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 of Water 

Environmental Protection Washington, DC 

Agency 

TECHNICAL SUPPORT DOCUMENT FOR 

WATER QUALITY-BASED TOXlCS CONTROL 

This copy represents 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 criteria for metals: The Agency has issued “Interim Guidance Interpretation 

and Implementation Aquatic Life Criteria for Metals.” The interim guidance supersedes criteria document 

statements expressing criteria in terms of a acid soluble analytical method and also the metals 

discusses of Section 5.7.3. The availability of this document appeared in the 

June 5, 1992 Federal Register (Vol. 57, No. 109, pg. 24407). 

March 1991 

Office of Water Enforcement and Permits 

Office of Water Regulations and Standards 

U.S. Environmental Protection Agency 

Washington, DC 20460 
FOREWORD 

The U.S. Environmental Protection Agency (EPA) and the State pollution control agencies have been charged 

with enforcing the laws regarding pollution of the natural environment. Environmental pollution is an urgent and 

continuing problem and, consequently, the laws grant considerable discretion to the control authorities to define 

environmental goals and develop the means to attain them. Establishing environmentally protective levels and 

incorporating them in a decisionmaking process entails a considerable amount of scientific knowledge and 

judgment. One area where scientific knowledge is rapidly changing concerns the discharge of toxic pollutants to 

the Nation’s surface waters. 

This document provides technical guidance for assessing and regulating the discharge of toxic substances to the 

waters of the United States. It was issued in support of EPA regulations and policy initiatives involving the 

application of biological and chemical assessment techniques to control toxic pollution to surface waters. This 

document is agency guidance only. It does not establish or affect legal rights or obligations. It does not establish 

a binding norm and is not finally determinative of the issues addressed. Agency decisions in any particular case 

will be made applying the law and regulations on the basis of specific facts when permits are issued or regulations 

promulgated. 

This document is expected to be revised periodically to reflect advances in this rapidly evolving area. Comments 

from users will be welcomed. Send comments to U.S. EPA, Office of Water Enforcement and Permits, 401 M 

Street, SW, Mailcode EN366, Washington, DC 20460. 

James R. Elder, Director 

Office of Water Enforcement and Permits 

Martha G. Prothro, Director 

Office of Water Regulations and Standards 

iii 
TABLE OF CONTENTS 

Section Page 

Foreword ................................. iii 

Acknowledgment ....................................................................... xiii 

Executive Summary ..................................................................... xiv 

List of Abbreviations. .................................................................... xvii 

Glossary .......................................................................... xix 

Introduction .......................................................................... xxiii 

1. APPROACHES TO WATER QUALITY-BASED TOXICS CONTROL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 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 

Receiving Water Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 

Chemical-Specific Analytical Method Precision . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 

1.3 WHOLE EFFLUENT APPROACH FOR AQUATIC 

LIFE PROTECTION ........................................................... 4 

1.3.1 

1.3.2 

1.3.3 

1.3.4 

Toxic units ............................................................ 6 

Correlation of Whole Effluent Toxicity Measurements to 

Actual Receiving Water Impact . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 

Toxicity Test Method Precision. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 

Considerations Involved When Implementing the Whole 

Effluent Toxicity Approach. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 

1.4 BIOLOGICAL CRITERIA/BIOASSESSMENT AND BIOSURVEY 

APPROACH FOR AQUATIC LIFE PROTECTION. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18 

1.4.1 

1.4.2 

Use of Biosurveys and Bioassessments in Water Qualitybased Toxics Control. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18 

Conducting Biosurveys . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ... 19 

1.5 INTEGRATION OF THE WHOLE EFFLUENT, CHEMICALSPECIFIC, AND BIOASSESSMENT APPROACHES .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 

1.5.1 Capabilities and Limitations of the Chemical-Specific Approach ..................... 20 

1.5.2 Capabilities and Limitations of the Whole Effluent Approach ....................... 21 

1.5.3 Capabilities and Limitations of the Bioassessment Approach ....................... 22 

V 
Section 

Table of Contents (Continued) 

Page 

1.6 OTHER FACTORS INFLUENCING WATER 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 HEALTH PROTECTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24 

1.7.1 Types of Health Effects ................................................................................................................ 25 

REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...26 

2. WATER QUALITY CRlTERlA AND STANDARDS. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29 

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

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

2.1.2 Water Quality Standards and State Toxics 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 QUALITY CRITERIA FOR AQUATIC LIFE PROTECTION. . . . . . . . . . . . . . . . . . . . . . . . . . . 34 

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

2.3.2 Magnitude for Single Chemicals ........................................... 34 

2.3.3 Magnitude for Whole Effluent Toxicity ....................................... 35 

2.3.4 Duration for Single Chemicals and Whole Effluent Toxicity ........................ 35 

2.3.5 Frequency for Single Chemicals and Whole Effluent Toxicity ....................... 36 

2.4 WATER QUALITY CRITERIA FOR HUMAN HEALTH PROTECTION . . . . . . . . . . . . . . . . . . . . . . . . 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 Exposure Considerations ........................................... 37 

Fish Consumption 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 

of lRlS Values ....................................................... ..40 

Deriving Reference Tissue Concentrations for Monitoring 

Fish Tissue ......................................................... ..41 

vi 
Table of Contents (Continued) 

Section Page 

2.5. BIOLOGICAL CRITERIA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41 

2.5.1 Regulatory Bases for Bio criteria. ............................................ 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 

NPDES Regulation Requirements ........................................... 47 

Background for Toxic Effects Assessments on Aquatic Life 

and Human Health ..................................................... 48 

General Considerations in Effluent Characterization. ............................. 49 

3.2 DETERMINING THE NEED FOR PERMIT LIMITS WITHOUT 

EFFLUENT MONITORING DATA FOR A SPECIFIC FACILITY. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50 

3.3 DETERMINING THE NEED FOR PERMIT LIMITS WITH 

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 Effluent Toxicity ............................. 53 

Use of Toxicity Testing in Multiple-source Discharge 

Situations.. .......................................................... 59 

Ambient Toxicity Testing ................................................. 61 

Special Considerations for Discharges to Marine and 

Estuarine Environments .................................................. 61 

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

Effluent Characterization for Specific Chemicals ................................ 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 TOTAL MAXIMUM DAILY LOADS AND WASTELOAD 

ALLOCATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67 

4.2.1 

4.2.2 

Total Maximum Daily Loads .............................................. 67 

Wasteload Allocation 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 Size of Mixing Zones ....................................... 71 

Prevention of Lethality to Passing Organisms .................................. 71 

Prevention of Bioaccumulation Problems for Human Health ....................... 72 

4.4 MIXING ZONE ANALYSES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72 

4.4.1 

4.4.2 

4.4.3 

4.4.4 

4.4.5 

General Recommendations for Outfall Design, ................................. 73 

Critical Design Periods for Waterbodies ...................................... 73 

General Recommendations for Tracer Studies. ................................. 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 

Specific Model Recommendations .......................................... 83 

Effluent Toxicity Modeling. ............................................... 85 

4.6 HUMAN HEALTH ........................................................... 87 

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

4.6.2 Determining the TMDL Based on 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 BASIC PRINCIPLES OF EFFLUENT VARIABILITY. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93 

5.2.1 Variations in Effluent Quality .............................................. 93 

5.2.2 Statistical Parameters and Relationship to Permit Limits ........................... 95 

5.2.3 Expression of Permit Limits ............................................... 96 

5.3 ENSURING CONSISTENCY WITH THE WASTELOAD 

ALLOCATlON . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96 

5.3.1 

5.3.2 

Statistical Considerations of WLAs .......................................... 96 

Types of Water Quality Models and Model Outputs ............................. 96 

5.4 PERMIT LIMIT DERIVATION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98 

5.4.1 

5.4.2 

5.4.3 

5.4.4 

EPA Recommendations for Permitting for Aquatic Life 

Protection ............................................................ 98 

Other Approaches to Permitting for Aquatic Life . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103 

