
0.34
<0.001
0.88
0.003
117648.20
55
6
Rejected
Crassostrea virginica
Daphnia magna
0.44
1.54
116
0.28
<0.001
1.08
0.08
137171.43
58
6
Rejected
Crassostrea virginica
Lampsilis siliquoidea
0.82
-0.28
3
0.95
0.0041
0.06
30.00
22000.00
100
4
Accepted
Crassostrea virginica
Oncorhynchus mykiss
0.59
0.97
120
0.5
<0.001
0.68
0.02
570000.00
68
6
Rejected
Crassostrea virginica
Pimephales promelas
0.75
0.44
24
0.61
<0.001
0.68
1.24
206300.75
69
6
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
<0.001
0.67
0.003
182000.00
64
6
Rejected
Cyprinodon variegatus
Daphnia magna
0.53
1.79
84
0.49
<0.001
0.72
0.08
304000.00
64
6
Rejected
Cyprinodon variegatus
Lampsilis siliquoidea
0.72
0.76
1
0.99
0.0392
0
30.00
22000.00
NA
6
Accepted qualitatively
Cyprinodon variegatus
Oncorhynchus mykiss
0.75
0.9
87
0.65
<0.001
0.56
0.82
12700000.00
75
4
Accepted
Cyprinodon variegatus
Pimephales promelas
0.69
0.98
24
0.74
<0.001
0.43
2.27
16500000.00
77
4
Accepted
Farfantepenaeus duorarum
Americamysis bahia
1.03
0.06
6
0.81
0.0022
0.55
0.01
720.00
50
4
Accepted
Farfantepenaeus duorarum
Daphnia magna
1.08
0.14
16
0.76
<0.001
1.32
0.04
65686.02
44
5
Rejected
Farfantepenaeus duorarum
Oncorhynchus mykiss
1.2
-1.36
15
0.72
<0.001
1.54
0.57
221000.00
47
6
Rejected
Fenneropenaeus merguiensis
Daphnia magna
0.82
1.43
4
0.66
0.0473
0.4
5.00
1251.41
67
5
Accepted
Gasterosteus aculeatus
Oncorhynchus mykiss
1.05
0.29
4
0.9
0.0038
0.18
0.61
890.00
83
4
Accepted
Hydroides elegans
Daphnia magna
0.49
1.59
2
0.96
0.0182
0.01
5.00
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
Litopenaeus stylirostris
Americamysis bahia
1.04
0.01
5
0.6
0.0401
0.29
0.58
24.09
57
4
Accepted
Menidia menidia
Oncorhynchus mykiss
1.28
-1.4
3
0.94
0.005
0.23
11.24
91000.00
60
4
Accepted qualitatively
Menidia peninsulae
Americamysis bahia
0.63
0.91
3
0.88
0.0162
0.32
0.01
1160.00
80
6
Accepted qualitatively
Menidia peninsulae
Oncorhynchus mykiss
1.01
-0.36
2
0.91
0.0421
0.35
0.82
1600.00
50
4
Accepted qualitatively
L-15
-------
rmliilril S|HTir-.
Slll l iiliiili- S|iriir-
Mn|ir
lllUTir|il
1 h-siivr*
Ml
livriliilll
(\-2)
K-
|l-\ ฆline*
Menu
Sl|ll;liv
I'.i rur
(MSI.)
S||ITii!i;ilr
Mnilrl
Minimum
\ uliir
(M!i 1 )
SlIITii-iiilr
Mi ii Ul
Miixiiiiiiin
\ lllllc (MU 1 )
( riisN
\ iiliiLiliim
Si K\l'NN
1 jMiiiciiiiii
Divlmur
1 M- in ( rilcrhi
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
Oncorhynchus mykiss
1.44
-0.37
3
0.89
0.0144
0.12
0.82
29.18
100
4
Accepted qualitatively
Tigriopus japonicus
Pimephales promelas
0.81
1.12
5
0.76
0.0103
0.11
195.14
27000.00
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
NA = Not Available.
L-16
-------
Table L-4. ICE-Estimated Species Sensitivity to PFOS.
Values in bold and underlined are used for SMAV.
C 0111111011 Nsimc
Scientific
SiiiToป:ilc
Input
I nit
Ksl i 111:1 led
Toxicity (mji/l.)
05% Confidence
Intenills
SMAV
Calanoid copepod
Acartia tonsa
Daphnia magna
Ug/L
13.12abc
(0.66 - 259.64)
NA
Amphipod
Allorchestes compressa
Daphnia magna
mg/L
1072.28
(323.49 - 3554.23)
50.94
Pimephales promelas
Ug/L
2.42
(1.29 -4.54)
Mysid
Americamysis bahia
Daphnia magna
Ug/L
9.22
(5.22 - 16.29)
9.22
Oncorhynchus mykiss
Ug/L
1.17ฐ
(0.7 - 1.96)
Pimephales promelas
Ug/L
0.36ฐ
(0.14-0.96)
Thicklip mullet
Chelon labrosus
Lampsilis siliquoidea
mg/L
1144.93ab
(126.12 - 10393.7)
NA
Bigscale mullet
Chelon macrolepis
Pimephales promelas
Ug/L
61.79ab
(4.94-772.16)
NA
Eastern oyster
Crassostrea virginica
Americamysis bahia
Ug/L
2.52ฐ
(1.45 -4.37)
1.89
Daphnia magna
Ug/L
4.31ฐ
(2.02 - 9.2)
Lampsilis siliquoidea
Ug/L
1.56
(0.44 -5.55)
Oncorhynchus mykiss
Ug/L
2.01ฐ
(1.3-3.1)
Pimephales promelas
Ug/L
2.28
(0.78 - 6.67)
Leon springs pupfish
Cyprinodon bovinus
Oncorhynchus mykiss
mg/L
21.5T
(3.2 -236.94)
1.82
Pimephales promelas
Ug/L
1.82
(0.78 - 4.24)
Sheepshead minnow
Cyprinodon variegatus
Americamysis bahia
Ug/L
9.87ฐ
(5.58 - 17.46)
5.77
Daphnia magna
Ug/L
20.32ฐ
(9.75 - 42.39)
Lampsilis siliquoidea
Ug/L
6.76a
(0.56 - 81.92)
Oncorhynchus mykiss
Ug/L
7.08
(4.53 - 11.06)
Pimephales promelas
Ug/L
4.7
(2.32 -9.52)
Pink shrimp
Farfantepenaeus duorarum
Americamysis bahia
mg/L
6.02
(1.34 -26.97)
6.02
Daphnia magna
Ug/L
173.22ac
(14.83 -2023.16)
Oncorhynchus mykiss
Ug/L
2.12ฐ
(0.38- 11.71)
Banana prawn
Fenneropenaeus merguiensis
Daphnia magna
mg/L
722.81
(131.83 - 3963)
722.81
Threespine stickleback
Gasterosteus aculeatus
Oncorhynchus mykiss
mg/L
16.46
(5.22 - 51.84)
16.46
Polychaete
Hydroides elegans
Daphnia magna
Ug/L
8.45bc
(1.31 - 54.56)
NA
Oncorhynchus mykiss
Ug/L
1.28ฐ
(0.89 - 1.83)
Blue shrimp
Litopenaeus stylirostris
Americamysis bahia
mg/L
5.41
(1.59- 18.41)
5.41
Atlantic silverside
Menidia menidia
Oncorhynchus mykiss
Ug/L
3.9T
(0.52 - 30.32)
NA
Tidewater silverside
Menidia peninsulae
Americamysis bahia
mg/L
22.65d
(3.47 - 147.72)
NA
L-17
-------
Common \:ime
Scientific
Siiitoป;iIc
Input
I nit
Kstiniiitcd
Toxicity (mป/l.)
05% Confidence
In ton :ils (mป/l.)
S\1 AY
Oncorhynchus mykiss
mg/L
3.35a
(0.1 - 118.6)
Mysid
Metamysidopsis insularis
Daphnia magna
mg/L
258.03
(48.24 - 1380.1)
156.17
Lampsilis siliquoidea
|ig/L
94.52
(27.87 - 320.53)
Striped mullet
Mugil cephalus
Oncorhynchus mykiss
mg/L
7.66d
(2.17-27.01)
NA
Harpacticoid copepod
Tigriopus japonicus
Pimephales promelas
Ug/L
18.04
(7.2 - 45.24)
18.04
Harpacticoid copepod
Tisbe battagliai
Daphnia magna
mg/L
550.44
(107.35 -2822.37)
550.44
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-18
-------
L.2.4 Derivation of Acute Water Quality Benchmark for Estuarine/Marine Water
The web-ICE predicted acute dataset for PFOS contains 15 genera, representing the eight
MDR groups that would be necessary for developing an estuarine/marine criterion. The EPA
fulfilled these eight MDRs by integrating the acceptable quantitative study data (discussed in
Section 3.1.1.2) with data derived using web-ICE to support calculating a protective benchmark.
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 criteria values for PFOS. GMAVs for the four most sensitive genera were within a
factor of 1.7 of each other (Table L-6). The estuarine/marine FAV (the 5th percentile of the genus
sensitivity distribution) for PFOS is 1.096 mg/L (Table L-6). The FAV is lower than all of the
GMAVs for both the tested species and for values derived using web-ICE. The FAV was then
divided by two to obtain a concentration yielding a minimal effects acute benchmark. The
FAV/2, which is the estuarine/marine acute water column benchmark magnitude, is 0.55 mg/L
PFOS (rounded to two significant figures) and is expected to be protective of 95% of
estuarine/marine genera potentially exposed to PFOS 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). This acute benchmark for estuarine/marine aquatic life is greater than the
recommended acute freshwater criterion (0.071 mg/L), suggesting that estuarine/marine species
may be less acutely sensitive to PFOS and emphasizing the importance of having a separate
benchmark value for the protection of estuarine/marine aquatic life.
L-19
-------
Table L-5. Ranked Estuarine/Marine Genus Mean Acute Values.
MDR
(roup
Nsinu'
Species (lil'cslauc)
SM.W
(iM.W
Rank
Pcrcenlile
D
Mediterranean
mussel
Mytilus galloprovincialis
1.1
1.1
1
0.06
F
Purple sea urchin
Strongylocentrotus purpuratus
1.7
1.7
2
0.13
E
Sea urchin
Paracentrotus lividus
1.795
1.795
3
0.19
D
Eastern oyster
Crassostrea virginica
1.89
1.89
4
0.25
C
Mysid
Americamysis bahia
4.914
4.914
5
0.31
A
Leon springs pupfish
Cyprinodon bovinus
1.82
5.225
6
0.38
Sheepshead minnow
Cyprinodon variegatus
>15
F
Blue shrimp
Litopenaeus stylirostris
5.41
5.41
7
0.44
F
Pink shrimp
Farfantepenaeus duorarum
6.02
6.02
8
0.50
C
Mysid
Siriella armata
6.9
6.9
9
0.56
B
Threespine stickleback
Gasterosteus aculeatus
16.46
16.46
10
0.63
G
Harpacticoid copepod
Tigriopus japonicus
18.04
18.04
11
0.69
E
Amphipod
Allorchestes compressa
50.94
50.94
12
0.75
C
Mysid
Metamysidopsis insularis
156.2
156.2
13
0.81
H
Harpacticoid copepod
Tisbe battagliai
550.4
550.4
14
0.88
F
Banana prawn
Fenneropenaeus merguiensis
722.8
722.8
15
0.94
MDR Groups
a. Family in the phylum Chordata
b. Family in the phylum Chordata
c. Either the Mysidae or Panaeidae 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-20
-------
Table L-6. Estuarine/Marine Final Acute Value and Protective Aquatic Acute Benchmark.
Bold values represent genera for which empirical toxicity data were available.
Calculated Estuarine/Marine FAY based on 4 lowest values; n=15
Rank
Genus
GMAV
(mg/L)
ln(GMAV)
ln(GMAV)2
P=R/(N+1)
sqrt(P)
1
Mytilus
1.1
0.10
0.01
0.063
0.250
2
Strongylocentrotus
1.7
0.53
0.28
0.125
0.354
3
Paracentrotus
1.795
0.59
0.34
0.188
0.433
4
Crassostrea
1.89
0.64
0.41
0.250
0.500
E (Sum):
1.85
1.04
0.63
1.54
S2 =
L =
A =
FAV =
PVAL=
5.32 S = slope
-0.424 L = X-axis intercept
0.092 A = InFAV
1.096 P = cumulative probability
0.55 mg/L PFOS (rounded to two significant figures)
0.90
s
2 0.80
| 0.70
sa v
a
oi
Q
s 0.40 4
a
o
u
ji 0.30
0.10
0.00
0.1
Femieropenaeus ~
~ Tisbe
-ฆ
~ Metamysidopsis
~ Allorchestes
-ฆ
~ Tigriopus
O Gasterosteus
ฆ Siriella
~ Farfantepenaeus
~ Litopenaeus
Fish (Empirical and WeblCE)
# Cyprinodon
O Fish (WeblCE)
-ฆ
ฆ Americamysis
ฆ Invertebrate (Empirical)
A Crassostrea
~ Invertebrate (WeblCE)
ฆ Paracentrotus
~ Mollusk (Empirical)
ฆ Strongylocentrotus
A Mollusk (WeblCE)
~ Mytilus
Acute Benchmark
i iii
i i i i i i i i i i i i i i i i
i i i i i i I i i i i i i i i I
1 10 100
Genus Mean Acute Value (mg/L PFOS)
1000
Figure L-2. Ranked Estuarine/Marine Acute PFOS GMAVs used for the Aquatic Life
Acute Benchmark Calculation.
