### United States Environmental Protection Agency Office Of Water (EN-336) EPA/505/2-90-001 PB91-127415 March 1991 

# EPA Technical Support Document For Water Quality-based Toxics Control

### United States Office ofWater Environmental Protection 

### Washington, DC Agency



# TECHNICAL SUPPORT DOCUMENT FOR WATER QUALITY-BASED TOXlCS CONTROL

This 

## copyrepresents the second printing of this document. Changes made to this document reflect corrections of typographical errors and the Wowing update of the interim 

## guidance on 

## criteriafor metals: TheAgency 

## has issued“Interim 

## Guidance Interpretation 

## andImplementation Aquatic Life

## Criteriafor 

## Metals.” The interim guidance supersedescriteria document 

## statements expressing 

## criteria in 

## terms ofa acid soluble 

## analytical method 

## and also the 

## metals discusses of Section 

## 5.7.3. Theavailability of this document appeared in the June 5, 1992Federal Register(Vol. 57, No. 109, pg. 24407).

March 1991 

### Office ofWater Enforcement and Permits Office ofWater Regulations and Standards 

## U.S. Environmental Protection Agency Washington, DC20460

### FOREWORD
The U.S.EnvironmentalProtectionAgency(EPA) and the Statepollution control agencies havebeencharged with enforcingthelawsregardingpollutionofthe naturalenvironment. Environmental pollution isanurgentand continuing problemand,consequently, the laws grantconsiderable discretion tothecontrolauthorities to define environmentalgoalsanddevelopthe means to attainthem. Establishing environmentallyprotectivelevelsand incorporating them in a decisionmakingprocess entailsa considerable amount of scientificknowledgeand judgment. Onearea wherescientificknowledge israpidlychangingconcerns thedischarge oftoxic pollutantsto the Nation’ssurface waters. Thisdocumentprovides technicalguidanceforassessing andregulatingthedischarge oftoxic substances to the watersof the United States. It was issuedin support of EPA regulationsand policy initiatives involving the applicationof biologicalandchemicalassessment techniquesto control toxic pollution to surface waters. This documentisagency guidanceonly. Itdoesnot establish oraffectlegalrightsorobligations. Itdoesnot establish abinding normand isnotfinallydeterminativeofthe issues addressed. Agencydecisions in any particularcase will bemadeapplyingthe lawandregulations onthe basis ofspecific facts whenpermitsareissued or regulations promulgated. Thisdocumentisexpectedto berevised periodically to reflectadvances inthis rapidlyevolvingarea. Comments from users will bewelcomed. Sendcommentsto U.S.EPA, OfficeofWaterEnforcement and Permits, 401 M Street,SW, MailcodeEN366, Washington,DC20460.

James R.Elder,Director OfficeofWaterEnforcement and Permits MarthaG. Prothro,Director OfficeofWaterRegulations andStandards

iii

### TABLE OFCONTENTS Section Page Foreword

................................. iii Acknowledgment

....................................................................... xiii Executive Summary

..................................................................... xiv Listof Abbreviations.

.................................................................... xvii Glossary

.......................................................................... xix Introduction

.......................................................................... xxiii
1. APPROACHES TO WATER QUALITY-BASED TOXICSCONTROL

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 1.1 INTRODUCTION

............................................................ 1 1.2 CHEMICAL-SPECIFIC APPROACH FOR AQUATIC LIFE PROTECTION

........................................................... 1
1.2.1
1.2.2 Correlation of Chemical-specific Measurements to Actual ReceivingWater Impacts

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 Chemical-SpecificAnalytical Method Precision

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
1.3 WHOLEEFFLUENT APPROACH FOR AQUATIC LIFE PROTECTION

........................................................... 4
1.3.1 1.3.2
1.3.3 1.3.4 Toxic units

............................................................ 6 Correlation of Whole EffluentToxicity Measurements to Actual ReceivingWater Impact

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 Toxicity Test Method Precision.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 Considerations Involved When Implementing the Whole EffluentToxicity Approach.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11
1.4 BIOLOGICALCRITERIA/BIOASSESSMENT AND BIOSURVEY APPROACH FORAQUATIC LIFEPROTECTION.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18
1.4.1
1.4.2 Useof Biosurveysand Bioassessments in Water QualitybasedToxics Control.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18 Conducting Biosurveys

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...19
1.5 INTEGRATIONOFTHEWHOLEEFFLUENT,CHEMICALSPECIFIC, AND BIOASSESSMENT APPROACHES

.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20
1.5.1 Capabilities and Limitations of the Chemical-SpecificApproach

..................... 20 1.5.2 Capabilities and Limitations of the Whole EffluentApproach

....................... 21 1.5.3 Capabilities and Limitations of the Bioassessment Approach

....................... 22

V

### Section Table of Contents (Continued) Page

1.6 OTHERFACTORS INFLUENCINGWATER QUALITY-BASED TOXICS CONTROL

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23
1.6.1 Persistence

........................................................... 23 1.6.2 Additivity, Antagonism, and Synergism

...................................... 24 1.6.3 Test Interferences

...................................................... 24
1.7 HUMAN HEALTHPROTECTION

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24
1.7.1 Types ofHealth Effects

................................................................................................................ 25
REFERENCES

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...26
2. WATER QUALITYCRlTERlA AND STANDARDS.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29
2.1 INTRODUCTION

........................................................... 29
2.1.1 Overview of Water Quality Standards.

....................................... 29 2.1.2 Water Quality Standardsand StateToxics Control Programs

....................... 30
2.2 GENERAL CONSIDERATIONS.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31
2.2.1 Magnitude, Duration, and Frequency.

....................................... 31 2.2.2 Mixing Zones..

....................................................... 33
2.3 WATER QUALITYCRITERIA FOR AQUATICLIFE PROTECTION.

. . . . . . . . . . . . . . . . . . . . . . . . . . 34
2.3.1 Development Process for Criteria

........................................... 34 2.3.2 Magnitude forSingle Chemicals

........................................... 34 2.3.3 Magnitude for Whole EffluentToxicity

....................................... 35 2.3.4 Duration for Single Chemicalsand Whole EffluentToxicity

........................ 35 2.3.5 Frequencyfor Single Chemicalsand Whole EffluentToxicity

....................... 36
2.4 WATER QUALITYCRITERIA FORHUMAN HEALTHPROTECTION

. . . . . . . . . . . . . . . . . . . . . . . . 36
2.4.1 2.4.2 2.4.3 2.4.4 2.4.5
2.4.6
2.4.7 2.4.8 2.4.9
2.4.10 Overview

............................................................ 36 Magnitude and Duration

............................................... ..37 Human ExposureConsiderations

........................................... 37 FishConsumption Values.

................................................ 37 Bioaccumulation Considerations for Reference Ambient Concentration Development

.............................................. 38 Updating Human Health Criteria and Generating RACs Using IRIS

............................................................ 38 Calculating RACs for Non-carcinogens

....................................... 39 Calculating RACs for Carcinogens

.......................................... 40 Deriving Quantitative Risk Assessments in the Absence oflRlS Values

....................................................... ..40 Deriving Reference TissueConcentrations for Monitoring Fish Tissue

......................................................... ..41

vi

### Table of Contents (Continued) Section Page

2.5. BIOLOGICALCRITERIA.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41
2.5.1 Regulatory Bases for Biocriteria.

............................................ 41 2.5.2 Development and Implementation of Biocriteria

................................ 41
2.6. SEDIMENTCRITERIA

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...42
2.6.1 2.6.2 2.6.3 2.6.4 Current Developments in Sediment Criteria

................................... 42 Approach to Sediment Criteria Development

.................................. 42 Application of Sediment Criteria.

........................................... 43 Sediment Criteria Status

................................................. 43
REFERENCES

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...45
3. EFFLUENT CHARACTERlZATlON

....................................................... 47 3.1 lNTRODUCTlON...........................................................47
3.1.1 3.1.2
3.1.3 NPDESRegulation Requirements

........................................... 47 Background for Toxic Effects Assessments on Aquatic Life and Human Health

..................................................... 48 General Considerations in Effluent Characterization.

............................. 49
3.2 DETERMINING THENEEDFORPERMITLIMITSWITHOUT EFFLUENT MONITORING DATAFOR A SPECIFIC FACILITY.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . 50 3.3 DETERMINING THENEEDFORPERMITLIMITSWITH EFFLUENT MONITORING DATA,

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51
3.3.1 3.3.2
3.3.3 3.3.4
3.3.5 3.3.6
3.3.7 3.3.8 3.3.9 3.3.10 General Considerations

.................................................. 51 Addressing Uncertainty in Effluent Characterization by Generating Effluent Monitoring Data

........................................ 52 Effluent Characterization for Whole EffluentToxicity

............................. 53 Useof Toxicity Testing in Multiple-source Discharge Situations..

.......................................................... 59 Ambient Toxicity Testing

................................................. 61 SpecialConsiderations for Dischargesto Marine and EstuarineEnvironments

.................................................. 61 Using a Chemical-specific Limit to Control Toxicity

.............................. 61 Effluent Characterization for SpecificChemicals

................................ 62 Effluent Characterization for Bioconcentratable Pollutants

......................... 64 Analytical Considerations for Chemical.

...................................... 65
REFERENCES

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...66

vii

### Table of Contents (Continued) Section Page

4. EXPOSURE AND WASTELOAD ALLOCATION

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67 4.1 INTRODUCTION

........................................................... 67 4.2 TOTALMAXIMUM DAILYLOADSAND WASTELOAD ALLOCATIONS

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67
4.2.1 4.2.2 Total Maximum Daily Loads

.............................................. 67 WasteloadAllocation Schemes.

............................................ 69
4.3 INCOMPLETELY MIXED, DISCHARGE RECEIVING WATER SITUATIONS.

. . . . . . . . . . . . . . . . . . . 69
4.3.1 4.3.2 4.3.3 4.3.4 Determination of Mixing Zone Boundaries

.................................... 70 Minimizing the Sizeof Mixing Zones

....................................... 71 Prevention of Lethality to PassingOrganisms

.................................. 71 Prevention of Bioaccumulation Problemsfor Human Health

....................... 72
4.4 MIXING ZONEANALYSES

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72
4.4.1 4.4.2 4.4.3 4.4.4 4.4.5 General Recommendations for Outfall Design,

................................. 73 Critical Design Periodsfor Waterbodies

...................................... 73 General Recommendations for TracerStudies.

................................. 74 Discharge-induced Mixing.

............................................... 75 Ambient-induced Mixing.

................................................ 77
4.5 COMPLETELY MIXED DISCHARGE-RECEIVING WATER SITUATIONS

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78
4.5.1 45.2
4.5.3 4.5.4 4.5.5 Wasteload Modeling Techniques

........................................... 78 Calculating the Allowable Effluent Concentration Distribution and the Return Period..

................................................. 82 General Recommendations for Model Selection

................................ 83 SpecificModel Recommendations

.......................................... 83 EffluentToxicity Modeling.

............................................... 85
4.6 HUMAN HEALTH

........................................................... 87
4.6.1 Human Health Considerations

............................................. 87 4.6.2 Determining the TMDL Basedon Human Health Toxicants

........................ 87
REFERENCES

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90
5. PERMIT REQUIREMENTS

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93 5.1 INTRODUCTION

........................................................... 93
5.1.1 Regulatory Requirements.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

viii

### Table of Contents (Continued) Section Page

5.2 BASICPRINCIPLES OF EFFLUENT VARIABILITY.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93
5.2.1 Variations in EffluentQuality

.............................................. 93 5.2.2 Statistical Parametersand Relationship to Permit Limits

........................... 95 5.2.3 Expressionof Permit Limits

............................................... 96
5.3 ENSURING CONSISTENCY WITHTHEWASTELOAD ALLOCATlON

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96
5.3.1 5.3.2 StatisticalConsiderations of WLAs

.......................................... 96 Typesof Water Quality Models and Model Outputs

............................. 96
5.4 PERMITLIMIT DERIVATION

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98
5.4.1
5.4.2 5.4.3 5.4.4 EPARecommendations for Permitting for Aquatic Life Protection

............................................................ 98 Other Approaches to Permitting for Aquatic Life

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103 SpecialPermitting Requirements

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 104 EPARecommendations for Permitting for Human Health Protection

........................................................... 104
5.5 SPECIAL CONSIDERATIONS IN USE OF STATISTICAL PERMIT LIMIT DERIVATION TECHNIQUES.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 105
5.5.1 Effectof Changes of StatisticalParameterson Permit Limits

....................... 105 5.5.2 Coefficient of Variation

................................................. 106 5.5.3 Number of Samples

................................................... 107 5.5.4 Probability Basis

.................................................... ..110
5.6 PERMlTDOCUMENTATION.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110 5.7 EXPRESSING LIMITSAND DEVELOPING MONITORING REQUIREMENTS

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110
5.7.1 5.7.2 5.7.3 5.7.4 5.7.5 5.7.6 5.7.7 Mass-basedEffluent Limits

............................................... 110 EnergyConservation

................................................... 111 Considerations in the Useof Chemical specific Limits

........................... 111 Considerations in the Useof Whole EffluentToxicity Limits

....................... 112 Selectionof Monitoring Frequencies.

....................................... 113 Analytical Variability

................................................... 113 Antibacksliding

....................................................... 113
5.8 TOXICITY REDUCTIONEVALUATIONS

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114
5.8.1 TRE Guidance Documents

............................................... 114 5.8.2 Recommended Approach for Conducting TREs

................................ 114 5.8.3 CircumstancesWarranting aTRE

.......................................... 117 5.8.4 Mechanismsfor ReceivingTREs

........................................... 118
REFERENCES

................................................................... 121

ix

### Table of Contents (Continued) Section Page

6. COMPLIANCEMONITORING AND ENFORCEMENT

.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123 6.1 INTRODUCTION

.......................................................... 123 6.2 PERMITREQUIREMENTS..

................................................ ..123 6.3 COMPLIANCEMONITORING

................................................ 123
6.3.1 Self-monitoring Reports

................................................. 123 6.3.2 Discharge Monitoring Reports/Quality Assurance

.............................. 124 6.3.3 Inspections

.......................................................... 124
6.4 VIOLATION REVIEW.

....................................................... 124 6.5 ENFORCEMENT

........................................................ ..125 6.6 REPORTING OFVIOLATIONS

................................................. 126
REFERENCES

................................................................... 127
7. CASEEXAMPLES

.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...129 7.1 INTRODUCTION

.......................................................... 129
7.2 CASE1: INDUSTRIAL DISCHARGE

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 129
7.2.1 7.2.2 7.2.3 7.2.4 7.2.5 7.2.6 7.2.7 General SiteDescription and Information

.................................... 129 Effluent Characterization for SpecificChemicals

............................... 129 Effluent Characterization for Whole EffluentToxicity

............................ 131 Determine WasteloadAllocations.

......................................... 132 Develop Permit Limits

.................................................. 132 Determining and Expressingthe Controlling Effluent limit

....................... 133 Comparing Different Limit Development Methods

............................. 133
7.3 CASE 2: POTWDISCHARGE

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 134
7.3.1 7.3.2 7.3.3 7.3.4 7.3.5 7.3.6 7.3.7 General Site Description and Information

.................................... 134 Effluent Characterization for SpecificChemicals

............................... 134 Effluent Characterization for Whole EffluentToxicity

............................ 136 Determine WasteloadAllocations.

......................................... 136 Develop Permit Limits

.................................................. 136 Determining and Expressingthe Controlling Effluent Limits.

...................... 137 Comparing Different Limit Development Methods

............................. 137
7.4 CASE 3: MULTIPLEDISCHARGERS INTO THESAMEREACH

. . . . . . . . . . . . . . . . . . . . . . . . . 137
7.4.1 7.4.2 7.3.3 Effluent Characterization

................................................ 137 TMDLs and WLAs

................................................... ..138 Permit Limit Development.

.............................................. 139
INDEX

.............................................................................. 140

X

### APPENDICES Section
Appendix A-1: Appendix A-2: Appendix A-3: Appendix B-1: Appendix B-2: Appendix B-3: Appendix B-4:
Appendix B-5: Appendix B-6: Appendix C: Appendix D: Appendix E: Appendix F: Appendix C:
Appendix H:
Appendix I: Page
Toxicity Test PrecisionData

................................................ A-1 Effluent Variability Data

.................................................. A-2-1 Acute to Chronic Ratio Data

.............................................. A-3-1 Summary of CleanWaterAct Provisions

...................................... B-1-1 Policiesfor Toxics Control

................................................ B-2-1 Regulationsfor Toxics Control

............................................. B-3-1 Whole-Effluent Toxicity Permitting Principlesand Enforcement Strategy

............................................................. B-4-1 Quality Control FactSheets

............................................... B-5-1 CaseDecisionson Whole-Effluent Toxicity

.................................... B-6-1 Ambient Toxicity Testing and DataAnalysis.

.................................... C-1 Duration and Frequency

.................................................. D-1 Lognormal Distribution and Permit Limit Derivations

.............................. E-1 Sampling..

............................................................ F-1 The Development of a Biological Indicator Approach to Water Quality-based Human Health Toxics Control.

................................... G-1 ReferenceDose (RfD): Description and Usein Health Risk Assessments

.............................. H-1 ChemicalsAvailable in IRIS.

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . l-1

xi

### ACKNOWLEDGEMENT Permit Requirements
Thepreparation oftherevised Technical Support Document forWater Quality-based Toxic Control beganwith a3-dayconference heldin Williamsburg,Virginia, in December1988. Representatives of EPA Headquarters and Regions, States, privateindustry, municipalities, academia,and various interest groups attended this meetingand providedvaluableinput. The principalauthorsof thisdocumentwere BillSwietlik, James Taft,Jacqueline Romney, Kathryn Smith,JohnCannell, RobertWood,James Pendergast, andRick Brandes ofthePermits Division; Sheila Frace andMargarete Heberof the EnforcementDivision, ElizabethSoutherlandand RichardHealyofthe Assessment andWatershed ProtectionDivision;and CharlesDeb, WarrenBanks, and Robert April of the CriteriaandStandards Division. Listed belowarethe contributors to specific chapters ofthisdocument.
Approaches toWater Quality-based Toxics Control MargaretHeber,U.S.EPA, Enforcement Division KathrynSmith,U.S.EPA, Permits Division

Water Quality Criteria and Standards Charles Delos,U.S.EPA, CriteriaandStandards Division WarrenBanks, U.S.EPA, CriteriaandStandards Division KathyBarylski, U.S.EPA, CriteriaandStandards Division Robert April, U.S.EPA, CriteriaandStandards Division DavidMoon, U.S.EPA, CriteriaandStandards Division Jacqueline Romney, U.S.EPA, Permits Division Appendices

Effluent Characterization Robert Wood,U.S.EPA, Permits Division BillSwietlik,U.S.EPA, Permits Division James Pendergast, U.S.EPA, Permits Division

Exposure and Wasteload Allocation ElizabethSoutherland,U.S.EPA, Assessment and Watershed Protection Richard Healy,U.S. EPA, Assessment andWatershed Protection BillSwietlik,U.S.EPA, Permits Division James Pendergast, U.S.EPA, Permits Division JohnCannell,U.S.EPA, Permits Division

Compliance Monitoring and Enforcement SheilaFrace, U.S.EPA, Enforcement Division TheodoreCoopwood,U.S.EPA, Enforcement Division

Human Health Component ofAllChapters JohnCannell,U.S.EPA, Permits Division KatherineDowell,U.S.EPA, Permits Division WilliamMorrow, U.S.EPA, Permits Division

Case Examples Workgroup BillSwietlik,U.S.EPA, Permits Division James Pendergast, U.S.EPA, Permits Division Charles Delos,U.S.EPA, Criteriaand Standards Division Jacqueline Romney, U.S.EPA, Permits Division

AppendixA: MargaretHeber,U.S.EPA, Enforcement Division AppendixB: U.S.EPA, Permits Division AppendixC: U.S.EPA, Permits Division AppendixD: Nelson Thomas, U.S.EPA, ERL/ORD, Duluth, MN Appendix E: HenryKahnand MarlaSmith,U.S.EPA, Analysis and Evaluation Division AppendixF: U.S.EPA, Permits Division AppendixC: U.S.EPA, Permits Division AppendixH: U.S.EPA, RfD Workgroup

xiii

### EXECUTIVE SUMMARY
The revisedTechnical Support Document for Water Quality-based Toxics Control(TSD)providesStates and Regions with guidance on procedures for usein thewaterquality-based controloftoxic pollutants. It presents recommendations to regulatoryauthorities facedwith the taskof controlling the point sourcedischarge of toxic pollutantsto the Nation’swaters. Thedocument provides guidancefor eachstepin the waterquality-based toxicscontrol process from standards developmentto compliancemonitoring. Both human healthand aquatictoxicity issues areincorporated into the discussions throughout the document. Theoverallapproachin this revised documentprovides additionalexplanations andrationales based on accumulated experience anddataforthe variousrecommendations that were made in the original TSD. Thefollowing isabriefsynopsis ofthe guidanceprovidedin the TSD. Approachesto WaterQuality-basedToxicsControl TheEnvironmentalProtection Agency’s(EPA) surface toxicscontrol regulation,54 FR 23868, June2, 1989,established specific requirementsthat the “integrated” approachbe usedin water quality-based toxicscontrol. The“integrated” approachconsists ofwholeeffluentandchemical-specific approaches asa means of protecting aquatic life and human health. Astechniquesare madeavailable for implementingbiocriteria,theytoo shouldbe integratedinto thewaterquality-based toxicscontrol,thuscreating atriad ofapproaches:whole effluent,chemical-specific, and biological assessments.Eachapproach has its limitations and thus,exclusive useofoneapproach alonecannotensurerequired protection of aquatic life and human health. Theadvantages/ disadvantages of each approachand how the integrated approachcreates an effective toxicscontrol programarediscussed in thetext. Thewhole effluentapproachto toxicscontrol involves the useof toxicity testsandwaterquality criteriafor the parameter“toxicity” toassess andcontroltheaggregate toxicity ofeffluents.New references and informationin supportofthewholeeffluenttoxicityassessment andcontrolapproachhavebeenincludedin Chap ter 1andassociated appendices (e.g.,precision data,justifications foracute-to-chronicratio recommendations, informationon analytical variabilityintoxicitytesting).Thechemical-specific approach to aquatic life toxics control relieson numeric water quality criteria in Statestandards and interpretationsof Statenarrative standards to assess andcontrolspecific toxicantsindividually. Water Quality Standardsand Criteria Wherespecific numerical criteria for a chemical or biological parameter(suchastoxicity) areabsent,compliancewith water quality standards mustbebased on the generalnarrativecriteria and on protection of the designateduses.Formanypollutants, EPA’s recommendedcriteria may be used,or criteria may be developedusingdatafrom the IntegratedRiskInformationSystem, or dataon the toxicologicaleffectsof the pollutant found eitherin the literatureor requiredofadischarger. Aquaticimpactsoccurnot onlyfromthe magnitudeof a pollutant, but alsofromthedurationandfrequency with which criteria areexceeded. EPA’s recommended aquaticlifecriteriafor both individual toxicantsand whole effluenttoxicity arespecifiedas two numbers: the criterioncontinuousconcentrationisapplied asa 4-dayaverageconcentration;and the criterion maximum concentrationisappliedasan 1-houraverage concentration.The frequencywith which criteria areallowedto be exceededdependson site-specific factors asexplainedin thetext. Strictlyspeaking theterm“criteria” means EPA guidanceformally published under the authority of Section304(a)of the Clean Water Act. Thetoxicity levelrecommendations havenot beenso published. However,they representEPA’s carefullydeveloped technical recommendation, and soarereferredto in this document in thesame mannerasothercriteria. EPA’s recommendedcriteria for whole effluent toxicity are as follows: to protectaquaticlifeagainstchroniceffects, the ambienttoxicity shouldnot exceed1.0chronictoxic unit (TUc) to the mostsensitive of at leastthreedifferenttestspecies.Forprotection against acuteeffects, theambienttoxicity shouldnot exceed 0.3 acutetoxic units (TUc) to the mostsensitive of at leastthree differenttestspecies. EPA hasdevelopedrecommendedhuman health criteria,which arecalledreference ambient concentrations(RACs). In the absence of EPA’s recommended criteria,States maycalculateRACs basedon the equationsin the text. In addition, the needfor sediment andbiologicalcriteriain State waterqualitystandards is discussed. Effluent Characterization Thischaptercontainscompletelyrevised effluentcharacterization discussions andrecommendations.It includes streamlinedprocedures(ascomparedto the originalTSD) for predictingthe likely impactsof toxic effluentson aquatic life and human health. Recommendations areprovidedfor determining, eitherwith or without actualeffluentdata,whetheradischarge causes, hasthe reasonablepotential to cause,or contributesto an excursion aboveaState waterqualitystandard.These effluentcharacterization procedures canbeperformedin onestepanddo not include initial screeningfollowed by definitive data generation aswas recommended in theoriginalTSD. Therevised effluentcharacterization procedures forassessingpotential human health impacts now include control of bioaccumulative chemicals. Exposureand WasteloadAllocation Agoalofpermitwritersistodetermine whateffluentcomposition will protect aquatic organismsand human health. Exposure assessment includesananalysis of how muchofthewaterbodyis subjectto the exceedance of criteria,for how long, and how frequently. Thefirststepisto evaluate the effluentplumedispersion. If mixing is not rapidand completeand if Statestandards allowamixing zone,thewasteload allocationalsomustbebased

xivon a mixing zone analysis. Chapter 5 describes the means to assess dilution at the edge of a mixing zone. As with the original TSD, ambient criteria to control acute toxicity to aquatic life may be met within a short distance of the outfall. However, this provision is no longer restricted to outfalls that have a high-rate diffuser.
If mixing is rapid and complete, there are several models that can 

### beusedto assess exposure. Steady-state models 

### assume that the effluent concentration is constant and that the duration and frequency with which criteria are exceeded can be reflected entirely by selecting a design flow in the receiving water of appropriate averaging period and frequency.
Another means of modeling exposure is to use computer models that incorporate variability of the individual inputs (such as effluent flow and concentration, receiving water flow, temperature, background concentration, etc.). These models are termed dynamic models and are more accurate than steady-state models in reflecting or predicting exposure provided adequate data exist. The acceptable effluent condition derived using these models is expressed as the effluent long-term average and variance, which greatly simplifies derivation of permit limits. Three dynamic modeling approaches are described along with instructions for their use. Permit Requirements
The requirements of a wasteload allocation (WLA) must be translated into a permit limit in the wastewater discharge permit. In many cases permit limits will be different than the WLA to reflect different assumptions and means of expressing effluent quality. Three types of WLAs are identified, and recommendations are provided for deriving permit limits to properly enforce each type of WLA. Other permit-related issues such as permit documentation and how to express limitations are discussed. In addition, guidance for requiring and conducting toxicity reduction evaluations is presented. Compliance Monitoring
The compliance monitoring and enforcement process for water quality-based permits summarized in Chapter 6 is based on existing regulation and guidance. As with technology-based permits, any failure to meet a limit is a violation, and every violation must be reviewed to determine the appropriate response. Whole effluent toxicity monitoring and enforcement concepts embodied in the Compliance Monitoring and Enforcement Strategy for Toxics Control (January 19, 1989) have been added to this revision.

xv

## LIST OFABBREVIATIONS

AA ACR ADI AML ATC ATE AVS BAF BAT BCF BCT BMP BOD BPJ BPT
ccc CEAM CETTP CFR CHC CMC CTE
cv CWA DF DMR DO EC ECAO EMS EP EPA ERL FAV FDA FM atomic absorption acute-to-chronic ratio acceptable daily intake average monthly limit acceptable tissue concentration acute toxicity endpoint acid volatile sulfides bioaccumulation factor best available technology bioconcentration factor best conventional technology best management practice biochemical oxygen demand best professional judgment best practicable technology criteria continuous concentration Center for Exposure Assessment Modeling (EPA) Complex Effluent Toxicity Testing Program Code of Federal Regulations chemical of highest concern criteria maximum concentration chronic toxicity endpoint coefficient of variation Clean Water Act dilution factor discharge monitoring report dissolved oxygen effect concentration Environmental Criteria and Assessment Office Enforcement Management System equilibrium partitioning Environmental Protection Agency Environmental Research Laboratory (EPA) final acute value Food and Drug Administration food chain multipliers CC/MS gas chromatograph/mass spectrometer HHC human health criteria HPLC high-pressure liquid chromatography IC IRIS IA LC LOAEL LOEC LTA MCL MDL MERS ML NOAEL NOEC NPDES NTIS ONRW PCS POTW 

### PQL ql* 

### QA/QC QNCR QSAR RAC RfD RWC 

### SQC inhibition concentration Integrated Risk Information System (EPA) load allocation lethal concentration lowest observed adverse effect level lowest observed effect concentration long-term average maximum contaminant levels maximum daily limit Monticello Ecological Research Station minimum level
no observed adverse effect level
no observed effect concentration National Pollutant Discharge Elimination System National Technical Information Service outstanding national resource waters Permit Compliance System publicly owned treatment works practical quantitation limit
cancer potency factor quality assurance/quality control quarterly noncompliance report quantitative structure-activity relationships reference ambient concentration reference dose receiving water concentration sediment quality criteria STORET storage and retrieval of water quality information TIE TMDL TRE TSD TSS TTO TU TUa TUC 