Special Permitting Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 104 

EPA Recommendations for Permitting for Human Health 

Protection ........................................................... 104 

5.5 SPECIAL CONSIDERATIONS IN USE OF STATISTICAL PERMIT 

LIMIT DERIVATION TECHNIQUES. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 105 

5.5.1 Effect of Changes of Statistical Parameters on 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 PERMlT DOCUMENTATION. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110 

5.7 EXPRESSING LIMITS AND 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-based Effluent Limits ............................................... 110 

Energy Conservation ................................................... 111 

Considerations in the Use of Chemical specific Limits ........................... 111 

Considerations in the Use of Whole Effluent Toxicity Limits ....................... 112 

Selection of Monitoring Frequencies. ....................................... 113 

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

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

5.8 TOXICITY REDUCTION EVALUATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114 

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

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

5.8.3 Circumstances Warranting a TRE .......................................... 117 

5.8.4 Mechanisms for Receiving TREs ........................................... 118 

REFERENCES ................................................................... 121 

ix 
Table of Contents (Continued) 

Section Page 

6. COMPLIANCE MONITORING AND ENFORCEMENT .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123 

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

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

6.3 COMPLIANCE MONITORING ................................................ 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 OF VIOLATIONS ................................................. 126 

REFERENCES ................................................................... 127 

7. CASEEXAMPLES .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...129 

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

7.2 CASE 1: 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 Site Description and Information .................................... 129 

Effluent Characterization for Specific Chemicals ............................... 129 

Effluent Characterization for Whole Effluent Toxicity ............................ 131 

Determine Wasteload Allocations. ......................................... 132 

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

Determining and Expressing the Controlling Effluent limit ....................... 133 

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

7.3 CASE 2: POTW DISCHARGE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 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 Specific Chemicals ............................... 134 

Effluent Characterization for Whole Effluent Toxicity ............................ 136 

Determine Wasteload Allocations. ......................................... 136 

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

Determining and Expressing the Controlling Effluent Limits. ...................... 137 

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

7.4 CASE 3: MULTIPLE DISCHARGERS INTO THE SAME REACH . . . . . . . . . . . . . . . . . . . . . . . . . 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 Precision Data ................................................ A-1 

Effluent Variability Data .................................................. A-2-1 

Acute to Chronic Ratio Data .............................................. A-3-1 

Summary of Clean Water Act Provisions ...................................... B-1-1 

Policies for Toxics Control ................................................ B-2-1 

Regulations for Toxics Control ............................................. B-3-1 

Whole-Effluent Toxicity Permitting Principles and Enforcement 

Strategy ............................................................. B-4-1 

Quality Control Fact Sheets ............................................... B-5-1 

Case Decisions on Whole-Effluent Toxicity .................................... B-6-1 

Ambient Toxicity Testing and Data Analysis. .................................... 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 

Reference Dose (RfD): Description and Use in Health Risk 

Assessments .............................. H-1 

Chemicals Available in IRIS. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . l-1 

xi 
ACKNOWLEDGEMENT Permit Requirements 

The preparation of the revised Technical Support Document for Water 

Quality-based Toxic Control began with a 3-day conference held in 

Williamsburg, Virginia, in December 1988. Representatives of 

EPA Headquarters and Regions, States, private industry, municipalities, academia, and various interest groups attended this 

meeting and provided valuable input. The principal authors of 

this document were Bill Swietlik, James Taft, Jacqueline Romney, 

Kathryn Smith, John Cannell, Robert Wood, James Pendergast, 

and Rick Brandes of the Permits Division; Sheila Frace and Margarete 

Heber of the Enforcement Division, Elizabeth Southerland and 

Richard Healy of the Assessment and Watershed Protection Division; and Charles Deb, Warren Banks, and Robert April of the 

Criteria and Standards Division. Listed below are the contributors 

to specific chapters of this document. 

Approaches to Water Quality-based Toxics Control 

Margaret Heber, U.S. EPA, Enforcement Division 

Kathryn Smith, U.S. EPA, Permits Division 

Water Quality Criteria and Standards 

Charles Delos, U.S. EPA, Criteria and Standards Division 

Warren Banks, U.S. EPA, Criteria and Standards Division 

Kathy Barylski, U.S. EPA, Criteria and Standards Division 

Robert April, U.S. EPA, Criteria and Standards Division 

David Moon, U.S. EPA, Criteria and Standards Division 

Jacqueline Romney, U.S. EPA, Permits Division 

Appendices 

Effluent Characterization 

Robert Wood, U.S. EPA, Permits Division 

Bill Swietlik, U.S. EPA, Permits Division 

James Pendergast, U.S. EPA, Permits Division 

Exposure and Wasteload Allocation 

Elizabeth Southerland, U.S. EPA, Assessment and Watershed 

Protection 

Richard Healy, U.S. EPA, Assessment and Watershed Protection 

Bill Swietlik, U.S. EPA, Permits Division 

James Pendergast, U.S. EPA, Permits Division 

John Cannell, U.S. EPA, Permits Division 

Compliance Monitoring and Enforcement 

Sheila Frace, U.S. EPA, Enforcement Division 

Theodore Coopwood, U.S. EPA, Enforcement Division 

Human Health Component of All Chapters 

John Cannell, U.S. EPA, Permits Division 

Katherine Dowell, U.S. EPA, Permits Division 

William Morrow, U.S. EPA, Permits Division 

Case Examples Workgroup 

Bill Swietlik, U.S. EPA, Permits Division 

James Pendergast, U.S. EPA, Permits Division 

Charles Delos, U.S. EPA, Criteria and Standards Division 

Jacqueline Romney, U.S. EPA, Permits Division 

Appendix A: Margaret Heber, U.S. EPA, Enforcement Division 

Appendix B: U.S. EPA, Permits Division 

Appendix C: U.S. EPA, Permits Division 

Appendix D: Nelson Thomas, U.S. EPA, ERL/ORD, Duluth, MN 

Appendix E: Henry Kahn and Marla Smith, U.S. EPA, Analysis 

and Evaluation Division 

Appendix F: U.S. EPA, Permits Division 

Appendix C: U.S. EPA, Permits Division 

Appendix H: U.S. EPA, RfD Workgroup 

xiii 
EXECUTIVE SUMMARY 

The revised Technical Support Document for Water Quality-based 

Toxics Control (TSD) provides States and Regions with guidance 

on procedures for use in the water quality-based control of toxic 

pollutants. It presents recommendations to regulatory authorities 

faced with the task of controlling the point source discharge of 

toxic pollutants to the Nation’s waters. The document provides 

guidance for each step in the water quality-based toxics control 

process from standards development to compliance monitoring. 

Both human health and aquatic toxicity issues are incorporated 

into the discussions throughout the document. The overall approach in this revised document provides additional explanations 

and rationales based on accumulated experience and data for the 

various recommendations that were made in the original TSD. 

The following is a brief synopsis of the guidance provided in the 

TSD. 

Approaches to Water Quality-based Toxics Control 

The Environmental Protection Agency’s (EPA) surface toxics control regulation, 54 FR 23868, June 2, 1989, established specific 

requirements that the “integrated” approach be used in water 

quality-based toxics control. The “integrated” approach consists 

of whole effluent and chemical-specific approaches as a means of 

protecting aquatic life and human health. As techniques are 

made available for implementing biocriteria, they too should be 

integrated into the water quality-based toxics control, thus creating a triad of approaches: whole effluent, chemical-specific, and 

biological assessments. Each approach has its limitations and 

thus, exclusive use of one approach alone cannot ensure required 

protection of aquatic life and human health. The advantages/ 

disadvantages of each approach and how the integrated approach creates an effective toxics control program are discussed 

in the text. 