L-21
-------
L.2.5 Estuarine Marine/Benchmark Uncertainty
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 estimates from the
available and quantitatively acceptable models (see bolded and underlined values in Table L-4; n
=16), 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 92.23% to 530.04%). The ICE model with the lowest 95% CI range relative to the LC50
estimate (i.e., 92.23%) employed Oncorhynchus mykiss 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., 530.04%) employed Daphnia magna as the predictor species and
Fenneropenaeus merguiensis as the predicted species. Fifteen of the 16 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 PFOS
estuarine/marine benchmark.
Six of the 15 GMAVs used to derive the acute PFOS estuarine/marine benchmark were
based on empirical toxicity tests. The six GMAVs based on empirical data were not evenly
distributed across the GSD, with all empirical data falling below the 60th percentile of sensitivity
(Table L-2). Also, three of the four most sensitive GMAVs in the GSD (Figure L-2) were based
on empirical data and five of the six most sensitive GMAVs were based empirical acute values,
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.
L-22
-------
The estuarine/marine benchmark appears adequately protective based on the available
high quality empirical data (Appendix B.l). The acute PFOS estuarine/marine benchmark (i.e.,
0.55 mg/L) is two times lower than the lowest GMAV (i.e., 1.1 mg/L), which was based on
empirical data for Mytilus. 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 PFOS for a continuous exposure 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., 1.096 mg/L). The EPA identified three individual tests in Table H.l as meeting the
previous criteria:
1. Park et al. (2015) conducted a seven-day test with the mud crab, Macrophthalmus
japonicus. Exposures lasted seven days, but survival was also recorded at 96 hours. The
authors did not calculate an LCso, but at 96 hours there was 36% mortality in the highest
test concentration (i.e., 0.03 mg/L). Therefore, the 96-hour LCso was >0.03 mg/L. The
test was not used quantitatively because an LCso could not be calculated based on the
three exposure concentrations used. Overall, 36% mortality after 96 hours in the 0.03
mg/L treatment suggests this species may be sensitive to acute PFOS exposures relative
to the acute estuarine/marine benchmark. However, the source of the organisms (fish
market) could be problematic as there is no mention of potential previous exposure or
measures of PFOS in test organisms at any point during the experiment.
L-23
-------
2. Mhadhbi et al. (2012) conducted a 6-day test with the turbot, Schophthalmus maximus.
Endpoints included dead embryos, malformation, hatch success at 48 hours and larval
survival (missing heartbeat and a non-detached tail) at six days. The 6-day LCso of 0.11
mg/L PFOS was not acceptable for acute benchmark derivation because of the relatively
long exposure duration. Nevertheless, the 6-day LCso is nearly an order of magnitude
lower than the acute estuarine/marine benchmark final acute value (i.e., 1.096 mg/L) and
five times lower than the acute estuarine/marine benchmark, suggesting S. maximus is
sensitive to acute PFOS exposures at concentrations below the acute estuarine/marine
benchmark.
3. Jeon et al. (2010) performed a 6-day test on blackrock fish, Sebastes schlegeli. There
were no significant differences in total length, weight and survival (no mortality observed
in any of the exposures) over the 6 - day exposure. The NOEC (survival and growth) was
1 mg/L at each test salinity (10, 17.5, 25 and 34 ppt), which is less than the acute
estuarine/marine benchmark final acute value (i.e., 1.096 mg/L). The lack of effects
observed at 1.0 mg/L preclude this test from providing meaningful information about the
protectiveness of the acute estuarine/marine benchmark.
Results from Mhadhbi et al. (2012), which was determined to only be acceptable for
qualitative use, suggests S. maximus is sensitive to acute PFOS exposures at concentrations
below the acute estuarine/marine benchmark, but at an exposure duration that was 50% longer
than the standard 96- hour exposure duration from quantitatively acceptable tests. Additionally,
results of quantitatively acceptable empirical toxicity studies with estuarine/marine organisms do
not provide any evidence that the aquatic estuarine/marine community will experience
unacceptable acute effects at the acute estuarine/marine PFOS benchmark.
L-24
-------
L.2.6 ICE Regressions Supporting the Acute Estuarine/Marine Benchmark
Americamysis bahia
(Log LC50)
Figure L-3. Americamysis bahia (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
Li"
2 o
O
ฆฃ d.
o
d
O
*
*
*
*
~ s Ml
*
' * * *
*
Americamysis bahia
(Log LC50)
Figure L-4. Americamysis bahia (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
L-25
-------
CO
_CD
Oj
E _
P ฐ
o
o
a a
CD
E
to
CM
Americamysis bahia
[Log LC50)
Figure L-5. Americamysis bahia (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
CD
|s
ro _JI
d CD
o
~. _l
03 '
O
Danio rerio- embryo
(Log LC50)
Figure L-6. Danio rerio -embryo (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
L-26
-------
Danio rerio- embryo
[Log LC50)
Figure L-7. 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-8. Danio rerio - embryo (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
L-27
-------
Daphnia magna
(Log LC50)
Figure L-9. Daphnia magna (X-axis) and Amerieamysis bahia (Y-axis) regression model
used for ICE predicted values.
Daphnia magna
(Log LC50)
Figure L-10. Daphnia magna (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values.
L-28
-------
0
2
4
6
Daphnia magna
(Log LC50)
Figure L-ll. Daphnia magna (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values.
-2 0 2 4 6
Daphnia magna
(Log LC50)
Figure L-12. Daphnia magna (X-axis) and Oncorhynchus mykiss (Y-axis) regression model
used for ICE predicted values.
L-29
-------
0 2 4 6
Daphnia magna
(Log LC50)
Figure L-13. Daphnia magna (X-axis) and Pintephalespromelas (Y-axis) regression model
used for ICE predicted values.
2 3 4 5 6
Lampsilis siliquoidea
(Log LC50)
Figure L-14. Lampsilis siliquoidea (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
L-30
-------
23 4
5
6
Lampsilis siliquoidea
(Log LC50)
Figure L-15. 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-16. Lampsilis siliquoidea (X-axis) and Oncorhynchits mykiss (Y-axis) regression
model used for ICE predicted values.
L-31
-------
2 3 4 5 6
Lampsilis siliquoidea
(Log LC50)
Figure L-17. Lampsilis siliquoidea (X-axis) and Pimephalespromelas (Y-axis) regression
model used for ICE predicted values.
2 3 4 5 6
Ligumia recta
(Log LC50)
Figure L-18. Ligumia recta (X-axis) and Lampsilis siliquoidea (Y-axis) regression model
used for ICE predicted values.
L-32
-------
1
4
Lithobates catesbeianus
(Log LC50)
Figure L-19. Lithobates catesbeianus (X-axis) and Daphnia magna (Y-axis) regression
model used for ICE predicted values.
O
co Id"
5 Q
a
O
Lithobates catesbeianus
(Log LC50)
Figure L-20. Lithobates catesbeianus (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
L-33
-------
Lithobates catesbeianus
(Log LC50)
Figure L-21. Lithobates catesbeianus (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
CD
ra
-Ci
ป S 1
W5 Q
f 1
m ^
Oncorhynchus mykiss
(Log LC50)
Figure L-22. Oncorhynchus mykiss (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values.
L-34
-------
0 2 4 6
Oncorhynchus mykiss
(Log LC50)
Figure L-23. Oncorhynchus mykiss (X-axis) and Daphnia magna (Y-axis) regression model
used for ICE predicted values.
1 2 3 4 5
Oncorhynchus mykiss
(Log LC50)
Figure L-24. Oncorhynchus mykiss (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values.
L-35
-------
0
2
4
6
Oncorhynchus mykiss
(Log LC50)
Figure L-25. 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-26. Oncorhynchus mykiss (X-axis) and Pimephales promelas (Y-axis) regression
model used for ICE predicted values.
L-36
-------
Pimephales prornelas
(Log LC50)
Figure L-27. Pimephales prornelas (X-axis) and Americamysis bahia (Y-axis) regression
model used for ICE predicted values.
0 2 4 6
Pimephales prornelas
(Log LC50)
Figure L-28. Pimephales prornelas (X-axis) and Daphn'ui magna (Y-axis) regression model
used for ICE predicted values.
L-37
-------
3 4 5 6 7
Pimephales prornelas
(Log LC50)
Figure L-29. Pimephales promelas (X-axis) and Lampsilis siliquoidea (Y-axis) regression
model used for ICE predicted values.
Pimephales prornelas
(Log LC50)
Figure L-30. Pimephales promelas (X-axis) and Lithobates catesbeianus (Y-axis) regression
model used for ICE predicted values.
L-38
-------
Pimephales prornelas
(Log LC50)
Figure L-31. Pimephales promelas (X-axis) and Oncorhynchus mykiss (Y-axis) regression
model used for ICE predicted values.
2 3 4 5
Pimephales promelas
(Log LC50)
Figure L-32. Pimephales promelas (X-axis) and Xenopus laevis (Y-axis) regression model
used for ICE predicted values.
L-39
-------
2
3
4-
5
Xenopus laevis
[Log LC50)
Figure L-33. Xenopus laevis (X-axis) and Pimephales promelas (Y-axis) regression model
used for ICE predicted values.
L-40
-------
Appendix M Environmental Fate of PFOS in the Aquatic Environment
Natural degradation of PFOS has not been observed. As described above in Section 2.2
above, under environmental conditions, PFOS does not photolyze, hydrolyze, or biodegrade and
is thermally stable. For these reasons, PFOS is considered to be highly persistent in the
environment (Beach et al. 2006; OECD 2002).
M.1 Photolysis
PFOS does not appear to photolyze (OECD 2002). No experimental evidence of direct or
indirect photolysis was available (Hatfield 2001). The indirect photolytic half-life of PFOS using
an iron oxide photo-initiator matrix model was estimated to be > 3.7 years at 25ฐC. This half-life
was based on the analytical method of detection (Giesy et al. 2010).
M.2 Hydrolysis
No hydrolytic loss of PFOS was observed in a 49 day study under experimental
conditions of 50ฐC and pH conditions of 1.5, 5, 7, 9 or 11 (Hatfield 2001). Instead, the half-life
of PFOS was estimated to be >41 years at 25ฐC. However, this estimate was influenced by the
analytical limit of quantification and that no loss of PFOS was actually detected (Giesy et al.
2010).
M.3 Biodegradation
Several studies have demonstrated that PFOS does not biodegrade under aerobic or
anaerobic conditions (Gledhill and Markley 2000c; Gledhill and Markley 2000b; Gledhill and
Markley 2000a; Key et al. 1998; Laboratory 2002; Lange 2001; Remde and Debus 1996; Saez et
al. 2008). Results from a study conducted by Kurume Laboratory in 2002 showed no
biodegradation of PFOS after 28 days as measured by net oxygen demand, loss of total organic
carbon, and loss of parent material. Key et al. (1998) demonstrated that even under sulfur-
ic-1
-------
limiting conditions, PFOS did not degrade. Similarly, Saez et al. (2008) observed no PFOS
degradation under aerobic or anaerobic conditions in municipal sewage sludge. In contrast,
Schroder (2003) reported that PFOS was anaerobically degraded; however, the reported results
are uncertain as the results could likely be attributed to sorption and there was a lack of increased
fluoride concentrations reported (Fromel and Knepper 2010).
The persistence of PFOS has been attributed to the strong C-F bond. Additionally, there
have been limited indications that naturally occurring, defluorinating enzymes exist that can
break a C-F bond, which is likely due to the rarity of fluorinated molecules in nature (Fromel and
Knepper 2010). To date, no laboratory data exist that demonstrates the PFOS undergoes
significant biodegradation in environmental conditions (Beach et al. 2006; Giesy et al. 2010;
OECD 2002).
M.4 Thermal Stability
Based on carbon-sulfur (C-S) bond energy, which is weaker than the carbon-carbon (C-
C) or the C-F bond energies, PFOS is considered to have relatively low thermal stability. Thus,
PFOS would more easily breakdown under incineration conditions and would be nearly
completely destroyed when incinerated (Beach et al. 2006; Giesy et al. 2010).