### WQS WLA toxicity identification evaluation total maximum daily load toxicity reduction evaluation technical support document total suspended solids total toxic organics toxic unit acute toxic unit chronic toxic unit water quality standard wasteload allocation

xvii

### MODELING ABBREVIATIONS

ARM CHNTRN CETIS CIS CORMlX1 CTAP DESCON
DFLOW
DYNHYD4 DYNTOX EXAMS-II FCM2 FETRA FGETS FLOSTAT

HHDFLOW HSPF MEXAMS MlNTEQA2 MICH agriculturalrunoffmodel ChannelTransportModel ComplexEffluent Toxicity InformationSystem ChemicalInformationSystem CornellMixing ZoneExpertSystem Chemical TransportandAnalysis Program computer programthat estimates designconditions computer program that calculatesbiologically based designflows hydrodynamicmodel dynamictoxicsmodel Exposure Analysis ModelingSystem WASP FoodChainModel FiniteElement TransportModel FoodandGill Exchange ofToxicSubstances U.S.GeologicalSurveycomputer program that estimates the arithmetic meanflow and 7Q10of riversandstreams historicdailyflow program HydrologicSimulationProgram - FORTRAN MetalsExposure Analysis Modeling System EquilibriumMetalsSpeciation Model Michigan River Model NPS
PSY SARAH2
SERATRA SLSA TODAM
TOXIWASP TOXl4 TOXIC
UDKHDEN
ULINE UMERGE
UOUTPLM
UPLUME
WASP4 WASTOX WQAB FLOW Nonpoint SourceModelfor Urbanand Rural eas steady-state, two-dimensionalplumemodel Arsurface waterassessment modelfor backcalculating reductionsin biotic hazardous wastes Sediment ContaminantTransportModel SimplifiedLake/Stream Analysis TransportOne-Dimensional Degradation and Migration Model Chemical Transport and FateModel asubset ofWASP4 Toxic Organic Transport and Bioaccumulation Model three-dimensional model usedfor singleor multiple port diffusers uniformlineardensityflume model two-dimensionalmodelusedto analyze positively buoyantdischarge cooling tower plume modeladaptedfor marine discharges numericalmodelthat produces flux-averagedilutions waterqualityanalysis program Estuary and Stream QualityModel waterqualityanalysis system flowdatasubroutine

. . XVIII

### GLOSSARY

absolutetoxicity isthetoxicityoftheeffluentwithout considering dilution. acutemeans astimulussevere enoughto rapidlyinduceaneffect; in aquatictoxicity tests,aneffectobservedin 96 hours or less typically isconsideredacute. Whenreferringto aquatictoxicology or human health, an acuteaffectis not alwaysmeasured in termsof lethality. acute-to-chronic ratio (ACR)isthe ratioofthe acutetoxicity of

acutely aneffluentoratoxicantto itschronictoxicity. It isused asafactorfor estimatingchronictoxicity onthe basis of acutetoxicity data,orfor estimatingacutetoxicity on the basis of chronictoxicity data. toxic conditions are those acutely toxic to aquatic organismsfollowing their short-term exposurewithin anaffectedarea. acutetoxicity endpoints (ATE)aretoxicity test results,suchas anLC50 (96hours)andEC50 (48 hours),which describe astimulussevere enoughto rapidlyinduceaneffecton aquaticorganisms. additivity isthe characteristicpropertyofamixture oftoxicants that exhibitsatotal toxic effectequalto thearithmetic sumofthe effects ofthe individual toxicants. ambient toxicity is measuredby a toxicity test on a sample collectedfrom awaterbody. antagonism isthecharacteristic propertyofamixtureoftoxicants that exhibits a less-than-additive total toxic effect. antidegradation policiesare part of eachState’s water quality standards.These policiesaredesigned to protectwater qualityandprovideamethodofassessing activitiesthat mayimpactthe integrity ofthewaterbody. aquatic community isanassociation of interacting populations of aquaticorganismsin agivenwaterbodyor habitat. averaging period isthe periodoftime overwhich the receiving water concentration isaveragedfor comparisonwith criteria concentrations. This specification limits the duration of concentrationsabovethe criteria. bioaccumulation istheprocess bywhichacompoundistakenup byanaquaticorganism,both fromwaterandthrough food. bioaccumulation factor (BAF) is the ratio of a substance’s concentration intissueversus itsconcentration inambient water, in situationswhere the organismand the food chainareexposed. bioassay is a test usedto evaluate the relative potency of a chemicalor a mixture of chemicalsby comparing its effectonaliving organismwith the effectofastandard preparationon the sametype of organism. Bioassays frequently are usedin the pharmaceuticalindustry to evaluatethe potencyofvitaminsanddrugs. bioavailability isameasure ofthe physicochemical access that a toxicant has to the biologicalprocesses ofanorganism. Theless the bioavailabilityofatoxicant,the less itstoxic effecton an organism. bioconcentration is theprocess bywhichacompoundisabsorbed from water through gills or epithelial tissuesand is concentratedin the body. bioconcentration factor (BCF) is the ratio of a substance’s concentrationintissue versus itsconcentrationinwater, in situationswhere the food chain is not exposedor contaminated. For nonmetabolized substances,it represents equilibrium partitioning betweenwaterand organisms. biological assessment isanevaluationofthebiologicalcondition ofawaterbodyusingbiologicalsurveys andotherdirect measurements of residentbiota in surface waters. biological criteria, also known as biocriteria, are narrative expressions or numeric values of the biological characteristics of aquatic communities based on appropriate referenceconditions. Biological criteria serveasan index ofaquaticcommunity health. biological integrity isthe condition of the aquaticcommunity inhabiting unimpairedwaterbodies ofaspecified habitat asmeasuredbycommunity structureandfunction. biological monitoring, alsoknownasbiomonitoring, describes the livingorganisms inwaterqualitysurveillance used to indicate compliance with water quality standardsor effluent limits and to document water quality trends. Methodsof biological monitoring mayinclude,but are not limited to, toxicity testing suchasambienttoxicity testingorwhole effluenttoxicity testing. biological surveyor biosurvey isthe collecting,processing, and analyzing of a representativeportion of the resident aquaticcommunity to determine its structural and/or functional characteristics. biomagnification isthe process bywhichthe concentrationofa compound increases in speciesoccupying successive trophic levels. cancerpotency slopefactor (q1*) isanindicationofachemical’s human cancer-causingpotential derived using animal studiesorepidemiologicaldataon humanexposure. It is basedon extrapolating high-dose levelsover short periods oftimeto low-dose levels andalifetimeexposure periodthrough the useof a linearmodel. chronic means astimulusthat lingersorcontinuesforarelatively long periodoftime,oftenone-tenth ofthe lifespanor more. Chronic should be considereda relativeterm depending on the life span of an organism. The measurementof achroniceffectcanbereduced growth, reducedreproduction, etc., in addition to lethality. chronic toxicity endpoints (CTE) are results, such as a no observedeffectconcentration, lowestobservedeffect concentration, effect concentration, and inhibition concentration based on observations of reduced reproduction, growth, and/or survivalfrom life cycle, partial life cycle,and earlylife stagetestswith aquatic animalspecies.

xix

### coefficient of variation (CV)isastandardstatisticalmeasure of the relativevariation of a distribution or set of data, definedasthestandarddeviationdivided bythe mean. community component isageneraltermthat maypertaintothe biotic guild (fish, invertebrates,algae),the taxonomic category (order, family, genus, species),the feeding strategy (herbivore, omnivore, predator), or the organizationallevel(individual,population,assemblage) of abiological entity within the aquaticcommunity. completely mixed condition means nomeasurable difference in the concentrationofapollutant existsacross atransect ofthe waterbody (e.g.,doesnot varyby 5 percent). continuous simulation model isafateandtransportmodelthat

criteria

criteria usestime seriesinput datato predict receivingwater quality concentrationsin the samechronologicalorder asthat ofthe input variables. continuous concentration (CCC)is the EPA national water quality criteria recommendationfor the highest instreamconcentration of a toxicant or an effluentto which organismscan be exposedindefinitely without causingunacceptableeffect. maximum concentration (CMC) is the EPAnational water quality criteria recommendationfor the highest instreamconcentration of atoxicant or an effluent to which organismscan be exposedfor a brief period of time without causinganacuteeffect. critical life stage isthe periodoftime in anorganism’slifespan in which it isthe most susceptibleto adverseeffects causedby exposureto toxicants, usuallyduring early development (egg, embryo, larvae). Chronic toxicity testsareoften run on critical lifestages to replacelong duration, life-cycle tests since the most toxic effect usuallyoccursduring the critical lifestage. designflow istheflow used forsteady-state wasteload allocation modeling. designated uses arethose uses specified inwater quality standards foreachwaterbodyorsegmentwhetheror nottheyare being attained. discharge length scaleisthe squareroot of the cross-sectional areaofanydischargeoutlet. diversity isthe number and abundanceof biological taxa in a specifiedlocation. effect concentration (EC) is a point estimateof the toxicant concentrationthat would cause an observable adverse effect (such as death, immobilization, or serious incapacitation) in a given percentage of the test organisms. equilibrium partitioning (EP) is a method for generating sediment criteria thatfocuses onthechemical interaction betweensedimentsand contaminants. final acutevalue (FAV) isanestimateoftheconcentrationofthe toxicant correspondingto a cumulative probability of 0.05 in theacutetoxicity values forallgeneraforwhich acceptableacute tests have been conducted on the toxicant. frequency is how often criteria can be exceeded without unacceptablyaffectingthe community. genotoxic istheability ofasubstance to damageanorganism’s geneticmaterial(DNA). harmonic meanflow isthe numberofdailyflow measurements dividedbythesumofthe reciprocals oftheflows. That is,it isthe reciprocalofthe meanof reciprocals. inhibition concentration (1C) isapoint estimateofthetoxicant concentration that would cause a givenpercentreduction (e.g.,lC25) inanonlethalbiologicalmeasurement ofthe testorganisms, suchasreproduction or growth. lethal concentration is the point estimate of the toxicant concentrationthatwould belethaltoagivenpercentage ofthe testorganismsduring aspecificperiod. lipophilic isa high affinityfor lipids (fats). load allocations (LA)arethe portion ofareceivingwater’stotal maximumdailytoadthat isattributed eithertooneofits existing or future nonpoint sources of pollution or to naturalbackgroundsources. tognormal probabilistic dilution model calculates the probability distribution of receiving water quality concentrations from the lognormal probability distributionsofthe input variables. log P (also expressed as log kowor as n-octanal/water partition coefficient) istheratio,inatwo-phasesystem of n-octanol and water at equilibrium, of the concentrationof achemicalin the n-octanol phaseto that in thewaterphase. lowest observed adverse effect level (LOAEL)is the lowest concentrationof an effluentor toxicant that resultsin statistically significant adverse healtheffects as

observed in chronicor subchronichuman epidemiology studies or animalexposure. magnitude ishow much afapollutant (or pollutant parameter suchastoxicity), expressed asa concentrationortoxic unit isallowable. minimum level (ML) refersto the level at which the entire analyticalsystemgivesrecognizablemass spectraand acceptable calibration points when analyzing for pollutants of concern. This level correspondsto the lowest pointatwhichthecalibrationcurveisdetermined. mixing zone isanareawhereaneffluent dischargeundergoes initial dilution and isextendedto coverthe secondary mixing in theambientwaterbody. A mixing zoneisan allocatedimpactzonewherewaterquality criteriacan be exceededas long asacutely toxic conditions are prevented. Monte Carlosimulation isastochastic modelingtechniquethat involvesthe randomselectionof setsof input datafor use in repetitive model runs in order to predict the probability distributions of receiving water quality concentrations.no observed adverse effect level (NOAEL) isa tested dose of an effluent or a toxicant below which no adverse biological effects are observed, as identified from chronic or subchronic human epidemiology studies or animal exposure studies.
no observed effect concentration (NOEC) isthe highest tested concentration of an effluent or a toxicant at which no adverse effects are observed on the aquatic test organisms at a specific time of observation. Determined using hypothesis testing. nonthreshold effects are associated with exposure to chemicals that have no safe exposure levels (i.e., cancer). permit averaging period is the duration of time over which a permit limit is calculated (days, weeks, or months). persistent pollutant is not subject to decay, degradation, transformation, volatilization, hydrolysis, or photolysis. priority pollutants are those pollutants listed by the Administrator under CWA Section 307(a). probability is a number expressing the likelihood of occurrence of a specific event, such as the ratio of the number of outcomes that will produce a given event to the total number of possible outcomes. probability distribution is a mathematical representation of the probabilities that a given variable will have various values. practical quantitation limit (PQL) is a correction factor, sometimes arbitrarily defined, used to account for uncertainty in measurement precision. reasonable potential is where an effluent is projected or calculated to cause an excursion above a water quality standard based on a number of factors including, as a minimum, the four factors listed in 40 CFR 122.44(d)(l)(ii). receiving water concentration (RWC) isthe concentration of a toxicant or the parameter toxicity in the receiving water after mixing (formerly termed “instream waste concentration” [IWC]).
recurrence interval isthe average number of years within that a variable will be less than or equal to a specified value. This term is synonymous with return period. reference ambient concentration (RAC) isthe concentration of a chemical in water that will not cause adverse impacts to human health. RAC is expressed in units of mg/l. reference tissue concentration (RTC) is the concentration of a chemical in edible fish or shellfish tissue that will not cause adverse impacts to human health when ingested. RTC is expressed in units of mg/kg. reference dose (RfD) is an estimate of the daily exposure to human population that is likely to be without an appreciable risk of deleterious effect during a lifetime; derived from nonobserved adverse effect level or lowest observed adverse effect level. relative toxicity isthe toxicity of the effluent when it ismixed with the receiving water, or a dilution water of similar composition for toxicity testing. slug flow sampling is a monitoring procedure that follows the same slug of wastewater throughout its transport in the receiving water. Water quality samples are collected at receiving water stations, tributary inflows, and point source discharges only when a dye slug or tracer passes that point. steady-state model is a fate and transport model that uses constant values of input variables to predict constant values of receiving water quality concentrations. STORET is EPA’s computerized water quality data base that includes physical, chemical, and biological data measured in waterbodies throughout the United States. sublethal means a stimulus below the level that causes death. synergism isthe characteristic property of a mixture of toxicants that exhibits a greater-than-additive total toxic effect. threshold effects result from chemicals that have a safe level (i.e., acute, subacute, or chronic human health effects). total maximum daily load (TMDL) is the sum of the individual wasteload allocations and load allocations. A margin of safety is included with the two types of allocations so that any additional loading, regardless of source, would not produce a violation of water quality standards. toxicity identification evaluation (TIE) is a set of procedures to identify the specific chemicals responsible for effluent toxicity. toxicity reduction evaluation (TRE) is a site-specific study conducted in a stepwise process designed to identify the causative agents of effluent toxicity, isolate the sources of toxicity, evaluate the effectiveness of toxicity control options, and then confirm the reduction in effluent toxicity. toxicity test isa procedure to determine the toxicity of a chemical or an effluent using living organisms. A toxicity test measures the degree of effect on exposed test organisms of a specific chemical or effluent. toxics are those pollutants that have a toxic effect on living organisms. TheCWASection 307(a)“priority” pollutants are a subset of this group of pollutants. toxic pollutants are those pollutants listed by the Administrator under CWA Section 307(a). toxic units (TUs) are a measure of toxicity in an effluent as determined by the acute toxicity units or chronic toxicity units measured. toxic unit acute (TU,) is the reciprocal of the effluent concentration that causes 50 percent of the organisms to die by the end of the acute exposure period (i.e., 100 GO). toxic unit chronic (TU,) is the reciprocal of the effluent concentration that causes no observable effect on the test organisms by the end of the chronic exposure period (i.e., 1OO/NOEC). water quality assessment is an evaluation of the condition of a waterbody using biological surveys, chemical-specific analyses of pollutants in waterbodies, and toxicity tests.

xxiwasteload allocatlon @WA) isthe portion of a receiving water’s total maximum daily load that is allocated to one of its existing or future point sources of pollution. water quality criteria are comprised of numeric and narrative criteria. Numericcriteriaarescientificallyderivedambient concentrations developed by EPA or States for various pollutants of concern to protect human health and aquatic life. Narrative criteria arestatements that describe the desired water quality goal. water quality limited characterizes a stream segment in which it is known that water does not meet applicable water quality standards, and/or is not expected to meet applicablewaterqualitystandardsevenafterapplication of technology-based effluent limitations. water quality standard is a law or regulation that consists of the beneficial designated use or uses of a waterbody, the numeric and narrative water quality criteria that are necessary to protect the use or uses of that particular waterbody, and an antidegradation statement. whole effluent toxicity is the total toxic effect of an effluent measured directly with a toxicity test.

xxii

### INTRODUCTION Purpose Thepurposeofthis revisedTechnical Support Document (TSD) for Water Quality-based Toxics Control isto providethe mostcurrent proceduralrecommendations and guidancefor identifying,analyzing,and controlling adverse waterquality impactscausedby toxic discharges to the surface watersofthe United States.The originalTSD waspublishedin September 1985. Since then, the CleanWaterAct(CWA) wasamendedin 1987with anemphasis on controlling toxic pollutants. New policiesand regulations have been promulgatedand a vastamount of knowledgeand experiencedhasbeengainedin controlling toxic pollutants. Because ofthesechanges,EPA revised and updatedtheTSD. Thisguidancedocumentisintendedto supportthe implementation ofthe CWA waterquality-based approachto toxicscontrol. Assuch,the recommendations andguidancefound inthisdocument arenot binding and shouldbe usedby regulatoryauthoritieswith discretion. The guidance in this document hasbeen developed as

the mostcurrentrepresentation ofknowledgeinthe fieldof assessment andcontrol oftoxic discharges.Some ofthe guidance in this document is basedon ongoing research and development(bioaccumulationmethods,Chapter3)andshould not beuseduntil theprocedures arefinalized. Background The EPA surface watertoxics control program, representeddiagrammaticallyin the figure, relieson portions of the national pretreatment program, the effluent limitations guidelines program,thesludgeprogram, thecombinedsewer overflowprogram, the stormwatermanagementprogram,the 304(l) program,the waterqualitystandards program,andthe NationalPollutantDischargeElimination System (NPDES) program.States areauthorized by EPA to implementcertainportionsofthe nationaltoxicscontrol program,suchasthe NPDES program. Scientific andtechnical guidance is developedand published by EPA to assist the States.EPA isrequiredbythe CWA andfederalregulations to play an oversightrole to ensurethat States authorizedto implement variousprogram requirements do soin accordance with federal regulations. States aregivendiscretionintheCWA to establish andimplement water quality standards. Assuch,there may be differences in toxicscontrol programs betweenStates.EPA’s oversightroleisto ensurethat eachState’sprogram istechnicallysound and that eachState fully implementsitsprogram. Throughoutthe evolutionofthetoxicscontrol program,EPA has providedguidanceconcerningnewprograminitiatives,statutory developments, and regulatoryrequirements. In 1980, EPAemphasized in itspreamble to NPDES regulations(45 FR 33520)that NPDES permitlimitationsmustreflect the moststringentoftechnology-based, water quality-basedcontrols,or other standards requiredbythe CWA(e.g.,oceandischargerequirementsunder Section403 and toxics standardsor prohibition under Section 307[a]). EPA reiteratedthe significanceof surface water toxics control in 1984 through the publication of its national policy statement entitled,“Policyforthe Development ofWaterQualityBased PermitLimitations forToxicPollutants”(49 FR 9016,March 9, 1984). EPA recommended the useof“biologicaltechniques as a complement to chemical-specific analyses to assess effluent discharges and expresspermit limitations” (49 FR 9017). The preambleto additional regulationspromulgatedin 1984 (49 FR 37998) stressed the importanceof establishingeffluent limitations in NPDES permitsto control toxic pollutants. Regulatory provisions promulgated onJune 2,1989(54FR 23868),clarifyEPA’s surface watertoxicscontrolprogramandtheuseofwholeeffluent toxicity, and implement CWA Section 304(l) concerning the identification ofimpaired waters andthedevelopment ofindividual controlstrategies. The control of toxic dischargesto the Nation’s waters is an important objectiveof the CWA. To effectivelyaccomplishthis objective,EPA recommends the useofan integratedwaterquality-basedapproachfor controlling toxic discharges. EPA’sintegrated“standards to permits”approach,illustratedin thefigure, startswith water quality criteria,objectives,and standardsand results in NPDES permit limitsto controltoxic pollutantsthrough the use of both chemical-specific and whole effluent toxicity limitations. Limitationsareessential for controlling the discharge of toxic pollutantsto the Nation’swater. Once NPDES permit limitsareset,complianceisessential.Compliancecanbeascertainedbycontinual routine monitoring ofeffluentquality. Water quality-based effluent limitationswhen developedin accordance with the procedures in thisdocument,will protectwaterquality andpreventtheviolationof State waterquality standards.

. . . XXIII

### Overviewof the Water Quality-based“Standardsto Permits”Process for ToxicsControl

xxiv

# 1. APPROACHES TO WATER QUALITY-BASED TOXICS CONTROL

1.1 INTRODUCTION

In this chapter, basic principles are presented that cover the protection of aquatic life and the protection of human health from impacts caused by the release of toxics to the Nation’s surface waters. Protection against toxic releases iscalled for under Section 101(a)(3) of the Clean Water Act (CWA), which states that “it is the national policy that the discharge of toxic pollutants in toxic amounts be prohibited.” In addition, CWA Section 303(c) requires States to develop water quality standards to protect the public health or welfare, enhance the quality of water, and serve the purposes of the CWA. The control of the discharge of toxics is a paramount objective of the National Pollutant Discharge Elimination System (NPDES) and water quality standards programs. The CWA and Environmental Protection Agency (EPA) regulations (described in Appendices B-1 and B-4, respectively) authorize and require the use of the “integrated strategy” to achieve and maintain water quality standards. In addition, EPApolicy and guidance have long advocated this approach (see Appendices B-2 and B-3). For the protection of aquatic life, the integrated strategy involves the use of three control approaches: the chemical-specific control approach, the whole effluent toxicity control approach, and the biological criteria/bioassessment and biosurvey approach. However, for the protection of human health, technical constraints do not yet allow for full reliance on an integrated strategy, and thus primarily chemical-specific assessment and control techniques should be employed.
The integrated approach to water quality-based toxics control, including the use of toxicity testing and whole effluent toxicity limits, chemical-specific testing and limits, and biological criteria using bioassessments/biosurveys, relies on the water quality standards that each State has adopted. All States have water quality standards consisting of both chemical-specific numeric criteria for individual pollutants, and narrative “free from toxics in toxic amounts” criteria. Currently, a few States have incorporated biological criteria into water quality standards.
The narrative water quality criteria in all States generally require that the State waters be free from oil, scum, floating debris, materials that will cause odors, materials that are unsightly or deleterious, materials that will cause a nuisance, or substances in concentrations that are toxic to aquatic life, wildlife, or human The use of toxicity testing and whole effluent toxicity health. limits is based upon a State’s narrative water quality criterion and/ or in some cases, a State numeric criterion for toxicity. Chemical-specific numeric criteria have been adopted by each State. In many cases, States have adopted EPA-recommended water quality criteria as a part of their water quality standards [1, 2]. (See Chapter 2, Water Quality Criteria and Standards, for further information.) These State-adopted numeric chemical criteria provide the basis upon which specific chemicals can be limited in permits. Where States have not developed chemicalspecific numeric criteria, States may interpret their narrative standards for specific chemicals by using EPA criteria updated with current quantitative risk values. Biological criteria provide a direct measure of ambient aquatic life and overall biological integrity in a waterbody. Biological criteria constitute one basis for limits that will protect the biological integrity of a surface water.
The integrated approach must include the control of toxics through implementation of the narrative “no toxics” criterion and/or numeric criteria for the parameter toxicity, the control of individual pollutants for which specific chemical water quality criteria exist in a State’s standards, as well as use of biological criteria. Reliance solely on the chemical-specific numeric criteria or the narrative criterion or biological criteria would result in only a partially effective State toxics control program. In the discussion that follows, each control approach is described in greater detail as well as how each of the approaches complement the other two by providing additional information for the protection of water quality.



## 1.2 CHEMICAL-SPECIFIC APPROACH FORAQUATIC LIFE PROTECTION

The chemical-specific approach to toxics control for the protection of aquatic life uses specific chemical effluent limits in NPDES permits to control the discharge of toxics. These limits are developed from laboratory-derived, biologically based numeric water quality criteria adopted within a State’s water quality standards. Water quality criteria are adopted by a State for the protection of the designated uses of the receiving water. Chemical-specific water quality-based limits in NPDES permits involve a site-specific evaluation of the discharge and its effect upon the receiving water. This may include collection of effluent and receiving water data and result in the development of a wasteload allocation (WLA) and a total maximum daily load (TMDL) through modeling, a mixing zone analysis, and the calculation of permit limits. Once a numeric water quality criterion is adopted, chemical-specific limits must be developed in NPDES permits to ensure that a permittee’s discharge does not exceed acute or chronic water quality criteria for the pollutant in a receiving water if there is a reasonable potential for that discharge to cause or contribute to excursions of the criterion. These steps are discussed in Chapters 3, 4, and 5.
EPA water quality criteria for the protection of aquatic life are developed under the requirements of CWA Section 304(a)(1) and are published by EPA in separate criteria documents and summarized in the Quality Criteria for Water [1]. Water quality criteria are derived scientifically and attempt to consider a wide range of toxic endpoints including acute and chronic impacts and

1

### bioaccumulation. Each criteriaconsists of two values-an acute and a chronic value. Criteria are developed using the latest scientificknowledgeon the kindandextentof identifiableeffects on organisms, suchasplankton,fish,shellfish, wildlife,and plant life,which maybe expectedfrom the presence of pollutants in anybodyofwater. Waterqualitycriteriaalsoreflectthe concentration and dispersal of pollutants,or their byproducts,through biological,physical,and chemicalprocesses, and the effectsof pollutantson biological community diversity,productivity, and stabilityofthereceiving water[1] Theycanbeused toassess and control avarietyofwaterquality impacts. Chapter2 providesa more detailed discussionof the derivation of numeric criteria. Recommendations for usingchemical-specific datato determine which individual toxicants need to be controlled arefound in Chapter3. Legalrequirements, including chemical-specific limits in permits,arefound in Chapter5.

1.2.1 Correlation ofChemical-specific Measurements toActual Receiving Water impacts EPA hasconducteda series of studiesto determinewhether its waterquality criteriaconcentrations areprotectiveofaquaticlife in receivingwater systems. The first study was conducted at ShaylerRun, Ohio, to evaluatethe applicability of laboratorygenerated toxicity datato a naturalstreamartificiallydosedwith copperto providesteadyconcentrations[3]. The resultsof the study indicate that severalcharacteristics of site-specific water qualityaffect thetoxicity ofcopper. Theresults alsoindicatethat avoidance ofelevated concentrationareas by instream organisms can produce observableecological changesat concentrations belowthosefound to be harmfulin laboratorytoxicity tests. No instreameffects wereobserved at continuousexposureconcentrationsnearEPA’s currentchroniccriterion,appliedatthewater hardness of Shayler Run. Studiesperformedon experimentalstreams at EPA’s Monticello Ecological Research Station(MERS) indicategood agreementbetween EPA’s criteriaconcentrations and the instreamconcentrations producingaquaticlifeeffects understeadyexposureconditions [4-13]. EPA’s waterquality criteriaarenot thresholdlevels abovewhich definitemeasurable instreameffects arealwaysexpected. Rather,the criteriaembody conservative assumptions suchthat smallexcursions abovethe criteriashould not resultin measurable environmental impactsupon the biota. The data indicatethat iftheambientwaterqualitycriteriaaremet,thenthe biota in the receivingwatersystem will be protectedfrom unacceptableimpacts caused bythechemical ofconcern.Thestudies conductedbyMERS are described in greater detailin Box1-1and Tables1-1and 1-2.

1.2.2 Chemical-specific Analytical Method Precision Tables1-3to 1-5illustratethetypesof precisioncommonlyseen in inorganic,organic,and nonmetalinorganicchemicalanalyses that areroutinely used fordeterminingconcentrations ofspecific pollutantsineffluents.These tables showtheobserved variability. Thevariability of chemicalmeasurements increases asone approaches thelimit ofdetectability forachemical.Table1-3shows the interlaboratoryprecisionof 10 metals. The coefficientof variation(CV),definedasthe standarddeviationdivided by the mean x 100, for these analysesranges from 18percent to 129percent[15]. Table1-4showsthe interlaboratoryprecision

Box 1-1. Correlation of Chemical-specific Criteria to Instream Impacts In studying the field applicability of EPA’s water quality criteria in freshwatersystems,MERS (Monticello EcologicalResearch Station) conducted studiesin experimentalstreams[4-14] to determinethe level of protection provided by the individual chemicalcriteria. Each of the streams wasone-quartermile long with alternating mud-bottomedpoolsand rockyriffles. Fishwerestockedinto the streams to a known population densitywhile otherplantsandanimals werethe resultof naturalcolonization. Thechemicals studiedwereammonia,chlorine,chlorinecombinedwith ammonia,selenium, andpentachlorophenol. Some studies wereconductedduring asummer(pentachlorophenol) whileotherscontinuedfor more than 2years(seleniumIV). Tables1-1and 1-2showsampledataon ammoniaandammoniacombinedwith chlorine. In all experiments, the streams weredosedcontinuouslywith the chemical(s) beingstudiedand the biologicaleffects weredeterminedstatisticallyby a comparison to the control streams.Theconcentrationat which biologicaleffects occurredwerethen comparedto the EPA criteriacontinuousconcentration(CCC)for thatcompound. Withthe exceptionof chlorine in the presence ofammonia,the datafromthe otherexperimentsindicatethat slight or no effects werefound in the streams at the CCC. This indicatesthat the CCCis providing chronic protectionat the recommended concentrationfor that particularchemical. In the case of chlorinecombined with ammonia,asubstantial impactwas found,but onlyononespecies, thechannelcatfish. Because theCCCis designed to protect most,but not allofthe species allofthetime (see discussion in Chapter2on EPA Ambient WaterQualityCriteria),slight impactsmaybeexpectedundercontinuousexposure conditions.