The whole effluent approach to toxics control involves the use of 

toxicity tests and water quality criteria for the parameter “toxicity” to assess and control the aggregate toxicity of effluents. New 

references and information in support of the whole effluent toxicity assessment and control approach have been included in Chap 

ter 1 and associated appendices (e.g., precision data, justifications 

for acute-to-chronic ratio recommendations, information on analytical variability in toxicity testing). The chemical-specific approach 

to aquatic life toxics control relies on numeric water quality 

criteria in State standards and interpretations of State narrative 

standards to assess and control specific toxicants individually. 

Water Quality Standards and Criteria 

Where specific numerical criteria for a chemical or biological 

parameter (such as toxicity) are absent, compliance with water 

quality standards must be based on the general narrative criteria 

and on protection of the designated uses. For many pollutants, 

EPA’s recommended criteria may be used, or criteria may be 

developed using data from the Integrated Risk Information System, or data on the toxicological effects of the pollutant found 

either in the literature or required of a discharger. 

Aquatic impacts occur not only from the magnitude of a pollutant, but also from the duration and frequency with which criteria 

are exceeded. EPA’s recommended aquatic life criteria for both 

individual toxicants and whole effluent toxicity are specified as 

two numbers: the criterion continuous concentration is applied 

as a 4-day average concentration; and the criterion maximum 

concentration is applied as an 1-hour average concentration. The 

frequency with which criteria are allowed to be exceeded depends on site-specific factors as explained in the text. 

Strictly speaking the term “criteria” means EPA guidance formally 

published under the authority of Section 304(a) of the Clean 

Water Act. The toxicity level recommendations have not been so 

published. However, they represent EPA’s carefully developed 

technical recommendation, and so are referred to in this document in the same manner as other criteria. 

EPA’s recommended criteria for whole effluent toxicity are as 

follows: to protect aquatic life against chronic effects, the ambient toxicity should not exceed 1.0 chronic toxic unit (TUc) to the 

most sensitive of at least three different test species. For protection against acute effects, the ambient toxicity should not exceed 

0.3 acute toxic units (TUc) to the most sensitive of at least three 

different test species. 

EPA has developed recommended human health criteria, which 

are called reference ambient concentrations (RACs). In the absence of EPA’s recommended criteria, States may calculate RACs 

based on the equations in the text. In addition, the need for 

sediment and biological criteria in State water quality standards is 

discussed. 

Effluent Characterization 

This chapter contains completely revised effluent characterization 

discussions and recommendations. It includes streamlined procedures (as compared to the original TSD) for predicting the likely 

impacts of toxic effluents on aquatic life and human health. 

Recommendations are provided for determining, either with or 

without actual effluent data, whether a discharge causes, has the 

reasonable potential to cause, or contributes to an excursion 

above a State water quality standard. These effluent characterization procedures can be performed in one step and do not include 

initial screening followed by definitive data generation as was 

recommended in the original TSD. 

The revised effluent characterization procedures for assessing potential human health impacts now include control of 

bioaccumulative chemicals. 

Exposure and Wasteload Allocation 

A goal of permit writers is to determine what effluent composition 

will protect aquatic organisms and human health. Exposure 

assessment includes an analysis of how much of the waterbody is 

subject to the exceedance of criteria, for how long, and how 

frequently. The first step is to evaluate the effluent plume dispersion. If mixing is not rapid and complete and if State standards 

allow a mixing zone, the wasteload allocation also must be based 

xiv 
on 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 

be used to 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 OF ABBREVIATIONS 

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 

CORMlX 1 

CTAP 

DESCON 

DFLOW 

DYNHYD4 

DYNTOX 

EXAMS-II 

FCM2 

FETRA 

FGETS 

FLOSTAT 

HHDFLOW 

HSPF 

MEXAMS 

MlNTEQA2 

MICH 

agricultural runoff model 

Channel Transport Model 

Complex Effluent Toxicity Information System 

Chemical Information System 

Cornell Mixing Zone Expert System 

Chemical Transport and Analysis Program 

computer program that estimates design conditions 

computer program that calculates biologically 

based design flows 

hydrodynamic model 

dynamic toxics model 

Exposure Analysis Modeling System 

WASP Food Chain Model 

Finite Element Transport Model 

Food and Gill Exchange of Toxic Substances 

U.S. Geological Survey computer program that 

estimates the arithmetic mean flow and 7Q10 of 

rivers and streams 

historic daily flow program 

Hydrologic Simulation Program - FORTRAN 

Metals Exposure Analysis Modeling System 

Equilibrium Metals Speciation Model 

Michigan River Model 

NPS 

PSY 

SARAH2 

SERATRA 

SLSA 

TODAM 

TOXIWASP 

TOXl4 

TOXIC 

UDKHDEN 

ULINE 

UMERGE 

UOUTPLM 

UPLUME 

WASP4 

WASTOX 

WQAB FLOW 

Nonpoint Source Model for Urban and Rural 

eas 

steady-state, two-dimensional plume model 

Arsurface water assessment model for back calculating reductions in biotic hazardous wastes 

Sediment Contaminant Transport Model 

Simplified Lake/Stream Analysis 

Transport One-Dimensional Degradation and Migration Model 

Chemical Transport and Fate Model 

a subset of WASP4 

Toxic Organic Transport and Bioaccumulation 

Model 

three-dimensional model used for single or multiple port diffusers 

uniform linear density flume model 

two-dimensional model used to analyze positively 

buoyant discharge 

cooling tower plume model adapted for marine 

discharges 

numerical model that produces flux-average dilutions 

water quality analysis program 

Estuary and Stream Quality Model 

water quality analysis system flow data subroutine 

. . XVIII 
GLOSSARY 

absolute toxicity is the toxicity of theeffluent without considering 

dilution. 

acute means a stimulus severe enough to rapidly induce an effect; 

in aquatic toxicity tests, an effect observed in 96 hours 

or less typically is considered acute. When referring to 

aquatic toxicology or human health, an acute affect is 

not always measured in terms of lethality. 

acute-to-chronic ratio (ACR) is the ratio of the acute toxicity of 

acutely 

an effluent or a toxicant to its chronic toxicity. It is used 

as a factor for estimating chronic toxicity on the basis of 

acute toxicity data, or for estimating acute toxicity on 

the basis of chronic toxicity data. 

toxic conditions are those acutely toxic to aquatic 

organisms following their short-term exposure within 

an affected area. 

acute toxicity endpoints (ATE) are toxicity test results, such as 

an LC50 (96 hours) and EC50 (48 hours), which describe 

a stimulus severe enough to rapidly induce an effect on 

aquatic organisms. 

additivity is the characteristic property of a mixture of toxicants 

that exhibits a total toxic effect equal to the arithmetic 

sum of the effects of the individual toxicants. 

ambient toxicity is measured by a toxicity test on a sample 

collected from a waterbody. 

antagonism is the characteristic property of a mixture of toxicants 

that exhibits a less-than-additive total toxic effect. 

antidegradation policies are part of each State’s water quality 

standards. These policies are designed to protect water 

quality and provide a method of assessing activities that 

may impact the integrity of the waterbody. 

aquatic community is an association of interacting populations 

of aquatic organisms in a given waterbody or habitat. 

averaging period is the period of time over which the receiving 

water concentration is averaged for comparison with 

criteria concentrations. This specification limits the 

duration of concentrations above the criteria. 

bioaccumulation is the process by which a compound is taken up 

by an aquatic organism, both from water and through 

food. 

bioaccumulation factor (BAF) is the ratio of a substance’s 

concentration in tissueversus its concentration in ambient 

water, in situations where the organism and the food 

chain are exposed. 

bioassay is a test used to evaluate the relative potency of a 

chemical or a mixture of chemicals by comparing its 

effect on a living organism with the effect of a standard 

preparation on the same type of organism. Bioassays 

frequently are used in the pharmaceutical industry to 

evaluate the potency of vitamins and drugs. 

bioavailability is a measure of the physicochemical access that a 

toxicant has to the biological processes of an organism. 