M.5 Adsorption/Desorption
In general, PFOS may adsorb to sediments (with a Kd greater than 1 mL/g; (Giesy et al.
2010)). However, this sorption to sediment is limited since PFOS has a Koc of 2.57, indicating
that PFOS is relatively mobile in water and the physicochemical characteristics of the sediment
ultimately influence the sorption of PFOS (Ahrens et al. 201 lb; Beach et al. 2006; Giesy et al.
2010; Higgins and Luthy 2006). Sediment characteristics have a strong influence on the
partitioning of PFOS (You et al. 2010). Specifically, organic content was found to have a
M-2
-------
significant influence on the partitioning of PFOS. Density of the sediment was also found to be
an important factor influencing partitioning (Ahrens et al. 201 lb). PFOS has a high affinity to
bind to organic carbon with log Koc values ranging between 2.57 and 3.8 cm3/g ((Higgins and
Luthy 2006) and (Ahrens et al. 2010); respectively). A sorption mechanism could be a salting-
out and calcium-bridging effect, as PFOS sorption to sediment increased with increased salinity,
pH, and calcium (You et al. 2010). Thus, the sorption of PFOS is a complicated process that is
partially dependent on other factors such as metal anion concentrations, pH, temperature, and
salinity; however, the strong relationship between PFOS concentrations and organic carbon in
soil, sediment, and sludge indicates that these other factors have a minor influence on PFOS
sorption (Ahrens et al. 201 lb; Chen et al. 2012; Higgins and Luthy 2006; You et al. 2010).
M-3
-------
Appendix N Occurrence of PFOS in Abiotic Media
N.l Summary of Measured Perfluorooctane Sulfonate Concentrations in
Surface Waters Across the United States.
Modified from: Jarvis et al. (2021).
Slsilc
\Y silcr hotly'
AriMimclic
Mesin PI OS
(oiuvnlnilion
(iiซ/l.)2
\lce of
PI OS
C oiuvnlnilion
(n a/i.)
Reference
Lake Erie
3.77
3
2.8 - 5.5
Sinclair et al.
(2006)
31.3
32.5
21.5-38.5
Boul anger et
al. (2004)
2.84
2.63
2.49-3.41
De Silva et
al. (2011)
4.5
4.2
4.0 - 5.3
Furdui et al.
(2008)
Lake Huron
2.25
1.96
0.239- 5.46
De Silva et
al. (2011)
1.73
1.5
1.2-2.7
Furdui et al.
(2007)
Lake
Michigan
2.03
2.03
0.93-3.13
Simcik and
Dorweiler
(2005)
2.00
1.96
1.73-2.36
De Silva et
al. (2011)
Lake Ontario
not provided
4.9
2.9-30
Sinclair et al.
(2006)
55.4
59.8
16.5-85.5
Boul anger et
al. (2004)
5.96
5.63
2.60-9.48
De Silva et
al. (2011)
8.69
6.6
3.6-37.6
Furdui et al.
(2008)
2.20
not provided
not provided
Houde 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
near Decatur
58,016
41,027
9- 150,000
OECD
(2002)
N-l
-------
AriMiim-nc
K;nii>e of
\lc:in PI OS
Medinn PI OS
PI OS
(oiuvnlrnlion
( oiicciili'iilion
( oiuvnlnilion
Slsito
\\ silerhodv1
(nii/l.):
(iiji/l.):
(nii/l.)
UclomuT
Waterbody in
Decatur
2.5< x <25
2.5< x <25
2.5< x <25
3M Company
Pond in
Decatur
111
111
111
(2001)
Waterbody in
Mobile
30.3
35.5
<25-41.5
3M Company
Pond in
Mobile
32.5
32.5
32.5
(2001)
Tennessee
River
(upstream of
Baker's
30.85
29.80
16.0-52.6
Hansen et al.
(2002)
Creek)
Tennessee
River
Hansen et al.
(2002)
(downstream
of Baker's
103.9
107.0
30.3 - 144
Creek)
California
Upper Silver
Creek
not provided
not provided
27-56
Plumlee et al.
(2008)
Coyote Creek
not provided
not provided
4.8-25
Animas River
<0.48
<0.48
<0.48
Arkansas
River
1.96
0.62
0.23 - 5.00
Arvada
Blunn
0.77
0.77
0.77
Reservoir
Barker
Reservoir
<0.49
<0.49
<0.49
Colorado
Bessemer
Ditch
14.0
14.0
14.0
CDPHE
Big
Thompson
River
3.90
3.90
3.90
(2020)
Blue River
1.20
1.20
1.20
Boulder
Feeder Canal
<0.45
<0.45
<0.45
Boyd Lake
1.00
1.00
1.00
Cache la
Poudre River
5.61
5.61
<0.45 - 11.0
Clear Creek
7.95
7.95
7.20 - 8.70
N-2
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
(nii/l.):
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
K;nii>e of
PI OS
( oiuvnlnilion
(nii/l.)
UclomuT
Colorado
River
0.67
0.66
0.65-0.69
Coon Creek
<0.48
<0.48
<0.48
Eagle River
0.68
0.68
0.68
East Plum
Creek
<0.43
<0.43
<0.43
Erie Lake
3.70
3.70
3.70
Fairmount
Reservoir
<2.50
<2.50
<2.50
Fountain
Creek
16.9
20.0
3.50 -24.0
Fraser River
1.00
1.00
1.00
Gore Creek
0.98
0.98
0.98
Gunnison
River
0.71
0.71
0.71
Horsetooth
Reservoir
0.51
0.51
0.51
Jackson
Creek
<0.44
<0.44
<0.44
Jerry Creek
<0.485
<0.485
<0.48 -
<0.49
Kannah
Creek
Flowline
<0.49
<0.49
<0.49
Lakewood
Reservoir
<0.45
<0.45
<0.45
Little
Fountain
Creek
<0.46
<0.46
<0.46
Maple Grove
Reservoir
10.0
10.0
10.0
Marstron
Reservoir
0.48
0.48
0.48
McBroom
Ditch
4.90
4.90
4.90
Mclellen
Reservoir
1.30
1.30
1.30
Mesa Creek
<0.49
<0.49
<0.49
Michigan
River
<0.46
<0.46
<0.46
N-3
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
K;nii>e of
PI OS
( oiuvnlnilion
(nii/l.)
UclomuT
Molina
Power Plant
Tail
<0.50
<0.50
<0.50
North Fork
Gunnison
River
<0.47
<0.47
<0.47
Purdy Mesa
Flowline
<0.49
<0.49
<0.49
Purgatoire
River
0.47
0.47
0.47
Ralston
Reservoir
<0.46
<0.46
<0.46
Rio Grande
<0.47
<0.47
<0.47
Roaring Fork
River
<0.50
<0.50
<0.50
San Juan
River
<0.44
<0.44
<0.44
Sand Creek
30.3
30.3
6.50 - 54.0
Severy Creek
<0.47
<0.47
<0.47
Somerville
Flowline
<0.48
<0.48
<0.48
South
Boulder
Creek
0.50
0.50
0.50
South Platte
River
10.5
11.5
3.80 - 16.0
St. Vrain
River
3.90
3.90
3.90
Strontia
Springs
<0.51
<0.51
<0.51
Taylor River
<0.45
<0.45
<0.45
Uncompahgr
e River
(delta)
0.54
0.54
0.54
Wei ton
Reservoir
2.60
2.60
2.60
White River
<0.46
<0.46
<0.46
Yampa River
<0.47
<0.47
<0.47
Delaware,
New Jersey,
Pennsylvania
Delaware
River
3.98
3.5
0.97 - 6.92
Pan et al.
(2018)
N-4
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
(nii/l.):
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
K;nii>e of
PI OS
( oiuvnlnilion
(nii/l-)
UclomuT
Florida
Waterbody in
Pensacola
16.29
2.5< x <25
<25 - 29
3M Company
(2001)
Pond in
Pensacola
2.5< x <25
2.5< x <25
2.5< x <25
Waterbody in
Port St. Lucie
50.83
2.5< x <25
<2.5-137.5
Small pond
in Port St.
Lucie3
9,784
1,945
1,830 -
48,200
Sarasota Bay
0.90
not provided
not provided
Houde et al.
(2006a)
Georgia
Waterbody in
Columbus
59.9
55
44.6-80
3M Company
(2001)
Pond in
Columbus
<2.5
<2.5
<2.5
Conasauga
River
162.1
192
<1.5-321
Konwick et
al. (2008)
Altamaha
River
2.63
2.6
2.6-2.7
Streams and
ponds in
Dalton
70.36
70.73
10.5-119.5
Oostanaula
River
150.3
151
148 - 152
Lasier et al.
(2011)
Louisiana
Waterbodies
(locations of
concern) near
Barksdale
A.F.B.
776.7
195.0
<10-7,070
Cochran
(2015);
Lanza et al.
(2017)
Reference
waterbodies
near
Barksdale
A.F.B.
<10
<10
<10
Michigan
Raisin River
3.5
3.5
3.5
Kannan et al.
(2005)
St Clair River
2.6
2
1.9-3.9
Si ski wit Lake
0.283
0.283
0.277-0.289
Scott et al.
(2010)
Minnesota
Upper
Mississippi
River
528.9
<2
<2-18,200
Newsted et
al. (2017)
N-5
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
(nii/l.):
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
K;nii>e of
PI OS
( oiuvnlnilion
(nii/l.)
UclomuT
Lake of the
Isles
2.47
2.47
2.47
Simcik and
Dorweiler
(2005)
Lake
Calhoun
50.4
50.4
50.4
Lake Harriet
22.1
22.1
22.1
Minnesota
River
9.21
9.21
9.21
Lake
Tettegouche
0.23
0.23
0.23
Lake
Nipisiquit
<0.27
<0.27
<0.27
Lake Loiten
<0.27
<0.27
<0.27
Little Trout
Lake
1.2
1.2
1.2
New Jersey
Echo Lake
Reservoir
<2
<2
<2
NJDEP
(2019)
Passaic River
13.1
13.1
13.0-13.2
Raritan River
6.9
6.9
6.9
Metedeconk
River
1.65
1.65
<2-2.8
Pine Lake
102
102
102
Horicon Lake
10
10
10
Little Pine
Lake
100
100
100
Mirror Lake
72.9
72.9
72.9
Woodbury
Creek
6.4
6.4
6.4
Fen wick
Creek
3.1
3.1
3.1
Cohansey
River
<2
<2
<2
Harbortown
Road
1.93
1.93
1.93
Zhang et al.
(2016)
Passaic River
4.59
4.07
0.244-9.99
New Mexico
Alamogordo
Domestic
Water Sys.
<
<1
<1
NMED
(2020)
Animas River
0.799
0.625
<0.89- 1.5
Canadian
River
0.848
0.9
<0.89- 1.2
N-6
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
K;nii>e of
PI OS
( oiuvnlnilion
(nii/l.)
UclomuT
Cloud
Country
Estates WUA
<0.93
<0.93
<0.93
Gila River
<0.93
<0.93
<0.93
Holloman
AFB Golf
Course Pond
1
1,220
1,220
1,220
Holloman
AFB Golf
Course Pond
2
878
878
878
Holloman
AFB Lagoon
G
310
310
310
Holloman
AFB Outfall
951
951
951
Holloman
AFB Sewage
Lagoon
2,200
2,200
2,200
Karr Canyon
Estates
<0.93
<0.93
<0.93
La Luz
MDWCA
<1.3
<1.3
<1.3
Lake
Holloman
4,033
4,500
1,700 - 5,900
Mountain
Orchard
MDWCA
<0.93
<0.93
<0.93
Pecos River
1.223
1.50
<0.94- 1.70
Rio Chama
<0.98
<0.98
<0.96 -<1
Rio Grande
1.052
0.474
<0.465 - 2.90
Rio Puerco
4.35
4.35
3.10 - 5.60
San Juan
River
<1.15
<1.15
<1.06-
<1.24
Tularosa
Water
System
0.723
0.723
<0.89 - 1.0
New York
Washington
Park Lake
1.67
1.77
<0.25-2.88
Kim and
Kannan
(2007)
Rensselaer
Lake
7.11
6.58
5.85-9.3
N-7
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
K;nii>e of
PI OS
( oiuvnlnilion
(nii/l.)
UclomuT
Iroquois Lake
not provided
not provided
not provided
Unnamed
lake 1 outside
Albany, NY
not provided
not provided
not provided
Unnamed
lake 2 outside
Albany, NY
not provided
not provided
not provided
Niagara
River
5.17
5.5
3.3-6.7
Finger Lakes
not provided
1.6
1.3-2.6
Lake
Onondaga
681
756
198- 1,090
Sinclair et al.