2Table l-3. Interlaboratory Precision of Inorganic Analysis at the Low End of the Measurement Detection Range [15] Table l-l. Effects in Streams Exposed to Ammonia [8-l 31

Indicator Effects Criteriaa 3Xb 1 9xc Analyte No. of Labs

Aluminum 37 Cadmium 63 Chromium 72 Copper 86 Iron 78 Lead 64 Manganese 55 Mercury 76 Silver 50 Zinc 62 Fish Fathead minnow Bluegill Channel catfish White sucker Rainbow trout Walleye I I 0 Od 0
+ 0 0 0 0 0
++ 0 0 0 ++ i ii+ 0
++
++ 46 129

79 18 118 Benthic Invertebrates 0 

### + ii

Zooplankton 0 

### + +

Notes a Criteria = 0.05 mg/l unionized ammonia (NH3) at average stream pH and temperature; 1.0 mgll total ammonia was added to reach this concentration; concentratcons of unionized ammonia varied daily and seasonally due to natural pH and temperature fluctuations. b 3X = Three times criteria concentration based on input of 3 mg/l total ammonia. c 9X = Nine times criteria concentration based on input of 9 mg/l total ammonia. d 0 = No difference from controls; +‘s represent gradation of differences from

controls ranging from slight(c) to dramatic (++++). Table 1-4. Interlaboratory Precision Ranges for Organic Chemical Analysis

EPA Document Referenced Chemical NO. abs % Data Discarded’ Table l-2. Effects in Streams Exposed to Ammonia and Chlorine [8-l 31

I Effects Benzene 4 Chlorobenzenes Ethyl benzene Toluene 23 Halocarbons 4 Halocarbons 11 Phenols 20 31-64

20 20 20

17 16-29 40-50 2045 38-64 38-69

16 ? 17 ? 22 -12-45

? 16-91 10

? ? 20

? ?

22 19

?

? 600/54-84-064

600/54-84-064

600/54-84-044

600/54-84-062

600/54-84-056 600/54-84-051 600/54&l-061

600/54-84-063 Indicator 4 ugila 35 ug/l -122 ug/l

FishChannel catfish Bluegill Benthic invertebrates Zooplankton Bacteria Periphyton Primary production Litter decomposition Aquatic plants 

## I I b ++ ++ 0
+ 0
++ 0 0
+ 0 +-I-+ 0 ii 0
+++ 0 0 ii 0 0 

### 0 i y 0 0 I + 0 Benzidine 3,3-Dichlorozidine 6 Pthalate esthers 3 Nitrosamines 24 Organochlorine Pesticides and PCBs 16 PNAs

l Discarded as outliers.

It is important to note that in many chemical analyses a decision may be made that certain anomalous data points, or outliers. are unusable and are not reported as valid data points. This type of data evaluation is made because in chemical analyses it is routine to repeat the analysis with the same sample and reference standard until an acceptable result ISobtained Notes a Average concentrations of TRC in presence of Zmg/l to 3mg/l total ammonia;

national criteria for chlorine = 11 ug/l. b 0 = No difference from controls; +‘s represent gradation of differences from

controls ranging from slight (+) to dramatic (++++).

3

### -- ~--

Table 1-5. Interlaboratory Precision of Nonmetal Inorganic Analyses Over the Measurement Range [15]

No. Lab Parameter CV 

### (%) Range

17 Alkalinity 4.9-l 4 >20 Residual chlorine 13-25

16 Ammonia nitrogen 15-58

6 Kjeldahl nitrogen, total 38-41 1s NO3 nitrogen 17-61

6 Total P 25-40 58 BOD 15-33 58 COD 6.9-34 21 TOC 4.6-70

associated with organic chemical analyses. The CVs range from 12 percent to 91 percent. Table l-5 demonstrates the interlaboratory precision of nonmetal inorganic analyses at the lower end of the measurement range. The CVs for this type of analyses range from 4.6 percent to 61 percent [15]. The data in Tables l-3 to l-5 reflect testing in reagent grade water. Actual CVs from testing effluents can be higher due to matrix effects. However, in 40 CFRPart 136 analytical methods, matrix effects are acknowledged.



## 1.3 WHOl.EEFFl.UENMPPROACHFORAQUATfCUFE PROTECTION

The whole effluent approach to toxics control for the protection of aquatic life involves the use of acute and chronic toxicity tests to measure the toxicity of wastewaters. Whole effluent toxicity is a useful parameter for assessing and protecting against impacts upon water quality and designated 

### uses caused by the aggregate toxic effect of the discharge of pollutants [16]. Whole effluent toxicity tests employ the use of standardized, surrogate freshwater or marine (depending upon the mixture of effluent and receiving water) plants, invertebrates, and vertebrates. EPAhas published extensive written protocols listing numerous marine and freshwater species for toxicity testing [ 17, 18, 191.
An acute toxicity test is defined as a test of 96-hours or less in duration in which lethality is the measured endpoint. A chronic toxicity 

### test is defined as a long-term test in which sublethal effects, such asfertilization, growth, and reproduction, are usually measured, in addition to lethality. Traditionally, chronic tests are full life-cycle tests or a shortened test of about 30 days known as an early life stage test. However, the duration 

### of most of the EPA chronic toxicity tests have been shortened to 7 days by focusing on the most sensitive life-cycle stages. For this reason the EPA chronic tests are called short-term chronic tests. 80x l-2 summarizes the short-term chronic tests currently recommended by EPA. The acute and short-term chronic methods recommended by EPA are presented in three methods manuals [17, 18, 19]. In a laboratory acute toxicity test, an effluent sample is collected, diluted, and placed in test chambers with the chosen test species. After 24, 48, 72, and 96 hours, the number of live organisms remaining in each test concentration and in a control is recorded. In a laboratory chronic toxicity test, an effluent sample is collected, diluted, and placed in test chambers. An example of a dilution series used in chronic or acute tests is 100, 50, 25, 12.5, and 6.25 percent, and a control. Test organisms are placed in these test chambers for specified periods of time. At various times during the exposure period, the organisms in each chamber are observed. In the short-term chronic tests, at test termination, the lowest effluent concentration that causes a significant adverse impact on the most sensitive endpoint for that test is calculated (this endpoint can be mortality, reduced fertilization, lower fecundity, reduced growth, etc.). In the acute tests, at test termination, the number of dead organisms are recorded and an LCso is cdlculated.
Dilution water is dn important part of toxicity testing. Dilution water may either be standard laboratory water and/or the receiving water. Sometimes the receiving water is used to dilute the effluent because it more closely simulates effluent/receiving water interactions. This may be especially important in the case of saline receiving waters. The salinity of the receiving water should be matched as closely as possible to the salinity in the test chambers (within the salinity range constraints of a particular method) for the purposes of conducting the tests. Quality control and quality assurance are an integral part of whole effluent toxicity testing. Use of a standard control water and a reference toxicant test are both recommended to ensure quality assurance in chronic testing. It is important to understand that each of the chronic tests has minimum criteria of acceptability for each endpoint that is measured in the controls (i.e., 80 percent survival and minimum criteria for growth, reproduction, and fertilization). The acute tests also have criteria of acceptability measured in the controls.
Acute toxicity endpoints (ATEs) commonly include 

### lethalconcentrations (LCs) and are described in terms of effluent concentrations. The LC is the concentration of toxicant at which a certain percentage of the test organisms die, e.g., the LC10 or LCSO. An exposure duration also is included in the endpoint such as 24, 48, 72, or 96 hours (e.g., 96-hour lC50). Commonly used chronic toxicity endpoints (CTEs) include the no observed effect concentration (NOEC), the lowest observed effect concentration (LOEC), and the effect concentration (EC). The NOEC is the highest concentration of toxicant, in terms of percent effluent, to which the test organisms are exposed that causes no observable adverse effect. The effects measured may include decreases in reproduction and growth, or lethality. The LOEC is the lowest concentration of toxicant to which the test organisms are exposed that 

### causes an observed effect. Again, the same effects are usually observed. The EC isthe toxicant concentration that would cause an adverse effect upon a certain percentage of the test organisms, (e.g., EC10 or EC&.
In chronic toxicity tests, the exposure duration in the EPAtesting protocols is almost always assumed to be the 7-day short-term period unless otherwise specified in the protocol. For example, the Cerioduphnia test must be continued until at least 60 percent

496 hours

1.5 hours

7-9 days

7 days

7 days

7-9 days 

### Box l-2. Short-term Chronic Toxicity Methods Species/Common Name Test Duration Test Endpoints
Freshwater Species
Ceriodaphnia dubio Cladoceran Approximately 7 days (until 60 percent of control have 3 broods) Survival, reproduction

Pimepholes promelos Fathead minnow 7 days Larval growth, survival

Pimephaies promelas Fathead minnow 7-9 days Embryo-larval survival, percent hatch, percent abnormality

5e/enustrum copricwnutum Freshwater algae Growth

Marine/Estuarine Species

Arbociu punctulota Sea urchin Fertilization

Chompiu paw/a Red macroalgae
Mysidopsis bahia Mysid Cystocarp production (fertilization)
Growth, survival, fecundity

Cyprinodon variegotus Sheepshead minnow Larval growth, survival

Cyprinodon voriegotus Sheepshead minnow Embryo-larval survival, percent hatch, percent abnormality

Menidia bery//ina Inland silverside 7 days Larval growth, survival

of the females produce three broods. This may require more or less than 7 days to occur.
It is useful to note that LCs and ECsare point estimates statistically derived from a mathematical model that assumes a continuous dose-response relationship. NOECs and LOECs, statistically determined using hypothesis testing, are not point estimates [18]. In order to overcome the difficulty in statistically deriving the NOEC using hypothesis testing, a new statistical procedure has been developed. This procedure, referred to as the inhibition concentration (IC), is a point estimate interpolated from the actual effluent concentrations at which measured effects occurred during a chronic test. The IC is an estimate of the toxicant concentration that would cause a given percent reduction in a biological measurement of the test organisms, including reproduction, growth, fertilization, or mortality. For example, an lC25 for reproduction would represent the effluent concentration at which a 25percent reduction in reproduction occurred.
Since the IC is a point estimate, a CV can be calculated. A CV cannot be calculated if hypothesis testing is used because results are only available for the effluent concentrations used. For thisreason, estimates of test precision cannot be calculated for NOECs derived by hypothesis testing.
The IC also is not dependent upon the selection of the effluent concentrations. In contrast, NOECs calculated by hypothesis testing are dependent upon the concentrations initially selected. For example, if a chronic test is conducted using 100, 50, 25, 12.5, and 6.25 percent effluent concentrations, and the LOEC exhibited by the data is at 25 percent effluent, the NOEC calculated by hypothesis testing is estimated to be the next lowest dilution, or 12.5 percent. However, the true NOEC value may lie somewhere between 25 percent and 12.5 percent effluent. Comparisons of both types of data indicate that an NOEC derived using the IC25 isapproximately the analogue of an NOEC derived using hypothesis testing (see Figure l-1). For the above reasons, if possible, the IC25 is the preferred statistical method for determining the NOEC.
Another important issue in conducting both acute and short-term chronic toxicity tests is the dilution series. The EPA methods manuals recommend six dilutions, including the control. The only exception to this is a toxicity test conducted on ambient receiving waters. Then, each ambient receiving water is compared statistically to the control without dilutions. It is not accurate to assume that two dilutions (the receiving water concentration [RWC] and control) are all that are ultimately necessary for determining compliance with a toxicity limit. If the toxicity tests are conducted with only the control and one effluent concentration (i.e., the RWC), the error and variabilitv associated with this type of statistical analysis is large [20]. *

20

10

k 4.3%
0 n=l ‘Cl0 26% n=6 

# ni 34.8% n=8

‘C20 For the above reasons, EPA recommends the use of five effluent concentrations and a control to determine the magnitude of toxicity. When conducting compliance monitoring, an option is to choose the five concentrations that bracket the RWC (two concentrations above and two below). This would result in the determination of compliance status as well as a statistically valid estimation of the NOEC. The information provided from the full dilution series would indicate how close the test endpoints are to the permit limit and how close to violating the limit the discharger is, and, if measured over time, the variability of the effluent.



### 1.3.1 TOXIC wts Since toxicity involves an inverse relationship to EC (the lower the EC, the higher the toxicity of the effluent), it is more understandable to translate concentration-based toxicity measurements into toxic units (TUs). In this way, the potential confusion involving the inverse relationship is overcome and the permit limit derivation process is better served. The number of toxic units in an effluent is defined as 100 divided by the EC measured: TU, = 100/LCsO TU, = 1OO/NOEC.
For example, an effluent with an acute toxicity of an LCso in 5 percent effluent is an effluent containing 20 T&s.
A very important aspect of toxic units is that two different types are used depending on whether acute or chronic aquatic toxicity is measured. The proper expressions for toxic units are TU, and

17.4% 17.4%

T”i ‘C25 GO ‘Go
Figure l-l. This figure represents the percentage of the time the mean NOEC was approximately equivalent to an lC10, ICTS, ICZO, ICa, IC3gl and ICs for all 23 effluent and reference toxicant data sets analyzed. The data sets included short-term chronic toxicity test for Ceriodaphnia dubia, Pimephaks pmmclos (fathead minnows), Arbocia punctulata (sea urchin), Cyprinodon variegatus (sheepshead minnows), and Champia panda (red algae) [21].

6TU,. TU, is the measurement of acute toxicity units and TU, is a measurement of chronic toxicity units. (See the glossary for a definition of these terms.) They are not the same measurement and should not be used interchangeably. Acute and chronic TUs make it easy to quantify the toxicity of an effluent and to specify water quality criteria based upon toxicity. For example, an effluent sample that contains 20 TUcs is twice as toxic as an effluent that contains 10 TUcs.

EPA conducted the Complex Effluent Toxicity Testing Program (CETTP) that examined sites in both freshwater and saltwater systems to investigate whether or not an evaluation of effluent toxicity, when adequately related to receiving water conditions (i.e., temperature, pH, salinity), can give a valid assessment of receiving system impacts on waters that support aquatic biota [22-251. Summaries of these site studies are provided in Box l-3 (freshwater) and Box l-4 (saltwater). In addition, three other studies, presented in Box l-3, were conducted to address this issue: a comparative investigation conducted by the University of Kentucky [26], a second study on the Trinity River in Texas conducted by the University of North Texas [27], and a third study conducted by the North Carolina Division of Environmental Management [28].
It is important to note that in these studies, different objectives were addressed. The CETTP freshwater studies attempted to correlate receiving water chronic toxicity measured by EPAtoxicity tests to instream observed impacts (Figure l-2). The CElTP saltwater studies compared effluent toxicity to ambient receiving water toxicity using dye studies to measure receiving water concentrations of effluent. The North Carolina study compared effluent toxicity to receiving water impact using Ceriodaphnio chronic toxicity tests and receiving stream benthic macroinvertebrates (Figure l-3). The Kentucky study examined the relationship between effluent toxicity tests and instream ecological parameters. The Trinity River study attempted to spatially compare the biological, physical, and chemical water quality and sediment quality of Trinity River reaches above and below the Dallas/Fort Worth area (Figure l-4). Together, these studies comprise a large data base specifically collected to determine the validity of toxicity tests to predict receiving water community impact. In order to address the correlation of effluent and ambient toxicity tests to receiving water impacts, EPA evaluated the results of the studies discussed above [29]. The results, when linked together, clearly show that if toxicity is present after considering dilution, impact will also be present.
Parkhurst et al., were requested by representatives of industrial and municipal discharges to critique the CETTP studies [30]. One major criticism was that the EPA study sites were not selected randomly and therefore the results of the studies cannot be extended to all waters. EPAagrees that the CElTP sites were not selected to represent a statistically valid sampling of all types of waterbodies in the United States. A representative sampling of receiving water would require assessment of more sites than EPA could study in a comprehensive manner. Such a sampling was beyond the capability of EPA’s resources. However, the CETTP and corresponding studies such as the Trinity River study [27] did show unequivocally that a strong correlation exists between toxicity and a biological impact.
EPAbelieves that it is reasonable to assume in the absence of data showing otherwise that this relationship is basically independent



### Box 1-3. Correlation of Toxicity Measurements to Receiving Water Impact (Freshwater)

EPA conducted eight freshwater site studies in which ambient toxicity was compared to the receiving water biological impact. These site studies were a part of the Complex Effluent Toxicity Testing Program (CETTP). Testing was done onsite concurrent with the field surveys. Sites exhibiting biological impacts in Oklahoma, Alabama, Maryland, West Virginia, Ohio, and Connecticut were included. Organisms were exposed to samples of water from various stations and tested for toxicity. Biological surveys (quantitative field sampling of fish, invertebrate, zooplankton, and periphyton communities in the receiving water areas upstream and downstream of the discharge points) were made at these stations at the same time the toxicity was tested to see how well the measured toxicity correlated to the health of the community. These studies have been reviewed and published in the EPA publication series [23, 31-381. Figure l-2 illustrates the data from the CElTP studies. A robust canonical correlation analysis was performed to determine whether or not statistically significant relationships existed between the ambient toxicity tests and instream biological response variables and to identify which variables played an important role in that relationship [29]. Influential variables were then used to classify stations as either impacted or not. Ceriodaphnio dubia productivity and/or Pimephales promelm weight were used as the basis for predicting impact. Fish richness was used to classify streams as impact observed or impact not observed.

7

### -



### Box 1-3. Correlation of Toxicity Measurements to Receiving Water Impact (Freshwater) (continued)

Classification was based on the relative performance of the stations on each stream in the study. Percentiles of the appropriate distribution (normal for toxicity variables, and Poisson for fish richness) were used to set cutoffs for classification. Two-way contingency tables representing stations as impact predicted or not, and impact observed or not were prepared from a variety of cutoffs (percentages). The exact test for independence was performed on each contingency table.
If toxicity test results were used to classify sites as impacted or not (predicted classification) and if a strong relationship does exist between ambient toxicity and biological response, then the classification of stations according to biological response should closely match the predicted classification. Hence, the errors in misclassification should be small. Figure l-2, developed using a 95 percent-95 percent cutoff, shows that false positives (impact predicted but none found} occurred at 7.5 percent of the 80 stations. The probability of getting no more than 7.5 percent false positives under the null hypothesis that there is no relationship between ambient toxicity and biological response is less than p=O.OOl. As discussed above, this is the only definitive error that can be identified in such comparisons. The correct or noncontradictory findings (no measured toxicity but observed impacts) were 92.5 percent of the stations. A variety of other cutoff criteria combinations were evaluated and the number of false positives remained in the 7 percent to 8 percent range. Therefore, a discharger’s chance of being charged incorrectly with causing instream toxicity is low if and onlv if dilution in the receiving water is considered.
A comparative time series study conducted on the Trinity River in Texas that used the same classification method as the CElTP studies also showed a strong relationship between ambient toxicity and instream biological response (Figure l-2). False positives (impact predicted but not observed) had a frequency of 8.3 percent. Overall there was a 91.7-percent accuracy of prediction or noncontradictory findings [ZS], and the probability of a false positive (impact predicted but not observed/impact predicted) ranged from 8 percent to 11 percent in these studies.
Another study conducted by the North Carolina Division of Environmental Management indicated the high accuracy of predicting receiving water impacts from whole effluent toxicity tests. Forty-three comparisons were made between freshwater flowing streams using the Ceriodophniu dubiu chronic test and a qualitative macroinvertebrate sampling. Overall there was 88 percent accuracy of prediction (Figure l-3) [28J.
In addition, another comparative study was conducted in the Kentucky River Basin [26]. This study consisted of a comparative ecological and toxicological investigation of a secondary wastewater treatment plant and 

### measured instream effects at 10 stations including reference sites. The principal objective of the study was to assess downstream persistence of aquatic contaminants, to quantify their effects on structure and function of aquatic communities, and to evaluate the fathead minnow embryo-larval test for measuring instream toxicity and estimating chronic effects on aquatic biota. The results of the study indicate a good predictive correlation between embryo-larval survival and independent ecological parameters, especially species richness of macroinvertebrates. The correlation coefficients for species richness and embryo-larval survival was 0.96, and for embryo-larval survival and diversity, it was 0.93. The estimated toxicity (LCT) correlated closely with the actual percent instream effluent dilution observed at the first downstream station at which no ecological impact was discernable. Using the statistical classification previously described in the CETTP and Trinity River studies, an analysis was conducted on the combined data sets of the CElTP, Trinity River, and Kentucky River Basin data. Because the North Carolina study was based on the Ceriodophnio dubio chronic test and a qualitative macroinvertebrate sampling, the data were not amenable to this type of statistical analysis. This combined analysis is illustrated in Figure l-5. The probability of getting no more than 9.4 percent false positives (impact predicted/impact not observed) when the null hypothesis (no relationship between ambient toxicity and biological response) is less than p=O.O028.

8

### Box 1-4. Correlation of Effluent Toxicity Measurements to Receiving Water Toxicity (Saltwater)

In saltwater systems, as in freshwater systems, receiving water impact should only be seen where receiving water waste concentrations are at or above the effect concentrations. Dilution in marine and estuarine systems may be greater due to large and/or complex mixing than most freshwater systems. As a result, there is a less likely chance for receiving water impacts to be observed in saltwater systems as predicted by toxicity tests. Figure l-6 illustrates the comparison between predictions of saltwater receiving water toxicity and whole effluent toxicity. Toxicity test data from 79 ambient stations (four study sites) were compared to effluent toxicity test results from an isolated discharge at each site. All receiving water toxicity to effluent toxicity correlations are based on dye studies conducted at each of the four sites to determine the actual dilution.
Most of the sites were selected because the discharge was isolated from other point sources and potential impacts from other point sources was anticipated to be negligible. Two of these studies indicated near-field effects, generally within the mixing zone. One study conducted at Fernandina Beach, Florida [25], showed impacts outside the proposed mixing zone. Results of another study (East Greenwich) indicated the existence of poor water quality well beyond the influence of the East Greenwich Sewage Treatment Plant and suggests that other sources (point or nonpoint) may contribute significantly [25, 39, 401. This condition may be typical in some of the more stressed estuaries.
In a total of 79 comparisons, 11 out of 15 (73 percent) of the receiving water samples predicted to be toxic were toxic. This constitutes 14 percent of the total comparisons. Toxicity was not predicted in the receiving water and toxicity was not seen in the receiving water 59 out of 64 times (92 percent). This constitutes 75 percent of the total comparisons.
In 5 percent of the total comparisons there was a false negative prediction, or the toxicity tests predicted no toxicity when the receiving water was toxic [24]. As previously discussed, toxicity is only one possible adverse influence. Since only toxicity is measured, a very high correlation should not be expected necessarily because receiving water biological impacts may be attributed to other sources or factors.
The results of the studies at these four sites indicates a 94 percent accuracy when using the marine and estuarine toxicity tests to predict receiving water impacts. In only 6 percent of the cases did effluent toxicity tests predict receiving water toxicity that was not present (false positive).

of waterbody type. Also, this was not the objective of the CETTP studies. The CETTP purpose was to determine if toxicity and impacts to biological communities are found concurrently in receiving waters. Therefore, EPAdisagrees that this is a reason to conclude that the CElTP studies failed to show the validity of toxicity tests to predict water quality impact.
Another criticism was the studies did not investigate replication of results over time. However, toxicity results cannot be expected to be replicated over time in waters where river flow and other timevariant factors change the degree of ambient toxicity. Indeed, the Kanawa River and Five-Mile Creek data showed that ambient toxicity did not occur at high river flows whereas it was found at low flows; this was an expected result. The objective of the CElTP studies was to see if impact was present when effluent toxicity exceeds the available effluent dilution. This objective was achieved by the studies. Another major criticism was the correlation between toxicity tests and biological impact relied extensively upon maximum impact responses and that correlation was poor when data from high flow events and lesser toxicity discharges (minimal impact responses) were added. EPAacknowledges that impact correlations will be higher where higher toxic impact occurs and lower where impacts are expected to be minimal. Such a response is expected given the complexity of ecosystems and that biological communities and species have different sensitivities to toxicants and may respond differently. Also, higher river dilution will reduce the potential instream impact from effluent toxicity. However, this observation does not disprove that the CElTP and other studies showed a statistically sound relationship to correlate toxicity to the existence of a biological ambient impact. Therefore, EPA still concludes that control of toxicity is a valid approach for protecting ambient water quality.
In addition, other studies confirm that effluent toxicity, when adequately related to ambient conditions, can give a valid assess9

### q Impact not predicted/ Impact observed

El Impact predicted! Impact observed


### q Impact not predicted/ Impact not observed



### n Impact predW& Impact not observed

Figure l-2. Comparison of Ambient Toxicity and lnstream Impact-EPA Study [23, 31-381 67 Impact not predicted/ Impact observed
Impact predeted/ Impact observed

0 Impact not predicted/ Impact not observed
Impact predctedl 

### n Impact not observed

Figure l-4. Comparison of Ambient Toxicity and lnstream Impact-Trinity River [29]

23 0%

No lnstream tomty 

### q predhcted. Impact noted
Instream toxlclty 

### q predated tmpact noted

No anstream toxmty f-J predeted. no Impact noted
lnstream toxwty 

### n predlcted. no Impact ITOted

Figure l-3. Comparison of Effluent Toxicity of Receiving Water Impact Using Ceriodophnio Chronic Toxicity Tests and

Freshwater Receiving Stream Benthic lnvertabrates at 43

Point Source Discharging Sites in North Carolina [28] 

### q Impact not predtiedl Impact observed
Impact predicted/ Impact observed

Cl Impact not predti& Impact not observed



### n Impact predicted’ Impact not observed

Figure l-5. Comparison of Ambient Toxicity and lnstream Impact-EPA Study, Trinity River Study, and Kentucky Study [26]

No amblent toxuty r;l prednted. toxlc!ty observed Predjcted amblent Q toxtoty. toxclty observed
No amblent tox~,ty 0 predzted. no tox~~ty observed
Pred~cled amb,ent 

### n tmutyno IOXlClty observed

Figure l-6. Comparison of Predictions of Receiving Water Toxicity Based on Effluent Toxicity and Ambient Receiving Water Testing in Saltwater Environments: 79 Ambient Stations and 4 Dischargers [24, 25, 39, 401

10ment of receiving water impact [3, 24, 26-29, 39, 411. These studies tested waters other than those studied under CElTP.
It is important to recognize 

### that toxicity caused by contaminants in the effluent, as measured by the whole effluent toxicity tests, is only one of many influences that determine the health of a biological community. Impact from toxics would only be suspected where effluent concentrations after dilution are at or above the toxicity effect concentrations. Influences from substrate differences and physical conditions, such as dissolved oxygen, temperature, channelization, flooding and weather cycles, also can affect the biological community adversely. These other types of influences may be better evaluated by using a bioassessment approach. However, the existence of these other factors concurrently with toxicity does not absolve a regulatory authority from controlling the discharge of toxicity if the State has established a designated use to protect aquatic biota.
The value of the toxicity test is its ability to assess the impact of discharged toxicants independent of effects from other factors. This allows regulatory authorities specifically to identify and control the portion of the impact caused by the discharge. Biological, physical, and chemical factors of the community can influence the actual effects that effluent toxicity may cause in the receiving water, and further emphasize the need for a totally integrated water quality-based approach.



### 1.3.3 Toxicity Test&that precision Like all measurements, toxicity tests exhibit variability. Toxicity test variability can be described in terms of two types of precision-“within” or intralaboratory precision, and round robin or interlaboratory precision. lntralaboratory precision isthe ability of trained laboratory personnel to obtain consistent results repeatedly when performing the same test on the same species using the same toxicant. Interlaboratory precision (or round robin tests) is a measure of how reproducible a method iswhen conducted by a large number of laboratories using the same method, species, and toxicant or effluent. Generally, intralaboratory results are less variable than interlaboratory results.
EPA believes that several toxicity test methods have a precision profile that can be reasonable to evaluate compliance with NPDES permits. The appropriateness of a given method can be determined in a permit proceeding or, in part, by rulemaking. EPA has proposed a range of whole effluent toxicity test procedures in 40 CfR 136 and may promulgate these methods soon. Current data, however, show 

### that the precision profiles of a number of whole effluent toxicity tests is similar to already approved chemicalspecific methods.
Research into the precision of whole effluent toxicity methods by various groups (including EPA) has shown that toxicity test procedures exhibit variability [17-l 8, 19, 42-49). In chronic toxicity tests, variability is measured close to the limit of detection because the endpoint of the test is already at the lower end of the biological method detection range (i.e., an NOEC). This is in contrast to acute toxicity tests where the test endpoint is normally calculated at midrange (i.e., LC50), but issometimes calculated at the lower end of the biological detection range (i.e., LC1). CVs cannot be calculated for NOEC endpoints determined using an analysis of variance (hypothesis testing) because this procedure does not produce a statistical point estimate. However, CVs can be calculated for NOECs if they are determined using the IC statistical procedure, and for EC and LC endpoints because they are all statistical point estimates.
To facilitate the comparability between different NOEC calculations using the lC25 and the analysis of variance (hypothesis testing), Appendices A-l and A-2 list NOEC results in terms of both. In some instances the lC25 could not be calculated based on statistical assumptions and available data. In addition, there are some instances where an lC25 cannot be calculated because there was no toxic effect. In these cases, the CV for a method and reference toxicant was calculated using only data where IC25scould be calculated.
A more detailed discussion of precision can be found in Box l-5. Tables l-6 and l-7 summarize the intralaboratory precision for all 10 EPAshort-term chronic whole effluent toxicity tests and some 

### acute toxicity tests. In addition, Table l-8 summarizes the interlaboratory precision for three chronic test species and two acute test species using a variety of different compounds.
In summary, whole effluent toxicity testing methods can represent practical tests that estimate potential receiving water impacts. Permit limits that are developed correctly from whole effluent toxicity tests should protect aquatic biota if the discharged effluent meets the limits. It is important not to confuse permit limit variability with toxicity test variability. Chapter 5 discusses permit limit variability.