The less the bioavailability of a toxicant, the less its toxic 

effect on an organism. 

bioconcentration is the process by which a compound is absorbed 

from water through gills or epithelial tissues and is 

concentrated in the body. 

bioconcentration factor (BCF) is the ratio of a substance’s 

concentration in tissue versus its concentration in water, 

in situations where the food chain is not exposed or 

contaminated. For nonmetabolized substances, it 

represents equilibrium partitioning between water and 

organisms. 

biological assessment is an evaluation of the biological condition 

of a waterbody using biological surveys and other direct 

measurements of resident biota 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 reference conditions. Biological criteria 

serve as an index of aquatic community health. 

biological integrity is the condition of the aquatic community 

inhabiting unimpaired waterbodies of a specified habitat 

as measured by community structure and function. 

biological monitoring, also known as biomonitoring, describes 

the living organisms in water quality surveillance used to 

indicate compliance with water quality standards or 

effluent limits and to document water quality trends. 

Methods of biological monitoring may include, but are 

not limited to, toxicity testing such as ambient toxicity 

testing or whole effluent toxicity testing. 

biological survey or biosurvey is the collecting, processing, and 

analyzing of a representative portion of the resident 

aquatic community to determine its structural and/or 

functional characteristics. 

biomagnification is the process by which the concentration of a 

compound increases in species occupying successive 

trophic levels. 

cancer potency slope factor (q1* ) is an indication of a chemical’s 

human cancer-causing potential derived using animal 

studies or epidemiological data on human exposure. It 

is based on extrapolating high-dose levels over short 

periods of time to low-dose levels and a lifetime exposure 

period through the use of a linear model. 

chronic means a stimulus that lingers or continues for a relatively 

long period of time, often one-tenth of the life span or 

more. Chronic should be considered a relative term 

depending on the life span of an organism. The 

measurementof a chronic effectcan be reduced growth, 

reduced reproduction, etc., in addition to lethality. 

chronic toxicity endpoints (CTE) are results, such as a no 

observed effect concentration, lowest observed effect 

concentration, effect concentration, and inhibition 

concentration based on observations of reduced 

reproduction, growth, and/or survival from life cycle, 

partial life cycle, and early life stage tests with aquatic 

animal species. 

xix 
coefficient of variation (CV) is a standard statistical measure of 

the relative variation of a distribution or set of data, 

defined as the standard deviation divided by the mean. 

community component is a general term that may pertain to the 

biotic guild (fish, invertebrates, algae), the taxonomic 

category (order, family, genus, species), the feeding 

strategy (herbivore, omnivore, predator), or the 

organizational level (individual, population, assemblage) 

of a biological entity within the aquatic community. 

completely mixed condition means no measurable difference in 

the concentration of a pollutant exists across a transect 

of the waterbody (e.g., does not vary by 5 percent). 

continuous simulation model is a fate and transport model that 

criteria 

criteria 

uses time series input data to predict receiving water 

quality concentrations in the same chronological order 

as that of the input variables. 

continuous concentration (CCC) is the EPA national 

water quality criteria recommendation for the highest 

instream concentration of a toxicant or an effluent to 

which organisms can be exposed indefinitely without 

causing unacceptable effect. 

maximum concentration (CMC) is the EPA national 

water quality criteria recommendation for the highest 

instream concentration of a toxicant or an effluent to 

which organisms can be exposed for a brief period of 

time without causing an acute effect. 

critical life stage is the period of time in an organism’s lifespan 

in which it is the most susceptible to adverse effects 

caused by exposure to toxicants, usually during early 

development (egg, embryo, larvae). Chronic toxicity 

tests are often run on critical life stages to replace long 

duration, life-cycle tests since the most toxic effect 

usually occurs during the critical life stage. 

design flow is the flow used for steady-state wasteload allocation 

modeling. 

designated uses are those uses specified in water quality standards 

for each waterbody or segment whether or not they are 

being attained. 

discharge length scale is the square root of the cross-sectional 

area of any discharge outlet. 

diversity is the number and abundance of biological taxa in a 

specified location. 

effect concentration (EC) is a point estimate of the toxicant 

concentration that 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 that focuses on the chemical interaction 

between sediments and contaminants. 

final acute value (FAV) is an estimate of the concentration of the 

toxicant corresponding to a cumulative probability of 

0.05 in the acute toxicity values for all genera for which 

acceptable acute tests have been conducted on the 

toxicant. 

frequency is how often criteria can be exceeded without 

unacceptably affecting the community. 

genotoxic is the ability of a substance to damage an organism’s 

genetic material (DNA). 

harmonic mean flow is the number of daily flow measurements 

divided by the sum of the reciprocals of the flows. That 

is, it is the reciprocal of the mean of reciprocals. 

inhibition concentration (1C) is a point estimate of the toxicant 

concentration that would cause given a percent reduction 

(e.g., lC25) in a nonlethal biological measurement of the 

test organisms, such as reproduction or growth. 

lethal concentration is the point estimate of the toxicant 

concentration thatwould be lethal to a given percentage 

of the test organisms during a specific period. 

lipophilic is a high affinity for lipids (fats). 

load allocations (LA) are the portion of a receiving water’s total 

maximum daily toad that is attributed either to one of its 

existing or future nonpoint sources of pollution or to 

natural background sources. 

tognormal probabilistic dilution model calculates the 

probability distribution of receiving water quality 

concentrations from the lognormal probability 

distributions of the input variables. 

log P (also expressed as log kow or as n-octanal/water 

partition coefficient) is the ratio, in a two-phasesystem 

of n-octanol and water at equilibrium, of the 

concentration of a chemical in the n-octanol phase to 

that in the water phase. 

lowest observed adverse effect level (LOAEL) is the lowest 

concentration of an effluent or toxicant that results in 

statistically significant adverse health effects as observed 

in chronic or subchronic human epidemiology studies 

or animal exposure. 

magnitude is how much af a pollutant (or pollutant parameter 

such as toxicity), expressed as a concentration or toxic 

unit is allowable. 

minimum level (ML) refers to the level at which the entire 

analytical system gives recognizable mass spectra and 

acceptable calibration points when analyzing for 

pollutants of concern. This level corresponds to the 

lowest point at which the calibration curve is determined. 

mixing zone is an area where an effluent discharge undergoes 

initial dilution and is extended to cover the secondary 

mixing in the ambient waterbody. A mixing zone is an 

allocated impact zone where water quality criteria can 

be exceeded as long as acutely toxic conditions are 

prevented. 

Monte Carlo simulation is a stochastic modeling technique that 

involves the random selection of sets of input data for 

use in repetitive model runs in order to predict the 

probability distributions of receiving water quality 

concentrations. 
no observed adverse effect level (NOAEL) is a 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) is the 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) is the concentration of a 

toxicant or the parameter toxicity in the receiving water 

after mixing (formerly termed “instream waste 

concentration” [IWC]). 

recurrence interval is the 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) is the 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 is the toxicity of the effluent when it is mixed 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 is the 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 is a 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., 1 OO/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. 

xxi 
wasteload allocatlon @WA) is the 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 

The purpose of this revised Technical Support Document (TSD) for 

Water Quality-based Toxics Control is to provide the most current 

procedural recommendations and guidance for identifying, analyzing, and controlling adverse water quality impacts caused by 

toxic discharges to the surface waters of the United States. The 

original TSD was published in September 1985. Since then, the 

Clean Water Act (CWA) was amended in 1987 with an emphasis 

on controlling toxic pollutants. New policies and regulations 

have been promulgated and a vast amount of knowledge and 

experienced has been gained in controlling toxic pollutants. Because of these changes, EPA revised and updated the TSD. 

This guidance document is intended to support the implementation of the CWA water quality-based approach to toxics control. 

As such, the recommendations and guidance found in this document are not binding and should be used by regulatory authorities with discretion. The guidance in this document has been 

developed as the most current representation of knowledge in the 

field of assessment and control of toxic discharges. Some of the 

guidance in this document is based on ongoing research and 

development (bioaccumulation methods, Chapter 3) and should 

not be used until the procedures are finalized. 