Lake Oneida
3.5
3.5
3.5
(2006)
Erie Canal
8.37
6.4
5.7 - 13
Hudson River
not provided
1.7
1.5-3.4
Lake
Champlain
not provided
2.7
0.8-7.7
Lower NY
Harbor
0.755
0.755
0.755
Zhang et al.
(2016)
Staten Island
1.66
1.66
1.66
Hudson River
1.81
1.81
0.79-2.84
North
Carolina
Cape Fear
River
31.2
28.9
<1 - 132
Nakayama et
al. (2007)
Narragansett
Bay
2.2
2.2
2.2
Benskin et al.
(2012)
Allen Cove
Inflow
1.20
1.20
1.20
Bristol
Harbor
0.508
0.46
0.437-0.626
Brook at Mill
Cove
9.80
9.80
9.80
Rhode Island
Buckeye
Brook
4.13
4.13
4.13
Zhang et al.
(2016)
Chickasheen
Brook
<0.05
<0.05
<0.05
EG Town
Dock
0.735
0.735
0.735
Fall River
0.238
0.238
0.238
Green Falls
River
0.291
0.291
0.29-0.292
Hunt River
1.48
1.48
1.48
N-8
-------
Slsito
\\ silerhodv1
AriMiim-nc
\lc:in PI OS
(oiuvnlrnlion
(nii/l.):
Medinn PI OS
( oiicciili'iilion
(iiji/l.):
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PI OS
( oiuvnlnilion
(nii/l.)
UclomuT
Mill Brook
3.94
3.94
3.94
Narrow River
0.298
0.264
0.176-0.488
Pawcatuck
River
0.561
0.561
0.509-0.612
Pawtuxet
River
2.19
2.19
2.19
Queens River
0.334
0.334
0.334
Sand Hill
Brook
1.82
1.82
1.82
Secret Lake -
Oak Hill
Brook
<0.05
<0.05
<0.05
Slack's
Tributary
0.777
0.777
0.777
South Ferry
Road Pier
0.161
0.161
0.161
Southern
Creek
3.74
3.74
3.74
Woonasquatu
cket River
14.6
14.6
5.87-23.2
South
Carolina
Charleston
Harbor
12.0
not provided
not provided
Houde et al.
(2006a)
Tennessee
Waterbody
near
Cleveland
2.5
-------
1 Name of Waterbody Sampled for PFOS. Name or description of waterbody above is consistent with that provided
in cited reference.
Calculation of arithmetic mean and median includes lower of Vi LOD or '/? LOQ, depending on information
provided. See full occurrence table in Appendix N for waterbody-specific details.
3 Study authors conducted additional sampling of this waterbody but were unable to detect the initial high PFOS
concentrations in any of the additional samples.
4 Reported as ng/g by the study authors.
N.2 PFOS occurrence and concentrations in the Great Lakes region
The Great Lakes are among the most widely studied waterbodies in the U.S. for PFOS
occurrence. However, occurrence data are still relatively limited for this system. Comparisons
across the Great Lake system indicate PFOS concentrations are higher in Lakes Erie and Ontario,
ranging between 2.8 and 38.5 ng/L and 2.6 and 85.5 ng/L, respectively (Figure 2-3) (Boulanger
et al. 2004; De Silva et al. 2011; Furdui et al. 2008; Sinclair et al. 2006), compared to the more
northern Great Lakes. These northern Great Lakes (i.e., Lakes Huron, Michigan, and Superior)
have a maximum observed concentration of 5.46 ng/L, which was observed in Lake Huron
(Remucal 2019). However, current measured PFOS concentrations were not from sampling sites
around urbanized areas (such as Chicago and Detroit) and may not be representative of the
potential sources of PFOS related to these areas. The measured concentrations of PFOS in the
surface waters of Lakes Huron and Michigan range between 0.24 and 5.46 ng/L (De Silva et al.
2011; Furdui et al. 2008) and 0.93 and 3.13 ng/L (De Silva et al. 2011; Simcik and Dorweiler
2005), respectively. In contrast, measured PFOS concentrations observed in Lake Superior were
considerably lower and range between 0.074 and 0.996 ng/L (De Silva et al. 2011; Furdui et al.
2008; Scott et al. 2010). The higher PFOS concentrations in Lakes Erie and Ontario are likely
due to higher levels of industrial activities and urbanization around these lakes (Boulanger et al.
2004; Remucal 2019) and could also be associated with the sampling locations. A mass balance
constructed for Lake Ontario by Boulanger et al. (2004) indicated wastewater effluent was the
N-10
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major source of PFOS to the lake. In contrast, inputs from Canadian tributaries and atmospheric
deposition of PFOS, and other PFAS that may be transformed into PFOS, were the major
contributing sources of PFOS to Lake Superior. Inputs from Canadian tributaries and
atmospheric deposition were estimated to contribute 57 and 32% of PFOS inputs into Lake
Superior, respectively (Scott et al. 2010).
N.3 PFOS occurrence and concentrations in the southeastern U.S.
Measured PFOS concentrations in southeastern U.S. surface waters were similar to those
measured in Lakes Erie and Ontario, with some of the highest observed concentrations occurring
in waterbodies near areas with PFOS manufacturing. In 2001, the 3M Company conducted a
multi-city study measuring PFOS concentrations across waterbodies with known manufacturing
and/or industrial uses of PFOS (3M Company 2001). In the 3M Company's 2001 report, PFOS
concentrations from sites with known PFOS discharges were compared to PFOS concentrations
measured in waterbodies with no known sources of any PFAS (3M Company 2001). In this
comparison study, cities with known PFOS exposure were Mobile and Decatur, Alabama,
Columbus, Georgia, and Pensacola, Florida. Measured PFOS concentrations ranged from not
detected (reported detection limit of 2.5 ng/L; 3M Company 2001) to 80 ng/L in the cities with
known PFOS discharges. These PFOS concentrations were compared to those measured in
control cities. These control cities were Cleveland, Tennessee and Port St. Lucie, Florida and
PFOS concentrations ranged from not detected to 137.5 ng/L (3M Company 2001). The PFOS
concentrations measured in Cleveland, Tennessee were below the limit of quantification (25
ng/L) and were lower than the PFOS concentrations observed in the cities with known PFOS
exposure, as was expected in the report for the control cities. However, PFOS concentrations
around Port St. Lucie, Florida, the other control city, were unexplainably similar to, and at times
N-ll
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higher than, the waterbodies with known PFOS discharges. The sources of PFOS near Port St.
Lucie, Florida remain unknown; however, observed PFOS concentration suggest the presence of
a potential manufacturing/industrial source or the use of AFFF in this area (3M Company 2001).
Water samples were collected from ponds near all of the sampling sites except those in
Cleveland, Tennessee. PFOS concentrations in these additional pond sites were similar to those
measured in Mobile, Alabama (ranging between 32 and 33 ng/L), lower than those observed in
Columbus, Georgia (as PFOS was not detected with a detection limit of 2.5 ng/L), and higher
than those measured in Decatur, Alabama (ranging between 108 and 111 ng/L) and in Port St.
Lucie, Florida (ranging between 1,830 and 48,200 ng/L). Samples collected from the pond site
near Port St. Lucie, Florida had some of the highest measured PFOS concentrations in publicly
available literature with the maximum concentration of 48,200 ng/L. In the report, the 3M
Company conducted additional sampling at the pond site in Port St. Lucie, Florida and
determined that the measured PFOS concentrations at this site were more variable than the initial
measurements indicated and were lower than the previous measurements, ranging between below
detection (i.e., <2.5 ng/L) and 2,340 ng/L. Aside from the samples collected in Port St. Lucie,
Florida, this report demonstrated that measured PFOS concentrations in surface waters tend to be
higher in areas with PFOS 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.
(2001) studied a stretch of the Tennessee River near Decatur, Alabama, and Konwick et al.
(2008) focused on the Conasauga River in Georgia, both areas with known PFOS discharge and
use. In Hansen et al. (2002), discharge from a fluorochemical manufacturing facility entered the
Tennessee River towards the middle of the study area. In contrast, Konwick et al. (2008)
N-12
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compared the PFOS concentrations measured in the Conasauga River with those from sites with
no known exposure along the Altamaha River. In both studies, mean PFOS concentrations were
higher in the study areas with PFOS sources. Specifically, Hansen et al. (2002) observed mean
PFOS concentrations upstream of the fluorochemical manufacturing facility were 30.85 ng/L
(ranging between 16.0 and 52.6 ng/L) and were 103.9 ng/L (ranging between 30.3 and 144 ng/L)
downstream of the fluorochemical manufacturing facility. Similarly, Konwick et al. (2008)
observed higher measured PFOS concentrations in the Conasauga River, which ranged from
below the limit of detection (i.e., 1.5 ng/L) to 321 ng/L, compared to those in the Altamaha
River, ranging between 2.6 and 2.7 ng/L. Consistent with the report from the 3M Company
summarized above, effluents from manufacturing facilities, WWTP, and carpet mill effluents
were determined to be the source of increased PFOS concentrations in both the Tennessee and
Conasauga Rivers (Hansen et al. 2002; Konwick et al. 2008; respectively). These PFOS
concentrations are relatively consistent with those measured in Alabama and Georgia as reported
by the 3M Company (3M Company 2001).
Nakayama et al. (2007) and Cochran (2015) measured PFAS, including PFOS, 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) detected PFOS in 97.5% of all samples above
the limit of quantification of 1 ng/L. PFOS concentrations in the Cape Fear Drainage Basin
ranged between <1 (the lower limit of quantification) and 132 ng/L with a mean concentration of
31.2 ng/L. As in other studies summarized above, lower PFAS concentrations, including PFOS,
were found in the upland tributaries and concentrations were highest in the middle reaches of the
Cape Fear Drainage Basin, nearer expected sources. Wastewater treatment plant effluents were
N-13
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identified as the source of PFAS to the study area. AFFF usage at the Department of Defense
base in Fayetteville, North Carolina and the land application of contaminated biosolids likely
contributed as well (Nakayama et al. 2007). Cochran (2015) detected PFOS in 79% of all water
samples collected and concentrations ranged between below the limit of quantification (i.e., 10
ng/L) and 7,070 ng/L, with an average concentration of 776.7 ng/L. PFOS concentrations
collected in Barksdale Air Force Base varied based on proximity to fire training areas. Cochran
(2015) attributed the evaluated PFOS concentrations to runoff and ground infiltration of AFFF
formerly used on the base during firefighting and/or training.
N.4 PFOS occurrence and concentrations in the midwestern U.S.
Similar PFOS concentrations were reported in the publicly available literature for
waterbodies in urban areas across the midwestern U.S., with lower PFOS concentrations reported
in remote areas in the same states (Newsted et al. 2017; Simcik and Dorweiler 2005). In
Minnesota, Simcik and Dorweiler (2005) observed PFOS concentrations ranged between 2.4 and
50.4 ng/L in urban areas near Minneapolis and between less than the limit of quantification (i.e.,
0.27 ng/L) and 1.2 ng/L in remote areas in northern Minnesota. Additionally, Newsted et al.
(2007) reported an average PFOS concentration of 528.9 ng/L (ranging between below limit of
quantification and 18,200 ng/L; limit of quantification not provided) in surface waters collected
from the Upper Mississippi River near the Minneapolis/St. Paul, Minnesota metropolitan area.
The source of PFOS at these urban sites was attributed to manufacturing (3M plant), runoff, and
wastewater discharge (Newsted et al. 2017; Simcik and Dorweiler 2005).
N.5 PFOS occurrence and concentrations in the northeastern U.S.
Several studies measured PFOS concentrations in surface waters in the northeastern U.S.
that are comparable to those reported in Minnesota (NJDEP 2019; Sinclair et al. 2006). Sinclair
N-14
-------
et al. (2006) measured PFOS in various waterbodies across New York state and observed a
median concentration of 756 ng/L in surface waters collected from the Superfund site at Lake
Onondaga (ranging between 198 and 1,090 ng/L; Table N. 1) and attributed these elevated
concentrations to several industries located along Lake Onondaga. All other observed
concentrations of PFOS in New York, including sites along the Niagara River, the Finger Lakes,
Lakes Oneida and Champlain, the Erie Canal, and the Hudson River, had lower median PFOS
concentrations ranging between 0.8 and 13 (Table N.l)(Sinclair et al. 2006).
The New Jersey Department of Environmental Protection (NJDEP) measured PFOS in
surface water samples collected from 14 different sites across New Jersey. PFOS concentrations
ranged from below the detection limit of 2.0 ng/L and 102 ng/L (NJDEP 2019). Individual
samples collected along Pine, Little Pine, and Mirror Lakes had measured PFOS concentrations
of 102, 100, and 72.9 ng/L, respectively. All other observed concentrations of PFOS in New
Jersey freshwaters were below 15 ng/L. NJDEP attributed the elevated concentrations of PFOS
observed at Pine, Little Pine, and Mirror Lakes to the use of AFFF in training and/or fire-fighting
on the Department of Defense (DoD) Joint Base McGuire-Lix-Lakehurst (NJDEP 2019).