### 1.3.4 conshfmti4uzs lnmhwd wlren kyhumthg the Whole E#lutml ToxkY@ AlDpvoacn An understanding of some basic considerations and toxicological principles is important in order to apply routinely the whole effluent approach to the assessment and control of municipal and industrial effluents. The following sections provide a more indepth discussion of each of these factors and principles. (Chapters 3 and 5 discuss specific details for characterizing an effluent and deriving permit limits.)


### Onsitenmus OfMe ToxMy Testing Comparisons of toxicity data between tests conducted onsite and tests conducted offsite on samples shipped to Environmental Research Laboratory (ERL)-Duluth and (ERL)-Narragansett via airfreight have, with a few exceptions, shown little variation. For many effluents, onsite or offsite test data do not appear to be significantly different. The major consideration is cost. Cost also should be weighed against data needs to make the onsiteioffsite determination.
For example, if the presence in the effluent of nonpersistent compounds (i.e., chlorine or other volatiles) issuspected or known, then the regulatory authority may want to conduct onsite testing. If it is not considered important to the analysis of toxic impact, offsite testing is as acceptable as onsite testing. In general, offsite testing would be acceptable for most effluents except those with volati les. When conducting flow-through toxicity tests which require a continuously pumped sample, onsite testing is strongly recommended. Regardless, cost considerations should not over11

### Box 1-5. Toxicity Test Method Precision

Precision can be described by the mean and relative standard deviation (percent coefficient of variation, or CV=standard deviation/mean x 100) of the calculated endpoints from the replicated toxicity tests. Several factors can affect the precision of the test, including test organism age, condition, sensitivity, temperature control, salinity, pH control, handling and feeding of the test organisms, and the training of laboratory personnel. For these reasons, it is recommended that trained laboratory personnel carefully conduct the tests in strict accordance with the test manuals for acute and chronic toxicity testing. In addition, acute and chronic toxicity testing quality assurance practices should be fully performed. Simple quality assurance procedures, which are described at the beginning of each manual, include:

l Single laboratory precision determinations, using reference toxicants, on each of the tests procedures to determine the ability of the laboratory personnel to obtain consistent, precise results. These determinations should be made before attempting to measure effluent toxicity, and routinely confirmed as long as routine whole effluent toxicity tests are being conducted.

l Use of reference toxicants to routinely evaluate the quality and sensitivity of the test organisms to be used in each test.

l Development of “control charts” should be prepared for each reference toxicant/organism/protocol combination to determine if the results are within prescribed limits. The control chart consists of successive data added with each reference toxicant test, and is the basis for evaluating data once the control chart” is established.

l The minimum criteria of test acceptability specific for each protocol. Guidelines for recommended quality assurance practices are found in each manual [17, 18, 191. Within-laboratory precision data are routinely calculated on a minimum of two reference toxicants as part of the EPA methods development process. These data have been established for each of the four EPA freshwater chronic methods and each of the six marine/estuarine chronic methods. Within-laboratory precision is detailed at the end of each of the methods sections in the methods manuals [17, 18, 191 and is summarized in Appendix A (Tables A-l -1 to A-l -18 for the marine/estuarine methods and Tables A-l -19 to A-l -31 for the freshwater methods) and summarized in Tables l-6 and l-7. lntralaboratory precision data also are presented for acute toxicity tests and are summarized in Table l-8. Each laboratory should be establishing a reference toxicant “record,” including a control chart. EPA’s reference toxicant numbers are only meant to show precision of the methods within EPA laboratories and to serve as guidance for other laboratories. Each laboratory’s reference toxicant data will reflect conditions unique to that facility, including dilution water, culturing, etc. However, each laboratory’s reference toxicant CVs should reflect good repeatability.
The CVs may be calculated for acute LCso and chronic ECSu, ICzS, and lG,o data. A mean and range is given for the chronic no observed effect concentration (NOEC) precision data because an NOEC is not a point estimate and is dependent on the tightness of the concentration interval employed in the reference toxicant tests (i.e., the closer the NOEC concentration range the more precise the test isfor the reference toxicant). The closer the CV is to zero, the better. However, CVs should only be compared with the same test protocol/species tested against the same reference toxicant. Estimates of variability (CVs) should only be applied for specific protocols against a specific chemical using the same concentration intervals.
Reference toxicant data should be required for each of the methods stipulated by the permit authority as part of routine quality assurance/quality control (QA/QC) for checking the reliability of the tests conducted by the permittees. In addition, Criteria of Acceptability for each of the 10 chronic methods are listed in the methods manuals, and should be used as a check for whether the compliance data submitted is minimally acceptable [18, 191. (See Table 1 of each of the 4 freshwater methods and Table 2 of each of the 10 marine/estuarine methods entitled, “Summary of Recommended Effluent Toxicity Test Conditions.“)
To date, interlaboratory precision (round robin) tests have been completed for the 7-day Fathead Minnow Larval Survival and Growth Test, the Cladoceran. Ceriodophniu Survival and Reproduction Test, and the Sheepshead Minnow larval Survival and Growth Test. The results of these round robin studies show good reproducibility for these three methods. Results of the round robin testing will show greater variability (i.e., larger CVs) due to a larger number of variables introduced by many round robin laboratories participating. Researchers

12

### Box 1-5. Toxicity Test Method Precision (continued)

have found that a two- to threefold increase in CV values is acceptable with biological testing [46, 50, 511. Interlaboratory data also are presented from several acute toxicity tests [46]. The data from these round robin tests can be found in Appendix A (Tables A-l -5, A-l -23, A-l -24, A-l -27, A-l -28, and A-l -30) and are summarized in Table l-8.
Researchers agree that the precision of these tests is acceptable. Rue, Fava, and Crothe concluded that whole effluent toxicity test methods “are comparable to accepted analytical methodologies” [SO]. Another study by Crothe, Kimerle, and Malloch also concluded that when comparing ”

...CVs for select effluent toxicity test methods and commonly accepted analytical methods...the precision of both techniques issimilar” [51]. This has led the Agency to conclude ”

...that toxicity test methods, where properly followed, exhibit an acceptable range of variability” (see the discussion of toxicity testing requirements for POTWs, 55 FR30082 at 30112, July 24, 1990) 

## [=I.

ride the need to characterize adequately a given effluent and the factors unique to the discharge situation.


### Flow-throu#hvems Static andRenewalToxicity Testing Several factors should be considered in making the choice of toxicity test system. These include the type of toxicity being measured (i.e., is the effluent highly variable or not; is the discharge continuous or intermittent?); the amount of data needed (variable effluents may require more data); and, as between different systems that will provide adequate data, expense.
Two basic types of testing systems are available to measure effluent toxicity: flow-through systems and static systems. A flowthrough toxicity test is conducted using a diluter system and a continuous feed of effluent and dilution water. A static toxicity test is conducted in test chambers (without a serial diluter delivery system) into which effluent and diluent are added manually. Usually, only one effluent sample is collected and used at the beginning of a static test. A variation of the static procedure isthe renewal toxicity test. This test uses the same delivery system as that of a static test but the test solutions are changed, or renewed, on a predetermined schedule (i.e., every 24 hours). Fresh effluent samples generally are collected to renew the test solutions.
Online continuous flow-through testing can sample and measure “peaks” of toxicity should they occur during the testing period. In variable effluents, however, the test organisms would only be exposed to peak toxicity for periods proportional to the flowthrough rate, the duration of the peak in toxicity and length of the test. Static and static renewal tests also can measure peaks in effluent toxicity depending on the type of sampling used, and if the sampling occurs at the time of the toxicity peak.
If the effluent is highly variable and continuously discharged, either a flow-through or renewal test would be appropriate. If the effluent is highly variable with an intermittent discharge, a flowthrough or a renewal test also would be appropriate. However, the effluent sample collected for the renewal test should be a composite collected over the period of the discharge. If the effluent is not considered variable, such as a discharge from a 30- day retention basin, then a static or renewal test using a grab or 24-hour composite sample would be an appropriate test system. For a chronic toxicity test, a 24-hour composite effluent sample is most appropriate. For an acute test, four grab samples taken 6 hours apart or four 6-hour composite samples are most appropriate to measure the peaks of toxicity in an effluent.
Cost also is a factor. Flow-through tests are more resource intensive and require complex delivery systems. Consequently, less data can be generated per unit cost than with static or renewal testing. Where more data at less cost are desirable, static or renewal testing probably is more appropriate. Typically, more samples using renewal is preferable to fewer samples using flowthrough for the same total cost since this would allow better characterization of effluent variability.


### Grab Samplinflvems Composite Samplinfl The use of a grab sample or a composite sample is based upon the objectives of the test and an understanding of the long-term operations and schedules of the discharger. If the toxicity of the effluent is variable, grab samples collected during the peaks of effluent toxicity provide a measure of maximum toxic effect. Collection of grab samples may be necessary if there is little dispersion or mixing of the effluent in the receiving water. In these instances the peaks could persist in the receiving water. Although a grab sample has the potential of revealing the toxicity peak in an effluent, the sample has to be collected at the time of the toxicity spike. Therefore, in a variable effluent, the grab sample has a high probability of missing the toxicity peak. On the other hand, a 24-hour composite sample may more readily catch the toxicity peak(s), but the cornpositing process may tend to dilute the toxicity resulting in a misleading measure of the maximum toxicity of the effluent. Composited samples are, therefore, more appropriate for chronic tests where peak toxicity of short duration is of lesser concern. More detailed discussions of the type of toxicity tests and the best sampling methods are provided in the manuals for the acute and chronic, freshwater and marine toxicity testing procedures [17, 18, 191 and in Chapter 3. 

### Var/abMty There are three important sources of differences in a water quality impact analysis:

13Table 1-6. lntralaboratoty Precision of Chronic Whole Effluent Toxicity Test Methods
Test NOEC Mean Mean Water Method Range IC25 ww ICSO W%> Compound Used Cyprinodon voriegatus-Survival and Growth
>0.05 - 0.05 mg/l 0.07 41.8 0.13 40.8 Cop r AS 0.5 - 1.O mg/l’ 1.5 31.4 1.9 31.8 SDSr AS 31 - 125 ug/12 300.4 33.0 396.9 19.2 Copper NS 1.3 - 2.5 mg/l’ 2.2 27.6 2.6 35.3 SDS NS Embryo larval survival and teratogenicity

200 - 240 ug/12 ECIO EGO 202 2.0 - 4.0 mg/l’ 1.9
Menidia beryllina - Survival and Growth 31 - 125 ug/12 209.9 1.3 + 0 mg/l 1.3 Mysidopsis bohiu -Survival, Growth, and Fecundity ~0.3 - 5.0 mg/14 5.7 63 - 125 ug/ll 138.3
Arbacio punctulato -Fertilization 5.0 - 12.5 ug/l’ 23.5 1.2 - 3.3 mg/l’ 1.7 <6.1 - 24.4 ug/12 22.9 0.9 - 1.8 mg/l’ 2.58 Champia parvulo- Reproduction 0.5 - 1 .o ug/l’ 1.79 0.5 - 1 .o ug/ll 0.93 0.09 - 0.48 mg/12 0.31 0.15 - 0.60 mg/12 0.46 Pimephales promekx- Survival & Growth 128 - 256 ug/l’ -5 0.011 - 0.013 mg/ll -5 2.8 233.5 2.5 AS Copper 35 11.7 2.9 SDS AS

43.7 340.8 50.7 Copper NS 43.2 1.9 9.4 SDS NS

35.0 6.9 47.8 SDS NS 18.0 185.8 5.8 Copper NS

54.6 45.7 47.9 AS Copper 29.7 2.4 23.3 SDS AS 41.9 29.9 48.2 Copper NS 28.7 3.2 33.3 SDS NS

61.09 3.35 34.5 Copper NS 63 1.4 38.6 Copper AS/NS 69.0 0.36 37.0 SDS ASINS 62.3 0.75 22.92 SDS NS

- -5 - NAPCP6 Fw

- -5 - Cadmium Fw
Embryo larval survival and teratogenicity LCl 0.011 - 0.013 mg/l - - 0.0068 62 Cadmium Fw 0.011 - 0.013 mg/l - - 1.51 41.3 Diquat Fw Ceriodaphnia dubio - Reproduction 0.10 - 0.30 mg/l’ 0.22 41 .13 0.3 27.9 NAPCP Fw 0.25 - 1.OOmg/l 0.91 20.5 1.24 15.2 Sodium Chloride

Selenastrum Capricorn&urn - 96-hour Survival Go 2.1 - 2.8 g/l4 - - 2.4 10.2 Sodium Fw chloride

‘Difference of one test concentration. 2Difference of two test concentrations. 3Sodium dodecyl sulfate. 4Difference of four test concentrations. 5Raw data were unavailable, so IC25 and ICso could not be calculated. %odium pentachlorophenol. AS-artificial seawater. N&-natural seawater. FW-freshwater.
-: Data not available. Note: Data used in this table are found in Appendix A-l

14Table l-7. lntralaboratory Precision of Acute Whole Effluent Toxicity Test Methods

N (number of tests) CV(%) Compound Pimephales pfomelas* 12 40 NAPCP

(96-hour) 9 22 SDS 9 86 Cadmium

Daphnia pulex* 14 36 NAPCP

(48-hour) 10 43 SDS 9 21 Cadmium

Daphnia magna* 13 10 NAPCP

(48-hour) 8 29 SDS 8 72 Cadmium

‘Data taken from Draft 1990 Acute Manual.

Table l-8. Summary of Interlaboratory Variability Data for Whole Effluent Toxicity Test Methods [17, 18, 19,461

lC= CV(%)’

44.2

31 .o

41.1

- 

### -

29.0

20.5 Test Method Chronic 1. Cyprinodon variegatus 7-day growth and survival 2. Pirnephales promelas 7-day growth and survival

3. Ceriodaphnia dubia 7-day reproduction

4. Ceriodaphnia dubia 7-day reproduction 5. Ceriodaphnia dubia 7-day reproduction

6. Ceriodaphnia dubia 7-day reproduction NOEC Range

1 - 3.2% effluent2

~3.0 - 6.0 mg/12 potassium chromate

0.25 - 0.30 mg/l NAPCP3

6 - 12% effluent2

co.25 - 1.O mg/l sodium chloride 0.25-l .O mg/l sodium chloride

Acute
7. Cyprinodon variegatus 96-hour static 96-hour flow-through 96-hour static 96-hour flow-through 8. Mysidopsis bahia 96-hour static 96-hour flow-through 96-hour static 96-hour flow-through Toxicant LCCJJ CV(%)

endosulfan 37.7 endosulfan 46.2 silver nitrate 34.6 silver nitrate 50.1

endosulfan 59.5 endosulfan 51.9 silver nitrate 26.6 silver nitrate 22.3

’ CV-coefficient of variation. lThis represents a difference of one exposure concentration. 3NAPCP-Sodium pentachlorophenol.
-: Data unavailable. Note: Data summarized in this table were taken from Appendix A-l
15

### .

.

.

Each Effluent variability is caused by changes in the composition of the effluent. Virtually afl effluents vary in composition over time. Exposure variability is caused by changes in flow rates of both effluent and receiving water. There also are variable receiving water parameters that may be independent of flow, such as background toxicant levels, pH, salinity, tides, suspended solids, hardness, dissolved oxygen, and temperature, 

### that can be important in assessing impact. Species sensitivity differences are caused by the differences in response to toxicants between species. type of variability is discussed below.

Effluent variability is an important component in overall variability of water quality impact analyses and should be addressed adequately in permitting (see Chapter 5, Permit Requirements). Effluent variability can be addressed by designing proper sampling and testing procedures. Sampling measurements should be tailored to the toxic effect of concern (i.e., acute or chronic) and the need to design testing that accounts for effluent variability. Chapter 3, Effluent Characterization, describes recommendations for a testing frequency designed to assessvariable effluents. Appendix Fdetails suggested sampling procedures. Appendix A-2 demonstrates the types of effluent variability that may be seen in publicly owned treatment works (POTW) effluents as measured through toxicity testing of the effluents (see Appendix A-2, Tables A-2-l to A-2-9). The CVs (effluent variability) for POTW effluents are based on acute LCsn data that range from 19.6 percent to 42 percent effluent, and for IC25 chronic data that range from 52.8 percent to 101.3 percent, Also in Appendix A-2, Tables A-2-10 to A-2-12 show acute and short-term chronic effluent variability data from oil refineries on three species, fathead minnows, Ceriodaphnia, and mysids. The CVs associated with this effluent variability data range from 18.7 percent to 54 percent for the acute LCso data, and from 29.8 percent to 59.6 percent for the chronic NOEC data. Data on effluent variability in various types of manufacturing facilities are in Appendix A-2, Tables A-2- 13 to A-2-l 8. Acute toxicity test results show CVs for effluent variability ranging from 20.3 percent to >53.9 percent.
Tables A-2-6 to A-2-9 in Appendix A-2 illustrate the effluent variability of a POlW effluent over the course of a year in which gradual upgrading to full secondary treatment was occurring. Four saltwater short-term chronic toxicity tests were conducted on the POTW’s effluent using the sea urchin fertilization test (Arbacia punctulato), the red macroalga fertilization test (Charnpia parvu/a), the mysid 7-day growth, fecundity and survival test (Mysidopsis bahia), and the inland silverside 7-day larval growth and survival test (Menidia beryllina). The sea urchin and red macroalga tests were conducted daily during each of the four 7- day studies, and provide good examples of the daily variability of the effluent.
These results show that the effluents vary in toxicity and that any one effluent can exhibit significantly varying toxicity to different test species over time. The data also indicate that the effluents were rarely toxic below 10 percent effect concentration and were not toxic below 0.1 percent effect concentration. This information is discussed in Chapter 3, Recommendations for Testing the Toxicity of Effluents section.

Exposure variability is a complex factor that can be addressed in two ways. First, the simplest, easiest applied approach is to assume a steady state exposure condition (usually an estimate of presumed “worst case” exposure) using a critical receiving water flow or condition and a typical effluent flow.
A second method is to attempt to estimate or actually measure the variable exposure situation at the discharge site. This requires statistical analysis and some form of dynamic modeling. Chapter 4, Exposure and Wasteload Allocation, describes appropriate exposure assessment procedures for freshwater and saltwater systems.

One of the primary considerations in establishing a toxicity testing requirement for a discharger is requiring a suitable test species. Different species exhibit different sensitivities to toxicants. Often, differences of several orders of magnitude exist for a given individual toxicant between the least sensitive and the most sensitive species. This range varies greatly and can be narrow or wide depending on the individual toxicant involved.
Since the measured toxicity of an effluent will be caused by unknown toxic constituents, the relative sensitivities of various test species also will be unknown. Therefore, proper effluent toxicity analysis requires an assessment of a range of sensitivities of different test species to that effluent. A knowledge of the range is necessary so that the regulatory authority can protect aquatic organisms. The only way to assessthe range of sensitivities is to test a number of different species from different taxonomic groups, as in the development of the national ambient water quality criteria.
To provide sufficient information for making permitting decisions, EPA recommends a minimum number of three species, representing three different phyla (e.g., a fish, an invertebrate, and a plant) be used to test an effluent for toxicity. However, in some cases, the optimum number of species may be fewer or more depending upon such factors as how thoroughly the effluent has been characterized, the available receiving water dilution, the use classification and existing uses of the receiving water, as well as other special considerations. For example, if an effluent has been characterized as highly consistent, with little chance of variation due to batch processes, changes in raw materials or changes in treatment efficiency, then the use of the two most sensitive species, or even the one most sensitive species, may be appropriate as determined on a case-by-case basis.
Since whole effluents are complex mixtures of toxicants, generalizations about sensitive and nonsensitive species are difficult to make. For example, one generalization is that trout are considered sensitive organisms requiring high-quality water. However, this generalization may not apply in all cases; trout are very sensitive to oxygen depletion but may be relatively insensitive to

16certain toxicants. Another species, Daphnio magna, isvery sensitive when exposed to many toxicants, but relatively insensitive when exposured to the pesticide endrin. Bluegills are very resistant to metals, particularly copper. Conversely, bluegills are a sensitive test species for organophosphate pesticides. Figures l-7 to l-9 show the differences in species sensitivities to hexavalent chromium, dieldrin, and an effluent from a POlW, respectively [53]. The wide range between sensitivities for the different test species is shown. Comparing the figures shows that the fish, invertebrates, and algae shift relative sensitivities to the effluents/toxicants. The fish are less sensitive to chromium but more sensitive to dieldrin. For the cladocerans, the reverse istrue. The results of whole effluent tests using five marine/estuarine short-term chronic test methods also indicate that no species or test method isalways the most sensitive. In a total of 13 effluents tested onsite, Champia porvula was the most sensitive in 15 percent, Arbacia punctulata in 54 percent, mysids in 31 percent and fish in 15 percent of the cases 1241. Analysis of species sensitivity ranges found in the national ambient water quality criteria [1,2] indicates that if tests are conducted on three particular species (Daphnia magno, Pimepholes promelus. and Lepomis macrochirus), the most sensitive of the three will have an LCso within one order of magnitude of the most sensitive of all species tested [54]. This was found to be true for 71 of the 73 priority pollutants tested with four or more species. Sometimes, regulatory agencies require testing on representative resident species under the assumption that such tests are needed to assess impact to local biota. EPA considers it unnecessary to test resident species since standard test species have been shown to represent the sensitive range of all ecosystems analyzed [54]. Resident species toxicity testing is strongly discouraged unless it is required by State statute or some other legally binding factor, or it has been determined that a unique resident species would be far more protective of the receiving water than the EPA surrogate species. The use of other representative species should be subjected to strict quality assurance and quality control procedures and should follow rigorous test methodologies that are at least equivalent to EPA methods. Quality assurance procedures should account for the use of the same species, the same life stage and age of individuals, acclimation periods to avoid mortality due to collection, seasonal variations in populations, habitat requirements, health of the species cultured, as well as the use of reference toxicant tests and other standard procedures. To use a resident organism, a facility would have to develop a protocol to culture the organism and to assess intra- and interlaboratory variability. Such testing is more costly, more difficult, and potentially subject to more variability (disease, age, etc.) than standardized testing. In any case, organisms collected directly from the receiving water itself should never be used because existing impairment may mask any toxicity.


### Acoto-to-ChronicRatio The acute-to-chronic ratio (ACR) expresses the relationship between the concentration of whole effluent toxicity or a toxicant causing acute toxicity to a species (expressed as 

### anacutetoxicity endpoint such as an LC50) and the concentration of whole effluent toxicity or a toxicant causing chronic toxicity to the same species (expressed as a chronic toxicity endpoint such as an Plmephales promelas

_ . 7

,(/ Cenodaphora rehculata /

. . . . l ;/ ’

- . . . . . .

. . .

1 I -2.0 -1 0 0.0 1.0 2.0 Log LC5o in mg/l
Figure l-7. Log of LCsos of Freshwater Species Exposed to Hexavalent Chromium

Plmephafes promelas

Daphnrapolex

-3 0 -2.0 -1.0 0.0 Log LCw in mgF

Figure l-8. Log of LCSos of Freshwater Species Exposed to Dieldrin

8 5-

.- $ - r% Arbaoa punctu:ata
z 3- ii ,- Cenodaphnia rebculata f Pimephales promelas

z' 1- I 20 40 60 60 Log LC50 in mg/l as Percent Efiluent

Figure l-9. Log of LCsos of Freshwater and Saltwater Species Exposed to a POlW Effluent

NOEC or its equivalent, i.e., ACR=ATE/CTE or LCso/NOEC). An ACR is commonly used to extrapolate to a “chronic toxicity” concentration using exposure considerations and available acute toxicity data when chronic toxicity data for the species, chemical, or effluent of concern are unavailable. The ACR should be greater than one, since the ratio compares an acute effect concentration with a chronic effect concentration.
This parameter can be a source of uncertainty in predicting water quality impact because the ACR varies between species for a given chemical and, for any one species, between different toxicants. The latter is a reason why the ACR for a complex effluent may not be a constant. Regardless of this variability, when faced with a limited amount of chronic toxicity data, the regulatory authority

17must apply some ACR to an effluent or cb?mical (or decide to collect more data) when converting wasteload allocations to common terms in the permit limit derivation process described in Chapter 5.
The ACR also may be used in developing chronic toxicity limits where chronic toxicity is not measured directly, in order to minimize testing costs. Likewise, if the toxicity is for the most part manifested in reproduction, growth, etc. (i.e., nonlethal) endpoints, an acute test may not be appropriate for compliance monitoring. Where acute and chronic toxicity data are available, the ACR should be calculated directly for that specific effluent.
Data on 

### acute and chronic toxicity for complex effluents from different categories of dischargers (i.e., POTWs, oil refineries, and chemical manufacturers) show that ACRsfor whole effluents range from cf.0 to ~-50.0, with the majority of ACRs falling below 20 (see Appendix A-3). Acute to chronic ratios for oil refinery data from one plant, based on three species ranged from 1.49 to ~10.0. Acute to chronic ratios for a variety of chemical manufacturers, based on data from two species ranged from ~1 .O to ~50.0. Acute to chronic ratios for POTWs based on two species ranged from 1.4 to 16.1 (these data can be found in Appendix A3). Interestingly, this range of ACRs virtually is identical to ACRs generated on a number of wastewater dischargers in the State of Sao Paula, Brazil (Appendix A-3, Tables A-3-l and A-3-2). Although the 

### acuteand chronic toxicities measured in Brazil were proportionally higher (more toxic) than those measured in the United States, the ACRs were quite similar (Appendix A-3, Tables A-3-l to A-3-3).
EPA recommends that regulatory authorities use a measured ACR. In the absence of data to develop 

### an ACR, EPA’s data suggests that an ACR of 10 could be used (see Appendix A-3). This represents the upper 90th percentile of all the ACR data in Appendix A-3. Given the protective margin of safety inherent with the use of a critical flow for the calculation of a chronic receiving water waste concentration, an ACR of 10 should provide ample protection against chronic instream impacts.



## 1.4 BIGLOGICAL BIOSURVEY PRoTEcTlow CRITERIA/BIOASSESS#ENT APPROACH FORAQUATICAND LlFE

As illustrated in Figure l-10, ecological integrity is attainable when chemical, physical, and biological integrity occur simultaneously [SS]. Biological integrity is a good indicator of overall ecological integrity of aquatic environments because it can provide both a meaningful goal and a useful measure of environmental status that relates directly to the overall integrity of the Nation’s waters. To better protect the biological Integrity of aquatic communities, EPA recommends that States begin to develop and implement biological criteria in their water quallty standards. Biological criteria, or “biocriteria,” are numerical values or narrative statements that describe the reference biological integrity of aquatic communities inhabiting waters of a given designated aquatic life use. When formally adopted into State standards, biological criteria and aquatic life use designations serve as Figure l-10. The Elements of Ecological Integrity

direct, legal endpoints for determining aquatic life use nonattainment. Per Section 131 .l 1(b)(2) of the Water Quality Standards Regulation (40 CM Part 131), biological criteria can supplement existing chemical-specific criteria and provide an alternative to chemical-specific criteria where such criteria cannot be established. Biological criteria quantitatively are developed by identifying unimpaired or least-impacted reference waters that operationally represent best attainable conditions. Once candidate references are identified, integrated biological surveys (biosurveys) are used to characterize the resident community. 

### Because of the complexity of fully characterizing the biological integrity of an entire aquatic community, State standards should contain biological criteria that consider various components (measures of structure and/or function) of the larger aquatic community.
When biological criteria are incorporated into water quality programs, the biological integrity of surface waters may be directly evaluated and protected. Biological criteria also provide additional benefits by requiring an evaluation of physical integrity and providing a monitoring tool to assessthe effectiveness of current chemically based criteria. Table l-9 summarizes how biological criteria directly and indirectly protect the elements of ecological integrity [55].

A biological assessment, or “bioassessment,” is an evaluation of the biological condition of a waterbody using biological surveys and other direct measurements of resident biota in surface waters. A biological 

### survey, or “biosurvey,” consists of collecting, processing, and analyzing representative portions of a resident aquatic community to determine the community structure and function. Biosurveys and bioassessments can be used directly to evaluate the overall biological integrity (structure and/or functional characteristics) of an aquatic community. Deviations from the biological integrity of an aquatic community can be measured directly using bioassessments and biosurveys only when the impacted commu18Tabk l-9. Water Quality Programs That hcorporate Biological Criteria to Protect Elements of Ecological Integrity
Elements of Ecological Integrity

Chemical Integrity Directly Protects
Chemical-specific criteria (toxics) Whole effluent toxicity (toxics) Irtdirectty Protects

Biocriteria (identification of impairment) Physical Integrity

Biological Integrity Criteria for conventionals (pH, tempature, dissolved oxygen)
Biocriteria (biota response in surface water) BiocriCeria (habitat evaluation)
Chemical/whole effluent testing (biota response in laboratory)

nity is compared against a predetermined reference condition. Without proper quality controls (i.e., reference conditions), biosurveys tend to underestimate impairment. Biosurveys assess or detect the aggregate effect of impacts upon an aquatic communi?, where discharges are multiple, complex, and variable and where point, nonpoint, and stormwater discharges are all affecting the biological condition of the receiving water. The resident community integrates the effects of multiple stresses and sources on numerous interactive biological components over time. Because of this, biosurveys necessarily cannot measure the impacts of one particular effluent that is being discharged to Ihe receiving water. Chemical-specific analyses of pollutants known to impact aquatic life and whole effluent toxicity tests are predictive water quality assessment tools used to evaluate biological integrity. At the present time, biological surveys and biological assessments cannot be used as predictive water quality assessment tools. Biosurveys provide a useful monitor of both aggregate ecological impact and historical trends in the condition of an aquatic ecosystem. Biosurveys can detect aquatic life impacts that other available assessment methods may miss, such as impacts caused by pollutants that are difficult to identify chemically or characterize toxicologically, and impacts from complex or unanticipated exposures. Perhaps most importantly, biosurveys can detect impacts caused by habitat degradation such as channelization, sedimentation, and historical contamination that disrupt the interactive balance among community components. Biosurvey data should be applied towards:

l Refining use classifications among different types of aquatic systems and within a given type of use category.

l Defining and protecting existing aquatic life uses under State antidegradation policies as required by the water quality standards regulation.

l Classifying outstanding national resource waters. l Identifying where site-specific criteria modifications may be needed effectively to protect a waterbody.

l Improving use-attainability studies.

l Assessing impacts of certain nonpoint sources and, together with the chemical-specific and whole effluent toxicity approaches, assist in controlling them.

l Monitoring the ecological effects of regulatory action taken under CWA Sections 401, 402, and 301 (h).

l Evaluating the effectiveness and documenting the receiving water biological benefits of pollution controls.