Background 

The EPA surface water toxics control program, represented diagrammatically in the figure, relies on portions of the national 

pretreatment program, the effluent limitations guidelines program, the sludge program, the combined sewer overflow program, 

the stormwater management program, the 304(l) program, the 

water quality standards program, and the National Pollutant Discharge Elimination System (NPDES) program. States are authorized 

by EPA to implement certain portions of the national toxics control program, such as the NPDES program. Scientific and technical guidance is developed and published by EPA to assist the 

States. EPA is required by the CWA and federal regulations to play 

an oversight role to ensure that States authorized to implement 

various program requirements do so in accordance with federal 

regulations. 

States are given discretion in the CWA to establish and implement 

water quality standards. As such, there may be differences in 

toxics control programs between States. EPA’s oversight role is to 

ensure that each State’s program is technically sound and that 

each State fully implements its program. 

Throughout the evolution of the toxics control program, EPA has 

provided guidance concerning new program initiatives, statutory 

developments, and regulatory requirements. In 1980, EPA emphasized in its preamble to NPDES regulations (45 FR 33520) that 

NPDES permit limitations must reflect the most stringent of technology-based, water quality-based controls, or other standards 

required by the CWA (e.g., ocean discharge requirements under 

Section 403 and toxics standards or prohibition under Section 

307[a]). EPA reiterated the significance of surface water toxics 

control in 1984 through the publication of its national policy 

statement entitled, “Policy for the Development of Water QualityBased Permit Limitations for Toxic Pollutants” (49 FR 9016, March 

9, 1984). EPA recommended the use of “biological techniques as 

a complement to chemical-specific analyses to assess effluent 

discharges and express permit limitations” (49 FR 9017). The 

preamble to additional regulations promulgated in 1984 (49 FR 

37998) stressed the importance of establishing effluent limitations in NPDES permits to control toxic pollutants. Regulatory 

provisions promulgated on June 2, 1989 (54 FR 23868), clarify EPA’s 

surface water toxics control program and the use of whole effluent 

toxicity, and implement CWA Section 304(l) concerning the 

identification of impaired waters and the development of individual 

control strategies. 

The control of toxic discharges to the Nation’s waters is an 

important objective of the CWA. To effectively accomplish this 

objective, EPA recommends the use of an integrated water quality-based approach for controlling toxic discharges. EPA’s integrated “standards to permits” approach, illustrated in the figure, 

starts with water quality criteria, objectives, and standards and 

results in NPDES permit limits to control toxic pollutants through 

the use of both chemical-specific and whole effluent toxicity 

limitations. Limitations are essential for controlling the discharge 

of toxic pollutants to the Nation’s water. Once NPDES permit 

limits are set, compliance is essential. Compliance can be ascertained by continual routine monitoring of effluent quality. Water 

quality-based effluent limitations when developed in accordance 

with the procedures in this document, will protect water quality 

and prevent the violation of State water quality standards. 

. . . XXIII 
Overview of the Water Quality-based “Standards to Permits” Process for Toxics Control 

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 is called 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, EPA policy 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 

health. The use of toxicity testing and whole effluent toxicity 

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 FOR AQUATIC 

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 criteria consists of two values-an acute 

and a chronic value. Criteria are developed using the latest 

scientific knowledge on the kind and extent of identifiable effects 

on organisms, such as plankton, fish, shellfish, wildlife, and plant 

life, which may be expected from the presence of pollutants in 

any body of water. Water quality criteria also reflect the concentration and dispersal of pollutants, or their byproducts, through 

biological, physical, and chemical processes, and the effects of 

pollutants on biological community diversity, productivity, and 

stability of the receiving water [1] They can be used to assess and 

control a variety of water quality impacts. Chapter 2 provides a 

more detailed discussion of the derivation of numeric criteria. 

Recommendations for using chemical-specific data to determine 

which individual toxicants need to be controlled are found in 

Chapter 3. Legal requirements, including chemical-specific limits 

in permits, are found in Chapter 5. 

1.2.1 Correlation of Chemical-specific Measurements to Actual 

Receiving Water impacts 

EPA has conducted a series of studies to determine whether its 

water quality criteria concentrations are protective of aquatic life 

in receiving water systems. The first study was conducted at 

Shayler Run, Ohio, to evaluate the applicability of laboratorygenerated toxicity data to a natural stream artificially dosed with 

copper to provide steady concentrations [3]. The results of the 

study indicate that several characteristics of site-specific water 

quality affect the toxicity of copper. The results also indicate that 

avoidance of elevated concentration areas by instream organisms 

can produce observable ecological changes at concentrations 

below those found to be harmful in laboratory toxicity tests. No 

instream effects were observed at continuous exposure concentrations near EPA’s current chronic criterion, applied at the water 

hardness of Shayler Run. 

Studies performed on experimental streams at EPA’s Monticello 

Ecological Research Station (MERS) indicate good agreement between EPA’s criteria concentrations and the instream concentrations producing aquatic life effects under steady exposure conditions [4-13]. EPA’s water quality criteria are not threshold levels 

above which definite measurable instream effects are always expected. Rather, the criteria embody conservative assumptions 

such that small excursions above the criteria should not result in 

measurable environmental impacts upon the biota. The data 

indicate that if the ambient water quality criteria are met, then the 

biota in the receiving water system will be protected from unacceptable impacts caused by the chemical of concern. The studies 

conducted by MERS are described in greater detail in Box 1-1 and 

Tables 1-1 and 1-2. 

1.2.2 Chemical-specific Analytical Method Precision 

Tables 1-3 to 1-5 illustrate the types of precision commonly seen 

in inorganic, organic, and nonmetal inorganic chemical analyses 

that are routinely used for determining concentrations of specific 

pollutants in effluents. These tables show the observed variability. 

The variability of chemical measurements increases as one approaches the limit of detectability for a chemical. Table 1-3 shows 

the interlaboratory precision of 10 metals. The coefficient of 

variation (CV), defined as the standard deviation divided by the 

mean x 100, for these analyses ranges from 18 percent to 

129 percent [15]. Table 1-4 shows the interlaboratory precision 

Box 1-1. Correlation of Chemical-specific Criteria to Instream Impacts 

In studying the field applicability of EPA’s water quality criteria in freshwater systems, MERS (Monticello 

Ecological Research Station) conducted studies in experimental streams [4-14] to determine the level of 

protection provided by the individual chemical criteria. Each of the streams was one-quarter mile long with 

alternating mud-bottomed pools and rocky riffles. Fish were stocked into the streams to a known population 

density while other plants and animals were the result of natural colonization. 

The chemicals studied were ammonia, chlorine, chlorine combined with ammonia, selenium, and pentachlorophenol. Some studies were conducted during a summer (pentachlorophenol) while others continued for more 

than 2 years (selenium IV). Tables 1-1 and 1-2 show sample data on ammonia and ammonia combined with 

chlorine. In all experiments, the streams were dosed continuously with the chemical(s) being studied and the 

biological effects were determined statistically by a comparison to the control streams. The concentration at 

which biological effects occurred were then compared to the EPA criteria continuous concentration (CCC) for 

that compound. 

With the exception of chlorine in the presence of ammonia, the data from the other experiments indicate that 

slight or no effects were found in the streams at the CCC. This indicates that the CCC is providing chronic 

protection at the recommended concentration for that particular chemical. In the case of chlorine combined 

with ammonia, a substantial impact was found, but only on one species, the channel catfish. Because the CCC is 

designed to protect most, but not all of the species all of the time (see discussion in Chapter 2 on EPA Ambient 

Water Quality Criteria), slight impacts may be expected under continuous exposure conditions. 