N.6 PFOS occurrence and concentrations in the western U.S.
PFOS concentrations in surface waters of western U.S. states are consistent with the
lower-end concentrations (less than 100 ng/L) measured in eastern states; however, the
monitoring data for PFOS was limited in the western U.S. Plumlee et al. (2008) measured PFOS
concentrations in Coyote Creek and a tributary of Upper Silver Creek in San Jose, California and
found concentrations to be similar to those measured in eastern states. Concentrations of PFOS
in Coyote Creek ranged from 4.8 to 25 ng/L and concentrations in Upper Silver Creek ranged
from 27 to 56 ng/L. The source of PFOS to these aquatic systems was unknown, however,
N-15
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Plumlee et al. (2008) stated that a combination of atmospheric deposition of volatile precursors
and surface runoff were likely sources of PFOS to both Coyote and Upper Silver Creeks.
Lastly, Dinglasan-Panlilio et al. (2014) measured PFOS concentrations in surface waters
along the Puget Sound in Washington, as well as Clayoquot and Barkley Sounds in British
Columbia, Canada. PFOS concentrations measured by Dinglasan-Panlilio et al. (2014) were
lower than those observed from sites in eastern states (such as those summarized above for
Alabama, Florida, and North Carolina with known manufacturing and/or industrial use of
PFOS). Concentrations ranged from 0.2 to 5.9 ng/L in Puget Sound and 0.25 to 0.7 ng/L in
Clayoquot and Barkley Sound, British Columbia. These concentrations are consistent with those
reported in the publicly available literature for remote areas, such as in Minnesota (Simcik and
Dorweiler 2005) and in New York (Sinclair et al. 2006), as summarized above. The study
authors indicated specific regional sources and atmospheric deposition were likely PFOS sources
to these remote areas (Dinglasan-Panlilio et al. 2014).
N.7 Comparison of PFOS occurrence in the U.S. to global surface waters
Similar to surface waters in the U.S., generally PFOS and PFOA were the most
commonly detected PFAS in surface waters around the world (Ahrens 2011). On a global scale,
PFOS concentrations in surface waters generally range between picogram/liter and
nanogram/liter with some concentrations in the milligram/liter range. However, PFOS
occurrence data were limited for surface waters in Africa and South America. Based on the
currently available data, PFOS concentrations in the U.S. were relatively similar to those
reported in studies with sampling sites in other countries. Global surface water PFOS
concentrations reported in the public literature ranged between not detected and 2,100,000 ng/L
N-16
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(Jarvis et al. 2021). These global surface water concentrations are summarized in Jarvis et al.
(2021) to provide a comparison with those observed in the U.S.
Overall, the currently available data on PFOS occurrence in ambient surface waters show
the widespread distribution and variability of PFOS concentrations in surface waters around the
world and that surrounding land use has a large influence on PFOS concentrations in surface
waters. In general, urbanized areas with high population densities tended to have elevated PFOS
concentrations in surface waters (Jarvis et al. 2021). Like in the U.S., PFOS concentrations in
surface waters around the world vary widely and current information on the environmental
distribution of PFOS in surface waters around the world is relatively limited.
N.8 PFOS Occurrence and Detection in Aquatic Sediments
PFOS has been detected in sediments of aquatic environments across various countries
(Lau et al. 2007). Typically, in the U.S., soil and sediment measurements of PFOS occur in the
|ig/kg dry weight (dw) range with measured concentrations in the public literature ranging from
not detected (with a detection limit of 0.08 |ig/kg dw) to 31.38 |ig/kg dw (3M Company 2001;
Cochran 2015). Anderson et al. (2016), measured concentrations of PFAS in sediment across ten
U.S. Air Force bases where there is a known history of use of AFFF use and found that PFOS
concentrations were detected in 94% of samples. The median concentration of PFOS across all
sample sites was 31.0 |ig/kg, with a maximum concentration of 190,000 |ig/kg (Anderson et al.
2016). Arias et al. (2015) measured PFOS in sediment from an evaporation pond used to collect
the wastewater arising from fire-fighting exercises at an Australian military air base. Despite the
discontinued use of PFOS/PFOA-based foams six years earlier, the PFOS sediment
concentration was 38,000,000 |ig/kg, a million times higher than the average global values for
sediments (0.28 - 3.8 |ig/kg PFOS) reported by the authors (Arias et al. 2015).
N-17
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These observed concentrations were similar to other sediment concentrations in areas
with known perfluorinated chemical discharges and manufacturing. Lasier et al. (2011) measured
PFOS in sediment from the Coosa River, Georgia watershed, upstream, and downstream of a
land-application site of municipal/industrial wastewater with sediment concentrations ranging
from less than the method detection limit (MDL) to 1.73 |ig/kg dw upstream of the land-
application and 1.66 - 20.18 |ig/kg dwPFOS downstream. Giesy and Kannan (2001), as
presented in OECD (2002), measured PFOS in sediments collected from locations upstream and
downstream of the 3M facility in Decatur, Alabama. The two closest sites downstream of the 3M
facility had significantly greater concentrations (1,299 and 5,930 |ig/kg ww) than the two
upstream sites (-0.18 and 0.98 |ig/kg ww; OECD 2002).
Other sediment concentrations across the U.S. were much lower: <4 |ig/kg across sites in
Puget Sound, Washington, San Francisco and Monterey Bay California, the Niagara River in
New York, and Lake Michigan. These concentrations appeared to be similar to other sediment
concentrations across the globe (Table N-l).
Table N-l. Global Sediment Concentration of PFOS.
Location
PI-OS concentration
Uclcrcncc
Tokyo Bay, Japan
0.29-0.36 |ig/kg dw
Ahrens et al. (2010)
Ariake Sea, Japan
0.11 |ig/kg ww
Nakata et al. (2006)
Toronto, Canada
<0.1-2.2 |ig/kg ww
Vedagiri et al. (2018)
Lake Ontario, Canada
10 |ig/kg dw
ECCC (2018)
Lake Ontario, Niagara Basin
27-47 iig/kg
Meyers et al. (2012)
Lake Ontario, Mississauga Basin
4.4-19 |ig/kg
Meyers et al. (2012)
Lake Ontario, Rochester Basin
8.1-49 |ig/kg
Meyers et al. (2012)
Resolute Lake, Canada
24-85 |ig/kg ww
Butt et al. (2010)
Gufunes Bay, Iceland
<50 |ig/kg ww
Butt et al. (2010); Kallenborn (2004)
Faroe Islands
<50-0.11 |ig/kg
Butt et al. (2010); Kallenborn (2004)
Urban reservoir, Singapore
2.8-3.6 |ig/kg dw
Nguyen et al. (2016)
N-l 8
-------
N.9 PFOS Occurrence and Detection in Air and Rain
Air concentrations of PFOS in the atmosphere varied widely across the globe. 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 of 2006 (Kim and Kannan 2007). PFOS in the gas phase had a
mean concentration of 1.70 pg/m3 (range: 0.94-3.0) and the particulate phase had a mean
concentration of 0.64 pg/m3 (range: 0.35-1.16) (Kim and Kannan 2007). Kim and Kannan (2007)
also reported mean PFOS concentrations of 0.36 ng/L and 0.62 ng/L in rain and snow,
respectively.
Above Lake Ontario, concentrations of PFOS in the particulate phase measured in air
samples over the lake were higher than those observed by Kim and Kannan (2007) near Albany,
NY. The mean concentration of PFOS at Lake Ontario was 6.4 ฑ3.3 pg/m3 (Boulanger et al.
2005a), with a range of concentrations from not detected to 8.1 pg/m3 (Martin et al. 2010). In an
urban area in Minneapolis, Minnesota, PFOS was measured in both the particulate and gas
phase. PFOS in the particulate phase ranged from 2.1 - 7.9 pg/m3 and the gas phase ranged from
1.8-5.0 pg/m3 in across the five samples (MPCA/STS 2007).
In Canada, PFOS air concentrations measured in 2009 showed widespread distribution
with remote sites having similar concentrations as urban sites (ECCC 2018). Using passive
samplers, PFOS concentrations were detected in Toronto, Ontario (8 pg/m3), an agricultural site
in Saskatchewan (5 pg/m3), Whistler, British Columbia (4 pg/m3), and Alert, N Nunavut (2
pg/m3) (ECCC 2013).
Other reported concentrations of PFOS in air samples included Sydney, Florida (3.4
pg/m3), Tudor Hill, Bermuda (6.1 pg/m3), Malin Head, Ireland (3.3 pg/m3), and Hilo, Hawaii
(6.6 pg/m3) are similar to the concentrations reported in Canada (ECCC 2018) and Japan (Sasaki
et al. 2003). The annual geometric mean concentration of PFOS in air samples collected monthly
N-19
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from 2001-2002 in the town of Oyamazaki and Fukuchiyama City were 5.3 and 0.6 pg/m3,
respectively (Sasaki et al. 2003).
Across Europe, PFOS air concentrations were reported to be variable. In the particulate
phase PFOS concentrations ranged from <1.8-46 pg/m3. Most locations had low (-1-2 pg/m3)
to less than the reported Minimum Detection Limit (MDL) and included Hazelrigg, United
Kingdom, Kjeller Norway, and Mace Head, Ireland (Barber et al. 2007). The highest
concentrations were reported in Manchester, United Kingdom. Similarly, high concentrations
were reported for another urban area, 150 pg/m3 for Paris, France (ECCC 2018).
Even in the Arctic, PFOS, its precursors, and degradation products, have been detected in
air samples in Resolute Bay, Nunavut, Canada, during the summer of 2004 (Stock et al. 2007).
PFOS in the filter samples were 1-2 orders of magnitude greater than other compounds, with a
mean concentration of 5.9 pg/m3 (Butt et al. 2010). These concentrations are greater than PFOS
concentrations measured in the particle phase of air samples measured in Zeppelinstasjonen,
Svalbard, Norway (Butt et al. 2010). PFOS was measured in September and December, 2006 and
August and December, 2007, with mean concentrations of 0.11 pg/m3 (range: 0.03 - 0.50 pg/m3)
and 0.18 pg/m3 (range: 0.02 - 0.97 pg/m3), respectively (NILU 2007).
N.10 PFOS Occurrence and Detection in Groundwater
Similar to surface water, PFOS and PFOA are the dominant PFAS detected in
groundwater. Generally, PFOS concentrations tend to occur in the ng/L range, with some
elevated detections in the |ig/L range (Ahrens 2011; Xiao 2017). Concentrations of PFOS were
detected in groundwater samples across Minnesota in 2006 and 2007, approximately five years
after the 3M Corporation phased out PFOS production in Minnesota in 2002 (MPCA/STS 2007).
Data collected from shallow aquifers across Minnesota in both urban and agricultural areas were
N-20
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likely affected by a variety of different contamination sources (i.e., industrial and municipal
stormwater, pesticides, land application of contaminated biosolids and atmospheric deposition)
and indicated that perfluorinated chemicals are present in areas beyond the disposal sites and
aquifers associated with these disposal sites (MPCA/STS 2007). Groundwater samples of PFOS
ranged from < 0.00222 - 0.037 |ig/L across urban areas, with most of the perfluorinated
compound detections in the Twin Cities metro area (MPCA/STS 2007). Concentrations in rural
areas of Minnesota were all less than the analytical method reporting limit (0.025 |ig/L).
Detections of PFOS in groundwater have been associated with the use of AFFF and fire-
training locations (Ahrens 2011; Xiao 2017). The use of AFFF to suppress fires resulted in the
release of various PFAS into the environment as AFFF contains high levels of PFAS (Ahrens
2011; Moody and Field 2000). The use of AFFF in particular has been identified as an important
source of groundwater contamination with PFAS (Moody and Field 2000). This contamination is
often persistent, lasting for many years after the release (Moody and Field 2000; Xiao 2017). The
transformation of PFOS precursor compounds (see Section 2.3) by soil micro-organisms may be
a contributing source of PFOS in groundwater (Xiao 2017).
Groundwater samples from wells in the area of a known plume were measured in 1998
and 1999. Samples were taken at the Wurtsmith Air Force Base in northeastern Michigan, a base
where fire-training exercises were conducted from the 1950's until the base was decommissioned
in 1993. PFOS concentrations ranged from 4.0 to 110 |ig/L depending on the proximity to the
training pad, demonstrating that PFOS is still present in measurable quantities for at least five or
more years after the use of AFFF (Moody et al. 2003). These values are consistent with ten other
U.S. Air Force bases where there is a known historic use of AFFF to extinguish hydrocarbon-
based fires but were not active fire-training areas. Anderson et al. (2016) measured groundwater
N-21
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samples between March and September 2014 at the ten locations with PFOS concentrations
detected in 96% of samples. The median groundwater concentration of PFOS across all sites was
2.17 |ig/L, with a maximum concentration of 8,970 |ig/L (Anderson et al. 2016). Other reported
groundwater concentrations at other U.S. military installations summarized by Cousins et al.