### 1.42 4zkxwc6naBAs isthe case with all types of water quality monitoring programs, biosurveys should have clear data quality objectives, utilize consistent laboratory and field methods, and include quality assurance and quality control. Biosurveys should be tailored to the particular type of waterbody being assessed (e.g., wetland, lake, stream, river, or estuary) and should focus on aquatic community components that are representative of the larger ecosystem and that are practical to measure. Biosurveys should be coupled routinely with basic chemical and physical measurements and an objective evaluation of habitat quality.
EPA’s Office of Water and several State water quality programs have developed techniques as guidance to support biosurveys and bioassessments [56-621. The techniques are an excellent supplementary tool to whole effluent toxicity testing and chemical-specific techniques. However, it is important 

### that biosurveys include sampling of as many species at different trophic levels as possible to reveal accurately receiving water community impacts.
Excellent examples of biosurvey/bioassessment data collected and used in concert with ambient or effluent toxicity test data are the site studies described in Boxes l-3 and 1-4. The toxicity test results and the ambient biosurvey data were based on the recommended minimum of three trophic levels (a fish, invertebrate, and a plant) to give a good overall picture of 

### what was happening in

19the receiving water. Recommended methodologies for conducting biosurveys are included in References 56 through 62.



## 1.5 INTEGRATIONOF THE WHOLEEFFLUENT, CHEMICAL-SPECIFIC, AND BIOASSESSMENT APPROACHES

Section 101 (a) of the CWA states: “The objective of this Act is to restore and maintain the chemical, physical and biological integrity of the Nation’s waters.” Taken together, chemical, physical, and biological integrity define the overall ecological integrity of an aquatic ecosystem. Regulatory agencies should strive to fully integrate all three approaches since each has its respective capabilities and limitations. Table l-10 shows EPA guidance, State Implementation, and State application of each approach 1551. The information summarized in Box l-6, and discussed in detail below, explains how each approach complements the other and why no one of the approaches should be used alone.
A more detailed discussion of the capabilities and limitations of the three approaches is provided below.



### 1.5.1 l2qmMtie.s and Undtatiionsof the ChmdwkpecMc Awmach The principal capabilities of the chemical-specific approach are:

l At present, protection of human health only can be achieved by control of specific chemicals.

l A more complete understanding is available on the toxicology of specific chemicals. EPAacute ambient water quality criteria are based on protecting up to a minimum of eight different organisms including fish, invertebrates, and plants; a minimum of three organisms are used to develop chronic criteria. Considerable information is available in the scientific literature on toxicity caused by specific chemicals. Treatment systems are more easily designed to meet chemical requirements because more treatability data are available.
More information is available on the fate of a pollutant in receiving waters so that the pollutant fate can be conveniently predicted through modeling. Persistence and degradation can be factored into the evaluation.
Chemical analyses are sometimes lessexpensive than toxicity testing and biological surveys, if there are only a few toxicants present. This is more pertinent if only chlorine and ammonia are present in an effluent or ambient water.
This approach allows prediction of ecological impacts before they occur. NPDES permit limits can therefore be developed before an actual ecological impact occurs.
The principal limitations of the chemical-specific approach are:
All toxicants in complex was&waters are not known and, therefore, control requirements for all toxicants cannot be set. Toxicological information on these unknown pollutants is often unavailable.
The bioavailability of the toxicants at the discharge site are typically not assessed, and the interactions between toxicants (e.g., additivity, antagonism) are not measured or accounted for. Asa result, the controls may be either under protective or overly protective.
Direct biological receiving water impact and impairment is not typically measured. There is no way to ascertain directly if the chemical controls adequately are protecting aquatic life. Complete measurement of all individual toxicants, particularly where many are present in the mixture, can be expensive. Organic chemicals, in particular, can be costly to measure.

Table l-10. Process for Implementation of Water Quality Standards
Criteria EPA Guidance State Implementation State Application Chemical-Specific Pollutant-specific numeric criteria State Standards -use designation -numeric criteria -antidegradation Permit limits monitoring Best management practices Wasteload allocations

Narrative “Free Froms” Whole effluent toxicity guidance Water Quality Narrative Permit limits monitoring -no toxic amounts translator Wasteload allocation Best management practices Biological Biosurvey minimum requirement guidance State Standards -refined use -narrative/numeric criteria -antidegradation Permit conditions monitoring Best management practices Wasteload allocation

20

### . .-.-



### Box 1-6. Components of an Integrated Approach to Water Quality-based Toxics Control

Control ADDroach CaDabilities limitations Chemical-Specific -Human health protection -Complete toxicology -Straightforward treatability -Fate understood -Less expensive testing if only

a few toxicants are present -Prevents impacts
Whole effluent toxicity -Aggregate toxicity -Unknown toxicants addressed -6ioavailability measured -Accurate toxicology -Prevents impacts

Bioassessments -Measures actual receiving water effects -Historical trend analysis -Assessesquality above standards -Total effect of all sources, including unknown sources -Does not consider all toxics present -6ioavailability not measured -Interactions of mixtures (e.g., additivity) unaccounted for -Complete testing can be expensive -Direct biological impairment not measured
-No direct human health protection -Incomplete toxicology (few species may be tested) -No direct treatment -No persistency or sediment coverage -Conditions in ambient may be different -Incomplete knowledge of causative toxicant
-Critical flow effects not always assessed -Difficult to interpret impacts -Cause of impact not identified -No differentiation of sources -Impact has already occurred -No direct human health protection



### 1.5.2 Capabilltiss anti Limitations of the Wlrole EMwrf sures through ingestion of fish. This is particularly impor

### Approacn tant for carcinogens. The principal capabilities of whole effluent techniques are: l EPA’s water quality criteria are based on a minimum of eight different species for the acute criteria and three different species for the chronic criteria. Effluent aquatic toxicity commonly is measured with only one, two, or three species. For some toxicants a wider sensitivity range (more species) must be tested; particularly where the mode of toxicity action is specific (such as diazinon or some other pesticides). l The aggregate toxicity of all constituents in a complex effluent is measured, and toxic effect can be limited by limiting one parameter-whole effluent toxicity.

l Toxicity caused by compounds commonly not analyzed for in chemical tests is detected. Control of the toxicant is not dependent upon established toxicological information that may not yet be available for some pollutants.

l The bioavailability of the toxic constituents is assessed, and the effects of interactions of constituents are measured. Additivity, synergism, and antagonism between compounds in an effluent are addressed implicitly by whole effluent toxicity.

l The toxicity of the effluent or ambient water is measured directly for the species tested.
9 This approach allows prediction of ecological impacts before they occur. NPDES permit limits can therefore be developed before an actual ecological impact occurs.
The principal limitations of whole effluent techniques are:

l The approach only measures and controls toxicity to aquatic organisms. It does not protect human health from expol There is less knowledge on designing or manipulating treatment systems to treat the parameter toxicity. Investigate tools for identifying causative toxicants only have been recently developed and may not easily identify all causative toxicants. As a result, identification and proper control may be difficult and expensive.

l The whole effluent toxicity test directly measures only the immediate bioavailability of a toxicant; it cannot measure the persistence “downstream” and long-term cumulative toxicity of a compound. Thus, bioaccumulative chemicals necessarily are not assessed or limited. Toxicants can accumulate in sediment to toxic concentrations over a period of time.

l Where there are chemical/physical conditions present (pH changes, hardness changes, solids changes, salinity changes, photolysis, etc.) that act on toxicants in such a way as to

21“release” toxicity away from the discharge point, such toxicity may not be measured in the effluent. The opposite of this also is possible; toxicity may degrade rapidly so there is no trace of it away from the point of discharge. For example, the actual pH and temperature in an ambient water may be sufficiently low to preclude toxicity from ammonia whereas the higher pH and temperature of the toxicity test may induce toxicity from ammonia.
. It is not always clear which compound or mixture of compounds is causing toxicity in the mixture. The causative toxicant may be difficult to identify for control.



### 1.53 Capabilities and Limitations of the Bkssassm8nt 4m=h The principal capabilities of the bioassessment approach are:

l Biological communities reflect overall ecological integrity. Biosurvey results therefore directly assess the status of a waterbody. The status of a waterbody’s biological health may be of direct interest and more meaningful as a measure of a pollution-free environment.

l Biological communities integrate the effects of different pollutant stressors and thus provide a holistic measure of their aggregate impact. Biological assessments also measure stresses over long time periods and can measure historical trends and fluctuating environmental conditions.

l Biosurveys can identify previously unknown sources of impairment and may identify where site-specific chemical criteria are needed. Bioassessments can be useful in characterizing ecological impacts to a waterbody in multiple discharge situations.

l Bioassessments can characterize the ecological value of ambient waters that are in attainment of the standards. As such, bioassessments provide a means to determine compliance with State antidegradation requirements in standards.
The principal limitations of the bioassessment approach are:

l Bioassessments conducted at critical low flow conditions may be difficult to accomplish.

l Biosurvey data cannot fully characterize impairment until after suitable biocriteria are developed. Biosurvey data may not be sufficient to detect impairments without appropriate reference conditions.

l Bioassessments measure integrated impacts over long periods of time. Multiple factors can contribute to measured impacts. However, bioassessments cannot isolate the causative factor leading to the impairment nor predict future impairment.

l Bioassessments measure impact from any source and as such, the data bracketing a discharge used to assess impacts may be influenced by pollutant sources further upstream. Causes of biological impairment may not be assigned readily to any one discharger.

l Bioassessments identify water quality problems after they have occurred; they currently are not predictive of water quality problems. By design, bioassessments are limited in their ability to identify waters that are not impaired.

l The approach only measures biological impairments to aquatic organisms. It does not protect human health from exposures through ingestion of fish. By using all three approaches, a State will more thoroughly protect aquatic life. The chemical-specific approach provides a high accuracy of analysis of the individual chemical constituents, has been used by regulatory agencies, and is generally lowest in cost because of market availability. However, the level of protection of the chemical-specific approach can be low if toxicants are present in an effluent for which no chemical-specific criteria exists. In addition, some States have adopted very few criteria as a part of their water quality standards. On the other hand, whole effluent toxicity provides a high level of protection by measuring the aggregate effect of all toxicants. It provides accurate toxicology, but it can be higher in cost and has been historically less widely used by regulatory authorities. Bioassessments also provide a coverage of many biological impacts and allow for accurate historical trend analyses. However, bioassessments cost more and data interpretation can be difficult. Therefore, the integrated approach to water quality-based toxics control is essential for a strong toxics control program.
To more fully protect aquatic habitats and provide more comprehensive assessments of aquatic life use nonattainment, EPA recommends that States fully integrate chemical-specific, whole effluent, and bioassessment approaches into their water quality-based toxics control programs. It is EPA’s position that the concept of “independent application” be applied to water quality-based situations. Since each method has unique as well as overlapping attributes, sensitivities, and program applications, no single approach for detecting impact should be considered uniformly superior to any other approach. For example, the inability to detect receiving water impacts using a biosurvey alone is insufficient evidence to waive or relax a permit limit established using either of the other methods. The most protective results from each assessment conducted should be used in the effluent characterization process (see Chapter 3). The results of one assessment technique should not be used to contradict or overrule the results of the other(s). (For more information see Reference 55.)
Whenever there are discrepancies between the findings of the approaches, regulatory agencies may need to re-examine the findings to determine if simplifications or assumptions may have caused the difference. The State of Ohio found in 60 percent of the sites where they collected bioassessment data, a biological impact occurred when chemical-specific data predicted no impact. The reverse also can occur-biosurveys may not show any impact in a stream whereas effluent data modeled at low flow project an exceedance of a chemical-specific criterion. In this instance, the regulatory authority may need to consider a more detailed monitoring and modeling of chemical fate and transport

22(which could include probabilistic modeling) to determine if simplifications in dilution calculations projected higher concentrations than would be expected using the detailed model. The authority also would need to examine concurrently the sampling approach and analysis of the biosurvey data to determine if it appropriately characterized the water. Ifthere was still a difference, then the regulatory authority will need to use the more protective approach as the basis to determine necessary regulatory controls.



## 1.6 OTHER FACTORS INFLUENCING WATER~UAUTYBASED TOXICS CONTBOL

An understanding of the fate and behavior of both single toxicants and whole effluent toxicity after discharge can be important in the application of water quality-based toxics controls. Evaluating the combined effects of interacting toxic discharges also may be important in multiple discharge situations. When evaluating the receiving water behavior of toxicants and toxicity, factors such as toxicity degradation or persistence, and toxicant additivity, antagonism, and synergism are important. Ambient toxicity tests can give some indication of the importance of each of these factors:

l Toxicity Persistence-How long and to what extent (in terms of area), does effluent toxicity or the toxicity of a single toxicant persist after discharge? It is not reasonable to assume that in all cases the persistence of both individual toxic chemicals and effluent toxicity is conservative. For two effluents of equal initial toxicity, the aquatic effects of an effluent whose toxicity degrades rapidly will be different from an effluent whose toxicity persists.

l Additivity, Antagonism, and Synergism-When toxicants or effluents with toxic properties mix in the receiving water, what is their combined fate and toxic effects?

l Test Interferences-This includes pH, temperature, salinity, hardness, and metals.
Each of these factors is discussed below.



### 1.6.1 Pmktence As soon as an effluent mixes with receiving water its properties begin to change. The rate of change of toxicity in that effluent is a measure of its toxicity persistence or degradation. After mixing, the level of toxicity in the receiving water may either remain relatively constant (until further diluted), increase in toxicity due to transformation, or degrade due to fate processes (photodecomposition, microbial degradation) or compartmentalization processes (particulate adsorption and sediment deposition, volatilization).
One disadvantage of the chemical-specific approach is that the bioavailability of the toxicant after discharge is not measured. Onsite toxicity testing has indicated that the individual toxicants causing toxicity measured at discharge sites tend relatively to be persistent near the point of discharge [23, 31-381. However, persistence of individual chemicals can be modeled and the persistence of specific toxicants also can be accounted for in making impact predictions and setting controls. A procedure to determine whether or not an effluent’s toxicity is persistent has been developed by EPA [63]. The procedure describes the steps required to conduct a laboratory evaluation of the degradation of toxicity in complex effluents that are released to receiving waters by simplistically simulating a water body and discharge. EPA recommends this procedure be conducted where the interaction of sources of toxicants is critical to establishing controls.
This simple procedure is performed in a refrigerator-sized environmental chamber in the laboratory using commonly available glassware and shipped effluent samptes. Toxicity is measured using conventional 

### acuteor short-term chronic toxicity tests. The results are used to generate a toxicity degradation rate for the effluent under representative environmental conditions. The procedure has several applications, including measuring the decay of effluent toxicity in a stream or lake, and identifying the most important fate processes responsible for toxicity decay (which also may be useful in treatability or toxicity identification studies). Mixing zones designated by State water quality standards, or developed on a case-by-case basis, are typically small enough that toxicity evaluations need only consider near field situations. Continuous discharges continually can introduce toxic pollutants into a receiving water. Although these pollutants can decay over time, this decay will occur downstream or away from the discharge. The receiving water concentrations at the point of discharge continually are being refreshed. In these instances, toxicity can be considered conservative and persistent (nondecaying) in the near field. However, effluent toxicity can exhibit far field decay. Typical patterns of progressively decreasing downstream toxicity (similar to biochemical oxygen demand decay) have been observed in a number of freshwater situations [23, 31-381. This is of concern when evaluating the combined toxicity of sources located far apart. If there is reason to suspect that an effluent’s toxicity is not persistent, several techniques can be employed to measure changes of toxicity after discharge: Testing should be performed during various seasons of the year corresponding to various receiving water flow regimes. The toxicity test itself, when performed with dilution water immediately upstream or from an uncontaminated area nearby, is an analogue of the mixing and fate processes taking place in the receiving water. The types of rapid chemical reactions found in the mixing zone also can be expected to take place to a large extent when effluents and receiving waters are mixed for toxicity tests. The effects on toxicity persistence of varying physical/chemical conditions in the receiving water or in the effluent cannot, however, be accurately predicted from these results.
Ambient toxicity testing, as detailed in Appendix C, measures the ambient interactions of effluent and receiving water and can be used to assesstoxicity persistence. Toxicity persistence may present a more serious problem in estuarine or lake receiving waters where the toxicity is not flushed away rapidly. In one study, 

### on a POlW effluent being discharged into a small cove off of Narragansett Bay, the decay rate of the effluent was temperature-dependent and was reduced markedly during

23the winter. However, persistence of the effluent in the receiving water cove in the winter did present a problem because tidal flushing did not remove the toxicity [39].

For coastal discharges, certain toxic compounds are more often found to cause impacts in marine and estuarine environments [64]. Due to the physical and chemical processes that tend to trap pollutants in estuaries (sedimentation, salinity flux, etc.), the discharge of these compounds, at very low concentrations over a long period of time, may allow them to accumulate to toxic concentrations. For many of these compounds, applicable permit limits may need to be very stringent to avoid chronic toxicity problems due to the persistence of these compounds.


### 1.6.2 AwitMy, Antagonkm,andSpw@m
Where multiple toxic effluents are discharged to a receiving water, the resultant ambient toxicity is of interest. Since each effluent is composed of individual toxic substances, a mixture of the effluents in a receiving water produces a mixture of these individual pollutants (assuming conservative behavior). The overall ambient toxicity could be equal to the sum of each discharge’s toxicity (additivity), less than the sum (antagonism), or greater than the sum (synergism).

Alabaster and Lloyd (651 observed from their data that the combined acutely lethal toxicity to fish and other aquatic organisms is approximately the simple addition of the proportional contribution from each toxicant. The median value of the effect on fish is 0.95 of that predicted; the collective value for sewage effluents, river waters and a few industrial wastes is 0.85. The range for effluents, river wastes, and industrial wastes is 0.4 to 2.8. (Figure l-1 1 illustrates the data summary.)



# I 0 Sewage Effluents A Gas Liquor A Orllltng Flud

,, Regression Lme ot Potnts Sewage and Inaustria/ Wasles
‘. -1__ Efflu nt With Hi h lndustnal and esttctde % C%mponents 9951 ’ ’ ’ ’ ’ 1 . . . I I I I I #I,
04 1 2 4 8
Times as TOXICas Predicted from Summed Toxic Units Figure l-l 1. Data Summary on Additivity [65] In relation to chronic toxicity, for the growth of fish, Alabaster and Lloyd [65] conclude:

...in the few studies on the growth of fish, the joint effect of toxicants has been consistently less than additive which suggests that as concentrations of toxicants are reduced towards the levels of no effect, their potential for addition is also reduced. There appear to be no marked and consistent differences between the response of species to mixtures of toxicants.

Cases in which one effluent or pollutant parameter (such as total suspended solids) ameliorated the toxicity of another effluent pollutant (antagonism) have been observed. Testing procedures can be designed to measure such interactions. A description of such a procedure is found in “Recommended Multiple-Source Toxicity Test Procedures,” Box 3-3, Chapter 3.
Theoretically, under certain conditions, synergism, a greater than additive increase in toxicity upon mixing, can occur. However, field studies of effluent toxicity and laboratory experiments with specific chemicals imply that synergism would be an extremely rare phenomenon. It has not been observed during onsite effluent toxicity studies, and is not considered an important factor in the toxicological assessment of effluents.

In summary, the available information indicates that the combined effects of individual acutely toxic pollutants are from 0.4 to 2.8 times the effects predicted by adding the individual effects. The median combined effect is approximately additive. For this reason, EPA recommends in the absence of site-specific data that regulatory authorities consider combined acute toxicity to be additive. Since the data shows no such additivity for chronic toxicity, EPA recommends that chronic toxicity not be considered as additive.



### 1.6.3 kst intarfemnces Environmental conditions such as pH, temperature, salinity, hardness, and solids concentration can influence the toxicity test. For example, higher ambient solids concentrations provide more surfaces for toxicants to be adsorbed and can tend to reduce toxicity. In addition, toxicity caused by ammonia is controlled by the ambient pH and temperature. As a normal part of the whole effluent toxicity testing procedure, it is very important to replicate closely the “worst case” receiving water conditions in the testing conditions.

There may be a few unusual situations where the pH, temperature, hardness, salinity, and solids requirements of the testing procedures differ greatly from the worst environmental conditions for these parameters. In these situations, the effluent toxicity tests may either over or under predict the toxicity in the ambient receiving water. An example of this is where ammonia is present and the highest expected ambient water temperature is 20°C whereas the chronic toxicity test must be conducted at 25°C. Since a higher temperature causes more ammonia toxicity, the temperature requirements of the test may induce toxicity not found in the ambient water. In such an instance, the regulatory authority must look carefully at the test protocols and all the data collected to determine if the facility is actually contributing to toxicity in the ambient water. A toxicity identification evaluation

24may be necessary to make this determination. If this analysis shows a toxicity test result to be artificial due to environmental parameters, then that test should be overridden by subsequent valid toxicity tests conducted.



## 1.7 HUMAN HEALTH PROTECTION
Impacts on human health due to exposure to waterborne toxicants can occur through three primary exposure 

### routes: contact recreation, drinking water, and the ingestion of contaminated fish and shellfish tissues. Contact recreation may pose potential risks due to dermal absorption and incidental ingestion. Exposure through drinking water is a significant concern but can be mitigated for specific chemicals by applying drinking water criteria. The third exposure route, human consumption of contaminated aquatic life, is of primary concern in this document due to the potentially high concentrations achieved in fish and shellfish tissues from bioconcentration, and because no NPDES permitting controls exist between 

### tissuecontamination and human exposure. For these reasons, this document focuses on prevention of contaminated aquatic life from bioconcentration as the principal way to control human exposure to waterborne toxicants. Currently, the regulation of human health impacts typically are based only upon the control of individual chemicals. EPA human health water quality criteria protect against the consumption of contaminated water and aquatic life. There is no mechanism like the aquatic toxicity test to determine the effect of a chemical mixture like an effluent on human health. EPA is developing, however, a preliminary approach to analyzing effluents for bioaccumulation potential through the use of a whole effluent bioconcentration analysis followed by identification of individual bioconcentratable pollutants [66]. This procedure is described in Chapter 3. Once this method is reviewed (both internally and externally) and finalized, it will provide another way for regulatory authorities to assess bioconcentratable pollutants.



### 1.7.1 Tvpes oftfwlth Effixts Health effects from toxics are divided into two categories: nonthreshold effects, such as carcinogenicity, and threshold effects, such as acute, subacute, or chronic toxicity. Both terms are defined below.
EPA’sapproach to assessing the risks associated with nonthreshold human carcinogens is different from the approach for threshold toxicants due to the different mechanisms of action thought to be involved. In the case of carcinogens, the Agency assumes that a small number of molecular events can evoke changes in a single cell that can lead to uncontrolled cellular proliferation. This mechanism for carcinogenesis is referred to as “nonthreshold,” since there is essentially no level of exposure for such a chemical that does not pose a small, but finite, probability of generating a carcinogenic response. Cenotoxic pollutants are presumed to have no threshold level, but incremental risk levels can be determined based on the carcinogenic potency of the chemicals.
Threshold toxicants, on the other hand, are generally treated as if there is an identifiable exposure threshold (both for individuals and populations) below which effects are not observable. Threshold toxicants are chemicals that give rise to toxic endpoints other than cancer because of their effects on the function of various organ systems, Such chemicals are presumed to have safe exposure levels. This characteristic distinguishes threshold endpoints from nonthreshold endpoints. However, it should be noted that chemicals that cause cancer and mutations also commonly evoke other toxic effects (systemic toxicity). In the case of systemic toxicity, compensating and adaptive “defense” mechanisms exist that must be overcome before the toxic endpoint is manifested. For example, there could be a large number of cells performing the same or similar function whose population must be significantly altered before the effect is seen. The individual threshold hypothesis holds that a range of exposures from zero to some finite value can be tolerated by the organisms with essentially no chance of expression of the toxic effect. Currently, the control of toxicants that bioconcentrate in edible tissues is achieved in the NPDES program by limiting such poltutants individually. There are whole effluent tests that can measure a wastewater’s potential to cause carcinogenicity or mutagenicity (e.g., Ames test). However, the application of such data isexperimental because of the difficulty in establishing cause/effect relationships between exposure to wastewaters and human health problems. Therefore, at this time EPA recommends regulatory authorities focus on controls for bioconcentratable toxicants on a chemical-by-chemical control basis.
The remaining information regarding regulation of human health impacts is contained in the following chapters: Chapter 2, Water Quality Standards, discusses the development and updating of human health water quality criteria. Chapter 3, Effluent Characterization, discusses the evaluation of effluents for potential human health impacts. Chapter 4, Exposure and Wasteload Allocation, contains information on design conditions and averaging periods. Finally, Chapter 5, Permit Requirements, discusses the derivation of permit limits protective against human health impacts.

25

## CHAPTER 1 REFERENCES

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5. Hedke, SF., and J.W. Arthur. 1985. Evaluation of a SiteSpecific Water Quality Criterion for Pentachlorophenol Using Outdoor Experimental Streams. Aquatic Toxicology and Hazard Assessment: Seventh Symposium. ASTM STP 854. Ed. R.D. Cardwell, T. Purdy, and R.C. Bahner. American Society for Testing and Materials, Philadelphia, PA.
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10. Zischke, J.W., and I.W. Arthur. 1987. Effects of Elevated Ammonia Levels on Fingernail Clams (Musculium tronsversum) in Outdoor Experimental Streams. Arch. Environ. Cont. Tox. 16:225-31.

11. Herman&, R.O., K.N. Allen, and SF. Hedtke. In press. Toxicity and Fate of Total Residual Chlorine in Outdoor Experimental Streams. Proceedings of Eigth Conference on Water Chlorination.

12. Monticello Ecological Research Station. 1988. The lmpct of Chlorine/Ammonia on Ecosystem Structure and Function in Experimental Streams. Report to the Qffice of Water, U.S. EPA. March 1988. 13. Newman, R.M., ].A. Perry, E. Tam, and R.L. Crawford. 1987. Effects of Chronic Chlorine Exposure on Litter Processing in Outdoor Experimental Streams. freshwater Biology 18:415- 28.

14. Perry, J.A., N.H. Troelstrup, Jr., M. Newson, and B. Shelley. Generality. Wet. Sci. Tech, 19(11):55-71.
15. U.S. EPA. 1979. Methods for Chemical Analysis of Water and Wastes. Environmental Monitoring and Support Laboratory, Cincinnati, OH. EPA 600/4-79-020 (revised March 1983).
16. Wang, W. 1990. Toxicity Assessment of Pretreated Industrial Wastewaters Using Higher Plants. Res.1.Wuter Poll. Control Fed.62:853-60.

17. U.S. EPA. 1991. Methods for Measuring the Acute Toxicity of Effluents to Aquafic Organisms. 4th. ed. Qffice of Research and Development, Cincinnati, OH. EPA-600/4-90-027.
18. U.S. EPA. 1991. Short-Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Freshwater Organisms. 3d. ed. Qffice of Research and Development, Cincinnati, OH. EPA-600/4-91/002.
19. U.S. EPA. 1991. Short-Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Marine and Esiuarine Organisms. 2d. ed. Qffice of Research and Development, Cincinnati, OH. EPA-600/4-91/003.
20. Warren-Hicks, William J. 1990. Biological Test MethodsPerformance Considerations. Kilkelly Environ. Assoc. Report no. 1342-l 44-01 /01/29/90-02F.

21. Norberg-King, T.J. 1991. Calculations of ICp Values of ICI 5, IC20, IC25, IC30, and lC50 for Appendix A of the Revised Technical Support Document. Memorandum to M. Heber, U.S. EPA.

22. Bergman, H., R. Kimerle, and A.W. Maki, eds. 1985. Environmental Hazard Assessmenf of Effluents. Elmsford, NY: Pergamon Press, Inc.
23. U.S. EPA. 1987. Biomonitoring to Achieve Control of Toxic Effluents. EPA625/8-87/013.

24. Schimmel, SC., GE. Morrison, and M.A. Heber. 1989. Marine Complex Effluent Toxicity Program: Test Sensitivity, Repeatability, and Relevance to Receiving Water Impact. Env. Toxicol. and Chem. 5:739-46.

2625. Schimmel, SC., G.B. Thursby, M.A. Heber, and M.J. Chammas. 1989. Case Study of a Marine Discharge: Comparison of Effluent and Receiving Water Toxicity. In Aquatic Toxicology and Environmental t&e: Eleventh Volume. ASTM STP 1007. Ed. G.W. Suter II and M.A. Lewis. American Society for Testing and Materials, Philadelphia, PA.
26. Birge, W.]., J.A. Black, T.M. Short, and A.C. Westerman. 1989. A Comparative Ecological and Toxicological Investigation of a Secondary Wastewater Treatment Plant Effluent and Its Receiving Stream. Envi. Toxicol. and Chem. 8:437-50.