2 
Table 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 

Od I 0 

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 

Table l-2. Effects in Streams Exposed to Ammonia Discarded’ 

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 

Fish 

Channel 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 IS obtained 

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 CFR Part 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. EPA has 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 as fertilization, 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 lethal concentrations (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 is the toxicant concentration 

that would cause an adverse effect upon a certain percentage of 

the test organisms, (e.g., EC1 0 or EC&. 

In chronic toxicity tests, the exposure duration in the EPA testing 

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 

4 
96 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 ECs are 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 this 
reason, 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 is approximately 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, = 1 OO/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]. 

6 
TU,. 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 EPA toxicity 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. EPA agrees 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. 

EPA believes 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, EPA disagrees 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. EPA acknowledges 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 tmuty no 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 

10 
ment 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 is the 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 is when 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 is sometimes 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 IC25s could 

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 EPA short-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.) 

Onsite nmus 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) is suspected 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 is for 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 is similar” [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 FR 30082 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#h vems Static and Renewal Toxicity 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 is the 

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 Samplinfl vems 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: 

13 
Table 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 SDS r 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 Copper AS 

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 Copper AS 

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 .OO mg/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 

14 
Table 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 assess variable effluents. Appendix F details 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 assess the 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 

16 
certain toxicants. Another species, Daphnio magna, is very 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 is true. 

The results of whole effluent tests using five marine/estuarine 

short-term chronic test methods also indicate that no species or 

test method is always 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-Chronic Ratio 

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 an acute toxicity 

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 

Daphnra polex 

-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 Efi luent 

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 

17 
must 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 ACRs for 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 acute and 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 FOR AQUATIC 

AND 

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 assess the 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 commu18 
Tabk 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 is the 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 

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

1.5 INTEGRATION OF THE WHOLE EFFLUENT, 

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 Undtatiions of 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. EPA acute 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 less expensive 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. As a 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 

-Assesses quality 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 imporApproacn 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. If there 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 acute or 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 assess toxicity 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 

23 
the 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, and Spw@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 TOXIC as 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 

24 
may 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 tissue contamination 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 of tfwlth 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’s approach 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 is experimental 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 

1. U.S. EPA. 1987. Quality Criteria for Woter. Qffice of Wter 

Regulations and Standards, Washington, DC. EPA 440/5- 

86-001. 

2. U.S. EPA. 1983. Water Quo&y Standurds Hondbook. Office of 

Water Regulations and Standards, Washington, DC. 

3. Ceckler, J.R., W.B. Horning, T.N. Neiheisel, Q.H. Pickering, 

E.L. Robinson, and C.E. Stephan. 1976. Validity of fuborotory Tests for Predicting Copper Toxicity in Streams. EPA 600/ 

3-76-l 16. 

4. Hedke, S., C. West, K.N. Allen, T. Norberg-King, and D. 

Mount. 1986. The Toxicity of PCP to Aquatic Organisms 

Under Naturally Varying and Controlled Environmental 

Conditions. Env. Toxicol. Chem. 5531-42. 

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. 

6. Yount, I.D., and J.E. Richter. 1986. Effects of Pentachlorophenol on Periphyton Communities in Outdoor Experimental 

Streams. Arch. Environ. Contom. Toxicol. 15:51-60. 

7. Zischke. 1985. Effects of Pentachlorophenol on Invertebrates 

and Fish in Outdoor Experimental Channels. Aquatic Toxicology 7137-58. 

8. Arthur, J.W., C.W. West, K.N. Allen, and SF. Hedtke. 1987. 

Seasonal Toxicity of Ammonia to Five Fish and Nine Invertebrate Species. Bull. Environ. Cont. Tox. 38(2):324-31. 

9. Herman&, R.O., S.F. Hedtke, I.W. Arthur, R.W. Andrew, and 

J.C. Helgen. 1987. Ammonia Effects on Microinvertebrates 

and Fish in Outdoor Experimental Streams. Environ. Pollut. 

47~249-83. 

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. EPA 625/8-87/013. 

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. 

26 
25. 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 Results of EPA’s Complex 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 Response on 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 Tests for Predicting Biological Impct, Five Mile Creek, 

Birmingham, Alabama. U.S. EPA. EPA 600/8-85/015. 

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, West Virginia. U.S. EPA. 

EPA 600/3-85/071, 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 Effects of 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 Test Guidelines. 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 FR 30082. 

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. EPA 440/5-90-004, April 1990. 

56. Plafkin, J.L., et. al. 1989. Ropid Bioassessment Protocols for Use 

In Streams and Rivers. Office of Water Regulations and 

Standards. EPA 444!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 of Aquatic 

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. EPA 670/4-731001. 

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. EPA 905/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 Water Quality 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 CFR 131.10(a)). 

EPA’s regulations emphasize the uses specified in CWA Section 

101(a)(2), but do not preclude other beneficial uses and 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 Use Attainability 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 CFR 131.12 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 CFR 131.12(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 is existing, 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” (Section 131.12(a)(2)) protects the water quality in waters whose quality is better than that necessary to protect “fishable/ 

swimmable” uses of the waterbody. 40 CFR 131.12(a)(2) requires 

that certain procedures be followed and certain showings be 

made before lowering water quality in high-quality waters. These 

showings may be called an “antidegradation review.” In no case 

may water quality on a Tier II waterbody be lowered to the level at 

which existing uses are impaired. The Tier II protection usually is 

applied on a parameter-by-parameter basis (called the definitional approach to Tier II). This approach is applied on a case-bycase basis so that, if the level of any parameter is better than water 

quality standards for that waterbody, then an antidegradation 

review will be performed for any activity that could reduce the 

level of that parameter. 

Outstanding national resource waters (ONRWs) are provided the 

highest level of protection under the antidegradation policy (Tier 

Ill); no degradation is allowed. ONRWs include the highestquality waters of the United States. However, the ONRW 

antidegradation classification also offers special protection for 

waters of “exceptional ecological significance,” i.e., those 

waterbodies that are important, unique, or sensitive ecologically, 

but whose water quality, as measured by the traditional parameters such as dissolved oxygen or pH, may not be particularly 

high. Waters of exceptional ecological significance may also 

include waters whose characteristics cannot be described adequately by traditional parameters (such as wetlands and estuaries). 

States may, at their discretion, adopt certain policies in their 

standards affecting the application and implementation of standards. For example, policies concerning mixing zones, variances, 

low-flow exemptions, and schedules of compliance for water 

quality-based permit limits may be adopted. Although these are 

areas of State discretion, EPA retains authority to review and 

approve or disapprove such policies (see 40 CFR 131.13). Guidance on these subjects is available from EPA’s Office of Water 

Regulations and Standards, Criteria and Standards Division. 

2.1.2 Water Quality Standards and State Toxics Control 

Programs 

Applicable requirements for State adoption of water quality criteria for toxicants vary depending upon the toxicant. The reason 

for this is that the 1983 water quality standards regulation and the 

1987 amendments to the CWA (Pub. L. 100-4) include more 

specific requirements for the particular toxicants listed in CWA 

Section 307(a). For regulatory purposes, EPA has translated the 

65 compounds and families of compounds listed in Section 307(a) 

into 126 specific substances that EPA refers to as priority toxic 

pollutants. The 126 priority toxic pollutants are listed in Appendix 

A of 40 CFR Part 423. Because of the more specific requirements 

for priority toxic pollutants, it is convenient to organize the requirements applicable to State adoption of criteria for toxicants 

into three categories: 

• Requirements applicable to priority toxic pollutants that 

have been the subject of CWA Section 304(a)(1) criteria 

guidance 

• Requirements applicable to priority toxic pollutants that 

have not been the subject of CWA Section 304(a)(1) criteria 

guidance and 

• Requirements applicable to all other toxicants (i.e., 

nonpriority toxic pollutants). 