(2016) include: Tyndall Air Force Base (147 - 2,300 |ig/L; Schultz et al. 2004), Fallon Naval Air
station (< LOD - 380 |ig/L; Schultz et al. 2004) and Ellsworth Air Base (5 - 75 |ig/L; McGuire
et al. 2014). Similar concentrations are reported at other airports and bases globally, including at
a fire training area in Cologne, Germany (0.02 - 8.35 |ig/L; WeiB et al. 2012); air force base F18
in Sweden (< 0.001 - 42.2 |ig/L; Filipovic et al. 2015) and the Jersey airport in the United
Kingdom (10 - 98 |ig/L; Rumsby et al. 2009).
N.ll PFOS Occurrence and Detection in Ice
Very little information was provided about PFOS concentrations in ice. Saez et al. (2008)
found PFOS in a Russian Artie ice core sampled in 2007. The PFOS concentration reported was
0.0053 ng/L.
N-22
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Appendix O Bioaccumulation Factors (BAFs) Used to Calculate PFOS Tissue Values
O.l Summary Table of PFOS BAFs used to calculate tissue criteria and supplemental fish tissue values
Com 111011 N :1111c
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Tissue"1
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Location
Reference
common carp
Carassius auratus
Blood
11167
11167
high
17 Sites in six major rivers, Korea
Lam et al. (2014)
mandarin
Siniperca scherzeri
Blood
73612
73612
high
17 Sites in six major rivers, Korea
Lam et al. (2014)
lefteye flounder
Paralichthys olivaceus
Blood
5625
5625
medium
Ariake Bay
Taniyasu et al.
(2003)
crucian carp
Carassius carassius
Blood
80168
80168
high
Beijing Airport, China
Wang et al. (2016)
crucian carp
Carassius carassius
Blood
22484
22484
high
Gaobeidian Lake, China
Shi et al. (2020)
carp
Cyprinus carpio
Blood
84211
84211
medium
Lake Biwa
Taniyasu et al.
(2003)
bluegill
Lepomis macrochirus
Blood
11053
11053
medium
Lake Biwa
Taniyasu et al.
(2003)
largemouth bass
Micropterus salmoides
Blood
169737
169737
medium
Lake Biwa
Taniyasu et al.
(2003)
European perch
Perca fluviatilis
Blood
58000
58000
medium
Lake Halmsjon, near Stockholm,
Sweden
Wang et al. (2016)
black seabream
Acanthopagrus schlegeli
Blood
14138
14138
medium
Osaka Bay
Taniyasu et al.
(2003)
white croaker
Argyrosomus argentatus
Blood
19540
19540
medium
Osaka Bay
Taniyasu et al.
(2003)
Japanese scad
Trachurus japonicus
Blood
14138
14138
medium
Osaka Bay
Taniyasu et al.
(2003)
crucian carp
Carassius carassius
Blood
9638
9638
high
Tangxum Lake, China
Shi et al. (2015)
conger eel
Conger myriaster
Blood
3500
3500
medium
Tokyo Bay
Taniyasu et al.
(2003)
rockfish
Sebastes inermis
Blood
9423
9423
medium
Tokyo Bay
Taniyasu et al.
(2003)
Japanese stingfish
Sebastiscus marmoratus
Blood
5154
5154
medium
Tokyo Bay
Taniyasu et al.
(2003)
crucian carp
Carassius carassius
Blood
19999
19999
high
Xiaoqing River, China
Shi et al. (2015)
common carp
Cyprinus carpio
Blood
7244
7244
high
Xiaoqing River, China
Pan et al. (2017)
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kiinkiiii*
Locution
Re Terence
European perch
Perca fluviatilis
Liver
39000
39000
medium
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. (2015)
European chub
Leuciscus cephalus
Liver
19953
19953
high
Orge River, near Paris, France
Labadie and
Chevreuil (2011)
white croaker
Argyrosomus argentatus
Liver
1609
1609
medium
Osaka Bay
Taniyasu et al.
(2003)
common seabass
Lateolabrax japonicus
Liver
459.8
459.8
medium
Osaka Bay
Taniyasu et al.
(2003)
Japanese scad
Trachurus japonicus
Liver
1034
1034
medium
Osaka Bay
Taniyasu et al.
(2003)
crucian carp
Carassius auratus
Liver
19953
19953
high
Pearl River Delta, China
Panetal. (2014)
mud carp
Cirrhinus molitorella
Liver
25119
25119
medium
Pearl River Delta, China
Panetal. (2014)
leather catfish
Clarias fuscus
Liver
5012
5012
high
Pearl River Delta, China
Panetal. (2014)
grass carp
Ctenopharyngodon
idellus
Liver
39811
39811
high
Pearl River Delta, China
Panetal. (2014)
common carp
Cyprinus carpio
Liver
25119
25119
high
Pearl River Delta, China
Panetal. (2014)
chub
Hypophthalmichthys
molitrix
Liver
7943
7943
high
Pearl River Delta, China
Panetal. (2014)
snakehead
Ophicephalus argus
Liver
15849
15849
high
Pearl River Delta, China
Panetal. (2014)
bream
Parabramis pekinensis
Liver
3162
3162ฐ
high
Pearl River Delta, China
Panetal. (2014)
tilapia
Tilapia aurea
Liver
3162
3162ฐ
high
Pearl River Delta, China
Panetal. (2014)
chub
Leuciscus cephalus
Liver
4556
4556
medium
Roter Main, Upper Franconia,
Germany
Becker et al. (2010)
silver perch
Bidyanus bidyanus
Liver
26000
26000
high
Shoalhaven region, Australia
Terechovs et al.
(2019)
common shiner
Notropis cornutus
Liver
12589
12589
high
Spring/Etobicoke Creek, Toronto,
Canada
Awad et al. (2011)
sea mullet
Mugil cephalus
Liver
5000
5000
medium
Sydney Harbour, Australia
Thompson et al.
(2011)
crucian carp
Carassius carassius
Liver
4426
4426
high
Tangxum Lake, China
Shi et al. (2015)
common seabass
Lateolabrax japonicus
Liver
3269
3269
medium
Tokyo Bay
Taniyasu et al.
(2003)
flatfish
Pleuronectidae
Liver
6846
6846
medium
Tokyo Bay
Taniyasu et al.
(2003)
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Locution
Re Terence
rockfish
Sebastes inermis
Liver
2462
2462
medium
Tokyo Bay
Taniyasu et al.
(2003)
Japanese stingfish
Sebastiscus marmoratus
Liver
4423
4423
medium
Tokyo Bay
Taniyasu et al.
(2003)
crucian carp
Carassius carassius
Liver
9226
9226
high
Xiaoqing River, China
Shi et al. (2015)
common carp
Cyprinus carpio
Liver
4467
4467
high
Xiaoqing River, China
Pan et al. (2017)
crucian carp
Carassius carassius
Liver
10735
10735
high
Yubei River, China
Shi et al. (2020)
Mozambique
tilapia
Oreochromis
mossambicus
Muscle
17.44
17.44
medium
Matikulu, N2 Bridge
Fauconier et al.
(2020)
cape stumpnose
Rhabdosargus holubi
Muscle
8.718
8.718
medium
Matikulu, N2 Bridge
Fauconier et al.
(2020)
crucian carp
Carassius carassius
Muscle
50234
50234
high
Beijing Airport, China
Wang et al. (2016)
juvenile char
(muscle)
Salvelinus alpinus
Muscle
10800
20785
high
Char Lake, Canadian High Arctic
Lescord et al.
(2015)
adult char
(muscle)
Salvelinus alpinus
Muscle
40000
high
Char Lake, Canadian High Arctic
Lescord et al.
(2015)
crucian carp
Carassius carassius
Muscle
1130
1130
high
Gaobeidian Lake, China
Shi et al. (2020)
meagre
Argyrosomus regius
Muscle
2496
2496
high
Gironde estuary, SW France
Munoz et al. (2017)
common seabass
Dicentrarchus labrax
Muscle
3257
3257
high
Gironde estuary, SW France
Munoz et al. (2017)
spotted seabass
Dicentrarchus punctatus
Muscle
2535
3844
high
Gironde estuary, SW France
Munoz et al. (2017)
spotted seabass
Dicentrarchus punctatus
Muscle
5830
high
Gironde estuary, SW France
Munoz et al. (2017)
anchovy
Engraulis encrasicolus
Muscle
1761
1761
high
Gironde estuary, SW France
Munoz et al. (2017)
mullet
Liza ramada
Muscle
1226
1226
high
Gironde estuary, SW France
Munoz et al. (2017)
sprat
Sprattus sprattus
Muscle
808.7
808.7
medium
Gironde estuary, SW France
Munoz et al. (2017)
tilapia
tilapia
Muscle
245.0
256.4
medium
Key River, Taiwan
Lin et al. (2014)
tilapia
tilapia
Muscle
323.0
medium
Key River, Taiwan
Lin et al. (2014)
tilapia
tilapia
Muscle
213.0
medium
Key River, Taiwan
Lin et al. (2014)
European perch
Perca fluviatilis
Muscle
3400
3400
high
Lake Halmsjon, near Stockholm,
Sweden
Ahrens et al. (2015)
brown bullhead
Ameiurus nebulosus
Muscle
794.3
794.3
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
common carp
Cyprinus carpio
Muscle
7943
7943
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
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kiinkiiii*
Locution
Re Terence
northern pike
Esox lucius
Muscle
1000
1000
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
channel catfish
Ictalurus punctatus
Muscle
3162
3162
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
pumpkinseed
Lepomis gibbosus
Muscle
631.0
631.0
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
smallmouth bass
Micropterus dolomieu
Muscle
6310
6310
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
largemouth bass
Micropterus salmoides
Muscle
5012
5012
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
yellow perch
Percaflavescens
Muscle
794.3
794.3
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
white crappie
Pomoxis annularis
Muscle
1000
1000
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
black crappie
Pomoxis nigromaculatus
Muscle
1585
1585
medium
Lake Niapenco, Ontario, Canada.
Bhavsar et al.
(2016)
juvenile char
(muscle)
Salvelinus alpinus
Muscle
1878
1048
high
Meretta Lake, Canadian High Arctic
Lescord et al.
(2015)
adult char
(muscle)
Salvelinus alpinus
Muscle
585.4
high
Meretta Lake, Canadian High Arctic
Lescord et al.
(2015)
eel
Anguilla anguilla
Muscle
3236
3236
high
Netherlands
Kwadijk et al.
(2010)
European chub
Leuciscus cephalus
Muscle
2512
2512
high
Orge River, near Paris, France
Labadie and
Chevreuil (2011)
crucian carp
Carassius auratus
Muscle
1585
1585
high
Pearl River Delta, China
Panetal. (2014)
mud carp
Cirrhinus molitorella
Muscle
2512
2512
medium
Pearl River Delta, China
Panetal. (2014)
leather catfish
Clarias fuscus
Muscle
251.2
251.2C
high
Pearl River Delta, China
Panetal. (2014)
grass carp
Ctenopharyngodon
idellus
Muscle
2512
2512
high
Pearl River Delta, China
Panetal. (2014)
common carp
Cyprinus carpio
Muscle
1585
1585
high
Pearl River Delta, China
Panetal. (2014)
chub
Hypophthalmichthys
molitrix
Muscle
631.0
631.0
high
Pearl River Delta, China
Panetal. (2014)
snakehead
Ophicephalus argus
Muscle
398.1
398.1
high
Pearl River Delta, China
Panetal. (2014)
bream
Parabramis pekinensis
Muscle
398.1
398.1
high
Pearl River Delta, China
Panetal. (2014)
0-5
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kiinkiiii*
Locution
Re Terence
juvenile char
(whole body)
Salvelinus alpinus
WB
8889
8889
high
Small Lake, Canadian High Arctic
Lescord et al.
(2015)
common shiner
Notropis cornutus
WB
1995
1995
high
Spring/Etobicoke Creek, Toronto,
Canada
Awad et al. (2011)
crucian carp
Carassius carassius
WB
1963
1963
high
Tangxum Lake, China
Shi et al. (2015)
bleak
Alburnus alburnus
WB
251.2
251.2
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
(2017)
common carp
Cyprinus carpio
WB
1000
1000
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
(2017)
mullet
Liza sp.
WB
4.786
4.786
medium
Xerta, Ebro Delta, Spain
Pignotti et al.