27. Dickson, K.L., W.T. Waller, J.H. Kennedy, W.R. Arnold, W.P. Desmond, SD. Dyer, J.F.Hall, J.T. Knight, Jr., D. Malas, M.L. Martinez, S.L. Matzner. 1989. A Water Quality and Ecological Survey of the Trinity River, volume 7. Report conducted by Institute of Applied Sciences, University of N. Texas and Graduate Program in Environmental Sciences, University of Texas at Dallas.
28. Eagleson, K.W., D.L. Lenat, L. Ausley, and F.Winbome. 1990. Comparison of Measured lnstream Biological Responses with Responses Predicted by Ceriodaphnb Chronic Toxicity Tests. Env. Toxicol. and Chem. 9:1019-28.
29. Dickson, K.L., W.T. Waller, L.P. Ammann, and J.H. Kennedy. 1991. Examining the Relationship Between Ambient Toxicity and lnstreom Impact. Submitted to: Environmentat Toxicology and Chemistry.
30. Parkhurst, B.R., M.D. Marcus, and L.E. Noel. 1990. Review of the Resultsof EPA’sComplex Effluent Toxicity Testing Program. Utility Water Act Group.
31. Mount, O., N. Thomas, M. Barbour, T. Norberg, T. Roush, and R. Brandes. 1984. Effhmt and Ambient Toxicity Testing and lnstream Community Responseon the Ottawa River, Limo, Ohio. Permits Division, Washington, DC, Office of Research and Development, Duluth, MN. EPA-600/3-84-080, August 1984.
32. Mount, D.I., and T.]. Norberg-King, eds. 1985. Va/idity of Effluent ond Ambient Toxicity Tests for Predicting Biological Impact, Scippo Creek, Circleville, Ohio. U.S. EPA. EPA/600/3-85/ 044, June 1985.
33. Mount, D.I., et al., eds. 1985. Validity of Effluent and Ambient Toxicity Testsfor Predicting Biological Impct, Five Mile Creek, Birmingham, Alabama. U.S. EPA. EPA600/8-85/015.

34. Mount, D.I., A.E. Steen, and T. Norberg-King, eds. 1986. Validity of Effluent and Ambient Toxicity Tests for Predicting Siokqicol Impact, Bock River, Baltimore Harbor, Maryland. U.S. EPA. EPA&IO/&86/001, July 1986.
35. Mount, D.I., T. Norberg-King, and A.E. Steen, eds. 1986. Validity of Effluent and Ambient Toxicity Tests for Predicting Biological Impact, Nougatuck River, Waterbury, Connecticut. U.S. EPA. EPA 600/8-86/005, May 1986. 36. No&erg-King, T.J., and D.I. Mount, eds. 1986. Vaiidity of Effluent and Ambient Toxicity Tests for Predicting Biological Impact, Skeleton Creek, Enid, Oklahoma U.S. EPA. EPA 600/8- 86iOO2, March 1986.
37. Mount, D.I., A.E. Steen, and T. No&erg-King, eds. 1986. Validity of Effluent and Ambient Toxicity Tests for Predicting Biological Impact, Ohio River, Wheeling, WestVirginia. U.S. EPA. EPA600/3-85/071, March 1986.
38. Mount, D.I., and T. No&q-King, eds. 1986. Validity of Effluent and Ambient Toxicity Tests for Predicting Siol~ical Impact, Kanuwho River, Charleston, West Virginia. U.S. EPA. EPA 600/3-86/006, July 1986.
39. Dettmen, E.H., J.F. Paul, IS. Rosen, and C.J. Strobel. 1989. Transport, Fate, and Toxic Effectsof a Sewage Treatment Plant Effluent in u Rhode /s/and Estuary. U.S. EPA/ORD ERLNarragansett, Contribution No. 1003.
40. U.S. EPA. 1989. Biomonitoring for the Control of Toxicity in Effluent Discharges to the Marine Environment. EPA 625/8- 89/015, September 1989.
41. Nimmo, D.R., M. Dodson, P.H. Davies, J.C. Greene, and M.A. Kerr. 1990. Three Studies Using Ceriodaphnio to Detect Nonpoint Sources of Metals from Mine Drainage. Res. 1. Water Poll. Control Fed. 62:7-14.

42. Nebeker, A. 1982. Evaluation of a Dophnio magna, Renewal Life-Cycle Test Method with Silver and Endosulfan. Water Research 16:739-44.
43. Crothe, D., and R. Kimerle. 1985. Inter- and Intra-Laboratory Variability in Daphnia magna Effluent Toxicity Test Results. Env. Tox. and Chem. 4(2)189-92.

44. Qureshi, A.D., K.W. Flood, S.R.Thompson, SM. Junhurst, C.S. Inniss, and D.A. Rokosh. 1982. Comparison of a Luminescent Bacterial Test with Other Bioassays for Determining Toxicity of Pure Compounds and Complex Effluents. Ed. ].C. Pearson, et al. In Aquatic Toxicology Hazard Assessment: Fifth Conference. ASTM STP 766. American Society for Testing and Materials, Philadelphia, PA.
45. Strosher, M.T. 1984. A Comparison of Sio/ogical Testing Methods in Association with Chemical Analyses to Evaluate Toxicity of Waste Drilling Fluids in Alberta, volume I. Canadian Petroleum Association, Calgary, Alberta.
46. Schimmel, S.C. 1981. Results: lnterlabomtory Comparison of Acute Toxicity Tests Using Estuarine Animals. EPA-600/4-81- 003.
47. U.S. EPA. 1982. Pesticide Assessment Guidelines. Office of Pesticide Programs, Washington, DC. EPA/9-82-018-028.
48. U.S. EPA. 1982. Toxic Substances TestGuidelines. Office of Toxic Substances, Washington, DC. EPA/l 6-82-001-003.

27

### ---__ -----

49. Morrison, C,, E. Torello, R. Comeleo, R. Walsh, A. Kuhn, R. Burgess, M. Tagliabue, and W. Greene. 1989. Intralaboratory Precision of Saltwater Short-term Chronic Toxicity Tests. Res.1.W.P.C,F. 61 (11/12):1707-l 0.
50. Rue, W.J., !.A. Fava, and D.R. Grothe. 1988. A Review of Inter- and lntralaboratory Effluent Toxicity Test Method Variability. Aquatic Toxicology and rfozard Assessment: I Uth Volume. ASTM STP 971.
51. Crothe, D.R., R.A. Kimerle, and C.D. Malloch. 1990. A Perspective on Biological Assessments. Water Environment and Technology.
52. U.S. EPA. 1990. EPA Administered Permit Programs, the National Pollutant Discharge Elimination System, General Pretreatment Regulations for Existing and New Sources, Regulations to Enhance Control of Toxic Pollutant, and Hazardous Waste Discharges to Publicly Owned Treatment Works. July 24, 1990, 55 FR30082.
53. LeBlanc, C.A. 1984. Interspecies Relationships in Acute Toxicity of Chemicals to Aquatic Organisms. Env. Tax. and Chem. 3(1):47-60.
54. Kimerle, R.A., A.F. Werner, and W.]. Adams. 1984. Aquatic Hazard Evaluation Principles Applied to the Development of Water Quality Criteria. In Aquatic Toxicology and Hazard Assessment, Seventh Symposium. ASTM STP 854. Ed. R.D. Cardwell, R. Purdy, and R.C. Bahner. American Society for Testing and Materials, Philadelphia, PA.

55. U.S. EPA. Biological Criteriu, Notional Program Guidance for Surface Waters. EPA440/5-90-004, April 1990.
56. Plafkin, J.L.,et. al. 1989. Ropid Bioassessment Protocols for Use In Streams and Rivers. Office of Water Regulations and Standards. EPA444!4-89-001. 57. Karr, j.R., et. al. 1986. Assessing Biological integrity in Running Wuters: A Method and Its Rationale. Ill. Nat. t-list. Survey Special Publ. 5.
58. Ohio EPA. 1987. Biological Criteria for the Protection ofAquatic Life: Volumes 1, 2, and 3. Division of Water Quality Monitoring and Assessment. Columbus, OH.
59. Lenat, D.R. 1988. Water Quality Assessments of Streams Using a Qualitative Collection Method for Benthic Macroinvertebrates. 1.N. Am. Benthol. Sot. 7:222.
60. Schackleford, B. 1988. Rapid Bioassessments of Macroinvertebrute Communities: Biocriteria Development. Arkansas Dept. Poll. Contr. and Ecol., Little Rock, AR.
61. Maine DEP. 1987. Methods for Sio/ogica/ Sampling and Anafysis of M&e’s Waters. Maine Bureau of Water Quality Control.
62. Weber, Cl. 1973. Biological Field and Laboratory Methods for Measuring the Quality of Surface Waters and Effluents. EMSLCincinnati. EPA670/4-731001.
63. U.S. EPA. 1989. Method for Conducting Laboratory Toxicity Degradation Eva/u&ions with Complex Effluents. Battelle Report. March 1989.
64. U.S. EPA. 1989. Pollutants of Concern. Office of Marine and Estuarine Protection, Washington, DC.
65. Alabaster, J., and R. Lloyd, eds. 1982. Water Quality Criteria for Fish. 26 ed. Butterworths, London.
66. U.S. EPA. Draft 1991. Guidance on Assessment and Contrd of Bioconcentratoble Contaminants in Surface Wuters. Office of Water Enforcement and Permits, Washington, DC.



## ADDIROMl.REFERENCES

U.S. EPA. 1990. Introlaboratory Study of the Short-Term Chronic U.S. EPA. 1984. Technical Support Document for ConductTest wifh Mysidopsrs Buhia. Office of Marine and Estuarine ing Use Attainability Analysis. Office of Water Regulations Protection, Washington, DC. and Standards, Washington, DC.
U.S. EPA. 1988. Proceedings of the First Notional Workshop on Biological Criteria 1988. U.S. EPA, Region 5: Chicago, Illinois. EPA905/g-89/003.

28

# 2. WATER QUALITY CRITERIA AND STANDARDS

2.1 INTRODUCTION

The foundation of a water quality-based toxics control program consists of the State water quality standards applicable to the waterbody. The following discussion describes the regulatory and technical considerations for application of water quality standards.



### 2.1.1 Overview of WaterQuality Standards A water quality standard defines the water quality goals of a water body, or portion thereof, by designating the use or uses to be made of the water, by setting criteria necessary to protect the uses, and by establishing antidegradation policies and implementation procedures that serve to maintain and protect water quality. States adopt water quality standards to protect public health or welfare, enhance 

### the quality of water, and serve the purposes of the Clean Water Act (CWA). “Serve the purposes of the Act” (as defined in Sections 101(a), 101(a)(2), and 303(c) of the Act) means that water quality standards should (1) include provisions for restoring and maintaining chemical, physical, and biological integrity of State waters; (2) provide, wherever attainable, water quality for the protection and propagation of fish, shellfish, and wildlife and recreation in and on the water (“fishable/swimmable”); and (3) consider the use and value of State waters for public water supplies, propagation of fish and wildlife, recreation, agriculture and industrial purposes, and navigation.
The CWA describes various uses 

### of waters that are considered desirable and should be protected. These uses include public water supply, recreation, and propagation of fish and wildlife. The States are free to designate more specific uses (e.g., cold water and warm water aquatic life), or to designate uses not mentioned in the CWA, with the exception that waste transport and assimilation is not an acceptable designated use (see 40 CFR131.10(a)). EPA’s regulations emphasize the uses specified in CWA Section 101(a)(2), 

### but do not preclude other beneficial 

### usesand subcategories of uses as determined by the State.
When designating uses, States should give careful consideration to whether uses that will support the “fishable and swimmable” goal of Section 101(a)(2) are attainable. If the State does not designate uses in support of this goal, the State must perform a use attainability analysis under Section 131.10(j) of the standards regulation. States should designate uses for the waterbody that the State determines can be attained in the future. “Attainable uses” are those uses (based on the State’s system of water use classification) 

### that can be achieved when effluent limits under CWA Section 301(b)(1)(A) and (B) and Section 306 are implemented for point source discharges and when cost-effective and reasonable best management practices are implemented for nonpoint sources. The Water Quality Standards regulation specifies the conditions under which States may remove uses or establish subcategories of uses. Among these are that the State must provide opportunity for public hearing. In addition, uses that have been attained in the waterbody on or after November 28, 1975, whether or not they are included in the water quality standards, may not be removed unless a 

### use requiring more stringent criteria is added. These uses are the “existing uses” as defined in 40 CFR 131.3(e). Also, uses that are attainable, as defined above, may not be removed. Removal of a “fishable/ swimmable” use, or adoption of a subcategory of a “fishable/ swimmable” use that requires less stringent criteria, requires the State to conduct a use attainability analysis. Technical guidance on conducting use attainability analyses is available from EPA (e.g., Chapter 3 of 

### the Water Quality Standards Handbook (1983) [1], and Technical Support Manual: Waterbody Surveys and Assessments for Conducting UseAttainability Analyses (1983) [2].
In the Water Quality Standards regulation, Section 131.11 encourages States 

### to adopt both numeric and narrative criteria. Aquatic life criteria should protect against both short-term (acute) and long-term (chronic) effects. Numeric criteria particularly are important where the cause of toxicity is known or for protection against pollutants with potential human health impacts or bioaccumulation potential. Numeric water quality criteria also may be the best way to address nonpoint source pollution problems. Narrative criteria can be the basis for limiting toxicity in waste discharges where a specific pollutant can be identified as causing or contributing to the toxicity but there are no numeric criteria in the State standards or where toxicity cannot be traced to a particular pollutant. Section 131.11(a)(2) requires States to develop implementation procedures that explain how the State will ensure that narrative toxics criteria are met.
EPA’s water quality standards regulation requires each State 

### to adopt, as part of its water quality standards, an antidegradation policy consistent with 40 CFR131.12 and to identify the methods it will use for implementing the policy. Activities covered by the antidegradation policy and implementation methods include both point and nonpoint sources of pollution. Section 131.12 effectively sets out a three-tiered approach for the protection of water quality.
“Tier I” (40 CFR131.12(a)(1)) of antidegradation maintains and protects existing uses and the water quality necessary to protect these uses. An existing use can be established by demonstrating that fishing, swimming, or other uses have actually occurred since November 28, 1975, or that the water quality is suitable to allow such uses to occur, whether or not such uses are designated uses for the waterbody in question. (Compare Sections 131.3(e) and 131.3(f) of the existing regulation.) For example, in an area where shellfish are propagating and surviving in a biologically suitable habitat, the shellfish use isexisting, whether or not people are harvesting the shellfish. The aquatic life protection use is a broad category requiring further explanation, which may be found in the Water Quality Standards Handbook.



### 29

### “Tier II” (Section131.12(a)(2))protectsthewaterquality in waters whose qualityisbetterthanthat necessary to protect“fishable/ swimmable”uses ofthewaterbody.40 CFR 131.12(a)(2) requires that certain proceduresbe followed and certain showings be madebeforeloweringwaterqualityin high-qualitywaters.These showingsmaybecalledan “antidegradationreview.” In no case maywaterqualityonaTierIIwaterbodybeloweredtothe levelat which existinguses areimpaired. TheTierII protectionusuallyis applied on a parameter-by-parameter basis (calledthe definitionalapproach to TierII). Thisapproachisappliedonacase-bycase basis sothat, ifthelevelofanyparameter isbetterthanwater quality standards for that waterbody,then an antidegradation reviewwill be performedfor any activitythat could reducethe levelofthat parameter. Outstandingnationalresource waters(ONRWs) areprovidedthe highestlevelof protection undertheantidegradationpolicy(Tier Ill); no degradation is allowed. ONRWsinclude the highestquality waters of the United States. However, the ONRW antidegradation classification also offersspecialprotection for waters of “exceptional ecological significance,” i.e., those waterbodies that are important, unique,or sensitive ecologically, but whosewaterquality, asmeasured bythe traditional parameterssuchasdissolvedoxygen or pH, may not be particularly high. Watersof exceptional ecologicalsignificancemay also include waterswhose characteristics cannot be described adequately bytraditionalparameters (such as

wetlands andestuaries). Statesmay, at their discretion,adopt certain policiesin their standards affectingthe applicationand implementationof standards. Forexample,policies concerningmixingzones, variances, low-flow exemptions,and schedulesof compliancefor water quality-based permit limitsmaybeadopted. Althoughtheseare areasof Statediscretion, EPAretainsauthority to review and approveor disapprove suchpolicies(see40 CFR 131.13). Guidanceon thesesubjectsis availablefrom EPA’s Officeof Water Regulations and Standards, Criteriaand Standards Division.

2.1.2 Water Quality Standards and State Toxics Control Programs Applicablerequirements for State adoptionofwaterqualitycriteriafor toxicantsvarydepending upon the toxicant. The reason forthisisthatthe 1983waterqualitystandards regulationandthe 1987 amendmentsto the CWA(Pub. L. 100-4) include more specificrequirements for the particulartoxicants listed in CWA Section307(a). Forregulatorypurposes, EPA hastranslated the 65compounds andfamilies ofcompoundslistedin Section 307(a) into 126 specificsubstances that EPA refersto aspriority toxic pollutants.The126prioritytoxic pollutantsarelistedinAppendix Aof40 CFR Part423. Because ofthe morespecificrequirements for priority toxic pollutants, it is convenientto organizethe requirementsapplicableto Stateadoption of criteriafor toxicants into threecategories:

• Requirements applicableto priority toxic pollutants that have beenthe subjectof CWASection304(a)(1)criteria guidance

• Requirements applicableto priority toxic pollutants that havenot beenthesubject ofCWA Section 304(a)(1)criteria guidanceand • Requirements applicable to all other toxicants (i.e., nonprioritytoxic pollutants). The criteria requirementsapplicableto priority toxic pollutants (i.e.,thefirsttwo categories above),arespecified in CWA Section 303(c)(2)(B). On December2, 1988, EPA sent “Guidancefor StateImplementationof WaterQuality Standards for CWASection 303(c)(2)(B)”to eachof itsRegions andto eachState water pollution controlagency. Theguidancecontainedthreeoptions for implementing the new numericcriteria requirementsof the Act: (1)adoptStatewide numericcriteriainstandards forallthose priority toxic pollutants for which EPA has published national criteria; (2) adopt numericcriteriafor only those priority toxic pollutants and thosestreamsegments where the dischargeor presence ofthe pollutant could reasonably beexpectedto interferewith designated uses; or(3)adoptaspecific procedureinthe standards to “translate” the State’snarrative“free from toxics” standard to derivednumericcriteria. The transmittal memorandumfor the Section303(c)(2)(B)national guidanceexpresses the OfficeofWaterposition regarding priority toxic pollutants that may “reasonablybe expected” to interfere with designated uses.Thatmemorandum andguidance established a rebuttablepresumptionthat any information indicating that suchpollutantsaredischarged or presentin surface waters(now or in the future) issufficientjustificationto require adoptionorderivationofnumerical criteria.Thegoalisnotjustto identify pollutantsthat arealreadyimpactingsurface waters,but ratherto identifypollutantsthat maybeimpactingsurface waters now,or havethepotentialtodosointhefuture. Lack ofdetailed orwidespread monitoringdataisnot anacceptable basis to omit numerical(orderivednumerical)criteriafromwaterqualitystandardsunderOptions2and3. Even a limitedamountof monitoring data indicating the dischargeor presence of priority toxic pollutants in surface waters is sufficientbasisto concludethat numerical(orderivednumerical)criteriaarenecessary. WhereStatesselectan Option 2 or 3 approach, Statesmust include, as part of the rationalesupporting the adopted standards,the information usedin determiningwhich priority toxic pollutants requirecriteria. Wherethere isuncertaintyabout the need for criteriafor specificpriority toxic pollutants, the State shouldadopt(orderive)criteriaforsuchpollutantsso asto erron the sideof environmentalprotection and pollution prevention. Thisapproachisappropriategiventhegenerallack of monitoring data forprioritytoxicpollutants; itwillprovidemaximumprotection totheenvironmentbyanticipating,ratherthan reactingto,water quality problems. Forpriority toxic pollutantsforwhich EPA hasnot issued Section 304(a)(1)criteria guidance,CWASection303(c)(2)(B)requires States to adopt criteriabased on biologicalmonitoring or assessmentmethods.Thephrase “biological monitoringorassessment methods” includes(1) whole effluenttoxicity control methods, (2) biologicalcriteria methods,or (3) other methodsbasedon biological monitoring or assessment.The phrase“biological monitoring or assessment methods” in its broadestsensealso includescriteriadevelopedthrough translatorprocedures. This broadinterpretationofthat phraseisconsistent with EPA’s policy of applying chemical-specific, biological,andwhole effluenttoxicity methodsindependentlyin an integratedtoxicscontrol program. Italsoisconsistent with the intent of Congress to expand

30

### State standards programsbeyondchemical-specific approaches. WhereEPA hasnot issuedSection304(a)criteriaguidance,but availablelaboratorytoxicity (bioassay) dataaresufficientto supportderivationofchemical-specific criteria,States shouldconsider deriving and adopting numeric criteria for such priority toxic pollutants. This is particularly important where other componentsof a state’snarrativeCriterionimplementationprocedure (e.g.,whole effluenttoxicity controlsor biologicalcriteria) may not ensure full protectionofdesignated uses.For some pollutants, acombinationof chemical-specific andother approaches isnecessary(e.g., pollutantswhere bioaccumulationin fish tissueor waterconsumptionbyhumansisaprimaryconcern). Criteriarequirements applicableto toxicantsthat arenot priority toxic pollutants (i.e.,the third categoryabove),arespecifiedin the 1983waterqualitystandards regulation(see 40 CfR131.11). Undertheserequirements,Statesmust adopt criteriabasedon soundscientificrationalethat coversufficientparameters to protect designateduses. Both numeric and narrativecriteria are addressed bytheserequirements. Numericcriteriaarerequiredwheresuchcriteriaarenecessary to protectdesignateduses. Numericcriteriato protectaquaticlife should be developedto addressboth short-term (acute) and long-term(chronic)effects.Saltwater species, aswellasfreshwater species, mustadequatelybe protected. Adoptionof numeric criteriaisparticularlyimportantfortoxicantsknownto beimpairing surface watersandfortoxicantswith potential humanhealth impacts(e.g.,thosewith high bioaccumulationpotential). Human health should be protectedfrom exposureresultingfrom consumptionofwaterandfishorotheraquaticlife(e.g.,mussels, crayfish).Numericwaterqualitycriteriaalsoareusefulinaddressing nonpoint sourcepollution problems. In evaluatingwhetherchemical-specific numericcriteriafor toxicants are required, Statesshould considerwhether other approaches(such aswhole effluent toxicity criteria or biological controls)will ensurefull protectionof designateduses. Asmentionedabove,acombinationof independentapproaches maybe requiredin somecases to supportthe designated uses and complywith the requirements ofthe waterquality standardsregulation (e.g.,pollutantswherebioaccumulationinfishtissue orwater consumptionby humansisaprimaryconcern). Tosupplementnumericcriteriafortoxicants,all States alsohave adoptednarrativecriteriafortoxicants. Suchnarrativecriteriaare statements that describethe desiredwaterquality goal,suchas thefollowing: AllState watersmust,atalltimesandflows,befreefrom substances that aretoxicto humansoraquaticlife. EPA considers that the narrativecriteria apply to all designated usesat all flows unlessspecifiedotherwise in a State’swater quality standards. EPA alsobelieves that no acutelytoxic condition mayexistinanyState watersregardless ofdesignated use(54 FR

23875). Narrativecriteriacan bethe basis for establishingchemical-specificlimitsforwastedischarges wherea specificpollutant canbe identifiedascausingor contributing to thetoxicity andthe State has not adopted chemical-specific numeric criteria. Narrative criteriaalsocanbethe basis for establishing wholeeffluenttoxicitycontrolsrequiredby EPA regulations at40 CFR 122.44(d)(1)(v). To ensurethat narrativecriteriafor toxicants are attained,the water quality standardsregulation requires Statesto develop implementation procedures(see40 CFR131.11(a)(2)). Such implementationprocedures (Box2-1)shouldaddress all mechanisms used by the Stateto ensurethat narrative criteria are attained. Because implementationof chemical-specific numeric criteria is a key component of Statetoxics control programs, narrative criteria implementation proceduresmust describeor referencethe State’sproceduresto implement such chemicalspecificnumericcriteria(e.g.,procedures for establishingchemical-specific permitslimitsunderthe NPDES permitting program). Implementationprocedures alsomustaddress Stateprograms to control whole effluent toxicity and may addressprogramsto implement biological criteria,where such programshave been developedby the State. Implementationprocedurestherefore serveasumbrelladocumentsthat describehow the State’svarioustoxics control programsare integratedto ensureadequate protectionfor aquaticlife and human health and attainment of the narrativetoxics criterion. In essence, the procedureshould applythe“independentapplication” principle,which provides for independentevaluations ofattainmentofadesignatedusebased on chemical-specific, wholeeffluenttoxicity, andbiologicalcriteria methods(see Chapter1, Reference 56). EPA encourages, and mayultimately require,Stateimplementation procedures to providefor implementationof biologicalcriteria. However,the regulatorybasisfor requiring whole effluent toxicitycontrolsisclear.EPA regulations at40 CFR 122.44(d)(1)(v) require NPDES permitsto contain whole effluent toxicity limits whereapermitteehasbeenshownto cause, havethe reasonable potential to cause,or contribute to an in-streamexcursionof a narrativecriterion. Implementationof chemical-specific controls alsoisrequiredby EPA regulations at 40 CFR 122.44(d)(1). State implementation proceduresshould, at a minimum, specifyor reference methodsto beusedin implementingchemical-specific and whole effluent toxicity-basedcontrols, explain how these methodsareintegrated,andspecifyneededapplicationcriteria. In additionto EPA’s regulationat40 CFR Part131,EPA hasregulationsat 40 CFR 122.44that coverthe National Surface Water ToxicsControl Program.Theseregulationsintrinsicallyarelinked to the requirements to achieve waterquality standards, and specificallyaddress the control of pollutantsboth with and without numericcriteria. Forexample,Section122.44(d)(1)(vi)provides the permitting authority with severaloptions for establishing effluent limits when a Statedoes not havea chemical-specific numericcriteriaforapollutant presentin aneffluentataconcentration that causes or contributesto a violation of the State’s narrativecriteria.

2.2 GENERAL CONSIDERATIONS

2.2.1 Magnitude,Duration,andfrequency Asstatedearlier,criteria are specifications of water quality designedto ensureprotectionofthedesignated use. EPA criteriaare

31

### __---- -. --. - - .- --

Box 2-1. Components of an Ideal State Implementation Procedure

l Specific, scientifically defensible methods by which the State will implement its narrative toxics standard for all toxicants, including:

- Methods for &emical-specific criteria, including methods for applying chemical-specific criteria in permits, developing or modifying chemical-specific criteria via a “translator procedure” (defined and discussed below), and calculating site-specific criteria based on local water chemistry or biology

- Methods for developing and implementing whole effluent toxic&v criteria and/or controls

- Methods for developing and implementing bioloaical criteria.

l Integration of these methods in the State’s toxics control program (i.e., how the State will proceed when the specified methods produce conflicting or inconsistent results).

l Application criteria and information that are needed to apply numerical criteria, for example:

- Methods the State will use to identify thosepollutants to be regulated in a specific discharge

- An incremental cancer risk level for carcinogens

- Methods for identifying compliance thresholds inpermits where calculated limits are below detection

- Methods for selecting appropriate hardness, pl-l, and temperature variables for criteria expressed as functions

- Methods or policies controlling the size and in-zone quality of mixing zones

- Design flows to be used in translating chemical-specific numeric criteria for aquatic life and human health into permit limits

- Other methods and information that will be needed to apply standards on a case-by-case basis.

developed as national recommendations to assist States in developing their standards and to assist in interpreting narrative standards. EPAcriteria or guidance consist of three components:

l Magnitude-How much of a pollutant (or pollutant parameter such as toxicity), expressed as a concentration, isallowable.

l Duration-The period of time (averaging period) over which the instream concentration isaveraged for comparison with criteria concentrations. This specification limits the duration of concentrations above the criteria.

l Frequency-How often criteria can be exceeded.
A typical aquatic life water quality criteria statement contains a concentration, averaging period, and return frequency, stated in the following format:
The procedures described in the Guidelines for Deriving National Water Quality Criteria for the Protection ofAquatic Organisms and Their Uses indicate that, except possibly where a locally important species is very sensitive, 1]1 aquatic organisms and their uses should not be affected unacceptably if the four-day average concentration of (2J does not exceed 43. pg/L more than once every three years on the average and if the one-hour average concentration does not exceed mpg/L more than once every three years on the average.
In this example generic statement, the following terms are inserted at: (1) - either “freshwater” or “salhuater’ (2) - the name of the pollutant (3) - the lower of the chronic-effect or residue-based concentrations as the criterion continuous concentration (CCC) (4)- the acute effect-based criterion maximum concentration (CMC). Defining water quality criteria with an appropriate duration and frequency of excursions helps to ensure that criteria appropriately are considered in developing wasteload allocations (V/L&), which are then translated into permit requirements. Duration and frequency may be defined in the design stream flow appropriate to the criterion. However, in these cases, the State should provide an evaluation that the selected design stream flow approximates the recommended duration and frequency.