The criteria requirements applicable to priority toxic pollutants 

(i.e., the first two categories above), are specified in CWA Section 

303(c)(2)(B). On December 2, 1988, EPA sent “Guidance for 

State Implementation of Water Quality Standards for CWA Section 303(c)(2)(B)” to each of its Regions and to each State water 

pollution control agency. The guidance contained three options 

for implementing the new numeric criteria requirements of the 

Act: (1) adopt Statewide numeric criteria in standards for all those 

priority toxic pollutants for which EPA has published national 

criteria; (2) adopt numeric criteria for only those priority toxic 

pollutants and those stream segments where the discharge or 

presence of the pollutant could reasonably be expected to interfere with designated uses; or (3) adopt a specific procedure in the 

standards to “translate” the State’s narrative “free from toxics” 

standard to derived numeric criteria. 

The transmittal memorandum for the Section 303(c)(2)(B) national guidance expresses the Office of Water position regarding 

priority toxic pollutants that may “reasonably be expected” to 

interfere with designated uses. That memorandum and guidance 

established a rebuttable presumption that any information indicating that such pollutants are discharged or present in surface 

waters (now or in the future) is sufficient justification to require 

adoption or derivation of numerical criteria. The goal is not just to 

identify pollutants that are already impacting surface waters, but 

rather to identify pollutants that may be impacting surface waters 

now, or have the potential to do so in the future. Lack of detailed 

or widespread monitoring data is not an acceptable basis to omit 

numerical (or derived numerical) criteria from water quality standards under Options 2 and 3. Even a limited amount of monitoring data indicating the discharge or presence of priority toxic 

pollutants in surface waters is sufficient basis to conclude that 

numerical (or derived numerical) criteria are necessary. 

Where States select an Option 2 or 3 approach, States must 

include, as part of the rationale supporting the adopted standards, the information used in determining which priority toxic 

pollutants require criteria. Where there is uncertainty about the 

need for criteria for specific priority toxic pollutants, the State 

should adopt (or derive) criteria for such pollutants so as to err on 

the side of environmental protection and pollution prevention. 

This approach is appropriate given the general lack of monitoring 

data for priority toxic pollutants; it will provide maximum protection 

to the environment by anticipating, rather than reacting to, water 

quality problems. 

For priority toxic pollutants for which EPA has not issued Section 

304(a)(1) criteria guidance, CWA Section 303(c)(2)(B) requires 

States to adopt criteria based on biological monitoring or assessment methods. The phrase “biological monitoring or assessment 

methods” includes (1) whole effluent toxicity control methods, 

(2) biological criteria methods, or (3) other methods based on 

biological monitoring or assessment. The phrase “biological 

monitoring or assessment methods” in its broadest sense also 

includes criteria developed through translator procedures. This 

broad interpretation of that phrase is consistent with EPA’s policy 

of applying chemical-specific, biological, and whole effluent toxicity methods independently in an integrated toxics control program. It also is consistent with the intent of Congress to expand 

30 
State standards programs beyond chemical-specific approaches. 

Where EPA has not issued Section 304(a) criteria guidance, but 

available laboratory toxicity (bioassay) data are sufficient to support derivation of chemical-specific criteria, States should consider 

deriving and adopting numeric criteria for such priority toxic 

pollutants. This is particularly important where other components of a state’s narrative Criterion implementation procedure 

(e.g., whole effluent toxicity controls or biological criteria) may 

not ensure full protection of designated uses. For some pollutants, 

a combination of chemical-specific and other approaches is necessary (e.g., pollutants where bioaccumulation in fish tissue or 

water consumption by humans is a primary concern). 

Criteria requirements applicable to toxicants that are not priority 

toxic pollutants (i.e., the third category above), are specified in 

the 1983 water quality standards regulation (see 40 CfR 131.11). 

Under these requirements, States must adopt criteria based on 

sound scientific rationale that cover sufficient parameters to protect designated uses. Both numeric and narrative criteria are 

addressed by these requirements. 

Numeric criteria are required where such criteria are necessary to 

protect designated uses. Numeric criteria to protect aquatic life 

should be developed to address both short-term (acute) and 

long-term (chronic) effects. Saltwater species, as well as freshwater species, must adequately be protected. Adoption of numeric 

criteria is particularly important for toxicants known to be impairing surface waters and for toxicants with potential human health 

impacts (e.g., those with high bioaccumulation potential). Human health should be protected from exposure resulting from 

consumption of water and fish or other aquatic life (e.g., mussels, 

crayfish). Numeric water quality criteria also are useful in addressing nonpoint source pollution problems. 

In evaluating whether chemical-specific numeric criteria for toxicants are required, States should consider whether other approaches (such as whole effluent toxicity criteria or biological 

controls) will ensure full protection of designated uses. As mentioned above, a combination of independent approaches may be 

required in some cases to support the designated uses and comply with the requirements of the water quality standards regulation (e.g., pollutants where bioaccumulation in fish tissue or water 

consumption by humans is a primary concern). 

To supplement numeric criteria for toxicants, all States also have 

adopted narrative criteria for toxicants. Such narrative criteria are 

statements that describe the desired water quality goal, such as 

the following: 

All State waters must, at all times and flows, be free from 

substances that are toxic to humans or aquatic life. 

EPA considers that the narrative criteria apply to all designated 

uses at all flows unless specified otherwise in a State’s water 

quality standards. EPA also believes that no acutely toxic condition may exist in any State waters regardless of designated use (54 

FR 23875). 

Narrative criteria can be the basis for establishing chemical-specific limits for waste discharges where a specific pollutant can be 

identified as causing or contributing to the toxicity and the State 

has not adopted chemical-specific numeric criteria. Narrative 

criteria also can be the basis for establishing whole effluent toxicity controls required by EPA regulations at 40 CFR 122.44(d)(1)(v). 

To ensure that narrative criteria for toxicants are attained, the 

water quality standards regulation requires States to develop 

implementation procedures (see 40 CFR 131.11(a)(2)). Such 

implementation procedures (Box 2-1) should address all mechanisms used by the State to ensure that narrative criteria are 

attained. Because implementation of chemical-specific numeric 

criteria is a key component of State toxics control programs, 

narrative criteria implementation procedures must describe or 

reference the State’s procedures to implement such chemicalspecific numeric criteria (e.g., procedures for establishing chemical-specific permits limits under the NPDES permitting program). 

Implementation procedures also must address State programs to 

control whole effluent toxicity and may address programs to 

implement biological criteria, where such programs have been 

developed by the State. Implementation procedures therefore 

serve as umbrella documents that describe how the State’s various toxics control programs are integrated to ensure adequate 

protection for aquatic life and human health and attainment of 

the narrative toxics criterion. In essence, the procedure should 

apply the “independent application” principle, which provides for 

independent evaluations of attainment of a designated use based 

on chemical-specific, whole effluent toxicity, and biological criteria methods (see Chapter 1, Reference 56). 

EPA encourages, and may ultimately require, State implementation procedures to provide for implementation of biological criteria. However, the regulatory basis for requiring whole effluent 

toxicity controls is clear. EPA regulations at 40 CFR 122.44(d)(1)(v) 

require NPDES permits to contain whole effluent toxicity limits 

where a permittee has been shown to cause, have the reasonable 

potential to cause, or contribute to an in-stream excursion of a 

narrative criterion. Implementation of chemical-specific controls 

also is required by EPA regulations at 40 CFR 122.44(d)(1). State 

implementation procedures should, at a minimum, specify or 

reference methods to be used in implementing chemical-specific 

and whole effluent toxicity-based controls, explain how these 

methods are integrated, and specify needed application criteria. 

In addition to EPA’s regulation at 40 CFR Part 131, EPA has regulations at 40 CFR 122.44 that cover the National Surface Water 

Toxics Control Program. These regulations intrinsically are linked 

to the requirements to achieve water quality standards, and specifically address the control of pollutants both with and without 

numeric criteria. For example, Section 122.44(d)(1)(vi) provides 

the permitting authority with several options for establishing 

effluent limits when a State does not have a chemical-specific 

numeric criteria for a pollutant present in an effluent at a concentration that causes or contributes to a violation of the State’s 

narrative criteria. 