(2017)
roach
Rutilus rutilus
WB
199.5
199.5
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
100.0
100.0
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)
crucian carp
Carassius carassius
WB
2818
2818
high
Xiaoqing River, China
Shi et al. (2015)
mesozooplankton
Mesozooplankton
Invert
3450
3450
high
17 Sites in six major rivers, Korea
Lam et al. (2014)
microzooplankton
Microzooplankton
Invert
3017
3017
high
17 Sites in six major rivers, Korea
Lam et al. (2014)
chironomids
Diptera
Invert
550000
550000
high
9-Mile Lake, Canadian High Arctic
Lescord et al.
(2015)
zooplankton
zooplankton
Invert
100000
100000
high
9-Mile Lake, Canadian High Arctic
Lescord et al.
(2015)
snail
Gastropoda
Invert
15.26
15.26
medium
aMatikulu N2 Bridge
Fauconier et al.
(2020)
zooplankton
zooplankton
Invert
295.1
295.1
medium
Baltic Sea
Gebbink et al.
(2016)
chironomids
Diptera
Invert
280000
280000
high
Char Lake, Canadian High Arctic
Lescord et al.
(2015)
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kiinkiiii*
Locution
Re Terence
zooplankton
zooplankton
Invert
2400
2400
high
Char Lake, Canadian High Arctic
Lescord et al.
(2015)
ghost crab
Ocypode stimpsoni
Invert
3270
3270
medium
Fenglin - Xiamen Sea, China
Dai and Zheng
(2019)
copepods
Copepoda
Invert
3.400
68.50
high
Gironde Estuary, SW France
Munoz et al. (2019)
copepods
Copepoda
Invert
1380
medium
Gironde Estuary, SW France
Munoz et al. (2017)
brown shrimp
Crangon crangon
Invert
3.900
166.5
high
Gironde Estuary, SW France
Munoz et al. (2019)
brown shrimp
Crangon crangon
Invert
7110
medium
Gironde Estuary, SW France
Munoz et al. (2017)
oyster
Crassostrea gigas
Invert
122.0
122.0
high
Gironde Estuary, SW France
Munoz et al. (2017)
gam ma rids
Gammarus sp.
Invert
2380
2380
medium
Gironde Estuary, SW France
Munoz et al. (2017)
mysids
Mysidacea
Invert
3.900
117.8
high
Gironde Estuary, SW France
Munoz et al. (2017)
mysids
Mysidacea
Invert
3560
medium
Gironde Estuary, SW France
Munoz et al. (2017)
white shrimp
Palaemon longirostris
Invert
3.400
97.62
high
Gironde Estuary, SW France
Munoz et al. (2019)
white shrimp
Palaemon longirostris
Invert
2803
medium
Gironde Estuary, SW France
Munoz et al. (2017)
Pacific oyster
Crassostrea gigas
Invert
6430
6430
medium
Gulf Park - Xiamen Sea, China
Dai and Zheng
(2019)
snails
Bithynia tentaculata
Invert
128.4
128.4
high
Hogsmill River, Chertsey Bourne
River, Blackwater River
Wilkinson et al.
(2018)
amphipod
Gammarus pulex
Invert
118.0
118.0
high
Hogsmill River, Chertsey Bourne
River, Blackwater River
Wilkinson et al.
(2018)
Manila clam
Ruditapes philippinarum
Invert
3991
3991
high
Jiaozhou Bay, China
Cui et al. (2019)
orange-striped
hermit crab
Clibanarius infraspinatus
Invert
3879
3879
medium
Jimei Bridge - Xiamen Sea, China
Dai and Zheng
(2019)
Pacific oyster
Crassostrea gigas
Invert
4180
4180
medium
Jimei Bridge - Xiamen Sea, China
Dai and Zheng
(2019)
ghost crab
Ocypode stimpsoni
Invert
4240
4240
medium
Jimei Bridge - Xiamen Sea, China
Dai and Zheng
(2019)
diporeia
Diporeia hoyi
Invert
32000
32000
medium
Lake Ontario
Houde et al. (2008)
mysis
Mysis relicta
Invert
3000
3000
medium
Lake Ontario
Houde et al. (2008)
zooplankton
zooplankton
Invert
650.0
650.0
high
Lake Ontario
Houde et al. (2008)
worms
Capitellidae
Invert
913.93
913.93
high
Mai Po Marshes, Hong Kong
Loi et al. (2011)
gastropoda
Gastropoda
Invert
92.33
92.33
medium
Mai Po Marshes, Hong Kong
Loi et al. (2011)
sand prawn
Metapenaeus ensis
Invert
286.4
286.4
medium
Mai Po Marshes, Hong Kong
Loi et al. (2011)
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0.2 Summary of PFOS BAFs used to calculate tissue criteria and
supplemental fish tissue values
Field measured BAFs used to calculate fish and invertebrate PFOS tissue criteria (fish
muscle, fish whole body, and 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 P-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 P-3).
0-13
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0.3 PFOS BAFs References
Ahrens, L., K. Norstrom, T. Viktor, A.P. Cousins, 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, E. Braekevelt. 2011. Long-Term Environmental Fate of Perfluorinated Compounds
after Accidental Release at Toronto Airport. Environ. Sci. Technol. 45: 8081-8089.
Becker, A.M., S. Gerstmann, H. Frank. 2010. Perfluorooctanoic Acid and Perfluorooctane
Sulfonate in Two Fish Species Collected from the Roter Main River, Bayreuth, Germany.
Bulletin of Environmental Contamination and Toxicology 84: 132-135.
Bhavsar, S.P., C. Fowler, S. Day, S. Petro, N. Gandhi, S.B. Gewurtz, C. Hao, X. Zhao, K.G.
Drouillard, D. Morse. 2016. High levels, partitioning and fish consumption based water
guidelines of perfluoroalkyl acids downstream of a former firefighting training facility in
Canada. Environment International 94: 415-423.
Cui, W.J., J.X. Peng, Z.J. Tan, Y.X. Zhai, M M. Guo, and H.J. Mou. 2019. Pollution
characteriztics of perfluorinated alkyl substances (PFASs) in seawater, sediments, and biological
samples from Jiaozhou Bay, China. Huanjing Kexue 40(9): 3990-3999.
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, 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. Environment International 39: 19-26.
Fang, S., X. Chen, S. Zhao, Y. Zhang, W. Jiang, L. Yang, 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.
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, S.A. Mabury. 2007. Spatial Distribution of Perfluoroalkyl Contaminants in Lake Trout
from the Great Lakes. Environ. Sci. Technol. 41(5): 1554-1559.
0-14
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Precursors in the Baltic Sea Environment: Do Precursors Play a Role in Food Web Accumulation
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Houde M., G. Czub, J.M. Small, S. Backus, X. Wang, M. Alaee, D.C. Muir. 2008. Fractionation
and bioaccumulation of perfluorooctane sulfonate (PFOS) isomers in a Lake Ontario food web.
Environ. Sci. Technol. 42: 9397-9403.
Iwabuchi, K., N. Senzaki, S. Tsuda, H. Watanabe, I. Tamura, H. Takanobu, N. 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, T. Sakurai. 2018.
Bioaccumulation Patterns of Perfluoroalkyl Acids in an Estuary of the Ariake Sea, Japan.
Bulletin of Environmental Contamination and Toxicology 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. Environmental Science Process and 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.
Kwadijk, C., M.J.J. Kotterman, A. Koelmans. 2014. Partitioning of perfluorooctanesulfonate and
perfluorohexanesulfonate in the aquatic environment after an accidental release of Aqueous Film
Forming Foam at Schiphol Amsterdam Airport. Environ. Toxicol. Chem. 33: 1761-1765.
Labadie, P. and M. Chevreuil. 2011. Partitioning behaviour of perfluorinated alkyl contaminants
between water, sediment and fish in the Orge River (nearby Paris, France). Environmental
Pollution 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, H.-S. Cho. 2014. Perfluorinated alkyl substances in water,
sediment, plankton and fish from Korean rivers and lakes: A nationwide survey. Sci. Total
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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, T.-H. Yu. 2014. Occurrence of perfluorinated
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sediments, and biotissues. Environmental monitoring and assessment 186: 3265-3275.
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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.
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Selleslagh, C. Simonnet-Laprade, 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, 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, Q.-Q. Zhang. 2014. Bioaccumulation and risk assessment of
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region, South China. Ecotoxicology and Environmental Safety 107: 192-199.
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of perfluoroalkyl substances in Lake Chaohu, China. Int. J. Environ. Res. Public Health 16(10):
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on the Occurrence and Bioaccumulation of Hexafluoropropylene Oxide Trimer Acid: An
Emerging Concern. Environ. Sci. Technol. 51: 9553-9560.
Pignotti, E., G. Casas, M. Llorca, A. Tellbuscher, D. Almeida, E. Dinello, M. Farre, 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, S.E. Focardi. 2013. Perfluorinated compounds:
Levels, trophic web enrichments and human dietary intakes in transitional water ecosystems.
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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. Technol. 49:14156-14165.
Shi Y., R. Vestergren, T.H. Nost, Z. Zhou, Y. Cai. 2018. Probing the differential tissue
distribution and bioaccumulation behavior of per-and polyfluoroalkyl substances of varying
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52: 4592-4600.
Shi, Y., X. Song, Q. Ji, W. Li, S. He, Y. Cai. 2020. Tissue distribution and bioaccumulation of a
novel polyfluoroalkyl benzenesulfonate in crucian carp. Environment International 135: 105418.
0-16
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sulfonate and related perfluorinated organic compounds in water, fish, birds, and humans from
Japan. Environ. Sci. Technol. 37: 2634-2639.
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Nghiem. 2019. Occurrence and bioconcentration of micropollutants in Silver Perch (Bidyanus
bidyanus) in a reclaimed water reservoir. Sci. Total Environ. 650 ( 1): 585-593.
Thompson, J., A. Roach, G. Eaglesham, M.E. Bartkow, K. Edge, J.F. Mueller. 2011.
Perfluorinated alkyl acids in water, sediment and wildlife from Sydney Harbour and
surroundings. Mar. Pollut. Bull. 62(12): 2869-2875.
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Distribution, and Bioaccumulation Potential of Cyclic Perfluorinated Sulfonic Acids Isomers in
an Airport Impacted Ecosystem. Environ. Sci. Technol. 50: 10923-10932.
Wilkinson, J.L., P.S. Hooda, J. Swinden, J. Barker, S. Barton. 2018. Spatial (bio) accumulation
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sediment, aquatic plants and benthic organisms. Environ. Pollut. 234: 864-875.
Zhou, Z., Y. Shi, L. Xu, 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.
0-17
-------
Appendix P Translation of Chronic Water Column Criterion into Other
Fish Tissue Types (liver, blood, reproductive tissues)
The PFOS aquatic life criteria (summarized in Section 3.3) include chronic tissue criteria
for fish whole body, fish muscle, and invertebrate whole-body. Additional values for fish liver,
fish blood, and fish reproductive tissues were also calculated by transforming the chronic water
column criterion (i.e., 0.00025 mg/L) into representative tissue concentrations using tissue-
specific bioaccumulation factors (BAFs). Fish BAFs for liver, blood, and reproductive tissues
were identified following the same approaches used to identify fish whole body, muscle, and
inverterbrate whole body BAFs, which are described in detail in Section 2.11.3.1. Briefly, BAFs
were determined from field measurements and calculated using the equation:
BAF = ฃbiota ^Eq p_^
Cwater
Where:
CUota = PFOS concentration in organismal tissue(s)
Cwater = PFOS concentration in water
For further details on BAFs compilation and ranking, see Section 2.11.3.1 and Burkhard
(2021). BAFs based on reproductive tissues identified by Burkhard (2021) were further screened
to evaluate characteristics that influence reproductive tissue BAFs. These characteristics
included timing of sample collection and organism sex, age, length and weight. However, since
the data were limited, the influence of these characteristics could not be fully evaluated to
determine their potential influence on PFOS BAFs for reproductive tissues. Therefore,
characteristics of timing of sample collection and organism age, length or weight were currently
not considered to be influential given available data. Reproductive tissue BAFs were additionally
screened to ensure only BAFs based on adult females were considered, because female
P-l
-------
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 P-l.
Table P-l. Characteristics of adult fish sampled for the calculation of PFOS reproductive
tissue BAFs.
All sampled fish were adults, and all reproductive tissues identified as gonad. Weights, lengths, and BAFs are
Author
Species
Collection
Dale
11
Sex
Age
(\r.)
Weight
(g-\v\v)
Length
(cm)
liAl
(l/lvg)
Ahrens et al.
(2015)
European perch
(Perca fluviatilis)
10/12/2012
3
F
7, 8, 9
N.R.
N.R.
16,000
Becker et al.
(2010)
European chub
(Leuciscus
cephalus)
8/28/2007
6
N.R.
4
178.5
25.5
2,222
Labadie and
Chevreuil
(2011)
European chub
(Leuciscus
cephalus)
April 2010
5
3 M
2 F
N.R.