32

### 2.2.2 Mixlq Zones It is not always necessary to meet all water quality criteria within the discharge pipe to protect the integrity of the waterbody as a whole. Sometimes it is appropriate to allow for ambient concentrations above the criteria in small areas near outfalls. These areas are called mixing zones. Since these areas of impact, if disproportionately large, could potentially adversely impact the productivity of the waterbody, and have unanticipated ecological consequences, they should be carefully evaluated and appropriately limited in size. As our understanding of pollutant impacts on ecological systems evolves, there may be cases identified where no mixing zone is appropriate.
To ensure mixing zones do not impair the integrity of the waterbody, it should be determined that the mixing zone will not cause lethality to passing organisms and, considering likely pathways of exposure, that thereare no significant human health risks. One means to achieve these objectives is to limit the size of the area affected by the mixing zones.
For application of two-number aquatic life criteria, there may be up to two types of mixing zones (Figure 2-l). In the zone immediately surrounding the outfall, neither the acute nor the chronic criterion is met. The acute criterion is met at the edge of this zone. In the next mixing zone, the acute, but not the chronic, criterion is met. The chronic criterion is met at the edge of the second mixing zone.
In the general case, where a State has both acute and chronic aquatic life criteria, as well as human health criteria, independently established mixing zone specifications may apply to each of the three types of criteria. The acute mixing zone may be sized to prevent lethality to passing organisms, the chronic mixing zone

’ Chronic criteria met

Figure 2-1, Diagram of the Two Parts of the Mixing Zone sized to protect the ecology of the waterbody as a whole, and the health criteria mixing zone sized to prevent significant human risks. For any particular pollutant from any particular discharge, the magnitude, duration, frequency, and mixing zone associated with each of the three types of criteria will determine which one most limits the allowable discharge. Mixing zone allowances will increase the mass loadings of the pollutant to the waterbody, and decrease treatment requirements. They adversely impact immobile species, such as benthic communities, in the immediate vicinity of 

### the outfall. 

### Becauseof these and other factors, mixing zones must be applied carefully, so as not to impede progress toward the CWA goals of maintaining and improving water quality. EPA recommendations for allowances for mixing zones, and appropriate cautions about their use, are contained in this section.
The CWA allows mixing zones at the discretion of the State [l]. EPA recommends that States have a definitive statement in their standards on whether or not mixing zones are allowed. Where mixing zones provisions are part of the State standards, the State should describe the procedures for defining mixing zones.
To determine that a mixing zone is sized appropriately for aquatic life protection, water quality conditions within the mixing zone may be compared to laboratory-measured or predicted toxicity bench marks as follows:
It is not necessary to meet chronic criteria within the mixing zone, only at the edge of the mixing zone. Conditions within the mixing zone would thus 

### not be adequate to ensure survival, growth, and reproduction of all organisms that might otherwise attempt to reside continuously within the mixing zone.
If acute criteria (CMC derived from 48- to 96hour exposure tests) are met throughout the mixing zone, no lethality should result from temporary passage through the mixing zone. If acute criteria are exceeded no more than a few minutes in a parcel of water leaving an outfall (as assumed in deriving the Section 4.3.3 options for an outfall velocity of 3 m/set, and a size of 50 times the discharge length scale), this likewise assures no lethality to passing organisms.
If a full analysis of concentrations and hydraulic residence times within the mixing zone indicates that organisms drifting through the plume along the path of maximum exposure would not be exposed to conc.entrations exceeding the 

### acute criteria when averaged over the 1-hour (or appropriate site-specific) averaging period for acute criteria, then lethality to swimming or drifting organisms ordinarily should not be expected, even for rather fast-acting toxicants. In many situations, travel time through the acute mixing zone must be less than roughly 15 minutes if a 1-hour average exposure is not to exceed the acute criterion.
Where mixing zone toxicity is evaluated using the probit approach described in the water quality criteria “Bluebook” [3], or using models of toxicant accumula33tion and action in organisms (described by Mancini [4] or Erickson et al. [5]), the phenomenon of delayed mortality should be taken into account before judging the mixing zone concentrations to be safe.
The above recommendations assume that the effluent is repulsive, such that free-swimming organisms would avoid the mixing zones. While most toxic effluents are repulsive, caution is necessary in evaluating attractive mixing zones of known effluent toxicity, and denial of such mixing zones may well be appropriate. It also is important to ensure that concentration isopleths within any plume will not extend to restrict passage of swimming organisms into tributary streams.
In all cases, the size of the mixing zone and the area within certain concentration isopleths should be evaluated for their effect on the overall biological integrity of the waterbody. If the total area affected by elevated concentrations within all mixing zones combined is small compared to the total area of a waterbody (such as a river segment), then mixing zones are likely to have little effect on the integrity of the waterbody as a whole, provided that they do not impinge on unique or critical habitats. EPA has developed a multistep procedure for evaluating the overall acceptability of mixing zones [6].
For protection of human health, the presence of mixing zones should not result in significant health risks, when evaluated using reasonable assumptions about exposure pathways. Thus, where drinking water contaminants are a concern, mixing zones should not encroach on drinking water intakes. Where fish tissue residues are a concern (either because of measured or predicted residues), mixing zones should not be projected to result in significant health risks to average consumers of fish and shellfish, after considering exposure duration of the affected aquatic organisms in the mixing zone, and the patterns of fisheries use in the area.
While fish tissue contamination tends to be a far-field problem affecting entire waterbodies rather than a narrow-scale problem confined to mixing zones, restricting or eliminating mixing zones for bioaccumulative pollutants may be appropriate under conditions such as the following:

l Mixing zones should be restricted such that they do not encroach on areas often used for fish harvesting particularly of stationary specres such as shellfish.

l Mixing zones might be denied where such denial is used as a device to compensate for uncertainties in the protectiveness of the water quality criteria or uncertainties in the assimilative capacity of the waterbody.



## 2.3 WATER QUALITY CRITERIA FORAQUATIC LIFE PROTECTION

2.3.1 Devekyunentf%ocess for Criieria The development of national numerical water quality criteria for the protection of aquatic organisms isa complex process that uses information from many areas of aquatic toxicology. (See Reference 7 for a detailed discussion of this process.) After a decision is made that a national criterion is needed for a particular material, all available information concerning toxicity to, and bioaccumulation by, aquatic organisms is collected and reviewed for acceptability. If enough acceptable data for 48- to 96-hour toxicity tests on aquatic animals are available, they are used to derive the acute criterion. Ifsufficient data on the ratio of acute to chronic toxicity concentrations are available, they are used to derive the chronic or long-term exposure criteria. Ifjustified, one or both of the criteria may be related to another water quality characteristic, such as pti, temperature, or hardness. Separate criteria are developed for freshwaters and saltwaters.
The water quality standards regulation allows States to devetop numerical criteria or modify EPA’s recommended criteria to account for site-specific or other scientifically defensible factors. In cases where additional toxicological data are needed to modify or develop criteria, the discharger may be required to generate the data. Guidance on modifying national criteria is found in the handbook [l]. When a criterion must be developed for a chemical for which a national criterion has not been established, the regulatory authority should refer to the Guidelines for Deriving Criterio for Aquatic Lifeand Human He&h (see 45 FR79341, November 28, 1980, and 50 FR30784, july 29, 1985).



### 2.3.2 MagnitudeforSfngleChemicals Water quality criteria for aquatic life contain two expressions of allowable magnitude: a CMC to protect against acute (shortterm) effects and a CCC to protect against chronic (long-term) effects. EPA derives acute criteria from 48- to 96-hour tests of lethality or immobilization. EPA derives chronic criteria from longer-term (often greater than 28-day) tests that measure survival, growth, reproduction, or in some cases, bioconcentration.
Most State standards include numerical criteria for a limited number of individual toxic chemicals. Therefore, evaluation and control of toxic pollutants is based on maintenance of the designated use and often relies on the narrative criterion prohibiting toxic substances in toxic amounts. The adverse effects of concern will depend on the designated use and the chemical. Bioaccumulation of chemicals in aquatic organisms, toxicity to these organisms, the potential for additivity, antagonism, synergism, and persistence of the chemicals may be important. Available information on the toxic effects of the chemical is used when standards do not include specific numerical criteria. Such information can include EPA criteria documents, published literature reports, or studies conducted by the discharger.
As mentioned in Section 21.2, water quality-based controls may be based directly on the State’s technical determination of what concentration of a specific pollutant meets the State’s narrative “free from” toxics criterion. Although EPAwater quality standards regulation requires that the State’s process for implementing its narrative criterion be described in the State standards, there is no requirement that this concentration be adopted as a numerical criterion in State water quality standards prior to use in developing water quality-based controls and therefore a case-by-case interpretation of the narrative criterion may be necessary.

34

### 2.3.3 hbpitude for Mm/e Hfket Toxicity Criteria for toxicity in current State standards range from the narrative prohibition (e.g., no discharge of toxic chemicals in toxic amounts) to detailed requirements that specify the test species and the allowable toxicity level. At present, there are no national criteria developed under CWA Section 304(a) for whole effluent toxicity. Acute and chronic toxicity units (TUs) are a mechanism for quantifying instream toxicity using the whole effluent approach. The procedure to implement the narrative criteria using a whole effluent approach should specify the testing procedure, the duration of the tests (acute or chronic), the test species, and the frequency of testing required.
EPA’s recommended magnitudes for whole effluent toxicity are as follows (again, two expressions of allowable magnitude are used): a CMC to protect against acute (short-term) effects and a CCC to protect against chronic (long-term) effects. For acute protection, the CMC should be set at 0.3 acute toxic unit (TU,) to the most sensitive of at least three test species.
The selection of test species for testing the effluent is not critical provided species from ecologically diverse taxa are used (e.g., a fish, an invertebrate, and a plant). The factor of 0.3 is used to adjust the typical LCso endpoint of an acute toxicity test (50 percent mortality) to an LCT value (virtually no mortality). Specifically, a factor of 0.3 was found to include 91 percent of observed LCT to LCso ratios in 496 effluent toxicity tests as illustrated in Figure 2-2. This figure presents effluent toxicity data from many years of toxicity testing of both industrial and municipal effluents by the Environmental Services Division, U.S. EPA Region IV, Athens, Georgia



### 2 ! P t 130

120 4 110
100
90 1 125
-

66
-

- 8

d k d 42

-

# 1.

2

9 ': 5 0 - 8 + z d LCj/LC= Ratm

Figure 2-2. LCT to LC% Ratios for Effluent Toxicity Tests For chronic protection, the CCC should be set at 1.0 chronic toxic unit (TU,) to the most sensitive of at least three test species. The selection of test organisms is as described above. A 1.OTU, is applied at the edge of the mixing zone to prevent any chronic toxicity in the receiving water outside the mixing zone.


### 2.3.4 lhuathe forSiw#iehmhakwnd I#& EtYluent Toxicity The quality of an ambient water typically varies in response to variations of effluent quality, stream flow, and other factors. Organisms in the receiving water are not experiencing constant, steady exposure but rather are experiencing fluctuating exposures, including periods of high concentrations, which may have adverse effects. Thus, EPA’s criteria indicate a time period over which exposure is to be averaged, as well as a maximum concentration, thereby limiting the duration of exposure to elevated concentrations.
For acute criteria, EPA recommends an averaging period of 1 hour. That is, to protect against acute effects, the 1-hour average exposure should not exceed the CMC. The 1-hour acute averaging period was derived primarily from data on response time for toxicity to ammonia, a fast-acting toxicant. The l-hour averaging period is expected to be fully protective for the fastest-acting toxicants, and even more protective for slower-acting toxicants. Scientifically justifiable alternative (site-specific) averaging periods can be derived from (1) data relating toxic response to exposure time, if coupled with considerations of delayed mortality (mortality occurring after exposure has ended), or (2) models of toxicant uptake and action, such as presented by Erickson [S] and Mancini et al. [43.
In practice, 1-day periods are the shortest periods for which WfA modelers and enforcement personnel have adequate data. Attainment of the duration criterion can be ensured by paying particular attention to short-term effluent variability and requiring measures to control variability (e.g., installation of equalization basins) when needed.
For chronic criteria, EPA recommends an averaging period of 4 days. That is, the 4-day average exposure should not exceed the CCC. Different chronic averaging periods could be derived, depending on the nature of the pollutant and the toxic endpoint of concern (e.g., the rate of uptake and accumulation, and the mode of action).
The toxicity tests used to establish the national criteria are conducted using steady exposure to toxicants usually for at least 28 days. The test concentrations do not fluctuate as much astypically occurs instream. As the period of averaging increases, so too does the period of time the exposure concentrations can be above the criterion concentration without exceeding the average. The significant consideration involved in setting duration criteria is how long the exposure concentration can be above the criterion 

### concentration without unacceptably affecting the endpoint of the test (e.g., survival, growth, or reproduction). EPA selected the 4-day averaging period based on the shortest duration in which chronic effects are sometimes observed for certain species and toxicants, and thus should be fully protective even for the fastest-acting toxicants.

35

### - -- --__-- - - - - _-____


### 2.3.5 Fmpency for Single Chemkals and Wok? Effheat roxmty To predict or ascertain the attainment of criteria it is necessary to specify the allowable frequency for exceeding the criteria. This is because it is statistically impossible to project that criteria will never be exceeded. As ecological communities are naturally subjected to a series of stresses, the allowable frequency of pollutant stress may be set at a value that does not significantly increase the frequency or severity of all stresses combined.
EPA recommends a once in 3-year average frequency for excursions of both acute and chronic criteria. These recommendations apply to both chemical-specific and whole effluent approaches. However, the allowable frequency depends on sitespecific factors. To implement alternative frequencies, site-specific factors (see Appendix D) or other data or analyses should be taken into account. In all cases, the recommended frequency applies to actual ambient concentrations, and excludes the influence of measurement imprecision.
EPAestablished its recommended frequency as part of its Guidelines for Deriving Criteria, last issued in 1985 [8]. EPA selected the 3- year return interval with the intent of providing a degree of protection roughly equivalent to a 7410 design flow condition, and with some consideration of rates of ecological recovery from a variety of severe stresses. Because of the nature of the ecological recovery studies available, the severity of criteria excursions could not be related rigorously to the resulting ecological impacts. Nevertheless, EPA derives its criteria intending that a single marginal criteria excursion (i.e., a slight excursion over a l-hour period for acute or over a 4-day period for chronic) would result in little or no ecological effect and require little or no time for recovery. If the frequency of marginal criteria excursions is not high, it can be shown that the frequency of severe stresses, requiring measurable recovery periods, would be extremely small. EPAthus expects the 3-year return interval to provide a very high degree of protection.
Field studies indicate that many discharge situations are affected both by predictable and measurable discharges of toxicants and by unpredrctable spills of toxic substances. In most cases, the dischargers were unaware that spills were occurring. These spills are a second source of stress for the community and decrease recovery potential. An aggressive program to minimize, contain, and treat spills should be in place at any plant where the potential for spills exists.
The concentration, duration, and frequency provisions of the criteria are implemented through the development of WLAs and water quality-based effluent limits. As discussed in Chapter 4, the duration and frequency recommendations are implemented directly if a dynamic modeling approach is used to develop WLAs and permit limits. However, if a steady-state approach is used, a design condition is needed for the calculations.
For the protection of aquatic life, the duration and frequency recommendations provided above have been used to develop recommended design flows for steady-state modeling. Chapter 4 discusses these recommended design flows. Traditionally, most water quality-based permits for point source discharges had been tied to the 7-day, once in lo-year, low-flow conditions. The reason for this is that critical conditions for perennial point source discharges occur, in general, during the low-flow period. Currently, State laws and regulations generally state that water quality standards are applicable to the 7-day, 1Oyear low-flow or higher flow conditions.
It should be noted that EPA’s water quality criteria for aquatic life protection are applicable at all flow conditions, low as well as high. These criteria and their specified duration and frequency, if adopted into or used to interpret State water quality standards, may be used as the basis for total maximum daily load (TMDL) after considering seasonal flow and loading scenarios. The concentration, duration, and frequency provisions of EPA’s water quality criteria can be modified to account for site-specific conditions. As States have started using the new two-number water quality criteria for perennial as well as intermittent discharges such as combined sewer overflows, urban runoff, etc., their proper use in the context of the TMDL/WLA process needs to be emphasized.



## 2.4 WATER QUALITY CRITERIA FORHUMAN HEALTR PROTECTION

24.1 overriew There are a number of key elements of State water quality standards and implementation procedures relevant to human health protection. States must determine ambient standards for the two primary human exposure routes, fish consumption and drinking water. States must then establish whether mixing zones will apply, and, if so, determine the design conditions.
State standards or their implementation procedures often specify the risk level for carcinogens; methods for identifying compliance thresholds in permits where calculated limits are below detection; and methods for selecting appropriate hardness, pH, and temperature variables for criteria. However, if State standards do not specify these items, then the permitting authority must develop water quality-based effluent limits based upon either an interpretation of the State’s water quality standards or EPA’s criteria and procedures.
The purpose of the following section is to provide a review of EPA’s procedures used to develop assessments of human health effects in developing water quality criteria and reference ambient concentrations. A complete human health effects discussion is included in the (draft) Guidelines and Methodology Used in the Preparation of Health Effects Assessment Chapters of the Consent Decree Water Documents by EPA’s Environmentat Criteria and Assessment Office (ECAO). The procedures contained in the ECAO document are used in the development and updating of EPA water quality criteria and may be used in developing reference ambient concentrations (RACs) for those pollutants lacking EPA human health criteria. Although the same procedures are used to develop criteria and RAG, only those values that are subjected to the regulatory process of regional, State, and public comment can be considered “criteria.” RACs may be applied as site-specific interpretations of narrative standards and as a basis for permit limits under 40 CfR 122.44 (d)(l)(vi).

36

### ---__--

Procedures also are provided in this chapter to develop values called reference tissue concentrations (RTCs) that can be used in assessing or monitoring fish tissues for unacceptable residues.

Water quality criteria for human health contain only a single expression of allowable magnitude; a criterion concentration generally to protect against long-term (chronic) human health effects. Currently, national policy and prevailing opinion in the expert community dictate that the duration for human health criteria for carcinogens be derived assuming lifetime exposure, taken to be a 70-year time period. The duration of exposure assumed in deriving criteria for noncarcinogens is more complicated due to a wide variety of endpoints: some developmental (and thus age-specific and perhaps sex-specific), some lifetime, and some, such as organoleptic effects, not duration-related at all. Thus, appropriate durations depend on the individual noncarcinogenic pollutants and the endpoints or adverse effects being considered.

A complete human exposure evaluation for toxic pollutants of concern for bioaccumulation would not only encompass estimates of exposures due to fish consumption, but also exposure due to background concentrations and other exposure routes, including recreational and occupational contact, dietary intake from other than fish, inhalation of air, and drinking water. However, the focus of this document is on ingestion of contaminated fish tissue, a direct human exposure route of potentially significant risk. (For the human health sections in this document the term “fish” generally is used to mean both fish and shellfish.) The consumption of contaminated fish tissue is of serious concern since the presence of even extremely low ambient concentrations of bioaccu-mulative pollutants (sublethal to aquatic life) in surface waters, can result in residue concentrations in fish tissue that can pose a human health risk. Other exposure route information should be considered and incorporated in human exposure evaluations to the extent it is available.
Levels of actual human exposures from consuming contaminated fish vary depending upon a number of case-specific consumption factors. These factors include type of fish species consumed, type of fish tissue consumed, tissue lipid content, consumption rate and pattern, and food preparation practices. In addition, depending on the spatial variability in the fishery area, the behavior of the fish species, and the point of application of the RAC or criterion, the average exposure of fish may be only a small fraction of the expected exposure at the point of application of the criterion. If an effluent attracts fish, the average exposure might be greater than the expected exposure.
With shellfish, such as oysters, snails, and mussels, whole body tissue consumption commonly occurs, whereas with fish, muscle tissue and roe are most commonly eaten. This difference in the types of tissues consumed has implications for the amount of available bioaccumulative contaminants likely to be ingested. Whole body shellfish consumption presumably means ingestion of the entire burden of bioaccumulative contaminants. However, with most fish, selective cleaning and removal of internal organs, and sometimes body fat as well, from edible tissues, may result in removal of much of the lipid material in which bioaccumulative contaminants tend to concentrate.



### 24.4 FfshconsunrptlonV8hles EPA’s human health criteria have assumed a human body weight of 70 kg and the consumption of 0.0065 kg of fish and shellfish per day. Based on data collected in 1973-l 974, the national per capita consumption of freshwater and estuarine fish was estimated to average 6.5 g/day. Per capita consumption of all seafood (including marine species) was estimated to average 14.3 g/day. The 95th percentile for consumption of all seafood by individuals over a period of 1 month was estimated to be 42 g/day 191. The mean lipid content of fish tissue consumed in this study was estimated to be 3.0 percent [lo]. Currently, four levels of fish consumption are provided in the EPA guidance manual, Assessing Human Health Risk from Chemically Contaminated Fishand Shellfish. These are:

l 6.5 g/day to represent an estimate of average consumption of fish and shellfish from estuarine and freshwaters by the entire U.S. population [9]. This fish consumption level is based on the average of both consumers and nonconsumers of fish.

l 20 g/day to represent an estimate of the average consumption of fish and shellfish from marine, estuarine, and freshwaters by the U.S. population [l 11. This average fish consumption level also includes both consumers and nonconsumers of fish.

l 165 g/day to represent consumption of fish and shellfish from marine, estuarine, and freshwaters by the !B.!Xh percentile of the U.S. population consuming the most fish or seafood [12].

l 180 g/day to represent a “reasonable worst case” based on the assumption that some individuals would consume fish at a rate equal to the combined consumption of red meat, poultry, fish, and shellfish in the United States (EPA Risk Assessment Council assumption based on data from the U.S. Department of Agriculture Nationwide Food Consumption Survey of 1977-1978).
EPA currently is updating the national estuarine and freshwater fish and shellfish consumption default values and will provide a range of recommended national consumption values. This range will include mean values appropriate to the population at large, and values appropriate for those individuals who consume a relatively large proportion of fish in their diets (maximally exposed individuals). Many States use the EPA’s 6.5 giday consumption value. However, some States (e.g., Wisconsin, Louisiana, Illinois, and Arizona) use the above mentioned 20 g/day value. For salt waters Delaware uses another EPA value, 37 g/day [13]. In general, EPA recommends that the consumption values used in deriving RACs from the formulas in this chapter reflect the most current relevant and/ or site-specific information available.

37Table 2-l. Estimated Food Chain Multipliers

The ratio of the contaminant concentrations in fish tissue versus water is termed either the bioconcentration factor (BCF) or the bioaccumulation factor (BAF). Bioconcentration is defined as involving contaminant uptake from water only (not from food). Bioaccumulation is defined as involving contaminant uptake from both water and food. Under laboratory conditions, measurements of tissue/water partitioning generally are considered to involve uptake from water only. On the other hand, both process are likely to apply in the field since the entire food chain is exposed. Trophic Levels

Table 2-l shows the ratio of the BAF to the BCF as a function of the trophic level of the aquatic organism, and the log P (log octanol-water partition coefficient) of the chemical [14]. The BAF/BCF ratio ranges from 1 to 100, with the highest ratios applying to organisms in higher trophic levels, and to chemicals with log P close to 6.5. For chemicals with log P values greater than about 7, there is some uncertainty regarding the degree of bioaccumulation, but generally, trophic level effects appear to decrease due to slow transport kinetics of these chemicals in fish, the growth rate of the fish, and the chemical’s relatively low bioavailability.
Care must be taken in assigning the trophic level since certain fish species may inhabit one source area of contaminated food for only a portion of their life. Under such conditions of migration, fish would only receive a small portion of the chemical and never come into equilibrium. In addition, trophic level for a given fish species will vary with life stage and structure of the food chain.
In this document, bioaccumulation considerations are integrated into the RAC equations in Sections 2.4.7 and 2.4.8 by using food chain multipliers (FMs) with the BCF. The bioaccumulation and bioconcentration factors for a chemical are related as follows: log P 2 3 4 3.5 1.0 1.0 1.0 3.6 1.0 1.0 1 .o 3.7 1.0 1.0 1.0 3.8 1.0 1 .o 1.0 3.9 1.0 1.0 1.0 4.0 1.1 1.0 1.0 4.1 1.1 1.1 1.1 4.2 1.1 1.1 1.1 4.3 1.1 1.1 1.1 4.4 1.2 1.1 1.1 4.5 1.2 1.2 1.2 4.6 1.2 1.3 1.3 4.7 1.3 1.4 1.4 4.8 1.4 1.5 1.6 4.9 1.5 1.8 2.0 5.0 1.6 2.1 2.6 5.1 1.7 2.5 3.2 5.2 1.9 3.0 4.3 5.3 2.2 3.7 5.8 5.4 2.4 4.6 8.0 5.5 2.8 5.9 11 5.6 3.3 7.5 16 5.7 3.9 9.8 23 5.8 4.6 13 33 5.9 5.6 17 47 6.0 6.8 21 67 6.1 8.2 25 75 6.2 10 29 84 6.3 13 34 92 6.4 15 39 98 6.5 19 4.5 100 ~6.5 19.2’ 45’ 100. BAF = FM x BCF

l These recommended FMs are conservative estimates; FMs for log P By incorporating the FM and BCF terms into the RAC equations, values greater than 6.5 may range from the values given to as low as bioaccumulation is addressed. 0.1 for contaminants with very low bioavailability.

In this process, bioaccumulation considerations are included by incorporating the FM term with the BCF in calculating the RTCs and RACs. In Table 2-1, FM values derived from the work of Thomann 114, 151are listed according to log Pvalue and trophic level of the organism. Trophic level 4 organisms are typically the most desirable species for sport fishing and therefore, FMs for trophic level 4 generally should be used in the equations for calculating RTCs and RACs. In those vey rare situations where only lower trophic level organisms are found, e.g., possibly oyster beds, an FM for a lower trophic level may be used in calculating the RTCs and RAG. aped is problematic and subject to uncertainty. The option also is available to develop BAFs experimentally, but this will be extremely resource intensive if done on a site-specific basis with all the necessary experimental and quality controls.

Measured BAFs (especially for those chemicals with log P values above 6.5) reported in the literature should be used when available. To use experimentally measured BAFs in calculating the RAC or RTC, the (FM x BCF) term, is replaced by the BAF in the equations in Sections 2.4.7 and 2.4.8. Relatively few BAFs have been measured accurately and reported, and their application to sites other than the specific ecosystem where they were develEPA recommends using the most current risk information when updating criteria and generating RACs. The Integrated Risk Information System (IRIS) is an electronic online data base of the U.S. EPA that provides chemical-specific risk information on the relationship between chemical exposure and estimated human health effects [ 161. Riskassessment information contained in the IRIS, except as specifically noted, has been reviewed and agreed upon by an interdisciplinary group of scientists representing various program offices within the Agency and represent an Agencywide consensus. Risk assessment information and values are updated monthly and are approved for Agencywide use.

38

### _- - --

The IRIS is intended to make risk assessment information readily available to those individuals who must perform risk assessments and also to increase consistency among risk assessment/risk management decisions. The IRISisavailable to Federal and some State and local environmental agencies through the EPA’s electronic MAIL system and also is available to the public through the Public Health Network and TOXNET. Since IRIS is designed to be a publicly available data base, interested parties may submit studies or documents for consideration by the appropriate interdisciplinary review group for chemicals currently on the IRISor scheduled for review. Information regarding the submission of studies of chemicals may be obtained from the IRISInformation Submission Desk. In addition to chemical-specific summaries of hazard and dose-response assessments, the IRIS contains a series of sections identified by service codes that serve as a user’s guide as well as provide background documentation on methodology. Additional information is available from IRIS Users Support: 513/FTS 6847254.
The IRIScontains two types of quantitative risks values: reference dose (WD) and the carcinogenic potency estimate or slope factor. The RfD (formerly known as the acceptable daily intake or ADI) is the human health hazard assessment for noncarcinogenic (target organ) effects. The carcinogenic potency estimate (formerly known as ql*) represents the upper bound cancer causing potential resulting from lifetime exposure to a substance. The RfD or the oral carcinogenic potency estimate are used in Ihe derivation of an RAC. Appendix H contains the supporting information for derivation of RfDs.
EPA periodically updates risk assessment information including RfDs, cancer potency estimates, and related information on contaminant effects, and reports the current information on IRIS. Since the IRIScontains the Agency’s most recent quantitative risk assessment values, current IRISvalues should be used in developing new RACs. This means 

### that the 1980 human health criteria should be updated with the latest IRIS values. The procedure for deriiing an updated human health water quality criterion would require inserting the current RfD or carcinogenic potency estimate on the IRIS into the appropriate equation in Section 2.4.7 or 2.4.8. Figure 2-3 shows the procedure for determining an updated criterion or RAC using IRISdata. If a chemical has both carcinogenic and noncarcinogenic effects, i.e., both a cancer potency estimate and WD, the carcinogen RAC formula in Section 2.4.8 should be used as it will result in the more stringent RAC of the two.



### 2.4.7 CalculatiugRN& fur ubnciucinogens The RfD is an estimate of the daily exposure to the human population that is likely to be without appreciable risk of causing

Start 

# Q

Evaluate other * sources of data: HEAST. Risk Assistant, drinking water

MCLs, fish 

# Y YES YES 

# Y

L Calculate RAC advisory levels, FDA action levels,

YES

Figure 2-3. Procedure for Revising an EPA Human Health Criterion or Developing a Reference Ambient Concentration

39deleterious effects during a lifetime. The RfD is expressed in units of mg toxicant per kg human body weight per day.
RfDs are derived from the “no observed adverse effect level” (NOAEL) or the “lowest observed adverse effect level” (LOAEL) identified from chronic or subchronic human epidemiology studies or animal exposure (mammal LDSo) studies. [Note: LOAEL and NOAEL refer to animal and human toxicology and are there fore distinct from the aquatic toxicity terms “no observed effect concentration” (NOEC) and the “lowest observed effect concentration” (LOEC)]. Uncertainty factors are then applied to the NOAEL or LOAEL to account for uncertainties in the data associated with variability among individuals, extrapolation from nonhuman test species to humans, data on other than long-term exposures, and the use of an LOAEL [17]. An additional uncertainty may be applied to account for significant weakness or gaps in the data base.
The RfD is a threshold below which effects are unlikely to occur. While exposures above the RfD increase the probability of adverse effects, they do not produce a certainty of adverse effects. Similarly, while exposure at or below the RfD reduces the probability, it does not guarantee the absence of effects 

### in all persons. 

### The RfDs contained in the IRISare values that represent EPA’s consensus (and have uncertainty spanning perhaps an order of magnitude).
For noncarcinogenic effects, an updated criterion or an RAC can be derived using the following equation:
C or RAC (mg/l) = (RfD x wr) - (DT + IN) x WT
where c = RAC = RfD =
W-i= DT =
IN =
WI = FC = L =
FM = BCF = WI +[FCxLxFMxBCF) updated water quality criterion (mg/l) reference ambient concentration (mg/l) reference dose (mg toxicant/kg human body weight/ day) weight of an average human adult (70 kg) dietary exposure (other than fish) (mg toxicant/kg body human weight/day) inhalation exposure (mg toxicant/kg body human weight/day) average human adult water intake (2 liters/day) daily fish consumption (kg fish/day) ratio of lipid fraction of fish tissue consumed to 3 percent food chain multiplier (from Table 3-l) bioconcentration factor (mg toxicant/kg fish divided by mg toxicant/l water) for fish with 3 percent lipid.