2.2 GENERAL CONSIDERATIONS 

2.2.1 Magnitude, Duration, and frequency 

As stated earlier, criteria are specifications of water quality designed to ensure protection of the designated use. EPA criteria are 

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. EPA criteria or guidance consist of three components: 

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

l Duration-The period of time (averaging period) over which 

the instream concentration is averaged 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 of Aquatic 

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. Because of 

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 accumula33 
tion 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 FOR AQUATIC LIFE 

PROTECTION 

2.3.1 Devekyunent f%ocess for Criieria 

The development of national numerical water quality criteria for 

the protection of aquatic organisms is a 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. If sufficient data on the ratio of acute to 

chronic toxicity concentrations are available, they are used to 

derive the chronic or long-term exposure criteria. If justified, 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 Life and Human He&h (see 45 FR 79341, November 28, 1980, and 50 FR 30784, july 29, 1985). 

2.3.2 Magnitude for Sfngle Chemicals 

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 EPA water 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 .O TU, 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#ie hmhakwnd 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 as typically 

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. 

EPA established 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. 

EPA thus 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, 1 Oyear 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 FOR HUMAN 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 Ffsh consunrptlon V8hles 

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 Fish and 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. 

37 
Table 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, 151 are listed according to log P value 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. Risk assessment 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 IRIS is available 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 IRIS or scheduled 

for review. Information regarding the submission of studies of 

chemicals may be obtained from the IRIS Information 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 IRIS contains 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 IRIS contains the Agency’s most recent quantitative risk 

assessment values, current IRIS values 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 IRIS data. 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 Calculatiug RN& 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 Y MCLs, fish 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 

39 
deleterious 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 IRIS are 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 for Gucinugms 

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 ., 1 O- fc nsk \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 1 Oe6) 

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 1 Oms to 1 Oe7 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 IRIS are 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. Risk levels of 1 Oe5, 1 Om6, and 1 Om7 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 1 Oe6 may 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 (1 O-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 Risk Assessments in the Absence 

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 EPA risk 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 Deriving Refemce Tissue fducentratitms fur Monitoring 

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’s Assessing Human Health Risks from Chemically 

Contaminated Fish and 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 

RTCs can 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 Bases fur Bimriteda 

The primary statutory basis for EPA’s policy that States should 

develop biocriteria is found 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 CFR Part 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 . . .“) 

41 
and 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 Dmelopntent and ih@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 Cosront ilevel- 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, EPA also 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. 

42 
Equilibrium 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 Sedhnent Criteria 

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. EPA believes 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 

Scieuce Advisory Boiwd Review 

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. Water QuulkyStundards 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 Effects of 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 of Aquatic 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 for Assessing 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) 

Except as provided in this subparagraph, when the 

permitting authority determines, using the procedures 

in paragraph (d)(1)(ii) of this section, toxicity testing 

data, or other information, that a discharge causes, has 

the reasonable potential to cause, or contributes to an 

in-stream excursion above a narrative criterion within 

an applicable State water quality standard, the permit 

must contain effluent limits for whole effluent toxicity. 

Limits on whole effluent toxicity are not necessary where 

the permitting authority demonstrates in the fact sheet 

or statement of basis of the NPDES permit, using the 

procedures in paragraph (d)(1)(ii) of this section, that 

chemical-specific limits for the effluent are sufficient to 

attain and maintain applicable numeric and narrative 

State water quality standards. 

Under this regulation, permitting authorities need to investigate 

for the existence of whole effluent toxicity in effluents. If the 

permitting authority can demonstrate that control of specific 

chemicals is sufficient to control toxicity to the point of achieving 

compliance with the water quality criteria, then chemical-specific 

permit limits alone will be sufficient to comply with the regulation. 

40 CFR 122.44(d)(1)(vi) 

Where a State has not established a water quality criterion for a specific chemical pollutant that is present in 

an effluent at a concentration that causes, has the 

reasonable potential to cause, or contributes to an excursion above a narrative criterion within an applicable 

State water quality standard, the permitting authority 

must establish effluent limits using one or more of the 

following [three] options: . . . . 

Under this regulation, permitting authorities need to investigate 

for the existence of specific chemicals in effluents for which the 

State has not adopted numeric criteria, but which may be contributing to aquatic toxicity or impairment of human health. 

Narrative criteria apply when numeric criteria do not protect all 

the designated or existing uses. For example, the narrative 

criteria need to be used to protect human health if a State has 

only adopted a numeric criteria for protecting aquatic life. Conversely, the narrative criteria need to be used to protect aquatic 

life if a State has only adopted a numeric criteria for protecting 

human health. Once the permitting authority determines that 

one or more specific chemicals in an effluent must be controlled, 

the authorities can use EPA’s national criteria, develop their own 

criteria, or control the pollutant through use of an indicator 

pollutant, as provided in subparagraph (d)(1)(vi). In any case, 

the permitting authority will need to characterize the effluent in a 

manner consistent with the selected approach for controlling the 

pollutant. 

3.1.2 Background fur Toxic Effects Assessments on Aquatic 

life and Human Health 

Aquatic toxicity effects can be characterized by conducting a 

general assessment of the effluent, or by measuring effluent 

toxicity or concentrations of individual chemicals and comparing 

these measurements to the expected exposure concentrations in 

the receiving water. The “receiving water concentration” (RWC) 

is the measured or projected exposure concentration of a toxicant 

or the parameter toxicity (when dealing with the whole effluent 

toxicity) in the receiving water after mixing. The RWC is calculated at the edge of a mixing zone if such a zone is allowed by a 

State’s water quality standards. 

As with aquatic life protection, there are two possible approaches 

to characterizing effluents for human health effects: chemical-bychemical and whole effluent. However, only the chemical-bychemical approach currently is practical for assessing and controlling human health impacts. Appendix G discusses developing 

procedures for assessing human health impacts from whole effluents. 

A fundamental principle in the development of water qualitybased controls is that the RWC must be less than the criteria that 

comprise or characterize the water quality standards. With individual toxicants (or the parameter toxicity), the potential for 

toxicity in the receiving water is minimized where the RWC is less 

than the criterion continuous concentration (CCC), the criterion 

maximum concentration (CMC), and the reference ambient concentration (RAC). Toxicity becomes maximized where the RWC 

exceeds these criteria. Therefore, to prevent impacts to aquatic 

life or human health, the RWC of the parameter effluent 

toxicity or an individual toxicant (based on allowable dilution 

for the criterion) must be less than the most limiting of the 

applicable criterion, as indicated below. (The RAC as used 

throughout this chapter incorporates EPA human health criteria 

and State standards as well.) 

RWC < CCC (chronic aquatic life) 

RWC < CMC (acute aquatic life) 

RWC < RAC (human health) 

The water quality analyst will use the same basic components in 

the above-described relationship (i.e., critical receiving water flows, 

ambient criteria values, measures of effluent quality) for both 

effluent characterization and wasteload allocation (WLA) development, albeit from different perspectives. In the case of effluent 

characterization, the objective is to project receiving water concentrations based upon existing effluent quality to determine 

whether or not an excursion above ambient criteria occurs, or has 

the reasonable potential to occur. In developing WLAs, on the 

other hand, the objective is to fix the RWC at the desired criteria 

level and determine an allowable effluent loading that will not 

cause excursions above the criteria. 

Recommendations for projecting the RWC are described within 

this chapter. Chapter 4, Exposure Assessment and Wasteload 

Allocation, provides recommendations for determining allowable 

effluent loadings to achieve established ambient criteria and for 

calculating WLAs for establishing permit limits. The procedures 

described within Chapter 4 can also be used to calculate the 

dilution for analyses within Chapter 3. Chapter 5, Permit Requirements, describes the actual calculation of permit limits after effluent characterization and loadings, as well as WLAS, are complete. 

48