228.0 (M)
258.2 (F)
28.5 (M)
27.8 (F)
10,000
Shi et al.
(2015, 2018)
Crucian carp
(Carassius
carassius)
July 20141
30
24 F
6 M
N.R.
79.4 (F)
60.5 (M)
15.0 (F)
13.7 (M)
11,482
Shi et al.
(2015, 2018)
Crucian carp
(Carassius
carassius)
July 20142
13
9 F
4 M
N.R.
352.3 (F)
320.7 (M)
24.6 (F)
24.8 (M)
5,888
Shi et al.
(2020)
Crucian carp
(Carassius
carassius)
N.R.
303
N.R.
N.R.
N.R.
N.R.
7,990
Shi et al.
(2020)
Crucian carp
(Carassius
carassius)
N.R.
203
N.R.
N.R.
N.R.
N.R.
8,012
Wang et al.
(2016)
Crucian carp
(Carassius
carassius)
April 2014
8
N.R.
N.R.
(16.8-
65.1)5
(10.0-
14.7)5
25,645
N.R.= Not Reported
'Xiaoqing River, China
2TangxunLake, China
3Yubei River, China
4Gaobeidian Lake, China
5Range
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 3.2.3.1). Briefly, distributions of
BAFs used to derive additional tissue values were based on the lowest species-level BAF
P-2
-------
reported at a site. When more than one BAF was available for the same 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 P-2) and individual BAFs are provided in Appendix O.
Table P-2. Summary Stal
tistics for PF<
DS BAFs in Additional
7ish Tissues1.
Category
ii
(iconic! lie
Mean
IJAI
(1 ./kg-wel
weight)
Median
IJAI'"
-------
Table P-3. PFOS Concentrations for Additional Fish Tissue.1'2
Category
PI-'OS Concentration (inป/kป ww)
Liver
0.616
Blood
1.57
Reproductive Tissue
1.29
1 These PFOS concentrations are provided as supplemental information and are not intended to replace the PFOS fish tissue
criteria provided in Table .
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.
P-4
-------
Appendix Q Example Data Evaluation Records (DERs)
The PFOS toxicity literature evaluated and used to derive the PFOS aquatic life criteria
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).
Q-i
-------
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 the 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 PFOS exposure. Office of Water completed a DER for each species by
chemical combination from the PFOS studies identified by ECOTOX. Example DERs are
presented here to convey the meticulous level of evaluation, review, and documentation each
PFOS study identified by ECOTOX was subject to. Appendix Q.l shows an example fish DER
and Appendix Q.2 shows an example aquatic invertebrate DER.
Q-2
-------
Q.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 ofberyllium 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)
. , . . . . Bioaccumulation: steady state not
Dilution water not adequately characterized , ,
n J reached
Dermal or Injection Exposure Pathway
Review paper or previously published without modification
Q-3
-------
Other: (if any, list here)
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.
Q-4
-------
ABSTRACT: Copy and paste abstract from publication.
SUMMARY: Fill out and modify as needed.
Acute:
Species (IM'cshiiic)
Method'
1 CM
Duration
Chemical
/ Pnriu
pll
1 em p.
(ฐC)
lliirdness
(111ii/l. ilS
CaCO.0
til'
Salinilt
(Dl)ll
DOC
(niii/l.)
r.iTcci
Reported
r.iTcci
( onccnlralion
(niiป/l.)
Verified
I'.ITecl
C (inceiilr;ili
-------
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.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.
Pa ram el er
Trealmenl
Mean
Range
Dissolved
[1]
Oxygen
[2]
(% saturation
or mg/L)
j
j
[I]
Temperature
[2]
(Q
j
j
[I]
pH
[2]
j
j
Other (e.g.,
hardness,
salinity, DOC)
[1]
[2]
j
j
-------
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\ iiilion or
(oiiitii trillion
( <>iK'i'iilr;ilion
Nil in hoi' til'
\on-
Ik'low \on-
Siiindiird
1 iviilmonl
(iinils)
(iinils)
S;i in pies
Delccr'
Ik'kcl
r.rrorl
Kiinuo
Control
[11
[21
[31
[41
[51
[61
i
aNon-Detect: 0 = measured and detected; 1= measured and not detected; if not measured or reported enter as such
-------
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.
-------
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.
-------
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.
|Sl;iiul;inl
ISliindiird
|Mo;in
ISliindiird
ISliindiird
l)c\ iiilion
l)c\ iiilion
IlillCll
l)c\ iiilion
|Moiin
l)c\ iiilion or
IMciin
or
| Mciin
or
IV i veil I
or
1 iv;ilmonl
Nil in hoi' til'
M;iihI;ii-(I
Nil nihil- of
Siiindiird
IVrmK
Siiindiird
Sur\ i\ill
Siiindiird
(units)
S|):i\\iis|
I'.rrorl
I-'-UUNl
li-i-orl
lliilchl
I'.rrorl
Posll
I'.rrorl
Control
[1]
[2]
[31
[41
[51
[61
i
[ECx]
NOEC
LOEC
a Use superscript to identify the values reported to be significantly different from control.
-------
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
rn
[21
[31
[41
[51
[61
i
[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.
-------
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.
TF.ST 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 111
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:
-------
STUDY PARAMETERS: Provide information under Details and any relevant information of deficiencies in Remarks.
Complete for both Controlled Experiments and Field Studies/Observations.
5
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 [31
Control(s):
Treatment(s):
Number of Organisms per Replicate/
Treatment Group:
At least 10 organisms/treatment recommended [3]
At least 7 organisms/treatment acceptable [41
Control(s):
Treatment(s):
Exposure Pathway:
(i.e., water, sediment, gavage, or diet).
Note: all other pathways (e.g., dermal, single dose via
savage, 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
[21
Acute
Partial Life Cycle
Early Life Stage
Full Life Cycle
Other (please remark):
Test Concentrations (remember units):
Recommended test concentrations include at least three
concentrations other than the control; four or more will
provide a better statistical analysis [31
Nominal:
Measured:
Media measured in:
Observation Intervals:
Should be an appropriate number of observations
over the study to ensure water quality is being
properly maintained [41
-------
CONTROLLED EXPERIMENT STUDY PARAMETERS: Provide information under Details and any relevant
information of deficiencies in Remarks. Complete for Controlled Experiments only.
-*
Paramcler
Delails
Remarks
Acclimation/Holding:
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 111
Duration:
Identify number of individuals excluded from testing and/or
analysis (if 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) 111
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 121
Dilution Series (e.g., 0.5x, 0.6x, etc.):
-------
Paramcler
Delails
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 111
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 [31
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) [1,5]
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 111
-------
Piii'iiiiH'Icr
Doliiils
Koniiirks
.1
deeding:
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) 121
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 111
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.
-------
OBSERVATIONS: Provide information under Details and any relevant information in Remarks. This information should be
consistent with the Results Section in Part A.
Paramcler
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
Were replicate level concentration-response
data included in study publication (can be from
tables, graphs, or supplemental materials)?
specify endpoints in remarks
Yes No
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:
-------
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
-------
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.
Cum- II)
Spi'l'il'S
I'lidpuiiil
Tiv.i Inn-ill
Ki-|>lii;ili-
|Sl:i iul;i I'd
l)i-\ iiiiiim
hi*
Sl;i lld.i I'd
r.lTiil'l
#ซf
Sun ixiirs
V
k'
11 ฆ
Kl'SpilllM-
Kl'SpilllM-
l nil
Colli'
('(mi' units
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
-------
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.
-------
Q.2 Example Aquatic Invertebrate 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., 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.
{Ifyes, list file names of
Were other DERs completed for Companion Papers? Yes 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"
, , ,,, ,, , . , . . No Controls (for controlled experiments
Mixture (tor controlled experiments only) only)
Excessive Control Mortality (> 10% for acute and > 20% for chronic)
. , . . . . Bioaccumulation: steady state not
Dilution water not adequately characterized , ,
n J reached
Dermal or Injection Exposure Pathway
Review paper or previously published without modification
Q-23
-------
Other: (if any, list here)
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.
Q-24
-------
ABSTRACT: Copy and paste abstract from publication.
SUMMARY: Fill out and modify as needed.
Acute:
Species (lircsliiue)
Method'
1 CM
(liii'iilion
( hcmiciil
/ Piiriu
pll
Temp.
(ฐC)
Ihirdncss
(lllfi/l. ilS
CsiCO.0
or
S;ilini(\
(DDII
DOC
(niii/l.)
r.iTcci
Reported
I'.ITecl
( (iiicciilriilion
(miป/l.)
Verified
I'.ITecl
( oiicciiI r;it ion
(mป/l.)
Cliissificiilion
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 (lileslii^c)
Method-'
Tesl
dur;ilion
C hemic;il
/ Pu ri( \
pll
lemp.
(ฐC)
Ihirdness
(111 ii/1 :is
CaCO.o
or
S;ilini(\
(ppl)
l)()(
(mji/l.)
( hronic
Limits
Reported
Chronic
Vsilue
(inii/l. or
Verified
Chronic
\ ;ilue
(mii/l. or
Chronic
Value
I'.ndpoinl
Cliissificiilion
Quantitative /
Qualitative / Unused
a S=static, R=renewal, F=flow-through, U=unmeasured, M=measured, T=total, D=dissolved, Diet=dietary, MT=maternal transfer
Q-25
-------
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.
Pa ram el er
Trealmenl
Mean
Range
Dissolved
[1]
oxygen
[2]
(% saturation
or mg/L)
j
j
[I]
Temperature
[2]
(Q
j
j
[I]
pH
[2]
j
j
Other (e.g.,
hardness,
salinity, DOC)
[1]
[2]
j
j
Q-26
-------
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\ iiilion or
(oiiitii trillion
( <>iK'i'iilr;ilion
Nil in hoi' til'
\on-
Ik'low \on-
Siiindiird
1 iviilmonl
(iinils)
(iinils)
S;i in pies
Delccr'
Ik'kcl
r.rrorl
Kiinuo
Control
[11
[21
[31
[41
[51
[61
i
aNon-Detect: 0 = measured and detected; l=measured and not detected; if not measured or reported enter as such
Q-27
-------
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.
Q-28
-------
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.
Q-29
-------
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)e\ 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.
Q-30
-------
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.
|Me;iit Siihlolhiil
Response!
|Si;iiithirซl Do iiilion or
1 IVillllHMII
( ii nils)
Siiindiird l.rmr|
Control
[1]
[2]
[3]
[4]
[5]
[6]
./'
TECxl
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.
Q-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
Details
Remarks
Species:
Common Name:
Scientific Name:
Norili American s>pecies>".'
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:
Q-32
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STUDY PARAMETERS: Provide information under Details and any relevant information of deficiencies in Remarks.
Complete for both Con trolled Experiments and Field Studies Observations.
Pummel or
Delails
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 [51
Control(s):
Treatment(s):
Number of Organisms per Replicate/
Treatment Group:
At least 10 organisms/treatment recommended.
Control(s):
Treatment(s):
ฃ
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):
-
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 [41
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 [11
Q-33
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CONTROLLED EXPERIMENT STUDY PARAMETERS: Provide information under Details and any relevant
information of deficiencies in Remarks. Complete for Controlled Experiments only.
Parameter
Delails
Remarks
Acclimation/Holding:
Should be placed in a tank along with the water in
Duration:
Identify number of individuals excluded from testing and/or
analysis (if any):
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 111
Dissolved Oxygen (mg/L):
Health {any mortality observed?):
-
Acclimation followed published guidance?
Describe, if any
Yes No
-*
If yes, indicate which guidance:
i:
Test Vessel:
Test chambers should be loosely covered [1]
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 131
Substrate Used (if applicable)-.
Q-34
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Paramcler
Details
Remarks
*ฆ*
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 131
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 111
Yes No
Describe Preparation of Test
Concentrations (e.g., water exposure,
diet):
Q-35
-------
Parameter
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
Q-36
-------
*ฆ*
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) 121
Yes No
Lighting:
No specific requirements for 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 [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.
Q-37
-------
OBSERVATIONS: Provide information under Details and any relevant information in Remarks. This information should be
consistent with the Results Section in Part A.
Parameter
Delails
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 131
Yes No
Control survival (%):
Q-38
-------
Paramcler
Details
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:
Q-39
-------
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
Q-40
-------
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.
Cum- II)
Spi'l'il'S
I'lidpuiiil
Tiv.i Inn-ill
Ki-|>lii;ili-
|Sl:i iul;i I'd
l)i-\ iiiiiim
hi*
Sl;i lld.i I'd
r.lTiil'l
#ซf
Sun ixiirs
V
k'
11 ฆ
Kl'SpilllM-
kl'spilllsi-
I nil
Colli'
('(mi' units
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
Q-41
-------
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.
Q-42
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