If the receiving waterbody is not used as a drinking water source, the factor WI can be deleted. Where dietary and/or inhalation exposure values are unknown, these factors may be deleted from the above 

### calculation. For identified noncarcinogenic chemicals without known RfDs, extrapolation procedures can be used to estimate the RfD (see Appendix H). 

### 2.4.8 Calcm MCs forGucinugms Any human health criterion for a carcinogen is based on at least three interrelated considerations: potency, exposure, and risk characterization. States may make their own judgments on each of these factors within reasonable scientific bounds, but documentation to support their judgments must be clear and in the public record.
Maximum protection of human health from the potential effects of exposure to carcinogens via contaminated fish would require an RAC of zero. The zero level is based upon the assumption of nonthreshold effects (i.e., no safe level exists below which any increase in exposure does not result in an increase in the risk of cancer) for carcinogens. However, because safety does not require the absence of all risk, a numerical estimate of risk (in pgil) that corresponds to a given level of risk for a population of a specified size is selected instead. A cancer risk level is defined as the number of new cancers that may result in a po ulation of specified size due to an increase in exposure (e.g ., 1O- fcnsk \evel= 1 additional cancer in a population of l,OOO,OOO). Cancer risk is calculated by multiplying the experimentally derived cancer potency estimate by the concentration of the chemical in the fish and the average daily human consumption of contaminated fish. The risk for a specified population (e.g., 1,000,OOO people or 1Oe6) is then calculated by dividing the risk level by the specific cancer risk. EPA’s ambient water quality criteria documents provide risk levels ranging from 1Oms to 1Oe7 as examples.
When the cancer potency estimate, or slope factor (formerly known as the ql’), is derived using animal studies, high-dose exposures are extrapolated to low-dose concentrations and adjusted to a lifetime exposure period through the use of a linearized multistage model. The model 

### calculates the upper 95 percent confidence limit of the slope of a straight line that the model postulates to 

### occur at low doses. When based on human (epidemiological) data, the slope factor is based on the observed increase in cancer risk, and is not extrapolated. For deriving RACs for carcinogens, the oral cancer potency estimates or slope factors from the IRISare used.
It is important to note that cancer potency factors may overestimate actual risk. Such potency estimates are subject to great uncertainty due to two primary factors: (1) adequacy of the cancer data base (i.e., human versus animal data) and (2) limited information regarding the mechanism of cancer causation. The actual risk may be much lower, perhaps as low as zero, particularly for those chemicals for which human carcinogenicity information is lacking. Risklevels of 1Oe5,1Om6, and 1Om7 are often used by States as minimal risk levels in interpreting their standards. EPA considers risks to be additive, i.e., the risk from individual chemicals is not necessarily the overall risk from exposure to water. For example, an individual risk level of 1Oe6may yield a higher overall risk level if multiple carcinogenic chemicals are present.
For carcinogenic effects, the RAC can be determined by using the following equation:
C or RAC (mg/l) = (RL x W-l-) ql*[WI+FCxLx(FMxBCF)]

40

### - --

where C = updated water quality criterion (mg/l) RAC = reference ambient concentration (mg/l) RL = risk level (1O-x) WT = weight of an average human adult (70 kg) q1* = carcinogenic potency factor (kg day/mg) WI = average human adult water intake (2 liters/day) FC = daily fish consumption (kg fish/day) L = ratio of lipid fraction of fish tissue consumed to 3 percent FM = food chain multiplier (from Table 3-2) BCF = bioconcentration factor (mg toxicant/kg fish divided by mg toxicant/l water) for fish with 3 percent lipid.
If the receiving waterbody is not used as a drinking water source, the factor WI can be deleted. For identified carcinogenic chemicals without known cancer potency estimate values, extrapolation procedures can be used to estimate the cancer potency.



### 2.4.9 Deriving Uuantttative RiskAssessments in theAbsence of /MS Yahes The RfDs or cancer potency estimates comprise the existing dose factors for developing RAG. When IRIS data are unavailable, quantitative risk level information may be developed according to a State’s own procedures. Some States have established their own procedures whereby dose factors can be developed based 

### upon extrapolation of 

### acute and/or chronic animal data to concentrations of exposure protective of fish consumption by humans. Where no procedure exists, factors may be based upon extrapolation from mammalian or other data using IRIS documentation or information available from other EPArisk data bases. Also, where no other information or procedure exists, drinking water maximum contaminant levels (MCLs) or Food and Drug Administration (FDA) action levels may be used as guidance in developing numerical estimates.



### 2.4.19 DerivingRefemce Tissue fducentratitms furMonitoring Fkh Tissue Where fish tissue evaluations have been used for assessing human health risks, or, perhaps, used for additional routine monitoring where a chemical is below analytical detection limits, the following formulas may be used to calculate an RTC. Readers also should consult EPA’sAssessing Human Health Risksfrom Chemically Contaminated Fishand Shellfish [ 1 71,
The basic equations for deriving RTC (in mg/kg) use the same parameters as in equations 2.1 and 2.2, where BCF is normalized at 3.0 percent lipid:
For noncarcinogens: RTC (mg/kg) = (RFD x WTI - (DT + IN) x WT [wI/(BCF x FM x L)] + FC For carcinogens: RTC (mg/kg) = .-_ RLxWT ql l wI/‘(BCF x FM x L) + FC]
The above equations should be corrected for site-specific lipid content and bioaccumulation factors where data are available. Again, some States have established their own procedures whereby RTCscan be developed based upon extrapolation of acute and/or chronic animal data to safe concentrations protective of fish consumption by humans. Where additional risk information is needed, an RTC could be based upon other information such as drinking water MCLs or FDA action levels.



## 2.5 BIOLOGICAL CRITERIA

As discussed in Chapter 1, to fully protect aquatic habitats and provide more comprehensive assessments of aquatic life use attainment/nonattainment, States are to fully integrate chemicalspecific techniques, toxicity testing, biological surveys, and biocriteria into their water quality programs. In particular, the Agency’s policy is that States should develop and implement biological criteria in their water quality standards (see Chapter 1, Reference 55).



### 2.5.1 Reguiatmy 

### Basesfur Bimriteda The primary statutory basis for EPA’s policy that States should develop biocriteria isfound in Sections 101 (a) and 303(c)(2)(B) of the Water Quality Act of 1987. Section 101 (a) of the CWA gives the general authority for biological criteria. It establishes as the objective of the Act the restoration and maintenance of the chemical, physical, and biological integrity of the Nation’s waters. To meet this objective, water quality criteria should address biological integrity. Section 101 (a) includes the interim water quality goal for the protection and propagation of fish, shellfish, and wildlife.
Section 304 of the Act provides the legal basis for the development of informational criteria, including biological criteria. Specific directives for the development of regulatory biocriteria can be found in Section 303, which requires EPA to develop criteria based on biological assessment methods when numerical criteria are not established.
Once biocriteria formally are adopted into State standards, biocriteria and aquatic life use designations serve as direct, legal endpoints for determining a 

### quality life use attainment/ nonattainment. As stated in Section 131 .l 1(b)(2) of the Water Quality Standards Regulation (40 CFRPart 131), biocriteria should be used as a supplement to existing chemical-specific criteria and as criteria where such chemical-specific criteria have not been established. States are encouraged to implement and integrate all three approaches (biosurvey, chemical-specific, and toxicity testing methods) into their water quality programs, applying them in combination or independently (providing the most protective of the three methods is used) as site-specific conditions and assessment objectives dictate.
Section 304(a) directs EPA to develop and publish water quality criteria and information on methods for measuring water quality and establishing water quality criteria for toxic pollutants on bases other than pollutant-by-pollutant, including biological monitoring and assessment methods that assess:

l The effects of pollutants on aquatic community components (‘I.

. plankton, fish, shellfish, wildlife, plant life

. . .“)

41and community attributes (I’.

. . biological community diversity, productivity, and stability

. .“); in any body of water.

l Factors necessary ” . . to restore and maintain the chemical, physical, and biological integrity of all navigable waters

. ” for ”

. . the protection of shellfish, fish, and wildlife for classes and categories of receiving waters

. . . .I’



### 2.5.2 Dmelopntentandih@mwnMon ofBioctHerii.3 Biocriteria are numerical values or narrative expressions that describe the reference biological integrity of aquatic communities inhabiting unimpaired waters of a designated aquatic life use. The biological communities in these waters represent the best attainable conditions. The reference site conditions then become the basis for developing biocriteria for major surface water types (streams, rivers, lakes, wetlands, estuaries, or marine waters). Biologicat criteria support designated aquatic life use classifications for application in State standards. Each State develops its own designated use classification system based on the generic uses cited in the Act (e.g., protection and propagation of fish, shellfish, and wildlife). Designated uses are intentionally general. However, States may develop subcategories within use designations to refine and clarify the use class. Clarification of the use class is particularly helpful when a variety of surface waters with distinct characteristics fit within the same use class, or do not fit well into any category.
For example, subcategories of aquatic life uses may be on the basis of attainable habitat (e.g., cold versus warmwater communities dominates by bass versus catiish). Special uses also may be designated to protect particularly unique, sensitive, or valuable aquatic species, communities, or habitats.
Resident biota integrate multiple impacts over time and can detect impairment from known and unknown causes. Biocriteria can be used to verify improvement in water quality in response to regulatory efforts and detect continuing degradation of waters. They provide a framework for developing improved best management practices for nonpoint source impacts. Numeric criteria can provide effective monitoring criteria for inclusion in permits.
The assessment of the biological integrity should include measures of the structure and function of an aquatic community of species within a specified habitat. Expert knowledge of the system is required for the selection of appropriate biological components and measurement indices. The development and implementation of biological criteria requires:

l Selecting unimpaired (minimal impact) 

### surface waters to use as the reference condition for each designated use

l Measuring the structure and function of aquatic communities in reference surface waters to establish biological criteria

l Establishing a protocol to compare the biological criteria to biota in impacted waters to determine whether impairment has occurred. These elements serve as an interactive network that is particularly important during early development of biological criteria where rapid accumulation of information is effective for refining both designated uses and developing biological criteria values.



## 2.6 SEDIMENT CRITERIA

2.6.1 Cosrontilevel- in 6tWnent Criteria While ambient water quality criteria are playing an important role in assuring a healthy aquatic environment, they alone have not been sufficient to ensure appropriate levels of environmental protection. Sediment contamination, which can involve deposition of toxicants over long periods of time, is responsible for water quality impacts in some areas.
EPA has authority to pursue the development of sediment criteria in streams, lakes, and other waters of the United States under CWA Sections 104 and 304(a)(l) and (2) as follows:

l Section 104(n)(l) authorizes the Administrator to establish national programs that study the effects of pollution, including sedimentation, in estuaries on aquatic life.

l Section 304(a)(l) directs the Administrator to develop and publish criteria for water quality, including information on the factors affecting rates of organic and inorganic sedimentation for varying types of receiving waters.

l Section 304(a)(2) directs the Administrator to develop and publish information on, among other things, “the factors necessary for the protection and propagation of shellfish, fish, and wildlife for classes and categories of receiving waters...“
To the extent that sediment criteria could be developed that address the concerns of the Section 404(b)(l) guidelines for discharges of dredged or fill material under the CWA or the Marine Protection Research, and Sanctuaries Act, they also could be incorporated into those regulations.



### 2.6.2 &woact~ to ssdhnent Wteria Deve&wnent Over the past several years, sediment criteria development activities have centered on evaluating and developing the equilibrium partitioning approach for generating sediment criteria. The equilibrium partitioning approach focuses on predicting the chemical interaction between sediments and contaminants. Developing an understanding of the principal factors that influence the sediment/contaminant interactions will allow for predictions to be made as to what concentration of a contaminant benthic and other organisms may be exposed to. Chronic water quality criteria, or possibly other toxicological endpoints can then be used to predict potential biological effects. In addition to the development of sediment criteria, EPAalso is working to develop a standardized sediment toxicity test that could be used with or independently of sediment criteria and could be used to assess chronic effects in freshwater and marine water.

42Equilibrium partitioning (EqP) sediment quality criteria 

### (SQC) are the EPA’s best recommendation of the concentration of a substance in sediment that will not unacceptably affect benthic organisms or their uses. Methodologies for deriving effects based SQC vary for different classes of compounds. For non-ionic organic chemicals the methodology requires normalization to organic carbon, A methodology for deriving effects based sediment criteria for metal contaminants is under development and is expected to require normalization to acid volatile sulfide. EqP SQC values can be derived for varying degrees of uncertainty and levels of protection thus permitting use for ecosystem protection and remedial programs.



### 2.6.3 

### Applhtion of SedhnentCriteria SQC would provide a basis for making more informed decisions on the environmental impacts of contaminated sediments. Existing sediment assessment methodologies are limited in their ability to identify chemicals of concern, responsible parties, degree of contamination, and zones of impact. EPAbelieves that a comprehensive approach using SQC and biological test methods is preferred in order to make the most informed decisions.
Sediment criteria will be particularly valuable in site monitoring applications where sediment contaminant concentrations are gradually approaching a criteria over time. Sediment criteria also are valuable as a preventative tool to ensure that point and nonpoint sources of contamination are controlled to ensure uncontaminated. sediments remain uncontaminated. Also, comparison of field measurements to sediment criteria will be a reliable method for providing early warning of a potential problem. An early warning would provide an opportunity to take corrective action before adverse impacts occur. For the reasons mentioned above it has been identified that SQC are essential to resolving key contaminated sediment and source control issues in the Great Lakes.


### Specific 

### Appkatims Specific applications of sediment criteria are under development. The primary 

### use of EqP-based sediment criteria will be to assess risks associated with contaminants in sediments. The various offices and programs concerned with contaminated sediment have different regulatory mandates and thus, have different needs and areas for potential application of sediment criteria. Because each regulatory need is different, EqP-based sediment quality criteria designed specifically to meet the needs of one office or program may have to be implemented in different ways to meet the needs of another office or program.
One mode of application of EqP-based numerical SQC would be in a tiered approach. In such an application, when contaminants in sediments exceed the SQC, the sediments would be considered as causing unacceptable impacts. Further testing may or may not be required depending on site-specific conditions and the degree in which a criteria has been violated. (No additional testing would be required in locations where contamination significantly exceeds a criterion. Where sediment contaminant levels are close to a criteria, additional testing may be necessary.) Contaminants in a sediment at concentrations less than the sediment criteria would not be of concern. However, in some cases the sediment could not be considered safe because they may contain other contaminants above safe levels for which no sediment criteria exist. In addition, the synergistic, antagonistic, or additive effects of several contaminants in the sediments may be of concern.
Additional testing in other tiers of an evaluation approach, such as bioassays, could be required to determine if the sediment is safe. It is likely that such testing would incorporate site-specific considerations. Examples of specific applications of sediment criteria after they are developed are as follows:
Establish permit limits to ensure that uncontaminated sediments remain uncontaminated or sediments already contaminated have an opportunity to cleanse themselves. This would occur only after criteria and the means to tie point sources to sediment deposition are developed.
Establish target levels for nonpoint source causes of sediment contamination.
For remediation activities, SQC would be valuable in identifying:

- Remediation need

- Spatial extent of remediation area

- Benefits derived from remediation activities

- Responsible parties

- Impacts of depositing contaminated sediments in

water environments

- Success of remediation activities.
In tiered testing sediment evaluation processes, sediment criteria and biological testing procedures work very well together.



### 2.6.4 &?dinwnt &eria status 

### ScieuceAdvisoryBoiwdReview The Science Advisory Board has completed its review and issued a favorable report on the EqP for assessing sediment quality. The Subcommittee found the EqP “to have major strengths in its foundation in chemical theory, its ease of calculation, and its ability to make use of existing data... The conceptual basis of the approach is supported by the Subcommittee; however, its application at this time is limited.”
The Science Advisory Board also identified the need for “a better understanding of the uncertainty around the assumptions inherent in the approach, including assumptions of equilibrium, bioavailability, and kinetics, all critical to the application of the EqP.” An uncertainty analysis and a guidance document to assist in the regulatory application of developed criteria are under development and expected to be completed in 1991.

43

### Metbodologr for 

### Develophfi Sedfment Critaria for Metal co8tamin8nt8 EPA efforts at producing sediment criteria documents are being directed first toward phenanthrene, fluoranthene, DDT, dieldrin, acenaphthene and endrin. Efforts also are being directed to produce a guidance document, Application of Sediment Quality Criterio for the Protection of Aquatic Life, scheduled for release in 1991. EPA is proceeding with a methodology for developing sediment criteria for metal contaminants, with key work focused on identifying and understanding the role of acid volatile sulfides (AVS) in controlling the bioavailability of metal contaminants. A variety of field and laboratory verification studies are underway to add additional support to the methodology. Standard AVS sampling and analytical procedures are under development [l 81. Presentation of the metals methodology to the Science Advisory Board for review is scheduled for 1991.

44

## CHAPTER 2 REFERENCES

1.

2.

3.

4.

5

6.

7.

8.

9. 10. U.S.EPA 1984. WaterQuulkyStundards Handbook. Officeof Water Regulations and Standards (WH-585), Washington, DC. U.S. EPA. 1984. Technical Support Document for Conduciting

Use Attoinobility Studies. Office of Water Regulations and Standards (WH-SSS), Washington, DC. National Academy of Science. 1973. Water Quality Criteria

7 972. EPA-R3-73-033 or NTIS PB236199. Mancini, J.L. 1983. A Method for Calculating Effects on

Aquatic Organisms of Time-Varying Concentrations. Wuter Res. 17:1355-61. Erickson, R., C. Kleiner, J. Fiandt, and T. Highland. 1989.

Report on the feasibility of Predicting the Effectsof FluctuoGng Concentrations on Aquatic Organisms. U.S. EPA, ERL-Duluth. Brungs, W.A. 1986. Allocated Impact Zones for Areas of NonCompliance. U.S. EPA, Region 1. Water Management Division, Boston, MA. U.S. EPA. 1985. Guidelines for Deriving Numerical National

Water Quality Criteria for the Protection ofAquatic Organisms and Their Uses. NTIS PB85227049. U.S. EPA. 1987. integrated Risk lnformotion System. Volume

2, Chemical Files. Office of Health and Environmental Assessment. EPA/600/8-86/032b. March 1987b. Javitz, H.S. 1980. Letter to H. Kahn (EPA). SRI International. Stephan, C.E. 1980. Per Capita Consumption of NonMarine Fish and Shellfish. Memorandum to J. Stara. U.S. EPA, ERL-Duluth. 1 1. U.S. Department of Agriculture. 1984. Agricultural Stotirtics. U.S. DA, Washington, DC. 12. Finch, R. 1973. The MECCA Project: Effects of Regulatory Guidelines on the Intake of Mercury from Fish. Fisheries Bulletin 71:615-26. 13. U.S. EPA. 1989. Exposure 

### fuctors Handbook. OHEA, Washington, DC. EPA/600/8-89/043. 14. Thomann, R.V. 1989. Bioaccumulation Model of Organic Chemical Distribution in Aquatic Food Chains. Environ. Sci. Technol. 231699-707. 15. Thomann, R.V. 1987. A Statistical Model of Environmental Contaminants Using Variance Spectrum Analysis. Report to National Science Foundation. NTIS PB88-235130/A09. 16. U.S. EPA. 1987. /n&grated Risk lnformotion System. Volume I, Supportive Documentation. Office of Health and Environmental Assessment. EPA/600/8-86/032a. March 1987a. 17. U.S. EPA. 1988. Guidance Manual forAssessing Human Health issues from Chemically Contominoted Fish ond Shellfish. Submitted by Robert A. Pastorok, PTI Environmental Services, Bellevue, WA; for Battelle New England Marine Research Laboratory, Duxbury, MA. 18. University of Delaware Department of Civil Engineering. 1990. Development of an Analytical Method of the Determination of Acid Volatile Sulfide in Sediment. Submitted by Battelle to U.S. EPA,Criteria and Standards Division, Washington, DC.

45

# 3. EFFLUENT CHARACTERIZATION

3.1 INTRODUCTION

Once the applicable designated uses and water quality criteria for a waterbody are determined, the effluent must be characterized and the permitting authority must determine the need for permit limits to control the discharge. The purpose of effluent characterization is to determine whether the discharge causes, has the reasonable potential to cause, or contributes to an excursion of numeric or narrative water quality criteria. Once the permitting authority determines that a discharge causes, has the reasonable potential to cause, or contributes to the excursion of water quality criteria, the permitting authority must develop permit limits that will control the discharge. At a minimum, the permitting authority must make this determination at each permit reissuance. The effluent characterization procedures described in the following sections apply only to the water quality-based approach, not to end-of-the-pipe technology-based controls. Although many waterbodies receive discharges from only single point sources, permitting authorities will also occasionally encounter receiving waters where several dischargers are in close proximity. In such situations, the permitting authority may find that each discharger alone does not cause, have the reasonable potential to cause, or contribute to an excursion above water quality criteria. Yet, the dischargers may collectively 

### cause, have the reasonable potential to cause, or contribute to an excursion. Under these circumstances, limits must be developed for each discharger to protect against collective excursions of applicable water quality standards consistent with the Environmental Protection Agency’s (EPA) existing regulations in 40 CFR 122.44(d)(1)(ii) for controlling multiple discharges. The terms “cause,” "reasonable potential to cause,” and “contribute to” are the terms used in the National Pollutant Discharge Elimination System (NPDES) regulations for conditions under which water quality-based limits are required. Permitting authorities are required to consider each of these concepts when performing effluent characterizations.
This chapter is divided into two parts: Section 3.2, Determining the Need for Permit Limits Without Effluent Data, and Section 3.3, Determining the Need for Permit Limits With Effluent Data. Section 3.3 includes effluent characterization for whole effluent toxicity and for specific chemicals (including those for human health protection) and is based on the cumulative experience gained by EPA, States, publicly owned treatment works (POTWs), and industry when implementing the water quality-based approach to toxics control. The effluent bioconcentration evaluation procedures described in the section on human health are currently draft and are subject to further validation before being used. Until the procedures are fully developed, reviewed, and finalized, permitting authorities should not use them to characterize effluents. 

### 3.1.1 NPDES Regulation Requirements Effluent characterization is an essential step in determining the need for an NPDES permit limit. NPDES regulations under 40 CFR 122.44(d)(1) specify the minimum requirements and general types of analyses necessary for establishing permit limits. Each of these regulations is described below.
40 CFR 122.44(d)(1)(ii)
When determining whether a discharge causes, has the reasonable potential to cause, or contributes to an instream excursion above a narrative or numeric criteria within a State water quality standard, the permitting authority shall use procedures which account for existing controls on point and nonpoint sources of pollution, the variability of the pollutant or pollutant parameter in the effluent, the sensitivity of the species to toxicity testing (when evaluating whole effluent toxicity), and where appropriate, the dilution of the effluent in the receiving water.
This regulation requires at a minimum the consideration of each of these elements in determining the need for a limit.
40 CFR 122.44(d)(1)(iii)
When the permitting authority determines, using the procedures in paragraph (d)(1)(ii) of this section, that a discharge causes, has the reasonable potential to cause, or contributes to an in-stream excursion above the allowable ambient concentration of a State numeric criteria within a State water quality standard for an individual pollutant, the permit must contain effluent limits for that pollutant.
Under this regulation, permitting authorities need to investigate for the existence of pollutants in effluents if there is a numeric water quality criterion for that pollutant and to implement limits for those pollutants where necessary.
40 CFR 12244(d)(1)(iv)
When the permitting authority determines, using 

### the procedures in paragraph (d)(1)(ii) of this section, that a discharge causes, has the reasonable potential to cause, or contributes to an in-stream excursion above the numeric criterion for whole effluent toxicity, the permit must contain effluent limits for whole effluent toxicity.
Under this regulation, permitting authorities need to investigate for the existence of whole effluent toxicity in effluents if there is a numeric water quality criterion for that parameter and to implement whole effluent toxicity limits where necessary.



### 47

### 40 CFR 122.44(d)(1)(v) Exceptas provided in this subparagraph,when the permitting authority determines, usingthe procedures in paragraph(d)(1)(ii) of this section,toxicity testing data,orother information,that adischarge causes, has the reasonable potentialto cause, or contributesto an in-streamexcursionabovea narrativecriterionwithin anapplicableState waterquality standard,the permit mustcontain effluentlimitsforwhole effluenttoxicity. Limits onwholeeffluenttoxicityarenot necessary where the permitting authoritydemonstrates in thefactsheet or statementof basisof the NPDES permit, usingthe proceduresin paragraph(d)(1)(ii) of this section,that chemical-specific limitsforthe effluentaresufficientto attain and maintain applicablenumericand narrative State waterqualitystandards. Underthis regulation,permitting authoritiesneedto investigate for the existenceof whole effluent toxicity in effluents. If the permitting authority can demonstratethat control of specific chemicals issufficient to controltoxicityto thepoint ofachieving compliance with thewaterqualitycriteria,thenchemical-specific permit limitsalonewill besufficientto complywith the regulation. 40 CFR 122.44(d)(1)(vi) WhereaState hasnot established awaterqualitycriterion for a specificchemicalpollutant that ispresentin an effluent at a concentration that causes,hasthe reasonable potentialto cause, or contributesto an excursionabovea narrative criterionwithin anapplicable Statewaterquality standard,the permitting authority mustestablish effluentlimitsusingoneor moreofthe following [three]options:

. . . . Underthis regulation,permittingauthoritiesneedto investigate forthe existenceof specific chemicals in effluents forwhich the Statehasnot adopted numericcriteria,but which maybecontributing to aquatic toxicity or impairment of human health. Narrativecriteriaapplywhen numericcriteriado not protectall the designatedor existing uses. For example, the narrative criterianeedto be usedto protect human health if a Statehas only adoptedanumericcriteriafor protectingaquaticlife. Conversely, the narrativecriterianeedto be usedto protectaquatic life if a Statehasonly adopteda numericcriteriafor protecting human health. Oncethe permitting authority determines that oneor morespecific chemicals inaneffluentmustbecontrolled, theauthoritiescan useEPA’s nationalcriteria,developtheirown criteria, or control the pollutant through useof an indicator pollutant, asprovided in subparagraph (d)(1)(vi). In any case, thepermittingauthoritywill needto characterize theeffluentina mannerconsistent with theselected approach forcontrollingthe pollutant.

3.1.2 Background fur ToxicEffectsAssessments onAquatic life andHumanHealth Aquatic toxicity effectscan be characterized by conducting a general assessment of the effluent, or by measuringeffluent toxicity orconcentrations of individualchemicals andcomparing thesemeasurements to the expectedexposureconcentrationsin the receivingwater. The“receivingwaterconcentration” (RWC) isthe measured orprojected exposure concentrationofatoxicant or the parameter toxicity (when dealingwith thewhole effluent toxicity) in the receivingwateraftermixing. The RWC iscalculatedatthe edgeof a mixing zoneif suchazoneisallowedbya State’s waterqualitystandards. Aswith aquaticlifeprotection,therearetwo possible approaches to characterizing effluents for humanhealtheffects:chemical-bychemicaland whole effluent. However,only the chemical-bychemical approachcurrentlyispractical forassessing andcontrolling human health impacts. Appendix Gdiscusses developing procedures forassessing humanhealthimpacts fromwholeeffluents. A fundamentalprinciple in the developmentof water qualitybased controlsisthat the RWC mustbeless than the criteriathat comprise or characterize thewaterquality standards.With individual toxicants (or the parametertoxicity), the potential for toxicity in the receiving waterisminimizedwherethe RWC isless than the criterioncontinuousconcentration(CCC),the criterion maximumconcentration(CMC),andthe reference ambientconcentration(RAC).Toxicity becomes maximizedwherethe RWC exceeds these criteria.Therefore,to prevent impactsto aquatic life or human health, the RWCof the parameter effluent toxicity or an individual toxicant (basedon allowable dilution for the criterion) must be lessthan the most limiting of the applicable criterion, as indicated below. (The RACas used throughout this chapterincorporatesEPA human health criteria andState standards aswell.) RWC < CCC (chronic aquatic life) RWC <CMC (acuteaquatic life) RWC < RAC(human health) Thewaterqualityanalystwill usethe samebasiccomponentsin theabove-described relationship (i.e.,criticalreceiving waterflows, ambient criteria values,measures of effluent quality) for both effluentcharacterization andwasteload allocation(WLA)development, albeitfrom differentperspectives.In the caseof effluent characterization, the objectiveisto projectreceivingwaterconcentrationsbasedupon existing effluent quality to determine whetheror notanexcursionabove ambientcriteriaoccurs, orhas the reasonable potential to occur. In developingWLAs, on the other hand,theobjectiveisto fix the RWC at thedesiredcriteria leveland determinean allowableeffluent loading that will not cause excursions abovethe criteria. Recommendations for projectingthe RWC are described within this chapter. Chapter4, Exposure Assessment and Wasteload Allocation,providesrecommendations for determiningallowable effluent loadingsto achieveestablished ambientcriteriaand for calculatingWLAs for establishingpermit limits. Theprocedures describedwithin Chapter4 can also be usedto calculatethe dilution foranalyses within Chapter3. Chapter5, PermitRequirements,describes the actualcalculationof permitlimitsaftereffluent characterization and loadings,aswellasWLAS, arecomplete.

48